Chemical determinands

Volume 3 The Datasheets

Micro-organisms and Chemical and Physical Determinands

Contents

Introduction 1

Datasheets, Full Index, Alphabetical Listing 2

Datasheets Index, Sorted by Subject 7

2 Datasheets for Chemical and Physical Determinands 15

2.1 Inorganic determinands (except aesthetic determinands) 15 antimony 16 arsenic 18 asbestos 21 barium 24 beryllium 26 boron 28 bromate 31 cadmium 33 chloramines 36 chlorate 36 chlorine 39 chlorine dioxide 42 chlorite 45 chromium 48 copper 51 cyanide 54 cyanogen chloride 56 dichloramine 58 fluoride 60 iodine and iodide 63 lead 65 lithium 68 manganese 70 mercury 74 molybdenum 77 monochloramine 79 nickel 82 nitrate and nitrite 84 potassium permanganate 88 selenium 90 silver 93 tin 95 trichloramine 97 uranium 99

ii Draft Guidelines for Drinking-water Quality Management for New Zealand 2005

2.2 Organic chemicals (except aesthetic determinands and pesticides) 103 acrylamide 104 benzene 106 benzo[a]pyrene 108 bromochloroacetic acid 111 bromochloroacetonitrile 112 bromodichloromethane 114 bromoform 117 carbon tetrachloride 120 chloroacetones 122 chloroform 125 2-chlorophenol 128 chloropicrin 130 di(2-ethylhexyl)adipate (DEHA) 133 di(2-ethylhexyl)phthalate (DEHP) 135 dialkyltins 137 dibromoacetic acid 139 dibromoacetonitrile 141 dibromochloromethane 143 dichloroacetic acid 146 dichloroacetonitrile 149 1,2-dichlorobenzene 151 1,3-dichlorobenzene 153 1,4-dichlorobenzene 155 1,1-dichloroethane 157 1,2-dichloroethane 159 1,1-dichloroethene 161 1,2-dichloroethene 163 dichloromethane 165 2,4-dichlorophenol 168 dioxins 170 EDTA 172 epichlorohydrin 174 ethylbenzene 177 fluoranthene 179 formaldehyde 182 hexachlorobutadiene 185 monobromoacetic acid 187 monochloroacetic acid 188 monochlorobenzene 191 MX (3-chloro-4-dichloromethyl-5-hydroxy-2(5H)-furanone) 193 nitrilotriacetic acid 195 PCBs 197 polynuclear aromatic hydrocarbons 200 styrene 203 tetrachloroethene 206 toluene 208 tributyltin oxide 211 trichloroacetaldehyde 213 trichloroacetic acid 215 trichloroacetonitrile 217 trichlorobenzenes (total) 220 1,1,1-trichloroethane 222 trichloroethene 224 2,4,6-trichlorophenol 227 trihalomethanes 229 vinyl chloride 231 xylenes 234

Draft Guidelines for Drinking-water Quality Management for New Zealand 2005 iii

2.3 Pesticides 237 alachlor 238 aldicarb 240 aldrin/dieldrin 242 atrazine 244 azinphos methyl 247 bentazone 250 brodifacoum 252 bromacil 253 carbofuran 256 chlordane 258 chlorothalonil 261 chlorotoluron 263 chlorpyriphos 265 cyanazine 267 2,4-D 270 2,4-DB (4-(2,4-dichlorophenoxy)butyric acid) 273 DDT and its derivatives 275 diazinon 277 1,2-dibromo-3-chloropropane 279 1,2-dibromoethane 281 1,2-dichloropropane 283 1,3-dichloropropane 285 1,3-dichloropropene 286 dichlorprop 288 dimethoate 291 diquat 294 diuron 297 endosulfan 300 endrin 302 fenitrothion 304 fenoprop 305 glyphosate 308 heptachlor and heptachlor epoxide 310 hexachlorobenzene 312 hexazinone 314 isoproturon 317 lindane 319 malathion 321 MCPA 323 MCPB 325 mecoprop 327 metalaxyl 329 methamidophos 332 methomyl 333 methoxychlor 335 methyl parathion 337 metolachlor 339 metribuzin 342 molinate 345 oryzalin 346 oxadiazon 349 pendimethalin 350 pentachlorophenol 352 permethrin 355 phenylphenol 357 phorate 359 picloram 360 pirimiphos methyl 363 pirimisulfuron methyl 365 procymidone 368

iv Draft Guidelines for Drinking-water Quality Management for New Zealand 2005

propanil 369 propazine 371 propoxur 373 pyridate 375 pyriproxifen 376 quintozene 378 simazine 379 2,4,5-T 381 terbacil 384 terbuthylazine 386 thiabendazole 387 triclopyr 389 trifluralin 393 1080 395

2.4 Aesthetic determinands 397 aggressiveness 397 aluminium (Al3+) 399 ammonia (NH3 and NH4

+) 401 calcium (Ca2+) 403 chloride (Cl-) 404 colour 406 conductivity 407 hardness (total) 408 hydrogen sulphide (H2S) 410 iron (Fe2+ and Fe3+) 412 magnesium (Mg2+) 413 manganese 414 pH 415 sodium (Na+) 416 sulphate (SO4

2-) 418 taste and odour 419 temperature 421 total dissolved solids 422 turbidity 424 zinc (Zn2+) 425

Note: Determinands with possible health concerns but no MAV are listed in the inorganic, organic, pesticide or aesthetic datasheets, as appropriate.

Draft Guidelines for Drinking-water Quality Management for New Zealand 2005 v

Introduction Individual datasheets have been prepared for the microbiological, chemical and physical determinands listed in the Drinking-Water Standards for New Zealand 2005 (DWSNZ). These datasheets provide information on: • the sources of the determinands • how they are distributed in the environment • typical concentrations in drinking-water • how water can be treated to remove or inactivate them • how they can be analysed, and • a detailed explanation on how the MAV or guideline value was derived. This information should be useful for:

• water supply managers and drinking-water assessors trying to identify the source of a problem or how to solve it

• for analytical laboratories wanting to develop techniques to measure any of the determinands, and

• as a source of background information for the general public. All determinands appear in the full list. Some appear in more than one other list. For example, pesticides such as MCPB appear in the pesticide list and in the list of determinands that have a possible health concern, but no MAV; 2,4,6-trichlorophenol has a MAV, a GV and is a DBP so is listed according. The DWSNZ contains a list of synonyms for the names of the health significant chemical determinands and this should be consulted if the use of an alternative name for any of these determinands is suspected. These also appear in Appendix 6 (Volume 2 of the Guidelines). Note that the WHO Guidelines for Drinking-water Quality, 3rd edition, 2004 are available at: http://www.who.int/water_sanitation_health/dwq/gdwq3/en/print.html This includes their fact sheets, from which a lot of the information in these datasheets has been taken. For a list of pesticides registered in New Zealand, see: http://www.ermanz.govt.nz/hs/transfer-pesticides.asp Note: The analytical methods have not been updated as at September 2005.

Draft Guidelines for Drinking-water Quality Management for New Zealand 2005 1

Datasheets, Full Index, Alphabetical Listing

Entry See section

Acanthamoeba sp. 1.2 acrylamide 2.2 adenoviruses 1.4 Aeromonas 1.1 aggressiveness (ie, plumbosolvency) 2.4 alachlor 2.3 aldicarb 2.3 aldrin/dieldrin 2.3 aluminium 2.4 ammonia 2.4 anatoxin-a 1.3 anatoxin-a(s) 1.3 antimony 2.1 arsenic 2.1 asbestos 2.1 astroviruses 1.4 atrazine 2.3 azinphos methyl 2.3 Balantidium coli 1.2 barium 2.1 bentazone 2.3 benzene 2.2 benzo[a]pyrene 2.2 beryllium 2.1 boron 2.1 brodifacoum 2.3 bromacil 2.3 bromate 2.1 bromochloroacetic acid 2.2 bromochloroacetonitrile 2.2 bromodichloromethane 2.2 bromoform 2.2 cadmium 2.1 calcium 2.4 calciviruses (noroviruses) 1.4 Campylobacter 1.1 carbofuran 2.3 carbon tetrachloride 2.2 chloramines see mono, di and tri chlorate 2.1 chlordanes 2.3 chloride 2.4 chlorine dioxide 2.1 chlorine (free) 2.1 chlorite 2.1 chloroacetones 2.2

2 Draft Guidelines for Drinking-water Quality Management for New Zealand 2005

Entry See section

chloroform 2.2 2-chlorophenol 2.2 chloropicrin 2.2 chlorothalonil 2.3 chlorpyriphos 2.3 chlortoluron 2.3 chromium 2.1 coliforms, including faecal 1.1 coliphages 1.4 colour 2.4 copper 2.1 Cryptosporidium parvum 1.2 cyanide (total) 2.1 cyanazine 2.3 cyanobacteria 1.3 cyanogen chloride 2.1 Cyclospora 1.2 cylindrospermopsin 1.3 2,4-D 2.3 2,4-DB 2.3 DDT + isomers 2.3 di(2-ethylhexyl)adipate 2.2 di(2-ethylhexyl)phthalate 2.2 diazinon 2.3 1,2-dibromo-3-chloropropane 2.3 dibromoacetic acid 2.2 dibromoacetonitrile 2.2 dibromochloromethane 2.2 1,2-dibromoethane 2.2 dichloroacetic acid 2.2 dichloroacetonitrile 2.2 dichloramine 2.1 3,4-dichloroaniline see propanil 1,2-dichlorobenzene 2.2 1,3-dichlorobenzene 2.2 1,4-dichlorobenzene 2.2 1,2-dichloroethane 2.2 1,1-dichloroethane 2.2 1,1-dichloroethene 2.2 1,2-dichloroethene (cis and trans) 2.2 dichloromethane 2.2 2,4-dichlorophenol 2.4 1,2-dichloropropane 2.3 1,3-dichloropropane 2.3 1,3-dichloropropene (cis and trans) 2.3 dichlorprop 2.3 dieldrin see aldrin dimethoate 2.3 dioxins 2.2 diquat 2.3 diuron 2.3


Entry See section

EDTA 2.2

endosulfan 2.3 endotoxins 1.3 endrin 2.3 Entamoeba histolytica 1.2 enteroviruses 1.4 epichlorohydrin 2.2 Escherichia coli 1.1 ethylbenzene 2.2 ethylene dibromide see 1,2-dibromoethane fenitrothion 2.3 fenoprop 2.3 fluoranthene 2.2 fluoride 2.1 formaldehyde 2.2 Giardia intestinalis (lamblia) 1.2 glyphosate 2.3 hardness 2.4 helminths 1.5 hepatitis viruses 1.4 heptachlor and heptachlor epoxide 2.3 hexachlorobenzene 2.3 hexachlorobutadiene 2.2 hexazinone 2.3 homoanatoxin-a 1.3 hydrogen sulphide 2.4 iodine and iodide 2.1 iron 2.4 isoproturon 2.3 Isospora 1.2 Klebsiella 1.1 lead 2.1 Legionella 1.1 lindane 2.3 lithium 2.1 magnesium 2.4 malathion 2.3 manganese 2.1 MCPA 2.3 MCPB 2.3 mecoprop 2.3 mercury (total) 2.1 metalaxyl 2.3 methamidophos 2.3 methomyl 2.3 methoxychlor 2.3 methyl parathion 2.3 metolachlor 2.3 metribuzin 2.3 microcystins-(MC-LR toxicity equivalents) 1.3 Microsporidia 1.2


Entry See section molinate 2.3 molybdenum 2.1

monobromoacetic acid 2.2 monochloramine 2.1 monochloroacetic acid 2.2 monochlorobenzene 2.2 MX 2.2 Mycobacterium 1.1 Naegleria fowleri 1.2 nickel 2.1 nitrate 2.1 nitrilotriacetic acid (NTA) 2.2 nitrite 2.1 4-nitrophenol see methyl parathion nodularin 1.3 Norwalk virus see calciviruses odour see taste oryzalin 2.3 oxadiazon 2.3 PAH see polyaromatic hydrocarbons PCBs 2.3 pendimethalin 2.3 pentachloronitrobenzene see quintozene pentachlorophenol 2.3 permethrin 2.3 pH 2.4 phenylphenol 2.3 phorate 2.3 picloram 2.3 pirimiphos methyl 2.3 pirimsulfuron methyl 2.3 potassium permanganate 2.1 procymidone 2.3 propanil 2.3 propazine 2.3 propoxur 2.3 Pseudomonas aeruginosa 1.1 pyridate 2.3 pyriproxifen 2.3 quintozene 2.3 Rotavirus, para-rotaviruses and Reovirus 1.4 Salmonella 1.1 saxitoxins (as STX equivalents) 1.3 selenium 2.1 Shigella 1.1 silver 2.1 simazine 2.3 sodium 2.4 styrene 2.2 sulphate 2.4 2,4,5-T 2.3


Entry See section taste 2.4 temperature 2.4 terbacil 2.3

terbuthylazine 2.3 3,3�,4,4�-tetrachloroazobenzene see propanil tetrachloroethene 2.2 thiabendazole 2.3 tin 2.1 toluene 2.2 total dissolved solids 2.4 Toxoplasma 1.2 tributyltin oxide 2.2 trichloramine 2.1 trichloroacetaldehyde/chloral hydrate 2.2 trichloroacetic acid 2.2 trichloroacetonitrile 2.2 trichlorobenzenes (total) 2.2 1,1,1-trichloroethane 2.2 trichloroethene 2.2 2,4,6-trichlorophenol 2.2 triclopyr 2.3 trifluralin 2.3 trihalomethanes (THMs) see individual THMs turbidity 2.4 uranium 2.1 Vibrio 1.1 vinyl chloride 2.2 xylenes 2.2 Yersinia 1.1 zinc 2.4 1080 2.3


Datasheets Index, Sorted by Subject

1 Micro-organisms

1.1 Bacteria Aeromonas Campylobacter coliforms including faecal Escherichia coli Klebsiella Legionella Mycobacterium Pseudomonas aeruginosa Salmonella Shigella Vibrio Yersinia

1.2 Protozoa Acanthamoeba sp. Balantidium coli Cyclospora Cryptosporidium parvum Entamoeba histolytica Giardia intestinalis (lamblia) Isospora Microsporidia Naegleria fowleri Toxoplasma

1.3 Toxic algae and cyanotoxins Cyanobacteria anatoxin-a anatoxin-a(s) cylindrospermopsin endotoxins homoanatoxin-a microcystins-(MC-LR toxicity equivalents) nodularin saxitoxins (as STX equivalents)


1.4 Viruses adenoviruses astroviruses calciviruses (noroviruses) coliphages enteroviruses hepatitis viruses Norwalk virus see calciviruses Rotavirus, para-rotaviruses and Reovirus (Reoviridae)

1.5 Helminths helminth spp

2 Chemical and physical

2.1 Inorganic determinands (except aesthetic determinands) antimony arsenic asbestos barium beryllium boron bromate cadmium chloramines (see mono) chlorate chlorine (free) chlorine dioxide chlorite chromium copper cyanide (total) cyanogen chloride dichloramine fluoride iodine and iodide lead lithium manganese mercury molybdenum monochloramine nickel nitrate nitrite potassium permanganate selenium silver tin trichloramine uranium


2.2 Organic determinands (except aesthetic determinands), excluding pesticides acrylamide benzene benzo[a]pyrene bromochloroacetic acid bromochloroacetonitrile bromodichloromethane bromoform carbon tetrachloride chloroacetones chloroform 2-chlorophenol chloropicrin dialkyltins di(2-ethylhexyl)adipate di(2-ethylhexyl)phthalate dibromoacetic acid dibromoacetonitrile dibromochloromethane 1,2-dibromoethane dichloroacetic acid dichloroacetonitrile 1,2-dichlorobenzene 1,3-dichlorobenzene 1,4-dichlorobenzene 1,2-dichloroethane 1,1-dichloroethane 1,1-dichloroethene 1,2-dichloroethene (cis and trans) dichloromethane dioxins EDTA epichlorohydrin ethylbenzene fluoranthene formaldehyde hexachlorobutadiene monobromoacetic acid monochloroacetic acid monochlorobenzene MX nitrilotriacetic acid PCBs polyaromatic hydrocarbons (PAH): see benzo[a]pyrene and fluoranthene styrene tetrachloroethene toluene tributyltin oxide trichloroacetaldehyde/chloral hydrate trichloroacetic acid trichloroacetonitrile trichlorobenzenes (total)


1,1,1-trichloroethane trichloroethene 2,4,6-trichlorophenol trihalomethanes (THMs � see individuals) vinyl chloride xylenes

2.3 Pesticides alachlor aldicarb aldrin/dieldrin atrazine azinphos methyl bentazone brodifacoum bromacil carbofuran chlordanes chlorothalonil chlorpyriphos chlortoluron cyanazine 2,4-D 2,4-DB DDT + isomers diazinon 1,2-dibromo-3-chloropropane 1,2-dibromoethane 1,2-dichloropropane 1,3-dichloropropane 1,3-dichloropropene (cis and trans) dichlorprop dimethoate diquat diuron endosulfan endrin fenitrothion fenoprop glyphosate heptachlor and heptachlor epoxide hexachlorobenzene hexazinone isoproturon lindane malathion MCPA MCPB mecoprop metalaxyl methamidophos methomyl methoxychlor methyl parathion


metolachlor metribuzin molinate oryzalin oxadiazon pendimethalin pentachlorophenol permethrin phenylphenol phorate picloram pirimiphos methyl pirimsulfuron methyl procymidone propanil propazine propoxur pyridate pyriproxifen quintozene simazine 2,4,5-T terbacil terbuthylazine thiabendazole triclopyr trifluralin 1080

2.4 Aesthetic determinands aggressiveness (plumbosolvency) aluminium ammonia calcium chloride chlorine 2-chlorophenol colour copper 1,2-dichlorobenzene 1,4-dichlorobenzene 2,4-dichlorophenol ethylbenzene hardness hydrogen sulphide iron magnesium manganese monochloramine odour pH sodium styrene sulphate


taste temperature toluene total dissolved solids 1,2,3-trichlorobenzene 1,2,4-trichlorobenzene 1,2,5-trichlorobenzene 2,4,6-trichlorophenol turbidity xylenes zinc

3 Chemical determinands listed according to origin

3.1 Cyanotoxins anatoxin-a anatoxin-a(s) cylindrospermopsin endotoxins homoanatoxin-a microcystins-(MC-LR toxicity equivalents) nodularin saxitoxins (as STX equivalents)

3.2 Disinfectants chlorine chlorine dioxide iodine monochloramine

3.3 Disinfection by-products

(a) From treatment with ozone bromate bromoform dibromoacetic acid dibromoacetonitrile formaldehyde monobromoacetic acid

(b) From treatment with chlorine (gas), chlorine (hypochlorite), chloramines and chlorine dioxide

bromochloroacetic acid bromochloroacetonitrile bromodichloromethane bromoform carbon tetrachloride1 chlorate2 chlorite3


chloroacetones chloroform 2-chlorophenol chloropicrin cyanogen chloride dibromoacetonitrile dibromochloromethane dichloroacetic acid dichloroacetonitrile dichloramine 1,1-dichloroethane 1,2-dichloroethane dichloromethane 2,4-dichlorophenol formaldehyde monochloramine monochloroacetic acid monochlorobenzene MX tetrachloroethene trichloramine trichloroacetaldehyde trichloroacetic acid trichloroacetonitrile trichloroethene 2,4,6-trichlorophenol Notes: 1 Chlorine (gas and hypochlorite) only. 2 Chlorine (hypochlorite) and chlorine dioxide only. 3 Chlorine dioxide only.

(c) Trihalomethanes (THMs) bromodichloromethane bromoform chloroform dibromochloromethane

3.4 Determinands with possible sources in the distribution system acrylamide antimony asbestos benzo[a]pyrene chromium copper dialkyltins di(2-ethylhexyl)adipate di(2-ethylhexyl)phthalate formaldehyde iron lead nickel silver


toluene vinyl chloride xylenes zinc

3.5 Determinands with possible health concerns but no MAV asbestos brodifacoum bromochloroacetic acid bromochloroacetonitrile chlorine dioxide chloroacetones 2-chlorophenol chloropicrin chlorothalonil dialkyltins dibromoacetic acid dichloramine 1,3-dichlorobenzene 1,1-dichloroethane 2,4-dichlorophenol 1,3-dichloropropane dioxins fenitrothion glyphosate iodine and iodide MCPB methamidophos methomyl monobromoacetic acid MX PCBs (polychlorinated biphenyls) phorate potassium permanganate propoxur quintozene tin tributyltin oxide trichloramine


2 Datasheets for Chemical and Physical Determinands

2.1 Inorganic determinands (except aesthetic determinands)

antimony arsenic asbestos barium beryllium boron bromate cadmium chloramines (see mono) chlorate chlorine (free) chlorine dioxide chlorite chromium copper cyanide (total) cyanogen chloride dichloramine fluoride iodine and iodide lead lithium manganese mercury molybdenum monochloramine nickel nitrate nitrite potassium permanganate selenium silver tin trichloramine uranium


Antimony Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of antimony in drinking-water should not exceed 0.02 mg/L (20 µg/L). The maximum contaminant level (USEPA 2004) is 0.006 mg/L.

Sources to drinking-water

1 To source waters

Antimony can reach the aquatic environment from the weathering of minerals and rocks, run-off from soils and atmospheric deposition. Over 100 antimony containing minerals occur in nature. In New Zealand stibnite (Sb2S3), antimony sulphide, is the chief ore of antimony and is found in many quartz lodes, especially in the Reefton Goldfield and in Otago. Other examples of known major occurrence in New Zealand include near Russell, Reefton, Westland, Great Barrier Island and on the Coromandel Peninsula. Antimony can also get into water via the discharge of wastes from industries in which it is used. These include the production of semi-conductors, batteries, safety matches, electronic equipment, paint pigments, ceramic enamels, glass and pottery, plastics, ammunition primers, antifriction materials, cable sheathing, flame-proofing compounds and fireworks. Antimony is released into the atmosphere from coal-fired power plants and inorganic chemical plants. It is also found in gasolines.

2 From treatment processes

No known sources.

3 From the distribution system

The dissolution of antimony-tin solder, used in household plumbing, by corrosive water is a potential source of antimony from the distribution system.

Forms and fate in the environment The most common oxidation states of antimony are +3 and +5, both of which have been detected occasionally in natural source waters. In the normal redox range found in surface waters, most of the antimony present in the aquatic environment probably remains in solution. However, antimony has an affinity for clay and mineral surfaces and may co-precipitate with hydrous iron, manganese and aluminium. In fact, more than half of naturally-occurring antimony in sediments is bound to extractable iron and aluminium. In the atmosphere antimony may be present in gaseous, vapour and particulate forms. Smoking can result in antimony contamination of indoor air.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 897 zones, found antimony concentrations to range from �not detectable� (nd) to 0.012 mg/L, with the median concentration being �nd� (limit of detection = 0.0005 mg/L). WHO (2004) reports concentrations in surface water and groundwater normally in the range of 0.0001�0.0002 mg/L, and less than 0.005 mg/L in drinking-water.


Removal methods Coagulation/filtration and reverse osmosis are recommended methods for the removal of antimony from drinking-water. Where antimony enters the water post-treatment, ie, from dissolution of antimony-tin solder in the distribution system, control of the corrosiveness of the water will minimise the presence of antimony.

Analytical methods

Referee method

Electrothermal Atomic Absorption Spectrometric Method (APHA 3113) (pre-concentration may be necessary).

Some alternative methods

EPA Method 200.8, Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma � Mass Spectrometry.

Health considerations Antimony is not absorbed readily from the gastro-intestinal tract, with absorption ranges of <5% in cows to 15% in rats being reported. Its distribution is highest in spleen, liver and bone, and it is excreted in faeces and urine. Antimony can cross the placenta. Daily oral uptake of antimony appears to be signifcantly higher than exposure by inhalation, although total exposure from environmental sources, food and drinking-water is very low compared with occupational exposure. Polyacrylamide is used in food processing and exposure to acrylamide may also occur from this source. Acute antimony poisoning in humans may result in vomiting, diarrhoea and death. The effects of long-term human exposure to antimony have been investigated in a number of studies. In one study, six adult males who had worked in an antimony smelter for 2�13 years reported no adverse effects. However, other studies have reported increased blood pressure and heart irregularities, ulcers and increased incidence of spontaneous late abortions among female workers. One study, where workers were exposed to dust containing a mixture of antimony trioxide and antimony pentoxide for 9�31 years in an antimony smelting plant, resulted in reports of symptoms such as chronic coughing, bronchitis and emphysema, conjunctivitis, staining of front tooth surface, inactive tuberculosis and pleural adhesions. A dermatitis condition was observed in more than half the exposed workers. Animal studies have shown that antimony may cause sterility, fewer offspring and foetal damage. Mutagenic activity in tests with bacteria have been demonstrated using trivalent and pentavalent antimony salts. In addition they have been found to induce chromosomal aberrations in cultured human leucocytes and rat bone marrow cells. A study found that antimony (III) oxide and antimony ore concentrate caused lung tumours in female rats exposed by inhalation. In ingestion studies on rats and mice, antimony did not appear to cause tumours. The International Agency for Research on Cancer has concluded that, by inhalation exposure, antimony trioxide is possibly carcinogenic to humans (Group 2B) and antimony trisulphide is not classifiable as to its carcinogenicity to humans (Group 3).

Derivation of maximum acceptable value As there is limited evidence of the carcinogenicity of antimony to humans, a tolerable daily intake approach has been used for the derivation of the MAV. In a limited life-time study in which rats were


exposed to antimony in drinking-water at a single dose level of 0.43 mg antimony/kg body weight per day, reduced longevity and altered blood levels of glucose and cholesterol were observed. The incidence of benign or malignant tumours was not affected. This study has been used as the basis for the lowest-observable-adverse-effects level of antimony used in the derivation of the MAV. The MAV for antimony in drinking-water was derived as follows:

6 mg/kg body weight per day x 70 kg x 0.1 = 0.021 mg/L (rounded to 0.02 mg/L) 2 L per day x 1000

where:

• No Observed Adverse Effect Level = 6 mg/kg body weight per day for decreased body weight gain and reduced food and water intake in a 90-day study in which rats were administered potassium antimony tartrate in drinking-water

• average weight of an adult = 70 kg

• the proportion of tolerable daily intake assigned to the consumption of water = 0.1

• uncertainty factor = 1000 (100 for inter- and intraspecies variation and 10 for the short duration of the study)

• average amount of water consumed by an adult = 2 L per day.

References APHA. 1998. Standard Methods for the Examination of Water and Wastewater (20th ed). Washington: American Public Health Association, American Water Works Association, Water Environment Federation.

Australian Drinking Water Guidelines. NHMRC and AWRC, 1993 Draft, and the 2004 edition on http://www.nhmrc.gov.au/publications/synopses/eh19syn.htm

Health Canada. Guidelines for Canadian Drinking Water Quality � Supporting Documents. See http://www.hc-sc.gc.ca/ewh-semt/pubs/water-eau/doc_sup-appui/index_e.html

Kroehler CJ. 1990. What do the standards mean? Virginia Water Resources Centre.

NZGS. 1970. Minerals of New Zealand (Part A: Metallics) 38A.

USEPA. Method 200.8, Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma � Mass Spectrometry.

USEPA. 2004. 2004 Edition of the Drinking Water Standards and Health Advisories. Washington DC, USA: US Environmental Protection Agency. EPA 822-R-04-005. Available at: www.epa.gov/waterscience/criteria/drinking/standards/dwstandards.pdf

WHO. 2003. Antimony in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/74).

World Health Organization. 2004. Guidelines for Drinking-water Quality (3rd ed).

Arsenic Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of arsenic in drinking-water should not exceed 0.01 mg/L. The WHO guideline value is designated as provisional in view of the scientific uncertainties. The maximum contaminant level (USEPA 2004) is 0.01 mg/L.



1 To source waters

Arsenic can enter the aquatic environment by the weathering of minerals and rocks, run-off from soils, from geothermal fluids or atmospheric deposition. The mineralised zones of sulphitic ores probably contain the highest concentrations of arsenic although high levels of arsenic may also occur in some coals and peats. In New Zealand, arsenic occurs in greywacke and schists and in tertiary volcanics. In greywacke and schist it occurs as arsenopyrite and loellingite in gold-bearing lodes of the Reefton and Otago Goldfields. It also occurs in auriferous quartz lodes associated with volcanics in the Hauraki goldfield, especially in the Tokatea-Coromandel area. Geothermal fluids contain elevated concentrations of arsenic and water bodies such as the Waikato River subjected to their discharge have typically high arsenic concentrations. Arsenic can also be released to the aquatic environment via the discharge of wastes from industries in which it is used. Arsenic and its compounds are used in the production of semiconductors, pigments, for medical purposes, in glassmaking, in alloys with lead and copper, rodenticides, insecticides, herbicides, and as timber preservatives.


No known sources.


No known sources, despite the use of arsenical brasses.

Forms and fate in the environment The most common oxidation states of arsenic are +3 and +5 although it can also exist in the 0 and -3 states. Arsenic (V) is the stable form of arsenic in aerobic water while arsenic (III) is the predominant form of arsenic under anaerobic conditions such as in groundwaters. In surface waters, the majority of arsenic occurs in a soluble form which can be removed from the water by co-precipitation with hydrated iron and aluminium oxides, or adsorbed/chelated by suspended organic matter in water or humic substances in bottom sediments.

Typical concentrations in drinking-water Arsenic was routinely measured in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Of 1895 samples analysed between 1983 and 1989, 13 samples (1.3% of supplies) had concentrations equal to or exceeding the 1984 guideline value of 0.05 mg/L. The majority of drinking-water supplies in New Zealand have arsenic concentrations of less than 0.001 mg/L. However, supplies using source waters significantly contaminated with arsenic such as the Waikato River, that do not fully treat their water have been reported to contain up to 0.15 mg/L in reticulated water. The P2 Chemical Determinand Identification Programme, sampled from 342 zones, found arsenic concentrations to range from �not detectable� (nd) to 0.10 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). Levels in natural waters generally range between 0.001 and 0.002 mg/L, although concentrations may be elevated (up to 12 mg/L) in areas containing natural sources (WHO 2004).

Removal methods Conventional coagulation treatment with iron or aluminium can achieve good removal of arsenic. The effectiveness depends on: the oxidation state of the arsenic (trivalent arsenic should be converted to


pentavalent arsenic by oxidation with chlorine or potassium permanganate); the pH at which the process is carried out; and whether iron or aluminium is used as the coagulant. Lime-softening, ion exchange resins, and activated alumina can also be used to remove arsenic. The removal of arsenic from water by ion exchange and alumina depends upon the arsenic being present as the negatively-charged arsenate ion, AsO43-. This ion contains arsenic in the highest oxidation state (5+), and oxidation of any arsenic in the 3+ oxidation state is required if it is to be removed by these two processes.

Analytical methods

Referee method

Electrothermal Atomic Absorption Spectrometric Method (APHA 3113).


1 Hydride Generation/Atomic Absorption Spectrometric Method (APHA 3114B). 2 nductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations The health considerations apply mainly to the inorganic arsenic compounds. These are more likely to be present in drinking-water supplies than the organic compounds. Except for individuals who are occupationally exposed to arsenic, the most important route of exposure is through the oral intake of food and beverages. Ingested elemental arsenic is poorly absorbed and is largely eliminated unchanged. Soluble arsenic compounds are readily absorbed from the gastro-intestinal tract. Inorganic arsenic may accumulate in skin, bone and muscle. In humans, inorganic arsenic does not appear to cross the blood-brain barrier but transplacental transfer of arsenic has been reported. Early symptoms of acute arsenic intoxication include abdominal pain, vomiting, diarrhoea, pain in the muscles, weakness and flushing of the skin. Signs of chronic arsenicalism include dermal lesions, peripheral neuropathy, skin cancer and peripheral vascular disease. Arsenic does not appear to be mutagenic in bacterial and mammalian assays although it can induce chromosomal aberrations in a variety of cultured cell types, including human cells. Arsenic has not been demonstrated to be essential in humans. It is an important drinking-water contaminant, as it is one of the few substances shown to cause cancer in humans through consumption of drinking-water. There is overwhelming evidence from epidemiological studies that consumption of elevated levels of arsenic through drinking-water is causally related to the development of cancer at several sites, particularly skin, bladder and lung. In several parts of the world, arsenic-induced disease, including cancer, is a significant public health problem. Because trivalent inorganic arsenic has greater reactivity and toxicity than pentavalent inorganic arsenic, it is generally believed that the trivalent form is the carcinogen. However, there remains considerable uncertainty and controversy over both the mechanism of carcinogenicity and the shape of the dose-response curve at low intakes. Inorganic arsenic compounds are classified by IARC in Group 1 (carcinogenic to humans) on the basis of sufficient evidence for carcinogenicity in humans and limited evidence for carcinogenicity in animals.

Derivation of maximum acceptable value Data on the association between internal cancers and ingestion of arsenic in drinking-water are insufficient for quantitative assessment of risk. Instead, owing to the documented carcinogenicity of arsenic in the drinking-water of human populations, the lifetime risk of skin cancer has been estimated


using a multistage model that is both linear and quadratic in dose. On the basis of observations in a Taiwanese population ingesting arsenic contaminated drinking-water, the levels associated with lifetime skin cancer risks of 10-4, 10-5 and 10-6 are 1.7, 0.17 and 0.017 µg/L. These values may, however, overestimate the actual risk of skin cancer owing to the possible contribution of other factors to disease incidence in the Taiwanese population and to possible dose-dependent variations in the metabolism that could not be taken into consideration. Moreover, 1�14% of arsenic-induced skin cancers are fatal. WHO has established a provisional guideline value of 0.01 mg/L for arsenic in drinking-water. The estimated lifetime skin cancer risk associated with exposure to this concentration is six per 10,000 (6 x 10-4 or 6 x 10-6 � 8.4 x 10-5 lifetime risk of fatal skin cancers). The WHO provisional guideline value agrees with the value derived on the basis of the provisional maximum tolerable daily intake for inorganic arsenic of 2 µg/kg body weight, established by the Joint FAO/WHO Expert Committee on Food Additives in 1983, and assuming a 20% allocation to drinking-water. WHO stated in 2004 that there remains considerable uncertainty over the actual risks at low concentrations, and available data on mode of action do not provide a biological basis for using either linear or non-linear extrapolation. In view of the significant uncertainties surrounding the risk assessment for arsenic carcinogenicity, the practical quantifcation limit in the region of 0.001�0.01 mg/L and the practical difficulties in removing arsenic from drinking-water, the guideline value of 0.01 mg/L is retained. In view of the scientific uncertainties, the guideline value is designated as provisional.



Faust SD, Oly AM. 1983. Chemistry of Water Treatment. Butterworths.


IPCS. 2001. Arsenic and Arsenic Compounds. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 224).

Mattingley BI. 1991. New Zealand Drinking-water Surveillance Programme Data Review, 1983�89. DSIR Chemistry.


USEPA, Method 200.8, Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma � Mass Spectrometry.


WHO. 2003. Arsenic in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/75).


Asbestos Updated July 2005.


Maximum acceptable value There are insufficient data to derive a health based MAV for asbestos in drinking-water. The maximum contaminant level (USEPA 2004) is 7 million fibres per litre.


1 To source waters

Asbestos is a common term describing a variety of naturally formed hydrated silicates of which the two fundamental varieties are serpentine (chrysotile) and amphibole (actinolite, amosite, anthophyllite, crocidolite, tremolite). Chrysotile asbestos is formed in serpentine veins during the alteration and metamorphism of basic igneous rocks rich in ferromagnesium silicates. The amphibole asbestoses occur as a result of metamorphism, resulting, in particular, in the formation of schists and gneisses in association with limestones, argillites and igneous rocks. Examples in New Zealand of such mineral associations include the Dun Mountain ultramafics near Nelson, small lenses of mafic and ultramafic rocks at the base of the Routeburn Formation and in the Red Mountain Ultramafics (in the Livingstone Mountains and West Dome) in West Otago. Natural erosion of these rocks would result in the introduction of asbestos fibres to the aquatic environment. Asbestos fibres can also be introduced to raw waters from industrial effluents resulting from its use in the production of fire-proofing materials, heat-resistant textiles for fire-proof curtains, garments and gloves, building materials including insulation against heat and noise, floor and ceiling tiles, asphalt felts, coating and patching compounds, sheets and pipes. Asbestos is also used in brake linings and clutch facings because of its friction resistant qualities. Other uses include electrical insulation and certain paper products. Asbestos fibres may also be introduced to raw waters from sewage effluents dissolving any asbestos cement pipes through which they pass.


No known sources.


Asbestos fibres can be introduced to the treated water in the distribution system directly from the leaching of asbestos cement pipes by corrosive waters. Studies in the United States and Canada have reported typical asbestos fibre numbers in drinking-water of less than 1 MFL (million fibres per litre). Severe deterioration of asbestos cement pipes has been known to produce fibre numbers of up to 2000 MFL.

Forms and fate in the aquatic environment Not much is known about the properties of asbestos in the aquatic environment although the fate of asbestos fibres is thought to be affected by sedimentation, resuspension and migration processes.

Typical concentrations in drinking-water Asbestos has not been measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. However, in 1992 a survey was carried out of 18 water supplies (161 samples) to determine the extent to which asbestos was being released into the supplies through dissolution of asbestos cement pipes in the distribution system. In the majority of samples no asbestos fibres were detected. Chrysolite asbestos fibres were detected in 10 samples, and amphibole asbestos fibres were detected in one sample.


Removal techniques For situations where asbestos has entered the source water (weathering of ferromagnesium silicates), coagulation/flocculation with filtration will remove fibres. Asbestos can also arise from corrosion of asbestos cement lined pipes. Corrosion should be minimised by pH and lime correction, allowing a calcium carbonate deposit to line the pipes, or AC pipes should be replaced with plastic pipes.

Analytical methods

Referee method

A referee method cannot be selected for asbestos because a MAV has not been established and therefore the sensitivity required for the referee method is not known.

Some other alternative methods

No alternative methods can be recommended for asbestos for the above reasons. However, the following information may be useful: For initial screening purposes, asbestos analysis may be performed using Scanning Electron Microscopy-Energy Dispersive X-ray Analysis (SEM-EDXA). It should be noted that the results of analysis by SEM cannot be compared directly with results achieved using Transition Electron Microscopy (TEM). TEM gives greater than an order of magnitude improvement in resolution, enabling the identification and counting of finer fibres. Therefore, in samples where fibres are detected, the fibre count determined by SEM should be multiplied by 100. The method is capable of identifying fibres greater than five microns in length and greater than 0.15 microns in diameter. The detection limit for the method was estimated at around 300,000 fibres per litre, and the cost of using this technique is considerably less than the TEM method. Samples found to contain asbestos above 3.5 MFL should be confirmed using TEM. Asbestos can be analysed by transmission electron microscopy with identification by selected-area electron diffraction. This procedure is both costly and time consuming and is not suitable for routine analysis. The limit of detection is about 0.3 MFL.

Health considerations The health hazards associated with inhalation of asbestos have been recognised for a long time. They include asbestosis, cancer of the bronchial tubes, malignant mesothelioma and possible cancers of the gastrointestinal tract and larynx. Asbestos did not exhibit mutagenic activity in tests with bacteria, but has induced chromosomal aberrations, malignant transformation of mammalian cells in vitro, and various biochemical alterations associated with tumour promoters. The International Agency for Research on Cancer has concluded that asbestos is carcinogenic to humans by the inhalation route and has classified it in Group 1. Although well-studied, there has been little convincing evidence of the carcinogenicity of ingested asbestos in epidemiological studies of populations with drinking-water supplies containing high concentrations of asbestos. Limited data indicate that exposure to airborne asbestos released from tap water during showers or humidi fiation is negligible. Moreover, in extensive studies in animal species, asbestos has not consistently induced increases in the incidence of tumours of the gastrointestinal tract. The weight of the evidence shows that ingested asbestos is not hazardous to health.


Derivation of maximum acceptable value There is insufficent information to derive a health-based MAV for asbestos in drinking-water.

References Australian Drinking Water Guidelines. NHMRC and AWRC, 1993 Draft, and the 2004 edition on http://www.nhmrc.gov.au/publications/synopses/eh19syn.htm


Kroehler CJ. 1990. What do the Standards Mean? Virginia Water Resources Centre.

NZGS. 1970. Minerals of New Zealand (Part B: Non-Metallics), 38B.


WHO. 2003. Asbestos in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/2).


Barium

Maximum acceptable value Based on health considerations, the concentration of barium in drinking-water should not exceed 0.9 mg/L. The maximum contaminant level (USEPA 2004) is 2 mg/L.


1 To source waters

Barium does not occur freely in nature, but predominantly as the minerals barite (barium sulphate) and witherite (barium carbonate) in igneous rocks, sandstone and shale. It also occurs in ores of lead, zinc and silver and is often associated with fluorspar. Barium is also found in sedimentary rocks replacing the potassium in feldspar. The weathering of barite and witherite, as well as other barium-containing minerals releases barium to surface waters. In New Zealand the only major occurrence of barite is near Nelson. All other occurrences are very small with numerous veinlets in rocks of various ages, as vesicles in volcanic rocks and as concretions in Cretaceous and Tertiary sediments. Barium can also be found in raw water as the result of industrial discharges. Barium is used as a filler in paints, plastics and rubber products, in the production of glass, ceramics, photographic paper, soap, cement, metal alloys, flares, explosives, lubricating oils and drilling muds, cosmetics, diesel fuels, optical glasses, in the case hardening of steel, and as a rat poison. It may also enter water close to drilling operations where it is used as an oil drilling mud.


No known sources.


Barium is used in the manufacture of plastics from which guttering and down-piping are made. However, leaching tests have indicated the significant concentrations of barium are not released from the plastic.


Forms and fate in the environment Barium is present in water primarily from natural sources. It occurs in both the 0 and +2 oxidation states and its concentration in water is limited by the presence of sulphate and carbonate ions which cause it to precipitate. Generally barium is found in only trace amounts in surface waters because of this tendency to precipitate or to partake in adsorption or sedimentation processes.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 841 zones, found barium concentrations to range from �not detectable� (nd) to 0.082 mg/L, with the median concentration being 0.008 mg/L (limit of detection = 0.01 mg/L). WHO (2004) stated that concentrations in drinking-water are generally below 0.1 mg/L, although concentrations above 1 mg/L have been measured in drinking-water derived from groundwater.

Removal methods Cation exchange softening for the removal of barium is excellent. Barium is removed in preference to other major cations such as calcium and magnesium. Regeneration of the resin requires special attention because of the resin�s capacity and affinity for barium ions. Lime-softening at pH 9.5 to 11.5 effectively removes barium, and is pH dependent with optimum removal at pH 10.5. Reverse osmosis and electrodialysis also remove barium. Conventional alum and iron sulphate coagulation is not effective for the removal of barium, even though insoluble barium sulphate is expected to form. Two-stage coagulation is up to 80% effective, but is slow and very costly.

Analytical methods

Referee method



1 Flame Atomic Absorption Spectrometric Method (APHA 3111). 2 Inductively Coupled Plasma (ICP) Method (APHA 3120).

Health considerations Barium is not considered to be an essential element for human nutrition. The degree of absorption of barium from the gastro-intestinal tract depends on the solubility of the barium compound, the animal species, diet and age. Following absorption, barium is deposited in bone and teeth and the principal route of excretion is faecal. Food is the primary source of intake for the non-occupationally exposed population. However, where barium levels in water are high, drinking-water may contribute significantly to total intake. At high concentrations, barium causes vasoconstriction (constriction of blood vessels), peristalsis (contractions of the alimentary canal), convulsions and paralysis. Although an association between mortality from cardiovascular disease and the barium content of the drinking-water has been reported in an epidemiological study, these results were not confirmed in an analytical epidemiological study of individuals in the same population. Moreover, in a short-term study


in a small number of volunteers, there was no consistent indication of adverse cardiovascular effects following exposure to up to 10 mg/L in water. Long-term studies with rats have shown that relatively low doses of barium in drinking-water can result in significant and persistent increases in systolic blood pressure. This has significance to humans as an increase in systolic blood pressure can increase the risk of heart attack. There is no evidence that barium can cause an increase in the incidence of cancer. Barium chloride is not mutagenic in tests with bacteria and does not damage DNA. There is no evidence that barium is carcinogenic. Barium has been shown to cause nephropathy in laboratory animals, but the toxicological end-point of greatest concern to humans appears to be its potential to cause hypertension.

Derivation of maximum acceptable value The MAV for barium in drinking-water is a concentration of 0.7 mg/L. This value has been derived using the No-observed-adverse-effects level of 7.3 mg/L from the most sensitive epidemiological study conducted to date. This study reported no significant differences in blood pressure or the prevalence of cardiovascular disease between a population drinking-water containing a mean barium concentration of 7.3 mg/L and one ingesting water containing barium at 0.1 mg/L, and incorporating an uncertainty factor of 10 to account for intraspecies variation.




IPCS. 2001. Barium and Barium Compounds. Geneva: World Health Organization, International Programme on Chemical Safety (Concise International Chemical Assessment Document 33).

NZGS. 1970. Minerals of New Zealand (Part B: Non-Metallics), 38B.

Sorg TJ, Longsdon GS. 1980. Treatment technology to meet the Interim Primary Drinking Water Regulations for Inorganics: Part 5. JAWWA 72: 411�22.


WHO. 2003. Barium in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/76).


Beryllium Updated July 2005.

Maximum acceptable value (provisional) There are insufficient data to set a MAV for beryllium. DWSNZ 2005 retained the DWSNZ 2000 provisional MAV of 0.004 mg/L. WHO 2004 excluded beryllium from guideline derivation because it was considered unlikely to occur in drinking-water. The maximum contaminant level (USEPA 2004) is 0.004 mg/L.



1 To source waters

Beryllium enters natural waters through the weathering process and through atmospheric deposition. The combustion of fossil fuels is the major source of beryllium to the environment. Other less significant sources are slag and ash dumps. The principal beryllium containing mineral in New Zealand is the precious gemstone, beryl, which is very resistant to weathering. Major occurrence of beryl includes in granitic pegmatites in two places in New Zealand: Charleston and Stewart Island. In addition, beryllium replaces the silicon in feldspar minerals and it is estimated that 85�98% of the total crustal beryllium may be bound in these minerals. Beryllium may also be present in raw water from the discharge of industrial and municipal wastes. Beryllium is used in the production of light alloys, copper and brass, in the production of X-ray tubes and neon sign electrodes, and as a catalyst in the manufacture of organic chemicals. Beryllium has also been used experimentally in rocket and aircraft fuels and nuclear reactors.


No known sources.


No known sources.

Forms and fate in the environment Except for its metallic form, beryllium exists primarily in the +2 oxidation state. In aqueous solution beryllium does not exist as free Be(II) but as hydrated complexes, in particular beryllium hydroxide. In most aqueous environments, beryllium is present in particulate rather than dissolved form, primarily because of the insolubility of beryllium oxides.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 831 zones, found beryllium concentrations to range from �not detectable� (nd) to 0.012 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). Beryllium concentrations in drinking water overseas are generally very low, usually less than 0.001 mg/L.

Removal methods Coagulation/filtration, lime softening, activated alumina, ion exchange and reverse osmosis are methods that have been used for the removal of beryllium from drinking-water.

Analytical methods

Referee method

A referee method cannot be selected for beryllium because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for beryllium for the above reason. However, the following methods are used to analyse for beryllium:


1 Electrothermal Atomic Absorption Spectrometric Method (APHA 3113B). 2 Flame Atomic Absorption Spectrometry (APHA 3111). 3 Inductively Coupled Plasma Method (APHA 3120B). 4 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Beryllium appears to be poorly adsorbed from the gastrointestinal tract. No studies are available on the health effects on humans of beryllium following ingestion. As gastrointestinal absorption is poor, toxicity is expected to be low via this route. Inhalation is known to cause serious health effects, with long-term exposure resulting in pulmonary granulomatosis (a type of lung tumour). Beryllium interacts with DNA and causes gene mutations, chromosomal aberrations, and sister chromatid exchange in cultured mammalian somatic cells, although it is not mutagenic in bacterial test systems. Beryllium and beryllium compounds are classified as being probably carcinogenic to humans (Group 2A, limited evidence of carcinogenicity in humans and sufficient evidence in animals) by the International Agency for Research on Cancer on the basis of occupational exposure and inhalation studies in laboratory animals. There are no adequate studies by which to judge whether it is carcinogenic by oral exposure.

Derivation of maximum acceptable value There are no suitable oral data on which to base a toxicologically supportable MAV. However, the low concentrations of beryllium normally found in drinking-water seem unlikely to pose a hazard to consumers.



Canadian Water Quality Guidelines. April 1992.


NZGS. 1970. Minerals of New Zealand (Part A: Metallics), 38A.

USEPA, Method 200.8, Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma � Mass Spectrometry.



Boron Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of boron in drinking-water should not exceed 1.4 mg/L. In 2001 the Australian Drinking Water Guidelines considered that boron may be an essential trace


element for humans and based on an acceptable range of oral intake, a concentration of up to 4 mg/L in water would not pose a human health risk.


1 To source waters

The most common boron containing mineral is tourmaline which is present in igneous and some sedimentary rocks. The weathering of both of these rock types releases boron, which is then transported in solution. Soil leaching and volcanic activity may also add boron to water. Boron has been found in hot springs and brines at high concentrations, indicating that hydrothermal and geothermal fluids are also a source of boron. Boron may also be released to water from the discharge of industrial and domestic wastewaters, or in agricultural run-off. In industry, boron is used in fire retardants, borosilicate glass, enamels and antioxidants for soldering, detergents and in the photographic, cosmetic, leather, textile, paint and wood-processing industries. It is also used in the preparation of disinfectants and drugs and in some synthetic rocket fuels. Elemental boron is used to harden metals, in nuclear reactors for neutron absorption and in agriculture to improve crop yields.


No known sources.


No known sources.

Forms and fate in the environment The chemical behaviour of boron in the aquatic environment is poorly understood, but it is thought that the predominant species is boric acid which is moderately soluble in water and does not dissociate readily.

Typical concentrations in drinking-water Boron was routinely measured in New Zealand drinking-water supplies as part of the Department of Health three-yearly surveillance programme. Of 1904 samples analysed between 1983 and 1989, 35 samples (1.3% of supplies) contained concentrations equal to or exceeding the 1984 guideline value of 0.5 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 297 zones, found boron concentrations to range from �not detectable� (nd) to 11 mg/L, with the median concentration being �nd� (limit of detection = 0.06 mg/L).

Removal methods There are, at present, no economically feasible methods of removing boron from source waters, other than changing the source. Boron concentrations can be reduced by granular activated carbon, anion exchange or lime-softening.


Analytical methods

Referee method

Colorimetric Method, Azomethine-H Parts C, D (Boron in Waters, Effluents, Sewage and Some Solids, 1980, HMSO, UK).


1 Colorimetric Method (APHA 4500-B B).

Health considerations Boron, when administered as borates or boric acid, is rapidly and almost completely adsorbed from the gastrointestinal tract. Boron excretion occurs mainly through the kidney. Boron is present naturally in many food products, with high amounts found in foods of plant origin, especially fruits, leafy vegetables, nuts and legumes. It has been estimated that intake of boron from food is about 10 times that from water. Long term exposure of humans to boron compounds leads to mild gastrointestinal irritation. In short-term and long-term animal studies and in reproductive studies with rats, testicular atrophy was observed. Boric acid and borates were not mutagenic in various in vitro test systems. No increased tumour incidence was observed in long-term carcinogenicity studies in mice and rats. Acute boron poisoning has been reported after application of dressings, powders or ointments containing borax and boric acid to large areas of abraded skin and following ingestion. Symptoms of boron poisoning include gastrointestinal disturbances, skin eruptions, and central nervous system stimulation followed by depression. Tests for mutagenicity using bacteria and mammalian cells have been mostly negative. Neither boric acid nor borate induce chromosomal aberrations in mammalian cells.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for boron in drinking-water. Benchmark dose methodology (based on the influence of boron of foetal body weight affecting 5% of the animals, in a rat study) has been used to derive a tolerable daily intake value. This value has been used for the derivation of the MAV of boron in drinking-water. The MAV for boron in drinking-water was derived as follows:

10.3 mg/kg body weight per day x 70 kg x 0.2 = 1.4 mg/L 2 L per day x 50

where: • benchmark dose (5%) = 10.3 mg/kg body weight per day • average weight of an adult = 70 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.2 • uncertainty factor = 50 • average amount of water consumed by an adult = L per day. NB: Here the benchmark dose is used in place of the NOAEL.





HMSO. 1981. Boron in Waters, Effluents, Sewage and Some Solids, 1980: Methods for the examination of waters and associated materials. UK.


WHO. 2003. Boron in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/54).


Bromate Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of bromate in drinking-water should not exceed 0.01 mg/L. The WHO (2004) guideline value is provisional because of limitations in available analytical and treatment methods and uncertainties in the toxicological data. The maximum contaminant level (USEPA 2004) is 0.01 mg/L.


1 To source waters

Bromate is not a normal component of water. It is used in home hair permanent-wave neutralising solutions, as a food additive and to dissolve gold from its ore.


Disinfection/oxidation systems producing highly oxidising species, such as ozonation, are known to produce bromate in bromide-containing waters. Some bromate formation may also arise from the chlorination of bromide-containing waters through reaction between hypobromite and hypochlorite.


No known sources.

Forms and fate in the aquatic environment Because bromate is a strong oxidant, its chief fate is probably reaction with organic matter resulting in the formation of the bromide ion. Bromate probably does not volatilise and only slightly absorbs to soils and sediment.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme has not found bromate at detectable concentrations in the six zones sampled from supplies treated with ozone (limit of detection = 0.008 mg/L). Bromate has been reported in drinking-water with a variety of source water characteristics after ozonation at concentrations ranging from <0.002 to 0.29 mg/L, depending on bromide ion concentration, ozone dosage, pH, alkalinity and dissolved organic carbon (WHO 2004).


Removal methods Methods for the removal of bromate are still under investigation at present. Those that have shown some promise are: reduction to bromide by ferrous iron; ultraviolet irradiation; and ion-exchange, probably as a point of use unit. Some reduction in bromate production during ozonation can be achieved by keeping the pH as low as possible during treatment.

Analytical methods

Referee method

Ion chromatography (JAWWA 1992, 84(11), 88).


No alternative methods have been recommended for bromate because no methods meet the required criteria.

Health considerations Rats absorb bromate rapidly from their gastro-intestinal tract. Although bromate was not subsequently detected in tissue, bromide concentrations were increased significantly in plasma, red blood cells, pancreas, kidney, stomach and small intestine. Most cases of human poisoning from bromate are due to accidental or intentional ingestion of home permanent-wave solutions, which can contain 2�10% bromate. Toxic effects include nausea, abdominal pain and diarrhoea, central nervous system depression, and pulmonary oedema, most of which are reversible. Irreversible effects include kidney failure and deafness. In rats exposed to bromate in drinking-water for 15 months, adverse effects included inhibited body-weight gain and changes to the kidney. Kidney tumours have been reported in studies using male and female rats, but not with female mice. There is evidence that tumours occur only after a minimum total cumulative dose has been exceeded. Bromate exhibited mutagenic activity in tests using bacteria, and caused chromosomal aberrations in cultured mammalian cells. Some evidence of DNA damage has been reported in rats given potassium bromate. The International Agency for Research on Cancer has concluded that bromate is possibly carcinogenic to humans (Group 2B).

Derivation of maximum acceptable value WHO (2004) states: the upper-bound estimate of cancer potency for bromate is 0.19 per mg/kg of body weight per day, based on low-dose linear extrapolation (a one-stage Weibull time-to-tumour model was applied to the incidence of mesotheliomas, renal tubule tumours and thyroid follicular tumours in male rats given potassium bromate in drinkingwater, using the 12-, 26-, 52- and 77-week interim kill data). A health-based value of 0.002 mg/L is associated with the upper-bound excess cancer risk of 10-5. A similar conclusion may be reached through several other methods of extrapolation, leading to values in the range 0.002�0.006 mg/L.




Hautman DP, Bolyard M. 1992. Using ion chromatography to analyse inorganic disinfection by-products. JAWWA 84(11): 88.



WHO. 2003. Bromate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/78).


Cadmium Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of cadmium in drinking-water should not exceed 0.004 mg/L (4 µg/L). The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


1 To source waters

Cadmium can enter water from the weathering of rocks and minerals and run-off from soils. The only naturally-occurring cadmium compound of significance, the sulphide greenockite, CdS, which is fairly rare, is almost always associated with the polymetallic sulphide ores of zinc, lead and copper. Cadmium has a wide range of sources and may enter water in industrial and domestic discharges or from street and agricultural run-off. Domestic discharges generally contain high levels of cadmium. Its principal industrial uses are in electroplating other metals or alloys for corrosion protection, in solders and in amalgam used in dentistry. It is also used in the manufacture of pigments, nickel-cadmium storage batteries, electronic equipment, lubricants, photography supplies, glass, ceramics, biocides and as a stabiliser in plastics. It is likely to be present in waste discharged from fertiliser factories using phosphate ores containing cadmium. In agriculture, farm run-off containing these fertilisers are an important source of diffuse pollution by cadmium. Exhaust emission and tyre wear contribute a significant amount of cadmium to street run-off.


No known sources.


Cadmium may enter drinking-water from the dissolution of galvanised pipes in which it is an impurity associated with the zinc. It may also be present as a result of cadmium-containing solders in fittings, water heaters, water coolers and taps.

Forms and fate in the environment In fresh waters, cadmium exists principally as the free Cd(II) ion, cadmium chloride and cadmium carbonate. Adsorption is probably the most important process for removal of cadmium from the water column. Exchange of cadmium for calcium ions in the lattice structure of carbonate minerals can remove


cadmium from solution. In natural waters, co-precipitation with hydrous iron, aluminium and manganese oxides occurs. Alternatively, in waters of high organic content, adsorption of cadmium to humic substances and other organic complexing agents can be significant.

Typical concentrations in drinking-water Cadmium was routinely measured in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Typical concentrations of cadmium in New Zealand drinking-water supplies are <0.005 mg/L (<5 µg/L). The P2 Chemical Determinand Identification Programme, sampled from 898 zones, found cadmium concentrations to range from �not detectable� (nd) to 0.26 mg/L, with the median concentration being �nd� (limit of detection = 0.0005 mg/L).

Removal methods Lime-softening achieves good removal of cadmium, provided it is applied to hard waters. Ion exchange resins can remove cadmium, provided the resins are not overwhelmed by other cations such as calcium and magnesium. This form of treatment may be useful for the removal of heavy metals that have entered the water post-treatment. Adsorption of cadmium on to PAC, GAC and oxides of Mn(IV), Fe(III) and Al(III) has been reported. Chemical coagulation with aluminium and iron salts is limited as a viable option for the removal of soluble cadmium. The effectiveness of removal is dependent on the pH at which the process is carried out. In both cases, the effectiveness increases with increasing pH. In situations where the dissolution of poor-quality zinc from galvanized pipes is a source of cadmium, adjustment of the water chemistry to reduce its corrosiveness will minimise cadmium concentrations.

Analytical methods

Referee method



1 Inductively Coupled Plasma (ICP) Method (APHA 3120). 2 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Absorption of cadmium compounds is dependent on the solubility of the compounds. Cadmium accumulates primarily in the kidneys and has a long biological half-life in humans of about 10�35 years. The kidney is the main target organ for cadmium toxicity. In humans long-term exposure can cause kidney dysfunction leading to the excretion of protein in the urine. This may occur in about 10% of the population if the amount of cadmium exceeds 200 mg/kg. Other effects may include the formation of kidney stones and softening of the bones (osteomalacia). Itai-Itai disease has been reported in Japan among people exposed to cadmium via food and drinking-water. Symptoms were similar to osteomalacia accompanied by kidney dysfunction.


Evidence concerning the mutagenicity of cadmium is unclear with many tests reporting negative results although some report gene mutation and chromosome abnormalities in mammalian cells. However the positive results are reported as being weak and seen only at high concentrations. There is evidence for the carcinogenicity of cadmium by the inhalation route, and the International Agency for Research on Cancer has classified cadmium and cadmium compounds in Group 2A (probably carcinogenic to humans). However, there is no evidence of carcinogenicity by the oral route. Food is the main source of daily exposure to cadmium. The daily oral intake is 0.01�0.035 mg. Smoking is a significant additional source of cadmium exposure.

Derivation of maximum acceptable value As there is no evidence of carcinogenicity by the oral route and no evidence for the genotoxicity of cadmium, a provisional tolerable weekly intake (PTWI) approach has been used for the derivation of the MAV. Assuming an absorption rate for dietary cadmium of 5% and a daily excretion rate of 0.005% of body burden, the Joint FAO/WHO expert Committee on Food Additives concluded that, if levels of cadmium in the renal cortex are not to exceed 50 mg/kg, a total intake of cadmium should not exceed 0.001 mg/kg body weight per day. This total daily intake has been used to derive the MAV. The MAV for cadmium in drinking-water was derived as follows:

0.001 mg/kg body weight per day x 70 kg x 0.1 = 0.0035 mg/L (rounded to 0.004 mg/L) 2 L per day

where: • PTWI = 0.007 mg/kg, so the tolerable daily intake = 0.001 mg/kg body weight per day • average weight of an adult = 70 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.1 • average amount of water consumed by an adult = 2 L per day.




JECFA. 2000. Summary and Conclusions of the 55th meeting, Geneva, 6�15 June 2000. Geneva: World Health Organization, Joint FAO/WHO Expert Committee on Food Additives.


Sorg TJ, Csandady M, Longsdon GS. 1978. Treatment technology to meet the Interim Primary Drinking-water Regulations for Inorganics: Part 3. JAWWA 70(12): 680�891.

USEPA, Method 200.8, Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma- Mass Spectrometry.

USEPA. 2004. 2004 Edition of the Drinking Water Standards and Health Advisories. Washington DC, USA: US Environmental Protection Agency. EPA822-R-04-005. Available at: www.epa.gov/waterscience/criteria/drinking/standards/dwstandards.pdf



Chloramines Updated July 2005.

Monochloramine

Dichloramine

Trichloramine See individual entries.

Chlorate Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of chlorate in drinking-water should not exceed 0.8 mg/L. WHO (2004) stated that the guideline value for chlorate is designated as provisional because use of chlorine dioxide as a disinfectant may result in the guideline value being exceeded, and difficulties in meeting the guideline value must never be a reason for compromising adequate disinfection.


1 To source waters

Chlorate does not occur naturally but may enter source waters in industrial discharge from industries in which it is used or from agricultural practices. Chlorates have been used as herbicides and defoliants, and in the manufacture of dyes, matches and explosives. Chlorate is formed from disproportionation of chlorine dioxide which is used commercially as bleaching agent in paper production, paper pulp, and cleaning and tanning of leather.


Chlorate may appear in waters treated with chlorine dioxide, or in waters chlorinated with sodium or calcium hypochlorite. There is always a small amount of chlorate produced during the generation of chlorine dioxide. In addition, photochemical decomposition of chlorine dioxide can lead to chlorate formation in sections of the treatment process exposed to sunlight. Chlorate may also be created in chlorine dioxide-treated waters if chlorine comes in contact with the chlorine dioxide, or through disproportionation of the chlorine dioxide. Chlorate has also been detected in waters treated with hypochlorite. The older the solution the higher the chlorate concentration. Gas chlorinated systems do not show the presence of chlorate as the high hypochlorous acid/hypochlorite concentrations formed during injection are rapidly diluted.


No known sources.

Forms and fate in the environment Chlorine dioxide rapidly dissociates into chlorite, chlorate and chloride ion in treated water, chlorite being the predominant species. This reaction is influenced by the pH of the water.


The major route of environmental exposure to sodium chlorate is through drinking-water.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 130 zones, has found chlorate concentrations to range from �not detectable� (nd) to 2.2 mg/L, with the median concentration being �nd� (limit of detection = 0.05 mg/L).

Removal methods There is no satisfactory method for the removal of chlorate from water. Minimisation is best achieved by careful control of chlorine dioxide treatment conditions. pH values above about 9 and high carbonate conditions should be avoided after chlorine dioxide is added to the water. If high pH processes are used, the formation of chlorate can be minimised by minimising the concentration of chlorine dioxide during the high pH step. Treatment plants employing chlorine dioxide should be designed with a minimum of water surface area exposed to sunlight minimised to reduce chlorine dioxide loss through photo-decomposition. This precaution will also minimise chlorate formation through photo-decomposition. Granular activated carbon (GAC) removes chlorate by both chemical reduction and adsorption. Initial removal efficiencies are high, but they fall away very rapidly. Organics in the water reduce removal performance, and chlorate levels can actually be increased by the GAC. Where chlorination is carried out by hypochlorite dosing, the use of old hypochlorite solutions should be avoided.

Analytical methods

Referee method

Ion chromatography is the recommended method, having reported detection limits as low as 0.009 mg/L (9 µg/L) as chlorate. References include: Hautman and Bolyard (1992), USEPA Method 317.0 Rev 2 and USEPA Method 326.0.


The determination of individual chlorine-containing species in a chlorine dioxide-treated water presents a complex analytical problem. As a general comment, expensive instrumentation is required to obtain accurate results. Methods requiring simple apparatus (eg, amperometric titration) have been developed, and rely on the manipulation of pH to differentiate between the chlorine-containing species. Consequently, the final result involves the subtraction of a number of different readings, leading to a large uncertainty in the result. Considerable analytical skill and experience is required to obtain good precision and accuracy with these methods. Flow injection analysis provides a detection limit of approximately 0.1 mg/L as chlorate. Note: concentrations for chlorine dioxide, chlorite and chlorate determined by amperometric and DPD titrations are often expressed as mg/L of chlorine. This can lead to confusion, and when reporting results the units in which the results are expressed must be clearly stated, ie, as mg/L as Cl2 or mg/L as chlorite.

Health considerations Like chlorite, the primary concern with chlorate is oxidative damage to red blood cells. Also like chlorite, a chlorate dose of 0.036 mg/kg of body weight per day for 12 weeks did not result in any adverse effects in human volunteers. Although the database for chlorate is less extensive than that for chlorite, a recent well-conducted 90-day study in rats is available. A long-term study is in progress, which should provide more information on chronic exposure to chlorate (WHO 2004).


Chlorate is absorbed rapidly by the gastro-intestinal tract into blood plasma and distributed to the major organs. It is metabolised rapidly. Cases of chlorate poisoning are known due to its use as a herbicide. Symptoms include methaemoglobinaemia, anuria, abdominal pain and renal failure. Several studies have been carried out on human volunteers administered doses of sodium chlorate in drinking-water. The consumption of high doses of chlorate was associated with a change in serum urea nitrogen and corpuscular haemoglobin. No data are available on the mutagenic activity of chlorate.

Derivation of maximum acceptable value The WHO guideline value (2004) resulted from a recent well-conducted 90-day study in rats, based on thyroid gland colloid depletion at the next higher dose. The provisional MAV for chlorate in drinking-water was derived as follows:


where:

• NOAEL = 30 mg/kg/day, so the tolerable daily intake = 0.03 mg/kg body weight per day, based on an uncertainty factor of 1000 (10 each for inter- and intraspecies variation and 10 for the short duration of the study)








IPCS. 2000. Disinfectants and Disinfectant By-products. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 216). Available (529 pp) on http://whqlibdoc.who.int/ehc/WHO_EHC_216.pdf

Pfaff JD, Brockhoff CA. 1990. Determining inorganic disinfection by-products by ion chromatography. JAWWA 82(4): 192.

Themelis DG, et al. 1989. Determination of low concentrations of chlorite and chlorate ions by using a flow-injection system. Anal Chim Acta 255: 437.

USEPA Method 317.0 Rev 2.0 � Determination of Inorganic Oxyhalide Disinfection By-products in Drinking Water Using Ion Chromatography with the Addition of a Postcolumn Reagent for Trace Bromate Analysis.

USEPA Method 326.0 � Determination of Inorganic Oxyhalide Disinfection By-Products in Drinking Water Using Ion Chromatography Incorporating the Addition of a Suppressor Acidified Postcolumn Reagent for Trace Bromate Analysis. For the USEPA methods, see: http://www.epa.gov/safewater/methods/sourcalt.html


WHO. 2003. Chlorite and Chlorate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/86).


Chlorine Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of chlorine in drinking-water should not exceed 5 mg/L. Based on aesthetic considerations, the concentration in drinking water should not exceed 0.6�1.0 mg/L, but microbiological quality must not be compromised. The maximum contaminant level (USEPA 2004) is 4 mg/L.


1 To source waters

Chlorine may be present in source waters as a result of its discharge from industries in which it is used. These include the food industry where chlorine and hypochlorites are used for general sanitation and odour control, in the production of industrial and domestic disinfectants and bleaches, and in the synthesis of a large range of chemical compounds. Chlorine is widely used to disinfect drinking-water, cooling water, sewage, wastewater and swimming pool water.


Chlorine and hypochlorites are widely used as disinfectants for drinking-water supplies.


No known sources.

Forms and fate in the environment In environmental waters, chlorine exists as hypochlorous acid and hypochlorite ion. Both species exist in equilibrium, with their relative concentrations depending upon pH. Hypochlorous acid shows some volatility and some may be lost by aeration or water turbulence. Chlorine can react with ammonia or organic nitrogen compounds in the water to form chloramines.

Typical concentrations in drinking-water Free available chlorine concentrations in chlorinated waters should range from about 0.2 to 1 mg/L. However, the sampling point in the reticulation may have an effect on the residual measured; high levels may be found at points close to the treatment plant, while very low levels, or the absence of chlorine, may be found at the extremes of the distribution system. Free available chlorine levels in Australian waters have been found to range from 0.1 to 4 mg/L.

Removal methods If necessary, chlorine in water can be neutralised by the addition of a number of reducing agents, including sodium thiosulphate, sulphur dioxide, sodium bisulphite. Exposure of chlorinated water to sunlight will result in photodecomposition of the chlorine. The volatility of chlorine is pH dependent; it


is more easily lost from water in pH regions where hypochlorous acid is the predominant form. Some chlorine can be lost by aeration of water, but stripping is not an efficient way of removing chlorine.

Analytical methods

Referee method

DPD Ferrous Titrimetric Method (APHA 4500-Cl F). The limit of detection for this method is approximately 0.1 mg/L for field use, although lower levels can be determined under laboratory conditions and with care and experience. Analytical texts indicate that by manipulation of the conditions of the analysis, measurement of monochloramine, dichloramine and trichloramine can be made. These methods are of use when ammonia only is in the water being chlorinated. In most natural waters nitrogen-containing organic compounds are also present. Organic chloramines are formed from these compounds when chlorine reacts with them. Organic chloramines also produce colour during the DPD test and make attempts to differentiate between the different inorganic chloramines of little value. It is recommended, therefore, that only the total combined chlorine, ie, total chloramine concentration, is reported.


1 Amperometric Titration Method (APHA 4500-Cl D). While more accurate than the DPD methods, expensive equipment and a high degree of skill and care are required for this method. The limit of detection is better than 0.1 mg/L. The APHA method describes variations that will allow the determination of mono- and di-chloramine. Interferences due to organic chloramines may also cause interferences with these methods. Chlorine cannot be preserved in a sample. Analysis must be carried out as soon after sampling as possible, and if a delay between sampling and analysis is unavoidable the sample must be kept dark and chilled. 2 DPD Colorimetric Method (APHA 4500-Cl G). This method requires a spectrophotometer for the colorimetric measurements, although hand-held comparators do offer a cheaper, though less reliable variation for field use. The limit of detection with, instrumental assistance, is better than 0.05 mg/L. The limit of detection for the comparator depends on the colour disc in use. FAC concentrations as low as 0.1 mg/L should be detectable, but the accuracy of the method depends upon use of the correct lighting (natural lighting should be used with the sun behind the viewer), the individual�s ability to match colours and judge their intensity, and ensuing that readings are taken as soon after colour development as possible. Using a Nessleriser will enable concentrations of 0.05 mg/L to be read. 3 Syringaldazine (FACTS) Method (APHA 4500-Cl H). This colorimetric method is less prone to interference from combined chlorine than the DPD methods, but there can be problems with the solubility of the syringaldazine indicator. The limit of detection is 0.1 mg/L or better.

Health considerations Chlorine, or hypochlorites, are strong oxidising agents that readily react with organic molecules to produce a variety of chlorinated compounds. This reactivity makes it difficult to separate the effects of chlorine from those of its metabolites. In animal studies using a naturally-occurring non-radioactive chlorine isotope, chlorine was found to be absorbed rapidly by the gastrointestinal tract, and highest concentrations of the isotope were found in blood plasma.


It is assumed that the toxicity of aqueous solutions containing chlorine, hypochlorous acid or hypochlorite are similar since they are in dynamic equilibrium. Chlorine concentrations therefore refer to free available chlorine. Very few toxic effects have been associated with drinking-water containing high chlorine concentrations. In one report, 150 people drank water with 50 mg/L during a period of mains disinfection, with no adverse effects. Several instances have been reported where military personnel drank water with chlorine concentrations up to 32 mg/L for several months with no ill-effects. Momentary constriction of the throat, and mouth irritation were observed when the chlorine concentration exceeded 90 mg/L. Most people would refuse to drink water with a chlorine concentration over 25 mg/L. A number of studies have suggested an association between chlorination by-products and various cancers. This association has been consistent in relation to cancers of the bladder and rectum, but there are insufficient data to determine concentrations at which chlorination by-products might cause increased risk to human health. Long-term animal toxicity studies have shown no specific effects from the ingestion of chlorine. Chlorine, hypochlorous acid and hypochlorite did not act as carcinogens or tumour initiators. Assessment of the mutagenicity of chlorine is complicated by the reactivity of chlorine. Hypochlorite was found to be mutagenic in tests with one strain of bacteria but not with another. Chromosome aberrations were reported in tests with mammalian cells. The International Agency for Research on Cancer (IARC) has concluded that hypochlorites are not classifiable as to their carcinogenicity to humans, Group 3.

Derivation of maximum acceptable value The MAV for chlorine has been derived as follows:

15 mg/kg body weight per day x 70 kg x 1 = 5.25 mg/L, rounded to 5 mg/L 2 L per day x 100

where:

• no observable adverse effects level = 15 mg/kg body weight per day. This is from a study which reported the absence of toxicity in rodents that received chlorine as hypochlorite in drinking-water for up to two years

• average adult weight = 70 kg

• the proportion of the total intake of chlorine to drinking-water = 1

• the average quantity of water consumed by an adult = 2 L per day

• uncertainty factor = 100 for intra- and interspecies variation. It should be noted that this value is conservative as no adverse effect level was identified in this study. Most individuals are able to taste chlorine at the MAV.





WHO. 2003. Chlorite and Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/45).


Chlorine dioxide Updated July 2005.

Maximum acceptable value A MAV has not been established for chlorine dioxide because of the rapid breakdown of chlorine dioxide and because the chlorite MAV is adequately protective for potential toxicity from chlorine dioxide. The Australian Drinking Water Guidelines state: based on aesthetic considerations, the concentration in drinking water should not exceed 0.4 mg/L. Chlorine dioxide would not be a health consideration unless the concentration exceeded 1 mg/L. The maximum contaminant level (USEPA 2004) is 0.8 mg/L.


1 To source waters

Chlorine dioxide is used as a bleaching agent in paper production, paper pulp and for the cleaning and tanning of leather. However, in wastewaters from these industries it is generally chlorite which is detected because of its rapid formation from the reactions of chlorine dioxide with compounds in the waste.


Chlorine dioxide is used as a disinfectant for drinking-water supplies. It reacts with natural organic matter in the water, and other oxidizable materials to produce primarily chlorite and to a lesser extent chloride. Chlorate may also be present in these waters as a by-product of the reaction used to generate the chlorine dioxide.


No known source.

Forms and fate in the environment Chlorine dioxide is unlikely to be long-lived in the environment. This is partly because it reacts with organic matter in water, primarily to produce chlorite, and to some extent chloride, and partly because it will be photodecomposed readily by sunlight. The major photodecomposition products are: chloride, chlorate and molecular oxygen.

Typical concentrations in drinking-water Only one public water supply has used chlorine dioxide for water treatment on a long-term basis in New Zealand, and no data are available on typical chlorine dioxide concentrations.


Chlorine dioxide residuals in treated Australian waters typically have been between 0.2 and 0.4 mg/L. It is particularly effective in the control of manganese-reducing bacteria. The need to minimise chlorite production is a major factor that influences the doses of chlorine dioxide that can be used.

Removal methods Chlorine dioxide can be removed by the addition of chemical reducing agents, which include sodium thiosulphate, sulphur dioxide, sodium bisulphite, and ferrous chloride, to the water. The chemical reduction of chlorine dioxide can have unacceptable consequences on the water quality, because reduction often leads to the production of chlorite, and although the chlorite can be reduced to chloride under some conditions, the presence of oxygen in the water can produce chlorate. Decomposition of the chlorine dioxide with light could be considered, but chlorate, as well as chloride and oxygen, is formed in this process.

Analytical methods

Referee method

A referee method cannot be selected for chlorine dioxide because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for chlorine dioxide for the above reason. However, the following information may be useful: Chlorine dioxide cannot be preserved in a sample. Analysis must be carried out as soon after sampling as possible, and if a delay between sampling and analysis is unavoidable the sample must be kept dark and chilled. Aeration of the sample must be avoided. The analysis of chlorine dioxide in a water sample can be complex if chlorine and chloramines are present. The DPD and amperometric methods noted are extensions of the equivalent methods for chlorine. 1 Amperometric method (APHA 4500-ClO2 C).

This method requires more sophisticated equipment and greater skill on the part of the analyst than either of the other two methods. The uncertainty in the result can be larger than the result itself. 2 DPD method (APHA 4500-ClO2 D).

A number of pH adjustments have to be made with this method during the measurement, which adds to the difficulty of its use. The limit of detection is approximately 0.1 mg/L as Cl2, or 0.19 mg/L as ClO2. 3 Chlorophenol red method (Harp et al).

This is a relatively simple colorimetric method that utilises the quantitative reduction in the colour of chlorophenol red resulting from its reaction with chlorine dioxide. The method is specific for chlorine dioxide; other chlorine compounds do not react with chlorophenol red in this way. The limit of detection for this method has been reported to be 0.004 mg/L as ClO2, but these were under laboratory conditions during development, and the sensitivity would probably be less for field analysis. Some methods measure chlorine dioxide as chlorine equivalent. The conversion factor for chlorine dioxide expressed in mg of Cl2/L to chlorine dioxide expressed in mg of ClO2/L is:


Chlorine dioxide (as mg of Cl2/L) = chlorine dioxide (as mg of ClO2/L) x 2.6 This assumes the result in mg of Cl2/L was calculated to yield the full oxidising capability (5 electron transfer) of ClO2. The complication of knowing how the ClO2 concentration expressed as Cl2 was determined is a very good reason for ensuring that all these determinands are expressed in units of their own weight per volume, rather than equivalent weight of Cl2/volume, wherever possible.

Health considerations Chlorine dioxide is absorbed rapidly from the gastro-intestinal tract and no particular organ appears to concentrate the dose selectively following exposure. Animal studies have shown that following ingestion, chlorine dioxide is converted rapidly to the chloride ion, and to a lesser extent chlorite and chlorate. Excretion is primarily via urine. In a study with human volunteers, no adverse effects were observed after drinking-water containing chlorine dioxide concentrations up to 5 mg/L for periods of 12 weeks. In rats exposed from birth, high concentrations of chlorine dioxide may impair neurobehavioural and neurological development. Significant depression of thyroid hormones have been observed in rats, pigeons and monkeys exposed to doses of approximately 10 mg/kg body weight per day for long periods of time. No tumours were observed in rats following two-year exposures to chlorine dioxide in drinking-water. Chlorine dioxide was mutagenic in tests with one bacterium strain. It did not induce chromosomal aberrations in tests with mouse bone-marrow cells.

Derivation of maximum acceptable value A MAV has not been established for chlorine dioxide because of the rapid breakdown of chlorine dioxide and because the chlorite MAV is adequately protective for potential toxicity from chlorine dioxide. The taste and odour threshold for chlorine dioxide is 0.4 mg/L.



Harp DL, Klein RL, Schoonover DJ. 1981. JAWWA 73(7): 387.






Chlorite Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of chlorite in drinking-water should not exceed 0.8 mg/L. The guideline value for chlorite was designated as provisional because use of chlorine dioxide as a disinfectant may result in the chlorite guideline value being exceeded, and difficulties in meeting the guideline value must never be a reason for compromising adequate disinfection. The maximum contaminant level (USEPA 2004) is 1 mg/L.


1 To source waters

Chlorite may enter raw waters in industrial discharges when it is used in the production of paper, textiles and straw products and in the manufacture of waxes, shellacs, and varnishes.


The appearance of chlorite in treated water is associated with the use of chlorine dioxide. There are three reasons for chlorite appearing through the use of this treatment process. Chlorine dioxide is generated by the reaction of chlorine with chlorite. Poor control over the process, so that insufficient chlorine is added for complete reaction of the chlorite, may lead to excess chlorite being carried into the treated water. Chlorine dioxide itself will undergo a base-catalysed decomposition to form chlorite. The rate of the reaction is also increased by the presence of carbonate. It is therefore important that processes raising the pH above 9, and/or using carbonate, such as lime-soda softening, not be used after chlorine dioxide dosing. Chlorite is also produced from the chemical reduction of chlorine dioxide during the oxidation of organic matter.


No known sources.

Forms and fate in the environment Chlorine dioxide dissociates rapidly into chlorite, chlorate and chloride ion in treated water, chlorite being the predominant species. This reaction is influenced by the pH of the water. The major route of environmental exposure to sodium chlorite is through drinking-water.

Typical concentrations in drinking-water Chlorite was not routinely measured in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. However, a small programme was undertaken to monitor the chlorite and chlorate levels in a supply receiving water treated by chlorine dioxide. In ten samples the chlorite concentrations ranged from 0.6 to 2.1 mg/L as chlorite. Such a small data set does not provide a good basis for a comment on what might be typical in New Zealand; a number of factors, including raw water quality, will control the sorts of levels that will appear in a particular supply.


Removal methods It is preferable to avoid forming chlorite rather than trying to remove it once formed. This is best achieved by: removing as much organic matter from the water as possible before the chlorine dioxide is introduced; ensuring that the pH of the water is kept below 9; and avoiding high carbonate concentrations in the water. If high pH and high carbonate conditions are unavoidable, the chlorine dioxide concentration at this part of the process should be minimised. Some chemical reductants used in water treatment have been found to produce unacceptably high concentrations of chlorate when used to remove chlorite from water. Thiosulphate and ferrous iron appear to be reductants that could be used with minimal formation of chlorate. GAC removes chlorite by both adsorption and reduction, but the initial high efficiencies fall away rapidly.

Analytical methods

Referee method

The recommended method is ion chromatography, having a reported detection limit as low as 0.003 mg/L (3 µg/L) as chlorite. References include: Hautman and Bolyard (1992), USEPA Method 317.0 Rev 2 and USEPA Method 326.0.


The determination of individual chlorine-containing species in a chlorine dioxide-treated water can be complex. As a general comment, expensive instrumentation is required to obtain accurate results. Methods requiring simple apparatus have been developed, and rely on the manipulation of pH to differentiate between the chlorine-containing species. Consequently, the final result involves the subtraction of a number of different readings, leading to a large uncertainty. Considerable analytical skill and experience is required to obtain good precision and accuracy with these methods when concentrations less than 1 mg/L are of interest. Chlorite can be determined by: amperometric titration (APHA 4500-ClO2 D); DPD titration (APHA 4500-ClO2 E); polarography; specific ion electrode (using a chlorine-specific electrode); flow injection analysis; and ion chromatography. There is a range of accuracy and reliability within this set of methods. Flow injection analysis is also reliable but its reported detection limit of approximately 0.1 mg/L as chlorite is poorer than that of ion chromatography. NOTE: concentrations for chlorine dioxide, chlorite and chlorate determined by amperometric and DPD titrations are often expressed as mg/L as Cl2. This can lead to confusion, and when reporting results, the units in which the results are expressed must be stated clearly, ie, as mg/L as Cl2 or mg/L as chlorite.

Health considerations In a study with human volunteers, no adverse effects were observed after drinking-water with either chlorine dioxide or chlorite concentrations up to 5 mg/L for periods of 12 weeks. Chlorite affects red blood cells, resulting in methaemoglobin formation in cats and monkeys. The International Agency for Research on Cancer has concluded that chlorite is not classifiable as to its carcinogenicity to humans (Group 3).


Derivation of maximum acceptable value As chlorite is not classifiable as to its carcinogenicity to humans, a tolerable daily intake approach has been used for the derivation of the MAV. A two-generation study in rats, based on lower startle amplitude, decreased absolute brain weight in the F1 and F2 generations and altered liver weights in two generations. The MAV for chlorite in drinking-water was derived as follows:

2.9 mg/kg body weight per day x 70 kg x 0.8 = 0.812 mg/L (rounded to 0.8 mg/L) 2 L per day x 100

where: • no observable adverse effects level = 2.9 mg/kg body weight per day • average weight of an adult = 70 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.8 • average amount of water consumed by an adult = 2 L per day • uncertainty factor = 100 (10 each for inter- and intraspecies variation). The MAV is designated as provisional because use of chlorine dioxide as a disinfectant may result in the chlorite MAV being exceeded, and difficulties in meeting the MAV must never be a reason for compromising adequate disinfection.






Limoni B, et al. 1984. Determination of oxidants formed upon the disinfection of drinking-water with chlorine dioxide. J Environ Sci Health A19(8): 943.

Nokes CJ. 1992. Chlorine Dioxide in Potable Water Treatment: A monitoring study. Wellington: Department of Health.

Themelis DG, et al. 1989. Determination of low concentrations of chlorite and chlorate ions by using a flow-injection system. Anal Chim Acta 255: 437.

USEPA Method 317.0 Rev 2.0 � Determination of Inorganic Oxyhalide Disinfection By-products in Drinking Water Using Ion Chromatography with the Addition of a Postcolumn Reagent for Trace Bromate Analysis.

USEPA Method 326.0 � Determination of Inorganic Oxyhalide Disinfection By-Products in Drinking Water Using Ion Chromatography Incorporating the Addition of a Suppressor Acidified Postcolumn Reagent for Trace Bromate Analysis. For the USEPA methods, see: http://www.epa.gov/safewater/methods/sourcalt.html


WHO. 2003. Chlorite and Chlorate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/86).



Chromium Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of total chromium in drinking-water should not exceed 0.05 mg/L. WHO (2004) states that their guideline value is designated as provisional because of uncertainties in the toxicological database. The Australian Drinking Water Guidelines state �If the concentration of total chromium exceeds this value then a separate analysis for hexavalent chromium should be undertaken�. The maximum contaminant level for total chromium (USEPA 2004) is 0.1 mg/L.


1 To source waters

Chromium is present in most soils and rocks and it can enter water naturally from weathering and run-off from soils. Chromite, the main chromium containing mineral, is found in ultramafic rocks such as dunites, peridotites, pyroxenites and serpentinites, or as the detrital mineral from weathering of these rocks. Most of the known chromite in New Zealand occurs as disseminated grains in dunites, or segregated in small discontinuous lenses in dunites and serpentinites. Chromite is highly resistant to weathering. Examples of chromite occurring in ultramafic rocks include d�Urville Island; Croisilles Harbour, Nelson; Whangamoa, Nelson; Dun Mountain Area, Nelson; Red Hill, Wairau Valley; Red Mountain, Northwest Otago. Detrital chromite is found in the Wellsford-Silverdale area in Northland; from near Te Kuiti; Lee River near Nelson; and in beach sands between Greymouth and Hokitika. Chromium is used in a range of industries, especially leather tanning and timber treatment. Hexavalent chromium compounds are used in the metallurgical industry for stainless steel, chrome alloy and chromium metal production, in the chemical industry as oxidising agents in chrome plating, and in the production of other chromium compounds used in paints, dyes, explosives, ceramics and paper. Trivalent chromium salts are used in textile dyeing, in the ceramic and glass industry and in photography. Chromium compounds are also added to drilling muds to reduce corrosion fatigue of drilling pipes, in heating and cooling coils, in sprinkler systems and it is present in some fertilisers and pesticides.


No known sources.


It is possible that corrosion of metals in reticulation and plumbing systems could lead to levels of chromium at the tap being higher than those leaving the plant.

Forms and fate in the environment Chromium is present in the environment in the trivalent and hexavalent states although trivalent chromium occurs more commonly. Trivalent chromium is an essential trace element for humans, with food being the major source of intake. Hexavalent chromium occurs infrequently in nature. Its presence in water is generally the result of industrial and domestic chromium waste discharges. Hexavalent chromium is not considered to be an essential nutrient and harmful effects due to chromium have been attributed to this form.


Typical concentrations in drinking-water Chromium was measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Of 1904 samples analysed between 1983 and 1989, six samples (0.3%) were equal to or exceeded the MAV of 0.05 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 897 zones, found chromium concentrations to range from �not detectable� (nd) to 0.12 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L). Total chromium concentrations in drinking-water are usually less than 0.005 mg/L although concentrations between 0.06 mg/L to 0.12 mg/L have been reported overseas. In major Australian reticulated supplies concentrations of total chromium range up to 0.03 mg/L, with typical concentrations usually less than 0.005 mg/L.

Removal methods Cr(III) is more easily removed from drinking-water than Cr(VI). Very little removal of Cr(VI) occurs by coagulation with alum or ferric sulphate. However, nearly 100% removal can be achieved by coagulation with ferrous sulphate. In this process, the Cr(VI) is reduced to Cr(III) by the ferrous ions, then precipitated as insoluble chromium hydroxide. Some pH manipulation is required to obtain the optimum pH conditions for the different steps of reduction and precipitation. Lime-softening is effective for removal of Cr(III) but not Cr(VI). Some chromium removal can be achieved by the use of activated carbon, but the efficiency of removal depends strongly on the pH and the type of carbon used. Initial studies have shown that Cr(VI), in the form of the chromate ion, is removed efficiently by anion exchange techniques. To avoid chromium entering the water as the result of corrosion of metals in the reticulation, the corrosiveness of the water should be minimised.

Analytical methods

Referee method



1 Flame Atomic Absorption Spectrometric Method (APHA 3111). 2 Inductively Coupled Plasma Method (APHA 3120B). 3 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations The absorption of chromium after oral exposure is relatively low and depends on the speciation. Cr(VI) is more readily absorbed from the gastrointestinal tract than Cr(III) and is able to penetrate cellular membranes. In general, food appears to be the major source of intake.


There are no adequate toxicity studies available to provide a basis for a no observable adverse effects level. In a long-term carcinogenicity study in rats by the oral route with Cr(III), no increase in tumour incidence was observed. In rats, Cr(VI) is a carcinogen via the inhalation route, although the limited data available do not show evidence for carcinogenicity via the oral route. In epidemiological studies, an association has been found between exposure to Cr(VI) and lung cancer. The International Agency for Research on Cancer has classified Cr(VI) in Group 1 (carcinogenic to humans) and Cr(III) in Group 3 (not classifiable as to its carcinogenicity to humans). Cr(VI) compounds are positive in a wide range of in vitro and in vivo genotoxicity tests, whereas Cr(III) compounds are not. The mutagenic activity of Cr(VI) can be decreased or abolished by reducing agents, such as human gastric juices.

Derivation of maximum acceptable value The provisional MAV for chromium in drinking-water comes from the following WHO assessment. There are no adequate toxicity studies available to provide a basis for derivation a NOAEL. The guideline value was first proposed in 1958 for hexavalent chromium, based on health concerns, but was later changed to a guideline for total chromium because of difficulties in analysing for the hexavalent form only. Because of the carcinogenicity of Cr(VI) by the inhalation route and its genotoxicty, the 1984 MAV of 0.05 mg/L has been questioned, but the available toxicological data do not support the derivation of a new value. However, as a practical measure, 0.05 mg/L, which is considered to be unlikely to give rise to significant risks to health, has been retained as the provisional MAV until additional information becomes available and chromium can be re-evaluated.



Faust SD, Aly OM. 1983. Chemistry of Water Treatment. Butterworths.




Sorg TJ, Csandy M, Longsdon GS. 1978. Treatment technology to meet the Interim Primary Drinking Water Regulations for Inorganics: Part 3. JAWWA 70(12): 680�91.

USEPA, Method 200.8. Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma-Mass Spectrometry.


WHO. 2003. Chromium in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/4).



Copper Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of copper in drinking-water should not exceed 2 mg/L. Based on aesthetic considerations, the concentration of copper in drinking water should not exceed 1 mg/L. The action level (USEPA 2004) is 1.3 mg/L at the tap; they have a secondary drinking water regulation of 1.0 mg/L for copper.


1 To source waters

Copper occurs widely in nature in rocks and soils as sulphide and carbonate minerals. Weathering of these minerals releases copper to the aquatic environment. In New Zealand the typical geological settings for the occurrence of copper are associated with volcanic and metamorphic rocks. Areas likely to have high levels of copper include Northland, Great Barrier Island, Ruahine Ranges, Cape Colville, Coromandel Peninsula, Egmont National Park, Nelson, Stewart Island, Southland, Northwest Otago, South Westland and Fiordland. Copper is used in a range of industries including timber treatment, the manufacture of electrical wiring, electroplating, production of alloys (eg, bronze and brass) and items made from these alloys, photography, utensils, antifouling paint, pesticide formulations, and textiles. It is also used in construction, roofing materials and brass and copper plumbing. Copper salts may be added in small amounts to water supply reservoirs to suppress the growth of algae.


No known sources.


An important source of copper is from the dissolution of domestic water pipes and plumbing fixtures. A number of factors determine the extent to which copper is dissolved from pipes and fittings. The chemical factors include: pH, carbon dioxide concentration, dissolved oxygen concentration, chlorine residual, hardness and alkalinity. Studies have shown that high carbon dioxide content is a major factor in determining the corrosiveness of a water towards copper. In addition, the copper concentration in the water can be expected to be higher, if the water in contact with the copper remains stagnant.

Forms and fate in the environment The most common oxidation states of copper are +1 (cuprous) and +2 (cupric) with the cupric ion being dominant in aerated waters. Copper exhibits complex behaviour in the aquatic environment and may be present as the free cupric ion or complexed with inorganic or organic ligands. At pH levels and inorganic carbon concentrations characteristic of natural fresh waters, most of the soluble copper is present as complexes of cupric carbonate. Less than 1% of the total copper exists in the free ionic form in natural water. Copper has a high affinity for hydrous iron and manganese oxides, clays, carbonate minerals and organic matter. In reducing acidic environments, remobilisation of adsorbed or co-precipitated copper can occur. In the presence of soluble organic matter, adsorption of copper may be relatively ineffective, resulting in an increase in soluble forms in the water column.


Typical concentrations in drinking-water Copper was routinely measured in New Zealand drinking-water supplies as an aesthetic parameter as part of the Department of Health three yearly surveillance programme. Copper is now considered to be an inorganic parameter of health significance. Of 1143 samples analysed from 913 supplies, between 1983 and 1989, 2284 samples (31% of supplies) exceeded the then highest desirable 1984 MAV of 0.05 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 897 zones, found copper concentrations to range from �not detectable� (nd) to 10 mg/L, with the median concentration being 0.032 mg/L (limit of detection = 0.002 mg/L).

Removal methods The copper found in drinking-waters is usually not the result of the metal being present in the source water. It generally arises because of corrosion of fittings in the consumers� plumbing. Correct pH adjustment, removal of dissolved carbon dioxide, maintenance of a moderate alkalinity, avoidance of high chloride levels, and minimisation of aluminium carry-over from the coagulation process are ways in which a water�s tendency to dissolve copper can be reduced. In some instances, replacement of copper piping with plastic piping may be required to reduce severe copper problems. Flushing pipes before drinking-water is drawn from the tap can help to reduce high copper concentrations that arise from corrosive water being in contact with the piping for extended periods of time. Should copper be present in the raw water, alum coagulation at pH 7 is effective in its removal, provided the copper is bound to fulvic acid.

Analytical methods

Referee method



1 Flame Atomic Absorption Spectrometry (APHA 3111). 2 Inductively Coupled Plasma Method (APHA 3120B). 3 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Copper is an essential element, and the intake from food is normally 1�3 mg/day. In adults, the absorption and retention rates of copper depend on the daily intake; as a consequence, copper overload is unlikely. The absorption of copper by the gastro-intestinal tract is dependent on a number of factors including pH and copper speciation, but is probably 25�60% effective. Copper is stored in the brain and muscle tissue. High concentrations can also be found in the kidneys, heart and hair. Acute gastric irritation may be observed in some individuals at concentrations in drinking-water above 3 mg/L. In adults with Wilson�s disease, the copper regulatory mechanism is defective, and long-term ingestion can give rise to liver cirrhosis. Copper metabolism in infants, unlike that in adults, is not well-developed, and the liver of the newborn contains about 90% of the body burden, with much higher levels than in adults. Since 1984, there has been some concern regarding the possible involvement of copper from drinking-water in early childhood liver cirrhosis in bottle fed infants, although this has yet to be confirmed.


Copper was not found to be carcinogenic in tests with mice and dogs. The results of mutagenicity tests with different strains of bacteria are generally negative. Tests for mutagenicity using mammalian cells, both in vitro and in vivo, give predominantly positive results. Derivation of the Maximum acceptable value IPCS concluded that the upper limit of the acceptable range of oral intake in adults is uncertain but is most likely in the range of several (more than 2 or 3) but not many milligrams per day in adults. This evaluation was based solely on studies of gastrointestinal effects of copper-contaminated drinking-water. The available data on toxicity in animals were not considered helpful in establishing the upper limit of the acceptable range of oral intake due to uncertainty about an appropriate model for humans, but they help to establish a mode of action for the response. The data on the gastrointestinal effects of copper must be used with caution, since the effects observed are influenced by the concentration of ingested copper to a greater extent than the total mass or dose ingested in a 24-hour period. Recent studies have delineated the threshold for the effects of copper in drinking-water on the gastrointestinal tract, but there is still some uncertainty regarding the long-term effects of copper on sensitive populations, such as carriers of the gene for Wilson disease and other metabolic disorders of copper homeostasis. Basis of guideline: to be protective against acute gastrointestinal effects of copper and derivation provide an adequate margin of safety in populations with normal copper homeostasis. For adults with normal copper homeostasis, the MAV should permit consumption of two or three litres of water per day, use of a nutritional supplement, and copper from foods, without exceeding the tolerable upper intake level of 10 mg/day or eliciting an adverse gastrointestinal response. A concentration of 2 mg/L should contain a sufficient margin of safety for bottle-fed infants, because copper intake from other sources is usually low. The aesthetic guideline value for copper is 1 mg/L for the staining of laundry and sanitary ware.





IPCS. 1998. Copper. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 200).





WHO. 2003. Copper in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/88).



Cyanide Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of cyanide (as total CN) in drinking-water should not exceed 0.08 mg/L. The maximum contaminant level (USEPA 2004) is 0.2 mg/L.


1 To source waters

The natural decomposition of some plants that synthesise cyanoglycosides, and micro-organisms that produce free cyanide as a result of their metabolic processes, may release cyanide to source waters. Some microorganisms, such as the cyanobacterium Anacystis nidulans and the bacterium Chromobacterium violaceum, produce free cyanide. In addition, cyanide may also enter water via the discharge of wastes from industries in which it is used, and from agricultural practices. Cyanide is used in the electroplating, steel and chemical industries, in insecticides and rodenticides, and sodium cyanide is used in the extraction of gold and silver from low-grade ores. In uncontaminated water sources, free cyanide concentrations are usually less than 0.01 mg/L.


No known sources.


No known sources.

Forms and fate in the environment Cyanides are a diverse group of organic and inorganic compounds, characterised by the -C≡N group. The form of cyanide in water is dependent primarily on pH, and also on temperature, dissolved oxygen, salinity and the presence of other ions, complexing agents and sunlight.

Typical concentrations in drinking-water Cyanide was not measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. During the P2 Chemical Determinand Identification Programme, cyanide was found in only one zone, at a concentration equal to the limit of detection (0.01 mg/L). In major Australian reticulated supplies cyanide concentrations range up to 0.05 mg/L, with typical concentrations usually less than 0.02 mg/L.

Removal methods Chlorine oxidation, reverse osmosis, and ion exchange have been used for the removal of cyanide from drinking-water. Note that some cyanide-containing complexes are resistant to oxidation by chlorine. Cyanide in this form, however, is unlikely to present a health concern as it will be very strongly bound to the complexing metal ion. Ozone is also an effective oxidant.


Analytical methods

Referee method

Total Cyanide after Distillation (APHA 4500-CN C).


The following methods may be useful for the anlysis of cyanide: 1 Cyanides amenable to chlorination after distillation (APHA 4500-CN G).

After part of the sample is chlorinated to decompose the cyanides, both the chlorinated and untreated samples are distilled as described above. The difference in cyanide concentration between the two samples is expressed as the cyanides amenable to chlorination. Some organic compounds may oxidise or breakdown during chlorination giving higher results for cyanide after chlorination than before chlorination. 2 Cyanides amenable to chlorination without distillation (APHA 4500-CN H).

This method covers the determination of HCN and of CN complexes that are amenable to chlorination and also thiocyanates. The cyanides are oxidised by chloramine T after the sample has been heated. Addition of pyridine-barbituric acid reagent produces a red-blue colour which is measured photometrically at 578 nm. Thiocyanate interferes with this method and it is not applicable if the ratio of SCN to CN is greater than 3.

Health considerations Cyanide is highly toxic. It is readily absorbed by the gastrointestinal tract and metabolised to thiocyanate. Cyanide can deplete vitamin B12 levels and has been linked to an increased incidence of goitre and cretinism. Nutritionally inadequate people are most at risk. Effects on the thyroid and particularly the nervous system were observed in some populations in the tropics as a consequence of the chronic consumption of inadequately processed cassava containing high cyanide concentrations. No data are available on the carcinogenic properties of cyanide and tests for mutagenicity with different strains of bacteria have been mostly negative.

Derivation of maximum acceptable value There is a very limited number of toxicological studies suitable for deriving a MAV. There is, some indication in the literature that pigs may be more sensitive than rats. There is only one study available in which a clear effect level was observed, at 1.2 mg/kg body weight per day, in pigs exposed for 6 months. The effects observed were in behavioural patterns and serum biochemistry. This study has been used as the basis of the lowest-observable-adverse-effects level used in the derivation of the MAV. The MAV for cyanide in drinking-water was derived as follows:

1.2 mg/kg body weight per day x 70 kg x 0.2 = 0.084 mg/L (rounded to 0.08 mg/L) 2 L per day x 100

where:

• lowest observable adverse effect level = 1.2 mg/kg body weight per day in a six-month study in pigs




• uncertainty factor = 100 (for inter- and intraspecies variation; no additional factor for the use of a LOAEL was considered necessary because of doubts over the biological significance of the observed changes)







WHO. 2003. Cyanide in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/5).


Cyanogen chloride Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of cyanogen chloride in drinking-water should not exceed 0.08 mg/L.


1 To source waters

Cyanogen chloride may be found in raw water as an industrial contaminant. It is used in tear gas, in fumigant gases, and as a reagent in the synthesis of other compounds.


Cyanogen chloride may be formed as a by-product of chloramination or chlorination of water.


No known sources.

Forms and fate in the environment Cyanogen chloride is a volatile gas, only slightly soluble in water. Hydrolysis is an important removal mechanism for cyanogen chloride in water.


Typical concentrations in drinking-water Studies from the USA indicate that concentrations found in finished waters are below 0.01 mg/L. Concentrations tend to be higher in chloraminated waters than in chlorinated waters. Cyanogen chloride was not detected in any zones sampled during the P2 Chemical Determinand Identification Programme (limit of detection = 0.01 mg/L). WHO (2004) reported cyanogen chloride concentrations in drinking-water treated with chlorine and chloramine at 0.0004 and 0.0016 mg/L, respectively

Removal methods No information is available for methods to remove cyanogen chloride from contaminated source waters. As this compound arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with chlorine or chloramine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as the primary coagulant. NOM can also be removed by adsorption on to activated carbon, activated alumina or ion-exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated difficulties.

Analytical methods

Referee method

Cyanogen chloride method (APHA 4500-CN J).


No alternative methods have been recommended for cyanogen chloride because no methods meet the required criteria.

Health considerations Cyanogen chloride is rapidly metabolised to cyanide in the body. Effects of ingested cyanogen chloride in humans have not been reported. Cyanogen chloride was used as a nerve gas in World War I. A concentration of 48 ppm in air was lethal and inhalation of concentrations above 1 ppm causes irritation. No data are available on the carcinogenicity or mutagenicity of cyanogen chloride.


Derivation of maximum acceptable value There are few data on the oral toxicity of cyanogen chloride, and the proposed MAV is therefore based on cyanide. A MAV of 0.08 mg/L for cyanide as total cyanogenic compounds has been developed (see cyanide).



WHO. 2003. Cyanogen Chloride in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva,World Health Organization (WHO/SDE/WSH/03.04/51).


Dichloramine New entry July 2005. There are insufficient data to establish a MAV for dichloramine.


1 To source waters

Chloramines may be present in source waters as a result of their discharge from industries in which they are used. Their principal use is as intermediates in the manufacture of hydrazine.


Dichloramine can be formed in chlorinated water that contains ammonia and some organic nitrogen compounds. The concentration depends upon the pH and chlorine to nitrogen ratio. Ammonia may be intentionally added to the water to produce the chloramines as disinfectants.


It is possible that reactions of chlorine with nitrogenous material in the distribution system may produce chloramines.

Typical concentrations in drinking-water No typical value data are available for New Zealand. Chloramination is not used intentionally at present as a disinfectant in New Zealand, and the concentrations of inorganic chloramines present in waters depends upon the concentrations of inorganic and some organic nitrogen compounds present in the raw water, and control of the chlorination process.

Removal methods Chemical reducing agents, including sodium thiosulphate, sulphur dioxide, and sodium bisulphite can be used to remove dichloramine. Activated carbon adsorbs dichloramine.


Analytical methods

Referee method

DPD Ferrous Titrimetric Method (APHA 4500-Cl F). The limit of detection for this method is approximately 0.2 mg/L for field use, although lower levels can be determined under laboratory conditions and with care. Analytical texts indicate that by manipulation of the conditions of the analysis measurement of monochloramine, dichloramine and trichloramine can be made. These methods are of use when ammonia only is in the water being chlorinated. In most natural waters nitrogen-containing organic compounds are also present. Organic chloramines are formed from these compounds when chlorine reacts with them. Organic chloramines also produce colour during the DPD test and make attempts to differentiate between the different inorganic chloramines of little value. Unless investigating taste and odour problems, it is recommended that only the total combined chlorine (ie, total chloramine concentration) is reported. These methods measure dichloramine in terms of mg Cl as Cl2/L.


1 Amperometric Titration Method (APHA 4500-Cl D).

While more accurate than the DPD methods, expensive equipment and a high degree of skill and care are required for this method. The limit of detection is better than 0.1 mg/L. The APHA method describes variations that will allow the determination of mono- and di-chloramine. Interferences due to organic chloramines may also cause interferences with these methods. 2 DPD Colorimetric Method (APHA 4500-Cl G).

This method requires a spectrophotometer for the colorimetric measurements, although hand-held comparators do offer a cheaper, though less reliable variation for field use. The limit of detection (LOD), with instrumental assistance, is approximately 0.1 mg/L. The LOD for the comparator depends on the colour disc in use. Chloramine concentrations as low as 0.2 mg/L approximately should be detectable, but the accuracy of the method depends upon use of the correct lighting (natural lighting should be used with the sun behind the viewer), the individual�s ability to match colours and judge their intensity, and ensuing that readings are taken as soon after colour development as possible. The LOD may be about 0.10 mg/L when using a Nessleriser. The same comment on the usefulness of trying to determine the individual chloramine concentrations made for the referee method, also applies to this method.

Health considerations Studies have revealed equivocal evidence of carcinogenic activity of chloraminated drinking-water in female rats, as indicated by an increase in incidence of mononuclear cell leukaemia. Epidemiological studies did not report an association between ingestion of chloraminated drinking-water and increased urinary bladder mortality rates in humans. When tap-water containing chloramines was used for dialysis, acute haemolytic anaemia, characterised by denaturation of haemoglobin and lysis of red blood cells, was reported in haemodialysis patients.







Fluoride Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of fluoride in drinking-water should not exceed 1.5 mg/L. The maximum contaminant level (USEPA 2004) is 4 mg/L; their secondary drinking water regulation for fluoride is 2.0 mg/L.


1 To source waters

Fluoride is present in trace amounts in soils and rocks, where it averages about 0.3 g/kg of the Earth�s crust. It is most prevalent in active or inactive volcanic regions. The most common fluoride containing minerals include fluorite and apatite, while it is also present due to the replacement of the hydroxide in hornblende and some micas. The weathering of alkaline and siliceous, igneous and sedimentary rocks, especially shales, supplies fluorides to the aquatic environment. Major occurrences of fluorite in New Zealand are known near Nelson, in the Buller Gorge, at Greymouth and on Stewart Island. Fluoride species are also rich in volcanic and geothermal fluids. Fluorides are used in a variety of industrial processes, including metal plating, metal casting, welding, brazing and the manufacture of aluminium, steel, bricks, tiles, glass and ceramics, hydrofluoric acid and other fluorine chemicals, adhesives and metallurgical fluxes. They are also used in insecticides and herbicides. It is a by-product of the fertiliser industry. Fluoride is added to toothpaste, tooth powders, mouth washes and vitamin supplements. Concentrations in surface water are relatively low (<0.1�0.5 mg/L), while water from deeper wells may have quite high concentrations (1�10 mg/L) if the rock formations are fluoride-rich.

2 From the treatment processes

Fluoride is often added to drinking-water supplies to protect teeth against dental caries.


No known sources.

Forms and fate in the environment Fluorine, being the most reactive element of the halogen series, does not occur free in nature, but is present as fluoride ion. In general, most fluorides associated with monovalent cations are water soluble (eg, NaF, AgF, and KF) but those formed with divalent cations are usually insoluble. Groundwater usually contains higher concentrations than surface water.


Typical concentrations in drinking-water Fluoride was routinely measured in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Of 1904 samples analysed between 1983 and 1989, 10 samples (0.5%) exceeded the guideline range of 0.9-1.1 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 166 zones, has found naturally-occurring fluoride concentrations to range from �not detectable� (nd) to 1.8 mg/L, with the median concentration being �nd� (limit of detection = 0.1 mg/L).

Removal methods Defluoridation of waters with high natural levels of fluoride can be achieved through the passage of the water through bone char, granular tricalcium phosphate, ion exchange resins or activated alumina. Activated alumina has several advantages: specificity for fluoride; a high exchange capacity not affected by other anions; low cost compared to synthetic anion exchange resins; and easily regenerated with sodium hydroxide and dilute acid. Lime-softening may remove fluoride by precipitation as calcium fluoride. Alum flocculation may remove fluoride by adsorbing the fluoride onto the aluminium hydroxide floc. Where fluoride is deliberately added to water, dose control is important to maintain the concentration within the allowable limit.

Analytical methods

Referee method

Ion Selective Electrode Method (APHA 4500-F C).


1 SPADNS Method (APHA 4500-F D). 2 Ion Chromatography Method (APHA 4500-F F).

Health considerations Fluoride is probably an essential element for animals and humans although for humans this has not been demonstrated unequivocally. Soluble fluorides are rapidly absorbed in the gastrointestinal tract after uptake via drinking-water. Fluoride has an affinity for mineralising tissues of the body: in young people, bone and teeth; in older people, bone. Thus excretion is somewhat greater in adults because they have proportionally less mineralising tissue than do children. People with kidney impairment have a lower margin of safety for fluoride intake and limited data indicate that their fluoride retention may be up to three times normal. Fluoride is often added to water as the silicofluoride ion; it can also complex with aluminium to form the AlF6

3- ion. At the pH of the stomach, these ions dissociate, freeing up the fluoride ion. Low concentrations of fluoride produce beneficial effects on the teeth, especially in children. This protective effect for caries increases with concentration up to 2 mg fluoride/L drinking-water. The effectiveness of fluoride in dental disease prevention and health considerations has been reviewed extensively and fluoridation of public water supplies is recommended by many health authorities around the world, including WHO, as an important health measure. Mild dental fluorosis (mottling of the teeth) may occur at a prevalence of 12�33% at drinking-water concentrations between 0.9 and 1.2 mg/L. This may become manifest in areas with a temperate climate and with concentrations between 1.5 and 2 mg/L.


Skeletal fluorosis, characterised by hypermineralisation and consequently brittle bones, has occurred in association with high fluoride concentrations in drinking-water and also with occupational exposure to fluoride-containing dust. It generally occurs after prolonged exposure (several years) and is reversible. The US Environmental Protection Agency considers a concentration of 4 mg/L to be protective against crippling skeletal fluorosis. The results of several epidemiological studies, concerned with a possible adverse effect of fluoride in drinking-water on pregnancy, are inconclusive. Tests for mutagenicity with strains of bacteria have been negative. Chromosome aberrations have been reported in tests with mammalian cells but only at extremely high fluoride concentrations. In 1987, the International Agency for Research on Cancer classified inorganic fluorides in Group 3 (not classifiable as to its carcinogenicity to humans). Fluoride has also not been carcinogenic in more recent studies in laboratory animals. Virtually all foodstuffs contain traces of fluoride, in particular, high amounts can be found in tea leaves because of natural concentration by the tea plant. Total daily intake from all sources varies considerably, but has been estimated at 0.46 mg to 5.4 mg, with about 10% coming from unfluoridated drinking water.

Derivation of maximum acceptable value WHO considered that there is no new evidence to suggest that the MAV of 1.5 mg/L, set in 1984, needs to be revised. Concentrations above this value carry an increasing risk of dental fluorosis, and much higher concentrations lead to skeletal fluorosis. This value is higher than that recommended for artificial fluoridation of water supplies. The Ministry of Health recommends that the concentration of fluoride in fluoridated drinking-water supplies be between 0.7 and 1.0 mg/L. The minimum concentration of fluoride required for a protective effect against dental caries is about 0.5 mg/L and concentrations around 1 mg/L in temperate climates are optimal for the prevention of caries. Mottling of teeth due to dental fluorosis may occur at concentrations between 1.5 and 2 mg/L.





IPCS. 2002. Fluorides. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 227).


NZGS. 1970. Minerals of New Zealand (Part B: Non-metallics) 38B.


WHO. 2003. Fluoride in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/96).



Iodine and iodide Updated July 2005.

Maximum acceptable value There are insufficient data to derive a MAV for iodine. The Australian Drinking Water Guidelines state that based on health considerations, the concentration of iodide in drinking water should not exceed 0.1 mg/L; no guideline value was set for molecular iodine. Sources To Drinking-water

1 To source waters

Iodine may be present in source waters because of natural contamination. Iodine is naturally present in sea water, nitrate minerals and seaweed, mostly in the form of iodide salts. It may be present in water due to leaching from salt and mineral deposits. It may also enter source waters as a result of its use as an antiseptic, as a sanitising agent in hospitals and laboratories, in pharmaceuticals and in photographic developing materials.


Iodine is occasionally used for the emergency disinfection of drinking-water for field use but it is not used for disinfecting larger drinking-water supplies. Iodide is oxidised to iodine by strong disinfectants such as chlorine during water treatment. The chlorination of waters containing significant amounts of iodide would result in the formation of iodinated disinfection by-products; such waters are rare.


No known sources.

Forms and fate in the environment Iodine occurs naturally in water in the form of iodide. Iodide is largely oxidised to iodine during water treatment and following this it will be incorporated into organic matter.

Typical concentrations in drinking-water Iodine has not been monitored routinely in New Zealand drinking-water supplies. The mean concentration of total iodine in USA drinking-water is 4 ug/L (0.004 mg/L) and the maximum concentration is 18 ug/L, probably predominantly as iodide. Iodine has a taste threshold in water of about 0.15 mg/L.

Removal methods It is unlikely that the concentration of iodine in drinking-water would ever be high enough to justify treating to remove it. Chemical reducing agents, such as sodium thiosulphate, could be used to reduce the iodine to iodide, but disinfection of the water using oxidants (eg, chlorine) could not be used subsequently, as the iodide would be converted back to iodine again.


Analytical methods

Referee method

A referee method cannot be selected for iodine because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for iodine for the above reason. However, the following information may be useful: Iodine cannot be preserved in a sample. Analysis must be carried out as soon after sampling as possible, and if a delay between sampling and analysis is unavoidable the sample must be kept dark and chilled. Iodine can be analysed using the Leuco Crystal Violet Method (APHA 4500-I B). The limit of determination of this method is approximately 0.01 mg/L.

Health considerations Iodine is an essential trace element for humans and is used in the synthesis of thyroid hormones. Estimates of the dietary requirement for adult humans range from 80 to 150 mg/day; in many parts of the world, there are dietary deficiencies in iodine. Molecular iodine is rapidly converted to iodide following ingestion and is absorbed efficiently throughout the gastrointestinal tract. 70-80% of the total iodine content in the human body is found in the thyroid gland. Muscle and the eye also contain high iodide concentrations. High oral doses (more than 30 mg/kg body weight) of iodine can be lethal. Lower doses (3.3 mg/kg body weight) have been used to treat asthmatic patients without adverse effects. Chronic iodide exposure results in iodism. Symptoms resemble those of a sinus cold and may also include salivary gland swelling, gastrointestinal irritation, acneform skin, metallic or brassy taste, gingivitis, increased salivation, conjunctival irritation and oedema of the eyelids. Long-term consumption of iodinated drinking-water has not been associated with adverse health effects in humans. Prisoners� drinking-water containing up to 1 mg/L iodine for five years showed no signs of iodism or hypothyroidism, but some changes in uptake of iodine by the thyroid gland were observed. No data are available on the mutagenic activity of iodine.

Derivation of maximum acceptable value WHO (2004) did not establish a guideline value for iodine because available data suggest that derivation of a guideline value for iodine on the basis of information on the effects of iodide is inappropriate and there are few relevant data on the effects of iodine; also, because iodine is not recommended for longterm disinfection, lifetime exposure to iodine concentrations such as might occur from water disinfection is unlikely.



WHO. 2003. Iodine in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/46).



Lead Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of lead in drinking-water should not exceed 0.01 mg/L. The action level (USEPA 2004) is 0.015 mg/L at the tap.


1 To source waters

The principal natural pathway by which lead is released into the environment is weathering of sulphide ores, especially galena (PbS). In New Zealand, occurrences of galena are generally in intermediate volcanic rocks or associated with granitic rocks. Examples of major occurrences include on the Coromandel Peninsula; at Broadlands, Taupo; and around Nelson, Fiordland and Otago. Anthropogenic input of lead to the environment outweighs all natural sources. Lead reaches the aquatic environment through precipitation, fallout of lead dust, street run-off and industrial and municipal wastewater discharges. It has been estimated that in New Zealand, 99% of the lead emitted to the atmosphere comes from petrol combustion in vehicle engines. Lead is used in the manufacture of acid-storage batteries, alkyllead additive for petrol, construction materials, coatings and dyes, electronic equipment, plastics, veterinary medicines, fuels, radiation shielding, ammunition, corrosive-liquid containers, paints, glassware, fabricating storage tank liners, transporting radioactive materials, solder, piping, cable sheathing, roofing, and sound attenuators. It is also used in electroplating and the metallurgy industries.


No known sources.


Lead is rarely present in tap water as a result of its dissolution from natural sources; rather, its presence is primarily from household plumbing systems and fittings. Lead may be released into drinking-water from the dissolution of lead pipes and solders, and brass fittings in homes. The weathering of lead-based paints and fittings may also be a concern in rain-water supplies. It has also been found that the lead used as a stabiliser in uPVC pipes will leach from these pipes for several months after the pipes are installed. This has led to high elevated levels of lead in the water at the dead-ends of new subdivisions, until the leaching process has run its course. The amount of lead dissolved from plumbing systems depends, amongst other things, on pH, temperature, water hardness and standing time of the water.

Forms and fate in the environment The principal dissolved inorganic forms of lead are the free ion, and probably the carbonate and sulphate ion pairs. The importance of organic complexes is uncertain, but they may constitute a significant part of the dissolved lead in some waters. Soluble lead is removed from solution by association with sediments and suspended particulates, such as organic matter, hydrous oxides and clays.


Typical concentrations in drinking-water Lead was routinely measured in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Of 1900 samples analysed between 1983 and 1989, 15 samples (0.8%) had concentrations equal to or exceeding the MAV of 0.05 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 921 zones, found lead concentrations to range from �not detectable� (nd) to 0.98 mg/L, with the median concentration being 0.001 mg/L (limit of detection = 0.001 mg/L). WHO (2004) states that concentrations in drinking-water are generally below 0.005 mg/L, although much higher concentrations (above 0.1 mg/L) have been measured where lead fittings are present.

Removal methods The principal problem with lead in drinking-waters comes from corrosion of the distribution system (lead solder, pipes, weathering of lead paints). Corrosion can be minimised by:

• avoiding the use of lead plumbing

• calcium carbonate saturation, to seal lead surfaces with a non-toxic barrier of calcium carbonate

• pH and carbonate adjustment, to form a lead hydroxy-carbonate compound with minimum lead solubility. Correct pH and carbonate concentration are essential

• orthophosphate addition, to form a low solubility lead hydroxy-phosphate compound. Coagulation/flocculation with iron or aluminium salts has been shown to be effective in removing lead, especially where the lead is sorbed on to particles. Lime-softening removes lead in the form of Pb(OH)2. PAC and GAC have also been reported to be effective in removing lead.

Analytical methods

Referee method



1 Inductively Coupled Plasma Method (APHA 3120). 2 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Lead can be absorbed by the body through inhalation, ingestion or placental transfer. In adults, approximately 10% of ingested lead is absorbed. However, young children absorb 4�5 times as much and the biological half-life is considerably longer in children than in adults. After absorption, the lead is distributed in soft tissue such as the kidney, liver and bone marrow where it has a biological half-life in adults of less than 40 days, and in skeletal bone where it can persist for 20 to 30 years. Lead is a cumulative general poison. Infants, children up to six years of age, the foetus and pregnant women are the most susceptible to adverse health effects. Placental transfer of lead occurs in humans as early as the 12th week of gestation and continues throughout development. Lead can severely affect the central nervous system. Signs of acute intoxication include dullness, restlessness, irritability, poor attention span, headaches, muscle tremor, abdominal cramps, kidney damage, hallucinations, and loss of memory. Encephalopathy occurs at blood lead levels of 100�120 µg/dL in adults and 80�100 µg/dL in children. Signs of chronic toxicity may appear in adults


with blood lead levels of 50�80 µg/dL. Symptoms include tiredness, sleeplessness, irritability, headaches, joint pains and gastrointestinal symptoms. Many epidemiological studies have been carried out on the effects of lead exposure on the intellectual development of children. Although there are some conflicting results, on balance the studies demonstrate that exposure to lead can adversely affect intelligence. Research on primates has supported the results of the epidemiological studies, in that significant behavioural and cognitive effects have been observed following postnatal exposure. Other adverse effects associated with exposure to high amounts of lead include kidney damage, interference with the production of red blood cells, and interference with the metabolism of calcium needed for bone formation. Renal tumours have been induced in experimental animals exposed to high concentrations of lead compounds in the diet, and the International Agency for Research on Cancer has classified lead and inorganic lead compounds in Group 2B (possibly carcinogenic to humans). However, there is evidence from studies in humans that adverse neurotoxic effects other than cancer may occur at very low concentrations of lead and that a MAV derived on this basis would also be protective for carcinogenic effects. Owing to the decreasing use of lead-containing additives in petrol and of lead-containing solder in the food processing industry, concentrations in air and food are declining, and intake from drinking-water constitutes a greater proportion of total intake.

Derivation of maximum acceptable value As there is evidence from human studies that adverse effects other than cancer may occur at very low levels of lead, and that a guideline thus derived would also be protective for carcinogenic effects, it is considered appropriate to derive the MAV using a tolerable daily intake approach. In 1986, the Joint FAO/WHO Expert Committee on Food Additives established a provisional tolerable weekly intake (PTWI) of 25 µg lead/kg body weight (equivalent to 3.5 µg/kg body weight per day) for infants and children on the basis that lead is a cumulative poison and that there should be no accumulation of body burden of lead. The PTWI was based on metabolic studies in infants showing that a mean daily intake of 3�4 µg/kg body weight was not associated with an increase in blood lead levels or in the body burden of lead, whereas an intake of 5 µg/kg body weight or more resulted in lead retention. The MAV for lead in drinking-water has been derived using the case of a bottle fed baby. As infants are considered to be the most sensitive subgroup of the population, this MAV will also be protective for other age groups. The MAV was derived as follows:

0.0035 mg/kg body weight per day x 5 kg x 0.5 = 0.012 mg/L (rounded to 0.01 mg/L) 0.75 L per day where: • tolerable daily intake = 0.0035 mg/kg body weight per day • average weight of a bottle fed baby = 5 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.5 • average amount of water consumed by a bottle fed baby = 0.75 L per day.




AWWA. 1990. Water Quality and Treatment (4th ed). Pontius FW, McGraw, Hill (eds).





Sorg TJ, Csanady M, Logsdon GS. 1978. Treatment technology to meet the Interim Primary Drinking Water Regulations for Inorganics: Part 3. JAWWA 70(12): 680�91.

USEPA, Method 200.8, Determination of Trace Elements in Waters and Wastes by Inductively Coupled Plasma-Mass Spectrometry.


WHO. 2003. Lead in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/9).


Lithium New entry July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of lithium in drinking-water should not exceed 0.9 mg/L (900 µg/L).


1 To source waters

Lithium is an alkali metal that has properties similar to potassium and sodium. It is found in nature in nearly all igneous rocks, and in geothermally influenced water and natural brines. Lithium can enter water naturally through its dissolution from rocks and also through its disposal in wastewater at sites mining spodumene, lepidolite, petalite, amblygonite or triphylite (Merck & Co 1996). Lithium can also enter drinking-water through its use in a variety of industries including alloys for the aircraft industry, lithium carbonate and borate as ceramics, air-conditioning, welding and brazing fluxes (as lithium chloride and fluoride) and lubrication grease (as organo-lithium compounds). It is used in long-life batteries (including pacemakers and lifebuoys). In countries other than New Zealand (which is nuclear free) it is used in some nuclear applications as a coolant. Lithium salts an be used as a tracer. Lithium metal and lithium compounds are used by drug manufacturers in the synthesis of intermediates for important pharmaceuticals and in the production of lithium carbonate, which is widely used for treatment of clinical depression. Lithium metal is a key reagent in the production of Vitamin A and the hormones beta-methasone and ethinylestradiol. Lithium bromide and lithium carbonate are used in the process of steroidal synthesis.


Forms and fate in the environment Lithium does not occur in free form in nature and is highly soluble in water (it becomes hydrolysed and the metal presents a fire and explosion risk when exposed to water). The reaction forms lithium hydroxide and hydrogen gas. Lithium is a common element and is present in many plant and animal tissues. Daily intake of lithium is about 2 mg (Klaassen 1995).

Typical concentrations in drinking-water Lithium was included in Ministry of Health surveillance programmes until 1995, when the Drinking-water Standards for New Zealand 1995 (MoH 1995) were introduced. At this stage, analysis was halted because lithium was not listed as a health significant determinand in the 1995 DWSNZ. 7490 samples were collected as part of the Ministry of Health�s five-yearly surveillance programmes (ref: WINZ and STANLEY databases). The results ranged from nd (not detected) to 0.43 mg/L, with a median concentration of nd.

Removal methods No information is available on methods for removal of lithium in drinking-water.

Analytical methods

Referee method

Flame emission (APHA 3500-Li B).


Inductively coupled plasma � mass spectrometer (EPA 200.8).

Health considerations Lithium is readily absorbed from the gastrointestinal tract. Distribution in human organs is almost uniform. Excretion is chiefly through the kidneys, but some is eliminated in the faeces. In general the distribution of lithium in the body is quite similar to that of sodium and potassium, and it may compete with these elements at the renal tubular level, VDH (1997).

Acute poisoning

The oral toxicity of most lithium compounds is relatively low, oral LD50 values for several compounds and animal species range from 422�1165 mg/kg (Opresko 1995).

Chronic exposure

There is ample evidence that long-term exposure to lithium at low levels causes no serious adverse health effects. Because lithium is used routinely as a drug for the treatment of clinical depression, there is a large body of data on the human health effects of long term exposure to lithium. Lithium is administered therapeutically as the carbonate salt (lithium carbonate) in daily oral doses of 900�1800 mg/day for the treatment of manic and endogenous depression. The therapeutic use of lithium carbonate may produce unusual toxic responses. These include neuromuscular changes (eg, tremor, muscle hyperirritability, and ataxia (muscle inco-ordination), central nervous system changes (blackout spells, epileptic seizures, slurred speech, coma, psychosomatic


retardation), and increased thirst, cardiovascular changes (cardiac arrhythmia, hypertension, and circulatory collapse), gastrointestinal changes (anorexia, nausea, and vomiting), and renal damage (albuminuria and glycosuria) Klaassen 1995; Merck & Co 1996; VDH 1997). Lithium treatment is not recommended for patients with significant renal or cardiovascular disease, severe debilitation or dehydration, or sodium depilation, or to patients receiving other medications (eg, diuretics), because the risk of lithium toxicity is high in such patients. Data from lithium birth registries suggest an increase in cardiac and other abnormalities in infants born to women who have been prescribed treatment with lithium carbonate. Animal studies have shown lithium carbonate is a weak teratogen (causes birth defects). Increased frequency of cleft palate and faetal loss have been observed among the offspring of mice treated chronically during pregnancy with lithium carbonate. However, studies in rats, rabbits and monkeys gave no evidence of lithium-induced birth defects, VDH (1997). To protect human health, the USEPA estimated that a lithium concentration in a potable water supply should not exceed 0.7 mg/L (cited in VDH 1997). The International Agency for Research on Cancer (IARC) has not classified lithium for its ability to cause cancer.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for lithium in drinking-water, as follows:

NOAEL mg/kg body weight per day x 70 kg x proportion from d/w = 0.9 mg/L 2 L x uncertainty factor

where: • no observable adverse effect level = X mg/kg body weight per day • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = X% • uncertainty factor = X.


Klaassen C. 1995. Casarett and Doull�s Toxicology: The basic science of poisons (5th ed). McGraw-Hill, Health Professions Division.

Merck & Co. 1996. The Merck Index: An encyclopedia of chemicals, drugs and biologicals (12th ed). Whitehouse Station, New Jersey: Merck Research Laboratories, Division of Merck & Co Inc.

Opresko DM. 1995. Toxicity Summary for Lithium. Oak Ridge National Laboratory, US: Chemical Hazard Evaluation Group.

VDH. 1997. Factsheet on Lithium. Virginia Department of Health: Bureau of Toxic Substances.

WINZ and STANLEY databases. Ministry of Health drinking-water quality information bases. Held by ESR, Christchurch.

Manganese Updated July 2005.


Maximum acceptable value Based on health considerations, the concentration of manganese in drinking-water should not exceed 0.5 mg/L. Based on aesthetic considerations, the concentration of manganese in drinking-water should not exceed 0.04 mg/L. The USEPA has a secondary drinking water regulation of 0.05 mg/L for manganese.


1 To source waters

Manganese can reach the aquatic environment from the weathering of rocks and minerals and runoff from soils. Manganese is not an essential constituent of any of the more common silicate rock minerals, but it can substitute for iron, magnesium or calcium in silicate structures. Many igneous and metamorphic minerals contain manganese as a minor constituent. It is a significant constituent of basalt and many olivines and of pyroxene and amphibole. Small amounts are present in dolomite and limestone, substituting for calcium. Manganese may enter water from industrial discharges and agricultural runoff. Manganese and its compounds are used in the steel industry in the manufacture of metal alloys, in the manufacture of dry cell batteries, paints, varnishes, inks, dyes, glass, ceramics, matches, fire works and fertilisers. Manganese is also used in animal feeds. Bottom waters in lakes and reservoirs can become very low in dissolved oxygen. Under these conditions iron and manganese leach out from the sediments and begin to circulate through the water column, resulting in raw water concentrations that may be too high for the treatment process to handle. Uncontaminated rivers and streams generally have low concentrations of manganese, ranging from 0.001 mg/L to 0.6 mg/L.


No known sources.


There are situations in which manganese concentrations in the water at the consumers� taps can be higher than those entering the distribution system. This is not the result of manganese being dissolved from reticulation materials. It arises from manganese, in either soluble or insoluble form, passing into the distribution system. Insoluble manganese may settle out in pipes, and soluble manganese may be oxidized to insoluble forms, by oxygen or other chemical oxidants, such as chlorine, and also settle. Changes in water flows through the system may then resuspend the particulate manganese, which may lead to �black� water at the consumers� taps. Certain nuisance organisms concentrate manganese and give rise to taste, odour and turbidity problems in distributed water.

Forms and fate in the environment Manganese has three valence states in natural environments (2+, 3+, 4+). In the absence of dissolved oxygen Mn(II) predominates, otherwise it is readily oxidised to Mn(IV). In natural oxygenated waters, a substantial fraction of manganese is present in suspended form. In surface waters, divalent manganese will be oxidised to manganese dioxide which will undergo sedimentation. In the presence of complex-forming inorganic and organic compounds, the colloidal stability of manganese oxides will be enhanced. Alternatively, in areas of low dissolved oxygen or in anaerobic areas at low pH, soluble manganese forms may persist. Many of the groundwaters reported to carry large manganese concentrations are from thermal springs.


Typical concentrations in drinking-water Manganese was measured routinely in New Zealand drinking-water supplies as an aesthetic parameter as part of the Department of Health three yearly surveillance programme. It is now classified as an inorganic parameter of health significance. Of 1143 samples analysed from 913 supplies between 1983 and 1989, 91 samples (9.3% of supplies) were equal to or exceeded the highest desirable level (ie, GV) of 0.05 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 400 zones, found manganese concentrations to range from �not detectable� (nd) to 1.7 mg/L, with the median concentration being 0.002 mg/L (limit of detection = 0.001 mg/L).

Removal methods Oxidation of Mn(II) to insoluble Mn(IV) compounds is a commonly employed technique. Aeration uses oxygen from the air to achieve oxidation. This precipitates the manganese, which is either allowed to settle or is removed by filtration. The rate at which oxidation occurs is pH dependent, becoming faster as the pH is increased. Some Mn(II) is adsorbed on to higher oxidation states of Mn in slightly alkaline solution. A coating of higher oxides of manganese on filter granules acts to catalyse the removal of lower oxidation states. Organically-bound manganese is not removed by aeration. Other oxidising agents such as chlorine and potassium permanganate may be used for the removal of manganese. A pH of 8 is required for chlorine oxidation, and pH 7 to 8 is optimum for permanganate oxidation. Sand or anthracite filters can be used to filter the precipitated manganese from the water. Before the removal process becomes efficient, a coating of iron and manganese oxides must develop on the grains of the filter medium. Until this coating develops, removal may be poor. Natural zeolites (ion exchange materials) treated with manganese can be used as the filter medium. This medium is known as greensand, and requires periodic regeneration of the iron and manganese oxide coating. Mn(II) ions come into contact with the zeolite and are converted to the insoluble oxide which is filtered out by the filter bed. Natural zeolites or synthetic resins can also be used to remove manganese by a true ion-exchange process. Sodium attached to the zeolite is exchanged for Mn(II) ions in the incoming water. It is important that the water is free of oxygen that might lead to oxidation and precipitation of the manganese, as this will foul the zeolite. Manganese, in the absence of oxygen, can be removed effectively by raising the pH above approximately 10 to precipitate manganese hydroxide. This can be exploited if the lime-soda ash process is being used for hardness reduction.

Analytical methods

Referee method



1 Flame Atomic Absorption Spectrometric Method (APHA 3111). 2 Inductively Coupled Plasma Method (APHA 3120B). 3 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).


Health considerations Manganese is an essential trace element with an estimated daily nutritional requirement of 30�50 µg/kg body weight. The greatest exposure to manganese is usually from food. Its absorption rate can vary considerably according to actual intake, chemical form, and presence of other metals such as iron and copper. Typically, only about 3�8% of ingested manganese is absorbed by the gastro-intestinal tract. In infants and young animals, very high absorption rates of manganese have been observed. After absorption it is concentrated in the liver and eventually excreted in faeces. It has a relatively short biological half-life of 13 to 37 days in humans. Manganese deficiency affects bone, the brain and reproduction in a number of species. Evidence of manganese neurotoxicity has been seen in miners following prolonged exposure to manganese dusts. There is no convincing evidence of toxicity in humans associated with the consumption of manganese in drinking-water, but there are only limited studies available. By the oral route, manganese is often regarded as one of the least toxic elements. In one case, the symptoms associated with consuming drinking-water containing a manganese concentration of close to 28 µg/L included lethargy, increased muscle tone, tremor and mental disturbances. However, the concentrations of other metals were also high and the reported effects may not be due to manganese alone. Experiments with animals have shown no adverse effects, other than a change in appetite and a reduction in the metabolism of iron in haemoglobin synthesis. Some in vitro studies have reported mutagenic activity for manganese on mammalian cells and bacteria. There is no firm evidence that manganese is carcinogenic. Some studies indicate that it may, in fact, have an anti-carcinogenic effect.

Derivation of maximum acceptable value The MAV for manganese in drinking-water is based on the upper range value of manganese intake of 11 mg/day, identified using dietary surveys, at which there are no observed adverse effects (ie, considered a NOAEL), using an uncertainty factor of 3 to take into consideration the possible increased bioavailability of manganese from water. This results in a TDI of 0.06 mg/kg of body weight. The MAV was derived as follows:


where: • tolerable daily intake = 0.06 mg/kg body weight per day • average adult weight = 70 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.2 • average amount of water consumed per day = 2 L per day. The aesthetic guideline value for mangansese is 0.1 mg/L due to the fact that it deposits in water mains and causes discoloration when scoured out. Some water supplies may need to aim for a lower concentration to prevent the build-up. At concentrations exceeding 0.1 mg/L, manganese can impart an undesirable taste to water.







IPCS. 1999. Manganese and its Compounds. Geneva: World Health Organization, International Programme on Chemical Safety (Concise International Chemical Assessment Document 12).





WHO. 2003. Manganese in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/104).


Mercury Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of mercury in drinking-water should not exceed 0.002 mg/L (2 µg/L). The maximum contaminant level (USEPA 2004) is 0.002 mg/L.


1 To source waters

Mercury can enter the aquatic environment from the weathering of rocks and minerals and run-off from soils and due to geothermal activity. Cinnabar (HgS) is the most common mercury containing mineral. In New Zealand, mercury occurs in Permian-Triassic marine volcanic associations at Moumoukai, Hunua Ranges at Kakariki and Waihohine River, Wairarapa and Greenvale, Southland; in late Cenozoic Hydrothermal deposits at Puhipuhi and Ngawha Springs, Northland, Hauraki Goldfield and Lake Waikare, Lower Waikato; and in minor alluvial occurrences in Southland, Westland, Northland and the Hauraki Goldfields. High levels of mercury are also present in geothermal fluids and natural springs. Industrial discharges can also contribute to the mercury in water. Mercury is used in the paint industry in paint pigments and preservatives, including anti-fouling paints, in pulp and paper manufacture, and in the production of thermometers, electrical equipment (eg, mercury switches, batteries and fluorescent and mercury vapour lamps) dental amalgams and therapeutic medicinal compounds.


Apart from traces in chlorine gas, no known sources.



No known sources.

Forms and fate in the environment Mercury can exist in three oxidation states in the natural environment: 0, +1 and +2. The stable form of mercury in most natural water systems is the free metal. However, mercury in surface water open to the atmosphere is liable to escape by vaporisation. In freshwaters, it is common for the mercury to be adsorbed to suspended particulate matter and to the sediments.

Typical concentrations in drinking-water Mercury was not measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Highest levels of mercury in water are found in some thermal groundwaters and in water associated with mercury ores or in mining areas. The P2 Chemical Determinand Identification Programme, sampled from 114 zones, found mercury concentrations to range from �not detectable� (nd) to 0.0011 mg/L, with the median concentration being �nd� (limit of detection = 0.0005 mg/L). Concentrations of total mercury in natural water are generally so low that accurate analysis is dif. cult. Studies overseas have reported concentrations of less than 0.0005 mg/L, with some sources less than 0.00003 mg/L (30 ng/L). The highest value was 0.0055 mg/L from some wells in Japan. In major Australian reticulated supplies, the concentrations of total mercury range up to 0.001 mg/L, with typical concentrations usually less than 0.0001 mg/L.

Removal methods Inorganic and methyl mercury can be removed by filtration through GAC. Removal is better in the presence of ferric sulphate than in its absence, and removal is pH dependent with efficiency decreasing with increasing pH over the range 7 to 9. Co-precipitation of Hg(OH)2 with Fe(III) hydroxide is effective. Coagulation/flocculation with iron or aluminium salts removes mercury, to an extent determined by the turbidity. pH has very little effect on removal efficiency. Organic mercury is not removed by coagulation, nor is it removed by lime-softening.

Analytical methods

Referee method

Cold-Vapour Atomic Absorption Method (APHA 3112B).


No alternative methods have been recommended for mercury because no methods meet the required criteria.

Health considerations

Inorganic mercury

Absorption of inorganic mercury from drinking-water by the gastro-intestinal tract may be less than 15%. Inorganic mercury compounds are rapidly accumulated in the kidney, the main target organ for mercury toxicity. The biological half-life of mercury is very long, probably years.


Chronic exposure of workers to mercury has resulted in reported health effects including tremors, mental disturbances and gingivitis (inflammation of the mucous membrane surrounding the teeth). The main toxic effects are to the kidney, leading to kidney failure. In general, acute lethal toxic doses by ingestion of any form of mercury will result in symptoms including shock, cardiovascular collapse, acute renal failure and severe gastrointestinal damage. Various reports indicate that inorganic mercury binds to, and damages, mammalian DNA. Some evidence of carcinogenicity in rats has been reported.

Organic mercury

Organic mercury compounds are unlikely to be found in uncontaminated drinking-water. However, the toxic effects are more severe than those of inorganic mercury. Methylmercury compounds are almost completely absorbed by the gastrointestinal tract. Methylmercury has greater lipid solubility than inorganic mercury which permits it to cross biological membranes more easily, especially in the brain, spinal cord, peripheral nerves and the placenta. Methylmercury affects the central nervous system and the main effects of poisoning are irreversible neurological disorder and mental disability. Two major epidemics of methylmercury poisoning in Japan, known as Minimata disease, were caused by the industrial release of methyl and other mercury compounds. The mercury was accumulated by edible fish, which were then consumed by humans. A number of other countries have suffered epidemics associated with the consumption of bread prepared from cereals treated with mercury contaminated fungicide.

Derivation of maximum acceptable value Almost all mercury in uncontaminated drinking-water is thought to be in the form of Hg2+. Thus, intake of organic mercury compounds, especially the alkyl mercurials, is unlikely to pose a direct risk through ingestion of drinking-water. However, the conversion of methylmercury to inorganic mercury may occur. In 1972 the Joint FAO/WHO Expert Committee on Food Additives established a provisional tolerable weekly intake of 5 µg/kg of body weight of total mercury, of which no more than 3.3 µg/kg body weight should be present as methylmercury. In 1988, JECFA reassessed methylmercury, as new data were only then available on this compound, and confirmed the previously recommended provisional tolerable weekly intake (PTWI) of 3.3 µg/kg of body weight for the general population. It should be noted that pregnant women and nursing mothers are likely to be at a greater risk from the adverse effects of methylmercury. The available data were considered insufficient to recommend a specific methylmercury intake for this population group. To be on the conservative side, WHO has used the intake for methylmercury to derive a MAV for inorganic mercury in drinking-water. The MAV for mercury in drinking-water has been derived as follows:

0.00047 mg/kg body weight per day x 70 kg x 0.1 = 0.00165 mg/L (rounded to 0.002) 2 L per day

where: • tolerable daily intake = (0.003.3/7) 0.00047 mg/kg body weight per day • average weight of an adult = 70 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.1 • average amount of water consumed by an adult = 2 L per day.








WHO. 2003. Mercury in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/10).


Molybdenum Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of molybdenum in drinking-water should not exceed 0.08 mg/L.


1 To source waters

Molybdenum is widely distributed in nature, occurring chiefly as molybdenite and molybdates. The weathering of igneous and sedimentary rocks (especially shales) and run-off from soils constitutes an important natural source of molybdenum to the aquatic environment. In New Zealand disseminated molybdenite occurs in granitic rocks around Nelson and in quartz and pegmatite veins with molybdenite and chalcopyrite close to granitic contacts in the South Island (eg, Mt Radiant). Minor molybdenite-chalcopyrite mineralisation has also been identified in Westland, Southland, Fiordland and near Reefton. Molybdenum can also enter water via the discharge of industrial wastes from industries in which it is used and agricultural run-off. Molybdenum and its compounds are used in the manufacture of special steel alloys, electronic apparatus, glass, ceramics, pigments, as lubricants in oils and greases, and in fertilisers to overcome molybdenum deficiency in soils.


No known sources.


No known sources.

Forms and fate in the environment In natural waters molybdenum occurs in the +4 and +6 oxidation states as molybdenum sulphide (MoS2) and molybdate anion (MoO4

2-) respectively. In aerobic waters it is the soluble molybdates which are


stable. Adsorption and co-precipitation of the molybdate anion by hydrous oxides of iron and aluminium play primary roles in determining the aquatic fate of molybdenum.

Typical concentrations in drinking-water Molybdenum was not measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. The P2 Chemical Determinand Identification Programme, sampled from 831 zones, found molybdenum concentrations to range from �not detectable� (nd) to 0.003 mg/L, with the median concentration being �nd� (limit of detection = 0.005 mg/L).

Removal methods There are no published methods for the removal of molybdenum from drinking-water.

Analytical techniques

Referee method



1 Inductively Coupled Plasma Method (APHA 3120B). 2 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Molybdenum is considered to be an essential element, with an estimated daily requirement of about 0.075�0.25 mg for adults. Many foods contain significant amounts of molybdenum. Legumes, grains and liver have the highest concentrations and food is a significant source of intake. In humans, 30�70% of dietary molybdenum is absorbed from the gastro-intestinal tract and appears rapidly in the blood and most organs, and is excreted primarily in the urine. Molybdenum crosses the placental barrier. There is no apparent bioaccumulation of molybdenum in human tissues. Few data are available on the long- and short-term toxicity of molybdenum in humans. One study has linked high intake of molybdenum in food with gout-like symptoms, joint pains of the hands and legs and enlargement of the liver. No data are available on the carcinogenicity of molybdenum by the oral route. Tests for mutagenicity with bacteria have been inconclusive.

Derivation of maximum acceptable value In a two-year study of humans exposed through their drinking-water, a no observable adverse effects level of 0.2 mg/L was identified for molybdenum. There are some concerns about the quality of this study although it has been used for the derivation of the MAV. An uncertainty factor of 10 would normally be applied to reflect intraspecies variation. However, it is recognized that molybdenum is an essential element, and therefore a factor of 3 is considered to be adequate. This gives a MAV of 0.07 mg/L, which is in the same range as a value derived on the basis of the results of toxicological studies in animal species and is consistent with the essential daily requirement for molybdenum.







WHO. 2003. Molybdenum in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/11).


Monochloramine New entry July 2005.

Maximum acceptable value Based on health considerations, the concentration of monochloramine in drinking-water should not exceed 3.3 mg/L. WHO (2004) talks of mg/L monochloramine. Analytical techniques report chloramines as mg Cl as Cl2/L. The monochloramine MAV of 3.3 mg/L can be expressed as 4.5 mg/L Cl as Cl2. The maximum contaminant level (USEPA 2004) is 4 mg/L measured as chlorine.


1 To source waters

Monochloramine may be present in source waters as a result of discharge from industries in which it is used. Its principal use is as intermediates in the manufacture of hydrazine.


Monochloramine can be formed in chlorinated water that contains ammonia and some organic nitrogen compounds. The concentration of monochloramine depends upon the pH and chlorine to nitrogen ratio. Ammonia may be added intentionally to the water to produce monochloramine as a disinfectant.


It is possible that reactions of chlorine with nitrogenous material in the distribution system may produce monochloramine. Most individuals are able to taste chloramines at concentrations below 5 mg/L, and some at levels as low as 0.3 mg/L.

Forms and fate in the environment Monochloramine is persistent in the environment and its rate of disappearance is a function of pH and salinity: its half-life increases with increasing pH and decreases with increasing salinity. Monochloramine decomposes faster if discharged into receiving waters containing bromide, presumably by the formation of bromochloramine and decomposition of the dihalamine.


Typical concentrations in drinking-water No typical value data are available for New Zealand. Monochloramine is not intentionally used at present as a disinfectant in New Zealand, and the concentrations of inorganic chloramines present in waters depends upon the concentrations of inorganic and some organic nitrogen compounds present in the raw water, and control of the chlorination process.

Removal methods Chemical reducing agents, including sodium thiosulphate, sulphur dioxide, and sodium bisulphite can be used to remove monochloramine. Activated carbon adsorbs monochloramine.

Analytical methods

Referee method

DPD Ferrous Titrimetric Method (APHA 4500-Cl F). The limit of detection for this method is approximately 0.2 mg/L for field use, although lower levels can be determined under laboratory conditions and with care. Analytical texts indicate that by manipulation of the conditions of the analysis measurement of monochloramine, dichloramine and trichloramine can be made. These methods are of use when ammonia only is in the water being chlorinated. In most natural waters nitrogen-containing organic compounds are also present. Organic chloramines are formed from these compounds when chlorine reacts with them. Organic chloramines also produce colour during the DPD test and make attempts to differentiate between the different inorganic chloramines of little value. Unless investigating taste and odour problems, it is recommended that only the total combined chlorine (ie, total chloramine concentration) is reported. These methods measure monochloramine in terms of mg Cl as Cl2/L. The MAV of 3.3 mg/L monochloramine can be converted to mg Cl as Cl2/L based on molecular weights as follows:

3.3 mg/L monochloramine = 71/51.3 x 3 = 4.5 mg Cl as Cl2/L

where: • 71 is the molecular weight of chlorine (Cl2) • 51.5 is the molecular weight of monochloramine (NH2Cl).



While more accurate than the DPD methods, expensive equipment and a high degree of skill and care are required for this method. The limit of detection is better than 0.1 mg/L. The APHA method describes variations that will allow the determination of mono- and di-chloramine. Interferences due to organic chloramines may also cause interferences with these methods. 2. DPD Colorimetric Method (APHA 4500-Cl G)

This method requires a spectrophotometer for the colorimetric measurements, although hand-held comparators do offer a cheaper, though less reliable variation for field use. The limit of detection (LOD), with instrumental assistance, is approximately 0.1 mg/L. The LOD for the comparator depends on the colour disc in use. Chloramine concentrations as low as 0.2 mg/L approximately should be detectable, but the accuracy of the method depends upon use of the correct lighting (natural lighting should be used with the sun behind the viewer), the individual�s ability to match colours and judge their intensity, and ensuing that readings are taken as soon after colour development as possible. The LOD may be about 0.10 mg/L when using a Nessleriser.


The same comment on the usefulness of trying to determine the individual chloramine concentrations made for the referee method, also applies to this method.

Health considerations Monochloramine is readily absorbed by the gastrointestinal tract following oral administration in the rat. Following administration, the highest concentration was contained in the plasma. Monochloramine is metabolised to the chloride ion which is excreted mainly in the urine. Monochloramine has been reported to be weakly mutagenic with some bacteria but did not increase bone marrow chromosomal aberrations. Studies have revealed equivocal evidence of carcinogenic activity of chloraminated drinking-water in female rats, as indicated by an increase in incidence of mononuclear cell leukaemia. Epidemiological studies did not report an association between ingestion of chloraminated drinking-water and increased urinary bladder mortality rates in humans. When tap-water containing chloramines was used for dialysis, acute haemolytic anaemia, characterised by denaturation of haemoglobin and lysis of red blood cells, was reported in haemodialysis patients.

Derivation of maximum acceptable value The MAV for monochloramine has been derived using a tolerable daily intake approach as follows:

9.4 mg/kg body weight per day x 70 kg x 1 = 3.3 mg/L 2 L per day x 100

where:

• no observable adverse effects level = 9.4 mg/kg body weight per day. This is the highest dose administered to male rats in a two-year drinking-water study. It was chosen because of the probability that the lower body weights were caused by the unpalatability of the drinking-water

• average adult weight = 70 kg

• the proportion of the tolerable daily intake attributable to monochloramine = 1

• the average quantity of water consumed by an adult = 2 L per day

• uncertainty factor = 100 for intra- and interspecies variation. An additional uncertainty factor for possible carcinogenicity was not applied because equivocal cancer effects reported in the NTP study in only one species and in only one sex were within the range observed in historical controls.





WHO. 2003. Monochloramine in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/83).



Nickel Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of nickel in drinking-water should not exceed 0.02 mg/L. The WHO guideline value is considered provisional owing to uncertainties about the effect level for perinatal mortality.


1 To source waters

Nickel can enter the aquatic environment naturally from the weathering of rocks and minerals and run-off from soils. Nickel substitutes for iron in ferrous sulphides and also occurs in nickel-bearing laterites in ultramafic bedrock terranes. Industrial discharges can also contribute to the nickel in water. Nickel is used mainly in the production of stainless steel and other corrosion-resistant alloys, as a catalyst in industrial processes and in oil refining and in the manufacture of foods, baked goods, soft drinks, flavouring syrups and ice cream. Main releases to the environment are from the burning of fossil fuels and in waste discharges from electroplating industries.


No known sources.


Elevated nickel levels can arise in reticulated waters from the corrosion of nickel-plated fittings.

Forms and fate in the environment In aqueous solution nickel occurs mainly in the +2 oxidation state. It is present as relatively soluble salts in association with suspended solids and in combination with organic matter. Nickel co-precipitates with iron and manganese oxides and adsorbs to suspended organic matter.

Typical concentrations in drinking-water Nickel was not measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. The P2 Chemical Determinand Identification Programme, sampled from 900 zones, found nickel concentrations to range from �not detectable� (nd) to 0.70 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). In major Australian reticulated supplies, concentrations of nickel range up to 0.03 mg/L, with typical concentrations less than 0.01 mg/L.


Removal methods Nickel is rarely found in natural waters, partly because of the low solubility of some of its compounds. Nickel can be co-precipitated with iron and manganese oxides. Resins with chelating functional groups such as phosphoric acid and EDTA have a very high affinity for nickel. To avoid corrosion of nickel-plated fittings, the corrosiveness of the water should be minimised.

Analytical methods

Referee method



1 Inductively Coupled Plasma Method (APHA 3120B). 2 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Food is the dominant source of nickel exposure in the non-smoking, non-occupationally exposed population; water is generally a minor contributor to the total daily oral intake. Nickel is present in many foods. Highest concentrations occur in cocoa, soy beans and some cereals. It has been estimated that the average daily dietary intake is between 0.1 mg/day and 0.3 mg/day. In humans, absorption of soluble nickel from drinking-water may be 40 times higher than absorption of nickel from food. Nickel appears to be distributed to all organs, with primary accumulation in the kidneys, lungs and liver, with excretion occurring mainly through urine. Long-term exposure may result in toxic effects to the kidney. Nickel is able to pass through the human placenta. Several epidemiological studies have suggested a risk of nasal, sinus and lung cancer by inhalation of nickel. The International Agency for Research on Cancer concluded that inhaled nickel sulphate is carcinogenic to humans. There is no evidence that nickel is carcinogenic when ingested. Nickel is a common skin allergen and can cause dermatitis, particularly in adult women. Tests for mutagenicity with strains of bacteria have mostly been negative but gene mutations and chromosome aberrations have been reported in mammalian cells.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV. The relevant data base for deriving a no observable adverse effects level is limited. On the basis of a dietary study in rats in which altered organ to body weight ratios were observed, a NOAEL of 5 mg/kg body weight per day has been determined. The MAV for nickel in drinking-water was derived as follows:

5 mg/kg body weight per day x 70 kg x 0.1 = 0.0175 mg/L (rounded to 0.02 mg/L mg/L) 2 L per day x 1000


where:

• no observable adverse effect level = 5 mg/kg body weight per day from a dietary study in rats in which altered organ to body weight ratios were observed



• uncertainty factor = 1000 (100 for inter- and intraspecies variation and an extra factor of 10 to compensate for the lack of adequate studies on long-term exposure and reproductive effects, the lack of data on carcinogenicity by the oral route (although nickel, as both soluble and sparingly soluble compounds, is now considered as a human carcinogen in relation to pulmonary exposure), and a much higher intestinal absorption when taken on an empty stomach in drinking-water than when taken together with food

• average amount of water consumed by an adult = 2 L per day. This MAV should provide sufficient protection for individuals who are sensitive to nickel.







WHO. 2003. Nickel in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/55).


Nitrate and nitrite Updated July 2005.

Maximum acceptable value for nitrate (short-term) Based on health considerations, the concentration of nitrate (as NO3-) in drinking-water should not exceed 50 mg/L.

Maximum acceptable value for nitrite (short-term) Based on health considerations, the short-term concentration of nitrite (as NO2-) in drinking-water should not exceed 3 mg/L.

Maximum acceptable value for nitrite (long-term and provisional) Based on health considerations, the long-term concentration of nitrite (as NO2-) in drinking-water should not exceed 0.2 mg/L. The WHO guideline value for chronic effects of nitrite is considered provisional owing to uncertainty surrounding the relevance of the observed adverse health effects for humans and the susceptibility of humans compared with animals.


Maximum acceptable value for combined nitrate plus nitrite The sum of the ratios of the concentrations of each to its Maximum acceptable value (short-term) should not exceed 1. The maximum contaminant level for nitrate (USEPA 2004) is 10 mg/L as N, and 1 mg/L for nitrite as N, or a total of 10 mg/L.


1 To source waters

Nitrate and nitrite can enter the aquatic environment from the oxidation of vegetable and animal debris and animal excrement. Nitrate and nitrite can also enter water from agricultural, domestic and industrial discharges. Nitrate is used in chemical fertilisers, oxidising agents in the chemical industry, in the manufacture of glass, enamels for pottery, matches, pickling meat and in the production of explosives. A major source of nitrate is from municipal wastewaters and septic tanks. Nitrite is also used as a corrosion inhibitor in industry, and for curing meats.


The chlorination of raw waters containing significant amounts of ammonia or nitrite may lead to increases in nitrate through their oxidation. As 70% or more of the chlorine consumed during the oxidation of ammonia leads to N2 production, the increase in nitrate concentrations is likely to be small unless ammonia concentrations are high.


Chloramination may give rise to the formation of nitrite within the distribution system if the formation of chloramine is not sufficiently controlled. The formation of nitrite is as a consequence of microbial activity and may be intermittent. Nitrification in distribution systems can increase nitrite levels, usually by 0.2�1.5 mg/L.

Forms and fate in the environment Nitrate and nitrite are naturally occurring ions which make up part of the nitrogen cycle. Nitrate is the oxidised form of combined nitrogen found in natural waters and in dilute aqueous solutions is chemically unreactive. Under anaerobic conditions nitrate may be reduced to nitrite and ammonia. Nitrite is seldom present in surface waters at significant concentrations but may be present in ground waters. High nitrite levels are generally indicative of contamination. Incomplete nitrification of ammonia and denitrification of nitrate result in the biochemical production of nitrite which is generally present only under anaerobic conditions.

Typical concentrations in drinking-water

Nitrate

Nitrate was routinely measured in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Of 1908 samples analysed between 1983 and 1989, 14 samples (0.7%) contained concentrations equal to or exceeding the 1984 MAV of 10 mg/L (N). The P2 Chemical Determinand Identification Programme, sampled from 673 zones, found nitrate concentrations to range from �not detectable� (nd) to 30 mg/L as NO3-N, with the median concentration being 0.2 mg/L (limit of detection = 0.1 mg NO3-N/L).


Nitrite

Nitrite was not measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. The P2 Chemical Determinand Identification Programme, sampled from 227 zones, found nitrite concentrations to range from �not detectable� (nd) to 0.088 mg/L, with the median concentration being �nd� (limit of detection = 0.005 NO2-N mg/L).

Removal methods

Nitrate

Nitrate is not removed from water by classical methods of treatment. Ion exchange systems have been developed for removing nitrate, but dilution with water of lower nitrate concentration from another source is commonly used, where one is available.

Nitrite

Treatment of the water with an oxidising agent such as chlorine will convert the nitrite to nitrate. The nitrate can then be treated as explained for nitrate. The USEPA Maximum Concentration Level for nitrite indicates that the concentration at which it might be of concern is ten times less than the guideline for nitrate. The oxidation of high nitrite levels to nitrate therefore will not create an unacceptably high nitrate concentration in the water, unless the nitrate level is already high, or the nitrite level is extremely high.

Analytical methods

Nitrate

Referee method Cadmium Reduction Method (APHA-NO3-E). Some alternative methods 1 Ion Chromatography Method (APHA 4110). 2 Nitrate Electrode Method (APHA 4500-NO3 D).

Nitrite

Referee method Colorimetric Method (APHA 4500-NO2 B). Some alternative methods 1 Ion Chromatography Method (APHA 4110).

Health considerations Ingested nitrate is absorbed readily and completely from the upper small intestine. Nitrite may be absorbed directly from the stomach as well as from the small intestine. Sodium nitrite is used as a food preservative, especially in cured meats. The toxicity of nitrate in humans is thought to be due solely to its reduction to nitrite. The primary health concern regarding nitrate and nitrite is the formation of methaemoglobinaemia, so-called blue-baby syndrome. Nitrate is reduced to nitrite in the stomach of infants, and nitrite is able to oxidise haemoglobin (Hb) to methaemoglobin (metHb), which is unable to transport oxygen around the body. The reduced oxygen transport becomes clinically manifest when metHb concentrations reach 10% or more of normal Hb concentrations; the condition, called methaemoglobinaemia, causes cyanosis and, at


higher concentrations, asphyxia. The normal metHb level in infants under three months of age is less than 3%. Other susceptible groups include pregnant women and people with a deficiency of glucose-6-phosphate dehydrogenase or methaemoglobin reductase. Methaemoglobinaemia in infants also appears to be associated with simultaneous exposure to microbial contaminants (eg, Addison and Benjamin 2004). Nitrate is not mutagenic in bacteria and mammalian cells in vitro. Chromosomal aberrations were observed in the bone marrow of rats after oral nitrate uptake, but this could have been due to exogenous N-nitroso compound formation. Nitrite is mutagenic. The weight of evidence is strongly against there being an association between nitrite and nitrate exposure in humans and the risk of cancer.

Derivation of maximum acceptable values

Nitrate (short-term)

The MAV of 50 mg/L (as NO3) is to protect against methaemoglobinaemia in bottle-fed infants (short-term exposure). In epidemiological studies, methaemoglobinaemia was not reported in infants in areas where drinking-water consistently contained less than 50 mg of nitrate per litre. The epidemiological evidence for an association between dietary nitrate and cancer is insufficient, and the MAV for nitrate in drinking-water is established solely to prevent methaemoglobinaemia, which depends upon the conversion of nitrate to nitrite. Although bottle-fed babies are the most susceptible, occasional cases have been reported in some adult populations.

Nitrite (short-term)

The short-term MAV of 3 mg/L (as NO2) is to protect against methaemoglobinaemia in bottle-fed infants. Animal studies were inappropriate to establish a firm no observable adverse effect level for methaemoglobinaemia in rats. Therefore, a pragmatic approach was followed, accepting a relative potency for nitrite and nitrate with respect to methaemoglobin formation of 10:1 (on a molar basis), and a provisional MAV of of 3 mg/L has been adopted for nitrite.

Nitrite (long-term)

The 0.2 mg/L (as NO2) MAV for long-term exposure for chronic effects of nitrite is considered provisional owing to uncertainty surrounding the relevance of the observed adverse health effects for humans and the susceptibility of humans compared with animals. The occurrence of nitrite in the distribution system as a consequence of chloramine use will be intermittent, and average exposures over time should not exceed the provisional MAV. The nitrite MAV (long-term exposure) is based on allocation to drinking-water of 10% of JECFA ADI of 0.06 mg/kg of body weight per day, based on nitrite-induced morphological changes in the adrenals, heart and lungs in laboratory animal studies.

Nitrate : nitrite ratio

Because of the possibility of simultaneous occurrence of nitrite and nitrate in drinking-water, the sum of the ratio of the concentration of each to their short-term MAVs, as shown in the following formula, should not exceed 1: C(NO2) + C(NO3) ≤ 1 MAV(NO2) MAV(NO3) where C = concentration, and MAV = maximum acceptable value.

References Addison TM, Benjamin N. 2004. Nitrate and human health. Soil Use and Management 20: 98�104.


APHA. 1998. Standard Methods for the Examination of Water and Wastewater (20th ed). Washington: American Public Health Association, American Water Works Association, Water Environment Federation.






WHO. 2003. Nitrate and Nitrite in drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/56).


Potassium permanganate Updated July 2005.

Maximum acceptable value There are insufficient data to set a MAV for potassium permanganate.


1 To source waters

Potassium permanganate may enter source waters in discharges from industries in which it is used. It is used as an oxidising agent, in pharmaceutical manufacture (deodorant, disinfectant, detergent, gargling), as a bleaching agent, analytical reagent, dye and catalyst.


Potassium permanganate, as a disinfectant, should only be employed in emergencies, and then only on a small scale. However, it is more widely used as a preoxidant to help in the elimination of tastes and odours, or the oxidation of iron and manganese for their removal by physical processes.


No known sources.

Forms and fate in the environment Although it is a strong oxidizing agent at the pH values typical of environmental waters, the removal of potassium permanganate by oxidation-reduction reactions is likely to be relatively slow in comparision with the reactions of chlorine. As the result of its oxidation reactions, the permanganate ion is usually converted into highly insoluble, brown or brown-black manganese oxides.


Typical concentrations in drinking-water No information about the concentrations of potassium permanganate in drinking-water is available. The permanganate ion is highly unlikely to persist into the distribution system, unless as a result of a serious dosage error.

Removal methods Reaction with reducing reagents, such as sodium thiosulphate, sulphur dioxide, and sodium bisulphite, can be used to remove permanganate from water. However, because of the likely formation of insoluble products from these reactions, some means of physical removal of precipitated material will probably have to be used downstream of the reducing reagent addition.

Analytical methods

Referee method

A referee method cannot be selected for potassium permanganate because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for potassium permanganate for the above reason. However, the following information may be useful: Potassium permanganate cannot be preserved in a sample. Analysis should be carried out as soon after sampling as possible, and if a delay between sampling and analysis is unavoidable the sample should be kept chilled. No method for determining potassium permanganate appears in any of the standard texts for the analysis of water. However, by modification of APHA 3500-Mn D � Persulphate Method, satisfactory measurements probably could be made. This method is intended to allow the manganous ion concentration in the water to be measured by oxidising these ions to permanganate. If the oxidation step is omitted and the remainder of the method, using spectrophotometric absorption measurement at 525 nm, is used, permanganate concentrations as low as 0.05 mg/L as Mn, should be detectable.

Health considerations No information is available on the chronic health effects of ingestion of potassium permanganate. Symptoms of acute poisoning following ingestion of potassium permanganate include nausea, vomiting of a brownish coloured material, corrosion, oedema, liver and kidney damage and cardiovascular depression. The fatal dose is probably about 10 g and death may occur up to one month from the time of poisoning.

Derivation of maximum acceptable value There are insufficient data to set a MAV for potassium permanganate.


Reynolds JEF (ed). 1989. Martindale the Extra Pharmacopoeia (29th ed). London: The Pharmaceutical Press.


Selenium Updated July 2005.

Maximum acceptable value Based on health considerations, the concentration of selenium in drinking-water should not exceed 0.01 mg/L. The maximum contaminant level (USEPA 2004) is 0.05 mg/L.


1 To source waters

Selenium enters the aquatic environment naturally via the weathering of rocks and minerals, soil runoff and volcanic activity. Selenium occurs essentially in sulphide ores of heavy metals. Selenium is used in a wide variety of manufacturing industries including electronic, metallurgic, glass and ceramic, pigment, chemical and pharmaceutical. It may also be present in effluents from copper and lead refineries and in municipal sewage and is used as an animal remedy. Selenium is released from natural and human-made sources, with the main source being the burning of coal. Selenium is added as a trace element to some soils.


No known sources.


No known sources. Forms and Fate n the Environment Selenium exists in four oxidation states in the aquatic environment: 0, -2, +4 and +6. Dissolved selenium species are mainly in the +4 and +6 oxidation states while those in the -2 state decompose under aerobic conditions to form elemental selenium which is insoluble. Selenium (IV) is also reduced to insoluble elemental selenium under acidic and reducing conditions, whereas alkaline and oxidising conditions favour the formation of the stable Se(VI) compounds.

Typical concentrations in drinking-water Selenium was measured routinely in New Zealand drinking-water supplies as part of the Department of Health three yearly surveillance programme. Of 1779 samples analysed between 1983 and 1989, 21 (1.2%) had concentrations equal to or exceeding the MAV of 0.01 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 285 zones, found selenium concentrations ranged from �not detectable� (nd) to 0.009 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). In major Australian reticulated water supplies, selenium concentrations are less than 0.005 mg/L. Concentrations in drinking-water supplies overseas are generally below 0.01 mg/L but groundwater concentrations as high as 6 mg/L have been reported in the United States.


Removal methods The efficiency with which selenium can be removed from waters by conventional coagulation depends on the oxidation state of the selenium (IV and VI). Virtually no removal of selenium (VI) is achieved using either ferric salts or alum. Some studies have shown significant removal of selenium (IV) with ferric sulphate at pH values in the 6 to 7 region. However, contradictory data were obtained when the effect of coagulant dose was studied. Adsorption or ion-exchange on alumina columns has been shown to be effective in removing both these oxidation states, but removal efficiency depends on the concentrations of other anions that are preferentially taken up by alumina. Reverse osmosis can remove both oxidation states of selenium. Lime-softening and PAC or GAC have little effect on the concentration of either form of selenium.

Analytical methods

Referee method



1 Hydride Generation / Atomic Absorption Spectrometric Method (APHA 3114B). 2 Inductively Coupled Plasma (ICP) Method (APHA 3120). 3 Inductively Coupled Plasma � Mass Spectrometry (EPA Method 200.8).

Health considerations Selenium is an essential element for humans, and forms an integral part of the enzyme glutathione peroxidase and probably other proteins as well. Foodstuffs such as cereals, meat and fish are the principal source of selenium in the general population. Levels in food also vary greatly according to geographical area of production. Most selenium compounds are water soluble and are efficiently absorbed from the intestine. Selenium compounds appear to be of the same order of toxicity in humans and laboratory animals. There have been a number of reports of adverse effects caused by short- and long-term exposure to selenium, most of which have resulted from occupational exposure or accidental poisoning. The occurrence of acute or chronic nutritional toxicity is comparatively rare. Intakes above about 1 mg/day over prolonged periods may produce nail deformities characteristic of selenosis. Other characteristics of excess selenium intake include non-specific symptoms such as gastro-intestinal disturbances, dermatitis, dizziness, lassitude and a garlic odour to the breath. Long term toxicity in rats is characterised by depression of growth and liver pathology at levels of 0.03�0.4 mg selenium/kg body weight per day given in food. Tests for mutagenic activity using bacteria have reported both positive and negative results. Studies indicate that selenite can cause chromosome damage to mammalian cells. Except for selenium sulphide, which does not occur in drinking-water, experimental data do not indicate that selenium is carcinogenic. The International Agency for Research on Cancer has placed selenium and selenium compounds in Group 3 (not classifiable as to its carcinogenicity to humans).


Derivation of maximum acceptable value In humans, the toxic effects of long-term selenium exposure are manifested in nails, hair and liver. Based on Chinese data, clinical signs seem to occur at a daily intake above 0.8 mg. Daily intakes by Venezuelan children with clinical signs are estimated at about 0.7 mg, on the basis of their blood levels and Chinese data on blood level/intake relationships. Effects on synthesis of a liver protein were also seen in a group of patients with rheumatoid arthritis given 0.25 mg selenium/day. No clinical or biochemical signs of selenium toxicity were reported in a group of 142 persons with a mean daily intake of 0.24 mg (maximum 0.72 mg). Based on these data, a No-observable-adverse-effects level (NOAEL) in humans is estimated to be about 4 µg/kg body weight per day. The recommended daily intake of selenium is 1 µg/kg of body weight for adults. The MAV for selenium in drinking-water was derived as follows:


where: • no observable adverse effect level = 0.004 mg/kg body weight per day • average weight of an adult = 70 kg • the proportion of tolerable daily intake assigned to the consumption of water = 0.1 • average amount of water consumed by an adult = 2 L per day.








WHO. 2003. Selenium in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/13).



Silver Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of silver in drinking water should not exceed 0.1 mg/L. This provisional MAV is based on the guideline value in the Australian Drinking Water Guidelines. Note that WHO (2004) considered that there are no adequate data with which to derive a health-based guideline value for silver in drinking-water, and that where silver salts are used to maintain the bacteriological quality of drinking-water, levels of silver up to 0.1 mg/L can be tolerated without risk to health. The USEPA has a secondary drinking water regulation of 0.1 mg/L for silver.


1 To source waters

Silver enters the aquatic environment naturally via the weathering of rocks and minerals and soil runoff. In New Zealand, silver occurs almost always in association with gold. Typical geological settings include: associated with intermediate volcanics (eg, Hauraki Goldfield), associated with metamorphic rocks (eg, with most South Island gold-bearing reefs), in base-metal mineralisation near granite contacts (eg, at Richmond Hill, Copperstain Creek and Mt Radiant in Nelson), in some South Island ultrabasics around Nelson and in silver mineralisation at Puhipuhi, Northland. Silver also may be present in water from industrial discharges. Silver is used in photographic materials, in the manufacture of sterling and plated ware, jewellery, coins, electrical and electronic products (eg, batteries, contacts and conductors, brazing alloys and solders, catalysts, mirrors, fungicides, and in dental and medical supplies). Silver concentrations in natural source waters are generally very low, less than 0.0002 mg/L. Measurements of silver in Canadian surface waters (measured at NAQUADAT stations) indicated that the silver concentrations in 90% of the water samples were below detection limits (which ranged from 0.004 to 0.01 mg/L), and none was greater than 0.01 mg/L. Measurements at seven points along the length of the Ottawa River, Ontario, found concentrations of silver that ranged between 0.00001 and 0.00006 mg/L.


Soluble silver compounds may be used to disinfect drinking-water. At present, this method of disinfection is not known to be employed in any New Zealand drinking-water supplies.


Some point-of-use water filters contain activated carbon impregnated with silver.

Forms and fate in the environment In aqueous systems silver is predominantly present in the univalent state Ag(I). Under aerobic conditions Ag(I) is soluble and mobile but as the pH increases the silver tends to precipitate. Sorption and precipitation are the dominant mechanisms for controlling the transport and removal of silver in the aquatic environment.

Typical concentrations in drinking-water Silver was not routinely monitored in New Zealand drinking-water supplies and therefore no information is available about its concentration.


The P2 Chemical Determinand Identification Programme, sampled from 831 zones, found silver concentrations to range from �not detectable� (nd) to 0.0032 mg/L, with the median concentration being �nd� (limit of detection = 0.0005 mg/L). In a survey of Canadian tap water, silver was found in only 0.1 percent of 239 sampled waters; the detection limit for silver in this study, which used neutron activation analysis, was 0.000001 to 0.000005 mg/L. Silver concentrations in excess of 0.05 mg/L may occur when silver is used as an antimicrobial agent in water treatment devices (Smith et al 1977). Silver is used to treat potable water for Swiss ski resorts, German breweries, soft drink bottlers, British ships, Shell Oil Company tankers, drilling rigs, and over half the world�s airlines (USEPA 1980). Concentrations of 0.1 to 0.2 mg/L of silver are sufficient to ensure antimicrobial action.

Removal methods Natural turbidity influences the effectiveness of coagulation/flocculation with iron or aluminium salts to remove silver. Ferric sulphate is more effective than alum, and removal with alum is pH-dependent. Lime-softening is useful for the removal of silver when operating at a pH greater than 9. Silver has been removed effectively (86�92%) from secondary and industrial waste waters by cation exchange.

Analytical methods

Referee method

A referee method cannot be selected for silver because a MAV has not been established and therefore the sensitivity required for the referee method is not known. But we do now!


No alternative methods can be recommended for silver for the above reason. However, the following methods are used to analyse for silver: 1 Atomic Absorption Spectrometric Method (APHA 3111B and APHA 3113B).

The detection limits for silver by Flame AA and Furnace AA are 0.01 mg/L and 0.0002 mg/L (0.2 µg/L) respectively. 2 Inductively Coupled Plasma Method (APHA 3120B).

The upper limit for silver by ICP is 50 mg/L with an estimated detection limit of 0.007 mg/L (7 µg/L).

Health considerations Traces of silver can be found in most foods but it is not considered an essential trace element for mammals. Only a small percentage of silver is absorbed by the gastro-intestinal tract. Retention rates in humans and laboratory animals range between 0 and 10%. It is stored mainly in the liver and skin and is capable of binding to amino acids and proteins. The only obvious sign of silver intoxication is argyria, a condition is which skin and hair are heavily discoloured by silver in the tissues. An oral no observed adverse effect level for argyria in humans for a total lifetime intake of 10 g silver can be estimated from human case reports and long-term animal experiments.


No data are available on the carcinogenicity of silver. Silver salts are not mutagenic in tests with bacteria but can induce damage in mammalian DNA.

Derivation of maximum acceptable value The low levels of silver in drinking-water, generally below 0.005 mg/L (5 µg/L), are not relevant to human health with respect to argyria. On the other hand, special situations exist where silver salts may be used to maintain the bacterial quality of drinking-water. Higher levels of silver of up to 0.1 mg/L (this concentration gives a total dose over 70 years of half the human NOAEL of 10g) could be tolerated in such cases without risk to health. The guideline value for silver in drinking water was derived as follows:

0.4 mg per day x 0.5 = 0.1 mg/L 2 L per day

where: • 0.4 mg/day is derived from a lifetime no effect level of 10 g (Hill and Pillsbury 1939) • 0.5 is the proportion of total daily intake attributable to the consumption of drinking water • 2 L/day is the average amount of water consumed by an adult. No additional safety factors were used, as the calculation was based on a human no effect level. It is unlikely that silver concentrations in drinking water would ever reach a concentration that could cause adverse effects. Silver or silver salts should not be used as antimicrobial agents unless no other disinfectants are available.





Hill WR, Pillsbury DM. 1939. Argyria, the Pharmacology of Silver. Baltimore, USA: The Williams & Wilkins Co.


Smith DK, Thomas GH, Christison J, et al. 1977. Survey and Test Protocols for Point-of-Use Water Purifiers. Publication 77-EHD-8. Ottawa: Department of National Health and Welfare.

USEPA. 1980. Ambient Water Quality Criteria: Silver. NTIS Document No. PB81-117822. US Environmental Protection Agency, Environmental Criteria and Assessment Office.


WHO. 2003. Silver in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/14).


Tin Updated July 2005.


Maximum acceptable value Based on health considerations, the concentration of tin in drinking water should not exceed 1 mg/L. No health based MAV is proposed for tin in drinking-water because it is not hazardous to human health at concentrations normally found in drinking-water.


1 To source waters

Tin may enter the aquatic environment via the weathering of igneous rocks that contain the mineral cassiterite or tinstone. The major anthropogenic sources of inorganic tin to the environment are discharges from mining, refineries, food processing and packaging plants, steel manufacture and construction. Municipal sewage, containing tin from discarded household products, may discharge tin into the environment.


No known sources.


Tin may enter drinking-water through the dissolution of antimony-tin solder. Some water tanks are tinned.

Form and fate in the environment Tin can exist in several oxidation states, but in the aquatic environment will usually exist in the +4 oxidation state. Divalent tin may occur in anaerobic sediments and waters of low redox potential. Divalent tin compounds are soluble in water, but tin is usually oxidised to the less soluble +4 state. The most likely forms of +4 tin in natural water include the hydroxides and hydroxyoxides. In general, tin concentrations increase as follows: water< surface microlayer < sediments. Tin may be methylated in aqueous in the presence of some micro-organisms.

Typical concentrations in drinking water Tin was not routinely monitored in New Zealand drinking-water supplies and therefore no information is available about its concentration. The P2 Chemical Determinand Identification Programme, sampled from 897 zones, found tin concentrations to range from �not detectable� (nd) to 0.048 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L).

Removal methods Treatment of drinking-water to reduce the concentration of inorganic tin is unlikely to be required.

Analytical methods

Referee method

A referee method cannot be selected for tin because a MAV has not been established and therefore the sensitivity required for the referee method is not known.



No alternative methods can be recommended for tin for the above reason. However, the following methods are used to analyse for tin:

1 Flame Atomic Absorption Spectrometric Method (APHA 3111B) and APHA 3113B). The detection limit for tin by flame AA is 0.8 mg/L.

2 Furnace Atomic Absorption Spectrometric Method (APHA 3113B). The detection limit for tin by furnace AA is 0.005 mg/L.

Health considerations Tin and inorganic tin compounds are poorly absorbed from the gastrointestinal tract, do not accumulate in tissues, and are rapidly excreted, primarily in the faeces.

Interpretation of maximum acceptable value Because of the low toxicity of inorganic tin, a tentative guideline value could be derived three orders of magnitude higher than the normal tin concentration in drinking-water. Therefore, the presence of tin in drinking-water does not represent a hazard to human health. For this reason, the establishment of a numerical guideline value for inorganic tin is not deemed necessary.



World Health Organization. 1996. Guidelines for Drinking-water Quality (2nd ed). Volume 2: Health criteria and other supporting information.


Trichloramine Updated July 2005. There are insufficient data to establish a MAV for trichloramine.


1 To source waters

Chloramines may be present in source waters as a result of their discharge from industries in which they are used. Their principal use is as intermediates in the manufacture of hydrazine.


Trichloramine can be formed in chlorinated water that contains ammonia and some organic nitrogen compounds. The concentration depends upon the pH and chlorine to nitrogen ratio. Ammonia may be intentionally added to the water to produce the chloramines as disinfectants.


It is possible that reactions of chlorine with nitrogenous material in the distribution system may produce chloramines. Trichloramine is more likely to form at low pH.


Typical concentrations in drinking-water No typical value data are available for New Zealand. Chloramination is not used intentionally at present as a disinfectant in New Zealand, and the concentrations of inorganic chloramines present in waters depends upon the concentrations of inorganic and some organic nitrogen compounds present in the raw water, and control of the chlorination process.

Removal methods Chemical reducing agents, including sodium thiosulphate, sulphur dioxide, and sodium bisulphite can be used to remove trichloramine. Activated carbon adsorbs trichloramine. Trichloramine in water will not be present intentionally because of its irritant properties. Although mono- and dichloramine are volatile to some extent, trichloramine is very much more so and may be readily stripped from water by aeration. As dichloramine also has some undesirable properties, any loss of it during the process may also be welcomed. Loss of trichloramine can be expected from waters exposed to sunlight.

Analytical methods

Referee method

DPD Ferrous Titrimetric Method (APHA 4500-Cl F). The limit of detection for this method is approximately 0.2 mg/L for field use, although lower levels can be determined under laboratory conditions and with care. Analytical texts indicate that by manipulation of the conditions of the analysis measurement of monochloramine, dichloramine and trichloramine can be made. These methods are of use when ammonia only is in the water being chlorinated. In most natural waters nitrogen-containing organic compounds are also present. Organic chloramines are formed from these compounds when chlorine reacts with them. Organic chloramines also produce colour during the DPD test and make attempts to differentiate between the different inorganic chloramines of little value. Unless investigating taste and odour problems, it is recommended that only the total combined chlorine, ie, total chloramine concentration, is reported. These methods measure trichloramine in terms of mg Cl as Cl2/L.



While more accurate than the DPD methods, expensive equipment and a high degree of skill and care are required for this method. The limit of detection is better than 0.1 mg/L. The APHA method describes variations that will allow the determination of mono- and di-chloramine. Interferences due to organic chloramines may also cause interferences with these methods. 2 DPD Colorimetric Method (APHA 4500-Cl G)

This method requires a spectrophotometer for the colorimetric measurements, although hand-held comparators do offer a cheaper, though less reliable variation for field use. The limit of detection (LOD), with instrumental assistance, is approximately 0.1 mg/L. The LOD for the comparator depends on the colour disc in use. Chloramine concentrations as low as 0.2 mg/L approximately should be detectable, but the accuracy of the method depends upon use of the correct lighting (natural lighting should be used with the sun behind the viewer), the individual�s ability


to match colours and judge their intensity, and ensuing that readings are taken as soon after colour development as possible. The LOD may be about 0.10 mg/L when using a Nessleriser. The same comment on the usefulness of trying to determine the individual chloramine concentrations made for the referee method, also applies to this method.

Health considerations Studies have revealed equivocal evidence of carcinogenic activity of chloraminated drinking-water in female rats, as indicated by an increase in incidence of mononuclear cell leukaemia. Epidemiological studies did not report an association between ingestion of chloraminated drinking-water and increased urinary bladder mortality rates in humans. When tap-water containing chloramines was used for dialysis, acute haemolytic anaemia, characterised by denaturation of haemoglobin and lysis of red blood cells, was reported in haemodialysis patients. Trichloramine is an eye irritant, and will also irritate breathing passages. It is believed to be responsible for breathing difficulties experienced by asthmatics in some swimming pools.


Australian Drinking Water Guidelines, NHMRC and AWRC, 1993 Draft, and the 2004 edition on http://www.nhmrc.gov.au/publications/synopses/eh19syn.htm




Uranium Updated July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of uranium in drinking water should not exceed 0.02 mg/L. The guideline value is designated as provisional because of outstanding uncertainties regarding the toxicology and epidemiology of uranium as well as difficulties concerning its technical achievability in smaller supplies. A guideline value of up to 0.03 mg/L may be protective of kidney toxicity because of uncertainty regarding the clinical significance of changes observed in epidemiological studies. The MAV derived from chemical toxicity data is also protective of radiological effects. The maximum contaminant level (USEPA 2004) is 0.03 mg/L.



1 To source waters

Uranium can enter water naturally from the weathering of rocks, minerals and phosphate fertilisers. Geologically, uranium is a very mobile element. In New Zealand, uranium occurs in the non-marine Ohika Beds and Hawks Crag Breccia adjacent to the Paparoa Range of southwest Nelson. Uranium is also present in several granites and diorites on the West Coast. Uranium may also be discharged to water in wastes from industries in which it is used. Uranium is used in radio emissions shielding material, photographic emulsions, porcelains used in dentistry and optical lenses. In addition, alloys may contain uranium, and certain uranium compounds are used as catalysts in the chemical industry. Uranium is used as a radioactive tracer. Uranium is present in.the environment as a result of leaching from natural deposits and release from mill or mine tailings.


No known sources.


No known sources.

Forms and fate in the environment In oxidising environments the uranyl ion U(VI) is predominant whereas in reducing environments it is the uranous ion U(IV). Under conditions typical of natural waters, adsorption to hydrous ferric oxide in aerobic waters plays an important role in removing uranium from the aquatic environment.

Typical concentrations in drinking-water Uranium was not routinely monitored in New Zealand drinking-water supplies as part of the Department of Health three-yearly survey. The P2 Chemical Determinand Identification Programme, sampled from 828 zones, did not find uranium at detectable concentrations (limit of detection = 0.0005 mg/L). WHO (2004) states that levels in drinking-water are generally less than 0.001 mg/L; concentrations as high as 0.7 mg/L have been measured in private supplies in Canada.

Removal methods Anion exchange, lime softening and reverse osmosis have been recommended for the removal of uranium from drinking-water.

Analytical methods

Referee method

A referee method cannot be selected for uranium because a MAV has not been established and therefore the sensitivity required for the referee method is not known. But it has now.


No alternative methods can be recommended for uranium for the above reason. However, the following methods are used to analyse for uranium:


1 Fluorimetry (ASTM D2907-91). These test methods are applicable to the determination of uranium in water in the concentration range 0.005 to 50 mg/L. If interfering ions are present (small quantities of cadmium, chromium, cobalt, copper, iron, manganese, nickel, lead, platinium, silicon, thorium and zinc), an extraction method must be used. The direct method has a concentration range 0.005 to 2 mg/L, and the extraction method has a concentration range 0.04 to 50 mg/L. A fluorimeter is required having an excitation wavelength from 320 to 370 nm and measuring emission at 530 to 570 nm.

2 Colorimetric methods are available, but tend to be prone to interferences.

3 Inductively coupled plasma � mass spectrometry can be used for determining lower concentrations.


Chemical toxic effects

Absorption of dietary uranium by the gastro-intestinal tract has been estimated to be less than 1%. Highest uranium concentrations occur in the kidney and bone, with little in the liver. The overall biological half-life of uranium has been estimated at 6�12 months. In humans, the main toxic effect of short-term exposure to high concentrations of uranium is inflammation of the kidney. Little information is available on the effects of long-term exposure. One study, where 234 people drank contaminated water from wells with uranium concentrations up to 0.7 mg/L, reported no increase in the incidence of kidney disease or any other symptomatic complaint. In a number of studies carried out in rats, rabbits and dogs, most report that uranium has an effect on the kidney, although there are significant differences between species. No data are available on mutagenic effects in relation to uranium.

Radiological effects

Studies have shown high specific activity uranium isotopes to be carcinogenic in animals, causing malignant tumours in mice and bone sarcomas in rats. Similar studies using natural uranium (U-238) have not shown similar effects, possibly due to the lower radiation doses involved. Epidemiological data are inadequate to show whether exposure to uranium in drinking-water will lead to an increased risk of cancer.

Derivation of maximum acceptable value A TDI of 0.06 mg/kg of body weight per day was used, based on the application of an uncertainty factor of 100 (for inter- and intraspecies variation) to a LOAEL (equivalent to 0.06 mg of uranium per kg of body weight per day) for degenerative lesions in the proximal convoluted tubule of the kidney in male rats in a 91-day study in which uranyl nitrate hexahydrate was administered in drinking-water. It was considered unnecessary to apply an additional uncertainty factor for the use of a LOAEL instead of a NOAEL and the short length of the study because of the minimal degree of severity of the lesions and the short half-life of uranium in the kidney, with no indication that the severity of the renal lesions will be exacerbated following continued exposure. This is supported by data from epidemiological studies. The MAV for uranium in drinking-water was derived as follows:

0.06 mg/kg body weight per day x 70 kg x 0.8 = 0.0168 mg/L (rounded to 0.02 mg/L) 100 x 2 L per day

where: • lowest observable adverse effect level = 0.06 mg/kg body weight per day • average weight of an adult = 70 kg


• the proportion of tolerable daily intake assigned to the consumption of water = 0.8 • average amount of water consumed by an adult = 2 L per day.


ASTM D2907-91 (updated in 1997, but method withdrawn as at 2003).

Australian Drinking Water Guidelines, NHMRC and AWRC, 1993 Draft, and the 2004 edition on http://www.nhmrc.gov.au/publications/synopses/eh19syn.htm





WHO. 2003. Uranium in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/118).



2.2 Organic chemicals (except aesthetic determinands and pesticides)

acrylamide benzene benzo[a]pyrene bromochloroacetic acid bromochloroacetonitrile bromodichloromethane bromoform carbon tetrachloride chloroacetones chloroform 2-chlorophenol chloropicrin dialkyltins di(2-ethylhexyl)adipate di(2-ethylhexyl)phthalate dibromoacetic acid dibromoacetonitrile dibromochloromethane 1,2-dibromoethane dichloroacetic acid dichloroacetonitrile 1,2-dichlorobenzene 1,3-dichlorobenzene 1,4-dichlorobenzene 1,2-dichloroethane 1,1-dichloroethane 1,1-dichloroethene 1,2-dichloroethene (cis and trans) dichloromethane dioxins EDTA epichlorohydrin ethylbenzene fluoranthene formaldehyde hexachlorobutadiene monobromoacetic acid monochloroacetic acid monochlorobenzene MX nitrilotriacetic acid PCBs polyaromatic hydrocarbons (PAH): see benzo[a]pyrene and fluoranthene styrene tetrachloroethene toluene tributyltin oxide trichloroacetaldehyde/chloral hydrate trichloroacetic acid trichloroacetonitrile trichlorobenzenes (total)


1,1,1-trichloroethane trichloroethene 2,4,6-trichlorophenol trihalomethanes (THMs � see individuals) vinyl chloride xylenes

Acrylamide Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of acrylamide in drinking-water should not exceed 0.0006 mg/L (0.6 µg/L).


1 To source waters

Acrylamide may enter the aquatic environment as an industrial contaminant. It is used as a chemical intermediate or as a monomer in the production of polyacrylamide.


The most important source of drinking-water contamination by acrylamide is the use of polyacrylamide flocculants which are used for potable water treatment in New Zealand. Acrylamide occurs as a minor impurity in polyacrylamide polymers. A typical dose level of 1 mg/L of nonionic or anionic polyacrylamide results in an estimated maximum acrylamide concentration of 0.0005 mg/L (0.5 µg/L), with practical concentrations 2�3 times lower. Residual levels of acrylamide from the use of cationic polyacrylamides may be higher.


Acrylamide may enter drinking-water through the use of polyacrylamides as grouting agents in the construction of drinking-water reservoirs and wells.

Forms and fate in the environment Acrylamide is highly mobile in aqueous environments and readily leachable in soil. Its behaviour in subsurface soil, where most grout application takes place, has not been studied. Acrylamide is susceptible to biodegradation in soil and surface water. Non-biological hydrolysis may be important in natural water. Volatilisation is not an important removal process. Bioconcentration is not expected to be significant.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 114 zones, found acrylamide concentrations to range from �not detectable� (nd) to 0.00012 mg/L, with the median concentration being �nd� (limit of detection = 0.0001 mg/L).

Removal methods The most important source of drinking-water contamination by acrylamide is the use of polyacrylamide flocculants that contain residual levels of acrylamide monomer. Concentrations in drinking-water can be controlled by product specification.


Acrylamide can be removed by granular activated carbon, but is not removed effectively by powdered activated carbon. If the granular activated carbon is put at the end of the treatment process, acrylamide originating either in the source water or through the use of polyacrylamide flocculants for water treatment will be removed. Little is known about the rate at which acrylamide leaches from grouting agents in the reticulation.

Analytical methods

Referee method

HPLC/UVD. Determination of acrylamide monomer in waters and polymers, 1987, HMSO. 1988.


No alternative methods have been recommended for acrylamide because no methods meet the required criteria.

Health considerations Acrylamide is absorbed readily through the skin and following ingestion or inhalation. Absorbed acrylamide is distributed in body fluids and crosses the placental barrier. Acrylamide and its metabolites are accumulated in both nervous system tissues and blood as well as in liver, kidney and the male reproductive system. It is largely excreted as metabolites in urine and bile. Acrylamide is well-established as a cumulative neurotoxin. Humans exposed for a short time to groundwater contaminated with up to 400 mg/L of acrylamide showed effects including confusion, disorientation, memory disturbances and hallucinations. They recovered fully within four months. Many other cases of human exposure to acrylamide have been reported, generally dealing with dermal or inhalation exposure of workers in grouting operations or factories manufacturing acrylamide-based flocculants. Typical clinical symptoms were skin irritation, generalized fatigue, foot weakness and sensory changes, which reflect dysfunction of either the central or peripheral nervous system. In mutagenicity assays, acrylamide does not cause mutations in bacterial test systems, but does cause chromosome damage to mammalian cells in vitro and in vivo. In a long-term carcinogenicity study in rats exposed via drinking-water, acrylamide induced scrotal mesotheliomas. The International Agency for Research on Cancer has placed acrylamide in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value Based on the available information, it has been concluded that acrylamide is a genotoxic carcinogen. Therefore the risk evaluation was carried out using a non-threshold approach. On the basis of combined mammary, thyroid, and uterine tumours observed in female rats in a drinking-water study, and using the linearised multistage model, the MAV associated with an excess lifetime cancer risk of one per one hundred thousand (10-5) is estimated to be 0.0005 mg/L (0.5 µg/L).



Brown L, Rhead MM. 1979. Liquid chromatographic determination of acrylamide monomer in natural and polluted aqueous environments. Analyst 104: 391�9.

Department of Health, New Zealand. 1993. Health Risks of Synthetic Polymers and Monomers, and Methods of Monitoring and Control: A report prepared for the Department.

Johnson KA, Gorzinski SJ, Bodner KM, et al. 1986. Chronic toxicity and oncogenicity study on acrylamide incorporated in the drinking water of Fischer 344 rats. Toxicology and Applied Pharmacology 85: 154�68.

HMSO. 1988. Determination of Acrylamide Monomer in Waters and Polymers, 1987: Methods for the examination of waters and associated materials.

NZWWA. 1999. Standard for the Supply of Polyacrylamides for Use in Drinking-water Treatment. ISBN 1-877134-24-4. New Zealand Water and Wastes Association.

WHO. 2003. Acrylamide in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/71).


Benzene Revised July 2005

Maximum acceptable value Based on health considerations, the concentration of benzene in drinking-water should not exceed 0.01 mg/L. The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


1 To source waters

Benzene may be present as a natural component of petrol and motor vehicle emissions. Petrol spillages constitute the main source of benzene in the environment. In New Zealand petrol a limit of 4.2% (by volume) is allowable. Benzene may also enter source waters as an industrial contaminant since it is used by the chemical industry in the production of styrene, phenol and cyclohexane. Because of health risks, use as a solvent should decrease. There is however, a steady demand in New Zealand which will likely result in the continuing occurrence of benzene in the environment. In overseas studies benzene has been detected in the Rhine in Germany at approximately 0.0003 mg/L, and occasionally in groundwater supplies in the United States. Concentrations are usually less than 0.001 mg/L, but concentrations up to 0.18 mg/L have been detected in chemical plant effluent.


No known sources.


No known sources.

Forms and fate in the environment Volatilisation with subsequent oxidation is considered to be the primary fate of benzene entering the environment. In soil, benzene degrades under aerobic conditions only. In surface water, benzene rapidly volatilises to the air, biodegrades, or reacts with hydroxyl radicals.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 301 zones, found benzene concentrations to range from �not detectable� (nd) to 0.0036 mg/L, with the median concentration being �nd� (limit of detection = 0.0001 mg/L).

Removal methods Removal of benzene from contaminated source waters can be achieved through adsorption on to granular activated carbon and by air stripping.

Analytical methods

Referee method

Purge and Trap Capillary Column Gas Chromatographic/Mass Spectrometric Method (APHA 6210D, EPA 524.2).


1 Purge and Trap Capillary-Column Gas Chromatographic Method (APHA 6230D, EPA 502.2).

Health considerations Following ingestion, almost all of the benzene is absorbed from the gastro-intestinal tract. Benzene is also rapidly and efficiently absorbed following inhalation. Less than 1% is absorbed through the skin. Following absorption, benzene is widely distributed throughout the body, independent of the route of administration. It is metabolised predominantly into phenol by the liver, and also by bone marrow. Benzene has a low acute toxicity. The principal toxic effect following repeated exposure to low levels of benzene is in the blood and blood-forming tissues. Acute exposure of humans to high concentrations of benzene primarily affects the central nervous system. In fatal cases, extensive haemorrhages have been observed. There is considerable evidence that exposure to high benzene concentrations may eventually result in leukaemia. Although benzene does not induce mutations or DNA damage in standard bacterial assay systems, it has caused chromosomal aberrations in a variety of species in vivo. Benzene is carcinogenic in mice and rats after inhalation and oral exposure, producing malignant tumours at many sites. It is considered as a human carcinogen and is classified by the International Agency for Research on Cancer in Group 1 (carcinogenic to humans).

Derivation of maximum acceptable value Owing to the unequivocal evidence of the carcinogenicity of benzene to humans and animals and its documented chromosomal effects, quantitative risk assessment was used to calculate lifetime cancer risks. Based on a risk estimate using data on leukaemia from epidemiological studies involving inhalation exposure, it was calculated that a drinking-water concentration of 0.01 mg/L was associated with a lifetime excess cancer risk of one per one hundred thousand (10-5). As data on the carcinogenic risk to humans following ingestion of benzene are not available, risk estimates were also carried out on the basis of a two-year gavage study in rats and mice. The robust linear extrapolation model was used, as there was a statistical lack of fit of some of the data with the linearised multistage model. The estimated range of concentration in drinking-water corresponding to a excess lifetime cancer risk of 10-5, based on leukaemia and lymphomas in female mice and oral cavity squamous cell carcinomas in male rats, is 0.01-0.08 mg/L. The lower end of this estimate corresponds to


the estimate derived from epidemiological data, which formed the basis for the previous MAV of 0.01 mg/L associated with a 10-5 lifetime cancer risk. The MAV corresponding to an excess cancer risk of 10-5 is therefore 0.01 mg/L.




USEPA. 1988. Methods for the Determination of Organic Compounds in Drinking Water.

USEPA Draft Method 551. 1990. Determination of Chlorination Disinfection By-products and Chlorinated Solvents in Drinking Water by Liquid�Liquid Extraction and Gas Chromatography with Electron Capture Detection. Cincinnati, US: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).


WHO. 2003. Benzene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/24).


Benzo[a]pyrene New entry July 2005. (Also called benzo 3,4-pyrene.) (See also polynuclear aromatic hydrocarbons.)

Maximum acceptable value Based on health considerations, the concentration of benzo[a]pyrene in drinking-water should not exceed 0.0007 mg/L (0.7 µg/L). The maximum contaminant level (USEPA 2004) is 0.0002 mg/L.


1 To source waters

Polynuclear aromatic hydrocarbons are a large group of organic compounds formed from the incomplete combustion of organic matter, of which benzo[a]pyrene (BaP) has received the most extensive toxicological study. They have no industrial use but are formed naturally in forest fires, volcanic activity, or from anthropogenic activities such as domestic fires, vehicle emissions, coke ovens, the coal gas industry and related contaminated soils, and aluminium smelters. The principle route of entry to source water is via atmospheric deposition.


No known sources.



Treated water may be contaminated by leaching from coal-tar linings of water distribution systems. The presence of significant concentrations of BaP in drinking-water in the absence of very high concentrations of fluoranthene indicates the presence of coal-tar particles, which may arise from seriously deteriorating coal-tar pipe linings. It is recommended that the use of coal-tar-based and similar materials for pipe linings and coatings on storage tanks be discontinued.

Forms and fate in the environment Polynuclear aromatic hydrocarbons enter the environment through atmospheric deposition. Because of its low water solubility benzo[a]pyrene, adsorbs to sediments and suspended solids in aquatic systems. It does not hydrolyse. In some natural waters there are organisms capable of metabolising it, but in most waters it does not undergo biodegradation. Volatilisation may be important over periods exceeding one month, and if near the surface of waters, it is likely to undergo significant photodegradation. Losses through volatilisation, photodegradation and biodegradation are retarded by adsorption on to solids.

Typical concentrations in drinking-water The review of organic contaminants in New Zealand drinking-water supplies between 1987�1992 contained polynuclear aromatic hydrocarbons results from 217 samples, representing 204 supplies. Ten samples contained detectable concentrations of benzo[a]pyrene. Detected concentrations ranged from 0.00000004�0.00000086 mg/L (0.04�0.086 ng/L). The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find any benzo[a]pyrene concentrations (limit of detection = 0.0001 mg/L).

Removal methods Benzo[a]pyrene, like the polynuclear aromtaic hydrocarbons in general, is very insoluble in water and hence adsorbs readily to available surfaces. As a result, conventional coagulation/flocculation is able to achieve high levels of removal by removing particles to which it may be adsorbed, and by providing floc surfaces on to which benzo[a]pyrene in the bulk water may adsorb. Good removal can also be achieved by granular activated carbon. Chlorination can reduce benzo[a]pyrene concentrations by approximately 60% after three hour contact time. Oxidation by ozone results in a much more rapid reduction in benzo[a]pyrene concentration; 100% destruction is achieved in less than 15 minutes. As polynuclear aromatic hydrocarbons can be leached from coal-tar lined pipes, this surface covering should not be used for pipes in the treatment plant or reticulation.

Analytical methods

Referee method

Liquid�Solid Extraction and Capillary Column Gas Chromatography/Mass Spectrometry (EPA 525).


1. Liquid�Liquid Extraction and HPLC with Coupled Ultraviolet and Fluorescence Detection (EPA 550).

2. Liquid�Solid Extraction and HPLC with Coupled Ultraviolet and Fluorescence Detection (EPA 550.1).


Health considerations Benzo[a]pyrene is absorbed principally through the gastro-intestinal tract and the lungs. Absorbed benzo[a]pyrene is rapidly distributed to the organs and tissues and may be stored in mammary and fatty tissues. It crosses the placenta and is distributed in the developing tissue. Metabolism of benzo[a]pyrene occurs primarily in the liver. Benzo[a]pyrene metabolites are eliminated primarily in the faeces, with minor amounts excreted in urine. Human subjects skin-painted with benzo[a]pyrene developed skin lesions. Occupations associated with exposures to polynuclear aromatic hydrocarbons, of which benzo[a]pyrene is a component, have been associated clearly with human cancer. Benzo[a]pyrene has been shown to be mutagenic in tests with a strain of bacteria, however the diol-epoxide metabolite was considerably more mutagenic than the parent compound. The health effect of primary concern is carcinogenicity. Benzo[a]pyrene is one of the most potent carcinogens amongst the PAHs that have been tested to-date. The International Agency for Research on Cancer has classified benzo[a]pyrene in Group 2A (probably carcinogenic to humans).

Derivation of maximum acceptable value Benzo[a]pyrene appears to be a local carcinogen in that it induces tumours at the site of administration. Administration of benzo[a]pyrene in the diet of mice resulted in an increased incidence of forestomach tumours. Owing to the unusual protocol followed in this study which involved variable dosing patterns and age of sacrifice, these data could not be extrapolated accurately using the linearised multistage model normally applied to the derivation of the MAVs for carcinogens. However, a quantitative risk assessment was conducted using the two-stage birth-death mutation model. Using this model the estimated concentration of benzo[a]pyrene in drinking-water corresponding to an excess life-time cancer risk of one per 100,000 (10-5) is 0.0007 mg/L (0.7 µg/L).



1992. Organic Contaminants in New Zealand Potable Water Supplies: Report to Department of Health, New Zealand.


USEPA. National Primary Drinking Water Regulations: Technical Factsheet on polynuclear aromatic hydrocarbons (PAHs). This is available on: http://www.epa.gov/safewater/dwh/t-soc/pahs.html


Jolley RL, Bull RJ, Davis WP, et al (eds). 1985. Water Chlorination: Chemistry, environmental impact and health effects. Vol 5: 1515�[add page number]. Chelsea, Michigan: Lewis Publishers Inc.

Jolley RL, Condie LW, Johnson JD, et al (eds). 1990. Water Chlorination: Chemistry, environmental impact and health effects. Vol 6:, pp. 12�[add page number]. Chelsea, Michigan: Lewis Publishers, Inc.

WHO. 2003. Polynuclear Aromatic Hydrocarbons in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/59).



http://www.epa.gov/safewater/dwh/t-soc/dioxin.html

Bromochloroacetic acid New entry August 2005.

Maximum acceptable value There are insufficient data to derive MAVs for individual bromochloroacetic acid in drinking-water. WHO (2004) states that the available data relating to bromochloroacetic was considered inadequate to permit recommendation of a health-based guideline value.


1 To source waters

Brominated acetic acids are formed during disinfection of water which contains bromide ions and organic matter. Bromide ions occur naturally in surface water and groundwater and exhibit seasonal fluctuations in concentrations. Bromide ion concentrations can increase due to saltwater intrusion resulting from drought conditions, or due to pollution. Bromide is introduced into New Zealand surface waters usually by wind blown seaspray.


No known sources.


No known sources.

Forms and fate in the environment No information available.

Typical concentrations in drinking-water Brominated acetates generally are present in surface water and groundwater distribution systems at mean concentrations below 0.005 mg/L.

Removal methods Brominated acetic acids arise in waters as a disinfection by-product, so the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps.


Analytical methods

Referee method

[???]


[???]

Health considerations Data are limited on the oral toxicity of bromochloroacetic acid. Limited mutagenicity and genotoxicity data give generally positive results for bromochloroacetic acid. Data gaps include subchronic or chronic toxicity studies, multigeneration reproductive toxicity studies, standard developmental toxicity studies and carcinogenicity studies. The available data are considered inadequate to establish guideline values for these chemicals.

Derivation of maximum acceptable value The are insufficient data to derive a MAV for bromochloroacetic acid at this time.

References IPCS. 2000. Disinfectants and Disinfectant By-products. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 216).

WHO. 2003. Brominated Acetic Acids in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/79).


Bromochloroacetonitrile Revised July 2005.

Maximum acceptable value There are insufficient data to determine a MAV for bromochloroacetonitrile in drinking-water. WHO (2004) considered that the available data was inadequate to permit recommendation of a health-based guideline value.


1 To source waters

No known sources.


Brominated haloacetonitriles such as bromochloroacetonitrile are formed from organic precursors during chlorination of water containing bromide.



No known sources.

Form and fate in the environment Haloacetonitriles are reported to undergo hydrolysis in water, yielding nonvolatile products.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 209 zones, found bromochloroacetonitrile concentrations to range from �not detectable� (nd) to 0.005 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L).

Removal methods As bromochloroacetonitrile arises in waters as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.

Analytical methods

Referee method

A referee method cannot be selected for bromochloroacetonitrile because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for bromochloroacetonitrile for the above reason. However, the following information may be useful: Haloacetonitriles in water may be determined by solvent extraction with methyl tert-butyl ether and analysed by capillary column/electron capture/gas chromatography (EPA Method 551). Quantitation limits of 0.0004 mg/L (0.4 µg/L) are achievable. Interference may come from contaminated reagents or glassware.


Health considerations Haloacetonitriles are absorbed rapidly from the gastro-intestinal tract and metabolised to single carbon compounds. Insufficient data are available to determine whether haloacetonitriles can accumulate in specific organs. No data are available on the health effects of haloacetonitriles in humans. Bromochloroacetonitrile was a direct-acting mutagen in tests on bacteria and induced DNA damage (sister chromatid exchange and DNA strand breaks) in mammalian cells.

Derivation of maximum acceptable value There are insufficient data to determine a MAV for bromochloroacetonitrile in drinking-water.

References Australian Drinking-water Guidelines, NHMRC and AWRC, 1993 Draft.

IPCS. 2000. Disinfectants and Disinfectant By-products. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 216).


Bromodichloromethane Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of bromodichloromethane in drinking-water should not exceed 0.06 mg/L. The Australian Drinking Water Guidelines state that trihalomethane concentrations fluctuating occasionally (for a day or two annually) up to 1 mg/L are unlikely to pose a significant health risk. Action to reduce THMs is encouraged, but must not compromise disinfection, as nondisinfected water poses signficantly greater risk than THMs. The maximum contaminant level for total trihalomethanes (USEPA 2004) is 0.08 mg/L.


1 To source waters

Brominated trihalomethanes, such as bromodichloromethane, may occur in raw water as industrial contaminants. They are used as laboratory reagents, chemical intermediates, as fluids for mineral ore separation, as solvents for waxes, fats, and resins and as flame retardants.



Trihalomethanes, including bromodichloromethane, are most likely to be formed as by-products of the chlorination of drinking-water. Naturally-occurring bromide is oxidised by chlorine to form bromine (hypobromous acid and hypobromite ion) which reacts with organic matter, such as humic and fulvic acids, in the water, with the result that the trihalomethanes found in the water show varying degrees of bromine incorporation. When full bromine substitution occurs, bromoform is produced; mixed substitution of chlorine and bromine results in dibromochloromethane and bromodichloromethane. The concentration of trihalomethanes produced depend upon: pH, organic matter concentration, chlorine dose, bromide concentration, contact time and temperature.


No known sources.

Form and fate in the environment In air, brominated trihalomethanes may be removed through oxidation with atmospheric hydroxyl radicals. Volatilisation is a major removal mechanism for bromodichloromethane from water. Biodegradation occurs under anaerobic conditions. Hydrolysis is extremely low. Bioaccumulation in aquatic organisms may occur. Brominated trihalomethanes are expected to be mobile in soil.

Typical concentrations in drinking-water The review of organic contaminants in New Zealand drinking-water supplies between 1987�1992 contained bromodichloromethane results from 370 samples representing 157 chlorinated supplies. Bromodichloromethane was detected in 304 samples in concentrations ranging from 0.0002�0.076 mg/L (0.2�76 µg/L). The P2 Chemical Determinand Identification Programme sampled, from 511 zones, found bromodichloromethane concentrations to range from �not detectable� (nd) to 0.055 mg/L, with the median concentration being 0.0031 mg/L ((limit of detection = 0.002 mg/L).

Removal methods Bromodichloromethane can be removed from contaminated source waters by adsorption on to granular activated carbon, or by air stripping. Adsorption efficiency increases, and air stripping efficiency decreases, with increasing bromine substitution in the trihalomethane. However, as bromodichloromethane arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Trihalomethane concentrations in a chlorinated water increase with increasing pH. Concentrations in the finished water therefore can be reduced by ensuring that high pH levels are not present once the water is chlorinated.


Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Bromodichloromethane can be removed by adsorption on to granular activated carbon, or by air stripping. Adsorption efficiency increases, and air stripping efficiency decreases, with increasing bromine substitution in the trihalomethane. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.

Analytical methods

Referee method



1 Purge and Trap Capillary-Column Gas Chromatographic Method (APHA 6230D, EPA 502.2). 2 Liquid�Liquid Extraction and Gas Chromatography with Electron-Capture Detection (EPA 551).

Health considerations Available studies indicate that gastro-intestinal absorption is high for all trihalomethanes. They are fat soluble and accumulation is higher in tissues with high lipid content, including body fat, liver and kidneys. In a 90-day study in rats administered bromodichloromethane in drinking-water, mild to moderate histological changes in the liver and thyroid and a significant increase in the severity of hepatic lesions were observed at the highest dose. The International Agency for Research on Cancer has classified bromodichloromethane in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value Cancer risks have been estimated on the basis of increases in kidney tumour incidence in male mice, as these tumours give the most protective value. Hepatic tumours in female mice were not considered because of the possible role of the corn oil vehicle in induction of these tumours, although the estimated risks are within the same range. Using the linearised multistage model, the concentration of bromodichloromethane in drinking-water associated with an excess lifetime cancer risk of one per one hundred thousand (10-5) is 0.06 mg/L. This MAV is supported by a recently published study in rats that was not available for full evaluation.








WHO. 2003. Trihalomethanes in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/64).


Bromoform Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of bromoform in drinking-water should not exceed 0.1 mg/L. The maximum contaminant level for total trihalomethanes (USEPA 2004) is 0.08 mg/L.


1 To source waters

Brominated trihalomethanes may occur in raw water as industrial contaminants and from human activity. They have been used as laboratory reagents, chemical intermediates, as fluids for mineral ore separation, as solvents for waxes, fats, resins, and as flame retardants. Bromoform has been used as a sedative and cough depressant.


Brominated trihalomethanes are most likely to be formed as by-products of the chlorination of drinking-water. Naturally-occurring bromide is oxidised by chlorine to form bromine (hypobromous acid and hypobromite ion), which reacts with organic matter in the water, such as humic and fulvic acids, with the result that the trihalomethanes found in the water show varying degrees of bromine incorporation. When full bromine substitution occurs, bromoform is produced; mixed substitution of chlorine and bromine results in dibromochloromethane and bromodichloromethane. The concentration of trihalomethanes produced depend upon: pH, organic matter concentration, chlorine dose, bromide concentration, contact time and temperature.


No known sources.

Form and fate in the environment In air, brominated trihalomethanes may be removed by oxidation with atmospheric hydroxyl radicals. Volatilisation is a major removal mechanism for bromoform from water. Biodegradation occurs under anaerobic conditions. Hydrolysis is extremely low. Bioaccumulation in aquatic organisms may occur. Brominated trihalomethanes are expected to be mobile in soil.


Typical concentrations in drinking-water The review of organic contaminants in New Zealand drinking-water supplies between 1987�1992, contained bromoform results from 370 samples representing 157 chlorinated supplies. Bromoform was detected in 82 samples in concentrations ranging from 0.00033�0.029 mg/L (0.33�29 µg/L). The P2 Chemical Determinand Identification Programme, sampled from 511 zones, found bromoform concentrations to range from �not detectable� (nd) to 0.049 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L).

Removal methods Bromoform present in contaminated source waters may be removed by adsorption on to granular activated carbon, or by air stripping. Adsorption efficiency increases, and air stripping efficiency decreases, with increasing bromine substitution in the trihalomethane. However, as this compound arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Trihalomethane concentrations in chlorinated water increase with increasing pH. Finished water concentrations can therefore be reduced by ensuring that high pH levels are not present once the water is chlorinated. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Bromoform can be removed by adsorption on to granular activated carbon, or by air stripping. Adsorption efficiency increases, and air stripping efficiency decreases, with increasing bromine substitution in the trihalomethane. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants. Action to reduce THMs is encouraged, but must not compromise disinfection, as nondisinfected water poses significantly greater risk than THMs.

Analytical methods

Referee method





Health considerations Available studies indicate that gastro-intestinal absorption is high for all trihalomethanes. They are fat soluble and accumulation is higher in tissues with high lipid content, including body fat, liver and kidneys. In a 90-day study in rats administered bromoform in drinking-water, mild to moderate histological changes in the liver and thyroid and a significant increase in the hepatic lesions were observed at the highest dose. In an NTP bioassay, bromoform induced a small increase in relatively rare tumours of the large intestine in rats of both sexes but did not induce tumours in mice. Data from a variety of assays on the genotoxicity of bromoform are equivocal. In the past, orally administered bromoform was used as a sedative for children with whooping cough. Occasionally, instances of death were reported due to accidental overdose. The clinical signs in fatal cases were central nervous system depression followed by respiratory failure. The International Agency for Research on Cancer has classified bromoform in Group 3 (not classifiable as to its carcinogenicity in humans).

Derivation of maximum acceptable value A TDI of 17.9 mg/kg of body weight has been used, based on the absence of histopathological lesions in the liver in a well-conducted and well-documented 90-day study in rats, using an uncertainty factor of 1000 (100 for intra- and interspecies variation and 10 for possible carcinogenicity and short duration of exposure). The MAV for bromoform in drinking-water was derived as follows:

17.9 mg/kg body weight per day x 70 kg x 0.2 = 0.125 mg/L (rounded to 0.1 mg/L) 2 L x 1000

where: • tolerable daily intake (TDI) = 17.9 mg/kg body weight • average weight of an adult = 70 kg • proportion of tolerable daily intake allocated to drinking-water = 0.2 • average quantity of water consumed by an adult per day = 2 L • uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for possible carcinogenicity

and short duration of the study).








WHO. 2003. Trihalomethanes in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva,World Health Organization (WHO/SDE/WSH/03.04/64).


Carbon tetrachloride Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of carbon tetrachloride in drinking-water should not exceed 0.005 mg/L (5 µg/L). The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


1 To source waters

Carbon tetrachloride can be released to the aquatic environment as an industrial contaminant. It is used mainly in the production of chlorofluorocarbon refrigerants, foam blowing agents, and solvents. It is also used in the manufacture of paints and plastics, as a solvent in metal cleaning, and in fumigants. Carbon tetrachloride is listed as a controlled substance in the New Zealand Ozone Layer Protection Act, 1990 and can only be obtained with a permit. Its occurrence in the New Zealand environment should therefore decrease.


Carbon tetrachloride may appear in water as the result of its presence in the chlorine used to treat the water.


No known sources.

Forms and fate in the environment Most carbon tetrachloride released to the environment reaches the atmosphere, where it may be removed by photolysis. Carbon tetrachloride has an estimated half-life of 50 years in the atmosphere. It migrates readily from surface water to the atmosphere in a matter of days or weeks. Levels in anaerobic groundwater may remain elevated for months or years. Carbon tetrachloride is capable of adsorbing to organic matter in soils. Migration to groundwater is possible. Bioaccumulation has not been observed.

Typical concentrations in drinking-water The review of organic contaminants in New Zealand drinking-water supplies between 1987�1992 contained carbon tetrachloride results from 70 samples. Carbon tetrachloride was detected in three samples in concentrations ranging from 0.0004�0.0006 mg/L (0.4 � 0.6 µg/L). The P2 Chemical Determinand Identification Programme, sampled from 332 zones, found no carbon tetrachloride at detectable concentrations (limit of detection = 0.0005 mg/L).


Removal methods Carbon tetrachloride present in contaminated source waters can be removed by adsorption on to granular activated carbon or by air stripping. As carbon tetrachloride may appear in water as the result of its presence in the chlorine used to treat the water, reducing the organic loading of the water, and hence the chlorine demand, will minimise the amount of chlorine that has to be added, and therefore the amount of carbon tetrachloride appearing in the water.

Analytical methods

Referee method




Health considerations Carbon tetrachloride is absorbed readily from the gastro-intestinal tract, the respiratory tract and the skin. It is distributed to all major organs, with highest concentrations in fatty tissues. Carbon tetrachloride is thought to be metabolised in the liver to chloroform and other products, which binds to macromolecules, initiating lipid peroxidation and destroying cell membranes. Although available data on concentrations in food are limited, the intake from air is expected to be much greater than that from food or drinking-water. Workers exposed to 20�80 ppm carbon tetrachloride for 2�3 months experienced nausea, depression, dyspepsia and narcosis. Kidney and liver damage have been reported after short term exposures to 200 ppm. Similar effects have been reported following acute oral exposures. Death may result from ingestion of as little as 1.5 mL for an adult. Alcohol consumption enhances carbon tetrachloride-induced hepatic and renal effects in adults. Carbon tetrachloride does not exhibit any evidence of mutagenic activity in tests with bacteria or cultured liver cells. Carbon tetrachloride has been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). It has caused liver and other tumours in rats, mice and hamsters after oral, subcutaneous, and inhalation exposure. The length of time to the development of the first tumour has sometimes been short, within 12�16 weeks in some experiments.

Derivation of maximum acceptable value A tolerable daily intake (TDI) of 1.4 mg/kg of body weight was used, based on a NOAEL of 1 mg/kg of body weight per day for hepatotoxic effects in a 12-week oral gavage study in rats, incorporating a conversion factor of 5/7 for daily dosing and applying an uncertainty factor of 500 (100 for inter- and intraspecies variation, 10 for the duration of the study and a modifying factor of 0.5 because it was a bolus study). The MAV for carbon tetrachloride in drinking-water was derived as follows:

1 x (5/7) mg/kg body weight per day x 70 kg x 0.1 = 0.005 mg/L (5 µg/L) 2 L x 1000 x 0.5


where:

• no observable adverse effect level = 1 mg/kg body weight per day based a 12-week oral gavage study in rats (normalised for five days/week dosing in derivation)


• proportion of tolerable daily intake allocated to drinking-water = 0.1

• average quantity of water consumed by an adult per day = 2 L

• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for evidence of possibly non-genotoxic carcinogenicity). No additional factor for the short duration of the study was incorporated. It was considered to be unnecessary because the compound was administered in corn oil in the critical study and available data indicate that the toxicity following administration in water may be an order of magnitude less.




IPCS. 1999. Carbon Tetrachloride. World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 208). Geneva.

New Zealand Ozone Layer Protection Act 1990.




WHO. 2003. Carbon Tetrachloride in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva,World Health Organization (WHO/SDE/WSH/03.04/82).

White GC. 1986. Handbook of Chlorination (2nd ed). Van Nostrand Reinhold.


Chloroacetones Revised July 2005. dichloroacetones (1,1-dichloropropanone and 1,3-dichloropropanone) trichloroacetones (1,1,1-trichloropropanone and 1,1,3-trichloropropanone)

Maximum acceptable value There are insufficient data to determine a MAV for any of the chloroacetones in drinking-water.



1 To source waters

Dichloroacetones may be present in source waters as a result of their discharge from industries in which they are used. Chloroacetones are reagents in the synthesis of drugs, perfumes, insecticides and vinyl compounds.


Dichloroacetones may form as a result of chlorination from the reaction between chlorine and large organic molecules.


No known sources.

Forms and fate in the environment The chlorinated acetones undergo hydrolysis. The rate of this reaction depends on the number of chlorine atoms in the molecule and where they are positioned. 1,1-dichloroacetone is likely to undergo hydrolysis more rapidly than 1,3-dichloroacetone.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 491 zones, found dichloroacetone concentrations to range from �not detectable� (nd) to 0.0031 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L). 1,1-dichloroacetone concentrations are estimated to be less than 0.01 mg/L and usually less than 0.001 mg/L (WHO 2004). The P2 Chemical Determinand Identification Programme, sampled from 491 zones, found trichloroacetone concentrations to range from �not detectable� (nd) to 0.009 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L).

Removal methods Chloroacetones can be removed from contaminated source waters by adsorption on to granular activated carbon, or by air stripping. However, as chloroacetones arises in waters principally as a disinfection by-product, the preferred method for minimising their formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible.


Trihalomethane concentrations in chlorinated water increase with increasing pH. Concentrations in the finished water can therefore be reduced by ensuring that high pH levels are not present once the water is chlorinated. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Chloroacetones can be removed by adsorption on to granular activated carbon, or by air stripping. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.

Analytical methods

Referee method

A referee method cannot be selected for chloroacetones because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for chloroacetones for the above reason. However, the following information may be useful:

1 Chloroacetones in drinking-water may be analysed by purge and trap gas chromatography with mass spectrometry or electron capture detection (Method APHA 6232 or EPA Method 502.1). The limit of quantification is approximately 0.001 mg/L (1 µg/L). Interference may occur from impurities in the purge gas and organic compounds outgassing from the trap system.

2 A solvent extraction procedure with methyl tert-butyl ether (MTBE) may be used (EPA Method 551) and analysis by gas chromatography with mass spectrometry or electron capture detection. The detection limit is approximately 0.00002 mg/L (0.02 µg/L). Interference may occur from impurities in the reagents or glassware used for extraction.

Health considerations Studies with single doses of 1,1-dichloroacetone indicate that it affects the liver. A number of chlorinated acetones showed mutagenic activity. One carcinogenicity study concluded that 1,3-dichloroacetone is a tumour initiator in mouse skin. Acute oral toxicity studies in mice using 1,1-dichloropropanone and 1,3-dichloropropanone found no toxic effects with single doses of 130 mg/kg and 20 mg/kg respectively. No long-term toxicity studies have been reported.

Derivation of maximum acceptable value There are limited and insufficient data on the chloroacetones on which to propose a MAV for any of the chloroacetones.




USEPA Method 551. 1990. Determination of Chlorination Disinfection By-products and Chlorinated Solvents in Drinking Water by Liquid�Liquid Extraction and Gas Chromatography with Electron Capture Detection. Cincinnati, Ohio: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).

WHO. 2003. Chloroacetones in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/50).


Chloroform Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of chloroform in drinking-water should not exceed 0.2 mg/L. The maximum contaminant level for total trihalomethanes (USEPA 2004) is 0.08 mg/L.


1 To source waters

Chloroform was used primarily as the starting material in the manufacture of the refrigerant fluorocarbon-22. It is an important extraction solvent for resins, gums, and other products and therefore may enter raw water as an industrial contaminant.


Trihalomethanes, including chloroform, are most likely to be formed as by-products of the chlorination of drinking-water. Chlorine reacts with natural organic materials such as fulvic and humic acids to form chloroform, which is the most common trihalomethane. The amount of chloroform depends on temperature, pH, chlorine concentration, the concentration of organic matter, contact time and bromide concentration.


No known sources.

Form and fate in the environment Chloroform can be photo-oxidised in air. Volatilisation is a major removal mechanism for chloroform from water. Biodegradation in groundwater can occur but is slow, with a half-life ranging from weeks to years. Chloroform does not adsorb strongly to soil or sediments. Bioaccumulation in aquatic organisms may occur.

Typical concentrations in drinking-water The review of organic contaminants in New Zealand drinking-waters between 1987�1992 contained chloroform results from 370 samples representing 157 chlorinated supplies. Chloroform was detected in 288 samples in concentrations ranging from 0.0004�0.122 mg/L (0.4�122 µg/L).


The P2 Chemical Determinand Identification Programme, sampled from 511 zones, found chloroform concentrations to range from �not detectable� (nd) to 0.141 mg/L, with the median concentration being 0.0068 mg/L (limit of detection = 0.004 mg/L).

Removal methods Chloroform present in contaminated source waters can be removed by adsorption on to granular activated carbon, or by air stripping. However, as chloroform arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Trihalomethane concentrations in chlorinated water increase with increasing pH. Concentrations in the finished water can therefore be reduced by ensuring that high pH levels are not present once the water is chlorinated. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Chloroform can be removed by adsorption on to granular activated carbon, or by air stripping. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.

Analytical methods

Referee method




Health considerations Available studies indicate that gastro-intestinal absorption is high for all trihalomethanes and because of their high lipophilicity, accumulation is higher in tissues with high lipid content, including body fat, liver and kidneys.


Individuals may be exposed during showering to elevated concentrations from chlorinated tap water. Based on estimates of mean exposure from various media, the general population is exposed to chloroform principally in food, drinking-water and indoor air, in approximately equivalent amounts. Chloroform is a central nervous system depressant. It can also affect the liver and kidney functions. A fatal dose will result in respiratory or cardiac arrest. Workers exposed to chloroform by inhalation at levels between 0.1 and 1.2 g/m3 for one or more years reported symptoms including nausea, lassitude, dry mouth, flatulence, thirst, depression, irritability and the impression of scalding urine. In another incidence, workers inhaling chloroform at similar levels for 1-4 years had an increased incidence of viral hepatitis and enlarged liver. The genotoxicity of chloroform has been studied in a wide variety of assays, but the results are inconclusive. The weight of evidence for genotoxicity of chloroform is considered negative. The weight of evidence for liver tumours in mice is consistent with a threshold mechanism of induction. Although it is plausible that kidney tumours in rats may similarly be associated with a threshold mechanism, there are some limitations of the database in this regard. The most universally observed toxic effect of chloroform is damage to the centrilobular region of the liver. The severity of these effects per unit dose administered depends on the species, vehicle and method by which the chloroform is administered. The International Agency Responsible for Research on Cancer has classified chloroform as Class 2B.

Derivation of maximum acceptable value This MAV is supported by a 7.5-year study in dogs, in which a LOAEL of 15 mg/kg of body weight per day was observed for liver effects (applying an uncertainty factor of 1000 (100 for intra- and interspecies variation and 10 for the use of a LOAEL) and allocating 50% of the tolerable daily intake to drinking-water and correcting for six days per week dosing. The MAV for chloroform in drinking-water was derived as follows:

15 x (6/7) mg/kg body weight/day x 70 kg x 0.5 = 0.225 mg/L (rounded to 0.2 mg/L) 2 L x 1000

where:

• lowest observable adverse effect level = 15 mg/kg body weight per day, based on slight hepatotoxicity (increases in hepatic serum enzymes and fatty cysts) observed in beagle dogs ingesting 15 mg of chloroform per kg of body weight per day in toothpaste for 7.5 years (normalised for 6 days/week dosing)


• proportion of tolerable daily intake allocated to drinking-water = 0.5 based on estimates indicating that the general population is exposed to chloroform principally in food, drinking-water and indoor air in approximately equivalent amounts and that most of the chloroform in indoor air is present as a result of volatilization from drinking-water


• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for evidence of possibly non-genotoxic carcinogenicity). No additional factor for the short duration of the study was incorporated. It was considered to be unnecessary because the compound was administered in corn oil in the critical study and available data indicate that the toxicity following administration in water may be an order of magnitude less.










2-chlorophenol Revised July 2005.

Maximum acceptable value There are insufficient data to set a health based MAV for 2-chlorophenol in drinking-water. The Australian Drinking Water Guidelines have a health value of 0.3 mg/L.


1 To source waters

2-Chlorophenol may occur in raw water as an industrial contaminant, or through agricultural activity. It may be used as a precursor for the production of higher chlorophenols and dyestuffs. Chlorophenols are used commercially as preservatives, moth-proofing agents, germicides and anti-mildew agents.


Chlorophenols are most likely to occur in drinking-water as disinfection by-products through the reaction of naturally-occurring organic matter with chlorine.


No known sources.

Form and fate in the environment Because chlorinated phenols are moderately water-soluble, weakly acidic, and have low vapour pressures, it is anticipated that volatilisation does not play a significant role in removing these chemicals from water. Photolytic breakdown of dilute solutions of monochlorophenols has been reported. Sorption is not significant for monochlorophenols. Biodegradation appears to be the primary removal mechanism of chlorinated phenols from surface waters. Aquatic biota may bioconcentrate chlorinated phenols with bioconcentration factors increasing with increasing chlorine substitution.

Typical concentrations in drinking-water Only a few samples contained in the 1992 review of organic contaminants in New Zealand drinking-water supplies were analysed for 2-chlorophenol. All concentrations reported were below the detection limit of 0.01 mg/L.


Removal methods Chlorophenols can be removed from contaminated source waters by adsorption on to activated carbon. The effectiveness of the processes is pH dependent. Greater adsorption occurs as the pH is lowered. However, as this compound arises in New Zealand waters principally during water treatment as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. The formation of chlorophenols can be reduced by the use of chlorine dioxide in place of chlorine. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Chlorophenols can be removed by adsorption on to activated carbon. The effectiveness of the processes is pH dependent. Greater adsorption occurs as the pH is lowered. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.

Analytical methods

Referee method

A referee method cannot be selected for 2-chlorophenol because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for 2-chlorophenol for the above reason. However, the following information may be useful:

1 Chlorophenols in water can be solvent extracted with dichloromethane (Method APHA 6410) and analysed by gas chromatography with mass spectrometry detection (Method APHA 6410 or EPA 8270). The detection limit for this method is 0.003 mg/L (3 µg/L). Interference may come from contaminated reagents or glassware.

2 A more sensitive and specific method of analysis for chlorophenols is to solvent extract with dichloromethane and derivatise with pentafluorobenzyl ether and analyze by gas chromatography with electron capture detection (Method EPA 604 or APHA 6420). The limit of quantification for this method is 0.0006 mg/L (0.6 µg/L). The specificity of this method reduces the likelihood of interferences.


Health considerations Chlorophenols are well-absorbed after oral administration and they readily penetrate the skin. Chlorophenols do not appear to accumulate in body tissues in rats but are rapidly metabolised and eliminated from the body, principally in urine. Exposure to chlorophenols via tap water has been estimated to be less than 10% of total dietary exposure. There is a limited data base on the toxicity of 2-chlorophenol. One study which exposed rats to 50 mg/kg body weight per day for 10 weeks reported treatment related increases in conception rate, an increase in the number of still births and a decrease in the size of the litters. It has been suggested that 2-chlorophenol is a co-carcinogen.

Derivation of maximum acceptable value Because of the limited data base on the toxicity of 2-chlorophenol, no health-based MAV has been derived. Chlorophenols generally have very low taste and odour thresholds. The taste threshold in water for 2-chlorophenol is 0.0001 mg/L and its odour threshold is 0.001 mg/L. Micro-organisms in distribution systems may sometimes methylate chlorophenols to produce chlorinated anisoles, for which the odour threshold is considerably lower.






WHO. 2003. Chlorophenols in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/47).


Chloropicrin Revised July 2005. (Also called trichloronitromethane.)

Maximum acceptable value There are insufficient data to derive a MAV for chloropicrin in drinking-water.


1 To source waters

Chloropicrin may enter raw water as an industrial contaminant. It may be used as a reagent in the synthesis of organic chemicals, in the manufacture of methyl violet, as a fumigant for stored grain and as a chemical warfare agent.



Chloropicrin is formed in water by the reaction of chlorine with humic acids, amino acids, and nitrophenols. The presence of nitrate increases the formation of chloropicrin.


No known sources.

Form and fate in the environment Chloropicrin in water is reduced to chloroform when reducing agents are added to the water to remove excess chlorine. In the presence of light, it is degraded to carbon dioxide, chloride ion, and nitrate ion.

Typical concentrations in drinking-water No data are available on the concentration of chloropicrin in New Zealand drinking-water supplies. In a US study of 36 supplies suspected of having high disinfection by-product levels, the highest concentration of chloropicrin measured was 0.0056 mg/L (5.6 µg/L).

Removal methods No information is available on methods to remove chloropicrin from contaminated source waters. As chloropicrin arises in water principally as a disinfection by-product, the preferred method for minimising its concentration is to reduce the formation of natural organic matter (NOM) coming into contact with the chlorine. The presence of nitrate also increases its formation. Therefore the removal of nitrate, or the selection of a source with low nitrate concentrations will help to reduce chloropicrin formation. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.

Analytical methods

Referee method

A referee method cannot be selected for chloropicrin because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for chloropicrin because a MAV has not been established. However, the following information may be useful:


Chloropicrin in water may be analysed by solvent extraction with methyl tert-butyl ether (MTBE) and analysis by gas chromatography with electron capture detection (Method EPA Draft 551). The limit of determination is approximately 0.00002 mg/L (20 ng/L). APHA (1998) Method 5710D also uses MTBE and gas chromatography with electron capture detection.

Health considerations No long-term data are available on health effects in humans. Decreased survival and body weights have been reported following long-term oral exposure in laboratory animals. Chloropicrin has been shown to be mutagenic in bacterial tests and in in vitro assays in lymphocytes. Because of the high mortality in a carcinogenesis bioassay and the limited number of end-points examined in the 78-week toxicity study, the available data were considered inadequate to permit the establishment of a guideline value for chloropicrin. Studies investigating the chronic toxicity of chloropicrin in mice and rats found that there was an association between high dose levels (46 and 50 mg/kg body weight per day respectively) and accelerated mortality. In humans, inhalation of chloropicrin at 0.002 mg/L for 1 minute caused pulmonary effects. Studies to date do not permit an evaluation of the carcinogenicity of chloropicrin because of the short survival time of dosed animals. Chloropicrin exhibited mutagenic activity in some tests with bacteria, and with human lymphocytes in vitro.

Derivation of maximum acceptable value The are insufficient data to derive a MAV for chloropicrin at this time.




USEPA Draft Method 551. 1990. Determination of Chlorination Disinfection By-products and Chlorinated Solvents in Drinking Water by Liquid�Liquid Extraction and Gas Chromatography with Electron Capture Detection. Cincinnati, USA: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).

WHO. 2003. Chloropicrin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/52).



Di(2-ethylhexyl)adipate (DEHA) Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of di(2-ethylhexyl)adipate in drinking-water should not exceed 0.1 mg/L. The MAV is provisional because:

• WHO (2004) considered �a health-based value of 0.08 mg/L can be calculated for DEHA on the basis of a TDI of 0.28 mg/kg of body weight based on fetotoxicity in rats and allocating 1% of the TDI to drinking-water. However, because DEHA occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a health-based guideline value. The 1993 WHO Guidelines had proposed a health-based guideline value of 0.08 mg/litre for DEHA in drinking-water�

• the maximum contaminant level (USEPA 2004) is 0.4 mg/L.


1 To source waters

Di(2-ethylhexyl)adipate is found in raw water through human and industrial activity. It is used primarily as a plasticiser for synthetic resins such as PVC, so may leach from tubing, dishes, containers, etc. It is also used as a solvent, and for aircraft lubrication.


No known sources.


Di(2-ethylhexyl)adipate may leach from PVC pipes used in the distribution system.

Forms and fate in the environment Because of its low water solubility, di(2-ethylhexyl)adipate released into the environment would be expected to partition to solids (biota, sediment, soil). Biodegradation is likely to be a significant removal mechanism from the aquatic environment.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 348 zones, did not find any di(2-ethylhexyl)adipate at detectable concentrations (limit of detection = 0.002 mg/L). Overseas studies have detected DEHA at concentrations between 0.000001 mg/L (1 ng/L) to 0.0001 mg/L (100 ng/L) in treated drinking water.

Removal methods No information is available on processes that can be used to remove di(2-ethylhexyl)adipate from water.


Analytical methods

Referee method



1 Liquid�Liquid Extraction or Liquid�Solid Extraction and Gas Chromatography with Photoionisation Detection (EPA 506).

Health considerations As a consequence of its use in PVC films, food is the most important source of human exposure (up to 20 mg/day). DEHA is of low short-term toxicity; however, dietary levels above 6000 mg/kg of feed induce peroxisomal proliferation in the liver of rodents. This effect is often associated with the development of liver tumours. DEHA induced liver carcinomas in female mice at very high doses but not in male mice or rats. It is not genotoxic. IARC has placed DEHA in Group 3. Di(2-ethylhexyl)adipate is absorbed readily when given orally to rats and mice. It is distributed widely in the body, with highest levels of metabolites reported in fatty tissue, liver and kidney. Transplacental transport of di(2-ethylhexyl)adipate has been reported. The acute oral toxicity of di(2-ethylhexyl)adipate is low. No data exists on the effects of ingested di(2-ethylhexyl)adipate in humans. Short-term mouse and rat toxicity studies have demonstrated that high dietary levels of the compound induce liver toxicity, which is often associated with the development of liver tumours, particularly in female mice. Di(2-ethylhexyl)adipate has not exhibited mutagenic activity when applied to strains of bacteria or to mammalian cells. The International Agency for Research on Cancer has concluded that there is limited evidence that di(2-ethylhexyl)adipate is carcinogenic in mice. It is not classifiable as to its carcinogenicity in humans.

Derivation of maximum acceptable value Although di(2-ethylhexyl)adipate is carcinogenic in mice, the toxicity profile and lack of evidence of mutagenicity of di(2-ethylhexyl)adipate support the use of a tolerable daily intake approach to deriving a MAV for di(2-ethylhexyl)adipate in drinking-water. The MAV has been derived on the basis of a no observable adverse effects level determined in a 90-day toxicological study in rats. The MAV for di(2-ethylhexyl)adipate in drinking-water was derived as follows:

28 mg/kg body weight per day x 70 kg x 0.01 = 0.098 mg/L (rounded to 0.1 mg/L) 2 L x 100

where: • no observable adverse effect level = 28 mg/kg body weight per day in a fetotoxicity study in rats • average weight of an adult = 70 kg • proportion of tolerable daily intake allocated to drinking-water = 0.01 • average quantity of water consumed by an adult per day = 2 L • uncertainty factor = 100 for intra- and interspecies variation.




USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water. Supplement 1, Report No. EPA/600/4-90-020.


WHO. 2003. Di(2-ethylhexyl)adipate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/68).


Di(2-ethylhexyl)phthalate (DEHP) Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of di(2-ethylhexyl)phthalate in drinking-water should not exceed 0.009 mg/L (9 µg/L). The maximum contaminant level (USEPA 2004) is 0.006 mg/L.


1 To source waters

Di(2-ethylhexyl)phthalate is found in raw water through human and industrial activity. It is used primarily as a plasticiser in many flexible PVC products and in vinyl chloride co-polymer resins, and may leach from tubing, dishes, containers, etc. DEHP is the most widely used plasticiser. It is also used as a replacement for polychlorinated biphenyls in dielectric fluids for low voltage electrical capacitors. Found in surface water, groundwater and drinking-water in concentrations of a few micrograms per litre; in polluted surface water and groundwater, concentrations of hundreds of micrograms per litre have been reported.


No known sources.


Di(2-ethylhexyl)phthalate may leach from PVC pipes used in the distribution system.

Forms and fate in the environment Di(2-ethylhexyl)phthalate is insoluble in water but may be transported in the aquatic environment as complexes with humic substances or sorption on to particulate matter and biota. Volatilisation and photolysis are not expected to be important removal mechanisms of di(2-ethylhexyl)phthalate from the aquatic environment, but biodegradation and bioaccumulation will be significant.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 348 zones, found di(2-ethylhexyl)phthalate concentrations to range from �not detectable� (nd) to 0.007 mg/L (a single sample), with the median concentration being �nd� (limit of detection = 0.002 mg/L). Overseas studies have detected DEHP in drinking-water on a few occasions at concentrations from 0.00005 mg/L (50 ng/L) to 0.01 mg/L.

Removal methods No information is available on processes that can be used to remove di(2-ethylhexyl)phthalate from water.

Analytical methods

Referee method



1 Liquid�Liquid Extraction or Liquid�Solid Extraction and Gas Chromatography with Photoionisation Detection (EPA 506).

Health considerations In general, food will be the main exposure route. In rats, di(2-ethylhexyl)phthalate is well-absorbed from the gastro-intestinal tract after oral administration. Absorption is lower in humans and 11�25% of an ingested dose was found in urine. Mono(2-ethylhexyl)phthalate and its metabolites are extensively distributed throughout the body in rodents. The highest levels were found in the liver and fatty tissue. No or little accumulation occurs in rats. The acute oral toxicity of di(2-ethylhexyl)phthalate is low. Liver and testes appear to be the main target organs in di(2-ethylhexyl)phthalate toxicity. The most striking effect in short-term toxicity studies is the proliferation of hepatic peroxisomes. The available data suggest that primates, including humans, are far less sensitive to this effect than rodents. Tests on mice, rats, guinea pigs and ferrets have reported testicular effects, consisting of atrophy, tubular degeneration and inhibition or cessation of spermatogenesis. In similar studies, other effects reported included suppression of fertility, foetal mortality, foetal resorption, decreased foetal weight, neural tube effects and skeletal disorders. Di(2-ethylhexyl)phthalate has not exhibited mutagenic activity when applied to strains of bacteria or to mammalian cells. In long-term oral carcinogenicity studies, hepatocellular carcinomas were found in rats and mice. The International Agency for Research on Cancer concluded that di(2-ethylhexyl)phthalate is possibly carcinogenic to humans (Group 2B).


Derivation of maximum acceptable value Based on the absence of evidence for genotoxicity and the suggested relationship between prolonged proliferation of liver peroxisomes and the occurrence of heptacellular carcinomas, a tolerable daily intake approach was taken to derive the MAV. The MAV for di(2-ethylhexyl)phthalate in drinking-water was derived as follows:

2.5 mg/kg body weight/day x 70 kg x 0.01 = 0.00875 mg/L (rounded to 0.009 mg/L) 2 L x 100

where:

• no observable adverse effect level = 2.5 mg/kg body weight per day based on peroxisomal proliferation in the liver in rats




• uncertainty factor = 100 for intra- and interspecies variation (although the mechanism for heptacellular tumour induction is not fully resolved, using a NOAEL derived from the far most sensitive species with respect to the particularly sensitive end-point of peroxisomal proliferation justifies the use of this uncertainty value).


Canadian Water Quality Guidelines. 1992.


USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water. Supplement 1, Report No EPA/600/4-90-020.


WHO. 2003. Di(2-ethylhexyl)phthalate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/29).


Dialkyltins Revised July 2005.

Maximum acceptable value There are insufficient data to derive MAVs for individual dialkyltins in drinking-water.


1 To source waters

The group of compounds known as the organotins comprises a large number of compounds with different properties and applications. Of these, the dialkyl and tributyltin compounds are the ones most likely to be found in raw water. The di-substituted organotins are employed as stabilisers in plastics, including PVC water pipes.



No known sources.


Organotin stabilisers may leach from PVC pipes into water for a short time after installation.

Forms and fate in the environment Unknown quantities of organotins may be released into air from factories that produce polyurethane or PVC resins stabilised with organotins. There is little information given on the fate of organotins in the aquatic environment.

Typical concentrations in drinking-water No data are available on the concentration of dialkyltins in New Zealand drinking-water supplies. There is evidence for organotin stabilisers leaching from plastic pipes, with one overseas study reporting a concentration for dibutyltin sulphide of 0.1 mg/L in a static water that had been in contact with a plastic pipe.

Removal methods No information is available on processes that can be used to remove these compounds from water. Consumer exposure to these compounds could be reduced by ensuring that new piping, from which organometallic compounds are likely to leach, are filled with water, the pipes left filled for several days, and then flushed. The process should be repeated a number of times. Monitoring the levels of the metals at the beginning of each cycle will indicate the extent to which the leaching is still occurring.

Analytical methods

Referee method

A referee method cannot be selected for dialkyltins because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for dialkyltins because a MAV has not been established. However, the following information may be useful: The organotins can be analysed using a solvent extraction procedure (Greaves and Unger, 1988). They are extracted using a hexane-tropolone mixture and derivatised to form hexylbutyltins. Analysis is by gas chromatography with flame photometric detection. Limits of quantification are less than 0.000002 mg/L (2 ng/L).

Health considerations Available data suggest that organotins are poorly absorbed. Following oral administration in rodents they tend to be distributed primarily in the liver and kidney and excreted in the faeces. Dialkyltins are primarily immunotoxins and they appear to be of low general toxicity.


The principal effect on rats fed diets containing dioctyltin dichloride for six weeks was a reduction in thymus weight. Rats administered a mixture of octyltin trichloride and dioctyltin dichloride in the diet had highly significant increased frequency of primary tumours of the thymus. No evidence of mutagenicty was observed with dioctyltin dichloride and dibutyltin diacetate. Dibutyltin dichloride and dioctyltin dichloride have been reported to be positive in mammalian cell mutation assays in vitro in the absence of metabolic activation, and dibutyltin sulphide increased the incidence of chromosomal aberrations in rat bone marrow cells in vivo.

Derivation of maximum acceptable value There are insufficient data available to propose MAVs for individual dialkyltins.


Greaves J, Unger MA. 1988. Selected ion monitoring assay for tributyltin and its degradation products. Biomedical and Environmental Mass Spectrometry 15: 565�9.

WHO. 2003. Dialkyltins in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/109).


Dibromoacetic acid New entry August 2005.

Maximum acceptable value There are insufficient data to derive MAVs for individual dibromoacetic acid in drinking-water. WHO (2004) states that the available data relating to dibromoacetate was considered inadequate to permit recommendation of a health-based guideline value.


1 To source waters

Brominated acetic acids are formed during disinfection (with ozone) of water which contains bromide ions and organic matter. Bromide ions occur naturally in surface water and groundwater and exhibit seasonal fluctuations in concentrations. Bromide ion concentrations can increase due to saltwater intrusion resulting from drought conditions, or due to pollution. Bromide is introduced into New Zealand surface waters usually by wind blown seaspray.


No known sources.


No known sources.




Removal methods Brominated acetic acids arise in waters as a disinfection by-product, so the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the ozone. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps.

Analytical methods

Referee method

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[???]

Health considerations The database for dibromoacetic acid is considered inadequate for the derivation of a MAV. There are no systemic toxicity studies of subchronic duration or longer. The database also lacks suitable toxicokinetic studies, a carcinogenicity study, a developmental study in a second species and a multigeneration reproductive toxicity study (one has been conducted but is currently being evaluated by the USEPA). Available mutagenicity data suggest that dibromoacetate is genotoxic. Data gaps include subchronic or chronic toxicity studies, multigeneration reproductive toxicity studies, standard developmental toxicity studies and carcinogenicity studies. The available data are considered inadequate to establish guideline values for these chemicals.

Derivation of maximum acceptable value The are insufficient data to derive a MAV for dibromoacetic acid at this time.





Dibromoacetonitrile Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of dibromoacetonitrile in drinking-water should not exceed 0.08 mg/L.


1 To source waters

No known sources.


Halogenated acetonitriles are produced during water chlorination or chloramination from naturally occurring substances, including algae, fulvic acid and proteinaceous material. In general, increasing temperature and/or decreasing pH have been associated with increasing concentrations of halogenated acetonitriles. Ambient bromide levels appear to influence, to some degree, the speciation of halogenated acetonitrile compounds. Dichloroacetonitrile is by far the most predominant halogenated acetonitrile species detected in drinking-water.


No known sources.

Forms and fate in the environment Haloacetonitriles are reported to undergo hydrolysis in water, yielding nonvolatile products.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 209 zones, found dibromoacetonitrile concentrations to range from �not detectable� (nd) to 0.0041 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L).

Removal methods As dibromoacetonitrile arises in waters as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.


Analytical methods

Referee method

Liquid�Liquid Extraction and Gas Chromatography with Electron-Capture Detection (EPA 551).


No alternative methods have been recommended for dibromoacetonitrile because no methods meet the required criteria.

Health considerations Haloacetonitriles are rapidly absorbed from the gastro-intestinal tract and metabolised to single carbon compounds. Insufficient data are available to determine whether haloacetonitriles can accumulate in specific organs. Dibromoacetonitrile is currently under test (WHO 2004) for chronic toxicity in mice and rats. None of the available reproductive or developmental studies were adequate to use in the quantitative dose�response assessment. The data gap may be particularly relevant since cyanide, a metabolite of dibromoacetonitrile, induces male reproductive system toxicity, and due to uncertainty regarding the significance of the testes effects observed in the 14-day National Toxicology Program (NTP) rat study. No data are available on the health effects of haloacetonitriles in humans. Several short-term exposure studies of dibromoacetonitrile on rats have concluded that decreased body weight is the most sensitive end-point. Dibromoacetonitrile was a direct-acting mutagen in tests on bacteria and induced DNA damage (sister chromatid exchange and DNA strand breaks) in mammalian cells. IARC has concluded that dibromoacetonitrile is not classifiable as to its carcinogenicity in humans.

Derivation of maximum acceptable value A tolerable daily intake approach been used to derive the MAV for dibromoacetonitrile in drinking-water using a no observable adverse effects level determined for effects on body weight in a 90-day study in rats. The MAV for dibromoacetonitrile in drinking-water was derived as follows:

11.3 mg/kg body weight per day x 70 kg x 0.2 = 0. 0791 mg/L (rounded to 0.08 mg/L) 2 L x 1000

where:

• no observable adverse effect level = 11.3 mg/kg body weight per day for decreased body weight in male F344 rats in a 90-day drinking-water study




• uncertainty factor = 1000; 100 for intra- and interspecies variation and 10 for the short duration of the study.





USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water: Supplement 1. Report No EPA/600/4-90-020.

USEPA Method 551. 1990. Determination of Chlorination Disinfection By-products and Chlorinated Solvents in Drinking Water by Liquid�Liquid Extraction and Gas Chromatography with Electron Capture Detection. Cincinnati, USA: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).

WHO. 2003. Halogenated Acetonitriles in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/98).


Dibromochloromethane Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of dibromochloromethane in drinking-water should not exceed 0.15 mg/L. Action to reduce THMs is encouraged, but must not compromise disinfection, as non-disinfected water poses significantly greater risk than THMs. The maximum contaminant level for total trihalomethanes (USEPA 2004) is 0.08 mg/L.


1 To source waters

Brominated trihalomethanes, such as dibromochloromethane, may occur in raw water as an industrial contaminant and from human activity. They have been used as laboratory reagents, chemical intermediates, as fluids for mineral ore separation, as solvents for waxes, fats, and resins, and as flame retardants.


Trihalomethanes, including dibromochloromethane, are most likely to be formed as by-products of the chlorination of drinking-water. Naturally-occurring bromide is oxidised by chlorine to form bromine (hypobromous acid and hypobromite ion) which reacts with organic matter, such as humic and fulvic acids, in the water, with the result that the trihalomethanes found in the water show varying degrees of bromine incorporation. When full bromine substitution occurs, bromoform is produced; mixed substitution of chlorine and bromine results in dibromochloromethane and bromodichloromethane. The concentration of trihalomethanes produced depend upon: pH, organic matter concentration, chlorine dose, bromide concentration, contact time and temperature.


No known sources.


Form and fate in the environment In air, brominated trihalomethanes may be removed through oxidation with atmospheric hydroxyl radicals. Volatilisation is a major removal mechanism for bromodichloromethane from water. Biodegradation occurs under anaerobic conditions. Hydrolysis is extremely low. Bioaccumulation in aquatic organisms may occur. Brominated trihalomethanes are expected to be mobile in soil.

Typical concentrations in drinking-water The 1992 review of organic contaminants in New Zealand drinking-water supplies from 1987�1992 contained dibromochloromethane results from 370 samples representing 157 supplies. Dibromochloromethane was detected in 242 samples at concentrations ranging from 0.00023�0.029 mg/L. The P2 Chemical Determinand Identification Programme, sampled from 511 zones, found dibromochloromethane concentrations to range from �not detectable� (nd) to 0.028 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L).

Removal methods Dibromochloromethane present in contaminated source waters can be removed by adsorption on to granular activated carbon, or by air stripping. Adsorption efficiency increases, and air stripping efficiency decreases, with increasing bromine substitution in the trihalomethane. However, as dibromochloromethane arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Trihalomethane concentrations in chlorinated water increase with increasing pH. Concentrations in the finished water can therefore be reduced by ensuring that high pH levels are not present once the water is chlorinated. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Dibromochloromethane can be removed by adsorption on to granular activated carbon, or by air stripping. Adsorption efficiency increases, and air stripping efficiency decreases, with increasing bromine substitution in the trihalomethane. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.


Analytical methods

Referee method




Health considerations Available studies indicate that gastro-intestinal absorption is high for all trihalomethanes and because of their high lipophilicity, accumulation is higher in tissues with high lipid content, including body fat, liver and kidneys. In a 90-day study in rats administered dibromochloromethane in drinking-water, mild to moderate histological changes in the liver and thyroid, and a significant increase in the severity of hepatic lesions, were observed at the highest dose. Results of studies on the genotoxicity of trihalomethanes in bacteria have been inconsistent, with most reporting negative results. The International Agency for Research on Cancer has classified dibromochloromethane in Group 3 (not classifiable as to its carcinogenicity to humans).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV of dibromochloromethane in drinking-water. The no observable adverse effects level used for the derivation of the MAV was determined for the absence of histopathological effects in the liver in a well-conducted and well-documented 90-day study in rats. The MAV for dibromochloromethane in drinking-water was derived as follows:

30 x (5/7) mg/kg body weight per day x 70 kg x 0.2 = 0.15 mg/L 2 L x 1000

where:

• no observable adverse effect level = 30 mg/kg body weight per day based on the absence of histopathological lesions in the liver in a 90-day study on rats (normalised for 5 days/week dosing in derivation)




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the short duration of the study). An additional uncertainty factor for potential carcinogenicity was not applied because of the questions regarding mice liver tumours from corn oil vehicles and inconclusive evidence of genotoxicity.











Dichloroacetic acid Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of dichloroacetic acid in drinking-water should not exceed 0.05 mg/L. The WHO 2004 guideline value was designated as provisional because the data were insufficient to ensure that the value was technically achievable. The maximum contaminant level for the five haloacetic acids (USEPA 2004) is 0.06 mg/L.


1 To source waters

Dichloroacetic acid may occur in raw water as an industrial contaminant and through human activity. It is used as a chemical intermediate in the synthesis of organic materials, as an ingredient in pharmaceuticals and medicines, as a topical astringent, and as a fungicide.


Chlorinated acetic acids are formed from organic material during water chlorination.


No known sources.

Form and fate in the environment No information available.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 488 zones, found dichloroacetic acid concentrations to range from �not detectable� (nd) to 0.072 mg/L, with the median concentration being �nd� (limit of detection = 0.005 mg/L).

Removal methods No information is available for methods to remove dichloroacetic acid from contaminated source waters. As dichloroacetic acid arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. There is no information available as to how the chlorinated acetic acids may be removed from water, once formed.

Analytical methods

Referee method

Ion Exchange Liquid�Solid Extraction and Gas Chromatography with Electron Capture Detection (EPA 552.1).


1 Micro Liquid�Liquid Extraction Gas Chromatographic Method (APHA 6233B).

Health considerations Studies in rats suggest rapid intestinal absorption of dichloroacetic acid with levels in the liver and muscles increasing following administration. In humans, the average half-life of the parent compound in the plasma was 0.43 hours. Urinary excretion of unchanged dichloroacetate was negligible after eight hours, and cumulative excretion was less than 1% of the total dose in all subjects. In several bioassays, dichloroacetate has induced hepatic tumours in mice. Data on genotoxicity are inadequate. Diabetic patients treated orally with 3,4-dichloroacetate for 6�7 days experienced mild sedation, but no other clinical evidence of adverse effects was noted during, or immediately after treatment. Biochemical effects included significantly reduced fasting blood glucose levels, decreases in plasma lactate and alanine, decreased plasma cholesterol levels, decreased triglyceride levels, elevated plasma ketone bodies and elevated serum uric acid levels.


Dichloroacetate has been used to treat severe familial hypercholesterolaemia, resulting in significantly decreased total serum cholesterol levels. However, one patient experienced diminished deep tendon reflexes and decreased strength in all muscle groups of the lower extremities. After six months these conditions had improved, although serum cholesterol levels returned to their high levels.

Derivation of maximum acceptable value Because the evidence for the carcinogenicity of dichloroacetate is insufficient, a tolerable daily intake approach was used for the derivation of the MAV for dichloroacetate in drinking-water. A 75-week study in mice identified a no observable adverse effects level based on the absence of effects on the liver. The MAV for dichloroacetic acid in drinking-water was derived as follows:

7.6 mg/kg body weight per day x 70 kg x 0.2 = 0.05 mg/L 2 L x 1000

where:

• no observable adverse effect level = 7.6 mg/kg body weight per day based on a study in which no adverse effects were seen on the livers of mice exposed to dichloroacetate for 75 weeks




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for possible carcinogenicity).




USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water: Supplement 2. Report No EPA/600/R-92129.

USEPA Draft Method 552. 1990. Determination of Haloacetic Acids in Drinking Water by Liquid�Liquid Extraction, Derivatization, and Gas Chromatography with Electron Capture Detection. Cincinnati, USA: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).


WHO. 2003. Dichloroacetic Acid in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/121).



Dichloroacetonitrile Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of dichloroacetonitrile in drinking-water should not exceed 0.02 mg/L. The guideline value for dichloroacetonitrile is provisional due to limitations of the toxicological database (WHO 2004).


1 To source waters

No known sources.


Halogenated acetonitriles are produced during water chlorination or chloramination from naturally occurring substances, including algae, fulvic acid and proteinaceous material. In general, increasing temperature and/or decreasing pH have been associated with increasing concentrations of halogenated acetonitriles. Ambient bromide levels appear to influence, to some degree, the speciation of halogenated acetonitrile compounds. Dichloroacetonitrile is by far the most predominant halogenated acetonitrile species detected in drinking-water.


No known sources.

Forms and fate in the environment Haloacetonitriles are reported to undergo hydrolysis in water, yielding nonvolatile products.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 209 zones, found dichloroacetonitrile concentrations to range from �not detectable� (nd) to 0.0079 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L).

Removal methods As dichloroacetonitrile arises in waters as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.


Analytical methods

Referee method



No alternative methods have been recommended for dichloroacetonitrile because no methods meet the required criteria.

Health considerations Dichloroacetonitrile is well-absorbed from the gastrointestinal tract. Most is excreted in urine, with smaller amounts being eliminated in expired air and faeces. Its metabolites are detected in highest concentrations in liver, blood, muscle and skin. It may be formed in vivo following ingestion of chlorinated water. Studies of pregnant rats administered high levels of dichloroacetonitrile during gestation report significantly increased foetal resorptions and decreased foetal weight and size. Malformations of the cardiovascular, digestive and urogenital systems were observed. Assay results indicated that dichloroacetonitrile is mutagenic. The International Agency for Research on Cancer has concluded that dichloroacetonitrile is not classifiable as to its carcinogenicity to humans.

Derivation of maximum acceptable value Due to the lack of evidence of the carcinogenicity of dichloroacetonitrile, a tolerable daily intake approach has been used for the derivation of the MAV for dichloroacetonitrile in drinking-water. A lowest observable adverse effect level determined for adverse effects in a study in rats has been used as the basis of the derivation. The MAV for dichloroacetonitrile in drinking-water was derived as follows:


where:

• lowest observable adverse effect level = 8 mg/kg body weight per day for increased relative liver weight in male and female rats in a 90-day study




• uncertainty factor = 3000 (taking into consideration intra- and interspecies variation, the short duration of the study, the use of a minimal LOAEL and database deficiencies).

The MAV is designated as provisional because of the limitations of the database (ie, lack of long-term toxicity and carcinogenicity bioassays.








1,2-dichlorobenzene Revised July 2005. (Also called ortho-dichlorobenzene.)

Maximum acceptable value Based on health considerations, the concentration of 1,2-dichlorobenzene should not exceed 1.5 mg/L in drinking-water. The maximum contaminant level (USEPA 2004) is 0.6 mg/L.


1 To source waters

Dichlorobenzenes are used widely in industry and domestic products and therefore are likely to occur in source waters from industrial contamination and human activity. 1,2-dichlorobenzene is used primarily as a chemical intermediate for dyestuffs and pesticides.


No known sources.


No known sources.

Forms and fate in the environment The dichlorobenzenes are expected to adsorb readily to soils with high organic content and are not expected to leach appreciably into groundwater. In water the major processes for dichlorobenzene removal are likely to be adsorption to sediments and bioaccumulation in aquatic organisms. Evaporation from surface water may also be important. Dichlorobenzenes may be biodegraded under aerobic conditions but this is not likely to occur under the anaerobic conditions that may exist in lake sediments or groundwaters.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 294 zones, did not find any 1,2-dichlorobenzene at detectable concentrations (limit of detection = 0.0005 mg/L).


Removal methods Removal of 1,2-dichlorobenzene can be achieved by adsorption on to granular activated carbon, or air stripping.

Analytical methods

Referee method



1 Purge and Trap Capillary-Column Gas Chromatographic Method (APHA 6230D, EPA 502.2). 2 Liquid�Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (APHA 6410B).

Health considerations Sources of human exposure are predominantly air and food. Dichlorobenzenes are absorbed almost completely from the gastro-intestinal tract and distributed primarily to fat and fatty tissue because of their lipophilicity (solubility in fat) and to kidney, liver and lungs. They are metabolised rapidly by oxidation in the liver to chlorophenols and excreted in urine. Dichlorobenzenes are of low acute oral toxicity in experimental animals. The major target organs are the kidney and liver. Data concerning the health effects of exposure of humans to dichlorobenzenes are restricted to case reports of accidental exposure or misuse of the products. Reported acute effects following short-term exposure (all of which are reversible) include liver damage, blood disorders and disturbances to the immune system, the central nervous system, or the respiratory tract. Skin pigmentation and allergic dermatitis have followed skin contact. The balance of evidence suggests that 1,2-dichlorobenzene is not mutagenic in tests with bacteria, and there is no evidence for its carcinogenicity in rodents. The International Agency for Research on Cancer consider that 1,2-dichlorobenzene is not classifiable as to its carcinogenicity to humans (Group 3).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for 1,2-dichlorobenzene in drinking-water. The MAV has been derived on the basis of a no-observable-no-adverse effects level determined in a two-year mouse gavage study. The MAV for 1,2-dichlorobenzene in drinking-water was derived as follows:

60 x (5/7) mg/kg body weight/day x 70 kg x 0.1 = 1.5 mg/L 2 L/day x 100

where:

• no observable adverse effect level = 60 mg/kg body weight per day for tubular regeneration of the kidney, identified in a two-year mouse gavage study (normalised for five days/week dosing in the derivation)

• average weight of adult = 70 kg

• average quantity of water consumed by an adult = 2 L per day



• uncertainty factor = 100 (for intra- and interspecies variation). The taste threshold for 1,2-dichlorobenzene has been reported at 0.001 mg/L, and from 0.002 to 0.01 mg/L for odour.






WHO. 2003. Dichlorobenzenes in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva,World Health Organization (WHO/SDE/WSH/03.04/28).


1,3-dichlorobenzene Revised July 2005. (Also called meta-dichlorobenzene.)

Maximum acceptable value There are insufficient data to derive a health based MAV for 1,3-dichlorobenzene in drinking-water.


1 To source waters

Dichlorobenzenes are used widely in industry and domestic products and therefore are likely to occur in source waters from industrial contamination and human activity. 1,3-dichlorobenzene is a minor fumigant and insecticide and can be formed from incomplete combustion of waste.


No known sources.


No known sources.

Forms and fate in the environment The dichlorobenzenes are expected to adsorb readily to soils with high organic content and are not expected to leach appreciably into groundwater. In water, the major processes for dichlorobenzene removal are likely to be adsorption to sediments and bioaccumulation in aquatic organisms. Evaporation from surface water may also be important. Dichlorobenzenes may be biodegraded under aerobic


conditions but this is not likely to occur under the anaerobic conditions that may exist in lake sediments or groundwaters.

Typical concentrations in drinking-water No data are available on the concentration of 1,3-dichlorobenzene in New Zealand drinking-water supplies. In a US study of 685 groundwaters, 1,3-dichlorobenzene was detectable in 19 samples, with a maximum concentration of 0.237 mg/L being found.

Removal methods Removal of 1,3-dichlorobenzene can be achieved through adsorption on to activated carbon, or air stripping.

Analytical methods

Referee method

A referee method cannot be selected for 1,3-dichlorobenzene because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for 1,3-dichlorobenzene because a MAV has not been established. However, the following methods are used to analyse for 1,3-dichlorobenzene:

1 Purge and Trap Capillary Column Gas Chromatographic/Mass Spectrometric Method (APHA 6210D, EPA 524.2).


3 Liquid�Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (APHA 6410B).

Health considerations Sources of human exposure are predominantly air and food. Dichlorobenzenes are absorbed almost completely from the gastro-intestinal tract. Once ingested they are absorbed rapidly, primarily to fat and adipose tissue because of their lipophilicity, and to kidney, liver and lungs. They are rapidly metabolised by oxidation in the liver and excreted in urine. Dichlorobenzenes are of low acute oral toxicity in experimental animals, however no data are available on chronic toxicity for 1,3-dichlorobenzene. The major target organs are the kidney and liver. Data concerning the health effects of exposure of humans to dichlorobenzenes are restricted to case reports of accidental exposure or misuse of the products. Reported acute effects following short-term exposure (all of which are reversible) include liver damage, blood disorders and disturbances to the immune system, the central nervous system or the respiratory tract. Skin pigmentation and allergic dermatitis have followed skin contact. 1,3-dichlorobenzene shows no mutagenic activity in tests with bacteria.

Derivation of maximum acceptable value There are insufficient toxicological data to derive a MAV for this compound, but it should be noted that it is rarely found in drinking-water, and much less frequently than the other dichlorobenzenes.


The Australian Drinking-water Guidelines have set a maximum concentration guideline of 0.02 mg/L based on the aesthetic considerations of taste and odour.






1,4-dichlorobenzene Revised July 2005. (Also called para-dichlorobenzene.)

Maximum acceptable value Based on health considerations, the concentration of 1,4-dichlorobenzene in drinking-water should not exceed 0.4 mg/L. The maximum contaminant level (USEPA 2004) is 0.075 mg/L.


1 To source waters

Dichlorobenzenes are widely used in industry and domestic products and therefore are likely to occur in source waters from industrial contamination and human activity. 1,4-dichlorobenzene is often used in toilet blocks to deodorize air, and as a moth repellant.


No known sources.


No known sources.

Forms and fate in the environment The dichlorobenzenes are expected to adsorb readily to soils with high organic content and are not expected to leach appreciably into groundwater. In water, the major processes for dichlorobenzene removal are likely to be adsorption to sediments and bioaccumulation in aquatic organisms. Evaporation from surface water may also be important. Dichlorobenzenes may be biodegraded under aerobic conditions but this is not likely to occur under the anaerobic conditions that may exist in lake sediments or groundwaters.

Typical concentrations in drinking-water No data are available on the concentration of 1,4-dichlorobenzene in New Zealand drinking-water supplies. In a US study of 685 groundwaters, 1,4-dichlorobenzene was detected in 19 samples with a maximum concentration of 0.996 mg/L being reported.


Removal methods Removal of 1,4-dichlorobenzene can be achieved through adsorption on to activated carbon or air stripping. Good destruction by ozone has been reported.

Analytical methods

Referee method



1 Purge and Trap Capillary-Column Gas Chromatographic Method (APHA 6230D, EPA 502.2). 2 Liquid�Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (APHA 6410B).

Health considerations Sources of human exposure are predominantly air and food. Dichlorobenzenes are almost completely absorbed from the gastro-intestinal tract. Once ingested, they are absorbed rapidly, primarily to fat and adipose tissue because of their lipophilicity, and to kidney, liver and lungs. They are metabolised rapidly by oxidation in the liver and excreted in urine. Dichlorobenzenes are of low toxicity and the main target organs are the liver and kidneys. There is evidence that 1,4-dichlorobenzene increases the incidence of renal tumours in rats and the hepatocellular adenomas and carcinomas in mice after long-term exposure. The International Agency for Research on Cancer has placed 1,4-dichlorobenzene in Group 2B (possibly carcinogenic to humans). Data concerning the health effects of exposure of humans to dichlorobenzenes are restricted to case reports of accidental exposure or misuse of the products. Reported acute effects following short-term exposure (all of which are reversible) include liver damage, blood disorders and disturbances to the immune system, the central nervous system, or the respiratory tract. Skin pigmentation and allergic dermatitis have followed skin contact. 1,4-dichlorobenzene does not exhibit mutagenic activity in tests with bacteria or mammalian cells, and the relevance for humans of the tumours observed in animals is doubtful.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for 1,4-dichlorobenzene. It has been derived on the basis of a lowest-observable-no-adverse effects level determined in a two-year toxicological study in which rats were exposed to 1,4-dichloroethene by oral administration. The MAV for 1,4-dichlorobenzene in drinking-water was derived as follows:

150 x (5/7) mg/kg body weight/day x 70 kg x 0.1 = 0.375 mg/L (rounded to 0.4 mg/L) 2 L/day x 1000

where:

• lowest observable adverse effect level = 150 mg/kg body weight per day for kidney effects observed in a two-year rat gavage (normalised for five days/week dosing in derivation)





• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the use of a LOAEL in place of a NOAEL and because the toxic end-point is carcinogenicity.

Taste and odour thresholds for 1,4-dichlorobenzene have been reported at concentrations of 0.006 and 0.0003 mg/L respectively.



AWWA/Lyonnaise Des Eaux. 1987. Tastes and Odours in Drinking Water.






Revised July 2005.

1 To source waters

1,1-dichloroethane

Maximum acceptable value There are insufficient data to derive a MAV for 1,1-dichloroethane in drinking-water.


1,1-dichloroethane can be released to the aquatic environment as an industrial contaminant. It is used overseas in the commercial production of 1,1,1-trichloroethane and vinyl chloride, as a solvent in paints, and as a varnish and finish remover. It was used formerly as an anaesthetic. Occurrence in New Zealand waters is expected to be relatively low since vinyl chloride is not manufactured here, and 1,1-dichloroethane has not been identified, from a limited survey, as a solvent used by paint manufacturers in New Zealand.


No known sources.


No known sources.


Forms and fate in the environment 1,1-dichloroethane is volatile, so most of that released to the environment partitions to the atmosphere, where it is removed by photo-oxidation. Biodegradation is not expected to be significant in aquatic systems.

Typical concentrations in drinking-water No data are available on the concentration of 1,1-dichloroethane in New Zealand drinking-water supplies. Levels as high as 0.004 mg/L (4 µg/L) have been found in supplies in the USA.

Removal methods Some removal of 1,1-dichloroethane can be achieved by adsorption on to granular activated carbon, although the adsorption is relatively weak. Some removal is also achievable by air-stripping.

Analytical methods

Referee method

A referee method cannot be selected for 1,1-dichloroethane because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for 1,1-dichloroethane because a MAV has not been established. However, the following methods are used to analyse for 1,1-dichloroethane:

1. Purge and Trap Capillary Column Gas Chromatographic/Mass Spectrometric Method (APHA 6210D, EPA 524.2).

2. Purge and Trap Capillary-Column Gas Chromatographic Method (APHA 6230D, EPA 502.2).

Health considerations 1,1-dichloroethane is metabolised rapidly by mammals to acetic acid and a variety of chlorinated compounds. It is of relatively low acute toxicity, and limited data are available on its toxicity from short- and long-term studies. A 13-week inhalation study with 1,1-dichloroethane reported elevated blood-urea nitrogen concentrations in cats, but not in rats, rabbits or guinea pigs. No other adverse effects were observed. A 78-week feeding study reported a marginally significant increase in the incidence of tumours of the mammary glands of female rats. No statistically significant increase in tumours was observed in male rats, or male and female mice. 1,1-dichloroethane has exhibited mutagenic activity in tests with bacteria and mammalian cells. One carcinogenicity study in mice and rats provided no conclusive evidence of carcinogenicity, although there was some evidence for an increased incidence of haemangiosarcomas in treated animals. In humans, inhalation exposures to high concentrations of 1,1-dichloroethane causes central nervous system depression. It has been used as an anaesthetic until its use was discontinued because of the problems associated with heart rhythm.

Derivation of maximum acceptable value In view of the limited data base on toxicity and carcinogenicity no MAV for 1,1-dichloroethane in drinking-water is proposed.





AWWA/Lyonnaise Des Eaux. 1987. Tastes and Odours in Drinking Water.


WHO. 2003. 1,1-Dichloroethane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/19).


1,2-dichloroethane Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of 1,2-dichloroethane in drinking-water should not exceed 0.03 mg/L. The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


1 To source waters

1,2-dichloroethane can be released to the aquatic environment as a result of human and industrial activity. Overseas it is used mostly in the commercial production of vinyl chloride for use in the plastics industry, and for the manufacture of other chlorinated chemicals. Occurrence from these sources is likely to be low in New Zealand since all PVC is imported and only compounded here. It is also used as a solvent, and can be used as a lead scavenger in leaded petrol.


Chloroethanes may be formed in small amounts by the aqueous chlorination of effluents.


No known sources.

Forms and fate in the environment 1,2-dichloroethane is volatile, so most of that released to the environment partitions to the atmosphere. Biodegradation is not expected to be significant in aquatic systems. 1,2-dichloroethane may persist for long periods in groundwater, where volatiliation is restricted.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, found 1,2-dichloroethane concentrations to range from �not detectable� (nd) to 0.009 mg/L, with the median concentration being �nd� (limit of detection = 0.0005 mg/L).


Removal methods Some removal of 1,2-dichloroethane can be achieved by adsorption on to granular activated carbon, although the adsorption is relatively weak. Some removal is also achievable by air-stripping.

Analytical methods

Referee method


1,2-dichloroethane is absorbed readily through the lungs, skin and gastro-intestinal tract. It appears to be distributed readily following oral or inhalation exposure, with accumulation in the liver and kidneys. 1,2-dichloroethane crosses the blood-brain barrier and the placenta and it has been detected in human milk following occupational exposure. Excretion of absorbed 1,2-dichloroethane occurs rapidly, mainly in the urine and expired air.




In humans, acute oral exposure to 1,2-dichloroethane is reported to cause central nervous system, hepatic, gastro-intestinal, respiratory, renal and cardiovascular effects. Death following acute intoxication is most often attributed to cardiovascular or respiratory failure. Repeated inhalation exposures in the workplace result in anorexia, nausea, vomiting, weakness and fatigue, nervousness, epigastric pain and irritation of the respiratory tract and eyes.

1,2-dichloroethane has exhibited mutagenic activity in tests with different strains of bacteria, and its metabolites are known to be strongly mutagenic. A 13-week feeding and drinking study with 1,2-dichloroethane using rats and mice reported increased kidney and liver weights at high doses (4000 mg/L). No increase in the incidence of tumours or lesions was observed in mice or rats, but female rats exhibited an increase in the incidence of kidney lesions. Male rats fed 1,2-dichloroethane five times per week for 78 weeks were reported to have a significant increase in tumours of the forestomach and circulatory system. The same study reported tumours of the mammary glands in female rats.

The International Agency for Research on Cancer has classified 1,2-dichloroethane in Group 2B (possibly carcinogenic to humans). It has been shown to produce statistically significant increases in a number of tumour types in laboratory animals, including the relatively rare haemangiosarcoma, and the balance of evidence indicates that it is relatively genotoxic.

Derivation of maximum acceptable value As the balance of evidence indicates that 1,2-dichloroethane is potentially genotoxic and there are no suitable long-term studies on which to base a tolerable daily intake, the MAV for 1,2-dichloroethane in drinking-water was calculated using an extrapolation model. On the basis of haemangiosarcomas observed in male rats in a 78-week gavage study, and applying the linearised multistage model, it was determined that a concentration of 1,2-dichloroethane in drinking-water of 0.03 mg/L corresponds to a life-time risk of one additional cancer per 100,000 (10-5).






IPCS. 1995. 1,2-dichloroethane (2nd ed). Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 176).

IPCS. 1998. 1,2-dichloroethane. Geneva: World Health Organization, International Programme on Chemical Safety (Concise International Chemical Assessment Document 1).



WHO. 2003. 1,2-dichloroethane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/67).


1,1-dichloroethene Revised July 2005. (Also called vinylidene chloride.)

Maximum acceptable value Based on health considerations, the concentration of 1,1-dichloroethene in drinking-water should not exceed 0.03 mg/L. The maximum contaminant level (USEPA 2004) is 0.007 mg/L.


1 To source waters

1,1-Dichloroethene can be released to the aquatic environment as an industrial contaminant. It is used as a monomer to produce polyvinylidene chloride polymer plastics such as �Saran� and as an intermediate in the synthesis of other organic solvents. Potential for occurrence in New Zealand source waters is expected to be low since there is no polymerisation of the monomer in this country.


No known sources.


No known sources.

Forms and fate in the environment Most 1,1-dichloroethene released to the environment volatilises to the atmosphere, where it is oxidised rapidly. Photolysis is also expected to occur. Volatilisation is the major removal mechanism for


1,1-dichloroethene in surface waters and soils, and anaerobic biotransformation to vinyl chloride is expected to be important in groundwater.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find any 1,2-dichloroethene at detectable concentrations (limit of detection = 0.005 mg/L). Available overseas data indicate that the dichloroethenes are rarely found in drinking water. Studies in the United States have very occasionally reported DCEs in groundwater, usually from wells heavily contaminated with other chlorinated solvents.

Removal methods Some removal of 1,1-dichloroethene can be achieved by adsorption on to granular activated carbon, although the adsorption is relatively weak. Some removal is also achievable by air-stripping.

Analytical methods

Referee method





It is a central nervous system depressant and may cause liver and kidney toxicity in occupationally exposed humans. It causes liver and kidney injury in laboratory animals.

Following oral or inhalation exposure, 1,1-dichloroethene is absorbed almost completely, metabolised extensively, and excreted rapidly.

A long-term study, in which rats were exposed to 1,1-dichloroethene in their drinking-water for two years, reported minimal swelling to liver cells, but no other adverse effects. No changes were observed in tissues taken from dogs after 97 days of exposure. IARC has placed 1,1-dichloroethene in Group 3. It was found to be genotoxic in a number of test systems in vitro but was not active in the dominant lethal and micronucleus assays in vivo. It induced kidney tumours in mice in one inhalation study but was reported not to be carcinogenic in a number of other studies, including several in which it was given in drinking-water.

Derivation of maximum acceptable value As 1,1-dichloroethene is not classifiable as to its carcinogenicity to humans, a tolerable daily intake approach has been used for the derivation of the MAV. The MAV for 1,1-dichloroethene in drinking-water has been derived on the basis of a lowest-observable-adverse effects level determined in a two-year toxicological study in which rats were exposed to 1,1-dichloroethene in drinking-water. The MAV for 1,1-dichloroethene was derived as follows:

9 mg/kg body weight/day x 70 kg x 0.1 = 0.03 mg/L 2 L/day x 1000


where:

• lowest observable adverse effect level = 9 mg/kg body weight per day in a two-year drinking-water study (for increased incidence of hepatic lesions) in female rats




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the use of a LOAEL in place of a NOAEL and the potential for carcinogenicity).







WHO. 2003. 1,1-dichloroethene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/20).


1,2-dichloroethene Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of 1,2-dichloroethene in drinking-water should not exceed 0.06 mg/L. The maximum contaminant level (USEPA 2004) is 0.07 mg/L for the cis form and 0.1 mg/L for the trans form.


1 To source waters

1,2-dichloroethene (cis-, and trans- isomers) can be released to the aquatic environment as an industrial contaminant. It is used primarily as an intermediate in the synthesis of chlorinated solvents and compounds. It has also been used as an extraction solvent. The cis-form of 1,2-dichloroethene is more frequently found as a water contaminant. The presence of the two isomers, which are metabolites of other unsaturated halogenated hydrocarbons, may indicate the simultaneous presence of more toxic organochlorine chemicals, such as vinyl chloride. Accordingly, their presence indicates that more intensive monitoring should be conducted.


No known sources.



No known sources.

Forms and fate in the environment Removal of 1,2-dichloroethene from air is mainly through reaction with hydroxyl radicals. Volatilisation is expected to be the primary fate process in surface water and surface soils. It may leach through subsurface soils to groundwater. Anaerobic biodegradation may remove both isomers from groundwater.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find any 1,2-dichloroethene at detectable concentrations (limit of detection = 0.005 mg/L). Overseas it has been found in drinking-water supplies derived from groundwater at levels up to 0.12 mg/L (WHO 2004).

Removal methods Removal of 1,2-dichloroethene can be achieved by adsorption on to granular activated carbon, or by air stripping.

Analytical methods

Referee method




Health considerations There is little information on the adsorption, distribution, and excretion of 1,2-dichloroethene. However, based on analogy with 1,1-dichloroethene, it would be expected to be adsorbed readily, distributed mainly in the liver, kidney, and lung, and excreted rapidly. The cis-isomer is metabolised more readily than the trans-isomer in in vitro systems. Both isomers have been reported to cause increased serum alkaline phosphatase levels in rodents. In a three-month study in mice given the trans-isomer in drinking-water, there was a reported increase in serum alkaline phosphatase and reduced thymus and lung weights. It was also reported to cause transient immunological effects, the toxicological significance of which is unclear. Trans-1,2-dichloroethene also caused reduced kidney weights in rats, but at higher doses. Only one rat toxicity study is available for the cis-isomer. The toxic effects of the cis-isomer in rats were observed only at higher doses than for the trans-isomer in mice, but the toxicity was of a similar magnitude. Inhalation of high concentrations (9500 ppm and above) of 1,2-dichloroethane by humans causes central nervous system depression. Neurological effects have been reported following exposure to low levels of trans-1,2-dichloroethane, including nausea, drowsiness, fatigue and vertigo. A burning sensation of the eyes was also reported. The trans-isomer is reportedly about twice as potent a central nervous system depressant as the cis-isomer, which was previously used as an anaesthetic.


Derivation of maximum acceptable value Data on the trans-isomer were used to calculate a joint MAV for both isomers because toxicity for the trans-isomer occurred at a lower dose than for the cis-isomer and because data suggest that the mouse is a more sensitive species than the rat. The MAV for 1,2-dichloroethene (sum of cis and trans) in drinking-water was derived as follows:

17 mg/kg body weight/day x 70 kg x 0.1 = 0.0595 mg/L (rounded to 0.06 mg/L)

2 L/day x 1000

where:

• no observable adverse effect level = 17 mg/kg body weight per day, for increases in serum alkaline phosphatase levels and increased thymus weight, based on a 90-day study in mice administered trans-1,2-dichloroethene in drinking-water




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the short duration of the study).






WHO. 2003. 1,2-dichloroethene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/72).


Dichloromethane Revised July 2005. (Also called methylene chloride.)

Maximum acceptable value Based on health considerations, the concentration of dichloromethane in drinking-water should not exceed 0.02 mg/L. The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


1 To source waters

Dichloromethane is a widely used organic solvent found in paints, insecticides, degreasing and cleaning fluids, and paint strippers. It can be released to the aquatic environment via inadequate storage or disposal, or from the discharge of wastes from industries in which it is used.



Dichloromethane may be introduced into water during chlorination, as it is a recognised impurity in commercial chlorine. Although not commonly identified as a disinfection by-product, some published work indicates that dichloromethane may also be formed during chlorination of aqueous organic material.


No known sources.

Forms and fate in the environment Most dichloromethane released to water and soil will vaporise to air, where it can persist for up to 500 days. Dichloromethane biodegrades rapidly in water. In soil it biodegrades slightly and is highly mobile, leaching to groundwater.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find any dichloromethane at detectable concentrations (limit of detection = 0.008 mg/L).

Removal methods

When its appearance in water results from chlorination, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine.

Purge and Trap Capillary Column Gas Chromatographic/Mass Spectrometric Method (APHA


Overseas it is considered to arise predominantly from industrial contamination in waters, although there is evidence that dichloromethane may be formed as a disinfection by-product.

Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina or ion exchange resins, however these methods are generally more expensive than coagulation.

Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Dichloromethane can be removed by adsorption on to granular activated carbon, or by air stripping.

Analytical methods

Referee method

6210D, EPA 524.2).




Dichloromethane appears to be absorbed readily from the gastro-intestinal tract with distribution primarily to the liver. It is metabolised to carbon monoxide, carbon dioxide and formic acid. Animal data indicate that dichloromethane is excreted primarily through the lungs.

Exposure from drinking-water is likely to be insignificant compared with that from other sources.

Dichloromethane is of low acute toxicity. The primary effect associated with acute exposure is depression of the central nervous system. In humans, inhalation of a high concentration of dichloromethane has been associated with a variety of central nervous system effects, most notably narcosis. Acute exposure to levels of 300 ppm can impair sensory and motor functions. Epidemiological studies involving occupational exposure has failed to show a positive correlation between inhalation exposure and increased cancer incidence. Dichloromethane has exhibited mutagenic activity in various test systems. An inhalation study in mice provided conclusive evidence of carcinogenicity, whereas a drinking-water study provided only suggestive evidence. The International Agency for Research on Cancer has placed dichloromethane in Group 2B (possibly carcinogenic to humans), however, the balance of evidence suggests that it is not a genotoxic carcinogen and that genotoxic metabolites are not formed in relevant amounts in vivo.

Derivation of maximum acceptable value The balance of evidence suggests that dichloromethane is not genotoxic and therefore a tolerable daily intake approach has been taken for the derivation of the MAV for dichloromethane in drinking-water. The no observable adverse effect level used was for hepatoxic effects in a two-year drinking-water study in rats. The MAV for dichloromethane in drinking-water was derived as follows:


where:

• no observable adverse effect level = 6 mg/kg body weight per day for hepatotoxic effects in a two-year drinking-water study in rats




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 reflecting concern for carcinogenic potential).







WHO. 2003. Dichloromethane in drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/18).


2,4-dichlorophenol Revised July 2005.

Maximum acceptable value There are insufficient data to set a health based MAV for 2,4-dichlorophenol in drinking-water.


1 To source waters

2,4-dichlorophenol may occur in raw water as a result of its use as a pesticide. It may be used as a mothproofing agent, germicide and antiseptic, or in the production of the pesticide 2,4-D.




No known sources.

Form and fate in the environment Because chlorinated phenols are moderately water-soluble, weakly acidic, and have low vapour pressures, it is anticipated that volatilisation does not play a significant role in removing these chemicals from water. Photolysis of dichlorophenols appears to be minimal. Sorption is not significant for dichlorophenols. Biodegradation appears to be the primary removal mechanism of chlorinated phenols from surface waters. Aquatic biota may bioconcentrate chlorinated phenols with bioconcentration factors increasing with increasing chlorine substitution.

Typical concentrations in drinking-water No data are available on the concentration of 2,4-dichlorophenol in New Zealand drinking-water supplies. A Canadian study of 40 supplies found levels up to 0.000072 mg/L (72 ng/L or 0.072 µg/L) in chlorinated supplies. Levels ranged from 0.003-0.006 mg/L (3�6 µg/L) in a German survey of the Rhine area.

Removal methods Chlorophenols can be removed from contaminated source water by adsorption on to activated carbon. The effectiveness of the process is pH dependent. Greater adsorption occurs as the pH is lowered.

However, as this compound arises in New Zealand waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine.


Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. The formation of chlorophenols can be reduced largely by the use of chlorine dioxide in place of chlorine. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Chlorophenols can be removed by adsorption on to activated carbon. The effectiveness of the processes is pH dependent. Greater adsorption occurs as the pH is lowered. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.

A referee method cannot be selected for 2,4-dichlorophenol because a MAV has not been established and therefore the sensitivity required for the referee method is not known.

Analytical methods

Referee method


No alternative methods can be recommended for 2,4-dichlorophenol because a MAV has not been established. However, the following information may be useful:

1 Chlorophenols in water can be solvent extracted with dichloromethane (Method APHA 6410) and analysed by gas chromatography with mass spectrometry detection (Method APHA 6410 or EPA 8270). The method detection limit is 0.003 mg/L (3 µg/L). Interference may come from contaminated reagents or glassware.

2 A more sensitive and specific method of analysis for chlorophenols is to solvent extract with dichloromethane and derivatise with pentafluorobenzyl ether and analyse by gas chromatography with electron capture detection (Method EPA 604 or APHA 6420). The limit of quantification for this method is 0.0007 mg/L (0.7 µg/L). The specificity of this method reduces the likelihood of interferences.

Health considerations Chlorophenols are well-absorbed after oral administration and they readily penetrate the skin. Chlorophenols do not appear to accumulate in body tissues in rats but are rapidly metabolised and eliminated from the body, principally in urine. There is a limited data base on the toxicity of 2,4-dichlorophenol. Long-term studies over two years could not determine any dose-related effects using 2,4-dichlorophenol. Limited tests have reported no evidence of mutagenicity or carcinogenicity.


Derivation of maximum acceptable value Because of the limited data base on the toxicity of 2,4-dichlorophenol, no health-based MAV has been derived. The Australian Drinking-water Guidelines have a health value of 0.2 mg/L. Taste and odour thresholds for 2,4-dichlorophenol have been reported at 0.0003 mg/L and 0.04 mg/L respectively.




Sithole BB, Willimas DT. 1985. Halogenated phenols in water at 40 Canadian potable water treatment facilities. Journal of the Association of Official Analytical Chemists 69(5): 807�10.

Forms and fate in the environment





Dioxins New entry August 2005.

Maximum acceptable value No MAV has been derived because of the large number of compounds of differing toxicities in this group. WHO (2004) does not mention dioxins. The maximum contaminant level (USEPA 2004) is 0.00000008 mg/L (8 x 10-8).


1 To source waters

Dioxins can be produced by chlorination reactions with organic matter. As a result they are present as contaminants in chlorophenoxy herbicides (WHO 1996), and the timber preservative pentachlorophenol (Severn 1980), and in the effluent from wood pulp operations when chlorine is used as a bleach. There are also releases into the environment from incineration of municipal refuse, exhaust from cars fuelled with leaded petrol, fossil fuel combustion, disposal of industrial wastes, chlorophenol wood treatment, and accidental fires involving transformers containing PCBs.

The term dioxins covers a large number of structurally related compounds, but within this group one of the most toxic and environmentally stable compounds is 2,3,7,8-tetrachlorodibenzo-1,4-dioxin (TCDD), often termed simply dioxin. TCDD is the most studied member of the dioxin family, and much of the information about the environmental fate of the dioxins is based on what is known about it.


TCDD has a very low solubility in water, and consequently, if present in water, a large percentage of it is associated with sediments and suspended matter (IPCS 1989). Decomposition by sunlight can occur, but the rate of decomposition will decrease with water depth, and may contribute little to loss of TCDD attached to bottom sediments (USEPA b). The rate of decomposition decreases as the chlorine content of the dioxin increases, and is also influenced by the pattern of chlorine substitution within the compound and is also found to be faster in dioxins containing less chlorine substitution (IPCS 1989). Some loss may take place by volatilisation, but this too is expected to be hindered markedly by adsorption to sediment (USEPA b). TCDD is generally very resistant to biodegradation (IPCS 1989). In summary, sediments and soils are the main sinks for dioxins in the environment.

Typical concentrations in drinking-water The data on dioxin concentrations in drinking waters are very limited. Exposure to these compounds in the general population probably occurs mainly through the food chain (IPCS 1989). A set of three samples were taken from the Hamilton supply because it draws its water from the Waikato River downstream of timber processing operations. The raw water sample showed detectable concentrations (of the order of pg/L). Lower concentrations were found in the two samples from the distribution system. Once toxicities were taken into account, the toxicities of the dioxins in these samples were detemined to be not greatly different from purified laboratory water (Nokes 1992).

Removal methods The USEPA has approved granular activated carbon as a means of removing dioxins from drinking waters (USEPA a). While no information about the effectiveness of the combination of chemical coagulation, flocculation, sedimentation and filtration has been found, the tendency for dioxins to adsorb to particulate matter indicates that these processes should be able to lower dioxin concentrations. Dioxin reduction can be achieved by the use of UV irradiation and ozone together in the laboratory (Vollmuth and Niessner 1997). Concentrations start to decrease rapidly after the first two minutes of treatment.

Acute poisoning

Recommended analytical techniques

Referee method

N/A.


N/A.


The USEPA found that acute exposures at concentrations greater than the US maximum contaminant level (MCL) of 3 x 10-8 mg/L can cause liver damage, weight loss, atrophy of the thymus gland and immunosuppression. Acute effects also include persistent chloracne (IPCS 1989).

For short-term exposure, the USEPA considers a �safe� level for a 10 kg child consuming 1 litre of water per day to be 1 x 10-6 mg/L for a one-day exposure, and 1 x 10-7 mg/L for a 10-day exposure.


Chronic exposure

Dioxin has the potential to cause a number of reproductive health effects through long-term exposure, ranging from reduced fertility to birth defects. There is also some evidence for dioxins being capable of causing cancer at concentrations more than the USEPA�s MCL. IARC (International Agency for Research into Cancer) has concluded that TCDD is a human carcinogen, but that it is presently impossible to classify other dioxins with respect to human carcinogenicity (IARC 1997).

Derivation of maximum acceptable value There is inadequate data to enable derivation of a MAV or MAVs.

References IARC. 1997. Polychlorinated dibenzo-para-dioxins. International Agency for Research into Cancer Summary and Evaluation 69. World Health Organization, available on http://www.inchem.org/documents/iarc/iarc/iarc817.htm

IPCS. 1989. Polychlorinated dibenzo-para-dioxins and dibenzofurans. Environmental Health Criteria Monograph 88. Geneva: World Health Organization, International Programme on Chemical Safety.

Nokes CJ. 1992. Organic Contaminants in New Zealand Potable Water Supplies: A review of data from samples taken between May 1987 and May 1991. Report to the Department of Health.

Severn DJ. 1980. Assessment of human exposure and body burdens of chlorination by-products. In: Jolley RL, Brungs WA, Cumming RB, et al (eds). Water Chlorination: Environmental Impact and Health Effects 3. Ann Arbor, USA: Ann Arbor Science.

USEPA(a). National Primary Drinking Water Regulations, Consumer Factsheet on: DIOXIN (2,3,7,8-TCDD). Available on http://www.epa.gov/safewater/dwh/c-soc/dioxin.html

USEPA(b). National Primary Drinking Water Regulations, Technical Factsheet on: DIOXIN (2,3,7,8-TCDD). Available on http://www.epa.gov/OGDW/dwh/t-soc/dioxin.html


Vollmuht S, Niessner R. 1997. Degradation of polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans during the UV/ozone treatment of pentachlorophenol-containing water. Toxicological and Environmental Chemistry 61(1�4): 27�41.

WHO. 1996. Guidelines for Drinking-water Quality (2nd ed). Vol 2: Health criteria and other supporting information. Geneva: World Health Organization.

EDTA Revised July 2005. (Also called edetic acid or ethylenediamine tetraacetic acid.)

Maximum acceptable value Based on health considerations, the concentration of EDTA in drinking-water should not exceed 0.7 mg/L.


1 To source waters

EDTA may enter source water as an industrial contaminant or through agricultural activities. It is used widely in many industrial processes, in agriculture, in domestic products, including food additives, and in


http://www.epa.gov/OGDW/dwh/t-soc/dioxin.html

drugs for chelation therapy; it is the drug of choice to treat lead poisoning in humans and domestic animals. EDTA is also used in laundry detergents, cosmetics, photochemicals, water softening, electroplating, and in the production of textiles and paper.


No known sources.


No known sources.

Form and fate in the environment EDTA is only poorly degraded in the aquatic environment. It exists in the environment as metal complexes with the potential of mobilising potentially toxic heavy metals in water, thus increasing their concentration in water supplies. From overseas data this is not likely to be of health concern for the concentrations of heavy metals and EDTA found.




No data are available on the concentration of EDTA in New Zealand drinking-water supplies. Overseas data show EDTA to be present in surface waters generally at concentrations below 0.07 mg/L, although higher concentrations (0.9 mg/L) have been measured; detected in drinking-water prepared from surface waters at concentrations of 0.01�0.03 mg/L.

Removal methods No information is available on processes that remove EDTA from drinking-waters, although it could possibly be oxidised by ozone.

Analytical methods

Referee method

Reverse Phase Ion Pair Liquid Chromatography (Bergers and de Groot 1994).

No alternative methods have been recommended for EDTA because no methods meet the required criteria.

CaNa2EDTA is poorly absorbed from the gut. It is medically inert, and no accumulation in the body has been found. The long-term toxicity of EDTA is complicated by its ability to chelate essential and toxic metals, both in water and in animals. Toxicity data are therefore equivocal and difficult to interpret. Long-term feeding studies in rats and dogs gave no evidence of interference with mineral metabolism in either species. Adverse effects on mineral metabolism and nephrotoxicity was seen only after parenteral administration of high doses. High doses of EDTA tested on animals in the USA did not reveal any carcinogenicity.


A vast clinical experience with respect to the use of EDTA in the treatment of metal poisoning has demonstrated its safety in humans. The major human health problem due to oral exposure to EDTA appears to be zinc deficiency as a consequence of zinc complexed by EDTA. It has also been suggested that EDTA may enter kidney cells and, by interfering with zinc metabolism, exacerbate the toxicity of cadmium. Concern has been expressed over the ability of EDTA to complex, and therefore reduce the availability of, zinc. However, this is of significance only at elevated doses substantially in excess of those encountered in the environment.

Derivation of maximum acceptable value In 1973, the Joint FAO/WHO Expert Committee on Food Additives (JECFA) proposed an Allowable Daily Intake for calcium disodium edetate (CaNa2EDTA) as a food additive of 2.5 mg/kg body weight (1.9 mg/kg as the free acid). However, JECFA recommended that no sodium edetate should remain in food. The MAV was derived for EDTA (as the free acid) as follows:

1.9 mg/kg body weight per day x 70 kg x 0.01 = 0.665 mg/L (rounded to 0.7 mg/L) 2 L

where: • allowable daily intake = 1.9 mg/kg • average body weight = 70 kg • allocation of allowable daily intake to drinking-water = 0.01 • average quantity of water consumed per day = 2 L.


Bergers PJM, de Groot AC. 1994. The analysis of EDTA in water by HPLC. Wat Res 28(3): 639�42.

Fayyad M, Tutunji M, Taha Z. 1988. Indirect trace determination of EDTA in waters by potentiometric stripping analysis. Analytical Letters 21: 1425�32.

WHO. 2003. Edetic Acid (EDTA) in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/58).


Epichlorohydrin Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of epichlorohydrin in drinking-water should not exceed 0.0005 mg/L (0.5 µg/L). The guideline value is considered to be provisional because of the uncertainties surrounding the toxicity of epichlorohydrin and the use of a large uncertainty factor in deriving the WHO guideline value.



1 To source waters

Epichlorohydrin may enter raw water as a contaminant from a wide range of industrial uses. It is used mainly in the manufacture of glycerine and unmodified epoxy resins, and also in the manufacture of elastomers, water treatment resins, surfactants, ion exchange resins, plasticisers, dyestuffs, pharmaceutical products, oil emulsifiers, lubricants and adhesives.


Epichlorohydrin can enter drinking-water supplies through the use of polyamine flocculating agents containing epichlorohydrin, although this is not a significant source in New Zealand at this time (Gregor et al 1993). The New Zealand standard (NZWWA 1999) states that epichlorhydrin levels shall not exceed 5 mg/kg of active polymer.


Epichlorohydrin may enter drinking-water through the leaching of epichlorohydrin from epoxy resin coatings on pipes.

Form and fate in the environment Epichlorohydrin hydrolyses in water.

Typical concentrations in drinking-water No data are available on the concentration of epichlorohydrin in New Zealand drinking-water supplies or from overseas studies.

Removal methods Epichlorohydrin may be contained in flocculating agents, or in water treatment resins. Unacceptable levels of the substance might therefore be reduced by consideration of the chemicals and the dosage being used for treatment of the water. No information is available on processes that might reduce epichlorohydrin concentrations in water, although aeration is unlikely to be successful.

Analytical methods

Referee method

Gas Chromatography with Electron Capture Detection (Pesselman and Feit 1988).


No alternative methods have been recommended for epichlorohydrin because no methods meet the required criteria. However, the following information may be useful: Epichlorohydrin in water can be determined by a purge and trap gas chromatographic procedure with mass spectrometry (USEPA 1987), or flame ioniation detection (Standard JWWA 1989). The limit of quantification is 0.01 mg/L.


Health considerations Epichlorohydrin is rapidly and extensively absorbed following oral administration and may be absorbed following both inhalation and skin contact. Following oral administration and inhalation, epichlorohydrin metabolites are excreted rapidly in the urine and expelled air. Epichlorohydrin is a strong irritant and acutely toxic following oral, percutaneous, subcutaneous and respiratory exposure. Death is due to effects on the central nervous system and the respiratory centre. Acute toxic responses following skin contact in humans are characterised by an initial redness and itching or burning sensation. With time, the redness intensifies and the tissue becomes swollen and blistered. The initial symptoms following inhalation are local irritation, burning sensation of eyes and throat, swelling of the face, nausea, vomiting, and severe headache. In a case-study, long-term effects due primarily to damage of the liver and kidney were still present two years after exposure. In epichlorohydrin workers, increased incidences of chromatid and chromosomal breaks in peripheral lymphocytes and decreases of blood cell counts were observed. An epidemiological study was undertaken for 863 workers, with probable exposure to epichlorohydrin at two chemical plants. All cancer, leukaemia, and most other causes of death were related to estimated levels of exposure to epichlorohydrin. The most consistent relationship was between exposure level and heart disease. In animal studies, epichlorohydrin induced squamous cell carcinomas in the nasal cavity by inhalation and forestomach tumours by the oral route. It has been shown to be genotoxic in vitro and in vivo. The International Agency for Research on Cancer has placed epichlorohydrin in Group 2A (probably carcinogenic to humans).

Derivation of maximum acceptable value Although epichlorohydrin is a genotoxic carcinogen, the use of the linear multistage model for estimating cancer risk was considered inappropriate because tumours are seen only at the site of administration where epichlorohyrin is highly irritating. A tolerable daily intake approach was therefore taken for the derivation of the MAV for epichlorohydrin in drinking-water. The lowest observable adverse effect level used in the derivation was determined for forestomach hyperplasia in a two-year study in rats by gavage. The provisional MAV for epichlorohydrin in drinking-water was derived as follows:

2 x (5/7) mg/kg body weight per day x 70 kg x 0.1 = 0.0005 mg/L (0.5 µg/L) 2 L x 10,000

where:

• lowest observable adverse effect level = 2 mg/kg body weight/day for forestomach hyperplasia in a two-year study in rats by gavage (normalised for five days/week dosing in the derivation)




• uncertainty factor = 10000 (100 for intra- and interspecies variation and 10 reflecting carcinogenicity and 10 for the use of a LOAEL instead of a NOAEL).



ANSI/AWA. 1990. AWWA Standard for EPI-DMA Polyamines. B452-90 (includes addendum AWWA B452a-97).

Gregor JE, Simpson S, Andrew C. 1993. Health Risks of Synthetic Polymers and Monomers, and Methods of Monitoring and Control: A report prepared for the Department of Health.

NZWWA. 1999. Standard for the Supply of EPI-DMA Polyamines for Use in Drinking-water Treatment. ISBN 1-877134-26-0.

Pesselman RL, Feit MJ. 1988. Determination of residual epichlorohydrin and 3-chloropropanediol in water by gas chromatography with electron capture detection. Journal of Chromatography 439: 488�542.

USEPA. 1987. Health Advisory for Epichlorohydrin.

WHO. 2003. Epichlorohydrin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/94).


Ethylbenzene Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of ethylbenzene in drinking-water should not exceed 0.3 mg/L. The maximum contaminant level (USEPA 2004) is 0.7 mg/L.


1 To source waters

The primary source of ethylbenzene in the environment is the petroleum industry. Ethylbenzene occurs naturally as a component of crude oil, and is present in petrol. It is produced commercially by the alkylation of benzene with ethylene, and by fractionation of petroleum. It is a major component of commercial xylene and is used commercially in paints, in insecticides, and as a solvent. Therefore ethylbenzene may also be found in source waters as an industrial contaminant. It is also found in contaminated coal tar and coal gas industrial sites.


No known sources.


No known sources.

Form and fate in the environment Because of its high vapour pressure and low solubility, ethylbenzene will disperse in the atmosphere if released. Biodegradation of ethylbenzene in soil, activated sludge and water, under aerobic conditions, has been reported.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 301 zones, found ethylbenzene concentrations to range from �not detectable� (nd) to 0.0018 mg/L, with the median concentration being �nd� (limit of detection = 0.0005 mg/L).

Removal methods Ethylbenzene can be removed from water by adsorption on to granular activated carbon or by air stripping. Significant removal is also expected by ozonation.

Analytical methods

Referee method




Health considerations Ethylbenzene, in liquid form, is absorbed readily by humans via the skin and via the intestinal tract, and the vapour is readily absorbed when inhaled. It can be stored in fat and is metabolised to mandelic and phenylglyoxalic acids and excreted in urine. It can cross the placenta. Ethylbenzene has a low acute toxicity via the oral route. No data are available on the human health effects after oral exposure, and inhalation data are limited to short-term studies. A six-month feeding study using rats reported enlargement of the liver and kidney at high doses (400 mg/kg body weight per day). Liver effects were also observed in a number of inhalation studies. No longer-term studies are available. No carcinogenicity data on ethylbenzene are available and the compound has been found to be nonmutagenic in a number of tests.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of a MAV for ethylbenzene in drinking-water. A no observable adverse effect level was determined in a limited 6-month study in rats, based on observed heptatoxicity and neprotoxicty. The MAV for ethylbenzene in drinking-water was derived as follows:

136 x (5/7) mg/kg body weight per day x 70 kg x 0.1 = 0.34 mg/L (rounded to 0.3 mg/L) 2 L x 1000

where:

• no observable adverse effect level = 136 mg/kg body weight per day based on heptatoxicity and nephrotoxicty observed in a limited six-month study in rats (normalised for five days/week dosing in derivation)





• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the limited data base and short duration of the study).

Taste and odour thresholds for ethylbenzene have been reported at 0.08 and 0.002 mg/L repectively.



AWWA/Lyonnaise des Eaux. 1987. Tastes and Odours in Drinking Water.




WHO. 2003. Ethylbenzene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/26).


Fluoranthene New entry July 2005. (See also polynuclear aromatic hydrocarbons.)

Maximum acceptable value (provisional) Based on health considerations, the concentration of fluoranthene in drinking-water should not exceed 0.004 mg/L (4 µg/L). WHO (2004) states that under usual conditions the presence of fluoranthene in drinking-water does not represent a hazard to human health. For this reason, the establishment of a guideline value for fluoranthene is not deemed necessary.


1 To source waters

Polynuclear aromatic hydrocarbons (PAHs) are a class of diverse organic compounds containing two or more fused aromatic rings of carbon and hydrogen atoms. They are ubiquitous pollutants formed from the combustion of fossil fuels and are always found as a mixture of individual compounds. Owing to their low solubility and high affinity for particulate matter, PAHs are not usually found in water in notable concentrations. Only a small number of PAHs are produced commercially. Fluoranthene is one of these, it is used as an intermediate in the production of fluorescent dyes. PAHs in the environment are almost always derived from anthropogenic activities. The largest amount of PAHs enter the environment via the atmosphere from incomplete combustion processes, such as processing of crude oil and coal, industrial use of mineral oil products and coal (including related


contaminated soils), heating, fires, incineration of refuse, vehicle traffic, tobacco smoking and volcanic activities.


No known sources.

3 Distribution system

Fluoranthene is the most commonly detected PAH in drinking-water and is associated primarily with coal tar linings of cast iron or ductile iron distribution pipes.

Forms and fate in the environment PAHs reach the hydrosphere mainly by dry and wet deposition and road runoff but additionally from industrial wastes containing PAHs, and leaching from creosote-impregnated wood. PAHs are adsorbed strongly to the organic fraction of sediments and soils, and hence leaching of PAHs from the soil surface layer to groundwater is assumed to be negligible. However, their presence in groundwater has been reported, mainly at contaminated sites. In laboratory experiments with soil samples, the calculated half lives for selected PAHs varied widely, from about 100 days to a couple of years. For pure water, the photodegradation half-life appears to be in the range of hours (Mill et al 1981, Mill and Mabey 1985 both cited in WHO 1998), whereas the half life increases dramatically when sediment/water partitioning is taken into account (Zepp and Schlotzhauer 1979, cited in WHO 1998). In summary, it can be concluded that sediments and soils are the main sinks for PAHs in the environment.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 346 zones, found fluoranthene concentrations to range from �not detectable� (nd) to 0.0002 mg/L, with the median concentration being �nd� (limit of detection = 0.0002 mg/L).

Removal methods An investigation by the UK Dept for Environment (1993) into methods to reduce the rate of leaching of fluoranthene from the coal tar lining of distribution pipes found that of the remedial methods investigated only the partial covering of the lining material with woven nylon hose was shown to be effective. Specific information about the removal of fluoranthene by coagulation/flocculation is unavailable, however, PAHs are hydrophobic (Faust and Aly 1983) and removal by adsorption to floc is therefore likely to achieve good removal. Fluoranthene is removed very effectively by adsorption to activated carbon (Faust and Aly 1983). Oxidation by ozone at neutral and acidic pH values can be achieved (Camel and Bermond 1998).

Analytical methods

Referee method

Liquid�Solid Extraction Gas Chromatographic/Mass Spectrometric Method (EPA 525).



Liquid�Liquid Extraction High Performance Liquid Chromatographic Method (EPA 550). Liquid�Solid Extraction High Performance Liquid Chromatographic Method (EPA 550.1).

Health considerations PAHs are absorbed in experimental animals and humans through the pulmonary tract, the gastrointestinal tract, and the skin. Oral administration of fluoranthene to rats caused peak concentrations of these compounds in blood after 1�2 hours. In general, little is known about the metabolism of most PAHs, particularly in non-rodent species. PAH metabolites and their conjugates are excreted predominantly via the faeces and to a lesser extent in the urine. The excretion of urinary metabolites is a method used to assess internal human exposure to PAHs.

Acute poisoning

The oral LD50 for fluoranthene in the rat is about 2000 mg/kg of body weight (Smyth et al 1962, cited in WHO 1998).

Chronic exposure

The MAV is determined on the basis of health effects from chronic exposure. Male and female mice were given fluoranthene by gavage for 13 weeks at 0, 125, 250 or 500 mg/kg of body weight per day and then sacrificed and autopsied (USEPA 1988). All treated mice exhibited nephropathy, increased salivation, and increased liver enzyme levels in a dose-dependent manner. At doses of 250 and 500 mg/kg of body weight per day, statistically increased serum glutamate-pyruvate transaminase (SGPT) levels and increased absolute and relative liver weights were noted, as well as compound-related microscopic liver lesions (indicated by pigmentation) in 65 and 87.5% of the mice, respectively. Fluoranthene is classified by the International Agency for Research on Cancer (IARC) as Group 3: Unclassifiable as to carcinogenicity to humans.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for fluoranthene in drinking-water. The NOAEL was identified on the basis of increased serum glutimate pyruvate transaminase levels, kidney and liver pathology, and clinical and haematological changes, as follows: The provisional MAV for fluoranthene in drinking-water was derived as follows:

125 mg/kg body weight per day x 70 kg x 0.01 = 0.004 mg/L 2 L x 10,000

where:

• no observable adverse effect level = 125 mg /kg body weight per day for increased serum glutamate�pyruvate transaminase levels, kidney and liver pathology, and clinical and haematological changes in a 13-week oral gavage study in mice



• proportion of tolerable daily intake allocated to drinking-water = 1% (because there is significant exposure from food)

• uncertainty factor = 10,000 (100 for interspecies and intraspecies variation, 10 for the use of a subchronic study and inadequate database, and 10 because of clear evidence of co-carcinogenicity with benzo[a]pyrene in mouse skin-painting studies.


References Camel V, Bermond A. 1998. The use of ozone and associated oxidation processes in drinking water treatment. Wat Res 32(11): 3208�22.

Environment Canada. 1994. Canadian Environmental Protection Act Priority Substances List Assessment Report: Polycyclic aromatic hydrocarbons. Ottawa, Canada: Supply and Services.


Mill, et al. 1981. Photolysis of polycyclic aromatic hydrocarbons in water. Chemosphere 10: 1281�90.

Mill and Mabey. 1985. Photochemical transformations. In: Neely WB, Blau GE (eds). Environmental Exposure from Chemicals 1: 175�216. Boca Raton, USA: CRC Press.


Smyth HF, et al. 1962. Range finding toxicity data: List VI. Industrial Hygiene Journal March�April: 95�107.

UK Department of the Environment. 1993. PAHs in Drinking-water: Investigation of leaching (DWE 7102). Final report to the Department of the Environment.

USEPA. 1988. 13-week Mouse Oral Subchronic Toxicity Study (Fluoranthene). Muskegon, USA: Toxicity Research Laboratories (TRL study #042-008).


USEPA. National Primary Drinking Water Regulations, Technical Factsheet on: Polynuclear aromatic hydrocarbons (PAHs). Available on: http://www.epa.gov/safewater/dwh/t-soc/pahs.html



Zepp RG, PF Schlotzhauer. 1979. Photoreactivity of selected aromatic hydrocarbons in water. In: Jones PW, Leber P (eds). Polynuclear Aromatic Hydrocarbons: Third International Symposium on Chemistry and Biology � Carcinogenesis and Mutagenesis, 141�58. Ann Arbor, USA: Ann Arbor Science Publishers.

Formaldehyde Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of formaldehyde in drinking-water should not exceed 1 mg/L.


1 To source waters

Formaldehyde may enter raw water from a great variety of human activities. Its primary industrial use is in the production of urea-formaldehyde, phenolic, melamine, pentaerythritol, and polyacetal resins used in the building industry. It is also used in the production of intermediates in the chemical industry, such as acetylene chemicals and hexamethylene, and in cosmetics, fungicides, textiles and embalming fluids.


Formaldehyde is produced as a disinfection by-product through the reaction between ozone and naturally-occurring organic substances, such as humic and fulvic acids.




Leaching of formaldehyde may occur from polyacetal plastic fittings if these are used in the distribution system.

Forms and fate in the environment Formaldehyde is a reactive, water soluble, colourless gas. In surface water it is expected to oxidise to formic acid, and may polymerise to paraformaldehyde.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from four zones (supplies using ozone treatment), has not found any formaldehyde at detectable concentrations (limit of detection = 0.01 mg/L). Concentrations of up to 0.03 mg/L have been found in ozonated drinking-water (WHO 2004).

Removal methods There is no information available on methods of removing formaldehyde from water if it is present in the source water through contamination. However, as formaldehyde arises in waters predominantly as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the ozone. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all other treatment steps, ie, avoiding preoxidation wherever possible. Pilot studies have shown that biologically active media are able to reduce the concentration of formaldehyde in ozonated water. Biologically active media in conjunction with preozonation is therefore likely to offer a means of oxidising organic matter and removing the by-products of ozonation.

Analytical methods

Referee method

Dinitrophenylhydrazine Derivatization and High Performance Liquid Chromatography (EPA 554).


No alternative methods have been recommended for formaldehyde because no methods meet the required criteria.

Health considerations Formaldehyde is present in almost all common foods, and adult dietary intake is estimated at 11 mg/day. Drinking water would contribute less than 10% of total intake.


Ingested formaldehyde is absorbed readily by the gastro-intestinal tract. It is metabolised rapidly to formic acid and subsequently to carbon dioxide and water. It is distributed primarily in muscle, with lower levels in the intestines, liver and other tissues. Exposure of the skin to formaldehyde at levels higher than those encountered in drinking-water has been associated with irritation and allergic contact dermatitis. The presence of formaldehyde in some types of water filters has been associated with outbreaks of haemolytic anaemia in dialysis unit patients. A number of epidemiological studies have looked at the effects of inhalation of formaldehyde. No effects could be directly attributed to long-term occupational exposure, but studies among exposed workers have reported elevated incidences of a number of cancers including nasal, buccal, nasopharyngeal, skin, prostate and colon cancers. There was no evidence of tumour-promoting activity when formaldehyde was applied to mouse skin, but rats and mice exposed to formaldehyde by inhalation exhibited an increased incidence of carcinomas of the nasal cavity. Ingestion of formaldehyde in drinking-water for 2 years caused stomach irritation in rats. Formaldehyde has demonstrated mutagenic activity when applied to cells in vitro but not when applied in vivo. There is some evidence that formaldehyde is a carcinogen in humans exposed by inhalation. The International Agency for Research on Cancer has classified formaldehyde in Group 2A (probably carcinogenic to humans) by inhalation. The weight of evidence indicates that formaldehyde is not carcinogenic by the oral route.

Derivation of maximum acceptable value Although formaldehyde is considered to be probably carcinogenic to humans by inhalation, the weight of evidence indicates that formaldehyde is not carcinogenic by the oral route. Therefore a tolerable daily intake approach has been used for the derivation of a MAV for formaldehyde in drinking-water. A no observable adverse effect level determined from a two-year study in rats has been used as the basis of the derivation. The MAV for formaldehyde in drinking-water was derived as follows:

15 mg/kg body weight per day x 70 kg x 0.2 = 1.05 mg/L (rounded to 1 mg/L) 2 L x 100

where:

• no observable adverse effect level = 15 mg/kg body weight per day from a two-year study in rats (for a variety of effects, including increased relative kidney weights in females and an increased incidence of renal papillary necrosis in both sexes)




• uncertainty factor = 100 (for intra- and interspecies variation). No account was taken of potential carcinogenicity from the inhalation of formaldehyde from various indoor water uses, such as showering.




Merck Index (11th ed), 1989.


Whittle PJ, Rennie PJ. 1988. Determination of formaldehyde in river water by high performance liquid chromatography. Analyst 113: 665�6.

WHO. 2003. Formaldehyde in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/48).


Hexachlorobutadiene Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of hexachlorobutadiene in drinking-water should not exceed 0.0007 mg/L (0.7 µg/L).


1 To source waters

Hexachlorobutadiene may enter raw water as an industrial and agricultural contaminant. It is used as a solvent in chlorine gas production, an intermediate in the manufacture of rubber compounds, a lubricant, a gyroscopic fluid, a pesticide, and a fumigant in vineyards.


No known sources.


No known sources.

Fate and form in the environment The highest levels of hexachlorobutadiene have been found in air near factories producing tetrachloroethylene and trichloroethylene. In water it may not volatilise rapidly because of its low vapour pressure. Adsorption to soil particles is an important removal mechanism from water.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find hexachlorobutadiene at detectable concentrations (limit of detection = 0.0005 mg/L). Has been detected in surface water at concentrations of a few micrograms per litre and in drinking-water at concentrations below 0.0005 mg/L (WHO 2004).

Removal methods Trials have shown granular activated carbon to be effective in the removal of hexachlorobutadiene.


Analytical methods

Referee method




Health considerations In rats, absorption of hexachlorobutadiene is about 95% of the ingested dose and it is found in the blood, liver, brain, spleen, kidney and mesentery. It is metabolised in the gastro-intestinal tract and kidney to a number of water soluble metabolites, and excreted in urine. Long term intermittent human exposure has been reported to cause higher incidences of hypotension, myocardial dystrophy, nervous system and liver disorders, and respiratory tract lesions. The primary target organ for hexachlorobutadiene toxicity is the kidney. Kidney tumours were observed in a long-term oral study in rats. HCBD has not been shown to be carcinogenic by other routes of exposure. Tests for mutagenicity with different strains of bacteria have given positive and negative results. Some metabolites have given positive results. The International Agency for Research on Cancer has placed hexachlorobutadiene in Group 3 (not classifiable as to its carcinogenicity to humans).

Derivation of maximum acceptable value Based on the available metabolic and toxicological information available, a tolerable daily intake approach is considered most appropriate for the derivation of the MAV. A no observable adverse effect level determined for renal toxicity in a two-year feeding study in rats has been used for the basis of the derivation. The MAV for hexachlorobutadiene in drinking-water was derived as follows:

0.2 mg/kg body weight per day x 70 kg x 0.1 = 0.0007 mg/L (0.7 µg/L) 2 L x 1000

where:

• no observable adverse effect level = 0.2 mg/kg body weight per day for renal toxicity in a two-year feeding study in rats




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for limited evidence of carcinogenicity and the genotoxicity of some metabolites).

WHO (2004) says the practical quantification level for HCBD is of the order of 0.0002 mg/L, but concentrations in drinking-water can be controlled by specifying the HCBD content of products coming into contact with it.




IPCS. 1994. Hexachlorobutadiene. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 156).


WHO. 2003. Hexachlorobutadiene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/101).


Monobromoacetic acid New entry August 2005.

Maximum acceptable value There are insufficient data to derive MAVs for individual monobromoacetic acid in drinking-water. WHO (2004) states that the available data relating to monobromoacetate was considered inadequate to permit recommendation of a health-based guideline value.


1 To source waters

Brominated acetic acids are formed during disinfection (with ozone) of water which contains bromide ions and organic matter. Bromide ions occur naturally in surface water and groundwater and exhibit seasonal fluctuations in concentrations. Bromide ion concentrations can increase due to saltwater intrusion resulting from drought conditions, or due to pollution. Bromide is introduced into New Zealand surface waters usually by wind blown seaspray.


No known sources.


No known sources.




Removal methods Brominated acetic acids arise in waters as a disinfection by-product, so the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the ozone. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps.

Analytical methods

Referee method

[???]


[???]

Health considerations Data are limited on the oral toxicity of monobromoacetic acid. Limited mutagenicity and genotoxicity data give mixed results. Data gaps include subchronic or chronic toxicity studies, multigeneration reproductive toxicity studies, standard developmental toxicity studies and carcinogenicity studies. The available data are considered inadequate to establish guideline values for these chemicals.

Derivation of maximum acceptable value The are insufficient data to derive a MAV for monobromoacetic acid at this time.

Monochloroacetic acid




Revised July 2005. (Sometimes called chloroacetic acid.)

Maximum acceptable value Based on health considerations, the concentration of monochloroacetic acid in drinking-water should not exceed 0.02 mg/L. The maximum contaminant level for the five haloacetic acids (USEPA 2004) is 0.06 mg/L.




1 To source waters

Monochloroacetic acid may be used as an intermediate or reagent in the synthesis of a variety of chemicals and as a pre-emergence herbicide and therefore may enter raw water as an industrial and agricultural contaminant.


Chlorinated acetic acids are formed from natural organic material during water chlorination.


No known sources.

Form and fate in the environment There is little information available on the environmental fate of monochloroacetic acid.

The P2 Chemical Determinand Identification Programme, sampled from 488 zones, found monochloroacetic acid concentrations to range from �not detectable� (nd) to 0.010 mg/L, with the median concentration being �nd� (limit of detection = 0.005 mg/L). Based on preliminary data, concentrations of chloroacetic acids in Australian drinking waters range from 0.01 mg/L to 0.1 mg/L and concentrations reported overseas range up to 0.16 mg/L, and are typically about half the chloroform concentration (Australian Drinking-water Guidelines). Present in surface water-derived drinking-water at <0.002�0.0082 mg/L, mean 0.0021 mg/L (WHO 2004).

Removal methods No information is available on methods of removing monochloroacetic acid from contaminated source waters. As this compound arises predominantly in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible.

Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.


Analytical methods

Referee method

A referee method cannot be selected for monochloroacetic acid because a MAV has not been established and therefore the sensitivity required for the referee method is not known. But has now.


No alternative methods can be recommended for monochloroacetic acid because a MAV has not been established. However, the following information may be useful. Chloroacetic acids in water may be determined by solvent extraction with methyl tert-butyl ether, methylation and analysis by gas chromatography with electron capture detection (Method EPA 552). Limits of quantification are lower than 0.001 mg/L (1 µg/L).


Derivation of maximum acceptable value

Chloroacetic acids are probably absorbed rapidly after ingestion, but there are no data to confirm this assumption. Rats given monochloracetate subcutaneously had particularly high levels in liver and kidneys and approximately 50% of the dose was excreted in urine within less than one day. A short-term study of mice exposed to varying concentrations of monochloroacetic acid by gavage report decreased weight gain and increased liver weights at the highest dose, amongst females. No evidence of carcinogenicity of monochloroacetate was found in two-year gavage bioassays with rats and mice. Monochloroacetate has given mixed results in a limited number of mutagenicity assays and has been negative for clastogenicity in genotoxicity studies. IARC has not classified the carcinogenicity of monochloroacetic acid.

Based on the available metabolic and toxicological information available, a tolerable daily intake approach is considered most appropriate for the derivation of the MAV. A lowest observable adverse effect level found during a two-year feeding study on rat spleens has been used for the basis of the derivation. The MAV for monochloroacetic acid in drinking-water was derived as follows:


where:

• lowest observable adverse effect level = 3.5 mg/kg body weight per day from a study in which increased absolute and relative spleen weights were observed in male rats exposed to monochloroacetic acid in drinking-water for two years




• uncertainty factor = 1000 (100 for inter- and intraspecies variation and 10 for use of a minimal LOAEL instead of a NOAEL and database de.ciencies, including the lack of a multigeneration reproductive toxicity study).




NTP. 1992. Toxicology and Carcinogenesis Studies of Monochloroacetic Acid in F344/N rats and B6C3F1 Mice (Gavage Studies). National Toxicology Program, NTP TR 396, NIH Publication No 92-2851. USA: United States Department of Health and Human Services, Public Health Service, National Institute of Health.

USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking-Water, Supplement 1. Report No EPA/600/4-90-020.

USEPA Method 552. 1990. Determination of Haloacetic Acids in Drinking Water by Liquid�Liquid Extraction, Derivatization, and Gas Chromatography with Electron Capture Detection. Cincinnati, USA: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).

Based on health considerations, the concentration of monochlorobenzene in drinking-water should not exceed 0.3 mg/L. WHO (2004) states that because monochlorobenzene occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value. The maximum contaminant level (USEPA 2004) is 0.1 mg/L.


WHO. 2003. Monochloroacetic Acid in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/85).


Monochlorobenzene Revised July 2005.

Maximum acceptable value (provisional)


1 To source waters

Monochlorobenzene is used primarily as a solvent in pesticide formulations, as a degreasing agent, and as an intermediate in the synthesis of other halogenated organics, and therefore it may occur in the environment as an industrial and agricultural contaminant.


No known sources.


No known sources.


Fate and form in the environment Because monochlorobenzene is a volatile compound and is used as a solvent, most of that found in the environment is likely to be in the air, and inhalation is probably the major route of environmental exposure. Monochlorobenzene released into water and on to land will decrease in concentration mainly because of volatilisation into the atmosphere. Some biodegradation also occurs in water. Monochlorobenzene is relatively mobile in sandy soil and aquifer material and therefore may leach into groundwater. Bioconcentration is unlikely to occur in aquatic species.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 294 zones, did not find monochlorobenzene at detectable concentrations (limit of detection = 0.0005 mg/L). Monochlorobenzene has been detected in surface water, groundwater and drinking-water; mean concentrations were less than 0.001 mg/L in some potable water sources (maximum 0.005 mg/L) in Canada.

Removal methods Removal of monochlorobenzene can be achieved through adsorption on to activated carbon or air stripping.

Analytical methods

Referee method



Monochlorobenzene was not mutagenic in tests with bacteria, but may bind to RNA and DNA.


Health considerations Monochlorobenzene is absorbed readily after ingestion or inhalation and accumulates mainly in fatty tissue and the liver and kidney. It is metabolised to 4-chlorocatechol, which is excreted in urine. Monochlorobenzene is of low acute toxicity to experimental animals via the oral and inhalation routes and the major target organs are the liver and kidney.

In humans, the symptoms of monochlorobenzene toxicity resulting from poisoning and occupational exposure were central nervous system disturbances. Subjects occupationally exposed to monochlorobenzene for two years suffered headaches, dizziness and sleepiness.

There was evidence of an increase of liver tumours in male rats fed 120 mg/kg body weight per day of monochlorobenzene for two years. No increases were observed in female rats, or in male and female mice.


Derivation of maximum acceptable value A tolerable daily intake approach was used for the derivation of the MAV for monochlorobenzene in drinking-water. The no observable adverse effect level used as the basis for the derivation is for neoplastic nodules in the liver from a two-year study with rats and mice. The MAV for monochlorobenzene in drinking-water was derived as follows:


where:

• no observable adverse effect level = 60 mg/kg body weight per day for neoplastic nodules in the liver from a two-year study with rats and mice (normalised for five days/week dosing in derivation)




• uncertainty factor = 500 (100 for intra- and interspecies variation and 5 for limited evidence of carcinogenicity).

Taste and odour thresholds of 0.01�0.02 mg/L, and odour thresholds of 0.04�0.12 mg/L have been reported for monochlorobenzene.



References





WHO. 2003. Monochloroacetic Acid in Drinking-water. Background document for preparation of WHO Guidelines for drinking-water quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/85).

MX (3-chloro-4-dichloromethyl-5-hydroxy-2(5H)-furanone) Revised July 2005.

Maximum acceptable value Insufficient data are available to establish a MAV for MX in drinking-water.


1 To source waters

MX has no commercial use but may be present in the chlorinated effluents of pulp mills.



MX is formed by the reaction of chlorine with complex organic matter and therefore may be found in treated water as a by-product of the treatment process.

There is little information available on the environmental fate of MX.

Removal methods


No known sources.

Form and fate in the environment

Typical concentrations in drinking-water No data are available on the concentration of MX in New Zealand drinking-water supplies.

It has been identified in chlorinated humic acid solutions and drinking-water in Finland, the United Kingdom and the USA, and was found to be present in 37 water sources at levels of 0.000002�0.000067 mg/L (2�67 ng/L). Five drinking-water samples from different Japanese cities contained MX at concentrations ranging from <3 to 9 ng/L.

No information is available on methods for removing MX from contaminated source waters. As this compound arises in waters as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation.

Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. MX formation is minimised if the pH of the water is maintained above 7. The stability of MX is dependent on pH. Below pH 7 it is relatively stable but above pH 7 it rapidly breaks down.

Analytical methods

Referee method

A referee method cannot be selected for MX because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for MX because a MAV has not been established. However, the following information may be useful:


MX is difficult to detect because of low concentrations and the masking effects of other substances. MX can be analysed in drinking-water using XAD resins, high-pressure liquid chromatography, capillary column gas chromatography with mass spectrometry and selective ion monitoring. No detection limits are cited (Munch et al 1987; Hemming et al 1986).

Health considerations At least 40% of a dose of MX administered to rats was absorbed and about 5% was recovered in the liver, muscle, skin, kidneys and blood. MX is a potent mutagen in bacteria and in cells in vitro and has undergone a lifetime study in rats in which some tumorigenic responses were observed. These data indicate that MX induces thyroid and bile duct tumours. IARC has classified MX in Group 2B on the basis of rat tumorigenicity and its strong mutagenicity. MX has been reported to be an extremely strong mutagen in one strain of Salmonella typhimurium. Responses with other strains were generally positive, although not as strong. MX has induced significant increases in structural chromosomal aberrations in cultured mammalian cells.

Derivation of maximum acceptable value A health-based value of 0.0018 mg/L can be calculated for MX on the basis of the increase in cholangiomas and cholangiocarcinomas in female rats using the linearised multistage model (without a body surface area correction). However, this is significantly above the concentrations that would be found in drinking-water, and, in view of the analytical difficulties in measuring this compound at such low concentrations, it is considered unnecessary to propose a formal guideline value for MX in drinking-water (WHO 2004).


Hemming J, et al. 1986. Determination of the strong mutagen 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone in chlorinated drinking and humic waters. Chemosphere 15: 549�56.

WHO. 2003. MX in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/108).

Nitrilotriacetic acid


Munch JW, et al. 1987. Determination of a chlorinated furanone (MX) in treated water. In: Proceedings of the American Water Works Quality Technology Conference, Baltimore, MD. November, pp. 933�42.


Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of nitrilotriacetic acid in drinking-water should not exceed 0.2 mg/L.



1 To source waters

Nitrilotriacetic acid (NTA) may be present in raw water that has been contaminated with wastewater or industrial discharge. It is a chelating agent used in laundry detergents as a replacement for phosphate. It is also used in the treatment of boiler water to prevent scale formation, and in the photographic, metal plating, textile manufacturing, and paper and cellulose industries.


No known sources.


No known sources.

Fate and form in the environment Nitrilotriacetic acid mobilises heavy metals from aquatic sediments and is present in water primarily in the form of metal complexes, most of which degrade rapidly. Under certain conditions, nitrilotriacetic acid is broken down by photochemical and chemical reactions.

Typical concentrations in drinking-water No data are available on the concentration of nitrilotriacetic acid in New Zealand drinking-water supplies. The range of nitrilotriacetic acid measured in 70 Canadian drinking-waters was less than 0.0002�0.03 mg/L (0.2�30 µg/L). The average concentration of nitrilotriacetic acid detectable in the drinking-water of eight cities in the USA was 0.0021 mg/L (2.1 µg/L). Surface water concentrations in Europe have been found to be range up to 0.012 mg/L.

Removal methods No information is available on processes for removing nitrilotriacetic acid from water.

Analytical methods

Referee method

GC-NSD (Malaiyandi et al 1979; Aue et al 1972)


No alternative methods have been recommended for nitrilotriacetic acid because no methods meet the required criteria.

Health considerations Absorption of nitrilotriacetic acid from the gastro-intestinal tract is rapid and it does not appear to be metabolised by mammals. It is excreted in urine and accumulates in bone and possibly the kidneys. Nitrilotriacetic acid does not appear to be have high acute toxicity to mammals. There is little evidence regarding the toxicity of nitrilotriacetic acid in humans. Based on physical examination, blood chemistry analysis, and urinalysis, no adverse health effects were reported in a metabolism study in which volunteers ingested a single dose of 10 mg nitrilotriacetic acid.


Nitrilotriacetic acid has not been shown to be teratogenic or genotoxic in studies conducted to date but has induced urinary tract tumours in rats and mice at high doses. The induction of tumours in rodents is considered to be due to cytotoxicity resulting from the chelation of divalent cations such as zinc and calcium in the urinary tract, leading to the development of hyperplasia and neoplasia. The International Agency for Research on cancer has placed nitrilotriacetic acid in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value Because nitrilotriacetic acid induces tumours only after prolonged exposure to doses higher than those that produce nephrotoxicity, the MAV for nitrilotriacetic acid in drinking-water was determined using a tolerable daily intake approach. The no observable adverse effect level used in the derivation was determined for nephritis and nephrosis in a two-year study in rats. The MAV for nitrilotriacetic acid in drinking-water was derived as follows:

10 mg/kg body weight per day x 70 kg x 0.5 = 0.2 mg/L 2 L x 1000

where:

• no observable adverse effect level = 10 mg/kg body weight per day for nephritis and nephrosis in a two-year study in rats

The MAV was derived on the basis of a NOAEL for nephrotoxicity effects rather than induced tumours, because this occurred at higher doses. However, a larger uncertainty factor has been used to account for the evidence of urinary tumour induction at high doses.

(Also called polychlorinated biphenyls.)




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for carcinogenic potential at high doses).

References Aue W, Hastings CR, Gerhardt KO, et al. 1972. The determination of part-per-billion levels of citric and nitriloacetic acids in tap water and sewage effluents. Journal of Chromatography 72: 259�67.



Malaiyandi M, Williams DT, O�Grady R. 1979. A national survey of nitrilotriacetic acid in Canadian drinking-water. Environmental Science and Technology 13: 59�62.

WHO. 2003. Nitrilotriacetic Acid in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/30).


PCBs New entry August 2005.


Maximum acceptable value There are insufficient data to derive MAVs for any of the polychlorinated biphenyls in drinking-water. WHO (2004) does not mention PCBs. The maximum contaminant level (USEPA 2004) is 0.0005 mg/L.


1 To source waters

PCBs are used in hundreds of industrial and commercial applications including electrical, heat transfer, and hydraulic equipment; as plasticisers in paints, plastics and rubber products; in pigments, dyes and carbonless copy paper, and many other applications. They are not formed naturally. Their use is being restricted increasingly. At present, the major source of PCB exposure in the general environment is redistribution of PCBs previously introduced into the environment

Removal methods

.

Forms and fate in the environment PCBs have low solubilities, and as a result adsorb readily to sediments and suspended matter in the water (IPCS 1992). The strength of adsorption is greater for the more highly chlorinated PCBs, with the volatility of the compounds also being decreased by adsorption. Adsorption can immobilise PCBs for relatively long periods, but desorption does release the compounds back into the bulk water. Sediment thereby acts as a sink for these compounds, and as a possible reservoir. In water, hydrolysis and oxidation have little effect on PCBs; photolysis appears to be the only abiotic degradation process. The degradation of PCBs is dependent on their degree of chlorination: persistence increases with chlorine content (IPCS 1992). The stability of PCBs has resulted in their wide dispersion throughout the environment globally. Differences in volatilisation and decomposition rates between different members of the PCB family lead to changes in the composition of PCB mixtures in the environment.

Typical concentrations in drinking-water Globally, PCBs are present in rain and snow in the range 0.001 to 0.25 µg/L, and less than 0.001 µg/L in non-contaminated areas drinking-waters, although levels up to 0.005 µg/L have been reported (IPCS 1992). PCBs were included routinely on the Department of Health�s surveillance programme during the late 1980s and early 1990s. None were detected (detection limits approximately 0.0005 mg/L) (Nokes, 1992).

There is limited information available about the removal of PCBs from water. Physical data for adsorption to activated carbon show that this medium should be effective in removing PCBs (Faust and Aly 1983). The presence of humic acids in the water can reduce the adsorbance of PCBs by up to 71% (Pirbazari et al 1992). The removal of PCBs by coagulation ranges from 10�40% depending on the dose and the coagulant; alum is more effective than ferric chloride (Aly and Badawy 1986). No significant reduction in PCB levels is reported for experiments using ultraviolet light, ozone, and the combination of UV and ozone (Vollmuth and Niessner 1995). Reverse osmosis has been used to concentrate PCBs (Faust and Aly 1983), so RO should remove them from water, although its efficacy of is unknown.


Chlorination, although capable of reducing the concentrations of some isomers, can react with isomers of lower chlorine content to produce PCBs of higher molecular weight (Aly and Badawy 1986).


Referee method

A referee method cannot be selected for the PCBs because MAVs have not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for these PCBs because MAVs have not been established. However, the following information may be useful: ???????????????

Health considerations Exposure to PCBs is principally through food, including mother�s milk in the case of babies.

Acute poisoning

Occupational exposures have led to skin rashes, and after exposure to high concentrations of PCBs, itiching, burning sensations, irritation of the conjunctivae and chloracne have been found, the latter being the most prevalent (IPCS 1992). As PCBs contain contaminants it is difficult to determine the extent to which these signs and symptoms arise from PCBs themselves and those from polychlorinated dibenzofuran contaminants.

Chronic exposure

PCBs also have effects on the immune, reproductive, nervous, and endocrine systems.

Nokes CJ. 1992. Organic Contaminants in New Zealand Potable Water Supplies: A review of data from samples taken between May 1987 and May 1991. Report to the Department of Health.

Both IARC (IARC 1987) and the USEPA (USEPA) have concluded that PCBs are probable human carcinogens. The most carcinogenic PCB mixtures are believed to be those bound to sediments. There is evidence that PCBs are linked to cancer in the liver (USEPA), the gall bladder and the gastrointestinal tract (IPCS 1992), and to the production of malignant melanomas. The PCB in question influences which organ is the target.

Derivation of maximum acceptable value There are insufficient data to derive MAVs for any of the polychlorinated biphenyls in drinking-water.

References Aly OA, Badawy MI. 1986. PCB removal by conventional water treatment: effect of chemical coagulation and chlorination. Bulletin for Environmental Contamination and Toxicology 36(6): 929�34.


IARC. 1987. Polychlorinated Biphenyls. International Agency for Research into Cancer, Summary and Evaluation Supplement 7. World Health Organization, available on http://www.inchem.org/documents/iarc/iarc/iarc936.htm

IPCS. 1992. Polychlorinated Biphenyls and Terphenyls. Environmental Health Criteria Monograph 1992. Geneva: World Health Organization, International Programme on Chemical Safety.

Pirabazri M, et al. 1992. Evaluating GAC adsorbers for the removal of PCBs and toxaphene. J Am Wat Wks Assn 84(2): 83�90.


USEPA. Health Effects of PCBs. Office of Pollution Prevention and Toxics. Available on http://www.epa.gov/opptintr/pcb/effects.htm


Polynuclear aromatic hydrocarbons

1 To source waters

Vollmuht S, Niessner R. 1995. Degradation of PCDD, PCDF, PAH, PCB and chlorinated phenols during the destruction-treatment of landfill seepage water in a laboratory model reactor (UV, ozone and UV/ozone). Chemosphere 30(12): 2317�31.

Revised July 2005. (Sometimes called polyaromatic hydrocarbons or PAHs.)

benzo 1,12-perylene (benzo[g,h,i]perylene benzo 11,12-fluoranthene (benzo[k]fluoranthene) benzo 3,4-fluoranthene (benzo[b]fluoranthene) benzo 3,4-pyrene (benzo[a]pyrene) see separate listing benzo[a]anthracene chrysene dibenz[a,h]anthracene fluoranthene see separate listing indeno[1,2,3-c,d]pyrene naphthalene

Maximum acceptable value There are insufficient data to derive MAVs for any of the polynuclear aromatic hydrocarbons in drinking-water other than benzo[a]pyrene and fluoranthene


Polycyclic aromatic hydrocarbons are a large group of organic compounds formed from the incomplete combustion of organic matter. They have no industrial use but are formed naturally in forest fires, and volcanic activity or from anthropogenic activities such as domestic fires, vehicle emissions, coke ovens, coal gas manufacture (including related contaminated soils), and aluminium smelters. The principle route of entry to source water is via atmospheric deposition. The United States Environmental Protection Agency has identified 16 priority pollutant polynuclear aromatic hydrocarbons.


No known sources.


Treated water may be contaminated by leaching from coal-tar liners in water distribution systems.

Forms and fate in the environment Polynuclear aromatic hydrocarbons enter the environment through atmospheric deposition. Because of their low water solubility most polynuclear aromatic hydrocarbons are adsorbed to sediments and suspended solids in aquatic systems. Volatilisation may be important over periods exceeding one month. Most polynuclear aromatic hydrocarbons are susceptible to aqueous photolysis. Polynuclear aromatic hydrocarbons of three or fewer fused aromatic rings are biodegraded but for the larger polynuclear


aromatic hydrocarbons this is minimal. Polynuclear aromatic hydrocarbons are adsorbed but not greatly accumulated by aquatic biota.

Typical concentrations in drinking-water The review of organic contaminants in New Zealand drinking-water supplies between 1987�1992 contained polynuclear aromatic hydrocarbons results from 217 samples, representing 204 supplies. The concentrations of the individual polynuclear aromatic hydrocarbons are summarised below.

Benzo[a]anthracene

Benzo[a]anthracene was detected in four of the samples with concentrations ranging from 0.0000012�0.0000031 mg/L (1.2�3.1 ng/L).

Benzo 1,12-perylene (benzo[g,h,i]perylene

The concentration of benzo 11,12-perylene was less than 0.00000046 mg/L (0.46 ng/L) in all samples.

Benzo 11,12-fluoranthene (benzo[k]fluoranthene

Twenty seven samples contained detectable concentrations of benzo[k]fluoranthene. Detected concentrations ranged from 0.00000003�0.00000042 mg/L (0.03�0.42 ng/L).

Benzo 3,4-fluoranthene (benzo[b]fluoranthene

Sixteen samples contained detectable concentrations of benzo[b]fluoranthene. Detected concentrations ranged from 0.0000001�0.00000099 mg/L (0.1�0.99 ng/L).

Chrysene

Chrysene was not detected (less than 0.0000023 mg/L (2.3 ng/L)) in any of these samples.

Indeno[1,2,3-c,d]pyrene

No data are available on the concentration of indeno[1,2,3-c,d]pyrene in New Zealand drinking-water supplies. Indeno [1,2,3-c,d] pyrene has been detected in some Canadian water supplies (detection limits 0.000001�0.000006 mg/L (1�6 ng/L)).

Removal methods Polynuclear aromatic hydrocarbons are very insoluble in water and hence adsorb readily to available surfaces. As a result, conventional coagulation/flocculation is able to achieve high levels of removal by removing particles to which the polynuclear aromatic hydrocarbons are adsorbed, and by providing floc surfaces on to which polynuclear aromatic hydrocarbons in the bulk water may adsorb. Good removal can also be achieved by granular activated carbon. Chlorination can reduce polynuclear aromatic hydrocarbons concentrations either through the production of oxidation products, or the formation of chlorinated polynuclear aromatic hydrocarbons. As polynuclear aromatic hydrocarbons can be leached from coal-tar lined pipes, this surface covering should not be used for pipes in the treatment plant or reticulation.


Analytical methods

Referee method

A referee method cannot be selected for these PAHs because MAVs have not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for these PAHs because MAVs have not been established. However, the following information may be useful:

1 Liquid�Solid Extraction and Capillary Column Gas Chromatography/Mass Spectrometry (EPA 525).

2 Liquid�Liquid Extraction and HPLC with Coupled Ultraviolet and Fluorescence Detection (EPA 550).

3 Liquid�Solid Extraction and HPLC with Coupled Ultraviolet and Fluorescence Detection (EPA 550.1).

Health considerations PAHs have been detected in a variety of foods as a result of the deposition of airborne PAHs and in fish from contaminated waters. PAHs are also formed during some methods of food preparation, such as charbroilling, grilling, roasting, frying or baking. For the general population, the major routes of exposure to PAHs are from food and ambient and indoor air. The use of open fires for heating and cooking may increase PAH exposure. There have been few studies on the human health effects of polynuclear aromatic hydrocarbons. Cases of accidental poisoning with naphthalene, resulting in death by acute haemolytic anaemia, have been reported. Occupations associated with exposures to polynuclear aromatic hydrocarbons, of which benzo[a]pyrene is a component, have been associated clearly with human cancer. Benzo[a]pyrene is absorbed principally through the gastro-intestinal tract and the lungs. The rate of absorption of different polynuclear aromatic hydrocarbons is influenced by their lipid solubilities and the content of polyunsaturated fatty acids in the diet. Most of the toxicological literature deals with benzo[a]pyrene. Few studies are available for the other polynuclear aromatic hydrocarbons. The health effect of primary concern is carcinogenicity. Many polynuclear aromatic hydrocarbon-containing mixtures have been associated with increased incidence of cancer, but the contribution of each of the individual components to the overall carcinogenic potency is difficult to assess. The relative carcinogenic potencies of various polynuclear aromatic hydrocarbons, based on bioassays by several routes of administration and related toxicological data, have been ranked in decreasing order as follows: dibenz[a,h]anthracene, benzo[a]pyrene, anthanthrene, indeno[1,2,3-cd]pyrene, benz[a]anthracene, benzo[b]fluoranthene, pyrene, benzo[k]fluoranthene, benzo[j]fluoranthene, cyclopentadieno[c,d]pyrene, benzo[g,h,i]perylene, chrysene and benzo[e]pyrene. The US Environmental Protection Agency has determined that acenaphthene, anthracene, fluoranthene, fluorene and pyrene are not classifiable as to human carcinogenicity based on no human data and inadequate data from animal bioassays.

Derivation of maximum acceptable value Benzo[a]pyrene and fluoranthene are listed separately. There are insufficient data to derive MAVs for any of the other polynuclear aromatic hydrocarbons in drinking-water.


The only time that WHO had a guideline value for other than benzo[a]pyrene was in their 1971 International Standards, which stated that some PAHs are known to be carcinogenic and that the concentrations of six representative PAH compounds (fluoranthene, 3,4-benzfluoranthene, 11,12-benzfluoranthene, 3,4-benzpyrene, 1,12-benzpyrene and indeno[1,2,3-cd]pyrene) should not in general exceed 0.0002 mg/L.


Organic Contaminants in New Zealand Potable Water Supplies. Report to Department of Health, New Zealand, 1992.


USEPA. National Primary Drinking Water Regulations, Technical Factsheet on: Polynuclear aromatic hydrocarbons (PAHs). Available on: http://www.epa.gov/safewater/dwh/t-soc/pahs.html

Jolley RL, Bull RJ, Davis WP, et al (eds). 1985. Water Chlorination: Chemistry, environmental impact and health effects 5: 1515�[add page number]. Chelsea, USA: Lewis Publishers Inc.

Jolley RL, Condie LW, Johnson JD, et al (eds). 1990. Water Chlorination: Chemistry, environmental impact and health effects 6: 12�[add page number]. Chelsea, USA: Lewis Publishers Inc.



Styrene Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of styrene in drinking-water should not exceed 0.03 mg/L. The maximum contaminant level (USEPA 2004) is 0.1 mg/L.


1 To source waters

Styrene may occur in source waters as a result of industrial contamination or the breakdown of discarded polystyrene products in landfills. Styrene is used extensively in New Zealand in the production of resins for the fibreglass industry. Polystyrene is not manufactured here but is moulded into many products. Forest fires may contribute to atmospheric concentrations of styrene.


No known sources.


No known sources.



Form and fate in the environment Styrene is expected to be volatilised readily to the atmosphere where it will react with hydroxyl radicals and ozone. In surface water styrene will oxidise and form peroxides or aldehydes with a penetrating odour. It is biodegraded and not bioaccumulated.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 301 zones, did not find styrene at detectable concentrations (limit of detection = 0.0005 mg/L).

Styrene has a low acute toxicity.

Styrene has been detected in drinking-water and surface water at concentrations below 0.001 mg/L (WHO 2004).

Removal methods Granular activated carbon can be used to remove styrene from water. It can be oxidized also by ozone. The oxidation products (aldehydes, ketones and benzoic acid) may themselves need to be removed if they are present in sufficient quantities.

Analytical methods

Referee method




Health considerations The daily exposure to styrene has been estimated to be 0.04 mg per person, with smokers receiving a higher dose. Uptake of styrene is rapid and it is distributed widely in the whole body, with a preference for lipids. It is metabolised by a number of tissues and organs to styrene-7,8-oxide. Elimination of styrene from lipid depots is less rapid (half-life 2�4 days) than from other tissues and no tendency for long-term accumulation exists. More than 90% of the oral dose is excreted rapidly as metabolites in urine.

Short-term controlled studies in volunteers exposed to styrene by inhalation at concentrations above 210 mg/m3 in air showed that it can cause irritation of mucous membranes of eyes, nose and/or respiratory tract and depression of the central nervous system. No chromosomal aberrations in peripheral lymphocytes could be detected in workers occupationally exposed to low concentrations of styrene, but significantly elevated frequencies of chromosomal aberrations were observed in peripheral lymphocytes of workers occupationally exposed to much higher styrene concentrations. Based on the available data, the International Agency for Research on Cancer has classified styrene in Group 2B (possibly carcinogenic to humans). It possesses mutagenic properties in in vitro systems with metabolic activation only. In vivo studies showed positive effects only at high doses. The available data suggest that the carcinogenic effects of styrene are due to the formation of the carcinogenic metabolite


styrene-7,8-oxide as a consequence of overloading the detoxication mechanisms after exposure to high styrene levels.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of a MAV for styrene in drinking-water. A no observable adverse effect level determined for reduced body weight in a two-year drinking-water study in rats has been used as the basis of the derivation. The MAV for styrene in drinking-water was derived as follows:


where:

• no observable adverse effect level = 7.7 mg/kg body weight per day for reduced body weight from a two-year drinking-water study in rats




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for carcinogenicity and genotoxicity of the reactive intermediate styrene-7,8-oxide).

The odour threshold for styrene has been reported at a concentration as low as 0.004 mg/L. The taste threshold of styrene in water at 40°C ranges from 0.02 mg/L to 2.6 mg/L, depending on individual sensitivities.




Irving Sax N (ed). 1987. Hazardous Chemicals Information Annual No. 2.



WHO. 2003. Styrene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/27).



Tetrachloroethene Revised July 2005. (Also known as tetrachloroethylene or perchloroethylene.)

Based on health considerations, the concentration of tetrachloroethene in drinking-water should not exceed 0.05 mg/L. The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


Tetrachloroethene may be formed from naturally occurring organic precursors during the chlorination of drinking-water.

Maximum acceptable value

1 To source waters

Tetrachloroethene can be released to the aquatic environment as an industrial contaminant. It is used as a solvent in the dry-cleaning industry. It is also used as a degreasing solvent in metal industries, as a heat transfer medium and in the manufacture of fluorohydrocarbons.



No known sources.

Forms and fate in the environment Most tetrachloroethene released to the environment is found in the atmosphere. In water, tetrachloroethene does not readily undergo hydrolysis or photolysis, but is biodegraded to dichloroethene, vinyl chloride, and ethylene. Tetrachloroethene can persist in waters where volatilisation cannot occur. It volatilises less readily from soil than from water. It is expected to be fairly mobile in soils and hence is likely to leach into groundwater. Tetrachloroethene does not appear to bioaccumulate in animals or food chains.

Typical concentrations in drinking-water The 1992 review of organic contaminants in New Zealand drinking-water supplies from 1987�1992 contained tetrachloroethene results from 52 samples. It was found in only one of these samples, at a concentration of 0.0017 mg/L.

The P2 Chemical Determinand Identification Programme, sampled from 505 zones, found tetrachloroethene concentrations to range from �not detectable� (nd) to 0.004 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). Concentrations in drinking-water are generally below 0.003 mg/L, although much higher levels have been detected in groundwater (0.023 mg/L), and 1 mg/L in contaminated groundwater (WHO 2004).

Removal methods Tetrachloroethene present in contaminated source waters can be removed by adsorption on to granular activated carbon, or by air stripping. Although considered to arise predominantly from industrial contamination in overseas waters, there is evidence that tetrachloroethene is also formed as a disinfection by-product.


When its appearance in a water results from chlorination, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.

Tetrachloroethene, once present as a disinfection by-product, can be removed by adsorption on to granular activated carbon, or by air stripping.

Analytical methods

Referee method




Health considerations Results from animal studies indicate that tetrachloroethene is absorbed rapidly and completely from the gastro-intestinal tract. It is eliminated primarily by the lungs. In the body it is metabolised to trichloroacetic acid. The most notable acute effect of short-term exposure is depression of the central nervous system. Longer-term studies of up to three months, using mice and rats, reported weight loss, and some evidence of liver and kidney toxicity at high doses. Inhalation exposures have resulted in maternal and foetal toxicity in mice, rats and rabbits. In humans, oral doses of 4.2�6 g tetrachloroethene administered to patients to control worm infections caused central nervous system effects, such as inebriation, perceptual distortion and exhilaration. Several developmental effects, such as eye, ear, central nervous system, chromosomal, and oral cleft anomalies, were associated with exposure to tetrachloroethene and other solvents in contaminated drinking-water supplies. Inhalation exposures have been associated with reproductive effects in female dry cleaners, including menstrual disorders and spontaneous abortions. IARC has classified tetrachloroethene in Group 2A (probably carcinogenic to humans). It reportedly produces liver tumours in mice, with some evidence of mononuclear cell leukaemia in rats and kidney tumours in male rats. The overall evidence from studies conducted to assess the genotoxicity of tetrachloroethene, including induction of single-strand DNA breaks, mutation in germ cells, and chromosomal aberrations in vitro and in vivo, indicates that tetrachloroethene is not genotoxic.


Derivation of maximum acceptable value In view of the overall evidence for non-genotoxicity and evidence for a saturable metabolic pathway leading to kidney tumours in rats, it is considered appropriate to use a tolerable daily intake approach for the derivation of a MAV for tetrachloroethene in drinking-water. The no observable adverse effect level used as the basis of the derivation was indicated by a six-week gavage study in male mice and a 90-day drinking-water study in male and female rats.


where:

• no observable adverse effect level = 14 mg/kg body weight per day indicated by a six-week gavage study of hepatotoxic effects in male mice and a 90-day drinking-water study in male and female rats




The odour threshold in water is 0.3 mg/L.


• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for carcinogenic potential).






[Editors/authors?]. 1980. Water Chlorination: Chemistry, environmental impact and health effects 3: 1007�[add page number]. Chelsea, USA: Lewis Publishers Inc.

Jolley RL, Bull RJ, Davis WP, et al (eds). 1985. Water Chlorination: Chemistry, environmental impact and health effects 5: 1381�[add page number]. Chelsea: USA: Lewis Publishers Inc.


Toluene

WHO. 2003. Tetrachloroethene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/23).

Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of toluene in drinking-water should not exceed 0.8 mg/L. The maximum contaminant level (USEPA 2004) is 1 mg/L.



1 To source waters

Toluene may be found in source waters due to industrial discharges as a result of human activity or from natural sources. It occurs naturally as a component of crude oil, and is present in petrol. It is produced during petroleum refining and also occurs in natural gas, emissions from volcanoes, forest fires and cigarettes. Commercially it is used as a solvent, especially for paints, coatings, gums, oils, and resins. The average concentration of toluene in the Rhine River in the Netherlands is approximately 0.002 mg/L.


No known sources.


Toluene can leach from synthetic coating materials used to protect drinking-water storage tanks.

Forms and fate in the environment In surface water, toluene volatilises rapidly to the air. Biodegradation and sorption are less important for the removal of toluene from surface waters.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 301 zones, found toluene concentrations to range from �not detectable� (nd) to 0.020 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). Approximately 1% of all groundwater supplies in the United States have concentrations greater than 0.0005 mg/L. Concentrations of a few micrograms per litre have been found in surface water, groundwater and drinking-water; point emissions can lead to higher concentrations in groundwater (up to 1 mg/litre). It may also penetrate plastic pipes from contaminated soil.

Removal methods Toluene can be removed from water by adsorption on to granular activated carbon, or by air stripping. Significant oxidation can also be expected by ozonation, although oxidation products may themselves be a concern if present in high enough quantities. Toluene may leach from some compounds used to seal water reservoirs. This possible source may need to be considered in the event of unacceptable concentrations of toluene appearing in a supply.

Analytical methods

Referee method





Health considerations In humans, toluene is absorbed readily from the gastro-intestinal tract after oral intake, and is distributed preferentially in adipose tissue, successively followed by adrenals, kidneys, liver, and brain. It is metabolised rapidly in the liver to benzyl alcohol, benzoic acid, and to a lesser extent, phenols. Toluene has a low acute toxicity via the oral route. Virtually all data available for humans refer to exposure to toluene via the inhalational route. Upon acute exposure, the predominant effects were impairment of the central nervous system and irritation of mucous membranes. Fatigue and drowsiness were the most sensitive effects. The toxic effects of toluene after long-term exposure are basically the same. Controlled long-term studies via the oral or inhalational route are lacking. The acute oral toxicity is low. Toluene exerts embryotoxic and fetotoxic effects, but there is no clear evidence of teratogenic activity in laboratory animals and humans. Toluene generally did not exhibit mutagenic activity in tests on bacteria, yeast cells, and mammalian cells in vitro. Available evidence suggests that toluene should not be regarded as an initiating carcinogen and the International Agency for Research on Cancer has classified it in Group 3 (not classifiable as to its carcinogenicity in humans). Epidemiological studies on the occurrence of cancer as a consequence of exposure of human populations to toluene alone are lacking.

Derivation of maximum acceptable value A tolerable daily intake approach has been used to derive the MAV for toluene in drinking-water. The lowest observable adverse effect level used in the derivation is for hepatotoxicity effects in mice determined from a 13-week gavage study. The MAV for toluene in drinking-water was derived as follows:


where:

• lowest observable adverse effect level = 312 mg/kg body weight per day for hepatotoxicity in mice (normalised for five days/week dosing in derivation)




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the short duration of the study and use of a LOAEL instead of a NOAEL).

The taste threshold for toluene has been reported at concentrations between 0.04 and 0.12 mg/L, and the odour threshold at about 0.024 to 0.17 mg/L.






No information is available on processes that can be used to remove tributyltin oxide from water.


WHO. 2003. Toluene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/116).


Tributyltin oxide Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of tributyltin oxide in drinking-water should not exceed 0.002 mg/L (2 µg/L). The WHO (2004) did not establish a guideline value for tributyl tin because it is considered unlikely to occur in drinking-water.


1 To source waters

The group of compounds known as the organotins comprises a large number of compounds with different properties and applications. Of these the dialkyl and tributyltin compounds are the ones most likely to be found in raw water. The trisubstituted compounds are used in the preservation of materials (wood, stone, textiles), as biocides, and disinfectants. Tributyltins may be found in raw water and sediment as a result of their use as antifouling agents. The use of tributyl-organotin compounds, particularly tributyltin oxide, in antifouling paints has now been banned in a number of countries because it is extremely toxic to aquatic life. Tributyltin is also used as a biocide in boiler waters.

2 From the treatment process

No known sources.


No known sources.

Forms and fate in the environment There is little information available on the fate of organotins in the aquatic environment. Tributyltin is thought to accumulate in aquatic food chains.

Typical concentrations in drinking-water No data are available on the concentration of tributyltin in New Zealand drinking-water supplies. No overseas data are available for tributyltin concentrations in drinking-water.

Removal methods


Analytical methods

Referee method

Liquid�Liquid Extraction and Gas Chromatography with Flame Photometric Detection (Greaves and Unger 1988)


No alternative methods have been recommended for tributyltin because no methods meet the required criteria.

Available data suggest that organotins are poorly absorbed and they tend to be distributed primarily in the liver and kidney following oral administration in rodents. Following oral administration it appears that the principal route of excretion of organotins is in the faeces.

Tributyltin oxide is not genotoxic. Although one carcinogenicity study was reported in which neoplastic changes were observed in endocrine organs, the significance of these changes is considered questionable. The most sensitive end-point appears to be immunotoxicity.


Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV of tributyltin oxide (TBTO) in drinking-water. The most sensitive end-point appears to be immunotoxicity in a 17-month feeding study in rats related to suppression of resistance to the nematode Trichinella spiralis. The significance of this finding to humans is not completely clear, but the no observable adverse effects level is consistent within an order of magnitude with other levels from long-term toxicity studies. The MAV for tributyltin oxide in drinking-water was derived as follows:


where:

World Health Organization. 1993. Guidelines for Drinking-Water Quality (2nd ed) Volume 1: Recommendations.

• no observable adverse effect level = 0.025 mg/kg body weight per day in a 17-month study in rats related to suppression of resistance to the nematode Trichinella spiralis




• uncertainty factor = 100 (for intra- and interspecies variation).


Greaves J, Unger MA. 1988. A selected ion monitoring assay for tributyltin and its degradation products. Biomedical and Environmental Mass Spectrometry 15: 565�9.

IPCS. 1990. Tributyltin compounds. Environmental Health Criteria, 116. World Health Organization, International Programme on Chemical Safety.



Trichloroacetaldehyde

Trichloroacetaldehyde may be formed as a by-product during chlorination of water containing natural organic matter.


Revised July 2005. (Also called chloral hydrate.)

Maximum acceptable value (provisional) Based on health considerations, the concentration of trichloroacetaldehyde in drinking-water should not exceed 0.01 mg/L. The WHO (2004) guideline value is designated as provisional because of limitations of the available database.


1 To source waters

Trichloroacetaldehyde may enter source waters from industrial discharges. Hydrated trichloroacetaldehyde (chloral hydrate) is used as a sedative and hypnotic in human and veterinary medicine.


No known sources.

Form and fate in the environment There is little information available on the environmental fate of trichloroacetaldehyde.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 511 zones, found trichloroacetaldehyde concentrations to range from �not detectable� (nd) to 0.030 mg/L, with the median concentration being �nd� (limit of detection = 0.002 mg/L). Found in drinking-water at concentrations of up to 0.10 mg/L (WHO 2004).

Removal methods No information is available on how trichloroacetaldehyde might be removed from contaminated source waters.

However, as this compound arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina orion exchange resins, however these methods are generally more expensive than coagulation.


Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible.

Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.



Analytical methods

Referee method


No alternative methods have been recommended for trichloroacetaldehyde because no methods meet the required criteria.

Health considerations Trichloroacetaldehyde is absorbed rapidly in humans with most of the dose being excreted in the urine as trichloroethanol glucuronide. Trichloroacetaldehyde has been used widely as a sedative or hypnotic drug in humans at recommended oral doses of 0.25�1.0 g. Concentrated solutions are quite irritating to the gastro-intestinal tract, and ingestion of undiluted preparations causes nausea and vomiting. Adverse effects in patients given either 0.5 or 1.0 g trichloroacetaldehyde included central nervous system depression, minor sensitivity reactions, gastro-intestinal disturbances and central nervous system excitement. Trichloroacetaldehyde induced arrhythmias have been described. The chronic use of trichloroacetaldehyde may result in development of intolerance, physical dependence, and addiction.

The carcinogenicity of trichloroacetaldehyde has been investigated in a two-year drinking-water using mice and it has been reported to cause liver tumours. Trichloroacetaldehyde has been shown to be mutagenic in short-term in vitro but does not bind to DNA. It caused disruption of chromosome segregation in cell division.

A tolerable daily intake approach has been used for the derivation of the MAV for trichloroacetaldehyde in drinking-water. The lowest observable adverse effect level used in the derivation is based on a study in which liver effects were seen in mice which received trichloroacetaldehyde in drinking-water for 90 days. The provisional MAV for trichloroacetaldehyde in drinking-water was derived as follows:

16 mg/kg body weight per day x 70 kg x 0.2 = 0.011 mg/L (rounded to 0.01 mg/L) 2 L x 10,000

where:

• lowest observable adverse effect level = 16 mg/kg body weight per day based on a study in which liver effects were seen in mice which received chloral hydrate in drinking-water for 90 days





• uncertainty factor = 10,000 (100 for intra- and interspecies variation and 10 for the short duration of the study and 10 for the use of a LOAEL instead of a NOAEL).


USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water, Supplement 1, Report No EPA/600/4-90-020.

WHO. 2003. Chloral Hydrate (Trichloroacetaldehyde) in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/49).


Trichloroacetic acid

Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of trichloroacetic acid in drinking-water should not exceed 0.2 mg/L. The maximum contaminant level for the five haloacetic acids (USEPA 2004) is 0.06 mg/L.


1 To source waters

Trichloroacetic acid may enter raw water through its industrial, agricultural and domestic use. It may be used as an intermediate in the synthesis of organic chemicals, laboratory reagent, herbicide, soil steriliser, and antiseptic.


Chlorinated acetic acids are formed from organic material during water chlorination.


No known sources.

Form and fate in the environment There is little information available on the environmental fate of trichloroacetic acid.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 488 zones, found trichloroacetic acid concentrations to range from �not detectable� (nd) to 0.115 mg/L, with the median concentration being �nd� (limit of detection = 0.005 mg/L). Detected in US groundwater and surface water distribution systems at mean concentrations of 0.0053 mg/L (range <0.001�0.08 mg/L) and 0.016 mg/L (range <0.001�0.174 mg/L), respectively; the maximum concentration was 0.20 mg/L, measured in chlorinated water in Australia (WHO 2004).


Removal methods No information is available on how trichloroacetic acid might be removed from contaminated source waters. However, as this compound arises in waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina or ion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.

Analytical methods

Referee method

Ion Exchange Liquid�Solid Extraction and Gas Chromatography with Electron Capture Detection (EPA 552.1).


1 Micro Liquid�Liquid Extraction Gas Chromatographic Method (APHA 6233B).

Health considerations Trichloroacetic acid is absorbed rapidly from the intestinal tract and metabolism occurs primarily in the liver. Trichloroacetic acid can be converted to carbon dioxide and chloride ion, or reduced to the aldehyde. Relatively small proportions of trichloroacetic acid are metabolised, and much of this compound is excreted unchanged in the urine. Trichloroacetic acid has been shown to induce tumours in the liver of mice. It has given mixed results in in vitro assays for mutations and chromosomal aberrations, and has been reported to cause chromosomal aberrations in in vivo studies. IARC has classified trichloroacetic acid in Group 3, not classifiable as to its carcinogenicity to humans. The weight of evidence indicates that trichloroacetic acid is not a genotoxic carcinogen.

Derivation of maximum acceptable value Because the weight of evidence on the carcinogenicity of trichloroacetic acid is restricted to one species, a tolerable daily intake approach has been used for the derivation of the MAV for trichloroacetic acid in drinking-water. The no observed adverse effect level used in the derivation is based on a study in which decreased body weight, increased liver serum enzyme activity and liver histopathology were seen in rats exposed to trichloroacetate in drinking-water for two years. The MAV for trichloroacetic acid in drinking-water was derived as follows:



where:


References


• no observable adverse effect level = 32.5 mg/kg body weight per day from a two-year study • average weight of an adult = 70 kg • proportion of tolerable daily intake allocated to drinking-water = 0.2

• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for database deficiencies, including the absence of a multigeneration reproductive study, the lack of a developmental study in a second species, and the absence of full histopathological data in a second species).

WHO (2004) emphasised that difficulties in meeting the guideline value must never be a reason for compromising adequate disinfection.



USEPA Method 552. 1990. Determination of Haloacetic Acids in Drinking Water by Liquid�Liquid Extraction, Derivatization, and Gas Chromatography with Electron Capture Detection. Cincinnati, USA: United States Environmental Protection Agency, Environmental Monitoring and Support Laboratory (EMSL).


WHO. 2003. Trichloroacetic Acid in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/120).


Trichloroacetonitrile Revised July 2005.

Maximum acceptable value WHO (2004) states that available data are insufficient to serve as a basis for derivation of a guideline value for trichloroacetonitrile.


1 To source waters

Trichloroacetonitrile has been used as an insecticide and therefore may enter raw water as a contaminant.


Haloacetonitriles such as trichloroacetonitrile are formed from organic precursors during chlorination or chloramination of drinking-water. In general, increasing temperature and/or decreasing pH have been associated with increasing concentrations of halogenated acetonitriles. Ambient bromide levels appear to influence, to some degree, the speciation of halogenated acetonitrile compounds. Dichloroacetonitrile is by far the most predominant halogenated acetonitrile species detected in drinking-water.



No known sources.

Form and fate in the environment Haloacetonitriles are reported to undergo hydrolysis in water, yielding nonvolatile products.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 209 zones, did not find trichloroacetonitrile at detectable concentrations (limit of detection = 0.0002 mg/L).

Trichloroacetonitrile concentrations are likely to be much less than 0.001 mg/L (WHO 2004).

Removal methods No information is available for methods to remove trichloroacetonitrile from contaminated source waters. However, as this compound arises in water principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine.

Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina or ion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products.

Analytical methods

Referee method



No alternative methods have been recommended for trichloroacetonitrile because no methods meet the required criteria.

Health considerations Haloacetonitriles are rapidly absorbed from the gastro-intestinal tract and metabolised to single carbon compounds. Insufficient data are available to determine whether haloacetonitriles can accumulate in specific organs. No data are available on the health effects of haloacetonitriles in humans. In a study in rats, trichloroacetonitrile decreased the percentage of females delivering litters and increased the percentage of foetal resorptions. Mean birth weights were reduced and postnatal survival was


significantly reduced. Another study reported numerous cardiovascular and urogenital malformations in surviving foetuses. IARC has concluded that dichloro-, dibromo-, bromochloro- and trichloroacetonitrile are not classifiable as to their carcinogenicity in humans. Dichloroacetonitrile and bromochloroacetonitrile have been shown to be mutagenic in bacterial assays, whereas results for dibromoacetonitrile and trichloroacetonitrile were negative. All four of these halogenated acetonitriles induced sister chromatid exchange and DNA strand breaks and adducts in mammalian cells in vitro but were negative in the mouse micronucleus test.

The 1995 Guidelines stated �Based on health considerations, the concentration of trichloroacetonitrile in drinking-water should not exceed 0.001 mg/L (1 µg/L).� This was a provisional MAV, and it still applied in the DWSNZ 2000.


A tolerable daily intake approach had been used for the derivation of the MAV of trichloroacetonitrile in drinking-water. The no observable adverse effects level used in the derivation was for decreases in foetal weight, and viability and cardiovascular and urogenital malformations in a teratology study in rats. The MAV had been designated as provisional owing to the lack of long-term studies in the available data base. The MAV for trichloroacetonitrile in drinking-water had been derived as follows:


where:

• no observable adverse effect level = 1 mg/kg body weight per day for foetal resorption, decreased foetal weight and viability and numerous cardiovascular and urogenital malformations from a study in which rats were dosed on gestation days 6�18




• uncertainty factor = 5000 (100 for intra- and interspecies variation and 10 for the severity of the effects at doses above the NOAEL and five for limitations of the data base (ie, lack of a 90-day study)).

WHO (2004) states that available data are insufficient to serve as a basis for derivation of a guideline value for trichloroacetonitrile. The previous provisional guideline value of 0.001 mg/L was based on a developmental toxicity study in which trichloroacetonitrile was administered by gavage in tricaprylin vehicle, and a recent re-evaluation judged this study to be unreliable in light of the finding in a more recent study that tricaprylin potentiates the developmental and teratogenic effects of halogenated acetonitriles and alters the spectrum of malformations in the fetuses of treated dams. This unreliability has resulted in trichloroacetonitrile now being listed in Table A2.2 of the DWSNZ, which is a list of determinands for which health concerns have been raised but for which no MAV has been set.






1,3,5-trichlorobenzene



Trichlorobenzenes (total) Revised July 2005.

1,2,3-trichlorobenzene 1,2,4-trichlorobenzene

Based on health considerations, the concentration of trichlorobenzenes (total) in drinking-water should not exceed 0.03 mg/L. WHO (2004) stated that because trichlorobenzenes occur at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a health-based guideline value. The maximum contaminant level for 1,2,4-trichlorobenzene (USEPA 2004) is 0.07 mg/L.


1 To source waters

Trichlorobenzenes may be found in raw water through industrial and agricultural activity. Industrial grade trichlorobenzene is more than 90% 1,2,4-trichlorobenzene with the remainder 1,2,3-trichlorobenzene and is used as an intermediate in chemical synthesis, a solvent, a coolant, a synthetic transfer oil, a lubricant, and a heat transfer medium; it is also used in polyester dyeing, in termite preparations, and as an insecticide.


No known sources.


No known sources.

Fate and form in the environment The trichlorobenzenes are expected to adsorb to soils of high organic content and are not expected to leach appreciably into groundwater. Trichlorobenzenes will not hydrolyse and are unlikely to biodegrade significantly. Some evaporation may occur from soil surfaces. In water, trichlorobenzenes are likely to adsorb to sediments and bioconcentrate in aquatic organisms. Evaporation from water may be a significant removal process.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 294 zones, did not find trichlorobenzene at detectable concentrations (limit of detection = 0.0005 mg/L).


Removal methods The concentration of trichlorobenzenes in a water can be reduced by adsorption on to granular activated carbon.

Trichlorobenzenes are of moderate acute toxicity. After short-term oral exposure, all three isomers show similar toxic effects, predominantly on the liver and thyroid.

Analytical methods

Referee method




Health considerations General population exposure will primarily result from air and food.

All three isomers of trichlorobenzenes are readily absorbed following oral administration in rats. In rats and rabbits the trichlorobenzenes are metabolised to trichlorophenols and mercapturic acids. High concentrations of the parent compound were found in fat, skin and liver, whereas high levels of metabolites were found in kidney and muscle.

Chronic and carcinogenicity studies via the oral route have not been carried out, but trichlorobenzenes did not exhibit mutagenic activity in tests with bacteria.

Trichlorobenzenes are moderately toxic to humans when ingested or inhaled. They are skin, eye and respiratory tract irritants. There has been one report of aplastic anaemia in a woman chronically exposed to 1,2,4-trichlorobenzene from washing work clothes.

Derivation of maximum acceptable value Long-term toxicity and carcinogenicity studies via the oral route have not been carried out, but the available data suggest that all three isomers are not genotoxic. Therefore, a tolerable daily intake approach has been used for the derivation of the MAV for trichlorobenzenes in drinking-water. The no observable adverse effect level used in the derivation is based on liver toxicity identified in a 13-week rat study. The MAV for trichlorobenzenes in drinking-water was derived as follows:


where:

• no observable adverse effect level = 7.7 mg/kg body weight per day for liver toxicity identified in a 13-week rat study





• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for the short duration of the study).

The MAV calculated would be 0.03 mg/L for each isomer, but because of the similarity in the toxicity of the trichlorobenzene isomers, the MAV is for total trichlorobenzenes.


Revised July 2005.

Odour thresholds of 0.01, 0.005-0.03, and 0.05 mg/L have been reported for 1,2,3-trichlorobenzene, 1,2,4-trichlorobenzene and 1,3,5-trichlorobenzene respectively.

References





WHO. 2003. Trichlorobenzenes in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/117).


1,1,1-trichloroethane

Maximum acceptable value (provisional) Based on health considerations, the concentration of 1,1,1-trichloroethane in drinking-water should not exceed 2 mg/L. WHO (2004) stated that because 1,1,1-trichloroethane occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value. The maximum contaminant level for 1,1,1-trichloroethane (USEPA 2004) is 0.2 mg/L and 0.005 mg/L for 1,1,2-trichloroethane.


1 To source waters

1,1,1-Trichloroethane can be released to the aquatic environment as an industrial contaminant. It is reported to be used widely overseas as a cleaning solvent, as a solvent for adhesives, coatings and textile dyes, as a coolant and lubricant in metal cutting oils, and as a component in inks and drain cleaners.

Trichloroethane is listed as a controlled substance in the New Zealand Ozone Layer Protection Act, 1990 and can be imported only under permit. There has been little demand for 1,1,1-trichloroethane for several years so its occurrence in the New Zealand environment should have been decreasing for some time.


No known sources.



No known sources.

Forms and fate in the environment 1,1,1-Trichloroethane is found mainly in the atmosphere, where it may act as a source of ClOx, which plays a vital role in ozone photochemistry. It is moderately soluble in water and can be anaerobically dechlorinated to 1,1-dichloroethane. A low soil adsorption coefficient suggests that it is mobile in soils and readily migrates to groundwaters. It does not bioaccumulate in animals.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, found 1,1,1-trichloroehane concentrations to range from �not detectable� (nd) to 0.003 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). In the United States, 1,1,1-trichloroethane has been found occasionally in water supplies at concentrations ranging from 0.0002 mg/L to 0.02 mg/L.

Removal methods Some removal can be achieved by adsorption on to granular activated carbon, although the adsorption is relatively weak. Some removal is also achievable by air-stripping. 1,1,1-trichloroethane has been found in only a small proportion of surface waters and groundwaters, usually at concentrations of less than 0.02 mg/L; higher concentrations (up to 0.15 mg/L) have been observed in a few instances.

Analytical methods

Referee method




Health considerations 1,1,1-trichloroethane is absorbed rapidly from the lungs and gastro-intestinal tract, but only small amounts (about 6% in humans) are metabolised. Large oral doses of 1,1,1-trichloroethane have produced nausea, vomiting and diarrhoea in humans. Acute inhalation exposures result in neurological effects. In fatalities resulting from inhalation, acute congestion of the lungs, fluid build-up and fatty deposits in the liver were reported. High concentrations of 1,1,1-trichloroethane in air can cause respiratory failure and problem with heart rhythm. Chronic exposure to low levels of 1,1,1-trichloroethane had no effect on serum and urine chemistry parameters, which are indicative of liver and kidney damage in humans. In animals, long-term studies have reported diminished body-weight gains at high doses (above 350 mg/kg body weight). Liver tumours were observed in mice, but not in rats fed 1,1,1-trichloroethane


for two-years. However, the study reported a high number of accidental deaths in both the control group and the study groups, and the results may be insignificant. IARC has placed 1,1,1-trichloroethane in Group 3 (not classifiable as to its carcinogenicity to humans). 1,1,1-trichloroethane does not appear to be mutagenic.

Derivation of maximum acceptable value A health-based MAV can be calculated for 1,1,1-trichloroethane based on changes in the kidney that were consistent with hyaline droplet nephropathy observed in a 13-week oral study in male rats, and taking into account the short duration of the study. The MAV for 1,1,1,-trichloroethane in drinking-water was derived as follows:

0.6 mg/kg body weight per day x 70 kg x 0.1 = 2.1 mg/L (rounded to 2 mg/L) 2 L

where: • the calculation used a TDI of 0.6 mg/kg of body weight • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 0.1.




New Zealand Ozone Layer Protection Act 1990.



WHO. 2003. 1,1,1-trichloroethane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/65).


Trichloroethene Revised July 2005. (Also called trichloroethylene.)

Maximum acceptable value (provisional) Based on health considerations, the concentration of trichloroethene in drinking-water should not exceed 0.08 mg/L. WHO (2004) states that the guideline value is designated as provisional because of deficiencies in the toxicological database. The maximum contaminant level (USEPA 2004) is 0.005 mg/L.



1 To source waters

Trichloroethene can be released to the aquatic environment as an industrial contaminant. It is used in drycleaning, for degreasing of metal parts, as a solvent for fats, waxes, resins, oils, rubber, paints, varnishes, and as an inhalation analgesic and anaesthetic. Its use in industrialised countries has declined sharply since 1970.


Trichloroethene may be formed from naturally organic precursors during chlorination of drinking-water.


No known sources.

Forms and fate in the environment It is expected that exposure to trichloroethene from air will be greater than that from food or drinking-water. Trichloroethene is readily released to the atmosphere, where it is highly reactive. In water, trichloroethene is removed by biodegradation, and possibly with some partitioning to sediment and suspended organic matter. Trichloroethene in anaerobic groundwater may degrade to more toxic compounds, including vinyl chloride. Trichloroethene is highly mobile in soil and may leach into groundwater supplies. There is low to moderate bioconcentration of trichloroethene in aquatic species.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 505 zones, found trichloroethene concentrations to range from �not detectable� (nd) to 0.003 mg/L, with the median concentration being �nd� (limit of detection = 0.001 mg/L). Found mostly in groundwater from which it is not lost to air; mean concentration of 0.0021 mg/L in a survey of drinking-water; also present in 24% of 158 non-random samples collected in a groundwater supply survey at a median level of 0.001 mg/L and a maximum of 0.13 mg/L (WHO 2004).

Removal methods Trichloroethene present in contaminated source waters may be removed by adsorption on to granular activated carbon, or air stripping. Although considered overseas to arise predominantly from waters containing industrial contamination, there is evidence that trichloroethene is also formed as a disinfection by-product. When its appearance in a water results from chlorination, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina or ion exchange resins, however these methods are generally more expensive than coagulation.


Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps, ie, avoiding prechlorination wherever possible. Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. Trichlorethene, once present as a disinfection by-product, can be removed by adsorption on to granular activated carbon, or by air stripping.

Analytical methods

Referee method




Health considerations At least 80% of ingested trichloroethene is absorbed and is distributed widely, with the highest concentration being in body fat. Inhalation studies with humans show that 40-75% of the retained dose is metabolised. Trichloroethene is eliminated with a biological half-life of 1.5 hours while the metabolites are excreted more slowly. Transplacental diffusion has been demonstrated in humans following inhalation. Humans exposed to high concentrations of trichloroethene have experienced central nervous system depression. Oral exposure of humans to 15�25 mL (21�35 g) of trichloroethene resulted in vomiting and abdominal pain, followed by transient unconsciousness. Humans exposed occupationally to trichloroethene had an increase in serum transaminases, which indicates damage to the liver parenchyma. Neurological abnormalities were associated with occupational exposure to 14�85 ppm trichloroethene for one month to 15 years, including decreased appetite, sleep disturbances, ataxia, vertigo, headache, and short-term memory loss. Trichloroethene has been classified by the International Agency for Research on Cancer in Group 3 (not classifiable as to its carcinogenicity to humans). Although it induces lung and liver tumours in mice, there are no conclusive data that it causes cancer in other species. Trichloroethene is a weakly active mutagen in bacteria and yeast species.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for trichloroethene in drinking-water. The lowest observable adverse effect level used in the derivation is for minor effects on relative liver weight in a six-week study in mice. The MAV for trichloroethene in drinking-water was derived as follows:


where:


• lowest observable adverse effect level = 100 mg/kg body weight per day for minor effects on relative liver weight in a six-week study in mice (normalised for five days/week dosing in derivation)




• uncertainty factor = 3000 (100 for intra- and interspecies variation and 10 for limited evidence of carcinogenicity, and 3 in view of the short duration of the study and the use of a LOAEL rather than a NOAEL).







[Editors/authors?]. 1980. Water Chlorination: Chemistry, environmental impact and health effects 3: 1007�[page number?]. Chelsea, USA: Lewis Publishers Inc

Jolley RL, Bull RJ, Davis WP, et al (eds). 1985. Water Chlorination: Chemistry, environmental impact and health effects 5: 1381�[page number?]. Chelsea, USA: Lewis Publishers Inc.

WHO. 2003. Trichloroethene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/22).


2,4,6-trichlorophenol Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of 2,4,6-trichlorophenol in drinking-water should not exceed 0.2 mg/L.


1 To source waters

2,4,6-trichlorophenol may occur in raw water as an industrial contaminant or from agricultural use. It may be used in the production of tetra-, and pentachlorophenol, as a germicide, glue and wood preservative, and an antimildew agent.





No known sources.

Forms and fate in the environment Because chlorinated phenols are moderately water-soluble, weakly acidic, and have low vapour pressures, it is anticipated that volatilisation does not play a significant role in removing these chemicals from water. The importance of photolysis of trichlorophenols in the natural environment is unknown. Sorption of trichlorophenol to organic-rich sediments is likely to be a significant removal mechanism for trichlorophenols from water. Biodegradation has been reported. Aquatic biota may bioconcentrate chlorinated phenols with bioconcentration factors increasing with increasing chlorine substitution.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 494 zones, did not find 2,4,6-trichlorophenol at detectable concentrations (limit of detection = 0.005 mg/L). 2,4,6-trichlorophenol concentrations are generally less than 0.001 mg/L (WHO 2004).

Removal methods Chlorophenols can be removed from contaminated source water by adsorption on to activated carbon. The effectiveness of the processes is pH dependent. Greater adsorption occurs as the pH is lowered. However, as this compound arises in New Zealand waters principally as a disinfection by-product, the preferred method for minimising its formation is to reduce the concentration of natural organic matter (NOM) coming into contact with the chlorine. Removal of NOM can be achieved by coagulation/flocculation with aluminium or iron salts. In some cases, adequate removal of NOM may be attained using organic polyelectrolytes as coagulants. NOM can also be removed by adsorption on to activated carbon, activated alumina or ion exchange resins, however these methods are generally more expensive than coagulation. Some reduction in disinfection by-product formation can be achieved by introducing the disinfectant into the water after the water has passed through all treatment steps (ie, avoiding prechlorination wherever possible). Chlorinated disinfection by-product formation can be reduced by the use of an alternative disinfectant such as ozone or chlorine dioxide, although these too have their associated disinfection by-products. The formation of chlorophenols can be reduced largely by the use of chlorine dioxide in place of chlorine. Where minimising disinfection by-product formation cannot reduce the concentration of disinfection by-products to a satisfactory level, methods to remove the disinfection by-products themselves may be considered. Chlorophenols can be removed by adsorption on to activated carbon. The effectiveness of the processes is pH dependent. Greater adsorption occurs as the pH is lowered. Note that the application of chlorine-containing disinfectants to activated carbon adsorbers should be avoided because of the unknown health effects of compounds formed through surface reactions between adsorbed contaminants and the disinfectants.


Analytical methods

Referee method

Micro Liquid�Liquid Extraction Gas Chromatographic Method (APHA 6233B).


1 Liquid�Liquid Extraction Gas Chromatographic Method (APHA 6420).

2 Liquid�Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (APHA 6410B).

3 Acetylation Liquid�Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (EPA 1653).

Health considerations Chlorophenols are well-absorbed after oral administration and they readily penetrate the skin. Chlorophenols do not appear to accumulate in body tissues in rats but are metabolised rapidly and eliminated from the body, principally in urine. 2,4,6-Trichlorophenol induced lymphomas and leukaemias in male rats and hepatic tumours in male and female mice. The compound has not been shown to be mutagenic in the Ames test but has shown weak mutagenic activity in other in vitro and in vivo studies. The International Agency for Research on Cancer has concluded that 2,4,6-trichlorophenol is possibly carcinogenic to humans (Group 2B).

Derivation of maximum acceptable value A MAV for 2,4,6-trichlorophenol has been derived by applying the linearised multistage model to leukaemias in male rats observed in a two-year feeding study. The hepatic tumours found in this study were not used in risk estimation, owing to the possible role of contaminants in their induction. The concentration in drinking-water associated with an excess lifetime cancer risk of one per one hundred thousand (10-5) is 0.2 mg/L. Taste and odour thresholds for 2,4,6-trichlorophenol have been reported at concentrations of 0.002 and 0.3 mg/L respectively.





Organic Contaminants in New Zealand Potable Water Supplies. Report to Department of Health, New Zealand, 1992.



Trihalomethanes Revised July 2005.


Trihalomethanes are a group of related organic chemicals that result from the chlorination of water containing natural organic matter precursors, or as industrial contaminants in the raw water. The four trihalomethanes of importance to drinking-water are: • bromodichloromethane • bromoform • chloroform • dibromochloromethane. Refer to the individual datasheets for specific details of these determinands. This datasheet discusses these determinands as a group.

Maximum acceptable value To account for additive toxicity, the sum of the ratio of the concentration of each trihalomethane to its respective MAV should not exceed 1. Action to reduce THMs is encouraged, but must not compromise disinfection, as nondisinfected water poses significantly greater risk than THMs.

Sources to drinking-water (Refer to individual datasheets.)

Form and fate in the environment

(Refer to individual datasheets.)

(Refer to individual datasheets.)

Typical concentrations in drinking-water THMs are rarely found in raw water but are often present in finished water; concentrations are generally below 0.10 mg/L. In most circumstances, chloroform is the dominant compound.

Removal methods (Refer to individual datasheets.)

Analytical methods (Refer to individual datasheets.)

Health considerations (Refer to individual datasheets.)

Interpretation of maximum acceptable value The sum of the ratio of the concentration of each trihalomethane to its respective guideline value should not exceed 1. The MAVs for the four trihalomethanes are: • bromoform 0.1 mg/L • dibromochloromethane 0.15 mg/L (DBCM) • bromodichloromethane 0.06 mg/L (BDCM) • chloroform 0.2 mg/L


The sum of the individual MAVs does not take account of the additive toxicity of trihalomethanes. The following fractionation approach can be taken to determine the total trihalomethane �concentration�:

Concentration (bromoform) + Concentration (DBCM) + Concentration (BDCM) + Concentration (chloroform) MAV (bromoform) MAV (DBMC) MAV (BDCM) MAV (chloroform) ≤ 1 For example, say a water sample contained: • 0.025 mg/L bromoform • 0.008 mg/L dibromochloromethane • 0.01 mg/L bromodichloromethane • 0.13 mg/L chloroform

then the calculation becomes: 0.025 + 0.008 + 0.01 + 0.13 0.1 0.15 0.06 0.2 = 0.25 + 0.053 + 0.167 + 0.65 = 1.12

(ie, the total THMs in this sample are than greater than the MAV, even though none of the individual THMs exceeded its MAV).





Vinyl chloride Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of vinyl chloride in drinking-water should not exceed 0.0003 mg/L (0.3 µg/L). The maximum contaminant level (USEPA 2004) is 0.002 mg/L.


1 To source waters

Vinyl chloride can be released to the aquatic environment as an industrial contaminant. It is used primarily for the production of polyvinyl chloride (PVC), which is used extensively in the plastics, rubber, paper and glass industries. Since PVC is imported and not polymerised in New Zealand, the occurrence of vinyl chloride in the New Zealand environment is not expected to be extensive. Vinyl chloride may also be used as a co-monomer with vinyl acetate or vinylidene chloride and as a raw material in the manufacture of 1,1,1-trichloroethane and monochloracetaldehye.


Chlorinated ethylenes such as vinyl chloride may be formed during the chlorination of water.



Vinyl chloride may be found in drinking-water from distribution systems constructed with PVC pipe or in water stored for long periods in PVC containers.

Forms and fate in the environment Volatilisation, followed by atmospheric oxidation, is considered to be the primary removal mechanism for vinyl chloride from the aquatic environment. When released to the ground, it does not absorb on to soil but migrates readily to groundwater, where it may be degraded to carbon dioxide and chloride ion, or it may be persistent for several years. Vinyl chloride has been reported to be a degradation product of trichloroethylene and tetrachloroethylene in groundwater.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find vinyl chloride at detectable concentrations (limit of detection = 0.001 mg/L). Rarely detected in surface waters, the concentrations measured generally not exceeding 0.010 mg/L; much higher concentrations found in groundwater and well water in contaminated areas; concentrations up to 0.010 mg/L detected in drinking-water (WHO 2004). It has occasionally been detected in drinking water supplies that use PVC pipes in the United States and Germany, with a maximum reported concentration of 0.01 mg/L.

Removal methods There are no published descriptions of trialed processes for reducing vinyl chloride concentrations in water. However, the constants involved in air/water equilibria indicate that it could be air-stripped readily from water.

Analytical methods

Referee method




Health considerations Vinyl chloride is absorbed readily following ingestion or inhalation. The highest concentrations of metabolites are found in the liver, kidneys and spleen. Vinyl chloride is metabolised to chloroethylene oxide and can rearrange spontaneously to chloroacetaldehyde; both substances are highly reactive and mutagenic. Low doses administered by gavage are metabolised and eliminated primarily in the urine, whereas a substantial proportion of higher doses are excreted unchanged via the lung. No significant accumulation of vinyl chloride occurs in the body. In rats it is estimated to have a biological half life of 20 minutes. Vinyl chloride has exhibited mutagenic activity in a variety of tests on bacteria and mammalian cells. Vinyl chloride is a narcotic agent, and loss of consciousness can occur at 25,000 mg/m3. Effects of chronic inhalation exposure include Raynaud�s phenomenon, a painful disorder of the hands. No data are available on oral exposure in humans.


There is sufficient evidence of carcinogenicity of vinyl chloride in humans from industrial populations exposed to high concentrations, and the International Agency for Research on Cancer has classified vinyl chloride in Group 1 (carcinogenic to humans). Causal association between vinyl chloride exposure and angiosarcoma is sufficiently proved, and some studies suggest that vinyl chloride can also be associated with heptacellular carcinoma, brain tumours, lung tumours and malignancies of the lymphatic and haemotopoietic tissues. Evidence indicates that vinyl chloride metabolites are genotoxic, interacting directly with DNA. DNA adducts formed by the reaction of DNA with a vinyl chloride metabolite have also been identified. Occupational exposure has resulted in chromosomal aberrations, micronuclei and sister chromatid exchanges; response levels were correlated with exposure levels. Animal studies show vinyl chloride to be a multi-site carcinogen. Vinyl chloride administered orally or by inhalation to mice, rats and hamsters produced tumours in the mammary glands, lungs, Zymbal gland, and skin, as well as angiosarcomas of the liver and other sites.

Derivation of maximum acceptable value The MAV of 0.0003 mg/L was based on an application of a linear extrapolation by drawing a straight line derivation between the dose, determined using a pharmocokinetic model, resulting in tumours in 10% of animals in rat bioassays involving oral exposure and the origin (zero dose), determining the value associated with the upper-bound risk of 10-5 and assuming a doubling of the risk for exposure from birth. The results of the linear extrapolation are nearly identical to those derived using the linearised multistage model. As vinyl chloride is a known human carcinogen, exposure to this compound should be avoided as far as practicable, and levels should be kept as low as technically feasible. Vinyl chloride is primarily of concern as a potential contaminant from some grades of PVC pipe and is best controlled by specification of material quality.





IPCS. 1999. Vinyl Chloride. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 215).



WHO. 2003. Vinyl Chloride in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/119).



Xylenes Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of xylenes in drinking-water should not exceed 0.6 mg/L. The maximum contaminant level (USEPA 2004) is 10 mg/L.


1 To source waters

Xylenes (ortho, meta, and para isomers) may occur in raw water as contaminants from industrial activity. Xylenes occur naturally as a component of crude oil, and are present in petrol. They are used in the manufacture of insecticides, pharmaceuticals, detergents, paints, inks and adhesives.

2 From treatment process

No known sources.


Xylene can enter water from solvents used in adhesives for bonding plastic drinking-water fittings. They may also leach from some compounds used to seal water reservoirs. Xyelenes can also penetrate plastic pipe from contaminated soil.

Forms and fate in the environment Releases of xylenes to the environment are largely to air because of their volatile nature. The distribution to water has been calculated to be less than 1%. When xylenes are released to surface water, they volatilise to air very rapidly. Xylenes degrade in air and are also biodegraded readily in soils and surface waters. They can be degraded in aerobic groundwater, but no biotransformation is expected under anaerobic conditions.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 301 zones, found meta- and para-xylene concentrations to range from �not detectable� (nd) to 0.022 mg/L, with the median concentration being �nd�, and ortho-xylene concentrations to range from �nd� to 0.01 mg/L, with the median concentration being �nd� (Limits of detection for all isomers = 0.0005mg/L). Concentrations of up to 0.008 mg/L have been reported in surface water, groundwater and drinking-water; levels of a few milligrams per litre were found in groundwater polluted by point emissions (WHO 2004).

Removal methods Xylenes can be removed from water by adsorption on to granular activated carbon or by air stripping. Xylenes may leach from some compounds used to seal water reservoirs and to bond plastic drinking-water fittings. This possible source may need to be considered in the event of unacceptable concentrations of xylenes appearing in a supply.


Analytical methods

Referee method




Health considerations Exposure to xylenes is mainly from air, and exposure is increased by smoking. Data on absorption of xylene after ingestion are not available. However, xylene isomers are readily absorbed after inhalation, with retention of 60�65% in humans. The few data available indicate rapid intake of the compound after intake and metabolism principally to methylbenzoic acid. Xylene can cross the placenta. No data are available for the effects of ingestion. In acute inhalation studies, irritation of eyes and throat were observed. After short-term exposure, abnormalities in psychometric functions were observed. Controlled studies of longer duration are lacking. A two-year study using rats and mice reported decreased growth at high doses (500 mg/kg body weight per day) but no xylene-related lesions. There was no evidence of carcinogenicity in oral and skin administration studies using rats and mice, and xylene was not mutagenic in tests using bacteria and mammalian cells. The International Agency for Research on Cancer has concluded that xylene is not classifiable as to its carcinogenicity in humans (Group 3).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for xylene in drinking-water. The no observable adverse effect level used in the derivation is based on decreased body weight in a 103-week gavage study in rats. The MAV for xylenes in drinking-water was derived as follows:


where:

• no observable adverse effect level = 250 mg/kg body weight per day based on decreased body weight in a 103 week gavage study in rats (normalised for 5 days/week dosing in the derivation)




• uncertainty factor = 1000 (100 for intra- and interspecies variation and 10 for limited toxicological end-point).

The odour threshold for xylene isomers has been reported to range between 0.02 and 1.8 mg/L, and at 0.3 mg/L they produce a detectable taste







WHO. 2003. Xylenes in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/25).



2.3 Pesticides alachlor aldicarb aldrin/dieldrin atrazine azinphos methyl bentazone brodifacoum bromacil carbofuran chlordanes chlorothalonil chlorpyriphos chlortoluron cyanazine 2,4-D 2,4-DB DDT + isomers diazinon 1,2-dibromo-3-chloropropane 1,2-dibromoethane 1,2-dichloropropane 1,3-dichloropropane 1,3-dichloropropene (cis and trans) dichlorprop dimethoate diquat diuron endosulfan endrin

glyphosate

fenitrothion fenoprop

heptachlor and heptachlor epoxide hexachlorobenzene hexazinone isoproturon lindane malathion MCPA MCPB mecoprop metalaxyl methamidophos methomyl methoxychlor methyl parathion metolachlor metribuzin molinate oryzalin oxadiazon pendimethalin pentachlorophenol


permethrin phenylphenol phorate

pyriproxifen

2,4,5-T

1080

Loss of alachlor from the soil is primarily by microbial degradation, with some volatilisation and photodegradation. The literature reports that its field half life ranges from 7 to 49 days, with 15 days being a recommended average value. It is metabolised rapidly in plants and does not bioaccumulate. Under certain conditions, alachlor can leach beyond the root zone into the groundwater. Many alachlor degradation products have been identi.ed in soil.

picloram pirimiphos methyl pirimsulfuron methyl procymidone propanil propazine propoxur pyridate

quintozene simazine

terbacil terbuthylazine thiabendazole triclopyr trifluralin

Alachlor Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of alachlor in drinking-water should not exceed 0.02 mg/L. The maximum contaminant level (USEPA 2004) is 0.002 mg/L.


1 To source waters

Alachlor may enter source waters as a result of its application as a herbicide, used for pre- and post emergence control of most annual grasses and broad leaved weeds in various crops. The total annual usage of alachlor in NZ in the late 1980s was 42,000 kg, 41,000 kg of which was used in the North Island. As at August 2005, alachlor is registered for use in New Zealand.


It has a water solubility of 240 mg/L and a sorption coefficient of 170 mL/g.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of alachlor (limit of detection = 0.0002 mg/L).


Has been detected in groundwater and surface water; has also been detected in drinking-water at levels below 0.002 mg/L (WHO 2004).

Removal methods Adsorption on to granular activated carbon can be used to remove alachlor from water, and significant reduction can also be achieved by ozonation.


Referee method

Liquid�Solid Extraction and Capillary Column Gas Chromatography with Mass Spectrometry (EPA 525).


1 Liquid�Liquid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 505). 2 Liquid�Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector

(EPA 507).

Health considerations Alachlor is absorbed through the gastro-intestinal tract of rats and it is distributed mainly to the blood, spleen, liver, kidney and heart. Long-term exposure studies in dogs and rats reported hepatotoxicity at all dose levels for both species. The rat study also reported highly significant levels of ocular lesions in the mid and high dose groups (>42 mg/kg body weight per day), identified as the uveal degeneration syndrome. The International Agency for Research on Cancer has not evaluated alachlor. On the basis of available data, evidence for the genotoxicity of alachlor is considered to be equivocal. However, a metabolite of alachlor has been shown to be mutagenic. Available data from two studies in rats indicate clearly that this compound is carcinogenic, causing benign and malignant tumours of the nasal turbinate, malignant stomach tumours, and benign thyroid tumours.

Derivation of maximum acceptable value In view of the data on carcinogenicity, the MAV for alachlor was calculated by applying the linearised multistage model to data on the incidence of nasal tumours in rats. The MAV in drinking-water associated with an excess lifetime cancer risk of one per one hundred thousand (10-5) is a concentration of 0.02 mg/L.

References AWWA. 1990. Water Quality and Treatment (4th ed). Pontius FW, McGraw, Hill (eds).

Frank R, et al. 1987. Survey of farm wells for pesticide residues, Southern Ontario, Canada, 1981�1982, 1984. Archives of Environmental Contamination and Toxicology 16: 1�8.

Miltner RJ, Baker DB, Speth TF, et al. 1989. Treatment of seasonal pesticides in surface waters. JAWWA 81(1): 43�52.




Wauchope RD, et al. 1992. The SCS/ARS/CES pesticide properties database for environmental decision making. Reviews of Environmental Contamination and Toxicology 123: 1�164.

Wilcock RJ. 1989. Patterns of pesticide use in New Zealand, Part 1: North Island 1985�1988. Water Quality Centre Publication No 15.

Wilcock RJ, Close M. 1990. Patterns of pesticide use in New Zealand, Part 2: South Island 1986�1989. Water Quality Centre Publication No 16.

WHO. 2003. Alachlor in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/31).




Aldicarb Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of aldicarb in drinking-water should not exceed 0.01 mg/L. The maximum contaminant level (USEPA 2004) is 0.003 mg/L.


1 To source waters

Aldicarb may enter source waters as a result of its application as a systemic carbamate insecticide used to control nematodes in soil, and insects and mites on a wide variety of crops. Aldicarb is not registered for use in New Zealand.

Forms and fate in the environment The sulphur atom in aldicarb is oxidised to sulphoxide and sulphone groups. Various oximes, nitriles, amides, acids and alcohols are also formed. Its field half life ranges from 7 to 80 days with 20 days being a recommended average value. Aldicarb and its degradation products are generally mobile in soil. It is very persistent in groundwaters; the half life for degradation to non-toxic products ranges from a few weeks to several years. The water solubility is 6000 mg/L and the sorption coefficient is 30 mL/g.

No data are available on the concentration of aldicarb in New Zealand drinking-water supplies. Aldicarb has been found in Canadian waters at concentrations up to 0.028 mg/L, and in waters from the USA in concentrations ranging from <0.01 to 0.5 mg/L.

Removal methods Adsorption on to granular activated carbon can be used to remove aldicarb from water.

Referee method

Reverse Phase High Performance Liquid Chromatography (EPA 531.1).



No alternative methods have been recommended for aldicarb because no methods meet the required criteria.

Health considerations Aldicarb is absorbed rapidly from the gastro-intestinal tract, the respiratory tract and the skin. It is metabolised rapidly and excreted. Aldicarb does not accumulate in tissues, but it appears to cross the placental barrier.

Aldicarb is highly acutely toxic in animals and it is one of the most acutely toxic pesticides. Poisoning in humans has resulted from ingestion of contaminated cucumbers and melons, the latter at a dose as low as 0.0021 mg/kg body weight. In humans, clinical symptoms of aldicarb intoxication include dizziness, weakness, diarrhoea, nausea, vomiting, abdominal pain, excessive perspiration, blurred vision, headache, muscular convulsions, temporary paralysis of the extremities, and dyspnea. Recovery is rapid, usually within six hours. The only toxic effect observed consistently with both long-term and single-dose administration of aldicarb in studies conducted to date is the rapidly reversible inhibition of acetylcholinesterase activity. The toxic effects of aldicarb appear to be dependent on the method and vehicle of administration.

The International Agency for Research on Cancer has concluded that aldicarb is not classifiable as to its carcinogenicity (Group 3).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for aldicarb in drinking-water. The No-observable-adverse-effects level (NOAEL) used for the MAV derivation is for acetylcholinesterase inhibition found in a 29-day study in rats administered drinking-water containing a 1:1 ratio of aldicarb sulfoxide and aldicarb sulfone. This study is considered to be the most relevant to the derivation of a drinking-water guideline because the rats were administered a ratio of aldicarb metabolites similar to that normally found in drinking-water. The MAV for aldicarb in drinking-water was derived as follows:

0.4 mg/kg body weight/day x 70 kg x 0.1 = 0.014 mg/L (rounded to 0.01 mg/L) 2 L/day x 100

where: • no observable adverse effect level = 0.4 mg/kg body weight per day • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 0.1 • uncertainty factor = 100 (for intra- and interspecies variation. No allowance has been made for the

short duration of the study in view of the extremely sensitive and rapidly reversible biological end-point used).


FAO/WHO. 1993. Pesticide Residues in Food � 1992. Rome: Food and Agriculture Organization of the United Nations, Joint FAO/WHO Meeting on Pesticide Residues (Report No 116).



Moye HA, Miles CJ. 1988. Aldicarb contamination of groundwater. Reviews of Environmental Contamination and Toxicology 105: 99.



[Authors?]. [Year?]. Environmental decision making. Reviews of Environmental Contamination and Toxicology 123: 1�164.

WHO. 2003. Aldicarb in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/72).


Zhong WZ, Lemley AT, Spalik J. 1984. Quantitative determination of ppb levels of carbamate pesticides in water by capillary gas chromatography. Journal of Chromatography 299: 269.

Aldrin/dieldrin Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of aldrin/dieldrin in drinking-water should not exceed 0.00004 mg/L (0.04 µg/L).

Sources to drinking-water Aldrin and dieldrin may enter source waters as a result of their use as highly effective insecticides for soil dwelling pests and for protection of wooden structures against termites and borer. The use of aldrin and dieldrin has been restricted severely, or banned, in many countries since the early 1970s. Aldrin and dieldrin are not currently used in New Zealand but they have been used in the past. Their registration was cancelled in 1989.

Forms and fate in the environment In soil, aldrin is removed by oxidation to dieldrin and by evaporation. In temperate countries like New Zealand, about 75% of aldrin is oxidised within a year. The further disappearance of dieldrin is very slow, with a half life of about five years. In living animals, aldrin is oxidised rapidly to dieldrin, which is oxidised in the liver and eliminated. The water solubilities for aldrin and dieldrin are 0.027 and 0.186 mg/L respectively.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 346 zones, found aldrin + dieldrin concentrations to range from �not detectable� (nd) to 0.00003 mg/L, with the median concentration being �nd� (limit of detection = 0.00001 mg/L). Aldrin and dieldrin have occasionally been detected in large Australian water storages where the maximum concentration reported is less than 0.001 mg/L, but the frequency of detection has dropped markedly since 1980.


Removal methods Adsorption on to granular activated carbon can be used to remove aldrin/dieldrin from water. Approximately 50% removal can be attained through conventional treatment using alum. Destruction can be achieved through oxidation by ozone.


Referee method

Liquid�Liquid Extraction and Gas Chromatography with Electron Capture Detector (APHA 6630C).


1 Liquid�Liquid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 505). Aldrin and dieldrin are listed under the Stockholm Convention on Persistent Organic Pollutants. Hence, monitoring may occur in addition to that required by drinking-water guidelines.

Health considerations Aldrin and dieldrin are absorbed by ingestion, inhalation and skin contact and tend to accumulate in adipose tissue. Aldrin and dieldrin can be metabolised from the adipose tissue compartment, causing an increase in blood level that results in toxic effects. The major metabolite is 9-hydroxy dieldrin. Both aldrin and dieldrin are highly toxic to humans. The target organs are the central nervous system and the liver. Severe cases of both accidental and occupational poisoning, and a number of fatalities have been reported. The lethal dose of dieldrin is estimated to be approximately 10 mg/kg body weight per day. The majority of individuals intoxicated with aldrin and/or dieldrin usually recover, and irreversible effects of residual pathology have not been reported.

The majority of studies on aldrin and dieldrin have not shown mutagenicity. The International Agency for Research on Cancer has classified dieldrin in Group 3 (not classifiable as to its carcinogenicity to humans). All the available information on aldrin and dieldrin taken together, including studies on humans, support the view that for practical purposes these chemicals make very little contribution to the incidence of cancer in humans.

Derivation of maximum acceptable value Due to the fact that dieldrin is considered to contribute very little, if any, to the incidence of cancer in humans, a tolerable daily intake approach has been used for the derivation of the MAV for aldrin/dieldrin in drinking-water. In 1977, the Joint FAO/WHO Meetings on Pesticide Residues (JMPR) recommended an ADI of 0.1 µg/kg of body weight (combined total for aldrin and dieldrin). This was based on a no observable adverse effect level of 1 mg/kg of diet in the dog and 0.5 mg/kg of diet in the rat, which are equivalent to 0.025 mg/kg of body weight per day in both species. JMPR applied an uncertainty factor of 250 based on concern about carcinogenicity observed in mice. WHO has reaffirmed this ADI which has been used for the derivation of the MAV for aldrin/dieldrin. The MAV for aldrin/dieldrin (ie, the sum of) in drinking-water was derived as follows:


where:


• no observable adverse effect level = 0.025 mg/kg body weight per day based on two studies on dogs and rats



• proportion of tolerable daily intake allocated to drinking-water = 0.01. Such a low percentage of the ADI was considered inappropriate for Australia, where the use of these pesticides has been severely restricted, so they adopted 0.1, which led to a 0.0003 mg/L limit

• uncertainty factor = 250 (based on concern about carcinogenicity observed in mice).



FAO/WHO. 1995. Pesticide residues in food � 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and WHO Toxicological and Environmental Core Assessment Groups. Rome, Food and Agriculture Organization of the United Nations (FAO Plant Production and Protection Paper 127).



USEPA. 1989. Drinking-water Health Advisory: Pesticides. United States Environmental Protection Agency Office of Drinking-water Health Advisors. Lewis Publishers.

WHO. 2003. Aldrin and Dieldrin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/73).


Atrazine Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of atrazine in drinking-water should not exceed 0.002 mg/L (2 µg/L). The maximum contaminant level (USEPA 2004) is 0.003 mg/L.


1 To source waters

Atrazine may enter source waters as the result of its use as a pre- and post-emergence herbicide for the control of annual grass and broad leaved weeds in various crops. It is also used in forestry and for non-selective weed control in non-crop areas. It has been reported in groundwater supplies at concentrations up to 0.002 mg/L in an area in Australia where atrazine was used over a 10-year period to suppress weed growth in irrigation channels (at application rates of 2�4 kg per hectare per year).


The total annual usage of atrazine in NZ in the late 1980s was 73,200 kg, mainly in the North Island. The highest usage of 11,100 kg was mainly on forestry in Taupo County which has an area of 393,000 ha. As at August 2005, atrazine is registered for use in New Zealand.

Forms and fate in the environment Atrazine can be degraded in surface water by photolysis and micro-organisms with half lives greater than 100 days. Hydrolysis and microbial degradation take place in the soil with half lives ranging from 18 to 120 days. The recommended average soil half life is 60 days. Degradation rates normally decrease with increasing depth and atrazine can be reasonably stable in groundwater. Atrazine and its dealkylated metabolites are moderately to very mobile in soils but the hydroxytriazine metabolites show low mobility and long persistence in soil. Degradation half lives of atrazine in soil ranged from 12 to 213 days over a wide geographical range of forestry sites in Australia; degradation rates were primarily dependent upon soil temperature (FHMG 2000). The water solubility is 33 mg/L and the sorption coefficient is 100 mL/g.

Typical concentrations in drinking-water Atrazine was not detected (approximately < 0.003 mg/L) in all of 230 samples from 212 drinking-water supplies sampled in New Zealand between 1988 and 1992. However, it has been found in groundwater in the Gisborne area (Close 1993). The well was sampled 3 times with levels of 0.037 mg/L, less than detection and 0.0021 mg/L (2.1 µg/L) being measured. Atrazine has also been found at low levels in several wells in the South Canterbury area and the Canterbury Regional Council is investigating the source of this atrazine. The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of atrazine (limit of detection = 0.0001 mg/L).

Removal methods Removal of atrazine can be achieved by adsorption on to granular activated carbon, and by ion exchange, ozone oxidation, and ultraviolet irradiation.


Referee method

Atrazine causes moderate irritation to rabbit skin, but is not appreciably irritating to the rabbit eye. It causes dermal sensitisation in the guinea pig.



1 Liquid�Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector (EPA 507).

Health considerations Atrazine appears to be absorbed readily from the gastro-intestinal tract. Following ingestion by rats, atrazine is retained mainly in red blood cells, liver, spleen and kidney. Most of the same metabolites found in soil can be found in degradation products in rats, with 2-chloro-4,6-diamino-1,3,5-triazine being the major urinary component.


An epidemiological study in northern Italy reported an increased relative risk of ovarian neoplasia (tumour formation) among women exposed to triazine herbicides. An 80% formulation of atrazine did not cause skin sensitisation upon repeated application to humans. The weight of evidence from a variety of genotoxicity assays indicates that atrazine is not genotoxic. There is evidence that atrazine can induce mammary tumours by hormonal changes in rats. It is highly probable that the mechanism for this process is non-genotoxic. No significant increase in neoplasia was observed in mice. The International Agency for Research on Cancer has concluded that atrazine is a possible human carcinogen (Group 2B). Desethyl atrazine (DEA) and desisopropyl atrazine (DIA) are chlorinated metabolites of atrazine and are well recognised in the toxicology literature. DEA and DIA were specifically tested in male and female rats, with significant results. Female rats demonstrated reduced maternal weight gain when exposed to 25 mg/kg-day over days 6 to 10 of gestation. Male rats had significantly reduced prostate weight and seminal vesicles after DEA and DIA exposure at 25 and 50 mg/kg-day (respectively) on postnatal days 23�53 (Toxnet 2004).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for atrazine in drinking-water. The no observable adverse effect level used in the derivation is based on a carcinogenicity study in rats. The MAV for atrazine in drinking-water was derived as follows:


where:

• no observable adverse effect level = 0.5 mg/kg body weight per day based on a carcinogenicity study in rats



• proportion of tolerable daily intake allocated to drinking-water = 0.1. The Australian Guidelines use a factor of 0.5, and derive a limit of 0.04 mg/L

• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 to reflect potential neoplasia).

From available monitoring data, it appears that the major metabolites of atrazine (desethylatrazine, desisopropylatrazine, diaminochlorotriazine, hydroxyatrazine) may constitute approximately 50% of the total atrazine-derived triazine compounds in some ground and surface water samples (Lerch et al 1998). This has been taken into account in deriving the guideline value.


Close ME. 1993. Assessment of pesticide contamination of groundwater in New Zealand, 2: Results of groundwater sampling. New Zealand Journal of Marine and Freshwater Research 27: 267�73.

FHMG (Forest Herbicide Management Group). 2000. Results from a Nationwide Study on the Risk of Atrazine Contamination to Surface Water and Groundwater Resulting from Routine Application by Australian Forest Growers. Draft report to the National Registration Authority for Agricultural and Veterinary Chemicals by the Forest Herbicide Management Group.



Lerch RN, Blanchard PE, Thurman EM. 1998. Contribution of hydroxylated atrazine degradation products to the total atrazine load in midwestern streams. Environmental Science and Technology 32: 40�8.

Smith VR. 1993. Groundwater Contamination by Organic Chemicals in Canterbury. Canterbury Regional Council Report 93(20).






Azinphos methyl is used as an organophosphorus triazine insecticide and acaricide (ticks). It is registered for use in New Zealand and is available as a wettable powder or suspension concentrate. Azinphos methyl is applied aerially and by ground methods. Trade names are: Azinphos Methyl 35F and 50W, Cotnion 350 FL and Gusathion M35.

TOXNET. 2004. TOXNET Databases Summary. US National Library of Medicine, National Institutes of Health, Department of Health and Human Services.


Wilcock RJ, Close M. 1990. Patterns of Pesticide Use in New Zealand, Part 2: South Island 1986�1989. Water Quality Centre Publication No 16.

WHO. 2003. Atrazine in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/32).

Azinphos methyl New entry, July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of azinphos methyl in drinking-water should not exceed 0.004 mg/L (4 µg/L). WHO (2004) did not develop a guideline value.


1 To source waters

No information is available on the annual usage of specific active ingredients in New Zealand, although azinphos methyl is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Azinphos methyl is moderately soluble in water: 33 mg/L at room temperature (Health Canada 1996). Because it is hydrolysed easily, azinphos methyl is not considered a significant leacher (USEPA 1985, cited in Health Canada 1996). The half life of azinphos methyl in laboratory and natural water systems was found to be 30�70 days at pH 5.1 to 8.4 (Weiss and Gakstatter 1977, cited in Health Canada 1996).


There is no information available regarding the greatest source of exposure to azinphos methyl for New Zealanders (ie, dermal contact, inhalation, diet: food, water).


Azinphos methyl is absorbed from the gastrointestinal tract.

No Ministry of Health drinking-water surveys have included azinphos methyl, so typical concentrations in New Zealand drinking-waters are unknown. No information is available about concentrations of azinphos methyl in groundwaters or surface waters. Azinphos methyl was not found in a survey of drinking-water samples from four Canadian provinces (detection limits ranged from 0.002 to 1 µg/L).

Removal methods Specific information about the removal of azinphos methyl from water is unavailable. However, oxidation of triazines (azinphos methyl is a member of this chemical family) by ozone is reported to be effective (Chiron et al 2000). The water chemistry, in particular the alkalinity and pH, will affect the oxidation rate. Use of activated carbon following ozonisation should be considered to adsorb oxidation products. Nanofiltration (membrane technology) in water with a low natural organic matter concentration is reported to remove approximately 50% of atrazine and simazine (Agbekodo et al 1996). The percentage is increased to 90�100% when 3.6 mg/L of natural organic matter is present. Similar results may be expected for azinphos methyl as it is from the same chemical family.

Adsorption on to activated carbon is expected to achieve some removal of azinphos methyl, although a guide to the efficiency of the process cannot be provided.


Referee method

Liquid/liquid extraction/gas chromatography-electron capture detector (EPA 8141A).


None recommended.

Health considerations Orally administered azinphos methyl has a biological half life of eight to nine hours with 90% of the dose being eliminated within 48 hours in the urine or faeces. Exposure can be from dietary risk, worker risk and drinking-water risk. With regards to drinking-water, acute risk is of most concern and chronic exposure does not appear to be of concern. However, surface and groundwater monitoring studies are needed to refine current models and monitoring estimates, (USEPA 1999b).

Acute poisoning

Potential symptoms of overexposure to azinphos-methyl are miosis, aching eyes, blurred vision, lacrimation (weeping) and rhinorrhea (nasal discharge), headache, chest tightness, wheezing, laryngeal spasms, salivation, cyanosis (bluish cast to skin), anorexia, nausea, vomiting and diarrhoea, sweating, twitching, paralysis and convulsions, low blood pressure, cardiac irregularities (Merck & Co 1996).


The acute oral LD50 for male guinea pigs is 80 mg/kg (RSocC 1987) which suggests a relatively high oral toxicity compared with other pesticides.

Chronic exposure

Azinphos methyl is an organophosphate pesticide. It has been determined that the organophosphates share a common mechanism of toxicity, the inhibition of cholinesterase levels (USEPA 1999a and WHO 1985). There is a data gap for studies of the effect of azinphos methyl on metabolism, teratology, reproduction and mutagenicity. The International Agency for Research on Cancer (IARC) has not classified azinphos methyl for its ability to cause cancer.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for azinphos methyl in drinking-water, as follows:


where: • no observable adverse effect level = 0.125 mg / kg body weight per day • average weight of adult = 70kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10% • uncertainty factor = 100.

References Agbekodo KM, Legube B, Dard S. 1996. Atrazine and simazine removal mechanisms by nanofiltration: influence of natural organic matter concentration. Wat Res 34(11): 2535�42.

Chiron S, Fernandez-Alba A, Rodriguez A, et al. 2000. Pesticide chemical oxidation: state-of-the-art. Wat Res 34(2): 366�77.

Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Azinphos methyl (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.


Merck & Co. 1996. The Merck Index: An encyclopedia of chemicals, drugs and biologicals (12th ed). Merck Research Laboratories Division of Merck & Co Inc.

Suntio L, Shiu W, Mackay D, et al. 1988. Critical review of Henry�s Law consants for pesticides. Rev Environ Contam Toxicol 103: 1.

USEPA. 1985. EPA draft final list of recommendations for chemicals in the National Survey for Pesticides in Groundwater (August 1985). Chem Regul Rep 9(34): 988.


USEPA. 1999a. Overview of Azinphos Methyl Revised Risk Assessment.

USEPA. 1999b. www.epa.gov/pesticides/op/azinphos/azmsum.htm

Weiss C, Gakstatter J. 1977. The decay of anti-cholinesterase activity of organic phosphorus insecticides on storage in water of different pH. Proc 2nd Int Water Pollut Res Conf, Tokyo, 1964. Cited in National Academy of Sciences. Drinking-water and Health. Washington DC, USA: US National Research Council.


http://www.epa.gov/pesticides/op/azinphos/azmsum.htm

WHO. 1985. Datasheet on Pesticides: No 59, azinphos methyl. World Health Organization. VBC/PDS/DS/85.59.

Bentazone Revised July 2005.

2 High Performance Liquid Chromatography with a Photoiodide Array Ultraviolet Detector (EPA 555).

Maximum acceptable value (provisional) Based on health considerations, the concentration of bentazone in drinking-water should not exceed 0.4 mg/L. WHO (2004) states that because bentazone occurs in drinking-water at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a health-based guideline value.


1 To source waters

Bentazone may enter source waters as a result of its application as a contact herbicide, used on winter and spring cereals. It is absorbed at the leaves. The total annual usage of bentazone in New Zealand in the late 1980s was 8670 kg with the highest use being 4000 kg in Ashburton county. As at August 2005, bentazone is registered for use in New Zealand.

Forms and fate in the environment Bentazone is mobile in a range of soil types and has been found in both surface and groundwaters overseas. The mechanism for degradation in soil is not known but the metabolite, 2-amino-N-isopropyl benzamide has been found. Half lives under optimal conditions range from 1.5 to 15 weeks depending on soil type. At temperatures below 10oC half lives are greater than 20 weeks. The recommended average soil half life is three weeks. The water solubility is 500 mg/L for the acid and 230,000 mg/L for the sodium salt, and the sorption coefficient is 34 mL/g.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of bentazone (limit of detection = 0.0001 mg/L).

Removal methods No information is available on technologies capable of removing bentazone from water.


Referee method

Liquid/Solid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 515.2).


1 Liquid�Liquid Extraction and Gas Chromatography with Electron Capture Detector (APHA 6640B).


Health considerations In animal studies, bentazone was absorbed rapidly from the gastro-intestinal tract and distributed via the bloodstream to various organs and tissues. Liver and kidneys exhibited the highest activity, but no penetration across the blood-brain barrier was observed. Up to ninety percent of the dose was excreted in the urine within 24 hours as unchanged bentazone.

The acute toxicity of bentazone appears to be moderate to low. Rats subjected to acute exposure exhibited poor muscle coordination, tremor and breathing difficulties.

No cases of human poisoning have been reported following bentazone exposure. Long-term studies conducted in rats and mice do not indicate a carcinogenic potential, and a variety of in vitro and in vivo assays indicate that bentazone is not genotoxic.

Derivation of maximum acceptable value The Joint FAO/WHO Meetings on Pesticide Residues (JMPR) evaluated bentazone in 1991 and the ADI they established has been used as the basis for the derivation of the MAV for bentazone in drinking-water shown below. The no observable adverse effect level used in the derivation is based upon haematological effects at higher doses, derived from a two-year dietary study in rats. The MAV for bentazone in drinking-water was derived as follows:

10 mg/kg body weight/day x 70 kg x 0.1 = 0.35 mg/L (rounded to 0.4 mg/L) 2 L/day x 100

where:

• no observable adverse effect level = 10 mg/kg body weight per day based upon haematological effects at higher doses, derived from a two-year dietary study in rats




• uncertainty factor = 100 (for inter and intra-species variation). The WHO (1993) Guidelines established a health-based guideline value of 0.03 mg/L for bentazone. This guideline value was amended to 0.3 mg/L in the addendum to the Guidelines, published in 1998, based on new information on the environmental behaviour of bentazone and exposure from food. However, because bentazone occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a health-based guideline value.



FAO/WHO. 1999. Pesticide residues in food � 1998: Evaluations � 1998: Part II � Toxicology. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/01.12).





Wilcock RJ, Close M. 1990. Patterns of Pesticide Use in New Zealand, Part 2: South Island 1986�1989. Water Quality Centre Publication No 16.

There are insufficient data to determine a MAV for brodifacoum in drinking-water. WHO (2004) does not mention brodifacoum.

WHO. 2003. Bentazone in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/77).


Brodifacoum New entry August 2005. (Also called super-warfarin.)



1 To source waters

Brodifacoum, an anticoagulant rodenticide, is stable in the solid form. Its solubility in water is less than 10 mg/L at 20°C and pH 7. It is is effective against rats and mice, including warfarin-resistant strains. It is used in agriculture and urban rodent control as ready-to-use baits of low concentration (usually 0.005% brodifacoum). Brodifacoum is registered for use in New Zealand.

Forms and fate in the environment Brodifacoum does not enter the atmosphere, because of its low volatility. It is practically insoluble in water. Brodifacoum is strongly bound on soil particles and is not taken up by plants. The rate of degradation is relatively slow and depends on soil type. Brodifacoum appears to bind rapidly in the soil with very slow desorption and without leaching.

Typical concentrations in drinking-water No information is available on concentrations in air, water, and soil. Being slightly soluble in water, its use cannot be a significant source of water contamination.


Removal methods Being strongly adsorbed to soil suggests treatment systems that remove particulate matter should remove brodifacoum.

Analytical methods

Referee method

A referee method cannot be selected for brodifacoum because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for bromochloroacetonitrile for the above reason. However, the following information may be useful: Analytical methods for the determination of brodifacoum include liquid chromatography with fluorescence detection and high-performance liquid chromatography, with detection limits of 0.001 mg/L and 0.002 mg/kg, respectively.

Health considerations Brodifacoum is absorbed through the gastrointestinal tract, skin, and respiratory system. The major route of elimination in different species after oral administration is through the faeces. The liver is the main organ of accumulation and storage. Brodifacoum has been found mainly as an unchanged compound. After a single oral dose to rats, liver concentrations remained high and relatively constant for 96 h. Elimination from the liver is slow and biphasic with an initial rapid phase lasting from 2 to 8 days after dosing and a slower terminal phase with an elimination half life of 130 days. In accidentally poisoned patients, the plasma half life was found to be approximately 16�36 days.

Derivation of maximum acceptable value There are insufficient data to determine a MAV for brodifacoum in drinking-water. Exposure of the general population to brodifacoum through air, drinking-water, or food is unlikely and does not constitute a significant health hazard. Poisoning incidents may occur in cases of massive intentional or unintentional ingestion, or prolonged skin contact during manufacture and formulation.

References IPCS. 1995. Anticoagulant rodenticides. Environmental Health Criteria 175. Geneva: World Health Organization.

IPCS INCHEM. International Programme on Chemical Safety (IPCS) and Canadian Centre for Occupational Health and Safety (CCOHS). Information available at: http://www.inchem.org/documents/hsg/hsg/hsg093.htm. Core site at www.inchem.org

Bromacil New entry July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of bromacil in drinking-water should not exceed 0.4 mg/L (400 µg/L).


http://www.inchem.org/documents/hsg/hsg/hsg093.htm

http://www.inchem.org/


1 To source waters

Bromacil is used as a broad spectrum herbicide to control weeds. It is currently registered for use in New Zealand and is available as a wettable powder. Registered formulations are of three types: Chemagro Terminex-A (which also contains amitrole and diuron), Hyvar X; and Krovar I DF, which also contains diuron.

Bromacil is stable to hydrolysis under normal environmental conditions. The primary routes of dissipation appear to be photolysis in water under alkaline conditions and microbial degradation in anaerobic soil. Bromacil�s persistence is demonstrated by half lives of 124 to 155 days in the field dissipation studies (USEPA 1996).

Bromacil is applied mainly by sprayers including boom, hand-held, knapsack, compressed air, tank-type, and power sprayers. Bromacil is also applied using aerosol, shaker, or sprinkler cans. No information is available on the annual usage of specific active ingredients in New Zealand, although bromacil is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Bromacil is a uracil compound that is very soluble in water: 815 mg/L, (Hort Research 2000), and has a mobility (as Koc) of 32, which suggests a moderate level of adsorption to organic soil.

There is no information available regarding the greatest source of exposure to bromacil for New Zealanders (ie, dermal contact, inhalation, diet: food, water). Based on international studies, people may be exposed to residues of bromacil through diet because it is applied to citrus crops.

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included bromacil, so typical concentrations in New Zealand drinking-waters are unknown. Bromacil has been detected on one occasion at one location in groundwater monitoring conducted by Environment Canterbury in and close to the Level Plain area in South Canterbury (Close et al 2001). In the Waikato region, bromacil has been detected in groundwater at four sites at concentrations of 0.00002�0.00637 mg/L (Hadfield and Smith 1999). Bromacil has been detected in groundwater in the Edendale area (Southland) at concentrations ranging between 0.00009 and 0.00024 mg/L (Hughes 2000). No information on levels of bromacil in drinking-water in other countries is available.

Removal methods No information is available on the removal of bromacil from water. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of bromacil, although a guide to the efficiency of the process cannot be provided.

Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.



Referee method

Liquid/liquid extraction/gas chromatography-nitrogen/phosphorus detector (EPA 507).


None recommended.

Health considerations A number of studies show that uracils, the class of compounds to which bromacil belongs, are absorbed into the body from the gut and excreted primarily in the urine (Extoxnet 1996).

Acute poisoning

In studies using laboratory animals, bromacil is slightly toxic by the oral, dermal, and inhalation routes and has been placed in Toxicity Category IV (the lowest of four categories) for these effects. The herbicide is irritating to the skin, eyes and respiratory tract. When as little as 100 mg/kg of the herbicide was fed to dogs, it caused vomiting, watering of the mouth, muscular weakness, excitability, diarrhea, and dilation of the pupils of the eyes. Rats that were fed single doses of bromacil experienced initial weight loss, paleness, exhaustion, and rapid breathing (Occupational Health Services Inc 1991, cited in PMEP 2001). Within four hours of being given 250 mg/kg of this, or a related material (isocil), sheep became bloated and walked with stilted gaits (Gosselin et al 1984, cited in PMEP 2001). The acute oral LD50 for rats is 5200 mg/kg (RSocC 1987) which suggests a relatively low oral toxicity compared with other pesticides.

Chronic exposure

In a chronic feeding study using beagle dogs, bromacil reduced body weight gain. In another chronic study using rats, effects in addition to reduced body weight gain included (1) increased incidence of thyroid cysts in the high dose males; (2) enlargement of the thymus in high dose females; and (3) dose-related incidence of thyroid tumours in the males. Bromacil demonstrates some evidence of causing developmental toxicity effects in rats and rabbits. These effects are likely to be due to maternal toxicity from exposure to bromacil rather than from specific developmental toxicity of bromacil. Therefore the USEPA does not consider bromacil a developmental toxicant (USEPA 1996). The International Agency for Research on Cancer has not classified bromacil, but USEPA has classified it as a Group C possible carcinogen based on increases in incidence of liver tumours in male mice, and positive trends in thyroid tumours in male rats, and, to a lesser extent, structural activity relationship to similar compounds.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for bromacil in drinking-water, as follows:


where: • no observable adverse effect level = 10 mg/kg body weight per day • average weight of adult = 70kg • average quantity of water consumed by an adult = 2 L per day


• proportion of tolerable daily intake allocated to drinking-water = 10% • uncertainty factor = 100.

References Close ME, Rosen MR, Smith VR. 2001. Fate and transport of nitrates and pesticides in New Zealand�s aquifers. In: Rosen MR, White PA (eds). Groundwaters of New Zealand, Chapter 8, pp. 185�220. New Zealand: The Hydrological Society.

Extoxnet. 1996. Pesticide Information Profile: Bromacil. Oregon State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

Gosselin RE, et al. 1984. Clinical Toxicology of Commercial Products (5th ed). Baltimore: Williams and Wilkins.

Hadfield J, Smith D. 1999. Pesticide Contamination of Groundwater in the Waikato Region. Environment Waikato Technical Report 1999/9.

Hort Research. 2000. Guidelines for the Sustainable Application of Agrichemicals. Version 1.1 CD-ROM. Palmerston North: Hort Research.

Hughes B. 2000. Edendale Pesticide Investigation Report 2000. Southland Regional Council Publication No 2000-14.

Occupational Health Services Inc. 1991. MSDS for Bromacil. Secaucus, USA: OHS Inc.

PMEP. 2001. Pesticide Information Profile: Bromacil. Cornell University, USA: PMEP.

Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Bromacil (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.


USEPA. 1996. EPA RED Facts: Bromacil: Prevention, pesticides and toxic substances. EPA-738-F-96-013.

Carbofuran Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of carbofuran in drinking-water should not exceed 0.008 mg/L (8 µg/L). The maximum contaminant level (USEPA 2004) is 0.04 mg/L.


1 To source waters

Carbofuran may enter source waters as a result of its use as a systemic acaricide, insecticide and nematocide. The total annual usage of carbofuran in New Zealand in the late 1980s was 1180 kg, all of it in the North Island. As at August 2005, carbofuran is not registered for use in New Zealand.

Forms and fate in the environment Carbofuran can dissipate from water by direct photolysis and photo-oxidation. It undergoes chemical and microbial degradation mainly through hydroxylation and hydrolysis. Carbofuran is mobile in soils and sediments. It has a half life in soil ranging from 1 to 37 weeks. The recommended average half life is seven weeks. The water solubility is 350 mg/L (700 mg/L in Agrochemicals Handbook) and the sorption coefficient is 22 mL/g.


Typical concentrations in drinking-water No data are available on the concentration of carbofuran in New Zealand drinking-water supplies. Studies from the USA have found groundwater concentrations of carbofuran up to 0.03 mg/L. Typical detectable concentrations were in the 0.001 to 0.005 mg/L range.

Removal methods Carbofuran has been found to decompose at high pH levels, such as during high pH softening. Hydrolysis products are produced. Alum coagulation does not remove carbofuran from water, but granular activated carbon should be able to control its concentration down to µg/L levels.


Referee method

Reverse Phase High Performance Liquid Chromatography (EPA 531.1).


No alternative methods have been recommended for carbofuran because no methods meet the required criteria.

Health considerations Animal studies indicate that carbofuran is absorbed rapidly and metabolised by hydroxylation and/or oxidation reactions. Elimination of carbofuran is rapid, in urine. Symptoms of carbofuran poisoning in humans resembles parathion intoxication except for diminished intensity and duration, particularly of the central nervous system. In a case of acute intoxication in a woman with a total dose of 60 mg carbofuran, slight cholinesterase inhibition was found, but the patient recovered completely within 72 hours. Several cases of adverse effects have been reported in individuals involved in the application and formulation of carbofuran. Symptoms included mild and reversible symptoms of acetylcholinesterase depression, such as malaise, hypersalivation, and vomiting. Symptoms following more severe poisoning included chest tightness, muscular twitching, convulsions and coma. Human volunteers administered 0.1 mg/kg body weight carbofuran orally showed symptoms of acetyl-cholinesterase depression, including salivation, diaphoresis (sweating), abdominal pain, drowsiness, dizziness, anxiety and vomiting. No symptoms were observed in volunteers administered 0.05 mg/kg body weight. The International Agency for Research on Cancer has not evaluated carbofuran. However, on the basis of available studies, this compound does not appear to be carcinogenic. Carbofuran does not have mutagenic activity.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for carbofuran in drinking-water. The no observable adverse effect level used in the derivation is based on acute (reversible) effects in dogs in a short-term (4-week) study conducted as an adjunct to a 13-week study in which inhibition of erythrocyte acetylcholinesterase activity was observed. The MAV for carbofuran in drinking-water was derived as follows:



where:

• no observable adverse effect level = 0.05 mg/kg body weight per day based on inhibition of acetylcholinesterase




• uncertainty factor = 30 (10 for intra-species variation and 3 for the steep dose-response curve.

References Health Canada. Guidelines for Canadian Drinking Water Quality � Supporting Documents. See http://www.hc-sc.gc.ca/ewh-semt/pubs/water-eau/doc_sup-appui/index_e.html

Miltner RJ, Baker DB, Speth TF, et al. 1989. Treatment of seasonal pesticides in surface waters. JAWWA 81(1): 43�52.


The Royal Society of Chemistry. 1987. The Agrochemicals Handbook (2nd ed).






Chlordane Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of chlordane in drinking-water should not exceed 0.0002 mg/L (0.2 µg/L). The maximum contaminant level (USEPA 2004) is 0.002 mg/L.


1 To source waters

Chlordane may enter source waters as the result of its application as a versatile broad-spectrum contact insecticide used mainly for non-agricultural purposes. It has also been used in the timber preservation industry. Chlordane is no longer registered for use of in New Zealand but it was used extensively in the past.


Forms and fate in the environment Chlordane is a mixture of isomers, mainly cis and trans chlordane. Technical chlordane contains at least 26 different compounds, including 60�75% chlordane isomers, heptachlor and nonachlor. Chlordane is very resistant to chemical and microbial degradation. It is very immobile and dissipation from soils is mainly due to volatilisation. The soil half life is 1 to 4 years. Chlordanes are unlikely to migrate to groundwater, where they have been found only rarely. Once in water bodies it is not removed by photodegradation, hydrolysis or biodegradation. The water solubility of chlordane is 0.1 mg/L.

Typical concentrations in drinking-water Chlordane was not detected (<0.00004 mg/L, ie, <0.04 µg/L) in all of 230 samples from 212 supplies sampled in New Zealand between 1988 and 1992. The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of chlordane (limit of detection = 0.00001 mg/L). In the United States, chlordane has been detected rarely in drinking water, and when found, concentrations were usually below 0.0001 mg/L.

Removal methods Specific information concerning the removal of chlordane from water is not available. However, its low solubility makes it likely that some removal by chemical coagulation is possible. Isotherm adsorption data also indicate that removal by adsorption on to activated carbon should be possible.


Referee method

Liquid/Liquid Extraction and Gas Chromatography with Electron Capture Detector (APHA 6630C).


1 Liquid/Liquid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 508). Chlordane is listed under the Stockholm Convention on Persistent Organic Pollutants. Hence, monitoring may occur in addition to that required by drinking-water guidelines.

Health considerations Animal studies indicate that a portion of cis-chlordane ingested is absorbed. Chlordane and its metabolites, mainly oxychlordane, are distributed quickly throughout the body and stored at the highest levels in adipose tissue. Oxychlordane has been detected in adipose tissue of the general human population. Chlordane, mainly as oxychlordane, has been detected in human milk. Chlordane is moderately toxic in acute exposure. In animals, acute exposure to chlordane produces ataxia, convulsions, respiratory failure and cyanosis. In experimental animals, prolonged exposure in the diet causes liver damage. Chlordane produces liver tumours in mice, but the weight of evidence indicates that it is not genotoxic. Chlordane can interfere with cell communication in vitro, a characteristic of many tumour promoters.


Humans exposed accidentally to chlordane by inhalation or ingestion reported neurological symptoms, including headache, dizziness, vision problems, loss of coordination, irritability, excitability, weakness, muscle twitching and convulsions. Following ingestion of drinking-water contaminated with chlordane at concentrations of up to 1.2 g/L, 13 people showed gastro-intestinal and/or neurological symptoms. The International Agency for Research on Cancer re-evaluated (in 1991) the evidence for carcinogenicity in humans associated with chlordane and heptachlor. Available studies were inadequate to evaluate an association between human exposure to chlordane/heptachlor and carcinogenicity. Chlordane has been classified in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value The Joint FAO/WHO Meetings on Pesticide Residues (JMPR) re-reviewed chlordane in 1986 and the results of this review have been used for the derivation of the MAV for chlordane in drinking-water. The no observable adverse effect level used in the derivation is based on a long-term dietary study in rats for increased liver weights, serum bilirubin levels and incidence of hepatocellular swelling. The MAV for chlordane (total isomers) in drinking-water was derived as follows:


where:

• no observable adverse effect level = 0.05 mg/kg body weight per day based on a long-term dietary study in rats




• uncertainty factor = 100 (for inter and intra-species variation).



FAO/WHO. 1995. Pesticide Residues in Food � 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and WHO Toxicological and Environmental Core Assessment Groups. Rome, Food and Agriculture Organization of the United Nations (FAO Plant Production and Protection Paper 127).

Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466�[Add page number].



WHO. 2003. Chlordane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization.



Chlorothalonil New entry August 2005.

Maximum acceptable value There are insufficient data to determine a MAV for chlorothalonil in drinking-water. A guideline value was not derived in WHO 2004 because it was considered that chlorothalonil was unlikely to occur in drinking-water.


1 To source waters

Chlorothalonil is an aromatic halogen compound, a member of the chloronitrile chemical family. Chlorothalonil is a broad-spectrum organochlorine pesticide (fungicide) used to control fungi that threaten vegetables, trees, small fruits, turf, ornamentals, and other agricultural crops. It is registered for use in New Zealand. Chlorothalonil is also used in anti-fouling paints.

Forms and fate in the environment In aerobic soils, the half life for chlorothalonil is from 1-3 months. Increased soil moisture or temperature increases chlorothalonil degradation. It is not degraded by sunlight on the soil surface. Chlorothalonil has high binding and low mobility in silty loam and silty clay loam soils, and has low binding and moderate mobility in sand. In studies conducted in water over ten weeks, chlorothalonil, at low levels, was generally stable. In very basic water (pH 9.0), about 65% of the chlorothalonil was degraded into two major metabolites after 10 weeks.

Typical concentrations in drinking-water Chlorothalonil was not found in any of 560 groundwater samples collected from 556 sites. Chlorothalonil was reported to be found in one surface water location in Michigan at 6.5 mg/L (USEPA 1987).

Removal methods No information available.


Referee method

A referee method cannot be selected for chlorothalonil because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for chlorothalonil for the above reason. However, the following information may be useful: ???????????????


Health considerations Chlorothalonil is excreted rapidly from the body, primarily unchanged. It is not thought to be stored in animal tissues. Rats and dogs fed very high doses for two years eliminated almost all of the chemical in urine, faeces, and expired air. After two years, the amount of the breakdown product found in the liver tissues was considered insignificant in both dogs and rats (USEPA 1987). At lower concentrations, chlorothalonil leaves the body within 24 hours. Residues have not been found in the tissues or milk of dairy cows (Vettorazzi 1979). Chlorothalonil is not very water soluble and does not store in fatty tissues. Its bioaccumulation factor is quite low, about 425 times the background water concentration.

Acute toxicity

Chlorothalonil and its metabolites are highly toxic to fish, aquatic invertebrates, and marine organisms. Fish, such as rainbow trout (LC50 of 0.25 mg/L) are noticeably affected even when chlorothalonil levels are low (less than 1 mg/L). Chlorothalonil is slightly toxic to mammals, but it can cause severe eye and skin irritation in certain formulations (Walker and Keith). Very high doses may cause a loss of muscle coordination, rapid breathing, nose bleeding, vomiting, and hyperactivity. Dermatitis, vaginal bleeding, bright yellow and/or bloody urine, and kidney tumors may also occur, followed by death. The oral LD50 is >10,000 mg/kg for rats, and 6000 mg/kg for mice. The acute dermal LD50 for both albino rabbits and albino rats is 10,000 mg/kg.

Chronic toxicity

In a number of tests of varying lengths of time, rats which were fed a range of doses of chlorothalonil generally showed no effects on physical appearance, behavior, or survival. Kidney changes such as enlargement were common. In a two-year dietary rat study, the lowest dose of chlorothalonil that produced no adverse effects in the animals was 60 ppm (3 mg/kg). Human eye and skin irritation is linked to chlorothalonil exposure. Fourteen out of 20 workers exposed to 0.5% chlorothalonil in a wood preservative developed dermatitis. All workers showed swelling and inflammation of the upper eyelids. Allergic skin responses have also been noted in vegetable and in horticultural workers. In a long-term rat study, high doses fed to both males and females did not affect reproduction or the litters that were produced. However, body weight gains for males and females of each generation were decreased. Administration of high doses of chlorothalonil to pregnant rabbits through the stomach during the sensitive period of gestation resulted in four of the nine mothers aborting. These studies suggest that chlorothalonil will not affect human reproduction except at very high doses. Female rats given high doses of chlorothalonil through the stomach during the sensitive period of gestation had normal fetuses, even though that dose was toxic to the mothers. One study of birth defects in rabbits showed no effects. Thus, chlorothalonil is expected to produce no birth defects in humans. Mutagenicity studies on various animals, bacteria, and plants indicate that chlorothalonil does not cause any chromosomal changes. The compound is therefore not expected to pose mutagenic risks to humans.


Chlorothalonil is a potential human carcinogen, known to affect the kidney, ureter, and bladder in experimental animals. Male and female rats fed chlorothalonil daily over a lifetime developed carcinogenic and benign kidney tumors at the higher doses. In another study, where mice were fed high daily doses of chlorothalonil for two years, females developed tumors in the fore-stomach area (attributed to irritation by the compound) and males developed carcinogenic and benign kidney tumors. However, this latter study was inconclusive as to the relationship between dose of chlorothalonil and the presence of cancer in the test animals. Chronic studies of rats and dogs fed high dietary levels show that chlorothalonil is toxic to the kidney. In addition to less urine output, changes in the kidney included enlargement, greenish-brown color, and development of small grains.

Derivation of maximum acceptable value There are insufficient data to determine a MAV for chlorothalonil in drinking-water.

References Extoxnet. A Pesticide Information Project of Co-operative Extension Offices of Cornell University, Oregon State University, the University of Idaho, and the University of California at Davis and the Institute for Environmental Toxicology, Michigan State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

USEPA. 1987. Chlorothalonil Health Advisory: Draft report. Environmental Protection Agency, Office of Drinking Water.

Vettorazzi G. 1979. International Regulatory Aspects for Pesticide Chemicals. Boca Raton, USA: CRC Press.

Walker MM, Keith LH. 1992. US Environmental Protection Agency�s Pesticide Fact Sheet Database. Chelsea, USA: Lewis Publishers.


Chlorotoluron Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of chlorotoluron in drinking-water should not exceed 0.04 mg/L.


1 To source waters

Chlorotoluron may enter source waters as a result of its application as a pre- and early post-emergence herbicide, used to control annual grasses and broadleaved weeds in winter cereals. Chlorotoluron has not been used, and is not registered for use, in New Zealand.

Forms and fate in the environment Chlorotoluron is degraded slowly in water and is quite persistent. Half lives in water range from 80 to greater than 200 days and half lives in soil range from 1 to several months. The water solubility of chlorotoluron is 70 mg/L.


Typical concentrations in drinking-water Chlorotoluron has not been used in New Zealand. It has been detected occasionally in waters in the UK at concentrations ranging from 0.00044 to 0.00058 mg/L (0.44 to 0.58 µg/L). Chlorotoluron has been detected frequently in German raw waters (ground and surface waters), and in levels up to 0.0012 mg/L (1.2 µg/L) in drainage water, soon after normal application on fields in Germany.

Removal methods Specific information concerning the removal of chlorotoluron from water is not available. However, phenylamide (or urea) pesticides, such as chlorotoluron, are reported to be broken down by chlorination. Slow sand filtration does not appear to remove this class of pesticide. Other phenylamide pesticides have been reported to be broken down by ozone.


Referee method

Liquid/Solid Extraction Gas Chromatographic/Mass Spectrometric Method (EPA 525.2).


An HPLC method may be suitable (Crathorne et al 1987).

Health considerations Chlorotoluron is absorbed readily and rapidly following ingestion. No evidence of accumulation of chlorotoluron in any particular organ or tissue has been reported. It is excreted rapidly in the urine in the form of metabolites. Chlorotoluron is of low acute, short-term and long-term exposures in animals, but it has been shown to cause adenomas and carcinomas of the kidney in male mice given high doses for two-years. No cases of human poisonings have been reported following chlorotoluron exposure. Chlorotoluron and its metabolites have shown no evidence of genotoxicity. Available evidence from animal studies suggests that chlorotoluron has a carcinogenic potential that is both species and sex specific. No information is available concerning the carcinogenicity of chlorotoluron to humans.

Derivation of maximum acceptable value As chlorotoluron and its metabolites show no evidence of genotoxicity, a tolerable daily intake approach was used for the derivation of the MAV for chlorotoluron in drinking-water. The no observable adverse effect level used in the derivation is based on a two-year feeding study of systemic effects in mice. The MAV for chlorotoluron in drinking-water was derived as follows:

11.3 mg/kg body weight/day x 70 kg x 0.1 = 0.04 mg/L 2 L/day x 1000


where:

• no observable adverse effect level = 11.3 mg/kg body weight per day based on a two-year feeding study in mice




• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 for evidence of carcinogenicity).

References Crathorne B, James CP, Stratford JA. 1987. HPLC Method for the Analysis of Chlorotoluron, Isoproturon and Linuron in Water. Medmenham, United Kingdom: Water Research Centre.


WHO. 2003. Chlorotoluron in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/33).


Chlorpyriphos Revised July 2005. (Also called chlorprifos.)

Maximum acceptable value Based on health considerations, the concentration of chlorpyriphos in drinking-water should not exceed 0.04 mg/L.


1 To source waters

Chlorpyrifos is a broad-spectrum organophosphorus insecticide used for the control of mosquitos, flies, various crop pests in soil and on foliage, household pests and aquatic larvae. Athough it is not recommended for addition to water for public health purposes by WHOPES, it may be used in some countries as an aquatic larvicide for the control of mosquito larvae. The total annual usage of chlorpyriphos in New Zealand in the late 1980s was 116,500 kg. As at August 2005, chlorpyrifos is registered for use in New Zealand.

Forms and fate in the environment Chlorpyrifos is strongly absorbed by soil and does not readily leach from it, degrading slowly by microbial action. In soil, chlorpyriphos is degraded slowly, with a half life of approximately 80 to 100 days, and it undergoes further degradation to organochlorine compounds and carbon dioxide. It has a low solubility in water and great tendency to partition from aqueous into organic phases in the environment. The water solubility is 0.4 mg/L and the sorption coefficient is 6070 mL/g.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 342 zones, did not find detectable concentrations of chlorpyriphos (limit of detection = 0.0002 mg/L). Detected in surface waters in USA, usually at concentrations below 0.0001 mg/L; also detected in groundwater in less than 1% of the wells tested, usually at concentrations below 0.00001 mg/L (WHO 2004).

Removal methods No data available; should be amenable to treatment by coagulation (10�20% removal), activated carbon adsorption, and ozonation.

Recommended analytical technqiues

Referee method

No referee method has been given for chlorpyriphos because no method meets the required criteria.


1 Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector or Flame Photometric Detector (HMSO 1986).

Health considerations Organophosphates are absorbed readily through the skin, and through the respiratory and gastrointestinal tracts. JMPR concluded that chlorpyrifos is unlikely to pose a carcinogenic risk to humans. Chlorpyrifos was not genotoxic in an adequate range of studies in vitro and in vivo. In long-term studies, inhibition of cholinesterase activity was the main toxicological finding in all species.

Derivation of maximum acceptable value The MAV for chlorpyriphos in drinking-water was derived as follows:

0.01 mg/kg x 70 kg x 0.1 = 0.035 mg/L (rounded to 0.04 mg/L) 2 L

where:

• acceptable daily intake = 0.01 mg/kg of body weight on the basis of a NOAEL of 1 mg/kg of body weight per day for inhibition of brain acetylcholinesterase activity in studies in mice, rats and dogs, using a 100-fold uncertainty factor

• acceptable daily intake = 0.01 mg/kg of body weight on the basis of a NOAEL of 0.1 mg/kg of body weight per day for inhibition of erythrocyte acetylcholinesterase activity in a study of human subjects exposed for nine days, using a 10-fold uncertainty factor


• poportion of acceptable daily intake allocated to drinking-water = 0.1

• average quantity of water consumed by an adult = 2 L/day.


References FAO/WHO. 2000. Pesticide residues in food � 1999 evaluations. Part II � Toxicological. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/00.4).


HMSO. 1986. Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector or Flame Photometric Detector (organophosphorus pesticides in river and drinking-water, tentative method 1980; and organophosphorus pesticides in sewage sludge: organophosphorus pesticides in river and drinking-water: an addition, 1985).

Royal Society of Chemistry. 1987. The Agrochemicals Handbook (2nd ed).

Martindale. 1994. The Extra Pharmacopoeia (30th ed). The Pharmaceutical Press.

WHO. 2003. Chlorpyrifos in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/87).


Cyanazine New entry July 2005.

Maximum acceptable value Based on health considerations, the concentration of cyanazine in drinking-water should not exceed 0.0007 mg/L (0.7 µg/L).


1 To source waters

Cyanazine is a member of the triazine family of herbicides. It is used as a pre- and post-emergence selective herbicide for the control of annual grasses and broadleaf weeds. Cyanazine is registered for use in New Zealand and is available as suspension concentrate or water dispersible granules. Its trade names are Bladex 50SC and Cy-pro 90DF. No information is available on the annual usage of specific active ingredients in New Zealand, although cyanazine is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Cyanazine is quite soluble in water: 171 mg/L (Merck & Co 1996). Cyanazine has a half life in soil of 14 days (Hort Research 2000). It can be degraded in soil and water by microorganisms and by hydrolysis. Four degradation products can be identified for cyanazine � the amide, two acids, and the amine. Aerobically and anaerobically aged cyanazine residues, primarily the amine degradation product, are intermediately mobile to mobile in sandy clay loam soil. The degradation products have all been identified in soil leachate, as has unaltered cyanazine (WHO 1998). There is no information available regarding the greatest source of exposure to cyanazine for New Zealanders (ie, dermal contact, inhalation, diet: food, water).


Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included cyanazine. In the New Zealand national pesticides surveys, conducted for groundwater every four years since 1990, cyanazine has been detected once, at a concentration of 0.001 mg/L. Monitoring conducted by Environment Canterbury has also detected cyanazine in groundwater at two locations in the Level Plain area in South Canterbury. At one location it has been detected in four monitoring rounds at concentrations ranging from 0.00007�0.00475 mg/L, whilst at the other location it has been detected once, at a concentration of 0.00004 mg/L (Close et al 2001). Cyanazine was detected in nine of 1128 samples of municipal and private water supplies in Quebec (1986), Ontario (1979 to 1986) and Alberta (1978 to 1986) (detection limits ranged from 0.000025 to 0.001 mg/L). Concentrations ranged from less than 0.0001 mg/L in Quebec water supplies to 0.004 mg/L in Ontario water supplies (Health Canada, 1989). Cyanazine has also been detected at trace levels in surface and groundwater in some US States (WHO 1998). Cyanazine has been found in groundwater in the Netherlands at concentrations above 0.0001 mg/L. It was not detected in surface water used as a source for drinking-water (Council of Europe 1993, cited in WHO 1998). Has been detected in surface water and groundwater, usually at concentrations of a few micrograms per litre, although levels as high as 1.3 and 3.5 mg/L have been measured in surface water and groundwater, respectively (WHO 2004).

Removal methods Oxidation of triazines by ozone is reported to be effective (Chiron et al 2000). The water chemistry, in particular the alkalinity and pH, will affect the oxidation rate. Use of activated carbon following ozonisation should be considered to adsorb oxidation products. Nanofiltration (membrane technology) in water with a low natural organic matter concentration is reported to remove approximately 50% of atrazine and simazine (Agbekodo et al 1996). The percentage is increased to 90�100% when 3.6 mg/L of natural organic matter is present. Similar results may be expected for cyanazine as it is from the same chemical family and of comparable molecular size. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of cyanazine, although a guide to the efficiency of the process cannot be provided.


Referee method

Liquid/liquid extraction/gas chromatography-electron capture detector (EPA 551.2).


None recommended.


Health considerations Cyanazine is rapidly absorbed from the gastrointestinal tract of experimental animals. Between 80 and 88% of doses of radioactively labelled cyanazine are eliminated from rats and dogs within four days, and within 21 days in cows. In rats, elimination in urine was almost equal to elimination in faeces. In dogs and cows, approximately one-half of the dose was eliminated in the urine, and about one-third was eliminated in the faeces. In cows, the amount of residues excreted daily was constant throughout the study period. Cyanazine was also detected in cow�s milk (WHO 1998).

Acute poisoning

The acute oral LD50 for rats is 182�334 mg/kg, mice 380 mg/kg, rabbits 141 mg/kg (RSocC 1987). These values suggest a relatively high acute oral toxicity compared with other pesticides. WHO (1996) has classified cyanazine as �moderately hazardous�. Poisoned animals have laboured breathing and blood in their saliva (Occupational Health Services, Material Safety Data Sheet on Cyanazine, 3/17/87 OHS: NY). The pesticide also causes inactivity and depression in laboratory animals.

Chronic exposure

On the basis of the available mutagenicity data on cyanazine, evidence for genotoxicity is equivocal. Cyanazine causes mammary gland tumours in Sprague-Dawley rats but not in mice. The mechanism of mammary gland tumour development in Sprague-Dawley rats is currently under investigation and may prove to be hormonal (WHO 1998). Cyanazine is also teratogenic (causes birth defects) in Fischer 344 rats at dose levels of 25 mg/kg of body weight per day and higher. The International Agency for Research on Cancer has not classified cyanazine for its ability to cause cancer. Atrazine, which has a chemical structure similar to that of cyanazine, has been found to increase the incidence of mammary tumours in rats and has been classified by IARC (1991) in Group 2B (agent is possibly carcinogenic to humans).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for cyanazine in drinking-water.The NOAEL was established on the basis of hyperactivity in male rats in a two-year toxicity/ carcinogenicity study. The MAV was derived as follows:


where:

• mo observable adverse effect level = 0.198 mg/kg body weight per day identified on the basis of hyperactivity in male rats in a two-year toxicity/carcinogenicity study

• average weight of adult = 70kg


• proportion of tolerable daily intake allocated to drinking-water = 10%

• uncertainty factor = 1000 (100 for interspecies and interspecies variation and 10 for limited evidence of carcinogenicity)




Close ME, Rosen MR, Smith VR. 2001. Fate and transport of nitrates and pesticides in New Zealand�s aquifers. In Rosen MR, White PA (eds). Groundwaters of New Zealand, Chapter 8, pp. 185�220. New Zealand: The Hydrological Society.

Council of Europe. 1993. Pesticides and Groundwater. Strasbourg: Council of Europe Press.

IARC. 1991. Atrazine. Occupational Exposures in Insecticide Application and Some Pesticides 53: 441�6. Lyon: International Agency for Research on Cancer. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans.


1 To source waters


Hort Research. 2000. Guidelines for the Sustainable Application of Agrichemicals. Version 1.1. CD-Rom. Palmerston North: Hort Research.


Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Cyanazine (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

WHO. 1996. The WHO Recommended Classification of Pesticides by Hazards and Guidelines to Classification. Geneva: World Health Organization, International Programme on Chemical Safety (WHO/PCS/96.3).

WHO. 1998. Guidelines for Drinking-water Quality (2nd ed). Addendum to Volume 2: Health criteria and other supporting information, pp. 165�75. Geneva: World Health Organization.

WHO. 2003. Cyanazine in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/60).


2,4-D Revised July 2005. (Also called 2,4-dichlorophenoxyacetic acid.)

Maximum acceptable value Based on health considerations, the concentration of 2,4-D in drinking-water should not exceed 0.04 mg/L. The maximum contaminant level (USEPA 2004) is 0.07 mg/L.


2,4-D may enter source waters as a result of its use as a systemic chlorophenoxy herbicide used widely in the control of broadleaf weeds. It is also used to control aquatic weeds. 2,4-D products are marketed as alkali salts, amine salts and ester formulations. The total annual usage in New Zealand of all forms of 2,4-D, including 2,4-DB, was 409,000 kg in the late 1980s, with nearly all the use being in the North Island. The highest usage in a county was 155,000 kg in Rangitikei. As at August 2005, 2,4-D is registered for use in New Zealand.

Forms and fate in the environment 2,4-D is microbially degraded in the environment with hydroxylation, decarboxylation, cleavage of the acid side chain and ring opening occurring. The half life in soils ranges from 2 to 14 days with a recommended average half life of 10 days. The half life in water ranges from 1 to several weeks.


Water solubility for 2,4-D acid ranges from 520 to 890 mg/L and the sorption coefficient is 20 mL/g. Water solubilities are much greater for the amine salts (18,000 to 4,000,000 mg/L) and much less for the esters (insoluble to 100 mg/L).

Typical concentrations in drinking-water Of 230 source water samples obtained from 212 supplies in New Zealand between 1988 and 1992, two samples contained detectable levels of 2,4-D. The concentrations were 0.0003 mg/L and 0.0022 mg/L (2.2 µg/L). In addition, 2,4-D was detected in two wells in the Te Puke area at concentrations between 0.00005�0.0001 mg/L (0.05-0.1 µg/L), and has also been found in surface waters. The P2 Chemical Determinand Identification Programme, sampled from 296 zones, did not find detectable concentrations of 2,4-D (limit of detection = 0.0001 mg/L).


No significant elevations were observed in sister chromatid exchanges or the frequency of chromosomal aberrations in forestry workers exposed to 2,4-D.

Levels in water overseas are usually below 0.0005 mg/L, although concentrations as high as 0.03 mg/L have been measured.

Removal methods No information is available on methods of removing 2,4-D from water. However, isotherm adsorption data indicate that removal by adsorption on to granular activated carbon should be possible. Removal of less than 100% has been reported with ozone. 0.001 mg/L should be achievable using GAC.


Referee method



1 Liquid/Liquid Extraction and Gas Chromatography with Electron Capture Detector (APHA 6640B).

Health considerations 2,4-D administered orally as the free acid or salt is absorbed rapidly and almost completely by humans. Animal studies have shown that after absorption it is distributed throughout the body, with highest concentrations in blood, kidney, liver, spleen and lung. Humans excrete most of the 2,4-D in urine. Symptoms of acute exposure to high doses of 2,4-D include effects on the gastro-intestinal tract such as nausea, vomiting and diarrhoea, direct myotoxic effects such as muscular weakness, stiffness, muscular spasms, and partial paralysis, effects on the kidney, pulmonary oedema, and effects on the central and peripheral nervous systems, including central nervous system depression, lethargy, slowed respiration, coma and death. Epidemiological results give limited evidence that occupational exposure to chlorophenoxy herbicides may cause cancer, and long-term studies in animals continue to show equivocal evidence of carcinogenicity, in one sex and species only.


The International Agency for Research on Cancer has classified chlorophenoxy herbicides in Group 2B (possibly carcinogenic to humans). However, based on the information available, it is not possible to ascertain the status of 2,4-D with respect to carcinogenicity, as almost all populations studied were exposed to a mixture of chlorophenoxy herbicides. In the only study in which exposure was clearly to 2,4-D only, the association was weak. JMPR has also concluded that 2,4-D and its salts and esters are not genotoxic.

Derivation of maximum acceptable value Because the data on the carcinogenic potential of 2,4-D are inadequate, and because 2,4-D has not been found to be genotoxic, the MAV was derived using a tolerable daily intake approach. The no observable adverse effect level used in the derivation is based on the effects of toxicity in dogs (for a variety of effects, including histopathological lesions in kidneys and liver), and a two-year study of toxicity and carcinogenicity in rats (for renal lesions). The MAV for the sum of 2,4-D and its salts and esters, expressed as 2,4-D, in drinking-water was derived as follows:


where:

• no observable adverse effect level = 1 mg/kg body weight per day for effects on the kidney in chronic studies in rats and mice



Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466.




References


Close ME. 1993. Assessment of pesticide contamination of groundwater in New Zealand, 2: Results of groundwater sampling. New Zealand Journal of Marine and Freshwater Research 27: 267�73.

FAO/WHO. 1997. Pesticide Residues in Food � 1996: Evaluations 1996: Part II � Toxicological. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/97.1).

Frank R, Logan L. 1988. Pesticide and industrial chemical residues at the mouth of the Grand, Saugeen and Thames Rivers, Ontario, Canada, 1981�85. Archives of Environmental Contamination and Toxicology 17: 741�[Add page number].


Que Hee SS, Sutherland RG. 1981. The Phenoxyalkanoic Herbicides. Vol I. Chemistry, analysis, and environmental pollution. Boca Raton, USA: CRC Press (Chemical Rubber Company Series in Pesticide Chemistry).



USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water. Supplement 2, Report No EPA/600/R-92129.

Chlorophenoxy herbicides not frequently found in drinking- water; when detected, concentrations are usually no greater than a few micrograms per litre (WHO 2004).

No information is available on methods of removing 2,4-DB from water, but chlorophenoxy acids have been reported to be oxidised by ozone. 0.1 mg/L should be achievable using GAC. Also see 2,4-D.





WHO. 1984. 2,4-Dichlorophenoxyacetic acid (2,4-D). Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 29).

WHO. 2003. 2,4-D in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/70).


2,4-DB (4-(2,4-dichlorophenoxy)butyric acid) Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of 2,4-DB in drinking-water should not exceed 0.1 mg/L.


1 To source waters

See 2,4-D. As at August 2005, 2,4-DB is registered for use in New Zealand.

Forms and fate in the environment 2,4-DB is degraded microbially in the environment with hydroxylation, decarboxylation, cleavage of the acid side chain and ring opening occurring. The half life in soils ranges from 2 to 14 days with a recommended average half life of 10 days. The half life in water ranges from one to several weeks.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 296 zones, did not find detectable concentrations of 2,4-DB (limit of detection = 0.0001 mg/L).

Removal methods



Referee method




In general, chlorophenoxy herbicides are absorbed rapidly from the gastro-intestinal tract and evenly distributed throughout the body. Accumulation in human tissues is not expected, and a steady-state level in the human body will be achieved within 3�5 days of exposure. Elimination occurs primarily in the urine, mostly in the unchanged form. Biological half lives of chlorophenoxy herbicides in mammals range from 10 to 33 hours. Metabolic conversions occur only at high doses. The salt and ester forms are hydrolysed rapidly and follow the same pharmacokinetic pathways as the free acid forms.



Short-term exposure studies on beagle dogs fed diets containing high doses of 2,4-DB reported effects including diarrhoea, inactivity, depression, weakness, cysts, increased mortality, reduced body weight and food consumption, haematological effects, abnormal blood chemistry and urinalysis, jaundice, increased relative thyroid, liver, spleen, and kidney weights and decreased relative testes weight. Long-term exposure studies on rats reported similar symptoms. Chlorophenoxy herbicides as a group, including 2,4-D and MCPA, have been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). However, based on the available data from studies on exposed populations and on animals, it is not possible to assess the carcinogenic potential of any specific chlorophenoxy herbicide. Therefore drinking-water guidelines for these compounds are based on a threshold approach for other toxic effects.

Derivation of maximum acceptable value As it is not possible to assess the carcinogenic potential to humans of any specific chlorophenoxy herbicide, a tolerable daily intake approach been used for the derivation of the MAV for 2,4-DB in drinking-water. The no observable adverse effect level used in the derivation is for effects on the body and organ weights, blood chemistry and haematological parameters in a two-year study in rats. The MAV for 2,4-DB in drinking-water was derived as follows:


where:

• no observable adverse effect level = 3 mg/kg body weight per day for effects on the body and organ weights, blood chemistry and haematological parameters in a two-year study in rats. The NOAEL used in the guideline value derivation is similar to the NOAEL of 2.5 mg/kg of body weight per day obtained in a short-term study in beagle dogs and the NOAEL for hepatocyte hypertrophy of 5 mg/kg of body weight per day obtained in a 3-month study in rats.





• uncertainty factor = 100 (for inter- and intra-species variation).


DDT is practically insoluble in water.





WHO. 2003. Chlorophenoxy herbicides (excluding 2,4-D and MCPA) in drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/44).


DDT and its derivatives Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of DDT and its derivatives in drinking-water should not exceed 0.001 mg/L (1 µg/L).


1 To source waters

DDT may enter source waters as a result of its use as a nonsystemic contact insecticide with a broad spectrum of activity.

DDT is not currently used in New Zealand, but was used extensively in the past, particularly on pasture to control grass grub and porina. There are still significant residues in soils in many areas. The registration of DDT was cancelled in New Zealand in 1989.

Forms and fate in the environment DDT and its metabolites are persistent in the environment and resistant to microbial degradation, although photochemical degradation does occur. The persistence of DDT in temperate climates is in the order of years. DDT is adsorbed readily to soils and sediments and most DDT that enters water bodies is firmly attached to soil particles. DDT is taken up readily by other organisms and bioconcentration is significant.

Typical concentrations in drinking-water DDT and its isomers were not detected in all of 230 samples from 212 supplies sampled in New Zealand between 1988 and 1992. Detection limits ranged from 0.00025 to 0.00004 mg/L (0.25 to 0.04 µg/L) for this class of compounds.


In a study of surface water supplies in the United States between 1964 and 1968, the highest concentration of DDT recorded was 0.0008 mg/L. In Germany, concentrations were even lower, averaging 0.00001 mg/L (10 ng/L).

Referee method

The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of DDT and its isomers (limit of detection = 0.0002 mg/L).

Removal methods Specific information concerning the removal of DDT and its isomers from water is unavailable. However, its low solubility and attraction to soil particles makes it likely that removal by chemical coagulation is possible. Isotherm adsorption data also indicate that removal by adsorption on to activated carbon should be possible.


Liquid/Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (APHA 6410B).



2 Liquid/Liquid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 508).

Health considerations Food is the major source of intake of DDT and related compounds for the general population. Absorption of small doses of DDT, such as those found in food residues, is virtually complete and is facilitated by the presence of fat in food. It is stored preferentially in fat. Like most species, humans convert DDT to DDE, which is stored even more avidly than the parent compound. A small amount of DDD may also be found in tissues. In humans, signs and symptoms reported following acute intoxication by DDT include nausea, vomiting, paraaethesia, dizziness, ataxia, confusion, tremor and, in severe cases, convulsions. Occupational exposure of workers over 25 years at an average dosage of 0.25 mg/kg body weight per day resulted in no reported adverse effects. From epidemiological observations of humans, there is no firm evidence that DDT has any reproductive or teratogenic effects. In most studies, DDT did not induce genotoxic effects in rodent or human cell systems, nor was it mutagenic in fungi or bacteria. DDT impaired reproduction in several species. All epidemiological studies in humans have indicated that DDT is not carcinogenic. The International Agency for Research on Cancer has concluded that, due to evidence of carcinogenicity in experimental animals, DDT is a possible human carcinogen (Group 2B).

Derivation of maximum acceptable value The MAV for DDT (plus its metabolites) was derived using the acceptable daily intake recommended by the Joint FAO/WHO Meetings on Pesticide Residues (JMPR) in 1984.


Because infants and children may be exposed to greater amounts of chemicals in relation to their body weight and because of concern over the bioaccumulation of DDT, a 10 kg child was used for the calculation of the MAV. The MAV for DDT and its derivatives in drinking-water was derived as follows:

1 mg/kg body weight/day x 10 kg x 0.01 = 0.001 mg/L (1 µg/L) 1 L/day x 100

where:

• acceptable daily intake is based on a NOAEL of 1 mg/kg of body weight per day for developmental toxicity in rats


The MAV exceeds the solubility of DDT of 0.001 mg/L (1 µg/L). However, some of the DDT may be adsorbed on to the small amount of particulate matter present in drinking-water, so that the MAV could be reached under certain circumstances.

• average weight of child = 10 kg

• average quantity of water consumed by an child = 1 L per day

• uncertainty factor = 100.

DDT is listed under the Stockholm Convention on Persistent Organic Pollutants. Hence, monitoring may occur in addition to that required by drinking-water guidelines.



FAO/WHO. 2001. Pesticide Residues in Food � 2000: Evaluations � 2000: Part II � Toxicology. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/01.3).



WHO. 2003. DDT and its Derivatives in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/89).


Diazinon Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of diazinon in drinking-water should not exceed 0.01 mg/L. WHO 2004 states that diazinon is unlikely to occur in drinking-water, so does not establish a guideline value.



1 To source waters

Diazinon may enter source waters as a result of its application as an insecticide, used to control a wide range of pests.

The MAV for diazinon was calculated by the New Zealand Ministry of Health as follows:

The total annual usage of diazinon in New Zealand in the late 1980s was 133,300 kg. As at August 2005, diazinon is registered for use in New Zealand.

Forms and fate in the environment Diazinon degrades in the field with a half life of approximately 40 days. The water solubility is 60 mg/L and the sorption coefficient is 1000 mL/g (estimate).

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 342 zones, did not find detectable concentrations of diazinon (limit of detection = 0.0003 mg/L).

Removal methods Reverse osmosis, granular activated carbon adsorption, and ozonation have been reported to remove diazinon from water with efficiencies ranging from 75 to 100%.


Referee method

No referee method has been given for diazinon because no method meets the required criteria.


1 Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector or Flame Photometric Detector (HMSO 1986).

Health considerations Diazinon is a reactive organophosphorus compound, and many of its toxic effects are similar to those produced by other substances of this class. Characteristic effects include inhibition of acetyl cholinesterase and central nervous system depression. Organophosphates are absorbed readily through the skin, and through the respiratory and gastrointestinal tracts. No information is available on the long-term health effects to humans of exposure to diazinon. The International Angency for Research on Cancer has not evaluated the carcinogenic potential of diazinon.



where: • acceptable daily intake = 0.002 mg/kg body weight • average weight of adult = 70 kg


• proportion of acceptable daily intake allocated to drinking-water = 0.2 • average quantity of water consumed by an adult = 2 L/day.


HMSO. 1986. Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector or Flame Photometric Detector (organophosphorus pesticides in river and drinking-water, tentative method 1980; and organophosphorus pesticides in sewage sludge: organophosphorus pesticides in river and drinking-water: an addition, 1985).

Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues, Geneva, 20�29 September 1993, FAO Plant Production and Protection Paper No 122 (Annex 1).

Royal Society of Chemistry. 1987. The Agrochemicals Handbook (2nd ed).


1,2-dibromo-3-chloropropane may enter source waters due to its use as a nematocidal fumigant. 1,2-dibromo-3-chloropropane (DBCP) has never been registered for use in New Zealand.

USEPA. 1989. Drinking Water Health Advisory: Pesticides. USEPA Office of Drinking Water Health Advisors, Lewis Publishers.


1,2-Dibromo-3-chloropropane Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of 1,2-dibromo-3-chloropropane in drinking-water should not exceed 0.001 mg/L (1 µg/L). The maximum contaminant level (USEPA 2004) is 0.0002 mg/L.


1 To source waters

Forms and fate in the environment 1,2-dibromo-3-chloropropane is expected to volatilise from surface water. It is highly persistent in soil and has been shown to remain there for more than two years. It is mobile in soil and may migrate to groundwater.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find detectable concentrations of 1,2-dibromo-3-chloropropane (limit of detection = 0.0005 mg/L). A limited survey overseas found levels of up to a few micrograms per litre in drinking-water.


Removal methods No information is available on technologies capable of removing 1,2-dibromo-3-chloropropane from water. However, isotherm adsorption data indicate that removal by adsorption on to granular activated carbon should be possible. WHO (2004) states that 0.001 mg/L should be achievable using air stripping followed by GAC.


Referee method


References



2. Liquid/Liquid Extraction and Gas Chromatography with Electron Capture Detector (APHA 6231B).

3. Liquid/Liquid Extraction and Gas Chromatography with Electron-Capture Detection (EPA 551).

Health considerations Absorption of dibromochloropropane is expected to be high following ingestion and distribution is primarily to the liver and kidneys. Dibromochloropropane can probably cross the placenta. Urine is the predominant route for elimination of metabolites. Workers occupationally exposed to dibromochloropropane were reported to have reduced spermatogenesis. This condition was reported to be reversible although permanent damage of germinal epithelium was reported in a follow-up of exposed workers. No chromosomal aberrations were identified in men who were affected, nor were there increases in abortions and malformations in offspring. No association was found between dibromochloropropane contamination in drinking-water and incidences of gastric cancer and leukaemia. On the basis of data from studies on rats and mice, dibromochloropropane was determined to be carcinogenic in both sexes by ingestion, inhalation and skin contact. It was also determined to be a reproductive toxicant in humans and several species of laboratory animals. The International Agency for Research on Cancer has classified dibromochloropropane in Group 2B (possible human carcinogen). Recent epidemiological evidence suggests an increase in cancer mortality in individuals exposed to high levels of dibromopropane.

Derivation of maximum acceptable value The linearised multistage model was applied to the data on the incidence of stomach, kidney and liver tumours in the male rat in a 104-week dietary study. The concentration of dibromochloropropane associated with an excess lifetime cancer risk of one per one hundred thousand (10-5) is 0.001 mg/L (1 µg/L). This MAV should be protective for the reproductive toxicity of dibromochloropropane. 1,2-dibromo-3-chloropropane has a taste and odour threshold in water of 0.01 mg/L.




USEPA. 1989. Drinking Water Health Advisory: Pesticides. United States Environmental Protection Agency Office of Drinking Water Health Advisors. Lewis Publishers.



USEPA. 1990. Methods for the Determination of Organic Compounds in Drinking Water. Supplement 1, Report No EPA/600/4-90-020.


WHO. 2003. 1,2-dibromo-3-chloropropane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/34).


1,2-Dibromoethane Revised July 2005. (Also called ethylene dibromide.)

Based on health considerations, the concentration of 1,2-dichloropropane in drinking-water should not exceed 0.0004 mg/L. WHO (2004) states that their guideline value is provisional due to serious limitations of the critical studies. The maximum contaminant level (USEPA 2004) is 0.00005 mg/L.

1 To source waters

Ethylene dibromide may enter source waters as a result of its use as an insecticidal fumigant, and as a lead scavenger in tetra-alkyl lead petrol and antiknock preparations. It is not currently registered as an insecticide in New Zealand but has been used in the past. It has been a problem overseas in groundwater as a result of petrol spills and agricultural use. Ethylene dibromide can also be used as a solvent.

Forms and fate in the environment Evaporation is an important removal mechanism for ethylene dibromide from surface water, with half lives between one and five days. Ethylene dibromide is very stable in groundwater, especially under anaerobic conditions, where half lives of around 20 years are estimated. Hydrolysis of ethylene dibromide to bromide, ethylene, ethylene glycol and carbon dioxide takes place in the soil. Evaporated ethylene dibromide in the atmosphere reacts with photochemically produced hydroxyl radicals with half lives of 32 days. The water solubility of ethylene dibromide is 4300 mg/L.

Typical concentrations in drinking-water No data are available on the concentration of ethylene dibromide in New Zealand drinking-water supplies. Overseas results indicate that in agricultural areas, ethylene dibromide is found in groundwaters at concentrations of 0.00001 mg/L (0.01 µg/L) to 0.015 mg/L. Detected in groundwater following its use as a soil fumigant at concentrations as high as 0.1 mg/L (WHO 2004).


Removal methods No information is available on methods of removing ethylene dibromide from water.


Referee method

Liquid/liquid extraction gas chromatographic APHA method 6231B uses a microextraction and capilliary columns. WHO (2004) states a limit of detection 0.01 mg/L by microextraction GC/MS; 0.03 mg/L by purge and trap GC with halogen-specific detector; 0.8 mg/litre by purge-and-trap capillary column GC with photoionization and electrolytic conductivity detectors in series.


Ethylene dibromide is also analysed by the purge and trap GC/MS, APHA methods 6200 B and C.

Health considerations Animal studies have shown that ethylene dibromide is absorbed readily following oral, inhalation and skin exposure. Following ingestion, the highest level of metabolites was found in the liver and kidney. 1,2-dibromoethane has induced an increased incidence of tumours at several sites in all carcinogenicity bioassays identified in which rats or mice were exposed to the compound by gavage, ingestion in drinking-water, dermal application and inhalation. However, many of these studies were characterized by high early mortality, limited histopathological examination, small group sizes or use of only one exposure level. Following long-term inhalation studies in mice and rats ethylene dibromide produced adenomas and carcinomas of the nasal cavity, haemangiosarcomas of the spleen and mammary tumours in both species. Ethylene dibromide induced skin and lung tumours in mice after skin application. In humans, prolonged contact with ethylene dibromide causes skin irritation. Long-term occupational exposure to ethylene dibromide affects semen quality. Statistically significant decreases in sperm count ejaculate, the percentage of viable and motile sperm, and increases in the proportion of sperm with morphological abnormalities were observed among the exposed men compared with controls. Ethylene dibromide induced sister chromatid exchange, mutations and unscheduled DNA synthesis in both human and rodent cells in vivo. In 1987, the International Agency for Research on Cancer concluded that the evidence for carcinogenicity in humans was inadequate, but that animal studies were sufficient to establish carcinogenicity and classed ethylene dibromide in Group 2A (probably carcinogenic to humans).

Derivation of maximum acceptable value The 0.0004 mg/L MAV is the lower end of the range (and thus more conservative estimate) of lifetime low-dose cancer risks calculated by linearised multistage modelling of the incidences of haemangiosarcomas and tumours in the stomach, liver, lung and adrenal cortex (adjusted for the observed high early mortality, where appropriate, and corrected for the expected rate of increase in tumour formation in rodents in a standard bioassay of 104 weeks) of rats and/or mice exposed to 1,2-dibromoethane by gavage.



IPCS. 1995. Report of the 1994 meeting of the Core Assessment Group. Geneva: World Health Organization, International Programme on Chemical Safety, Joint Meeting on Pesticides (WHO/PCS/95.7).

IPCS. 1996. 1,2-dibromoethane. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 177).


WHO. 2003. 1,2-dibromoethane in drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/66).


1,2-Dichloropropane Revised July 2005. (Also called propylene dichloride.)


Based on health considerations, the concentration of 1,2-dichloropropane in drinking-water should not exceed 0.05 mg/L. WHO (2004) states that the guideline value is provisional owing to limitations of the toxicological database. The maximum contaminant level (USEPA 2004) is 0.005 mg/L.


1 To source waters

1,2-dichloropropane may enter source waters as a result of its use as a soil fumigant, often in combination with 1,3-dichloropropene (D-D). It is also used for a variety of industrial purposes including as a chemical intermediate, as a lead scavenger for antiknock fluids, and in dry-cleaning and metal-degreasing solvents. It is not currently used in New Zealand as a pesticide. It was registered as Shell DD soil fumigant, but 1,2-dichloropropane does not appear on ERMA�s list of registered trade name pesticides as at August 2005.

Forms and fate in the environment 1,2-dichloropropane volatilises from surface waters and is degraded in air by photochemically produced hydroxyl radicals with a half life of 23 days or more. Little or no degradation in soil has been reported. It is mobile in soil and would migrate to groundwater.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find detectable concentrations of 1,2-dichloropropane (limit of detection = 0.0005 mg/L). Detected in groundwater and drinking-water, usually at concentrations below 0.020 mg/L, although levels as high as 0.44 mg/L have been measured in well water (WHO 2004).


Removal methods No information is available on technologies capable of removing 1,2-dichloropropane from water. WHO (2004) considers 0.001 mg/L should be achievable using GAC.


Referee method




Health considerations Animal studies have shown that 1,2-dichloropropane is absorbed readily from the gastro-intestinal tract and highest levels were detected in the liver, kidney and blood. Clinical symptoms following the ingestion of 1,2-dichloropropane in humans involve effects on the gastro-intestinal system (nausea, burning and vomiting), central nervous system (dizziness, disorientation, headache and coma), kidney failure and liver necrosis. Effects on the respiratory system, heart and blood have also been described. There is a relatively limited data base on the toxicity of 1,2-dichloropropane, but it is mutagenic in some short-term assays in vitro. When administered orally, 1,2-dichloropropane produced statistically significant increases in the incidence of hepatocellular adenomas and carcinomas in both sexes of mice. There was marginal evidence of carcinogenicity in female rats. The International Agency for Research on Cancer classified 1,2-dichloropropane as a Group 3 carcinogen (not classifiable as to its carcinogenicity to humans), as there are no human data and only limited data from animal studies.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for 1,2-dichloropropane in drinking-water. The lowest observable adverse effect level used in the derivation is based on a variety of systemic effects in a 13-week oral study in rats of changes in haematological parameters. The MAV for 1,2-dichloropropane in drinking-water was derived as follows:

100 x (5/7) mg/kg body weight/day x 70 kg x 0.1 = 0.05 mg/L 2 L/day x 5000

where:

• lowest observable adverse effect level = 100 mg/kg body weight per day on the basis of a variety of systemic effects in a 13-week oral study in rats (normalised for five days/week dosing in the derivation)





• uncertainty factor = 5000; 100 for inter- and intra-species variation; 10 for the use of a LOAEL instead of a NOAEL, and 5 to reflect limitations of the database, including the limited data on in vivo genotoxicity and use of a subchronic study).




WHO. 2003. 1,2-dichloropropane (1,2-DCP) in drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/61).


1,3-Dichloropropane Revised July 2005.

Maximum acceptable value There are insufficient data to derive a MAV for 1,3-dichloropropane in drinking-water. WHO (2004) states that the available data are considered insufficient to permit recommendation of a guideline value.


1 To source waters

1,3-dichloropropane may have entered source waters due to its presence as a by-product in D-D soil fumigant; this product does not appear on ERMA�s list of registered trade name pesticides as at August 2005. It is also used in the chemical synthesis industry. 1,3-dichloropropane has never been registered for use in New Zealand.

Forms and fate in the environment 1,3-dichloropropane volatilises from both soil and surface waters to the atmosphere where it can be degraded photochemically. It is mobile in soils.

Typical concentrations in drinking-water No data are available on the concentration of 1,3-dichloropropane in New Zealand drinking-water supplies, nor are data available on levels in drinking-waters overseas. Measurements in the Ohio River, USA, however, showed detected levels to be below 0.0008 mg/L (0.8 µg/L).

Removal methods No information is available on technologies capable of removing 1,2-dichloropropane from water. It should behave in a similar manner to 1,2-dichloropropane.



Referee method

A referee method cannot be selected for 1,3-dichloropropane because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for 1,3-dichloropropane for the above reason. However, the following information may be useful:

1 1,3-dichloropropane in drinking-water may be analysed by purge and trap gas chromatography with mass spectrometry detection (Method APHA 6210 or EPA Method 524.2). The detection limit is 0.0001 mg/L (0.1 µg/L).

Health considerations 1,3-dichloropropane is of low acute toxicity.

Based on health considerations, the concentration of 1,3-dichloropropene in drinking-water should not exceed 0.02 mg/L.

Short-term animal studies have shown that 1,3-dichloropropane induced mild dermatitis on the shaved skin of mice. Peripheral blood changes, including an increased number of white blood cells and reticulocytes (newly formed red blood cells) were observed in dermally exposed animals. There is some indication that 1,3-dichloropropane may be genotoxic in bacterial systems.

Derivation of maximum acceptable value No short-term, long-term, reproductive or developmental toxicity data pertinent to exposure via drinking-water could be located for this compound. The available data are considered to be insufficient to recommend a MAV for 1,3-dichloropropane in drinking-water.



WHO. 2003. 1,3-dichloropropane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/35).


1,3-Dichloropropene Revised July 2005.




1 To source waters

1,3-dichloropropene may enter source waters as a result of its application as a broad spectrum soil fumigant, used for nematode control. It is often used in combination with 1,2-dichloropropane. The commercial product is a mixture of the cis and trans isomers. It is was registered in New Zealand as Shell D-D soil fumigant, but does not appear on ERMA�s list of registered trade name pesticides as at August 2005.

Forms and fate in the environment 1,3-dichloropropene volatilises from both soil and surface waters to the atmosphere where it can be degraded photochemically. Hydrolysis and microbial degradation can also occur. Water solubility ranges from 2250 to 2800 mg/L and the sorption coefficient is 32 mL/g. The recommended average half life in soil is 10 days.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 332 zones, did not find detectable concentrations of 1,3-dichloropropene (limit of detection = 0.0005 mg/L). Has been found in surface water and groundwater at concentrationsof a few micrograms per litre (WHO 2004).

Removal methods No information is available on methods of removing 1,3-dichloropropene from water. However, isotherm adsorption data indicate that removal by adsorption on to granular activated carbon should be possible.




Referee method



1,3-dichloropropene is absorbed through the skin and respiratory and gastro-intestinal systems. Oral administration of 1,3-dichloropropene in rats resulted in approximately 90% absorption of the administered dose. 1,3-dichloropropene and its metabolites are excreted principally in urine. The only known human fatality following accidental ingestion of a D-D mixture of unknown dosage occurred within a few hours. Symptoms were abdominal pain, vomiting, muscle twitching and pulmonary oedema. Inhalation of 1,3-dichloropropene at concentrations above 1500 ppm resulted in gasping, coughing, substernal pain and respiratory distress. Mutagenicity tests on bacteria have indicated that 1,3-dichloropropene is a direct-acting mutagen.


It has been shown to produce forestomach tumours following long-term oral gavage exposure in rats and mice. Tumours were also found in the bladder and lung in female mice and liver in male rats. Long-term inhalation studies in the rat were negative, whereas inhalation studies in mice showed some benign lung tumours. The International Agency for Research on Cancer concluded that there was sufficient evidence for the carcinogenicity of 1,3-dichloropropene in experimental animals to classify it in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value Based on lung and bladder tumours observed in female mice in a two-year gavage study, and using the linearised multistage model, the concentration of 1,3-dichloropropene (the sum of cis and trans) associated with an excess lifetime cancer risk of one per 100,000 (10-5) is 0.02 mg/L.





WHO. 2003. 1,3-dichloropropene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/36).


Dichlorprop Revised July 2005. (Also called 2,4-DP or 2,4-dichlorophenoxypropionic acid.)

Maximum acceptable value Based on health considerations, the concentration of dichlorprop in drinking-water should not exceed 0.1 mg/L.


1 To source waters

Dichlorprop may enter source waters as a result of its use for the post-emergent control of annual and perennial broadleaved weeds; brush control; control of broadleaved aquatic weeds and chemical maintenance of embankments and roadside verges.

The total annual usage of dichlorprop in New Zealand in the late 1980s was 35,000 kg, all of it in the North Island with the greatest usage being in Rangitikei (9800 kg). As at August 2005, dichlorprop is registered for use in New Zealand.

Forms and fate in the environment The half life for degradation of dichlorprop to 2,4-dichlorophenol in soil is estimated to be 8�12 days and the recommended average half life is 10 days.


The water solubility of the acid is 350 mg/L with lower solubility for the ester (50 mg/L) and much higher solubility for salts (660,000 to 900,000 mg/L).

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 296 zones, did not find detectable concentrations of dichlorprop (limit of detection = 0.0001 mg/L). Chlorophenoxy herbicides not frequently found in drinking- water; when detected, concentrations are usually no greater than a few micrograms per litre (WHO 2004).

Removal methods No specific information is available on methods of removing dichlorprop from water, but chlorophenoxy acids have been reported to be oxidised by ozone.


Referee method





Health considerations In general, chlorophenoxy herbicides are absorbed rapidly from the gastro-intestinal tract and evenly distributed throughout the body. Accumulation in human tissues is not expected and a steady-state level in the human body will be achieved within 3�5 days of exposure. Elimination occurs primarily in the urine, mostly in the unchanged form. Biological half lives of chlorophenoxy herbicides in mammals range from 10 to 33 hours. Metabolic conversions occur only at high doses. The salt and ester forms are hydrolysed rapidly and follow the same pharmacokinetic pathways as the free acid forms. Dichlorprop has been shown to cross the placenta.

Rats fed diets containing high doses of dichlorprop had slight liver hypertrophy. Long-term exposure studies in rats reported symptoms including effects on the liver, kidneys and blood. In dietary studies in rats, slight liver hypertrophy was observed in a three-month study, and effects in a two-year study included hepatocellular swelling, mild anaemia, increased incidence of brown pigment in the kidneys (possibly indicative of slight degeneration of the tubular epithelium) and decreased urinary specific gravity and protein. Chlorophenoxy herbicides as a group, including 2,4-D and MCPA, have been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). However, based on the available data from studies on exposed populations and on animals, it is not possible to assess the carcinogenic potential of any specific chlorophenoxy herbicide. Therefore, drinking-water guidelines for these compounds are based on a threshold approach for other toxic effects.


Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for dichlorprop in drinking-water. The no observable adverse effect level used in the derivation is for renal toxicity based on a two-year study in rats. The MAV for dichlorprop in drinking-water was derived as follows:


where:


• no observable adverse effect level = 3.64 mg kg body weight per day for renal toxicity based on a two-year study in rats





Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466�[add page number].








WHO. 2003. Chlorophenoxy Herbicides (excluding 2,4-D and MCPA) in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/44).



Dimethoate New entry August 2005.

Maximum acceptable value Based on health considerations, the concentration of dimethoate in drinking-water should not exceed 0.007 mg/L (7 µg/L).


1 To source waters

Dimethoate, an organophosphate, acts by interfering with the activities of cholinesterase, an enzyme that is essential for the proper working of the nervous systems of both humans and insects. Dimethoate is used to kill mites and insects systemically and on contact. It is used on a wide range of plants and vegetables. Dimethoate has been administered to livestock for control of flies. It is available in aerosol spray, dust, emulsifiable concentrate, and ULV concentrate formulations. As at August 2005, dimethoate is registered for use in New Zealand.

Forms and fate in the environment Dimethoate is biodegradable. It undergoes rapid degradation in the environment and in sewage treatment plants. Because dimethoate is highly soluble in water and it adsorbs only very weakly to soil particles, it may be subject to considerable leaching. It may be subject to degradation by hydrolysis, especially in alkaline soils, and to evaporation from dry soil surfaces. Losses due to evaporation of 23 to 40% of applied dimethoate have been reported. Biodegradation may be significant, with 77% degradation reported for a non-sterile clay loam soil in two weeks reported. Dimethoate does not persist. Soil half lives of 4 to 16 days, or as high as 122 days have been reported. Half lives between 2.5 and 4 days were reported during drought and moderate rainfall conditions. Dimethoate breaks down faster in moist soils. It is broken down rapidly by most soil microorganisms. In water, dimethoate is not expected to adsorb to sediments or suspended particles, nor to bioaccumulate in aquatic organisms. It is subject to significant hydrolysis, especially in alkaline waters. Hydrolysis half lives of 3.7 and 118 days at pH 9 and pH 7, respectively, have been estimated. Photolysis and evaporation from open waters is not expected to be significant. The half life for dimethoate in raw river water was eight days, with disappearance possibly due to microbial action or chemical degradation.

Typical concentrations in drinking-water Detected at trace levels in a private well in Canada, but not detected in a Canadian survey of surface water or drinking- water supplies (WHO 2004).

Removal methods None reported.



Referee method

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[???]

Health considerations Dimethoate is moderately toxic by ingestion, inhalation and dermal absorption. As with all organophosphates, dimethoate is readily absorbed through the skin. Dimethoate is metabolised rapidly by mammals. Rats excreted about 60% of an administered dose in urine and expired air within 24 hours. In another study, rats given a single oral dose, excreted 50% in the urine and 25% in the feces within 24 hours. Nine days later, only 0.9 to 1.1% of the dose remained in the rats� tissues. Human volunteers excreted 76 to 100% of administered dimethoate within 24 hours.

Acute poisoning

The oral LD50 for technical dimethoate in rats is 60 to 387 mg/kg, 60 mg/kg in mice, 400 mg/kg in dogs, 200 mg/kg in hamsters, 300 mg/kg in rabbits, 350 mg/kg in guinea pigs, and 100 mg/kg in cats.

Repeated or prolonged exposure to organophosphates may result in the same effects as acute exposure, including the delayed symptoms. Other effects reported in workers repeatedly exposed include impaired memory and concentration, disorientation, severe depressions, irritability, confusion, headache, speech difficulties, delayed reaction times, nightmares, sleepwalking and drowsiness or insomnia. An influenza-like condition with headache, nausea, weakness, loss of appetite, and malaise has also been reported.

The four-hour LC50 for dimethoate in rats is 1.2 mg/L. Dimethoate is highly toxic to fish and to aquatic invertebrates. The 96-hour LC50 for dimethoate in rainbow trout is 6.2 ug/L. The 48-hour LC50 in Daphnia magna, a small freshwater crustacean, is 2.5 ug/L. There was no cholinesterase inhibition in an adult human who ingested 18 mg (about 0.26 mg/kg/day) of dimethoate/day for 21 days. No toxic effects and no cholinesterase inhibition were observed in individuals who ingested 2.5 mg/day (about 0.04 mg/kg/day) for four weeks. In another study with humans given oral doses of 5, 15, 30, 45 or 60 mg/day for 57 days, cholinesterase inhibition was observed only in the 30 mg/day or higher dosage groups.

Chronic exposure

When mice were given 60 ppm (9.5 to 10.5 mg/kg/day) dimethoate in their drinking water, there was decreased reproduction, pup survival, and growth rates of surviving pups. Adults in this study exhibited reduced weight gain, but their survival was not affected. In a 3-generation study with mice, 2.5 mg/kg/day did not decrease reproductive performance or pup survival. Once in the bloodstream, dimethoate may cross the placenta. Dimethoate is possibly a human teratogen. It was teratogenic in cats and rats. A dosage of 12 mg/kg/day given to pregnant cats increased the incidence of extra toes on kittens. The same dosage given to pregnant rats produced birth defects related to bone formation, runting and defects related to malfunction of the bladder. Dosages of 3 or 6 mg/kg/day were not teratogenic in cats or rats. The NOAEL for both cats and rats was 2.8 mg/kg/day. There were no teratogenic effects seen in the offspring of mice given 9.5�10.5 mg/kg/day dimethoate in their drinking water.


Dimethoate is possibly a mutagen. Mutagenic effects (dominant lethal) were more prominent in male mice given a single high dose of dimethoate than in male mice given one-twelfth of the same dose daily for 30 days. Dimethoate is not carcinogenic to rodents. JMPR concluded that although in vitro studies indicate that dimethoate has mutagenic potential, this potential does not appear to be expressed in vivo.

Derivation of maximum acceptable value An acceptable daily intake approach has been used for the derivation of the MAV for dimethoate in drinking-water. JMPR concluded that it was not appropriate to base the ADI on the results of the studies of volunteers, since the crucial end-point (reproductive performance) has not been assessed in humans. It was suggested that there may be a need to re-evaluate the toxicity of dimethoate after the periodic review of the residue and analytical aspects of dimethoate has been completed if it is determined that omethoate is a major residue. The MAV for dimethoate was derived as follows:



where:

• no observable adverse effect level = 1.2 mg dimethoate per kg body weight per day identified on the basis of reproductive performance in a study of reproductive toxicity in rats




• uncertainty factor = 500 to take into consideration concern regarding whether this could be a LOAEL.

References Extoxnet. A Pesticide Information Project of Co-operative Extension Offices of Cornell University, Oregon State University, the University of Idaho, and the University of California at Davis and the Institute for Environmental Toxicology, Michigan State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

FAO/WHO. 1997. Pesticide Residues in Food � 1996 Evaluations: Part II � Toxicological. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/97.1).


WHO. 2003. Dimethoate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/90).

Worthing CR (ed). 1987. The Pesticide Manual: A world compendium (8th ed). Croydon, England: The British Crop Protection Council.


Diquat New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of diquat in drinking-water should not exceed 0.01 mg/L (10 µg/L). WHO (2004) states that because diquat has rarely been found in drinking-water, it is not considered necessary to derive a guideline value. The maximum contaminant level (USEPA 2004) is 0.02 mg/L.

1 To source waters


Diquat is sold as diquat dibromide. It is a rapid acting, non-selective contact herbicide and crop desiccant. It is used to control crop weeds, and aquatic weeds (at or below 1 mg/L). Its use in the control of aquatic weeds presents a direct route of entry into surface water systems.

On a global basis, pre-harvest desiccation to aid the harvesting of seed and fodder crops accounts for the use of two-thirds of the global volume of diquat, whereas one-third of the diquat sold is used as a weed killer (WHO 1998). Diquat is registered for use in New Zealand and is available as liquids (ready to use) or as a soluble concentrate. The soluble concentrate is available with diquat as the single active ingredient (trade names Reglone and Torpedo), or with paraquat as a second active ingredient (trade name Preeglone). No information is available on the annual usage of specific active ingredients in New Zealand, although diquat is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Diquat is very soluble in water: 718 g/L (718,000 mg/L) (WHO 1998). It has a long half life in soil: 250 days (Hort Research 2000).

Diquat is adsorbed strongly to soil, is not taken up by plant roots, and is not metabolically degraded by plants. The rate of degradation in soil, although slow, was found to be sufficient to ensure that diquat residues would not accumulate indefinitely in soil but would reach a plateau level when the amount degraded each year was equal to the amount of new addition. In the presence of sunlight, rapid and extensive photochemical degradation occurs (WHO 1998). When diquat is added to natural waters to control aquatic weeds, residues in the water decline rapidly, owing mainly to the absorption of diquat into the aquatic plants, where it is bound firmly until the decaying weeds disintegrate into the bottom mud. The diquat is then irreversibly bound to the soil particles, leaving the water free of diquat residues. Half lives of diquat in natural waters are generally less than 48 hours (FAO/WHO 1995). Diquat will photodegrade in surface layers of water in 1�3 or more weeks when not adsorbed to particulate matter, (USEPA Technical Factsheet on diquat). There is no information available regarding the greatest source of exposure to diquat for New Zealanders (ie, dermal contact, inhalation, diet: food, water).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included diquat, and so typical concentrations in New Zealand drinking-waters are unknown.


Groundwater was analysed for diquat at two sites in Japan where the product had been used commercially for 5 and 15 years. No diquat was detected in the water, although the limit of detection was relatively high, being 0.1 mg/L (FAO/WHO 1995).

Nanofiltration and reverse osmosis

Removal methods Diquat is adsorbed by activated carbon (Faust and Aly 1983).

Diquat can be oxidised by ozone. This is achieved most effectively at more alkaline pH values, or by a combination of ozone and UV light. Both of these methods produce hydroxyl radicals which are strong oxidising agents (Haag and Yao 1992). Use of activated carbon following ozonisation should be considered to adsorb oxidation products.

may also provide a means of removing this compound from water, but no data are available to support this.

Diquat is an analog of paraquat, which is a pesticide of high toxicity to humans. Diquat is considerably less potent than paraquat, but nonetheless can cause severe acute and chronic poisoning. Paraquat is also registered for use in New Zealand.

Diquat poisoning is much less common than paraquat poisoning, so human reports and animal experimental data for diquat poisoning are less extensive than for paraquat (Reigart and Roberts 1999). Part of the reduced toxicity may be related to the fact that it is absorbed poorly from the gastrointestinal tract. A latency period of 24 hours is seen prior to visible acute toxic effects (Klaassen 1996).


Referee method

Liquid/solid extraction, HPLC with UV detection (EPA Method 549.2).


None recommended.


The mode of action of diquat is not understood.

Acute poisoning

Following acute, high-dose exposure or chronic exposure of animals to diquat, the major target organs were the gastrointestinal tract, the liver, and the kidneys. Oral LD50 values in various species are of the order of 100 to 400 mg/kg (Klaassen 1996), which suggests a moderate level of acute oral toxicity when compared with other pesticides. The acute lethal dose of diquat dibromide is considered to be 6�12 g for humans (WHO 1994). It is considered that diquat can form free radicals, and that the tissue necrosis is associated with the same mechanism(s) of superoxide-induced peroxidation as observed with paraquat, (Klaassen 1996).


In many human diquat poisoning cases (acute exposure), clinical signs of neurologic toxicity are the most important. These include nervousness, irritability, restlessness, combativeness, disorientation, nonsensical statements, inability to recognise friends or family members, and diminished reflexes. Neurologic effects may progress to coma, accompanied by tonic-clonic seizures, and result in the death of the patient (Vanholder et al 1981, Olson 1994, both cited in Reigart and Roberts 1999). Parkinsonism has also been reported following dermal exposure to diquat (Sechi et al 1992, cited in Reigart and Roberts 1999).

An acceptable daily intake approach has been used for the derivation of the MAV for diquat in drinking-water. The ADI was based on cataract formation at the next higher dose identified in a two-year study in rats and using an uncertainty factor of 100. The MAV was derived as follows:

The kidney is the principal excretory pathway for diquat absorbed into the body. Renal damage is therefore an important feature of poisonings.

Chronic exposure

Chronic feeding studies resulted in an increased incidence of cataracts in both dogs and rats (Klaassen 1996 and WHO 1998). Numerous reproductive and developmental toxicity studies have been conducted and have shown reduced maternal and foetal weight gain at certain doses (WHO 1998). The International Agency for Research on Cancer (IARC) has not classified diquat, but the USEPA considers that there is inadequate evidence to state whether diquat has the potential to cause cancer from a lifetime exposure in drinking-water (USEPA Technical Fact Sheet on diquat). No carcinogenic or tumorigenic potential was reported for diquat in long-term feeding studies in rats and dogs (Health Canada 1989).




where:

• no observable adverse effect level = 0.19 mg diquat ion per kg body weight per day identified on the basis of cataract formation at the next higher dose identified in a two-year study in rats




• uncertainty factor = 100. WHO (2004) states that it should be noted that the limit of detection of diquat in water is 0.001 mg/L, and its practical quantification limit is about 0.01 mg/L.

References FAO/WHO. 1994. Pesticide Residues in Food � 1993: Evaluations � 1993: Part II � Toxicology. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/94.4).

FAO/WHO. 1995. Pesticide Residues in Food � 1994: Evaluations � 1994: Part I � Residues. Rome: Food and Agriculture Organisation of the United Nations (FAO Plant Production and Protection Paper 131/1).

Klaassen C. 1996. Casarett and Doull�s Ttoxicology: The basic science of poisons (5th ed), pp. 673�4. McGraw-Hill.


Haag WR, Yao CCD. 1992. Rate constants for reaction of hydroxyl radicals with several drinking-water contaminants. Environ Sci Technol 26: 1005�13.


Vanholder R, Colardyn F, DeReuck J, et al. 1981. Diquat intoxication: Report of two cases and review of the literature. Am J Med 70: 1267�71.

1 To source waters


Olson K. 1994. Paraquat and diquat. In: Olson K, et al (eds). Poisoning and Drug Overdose (2nd ed), pp. 245�6. Norwalk CT: Appleton and Lange.

Reigart R, Roberts J. 1999. Recognition and Management of Pesticide Poisonings (5th ed) chapter 12, pp. 108�17. EPA, Office of Pesticide Programs.

Sechi G, Agnetti V, Piredda M, et al. 1992. Acute and persistent parkinsonism after use of diquat. Neurology 42: 261�3.



WHO. 1994. Paraquat and Diquat. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 39).

WHO. 1998. Guidelines for Drinking-water Quality (2nd ed). Addendum to Volume 2: Health criteria and other supporting information: diquat, pp. 209�17. Geneva: World Health Organization.

WHO. 2003. Diquat in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/91).


Diuron New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of diuron in drinking-water should not exceed 0.02 mg/L (20 µg/L). Diuron is not mentioned in WHO (2004).


Diuron is used as a pre-emergence herbicide. Its uses include control of vegetation in non-crop areas, including irrigation and drainage ditches. It is registered for use in New Zealand and is available in a wide variety of formulations, many containing other active ingredients including amitrole, bromacil, terbuthylazine and norflurazon. Trade names include: Agpro Diuron 800, Boundary, Chemagro Terminex-A, Fenican, Griffen Karmex 80 DF Herbicide. Diuron is also registered for use as an anti-fouling paint for protection of appliances/structures used in an aquatic environment and many of these formulations have copper compounds as additional active ingredients. Trade names include: Alloy Antifouling, Coppercoat Extra, Cruiser Superior, Interclene 165 BWA 900 Bright Red, Intersmooth Tin Free SPC.


No information is available on the annual usage of specific active ingredients in New Zealand, although diuron is understood to constitute approximately 60% of the Urea Derivatives class of pesticides used in the country (P Holland, personal communication).

Forms and fate in the environment Diuron is moderately soluble in water: 42 mg/L (Merck & Co 1996).


Pesticide monitoring of groundwater conducted by Environment Canterbury has detected diuron on four occasions at one location in the Level Plain area in South Canterbury. The concentrations ranged from 0.00008�0.0003 mg/L (Close et al 2001).

The USEPA has ranked diuron fairly high, as a Priority B chemical, with respect to potential for groundwater contamination, and it also rates highly in Agriculture Canada�s ranking of potential leaching agents (Health Canada 1989). It has a half life in soil of 65 days (Hort Research 2000), and a mobility (as Koc) of 530 was recorded from eight soils from Nelson, Marlborough and Hawkes Bay (Close et al 2001). This suggests that diuron is relatively easily adsorbed to organic soil and therefore not very mobile in those conditions. There is no information available regarding the greatest source of exposure to diuron for New Zealanders (eg, dermal contact, inhalation, diet: food, water).

No Ministry of Health drinking-water surveys have included diuron, so typical concentrations in New Zealand drinking-waters are unknown.

Removal methods Diuron is adsorbed by activated carbon (Faust and Aly 1983), and is degraded well by ozone (Camel and Bermond 1998). Use of activated carbon following ozonisation should be considered to adsorb oxidation products.

Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.


Referee method

Liquid/liquid extraction/liquid/solid extraction/high pressure liquid chromatography (EPA 553).


Liquid/liquid extraction/liquid/solid extraction/high pressure liquid chromatography�ultraviolet detection or high pressure liquid chromatography-mass spectrometer (EPA 8321B).

Health considerations Diuron is absorbed from the gastrointestinal and respiratory systems. In humans, it is metabolised within hours by hydroxylation and N-dealkylation, then excreted via the urine. Cows fed very low doses of diuron in their diets had small amounts of residues in whole milk. Cattle fed small amounts accumulated low levels of diuron in fat and muscle, liver, and kidney (Extoxnet 1996).


Acute poisoning

Diuron is of low acute toxicity. Juveniles and animals on protein-deficient diets are more susceptible than adults to the toxic effects of diuron, based on LD50 results (Hayes 1982, cited in Health Canada 1989). The acute oral LD50 for rats is 3400 mg/kg (RSocC 1987) which suggests a moderate to low acute oral toxicity when compared with other pesticides. Some signs of central nervous system depression have been noted at high levels of diuron exposure (Extoxnet 1996).

Chronic exposure

Chronic effects attributed to moderate to high doses of diuron over time include changes in blood chemistry, increased mortality, growth retardation, abnormal blood pigment and anemia. Daily low doses of diuron fed to female rats through three successive generations caused significantly decreased body weight of offspring in the second and third litters. The fertility rate remained unaffected. It is unlikely that diuron will cause reproductive effects in humans at expected levels of exposure. Diuron is teratogenic at high doses but does not appear to be mutagenic. Limited evidence indicates that low level exposures to diuron does not cause cancer. Low doses of diuron over extended periods of time can cause enlargement to the liver and the spleen (Extoxnet 1996). The International Agency for Research on Cancer has not classified diuron for its potential to cause cancer.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for diuron in drinking-water, as follows:


where: • no observable adverse effect level = 0.625 mg/kg body weight per day • average weight of adult = 70 kg

References

Extoxnet. 1996. Extension Toxicology Network Pesticide Information Profiles. A Pesticide Information Project of Co-operative Extension Offices of Cornell University, Oregon State University, the University of Idaho, and the University of California at Davis and the Institute for Environmental Toxicology, Michigan State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

• average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10% • uncertainty factor = 100.

Camel V, Bermond A. 1998. The use of ozone and associated oxidation processes in drinking water treatment. Wat Res 32(11): 3208�22.

Close ME, Rosen MR, Smith VR. 2001. Fate and transport of nitrates and pesticides in New Zealand�s aquifers. In: Rosen MR, White PA (eds). Groundwaters of New Zealand, Chapter 8, pp. 185�220. New Zealand: The Hydrological Society.


Hayes W Jr. 1982. Pesticides Studied in Man. Baltimore, USA: Williams and Wilkins.




Merck & Co. 1996. The Merck Index: An encyclopedia of chemicals, drugs and biologicals (12th edition). Merck Research Laboratories Division of Merck & Co Inc.

Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Diuron (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.


Endosulfan New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of endosulfan in drinking-water should not exceed 0.02 mg/L. WHO (2004) states that because endosulfan occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value.


1 To source waters

Endosulfan is a chlorinated hydrocarbon insecticide of the cyclodiene subgroup which acts as a contact poison in a wide variety of insects and mites. It can also be used as a wood preservative. It is used primarily on food crops like tea, fruits, vegetables and on grains. The commercial product is made up of a mixture of two separate parts (isomers): the alpha and beta configurations. Endosulfan is considered as a single (homogenous) product in this datasheet. Endosulfan contamination does not appear to be widespread in the aquatic environment, but the chemical has been found in agricultural runoff and rivers in industrialised areas where it is manufactured or formulated, as well as in surface water and groundwater samples collected from hazardous waste sites in the USA. Endosulfan is registered for use in New Zealand.

Forms and fate in the environment Endosulfan does not dissolve easily in water. It sticks to soil particles readily. Transport of this pesticide is most likely occur if endosulfan is attached to soil particles in surface runoff. Large amounts of endosulfan can be found in surface water near areas of application. It has also been found in surface water at very low concentrations and has been detected in the air at minute levels. It is has been found, but not quantified, in well water in California. It is not expected to pose a threat to groundwater.

In raw river water at room temperature and exposed to light, both isomers disappeared in four weeks. A breakdown product first appeared within the first week. The breakdown in water is faster (five weeks) under neutral conditions than at more acidic conditions (five months). Under strongly alkaline conditions the half life of the compound is one day. The two isomers have different degradation times in soil. The half life for the alpha isomer is 35 days and 150 days for the beta isomer under neutral conditions. These two isomers will persist longer under more acidic conditions. The compound is broken down in soil by fungi and by bacteria.


The breakdown product, endosulfan sulfate, has been observed in several field studies involving plants. The sulfate is more persistent than the parent compound, accounting for 90% of the residue in 11 weeks. Sulfate formation increases as temperatures increase. However, sunlight may play a role in the reaction, perhaps in starting the process. On most fruits and vegetables, 50% of the parent residue is lost within three to seven days.

Typical concentrations in drinking-water Surface water samples in the USA generally contain less than 0.001 mg/L endosulfan.

Removal methods No data available. However, being attracted to soil particles will enhance its removal in many water treatment processes.


Referee method

[???]


[???]

Health considerations The main source of exposure of the general population is food, but generally residues have been found to be well below the FAO/WHO maximum residue limits. Another important route of exposure to endosulfan for the general population is the use of tobacco products. Acute human exposure to organo-phosphate pesticides (including endosulfan) has been shown to result in the following symptoms: headache, giddiness, nervousness, blurred vision, weakness, nausea, cramps, diarrhea, and discomfort in the chest. Signs include sweating, miosis, tearing, salivation and other excessive respiratory tract secretion, vomiting, cyanosis, papilledema, uncontrollable muscle twitches followed by muscular weakness, convulsions, coma, loss of reflexes, and loss of sphincter control. The last four signs are seen only in severe cases but do not preclude a favourable outcome if treatment is prompt and energetic. Cardiac arrhythmias, various degrees of heart block, and cardiac arrest may occur (Hayes and Laws). Endosulfan is most likely to affect kidneys, liver, blood chemistry and the parathyroid gland. JMPR concluded that endosulfan is not genotoxic, and no carcinogenic effects were noted in long-term studies using mice and rats. The kidney is the target organ for toxicity. Several recent studies have shown that endosulfan, alone or in combination with other pesticides, may bind to estrogen receptors and perturb the endocrine system. Endosulfan is not classified as to its carcinogenicity to humans, and is not mutagenic in animal studies. Endosulfan exhibits weak oestrogenic activity, but no reproducible evidence of an effect in vivo.

The MAV for endosulfan in drinking-water was derived as follows:



where:


• allowable daily intake (ADI) = 0.006 mg/kg body weight based on results from a two-year dietary study of toxicity in rats, and supported by a 78-week study in mice, a one-year study in dogs, and a developmental toxicity study in rats



• average quantity of water consumed by an adult per day = 2 L.

References Extoxnet. Pesticide Information Profile. New York: PMEP, Cornell University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

FAO/WHO. 1999. Pesticide Residues in Food � 1998 Evaluations: Part II � Toxicological. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/99.18).

Hayes WJ Jr, Laws ERJ. 1991. Handbook of Pesticide Toxicology: Classes of Pesticides 2: 938. New York: Academic Press Inc.

1 To source waters

WHO. 2003. Endosulfan in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/92).


Endrin New entry August 2005.

Maximum acceptable value Based on health considerations, the concentration of endrin in drinking-water should not exceed 0.001 mg/L. The maximum contaminant level (USEPA 2004) is 0.002 mg/L.


Endrin is a broad-spectrum foliar insecticide that acts against a wide range of agricultural pests. It is also used as a rodenticide. Small amounts of endrin are present in food, but the total intake from food appears to be decreasing. Being a persistent organic pollutant, it is not currently registered for use in New Zealand.

Forms and fate in the environment The most important route of water contamination is surface run-off from soil and crops. Run-off is affected by numerous, complex factors, such as intensity of precipitation, irrigation practices, soil permeability, topographic relief, organic content of the soil, and the degree of vegetative cover. Soils of low permeability and low organic content allow copious run-off after heavy precipitation.

Although endrin has strong absorptive properties in soils such as clay and sandy loam, limited residues were found. Far greater retention was found in soils with a high organic content, in which it was adsorbed quickly and was difficult to remove. The degree to which endrin was retained in the soil depended not only on the soil type but on numerous other factors such as volatilisation, leaching, wind erosion, surface run-off, and crop uptake Its half-life in soil can be as long as 12 years.


Typical concentrations in drinking-water In an area of the US that was treated with endrin, the mean residue levels in water samples was 0.0004 mg/L (IPCS). Traces of endrin have been found in the drinking-water supplies of several countries (WHO 2004).

Removal methods 0.0002 mg/L should be achievable using GAC (WHO 2004).


Referee method

[???]


[???]

Health considerations Toxicological data are insufficient to indicate whether endrin is a carcinogenic hazard to humans.


Unlike dieldrin, its stereoisomer, endrin is metabolised rapidly by animals, and very little is accumulated in fat in comparison with compounds of similar chemical structure. The primary site of action of endrin is the central nervous system.

The MAV for endrin in drinking-water was derived as follows:


where:

• no observable adverse effect level = 0.025 mg/kg body weight per day based on a two-year study in dogs


• uncertainty factor = 100 for intra- and interspecies variation.



Endrin is listed under the Stockholm Convention on Persistent Organic Pollutants. Hence, monitoring may occur in addition to that required by drinking-water guidelines.

References FAO/WHO. 1995. Pesticide Residues in Food � 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment, and WHO Toxicological and Environmental Core Assessment Groups. Rome, Food and Agriculture Organization of the United Nations (FAO Plant Production and Protection Paper 127).


IPCS. 1992. Endrin. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 130). This is still available on the internet at: http://www.inchem.org/documents/ehc/ehc/ehc130.htm


WHO. 2003. Endrin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. World Health Organization, Geneva. (WHO/SDE/WSH/03.04/93).


Fenitrothion New entry August 2005

Maximum acceptable value There are insufficient data to determine a MAV for fenitrothion in drinking-water. WHO (2004) states that because fenitrothion occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value.


1 To source waters

Fenitrothion is mainly used in agriculture for controlling insects on rice, cereals, fruits, vegetables, stored grains and cotton and in forest areas. It is also used for the control of flies, mosquitos and cockroaches in public health programmes and/or indoor use. Fenitrothion is found in a variety of commercial insecticides. As at August 2005, fenitrothion is registered for use in New Zealand.

Form and fate in the environment Fenitrothion is stable in water only in the absence of sunlight or microbial contamination. In soil, biodegradation is the primary route of degradation, although photolysis may also play a role. Preliminary data indicates fenitrothion degrades fairly rapidly in soil with a half life of less than one week in non-sterile muck, sandy loam soils.

Fenitrothion residues detected in water were low (maximum 0.0013 mg/L) during the forest spray programme. Following the spraying of forests, water samples did not contain detectable amounts of fenitrothion; post-spray samples contained <0.00001 mg/L. Levels of fenitrothion residues in fruits, vegetables and cereal grains decline rapidly after treatment, with a half life of 1�2 days.

Typical concentrations in drinking-water No data available.


Analytical methods

Referee method

A referee method cannot be selected for chloroacetones because a MAV has not been established and therefore the sensitivity required for the referee method is not known.



No alternative methods can be recommended for chloroacetones for the above reason. However, the following information may be useful: ??????????

Health considerations On the basis of testing in an adequate range of studies in vitro and in vivo, JMPR concluded that fenitrothion is unlikely to be genotoxic. It also concluded that fenitrothion is unlikely to pose a carcinogenic risk to humans. In long-term studies of toxicity, inhibition of cholinesterase activity was the main toxicological finding in all species. When volunteers were given single oral doses ranging from 2.5 to 20 mg/person, the maximal concentration of p-nitro-m-cresol in the urine was reached within 12 hours, and nearly the entire amount discharged was eliminated during the first 24 hours.

Derivation of maximum acceptable value A MAV for fenitrothion in drinking-water could be derived as follows:


where:

• allowable daily intake (ADI) = 0.5 mg/kg body weight per day for inhibition of brain and erythrocyte cholinesterase activity in a two-year study of toxicity in rats, and supported by a NOAEL of 0.57 mg/kg of body weight per day for inhibition of brain and erythrocyte cholinesterase activity in a three-month study of ocular toxicity in rats, and a NOAEL of 0.65 mg/kg of body weight per day for reduced food consumption and body weight gain in a study of reproductive toxicity in rats






FAO/WHO. 2001. Pesticide Residues in Food � 2000 evaluations. Part II � Toxicological. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/01.3).

WHO. 2003. Fenitrothion in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/95).


Fenoprop Revised July 2005. (Also called silvex or 2,4,5-TP.)


Maximum acceptable value Based on health considerations, the concentration of fenoprop in drinking-water should not exceed 0.01 mg/L. The maximum contaminant level (USEPA 2004) is 0.05 mg/L.


1 To source waters

Fenoprop may enter source waters as a result of its use as a selective systemic hormone-type herbicide. It is absorbed by the roots and leaves and is used to control woody plants on non-crop areas and pastures and broadleaved weeds in a variety of crops. It is not currently registered in New Zealand but has been in the past.

Forms and fate in the environment In soil fenoprop is degraded to 2,4,5-trichlorophenol which is very resistant to further breakdown. Water solubility ranges from 150 to 176 mg/L.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 157 zones, did not find detectable concentrations of fenoprop (limit of detection = 0.0001 mg/L). Chlorophenoxy herbicides not frequently found in drinking-water; when detected, concentrations are usually no greater than a few micrograms per litre (WHO 2004).

Removal methods No specific information on methods of removing fenoprop from water is available. However, slow sand filtration can partially remove members of the chlorophenoxy acid pesticide family of which fenoprop is a member. Ozone has shown varying degrees of effectiveness in oxidising the chlorophenoxy acids.


Referee method




Health considerations In general, chlorophenoxy herbicides are absorbed rapidly from the gastro-intestinal tract and are distributed evenly throughout the body. Accumulation in human tissues is not expected, and a steady-state level in the human body will be achieved within 3�5 days of exposure. Elimination occurs primarily in the urine, mostly in the unchanged form. Biological half lives of chlorophenoxy herbicides in mammals range from 10 to 33 hours. Metabolic conversions occur only at high doses. The salt and ester forms are hydrolysed rapidly and follow the same pharmacokinetic pathways as the free acid forms. No adverse effects were reported following ingestion of a single dose of 1 mg fenoprop/kg body weight by eight human volunteers.


The results from short and long-term exposure studies in animals report depressed body weight gain and liver and kidney damage at elevated doses. Effects observed in long-term studies with beagle dogs given fenoprop in the diet include mild degeneration and necrosis of hepatocytes and fibroblastic proliferation in one study and severe liver pathology in another study. In rats, increased kidney weight was observed in two long-term dietary studies. Fenoprop has not exhibited mutagenic activity in a bacterial assay test. Chlorophenoxy herbicides as a group, including 2,4-D and MCPA, have been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). However, based on the available data from studies on exposed populations and on animals,it is not possible to assess the carcinogenic potential of any specific chlorophenoxy herbicide. Therefore, drinking-water guidelines for these compounds are based on a threshold approach for other toxic effects.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for fenoprop in drinking-water. The no observable adverse effect level used in the derivation is from a study in which beagle dogs were administered fenoprop in the diet for two-years. The MAV for fenoprop in drinking-water was derived as follows:


where:

• no observable adverse effect level = 0.9 mg/kg body weight per day for adverse effects on the liver in a study in which beagle dogs were administered fenoprop in the diet for two years




• uncertainty factor = 300 (100 for inter and intra-species variation and 3 for limitations of the data base).



Que Hee SS, Sutherland RG. 1981. The Phenoxyalkanoic Herbicides. Volume I. Chemistry, analysis, and environmental pollution. Boca Raton, USA: CRC Press (Chemical Rubber Company Series in Pesticide Chemistry).



WHO. 2003. Chlorophenoxy herbicides (excluding 2,4-D and MCPA) in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/44).



Glyphosate New entry August 2005.

Maximum acceptable value There are insufficient data to determine a MAV for glyphosate in drinking-water. WHO (2004) states that because glyphosate occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value. The maximum contaminant level (USEPA 2004) is 0.7 mg/L.


1 To source waters

Glyphosate is a broad-spectrum herbicide used in both agriculture and forestry, and for aquatic weed control. Glyphosate is usually formulated as an isopropylamine salt. A common trade name is Roundup. Glyphosate can enter surface and subsurface waters after direct use near aquatic environments or by runoff or leaching from terrestrial applications. As at August 2005, glyphosate is registered for use in New Zealand.

Forms and fate in the environment Microbial biodegradation of glyphosate occurs in soil, aquatic sediment and water, the major metabolite being aminomethylphosphonic acid (AMPA) (CAS No 1066-51-9). Glyphosate is chemically stable in water and is not subject to photochemical degradation. The low mobility of glyphosate in soil indicates minimal potential for the contamination of groundwater. Glyphosate is highly adsorbed on most soils, especially those with high organic content. The compound is so strongly attracted to the soil that little is expected to leach from the applied area. Microbes are primarily responsible for the breakdown of the product. The time it takes for half of the product to break down ranges from 1 to 174 days. Because glyphosate is bound so tightly to the soil, little is transferred by rain or irrigation water. One estimate showed less than two percent of the applied chemical lost to runoff. The herbicide could move when attached to soil particles in erosion run-off. Photodecomposition plays only a minor role in environmental breakdown.

Typical concentrations in drinking-water No information available.

Removal methods No information available. However, being so strongly adsorbed to soil suggests many water treatment processes, especially chemical coagulation processes, should remove glyphosate from the raw water.

Analytical methods

Referee method

A referee method cannot be selected for glyphosate because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for glyphosate for the above reason. However, the following information may be useful: ????????


Health considerations Glyphosate and AMPA have similar toxicological profiles, and both are considered to exhibit low toxicity. While it can be described as an organophosphorus compound, glyphosate is not an organophosphate ester but a phosphanoglycine, and it does not inhibit cholinesterase activity. Glyphosate is absorbed poorly from the digestive tract and is excreted largely unchanged by mammals. Ten days after treatment there were only minute amounts in the tissues of rats fed glyphosate for three weeks. Subchronic and chronic tests with glyphosate have been conducted with rats, dogs, mice, and rabbits in studies lasting from 21 days to two years. With few exceptions there were no treatment-related gross (easily observable) or cellular changes. In a chronic feeding study with rats, no toxic effects were observed in rats given doses as high as 31 mg/kg/day, the highest dose tested. No toxic effects were observed in a chronic feeding study with dogs fed up to 500 mg/kg/day, the highest dose tested. Mice fed glyphosate for 90 days exhibited reduced body weight gains. The lifetime administration of very high amounts of glyphosate produced only a slight reduction of body weight and some microscopic liver and kidney changes. Blood chemistry, cellular components, and organ function were not affected even at the highest doses. The USEPA has stated that there is sufficient evidence to conclude that glyphosate is not carcinogenic in humans.

Derivation of maximum acceptable value A health-based value of 0.9 mg/L could be derived, based on the group ADI for AMPA alone, or in combination with glyphosate of 0.3 mg/kg of body weight, based upon a NOAEL of 32 mg/kg of body weight per day, the highest dose tested, identified in a 26-month study of toxicity in rats fed technical-grade glyphosate and using an uncertainty factor of 100. Because of their low toxicity, the health-based value derived for AMPA alone, or in combination with glyphosate, is orders of magnitude higher than concentrations of glyphosate or AMPA normally found in drinking-water. Under usual conditions, therefore, the presence of glyphosate and AMPA in drinking-water does not represent a hazard to human health. For this reason, the establishment of a guideline value for glyphosate and AMPA is not deemed necessary.


FAO/WHO. 1998. Pesticide Residues in Food � 1997 Evaluations. Part II � Toxicological and environmental. Geneva: World Health Organization. Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/98.6).


IPCS. 1994. Glyphosate. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 159).

USEPA. 1987. Glyphosate Health Advisory. Environmental Protection Agency, Office of Drinking Water.

USEPA. 1992. Pesticide tolerance for glyphosate: US Environmental Protection Agency. Federal Register 57(49): 8739�40.



WHO. 2003. Glyphosate and AMPA in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/97).


Heptachlor and heptachlor epoxide Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of heptachlor in drinking-water should not exceed 0.00004 mg/L (0.04 µg/L). WHO (2004) states that because heptachlor and heptachlor epoxide occur at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value. The maximum contaminant level (USEPA 2004) is 0.0004 mg/L for heptachlor and 0.0002 mg/L for the epoxide.


1 To source waters

Heptachlor may enter source waters as a result of its application as a soil treatment and a seed treatment, or if it is applied directly to foliage to control a wide variety of insects. Heptachlor epoxide is an oxidation product of heptachlor. Heptachlor is not currently used in New Zealand. It was never fully registered and its highest status was provisional B, which was cancelled in the early 1970s. Heptachlor was used in Australia until September 1994 to protect wooden structures against termites. Its other former uses were withdrawn in the late 1970s and early 1980s.

Forms and fate in the environment Heptachlor is moderately persistent in soil where it is transformed mainly to its epoxide. It binds to soil particles and migrates slowly. Its half life in soil ranges from nine months to two years. Heptachlor epoxide is resistant to further chemical or biological changes in soil and also binds to soil and migrates slowly. The water solubility is 0.056 mg/L for heptachlor and 0.35 mg/L for heptachlor epoxide. Diet is likely to be the greatest source of exposure to heptachlor epoxide. Heptachlor and heptachlor epoxide have been found in many food classes. Human milk can be contaminated with heptachlor epoxide.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of heptachlor and heptachlor epoxide (limit of detection = 0.00001 mg/L). Heptachlor and heptachlor epoxide have been found in drinking-water at levels of nanograms per litre (WHO 2004).


Removal methods No information is available on methods of removing heptachlor or heptachlor epoxide from water. However, isotherm adsorption data also indicate that removal by adsorption on to granular activated carbon should be possible.


Referee method

Liquid/Liquid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 505).



2 Heptachlor can be extracted from water using liquid/liquid extraction, and analysed using gas chromatography with electron capture detection (APHA Method 6630C).

Health considerations Heptachlor is absorbed rapidly from the gastro-intestinal tract of rats following intragastric administration. Heptachlor is metabolised to heptachlor epoxide and other metabolites which are distributed throughout the body. The liver appears to be the target organ for heptachlor toxicity and oral administration of heptachlor enhanced the incidence of liver tumours induced in mice by oral administration of N-nitrosodiethylamine. Clinical case studies of humans subject to acute exposure of chlordane-containing heptachlor (via ingestion, skin or inhalation routes) report central nervous system effects such as irritability, salivation, laboured respiration, muscle tremors and convulsions. The International Agency for research on Cancer has classed heptachlor in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value The Joint FAO/WHO Meetings on Pesticide Residues (JMPR) have evaluated heptachlor on several occasions and in 1991 established as Acceptable Daily Intake (ADI) of 0.1 µg/L of body weight on the basis of a no-observable-adverse effect level of 0.025 mg/L. The MAV for heptachlor in drinking-water was derived as follows:

0.025 mg/kg body weight/day x 70 kg x 0.01 = 0.000044 mg/L (0.044 µg/L) 2 L/day x 200

(rounded to 0.00004 mg/L, or 0.04 µg/L, or 40 ng/L)

where:

• no observable adverse effect level = 0.025 mg/kg body weight per day from two studies in the dog




• uncertainty factor = 200 (100 for inter and intra-species variation and 2 for the inadequacy of the data base).





FAO/WHO. 1992. Pesticide Residues in Food � 1991: Evaluations � 1991. Part II. Toxicology. Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/92.52).

FAO/WHO. 1995. Pesticide Residues in Food � 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and WHO Toxicological and Environmental Core Assessment Groups. Rome, Food and Agriculture Organization of the United Nations (FAO Plant Production and Protection Paper 127).



WHO. 2003. Heptachlor and Heptachlor Epoxide in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/99).


Hexachlorobenzene Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of hexachlorobenzene in drinking-water should not exceed 0.0001 mg/L (0.1 µg/L). WHO (2004) states that because the health-based values are considerably higher than the concentrations at which HCB is detected in drinking-water (ie, sub-nanograms per litre), when it is detected, it is not considered necessary to establish a guideline value for HCB in drinking-water. The maximum contaminant level (USEPA 2004) is 0.001 mg/L.


1 To source waters

Hexachlorobenzene may enter source waters as a result of its application as a selective fungicide, used to control dwarf bunt in wheat. Hexachlorobenzene is not presently registered in New Zealand and has never reach full registration, although it did gain experimental use status. HCB is produced as a by-product or waste material in the production of tetrachloroethylene, trichloroethylene, carbon tetrachloride, chlorine, dimethyl tetrachloroterephthalate, vinyl chloride, atrazine, propazine, simazine, pentachloronitrobenzene (quintozene), and mirex. It is a contaminant in several pesticides including dimethyl tetrachlorophthalate and pentachloronitrobenzene. At present, it appears mainly as a by-product of several chemical processes or an impurity in some pesticides.

Forms and fate in the environment Hexachlorobenzene photolyses slowly in the atmosphere with a half life of about 80 days. It has low solubility in water (0.005 mg/L) and, despite its relatively low vapour pressure, volatilises from water at a


significant rate. Biotransformation in surface water, sludge, or soil suspensions is extremely slow. Hexachlorobenzene is adsorbed strongly by soil and sediments. It can bioaccumulate in aquatic organisms.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of hexchlorobenzene (limit of detection = 0.0001 mg/L). It has been detected only infrequently, and at very low concentrations (below 0.0001 mg/L), in drinking-water supplies (WHO 2004).

Removal methods No information is available on methods for removing hexachlorobenzene from water. However, isotherm adsorption data also indicate that removal by adsorption on to granular activated carbon should be possible.



Referee method

Liquid/Solid Extraction and Capillary Column Gas Chromatography/Mass Spectrometry (EPA 525).



Health considerations Animal studies have shown that following administration, hexachlorobenzene was detected in adipose tissue, bone marrow, skin, the Harderian gland, nasal mucosa and the preputial gland. It is metabolised slowly into lower chlorinated benzenes, chlorinated phenols and other metabolites, and excreted principally in faeces. Symptoms observed with acute exposure in animals include convulsions, tremors, weakness, loss of coordination, paralysis and pathological changes in organs. Short-term exposure studies in animals reported symptoms principally affecting the spleen, liver and kidneys. The International Agency for Research on Cancer has evaluated the evidence for carcinogenicity of hexachlorobenzene in animals and humans and have classed it in Group 2B (possibly carcinogenic to humans). Hexachlorobenzene has been shown to induce tumours in three animal species and at a variety of sites. No report of a direct association between hexachlorobenzene and human cancer is available. Hepatocellular carcinoma has been associated with porphyria (abnormal porphyrin metabolism). However, while this persisted at least 20 years after an epidemic of porphyria cutanea tarda in Turkey, caused by consumption of grain treated with hexachlorobenzene, no excess cancer occurrence was reported in this population 25 years after the accident.

Derivation of maximum acceptable value Hexachlorobenzene has been shown to induce tumours in three animal species and at a variety of sites. A linear low-dose extrapolation model was therefore used to calculate a reference dose. On the basis of


liver tumours observed in female rats in a two-year dietary study, and applying the linearised multistage model, the concentration of hexachlorobenzene associated with an excess lifetime cancer risk of one per one hundred thousand (10-5) is 0.001 mg/L (1 µg/L). Hexachlorobenzene is listed under the Stockholm Convention on Persistent Organic Pollutants.

References AWWA. 1990. Water Quality and Treatment (4th ed). Pontius FW, McGraw, Hill (eds).

IPCS. 1997. Hexachlorobenzene. Geneva: World Health Organization, International Programme on Chemical Safety (Environmental Health Criteria 195).



WHO. 2003. Hexachlorobenzene in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/100).


Hexazinone New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of hexazinone in drinking-water should not exceed 0.4 mg/L (400 µg/L). Hexazinone is not mentioned in WHO (2004).


1 To source waters

Hexazinone is used as a broad spectrum pre- and post- emergence triazine herbicide effective against woody and herbaceous weeds. It is registered for use in New Zealand and is available in a variety of formulations, some of which contain terbuthylazine or atrazine as additional active ingredients. Trade names are: Agpro Valzine 500, Agpro Valzine Extra, Forest Mix Special Blend Granular Herbic, Release Ultra, Release, Velgard, and Velpar. No information is available on the annual usage of specific active ingredients in New Zealand, although hexazinone is understood to have its largest use in forestry (P Holland, personal communication).

Forms and fate in the environment Hexazinone is very soluble in water: 330 g/L (330,000 mg/L) (Merck & Co 1996). Based on laboratory data and confirmed by field and forestry data, hexazinone appears to be persistent and mobile in soil and aquatic environments. Hexazinone was reported in runoff water up to six months post-treatment in a forestry dissipation study. Hexazinone has been detected in groundwater (at levels well below the Health Advisory) in Hawaii, Florida, Maine and North Carolina. Hexazinone also can contaminate surface water by spray drift at application, and for several months post-application via runoff (USEPA 1994). Close et al 1999 have reported a half life for hexazinone of 113 days and a mobility (Koc) of 50�60 in Horotiu sandy loam, which indicates a moderate level of adsorption to organic soil.


There is no information available regarding the greatest source of exposure to hexazinone for New Zealanders (ie, dermal contact, inhalation, diet: food, water). Based on international studies, the dietary risk posed by hexazinone is expected to be minimal. Exposure to workers and other applicators generally is not expected to pose undue risks, due to hexazinone�s overall low acute toxicity (USEPA 1994).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included hexazinone. In the New Zealand national pesticides surveys, conducted every four years since 1990, hexazinone has been detected in groundwaters twice, at concentrations of 0.0012 and 0.0023 mg/L. Pesticide monitoring of groundwater conducted by Environment Canterbury has detected hexazinone at two locations in the Level Plain area in South Canterbury. At the first location the concentration was 0.00018 mg/L, and at the second location the concentrations ranged from 0.00004�0.00079 mg/L (Close et al 2001). Hexazinone has been detected in groundwater at two sites in the Waikato region. Concentrations ranged from 0.00004�0.00024 mg/L (Hadfield and Smith 1999). Hexazinone has been detected in groundwater in the Edendale area (Southland) at concentrations ranging between 0.00004�0.00049 mg/L (Hughes 2000). In California hexazinone was detected in groundwater and reported in the 1995 well inventory report (DPR 1995). In a Canadian one-year study of hexazinone on short, gravity-irrigated runs in 1991 in southern Alberta, the herbicide was detected in 50% of runoff samples and about 27% of groundwater samples. No Canadian water quality guideline exists for hexazinone, but all detections were well below the US lifetime health advisory limit for drinking water of 0.21 mg/L (Minister of Public Works and Government Services, Canada 2000).

Removal methods Specific information about the removal of hexazinone from water is unavailable, however, oxidation of triazines (hexazinone is a member of this chemical family) by ozone is reported to be effective. The water chemistry, in particular the alkalinity and pH, will affect the oxidation rate (Chiron et al 2000). Use of activated carbon following ozonisation should be considered to adsorb oxidation products. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of hexazinone, although a guide to the efficiency of the process cannot be provided. Nanofiltration (membrane technology) in water with a low natural organic matter concentration is reported to remove approximately 50% of atrazine and simazine (Agbekodo et al 1996). The percentage is increased to 90�100% when 3.6 mg/L of natural organic matter is present. Similar results may be expected for hexazinone as it is from the same chemical family and of comparable size.


Referee method



None recommended.

Health considerations Almost all of a 14 mg/kg oral dose administered to rats was excreted in three to six days, with the majority in urine. In another study, animals fed 125 mg/kg for two weeks and then given a small single


dose, excreted almost all of the product within three days. Less than 1% of the parent hexazinone was detected in urine and faeces. There does not appear to be any significant tissue accumulation (USEPA 1987, cited in Extoxnet 2001).

Acute poisoning

In acute toxicity studies using laboratory animals, hexazinone has been shown to be a severe eye irritant and has been placed in Toxicity Category I (the highest of four categories) for primary eye irritation. It is slightly toxic through the acute oral route (Toxicity Category III) (USEPA 1994). It may irritate the eyes, nose and throat of humans (Extoxnet 2001). The acute oral LD50 for rats is 1690 mg/kg, for guinea pigs 860 mg/kg (RSocC 1987), which suggests a moderate acute oral toxicity when compared with other pesticides.

Chronic exposure

USEPA�s Office of Water has issued a lifetime Health Advisory which sets a maximum level of 0.21 mg/L allowable in drinking-water. The critical effect of chronic exposure to hexazinone is decreased body weight (USEPA 1998). The International Agency for Research on Cancer (IARC) has not classified hexazinone for its potential to cause cancer, but the USEPA has classified it as a Group 4 carcinogen, a chemical that is not classifiable as to human carcinogenicity (USEPA 1994).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for hexazinone in drinking-water, as follows:

10 mg/kg body weight per day x 70 kg x 0.1 = 0.35 mg/L, rounded to 0.4 mg/L 2 L x 100

where: • no observable adverse effect level = 10 mg /kg body weight per day • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10% • uncertainty factor = 100.



Close ME, Rosen MR, Smith VR. 2001. Fate and transport of nitrates and pesticides in New Zealand�s aquifers. In: Rosen MR, White PA (eds). Groundwaters of New Zealand, Chapter 8, pp. 185�220. New Zealand: The Hydrological Society.

Close ME, Pang L, Magesan GN, et al. 1999. Field study of pesticide leaching in an allophanic soil in New Zealand � model simulations using GLEAMS. In: Abstracts of the Workshop on Environmental Aspects of Pesticide Use. Hamilton, November 1999.

DPR. 1995. Well Inventory Report. California, USA: Department of Pesticide Regulation.

Extoxnet. 2001. Pesticide Information Profile: Hexazinone. New York: PMEP, Cornell University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/



Hughes B. 2000. Edendale Pesticide Investigation Report 2000. Southland Regional Council Publication No 2000�14.


Minister of Public Works and Government Services, Canada. 2000. In Coote DR, Gregorich LJ (eds). The Health of our Water: Toward sustainable agriculture in Canada. Research Branch, Agriculture and Agri-Food Canada. Publication 2020/E.

Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Hexazinone (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

USEPA. 1994. EPA RED Facts: Hexazinone. Office of Prevention, Pesticides and Toxic Substances (H-7508W). EPA-738-F-94-019.

USEPA. 1987. Health Advisory. Office of Drinking Water.


USEPA. 1998. Integrated Risk Information System (IRIS) Substance List: Hexazinone.

Isoproturon Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of isoproturon in drinking-water should not exceed 0.01 mg/L.


1 To source waters

Isoproturon may enter source waters as a result of its application as a selective systemic herbicide, used to control annual grasses and broadleaf weeds in cereals. It is not presently used in New Zealand, although as at August 2005, it is registered for use.

Forms and fate in the environment Isoproturon is mobile in soil and has been detected in surface and groundwater overseas. It is quite persistent in water and hydrolyses slowly with a half life of about 30 days. It undergoes degradation and hydrolysis in soil with a half life of about 40 days. The water solubility of isoproturon is 72 mg/L.

Typical concentrations in drinking-water No data are available on the concentration of isoproturon in New Zealand drinking-water supplies. In Germany, concentrations between 0.0001 and 0.000125 mg/L (0.1 and 0.125 µg/L) have been found in surface waters. Levels above 0.0001 mg/L (0.1 µg/L) have been found in drinking-waters in the UK. Has been detected in surface water and groundwater, usually at concentrations below 0.0001 mg/L; levels above 0.0001 mg/L have occasionally been detected in drinking-water (WHO 2004).


Removal methods No information on methods of removing isoproturon from water is available. However, chlorine has been reported to be effective in the break down of this family of pesticides. Slow sand filtration has no effect on the concentrations of these pesticides. 0.1 mg/L should be achievable using ozonation (WHO 2004).


Referee method

No referee method has been given for isoproturon because no method meets the required criteria.


No alternative methods have been recommended for isoproturon because no methods meet the required criteria. However, the following information may be useful: Isoproturon may be determined in water samples by separation with reverse-phase high-performance liquid chromatography and ultraviolet or electrochemical detection. Detection limits between 0.00001 and 0.0001 mg/L (10�100 ng/L) have been reported. High levels of phenoxyacidic herbicides may interfere with the determination of isoproturon.

Health considerations Isoproturon is absorbed readily and rapidly when given orally. Distribution is rapid, and no accumulation of isoproturon in any particular organ or tissue has been reported. Isoproturon is metabolised and excreted rapidly by the rat. Isoproturon is of low acute oral toxicity and low to moderate toxicity following short-and long-term exposures in mammals. It does not cause skin and eye irritation or sensitisation after repeated skin exposure. It does not possess significant genotoxic activity, but it causes marked enzyme induction and liver enlargement. Isoproturon caused an increase in hepatocellular tumours in male and female rats, but this was apparent at doses which also caused liver toxicity. Isoproturon appears to be a tumour promoter, rather than a complete carcinogen. Isoproturon has been in commercial use for a relatively short period, and so far no cases of human poisoning have been reported. Data on human health effects of isoproturon are limited to studies involving occupational exposures. One three-year study was carried out on a group of workers employed in various parts of the manufacturing process. Following urine and blood analysis, the authors reported no pathological abnormalities in the peripheral blood count or any indication of haemolytic anaemia.

Derivation of maximum acceptable value As isoproturon is considered to be a tumour promoter rather than a complete carcinogen, a tolerable daily intake approach has been used for the derivation of the MAV. The no observable adverse effect level used in the derivation of the MAV is from a study in dogs and a two-year feeding study in rats. The MAV for isoproturon in drinking-water was derived as follows:


where:

• no observable adverse effect level = 3 mg/kg body weight per day a 90-day study in dogs and a two-year feeding study in rats





• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 because there is evidence of nongenotoxic carcinogenicity in rats).

References Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466�[Add page number].

WHO. 2003. Isoproturon in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/37).


Lindane Revised July 2005. (Also called HCH-hexachlorocyclohexane or gammaHCH.)

Maximum acceptable value Based on health considerations, the concentration of lindane in drinking-water should not exceed 0.002 mg/L (2 µg/L). The maximum contaminant level (USEPA 2004) is 0.0002 mg/L.


1 To source waters

Lindane (the gamma isomer of hexachlorocyclohexane) may enter source waters as a result of its use as an insecticide on fruit and vegetable crops, for seed treatment, and in forestry. It is also used as a therapeutic pesticide (eg, treatment of scabies and head lice) in humans and animals. The total annual usage of lindane in New Zealand in the late 1980s was 1500 kg, all of it in the South Island. Usage before that was probably much greater than this amount. Its registration was cancelled in 1989.

Forms and fate in the environment Lindane can be degraded in soil with half lives ranging from 90 to 1400 days. The recommended average half life in soil is 400 days. The water solubility is low (7 mg/L) and leaching to groundwater is found rarely overseas. In surface waters lindane can be removed by evaporation.

Typical concentrations in drinking-water Lindane was not detected in 230 samples received from 212 New Zealand supplies between 1988 and 1992. The detection limit was approximately 0.00001 mg/L (0.01 µg/L). The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of lindane (limit of detection = 0.0001 mg/L). Has been detected in both surface water and groundwater, usually at concentrations below 0.0001 mg/L, although concentrations as high as 0.012 mg/L have been measured in wastewater-contaminated rivers (WHO 2004).


Removal methods Partial removal of lindane has been reported for slow sand filtration. However, isotherm adsorption data also indicate that removal by adsorption on to granular activated carbon should be possible.


Referee method






Health considerations Following oral administration, absorption of lindane is almost complete. In humans, the lindane content seems to increase with age, but no correlation has been established with levels or duration of exposure. Higher levels of the b-isomer are found in >80% of postmortem human adipose tissue samples. Lindane crosses the placenta and can also be present in human milk. Metabolism of lindane in animals and humans is via dehydrochlorination, dechlorination, dehydrogenation and oxidation. The final metabolites are isomers of dichlorophenol, trichlorophenol and tetrachlorophenol. The most commonly reported effects associated with oral or occupational exposure to lindane are neurophysiological and neuropsychological disorders and gastro-intestinal disturbances. Deaths of humans (usually children) have been reported following ingestion of lindane. In a study conducted in an Indian pesticide factory where handlers were directly exposed to hexachlorocyclohexane for 7�30 years, 94% of the handlers and 69% of non-handlers reported paraesthesia of the face and extremities. Headache and giddiness occurred in over 70% of the handlers and 40% of the non-handlers, compared with 7% of the control group. There is no evidence of genotoxicity from in vivo and in vitro short-term studies. Lindane causes liver tumours in mice given very high doses, but there is no evidence that this is a result of tumour promotion. JMPR has concluded that there was no evidence of genotoxicity. In the absence of genotoxicity. The International Agency for Research on Cancer has classed lindane in Group 2B (possibly carcinogenic to humans).

Derivation of maximum acceptable value Due to the lack of evidence of the carcinogenicity of lindane to humans, a tolerable daily intake approach has been used for the derivation of the MAV of lindane in drinking-water. The no observable adverse effect level used in the derivation is based on results of a liver and kidney toxicity/carcinogenicity short-term study in rats, in which an increased incidence of periacinar hepatocellular hypertrophy, increased liver and spleen weights, and increased mortality occurred at higher doses.

The MAV for lindane in drinking-water was derived as follows:



where:

• no observable adverse effect level = 0.5 mg/kg body weight per day based on liver and kidney toxicity observed in a short-term study in the rat




• uncertainty factor = 100 for inter and intra-species variation.




FAO/WHO. 2002. Pesticide Residues in Food � 2002. Rome. Food and Agriculture Organization of the United Nations. Joint FAO/WHO Meeting on Pesticide Residues (FAO Plant Production and Protection Paper 172).







WHO. 2003. Lindane in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/102).


Malathion, a non-systemic organophosphate insecticide, is commonly used to control mosquitos and a variety of insects that attack fruits, vegetables, landscaping plants and shrubs. It can also be found in other pesticide products used indoors, on pets to control ticks and insects and to control human head and

Malathion New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of malathion in drinking-water should not exceed 1 mg/L. WHO (2004) states that because malathion occurs at concentrations well below those at which toxic effects are observed, it is not considered necessary to derive a guideline value.


1 To source waters


body lice. Maldison is a common trade name. As at August 2005, maldison is registered for use in New Zealand.

Under least favourable conditions, ie, low pH and low organic content, malathion may persist in water with a half life of months or even years. However, under most conditions, the half life appears to be roughly 7�14 days.


In river water, the half life is less than one week, whereas malathion remained stable in distilled water for three weeks. Applied at 1 to 6 pounds/acre in ponds for mosquito control, it was effective for 2.5 to six weeks. The breakdown products in water are mono- and di-carboxylic acids.

Typical concentrations in drinking-water Malathion has been detected in surface water and drinking-water at concentrations below 0.002 mg/L (WHO 2004).


Analytical methods

Referee method

[???]


[???]

Health considerations Malathion is absorbed rapidly and effectively by practically all routes including the gastrointestinal tract, skin, mucous membranes, and lungs. Malathion inhibited cholinesterase activity in mice, rats and human volunteers. It increased the incidence of liver adenomas in mice when administered in the diet. Most of the evidence indicates that malathion is not genotoxic, although some studies indicate that it can produce chromosomal aberrations and sister chromatid exchange in vitro. JMPR has concluded that malathion is not genotoxic.

Derivation of maximum acceptable value The MAV for malathion in drinking-water was derived as follows:

29 mg/kg body weight/day x 70 kg x 0.1 = 1.015 mg/L (rounded to 1 mg/L) 2 L x 100

where:

• no observable adverse effect level = 29 mg/kg body weight per day, based on a two-year study of toxicity and carcinogenicity in rats, and supported by a NOAEL of 25 mg/kg of body weight per day in a developmental toxicity study in rabbits





• uncertainty factor = 100. However, intake of malathion from all sources is generally low and well below the ADI. As the chemical occurs in drinking-water at concentrations much lower than the health-based value, the presence of malathion in drinking-water under usual conditions is unlikely to represent a hazard to human health. For this reason, WHO (2004) considered it unnecessary to derive a guideline value for malathion in drinking-water.




WHO. 2003. Malathion in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/103).


MCPA Revised July 2005. (Also called 4-chloro-2-methylphenoxyacetic acid.)

Maximum acceptable value Based on health considerations, the concentration of MCPA in drinking-water should not exceed 0.002 mg/L (2 µg/L).


1 To source waters

MCPA may enter source waters as a result of its use as a systemic hormone-type selective herbicide which is readily absorbed by leaves and roots. It is used for the post-emergence control of annual and perennial weeds in cereals, grassland and turf. The total annual usage of MCPA in New Zealand in the late 1980s was 276,000 kg which was evenly split between the North and South Islands. As at August 2005, MCPA is registered for use in New Zealand.

Forms and fate in the environment MCPA can be expected to leach readily in most soils. It biodegrades with half lives ranging from seven to 60 days. The recommended average half life in soil is 25 days. It does not volatilise from surface waters, but undergoes photolysis and biodegradation with half lives between 10 and 24 days. The water solubility of MCPA is 825 mg/L for the acid with lower solubility (5 mg/L) for the ester and much higher solubility (270,000 to 900,000 mg/L) for the salts.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 296 zones, did not find detectable concentrations of MCPA (limit of detection = 0.0001 mg/L).

There are only limited and inconclusive data on the genotoxicity of MCPA. Data on humans and animals have been considered inadequate for an evaluation of carcinogenicity. The International Agency for Research on Cancer has evaluated chlorophenoxy herbicides and concluded that evidence of the carcinogenicity of chlorophenoxy herbicides is limited in humans and inadequate in animals (Group 2B). No adequate epidemiological data on exposure to MCPA alone are available. Recent carcinogenicity studies on rats and mice do not indicate that MCPA is carcinogenic.

Not frequently detected in drinking-water; has been measured in surface water and groundwater at concentrations below 0.00054 and 0.0055 mg/L, respectively (WHO 2004).

Removal methods After 50 to 80% oxidation of MCPA (500 mg/L) by ozone, no degradation products are detectable, indicating that ozonation may provide a practical means of removing this pesticide.


Referee method

High Performance Liquid Chromatography with a Photoiodide Array Ultraviolet Detector (EPA 555).



Health considerations Animal studies have shown that MCPA is absorbed readily from the gut of mice and detected in all organs tested. It is metabolised by the liver. In humans, 5% of the total dose was detected in the urine within 48 hours. Epidemiological investigations on MCPA have involved chlorophenoxyacetic weed killer producers and users, so exposure to this product generally is accompanied by exposure to 2,4-D, 2,4,5-T, mecoprop and dichlorprop. The International Agency for Research on Cancer carried out a comprehensive evaluation related to �professional exposure to chlorophenoxyacetic weed killers� which were considered to show �limited evidence� of carcinogenicity.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of a MAV for MCPA. The no observable adverse effect level used in the derivation is based on liver and kidney toxicity observed in a one-year feeding study in dogs. The MAV for MCPA in drinking-water was derived as follows:

0.15 mg/kg body weight/day x 70 kg x 0.1 = 0.002 mg/L (2 µg/L)

where:

2 L/day x 300

• no observable adverse effect level = 0.15 mg/kg body weight per day based on liver and kidney toxicity observed in a 1-year feeding study in dogs





• uncertainty factor = 300; 100 for inter and intra-species variation and 3 for the inadequacy of the data base.








WHO. 2003. MCPA in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality . Geneva: World Health Organization (WHO/SDE/WSH/03.04/38).


MCPB Revised July 2005. (Also called 4(2-methyl-4-chlorophenoxy)butyric acid.)

Maximum acceptable value (provisional) Based on health considerations, the concentration of MCPB in drinking-water should not exceed 0.03 mg/L (30 µg/L). WHO (2004) does not have a guideline value because MCPB is considered unlikely to occur in drinking-water.


1 To source waters

MCPB or 4-(2-methyl-4-chlorophenoxy)butyric acid, may enter source waters as a result of its use as a systemic hormone-type selective chlorophenoxy herbicide. It is absorbed readily by leaves and roots and is used for the control of annual and perennial weeds in cereals, grassland and turf. The total annual usage of MCPB in New Zealand in the late 1980s was 35,500 kg, with the majority of this being in the South Island. As at August 2005, MCPB is registered for use in New Zealand.


Forms and fate in the environment MCPB degrades in soil to form MCPA and 4-chloro-2-methylphenol. Its half life in soil ranges from six to 14 days unless the soil micro-organisms are acclimatised to the herbicide, in which case its half life can be less than one day. The water solubility of MCPB is 44 mg/L for the acid with lower solubility for the ester and much higher solubility for the salts.

Typical concentrations in drinking-water No data are available on the concentration of MCPB in New Zealand drinking-water supplies. It was undetected in 447 Canadian surfaces waters, but was found in Dutch surface waters at maximum concentrations ranging from 0.001 to 0.01 mg/L (1 to 10 µg/L).

Removal methods No specific information on methods of removing MCPB from water is available. Slow sand filtration can partially remove members of the chlorophenoxy acid pesticide family of which MCPB is a member. Ozone has shown varying degrees of effectiveness in oxidising the chlorophenoxy acids.


Referee method

[???]


[???] However, the following information may be useful:

1 Chlorophenoxy herbicides, including MCPB may be determined by a liquid-liquid extraction, chemical derivatisation, and analysis by gas chromatography with electron capture, electrolytic conductivity or mass spectrometry detection (Que Hee and Sutherland).

2 Alternatively, high-performance liquid chromatography with ultraviolet detection may be used for the quantitation. Detection limits range from 0.001 mg/L to 1 mg/L depending on the method of analysis. Interference may occur from impurities in the reagents or glassware used for extraction.

Health considerations In general, chlorophenoxy herbicides are absorbed rapidly from the gastro-intestinal tract and evenly distributed throughout the body. Accumulation in human tissues is not expected and a steady-state level in the human body will be achieved within 3�5 days of exposure. Elimination occurs primarily in the urine, mostly in the unchanged form. Biological half lives of chlorophenoxy herbicides in mammals range from 10 to 33 hours. Metabolic conversions occur only at high doses. The salt and ester forms are hydrolysed rapidly and follow the same pharmacokinetic pathways as the free acid forms. Chlorophenoxy herbicides as a group, including 2,4-D and MCPA, have been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). However, based on the available data from studies on exposed populations and on animals, it is not possible to assess the carcinogenic potential of any specific chlorophenoxy herbicide. Therefore, drinking-water guidelines for these compounds are based on a threshold approach for other toxic effects.

Derivation of maximum acceptable value The MAV for MCPB in drinking-water was derived as follows:


0.01 mg/kg body weight/day x 70 kg x 0.1 = 0.03 mg/L

2 L/day

An oral reference dose (RfD) of 0.01 mg/kg-day was set by the EPA for MCPB, based on a subchronic feeding study in which male beagles suffered from reproductive toxicity and other effects. In deriving the RfD, an uncertainty factor of 1000 was applied to account for inter and intra-species differences, as well as conversion from a subchronic to a chronic dose.12 [What does �12� relate to?]

References Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466�[Add page number].


USEPA. 2005. Integrated Risk Information System. US Environmental Protection Agency, the Office of Research and Development, and the National Center for Environmental Assessment.




World Health Organization. 1996. Guidelines for Drinking-water Quality (2nd ed). Volume 2: Health criteria and other supporting information.


Mecoprop Revised July 2005. (Also called MCPP.)

Maximum acceptable value Based on health considerations, the concentration of mecoprop in drinking-water should not exceed 0.01 mg/L.


1 To source waters

Mecoprop may enter source waters as a result of its use for the post-emergent control of broadleaved weeds in cereals, grassland and under fruit trees and vines, and the control of dock in pasture. The total annual usage of mecoprop in New Zealand in the late 1980s was 115,000 kg, mostly in the South Island. The highest usage was in the Ashburton County (33,000 kg). As at August 2005, mecoprop is registered for use in New Zealand.

Forms and fate in the environment Mecoprop is degraded in soil to 4-chloro-2-methylphenol with half lives of seven to 21 days. It is unlikely to leach to groundwater. Its water solubility is 620 mg/L for the acid and much higher (500,000 to 920,000 mg/L) for the alkali and amine salts.


Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 296 zones, did not find detectable concentrations of mecoprop (limit of detection = 0.0001 mg/L). Chlorophenoxy herbicides are not frequently found in drinking-water, but when detected, concentrations are usually no greater than a few micrograms per litre (WHO 2004).

Removal methods No specific information on methods of removing mecoprop from water is available. However, slow sand filtration can partially remove members of the chlorophenoxy acid pesticide family of which mecoprop is a member. Ozone has shown varying degrees of effectiveness in oxidising the chlorophenoxy acids.


Referee method

Liquid/Solid Extraction/Gas Chromatography with Electron Capture Detector (EPA 525.2).



2 Liquid/Solid Extraction/Gas Chromatography with Electron Capture Detector (EPA 552.1).

Health considerations In general, chlorophenoxy herbicides are absorbed rapidly from the gastro-intestinal tract and distributed evenly throughout the body. Accumulation in human tissues is not expected and a steady-state level in the human body will be achieved within 3�5 days of exposure. Elimination occurs primarily in the urine, mostly in the unchanged form. Biological half lives of chlorophenoxy herbicides in mammals range from 10 to 33 hours. Metabolic conversions occur only at high doses. The salt and ester forms are hydrolysed rapidly and follow the same pharmacokinetic pathways as the free acid forms. Mecoprop readily crosses the placental barrier. Short-term exposure studies on rats fed diets containing mecoprop reported blood and kidney effects. Long-term exposure studies in rats reported blood effects and increased relative kidney and liver weights. Symptoms described in case histories of humans suffering from acute poisonings by weedkiller solutions containing mecoprop include coma, fever, respiratory problems, myotonia (muscle stiffness), muscle cramps, skeletal muscle damage, electrocardiographic changes, decreased blood pressure, distended abdomen and rhabdomyolysis with renal failure. Chlorophenoxy herbicides as a group, including 2,4-D and MCPA, have been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). However, based on the available data from studies on exposed populations and on animals, it is not possible to assess the carcinogenic potential of any specific chlorophenoxy herbicide. Therefore, drinking-water guidelines for these compounds are based on a threshold approach for other toxic effects.

Derivation of maximum acceptable value Because it is not possible to assess the carcinogenic potential to humans of any specific chlorophenoxy herbicide, a tolerable daily intake approach has been used for the derivation of the MAV. The no


observable adverse effects level used in the derivation is based on increased kidney weight in one- and two-year studies in rats. The MAV for mecoprop in drinking-water was derived as follows:


where:

• no observable adverse effect level = 1 mg/kg body weight per day based on increased kidney weight in one- and two-year studies in rats




• uncertainty factor = 300 (100 for inter and intra-species variation and 3 for limitations of the data base).


Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466�[add page number].







Metalaxyl New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of metalaxyl in drinking-water should not exceed 0.1 mg/L (100 µg/L). Metalaxyl is not mentioned in WHO 2004.


1 To source waters

Metalaxyl is an agricultural fungicide. It is registered for use in New Zealand, and is available in a variety of formulations, many of which include other active ingredients such as mancozeb, thiabendazole,


cymoxanil and fludioxonil. Trade names are: Apron Combi, Apron TZ, Max MZ, Phytospear, Ridomil Gold 2.5G, Ridomil Gold MZ, Speartek and Wakil XL. Metalaxyl-m, an isomer of metalaxyl is also registered for use as a fungicide in New Zealand. Trade names are Apron XL, Ridomil Gold EC, Ridomil WG. No information is available on the annual usage of specific active ingredients in New Zealand, although metalaxyl is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).


Metalaxyl is very soluble in water: 7.1 g/L (7100 mg/L) (Merck & Co 1996). Metalaxyl is moderately stable under normal environmental conditions. It is photolytically stable in water when exposed to sunlight, with a half life of 400 days. Monitoring data demonstrate that metalaxyl has the potential to reach groundwater (USEPA 1994). Based on international studies, people may be exposed to residues of metalaxyl through the diet, but chronic dietary risk is minimal (USEPA 1994). Application and post-application risks to workers and others also are minimal because metalaxyl has no toxicological endpoints of concern. There is no information available regarding the greatest source of exposure to metalaxyl for New Zealanders (ie, dermal contact, inhalation, diet: food, water).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included metalaxyl, and typical concentrations in New Zealand drinking-waters are unknown. Metalaxyl has been detected in groundwater (NB: not necessarily drinking-water supplies) in five US states at levels typically reaching up to 0.003 mg/L. In order to reduce the possibility of groundwater contamination, USEPA is requiring a groundwater label advisory for metalaxyl end-use products (USEPA 1994).

Removal methods Metalaxyl has been reported to be decomposed completely by UV light in the presence of a titanium dioxide catalyst in the laboratory (Topalov et al 1999). This system has not been implemented in full-scale water treatment.

Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of metalaxyl, although a guide to the efficiency of the process cannot be provided.


Referee method

None given for metalaxyl in Table 12.3 of the DWSNZ Do they include metalaxyl-m, an isomer of metalaxyl � and is it in the MAV? Not in APHA 1998



None given for metalaxyl in Table 12.3 of the DWSNZ

Metalaxyl generally is of low acute toxicity but is a moderate eye irritant and has been placed in USEPA toxicity class III (Extoxnet 1996).

In a subchronic feeding study using rats, reduced food consumption and liver cell effects were noted at the highest dose tested. In a chronic toxicity study using beagle dogs, blood serum enzyme effects and increased liver weights were noted in the highest dose group. A study using rats resulted in liver effects (USEPA 1994). The liver is the primary target organ in animal systems (Extoxnet 1996).


Health considerations Studies with rats and goats showed rapid metabolism and excretion via the urine and faeces. Metalaxyl is metabolised to a variety of products before excretion (Extoxnet 1996).

Acute poisoning

The acute oral LD50 for rats is 669 mg/kg (RSocC 1987), which suggests a moderate acute oral toxicity when compared with other pesticides.

Chronic exposure

Although people may be exposed to residues of metalaxyl in many food commodities, USEPA describes the chronic dietary risk from all uses as minimal. Application and post-application risks to workers and others also are minimal because metalaxyl has no toxicological endpoints of concern.

The International Agency for Research on Cancer (IARC) has not classified metalaxyl, but the USEPA has classified it as a Group E carcinogen; that is, a chemical that does not show evidence of carcinogenicity for humans (USEPA 1994).

A tolerable daily intake approach has been used for the derivation of the MAV for metalaxyl in drinking-water, as follows:


Does this include metalaxyl-m?

where: • no observable adverse effect level = 3 mg /kg body weight per day • average weight of adult = 70 kg

Extoxnet. 1996. Pesticide Information Profile: Metalaxyl. Oregon State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

• average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10% • uncertainty factor = 100.

References



Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Metalaxyl (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

Topalov A, Molnar-Gabor D, Csanadi J. 1999. Photocatalytic oxidation of the fungicide metalaxyl dissolved in water over TiO2. Wat Res 33: 1371�6.

USEPA. 1994. EPA RED Facts: Metalaxyl. Prevention, Pesticides and Toxic Substances (7508W). EPA-738-F-94-013.

Methamidophos New entry August 2005.

Removal methods

Maximum acceptable value There are insufficient data to determine a MAV for methamidophos in drinking-water. WHO 2004 states that methamidophos is unlikely to occur in drinking-water so a guideline value is unnecessary.


1 To source waters

Methamidophos is a highly active, systemic, residual organophosphate insecticide/acaricide/avicide, used commonly on vegetables and fruit. As at August 2005, methamidophos is registered for use in New Zealand. Methamidophos is also a breakdown product of the organophosphate insecticide acephate (orthene).

Forms and fate in the environment In aerobic soils, the half-life of methamidophos is as follows: 1.9 days in silt, 4.8 days in loam, 6.1 days in sand, and 10-12 days in sandy loam. The half life of the chemical in water is 309 days at pH 5.0, 27 days at pH 7.0, and 3 days at pH 9.0. The chemical will break down in the presence of sunlight, and has a half life of 90 days in water at pH 5 when there is sunlight


No information available.


Referee method

A referee method cannot be selected for methamidophos because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for methamidophos for the above reason. However, the following information may be useful: [???]


Health considerations Methamidophos is a potent acetylcholinesterase inhibitor.

Extoxnet. Pesticide Information Profile. New York: PMEP, Cornell University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

Methomyl


Methamidophos is absorbed rapidly through the stomach, lungs and skin. It is eliminated primarily in the urine. Methamidophos is highly toxic via oral, dermal and inhalation routes of exposure. The oral doses of methamidophos that resulted in the mortality of half of the test organisms (LD50 values) are 21 and 16 mg/kg body weight for male and female rats respectively, 30�50 mg/kg body weight in guinea pigs, and 10�30 mg/kg body weight in rabbits. The primary target of organophosphate compounds is the nervous system. Some liver damage has been observed in rabbits. Reduced sperm count and reduced sperm viability have been observed in humans. Methamidophos has tested positive for genotoxicity, or ability to induce changes in chromosomes, in some tests and negative in others. It may be weakly mutagenic. There is no evidence of carcinogenicity in tests with rats and mice.

Derivation of maximum acceptable value There are limited and insufficient data on methamidophos on which to propose a MAV for drinking-water.

References


New entry August 2005.

Maximum acceptable value There are insufficient data to determine a MAV for methomyl in drinking-water. WHO 2004 states that methomyl is unlikely to occur in drinking-water so a guideline value is unnecessary.

1 To source waters

Methomyl, a carbamate, was introduced in 1966 as a broad spectrum insecticide. It is also used as an acaricide to control ticks and spiders. It is used for foliar treatment of vegetable, fruit and field crops, cotton, commercial ornamentals, and in and around poultry houses and dairies. It is also used as a fly bait. As at August 2005, methomyl is registered for use in New Zealand. Because of its high solubility in water (58,000 mg/L) and its soil half-life (33 days), methomyl may have potential for groundwater contamination.


Forms and fate in the environment It is very mobile in sandy loam and silty clay loam soils, but only slight leaching was observed in a silt loam and in a sandy soil. Adsorption of methomyl to soil particles is weak to moderate. Methomyl is degraded rapidly by soil microbes. The dissipation half life for methomyl in soil is reportedly three to six weeks. However, one month after methomyl treatment, test soil had traces of the insecticide and some of its breakdown byproducts, or metabolites. Methomyl residues are not expected to be found in treated soil after the growing season in which it is applied. Under aerobic conditions, methomyl has a soil half life of 30�45 days and degrades predominately to carbon dioxide. It is relatively stable to hydrolysis under neutral and acidic conditions. Under basic conditions, it degrades with a half life of 30 days. Under anaerobic conditions, acetonitrile is the major metabolite in the early stages of degradation, but carbon dioxide is the end product, with total conversion within eight days. Aqueous solutions of methomyl have been reported to decompose more rapidly on aeration, in sunlight, or in alkaline media. One study indicated a half life of six days for the insecticide in water. Its hydrolysis half life in soil, or the time that it takes for half of it to be broken down in groundwater, is estimated at over 25 weeks. In one experiment the hydrolysis half lives of methomyl in solutions at pHs of 6.0, 7.0 and 8.0 were 54, 38, and 20 weeks respectively. In pure water, the hydrolysis half life has been estimated to be 262 days. Methomyl is unlikely to bioconcentrate in aquatic systems.

Typical concentrations in drinking-water Methomyl has been detected at very low levels, 9 ppb and 1.2 ppb respectively, in groundwater in New York and New Jersey.

Removal methods



Referee method

A referee method cannot be selected for methomyl because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for methomyl for the above reason. However, the following information may be useful: [???]

Health considerations Methomyl is a highly toxic inhibitor of cholinesterase, an essential nervous system enzyme. Carbamates, the class of active ingredients in which methomyl is included, are absorbed quickly from the skin, lungs and gastrointestinal tract and are broken down and transformed in the liver. Although they do not appear to accumulate in any particular body tissue, they do alter many other enzyme systems besides the cholinesterases. Carbamates generally are excreted rapidly and do not accumulate in mammalian tissue. The LD50 for methomyl in rats is 12�48 mg/kg, in mice it is 10 mg/kg, and in guinea pigs it is 15 mg/kg.


Based on health considerations, the concentration of methoxychlor in drinking-water should not exceed 0.02 mg/L. The maximum contaminant level (USEPA 2004) is 0.04 mg/L.

Methoxychlor may be biodegraded anaerobically with half lives ranging from one week to two months, or aerobically with half lives of greater than three months. The recommended average half life in soil is four months. The main route of disappearance from the water phase is volatilisation with a half life in water of about 46 days.

In a 24 month study with rats fed doses of 0, 2.5, 5 or 20 mg/kg, the NOEL was 20 mg/kg. At 20 mg/kg, red blood cell counts and haemoglobin levels were reduced significantly in female rats. Based on a 5 mg/kg NOEL in a two-year feeding study with dogs, and utilising a 100-fold safety margin, the EPA has established an ADI (Acceptable Daily Intake) for methomyl of 0.025 mg/kg of body weight/day. Acetamide, a suspected oncogen, is a minor metabolite of methomyl.

Derivation of maximum acceptable value There are limited and insufficient data on methomyl on which to propose a MAV in drinking-water. The USEPA has established a lifetime health advisory level of 200 ppb (0.2 mg/L) for methomyl. Water containing methomyl at or below this level is acceptable for drinking every day over the course of one�s lifetime and does not pose any health risk.



Methoxychlor Revised July 2005.



1 To source waters

Methoxychlor may enter source waters as a result of its application as an insecticide used to protect vegetables, fruit trees, fodder cereals, and animals against a variety of pests. Methoxychlor is not currently registered for use in New Zealand.


The water solubility of methoxychlor is 0.1 mg/L and the sorption coefficient is 80,000 mL/g. It is unlikely to leach to groundwater. There is some potential for the accumulation of the parent compound and its metabolites in surface water sediments.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of methoxychlor (limit of detection = 0.0002 mg/L).


Detected occasionally in drinking-water, at concentrations as high as 0.3 mg/L in rural areas (WHO 2004).

Removal methods No information on methods of removing methoxychlor from water is available. However, isotherm adsorption data indicate that removal by adsorption on to granular activated carbon should be possible; 0.1 mg/L should be achievable using GAC (WHO 2004).


Referee method






Health considerations Although methoxychlor is absorbed from the gastrointestinal tract, it does not accumulate in mammalian tissue. Body stores built up under continuous exposure are cleared within a few weeks after cessation of exposure. It is metabolised to formaldehyde and phenolic metabolites and excreted in faeces. The main effects of single high exposures to methoxychlor are disturbances of glycogen metabolism and fatty degeneration of the organs. The main effect observed after long-term exposure tests on animals was growth retardation. In humans, a single dose of 2 mg/kg body weight was without effect on liver, testicles and small intestine. Doses of up to 2 mg/kg body weight per day administered orally to men and women over a period of 4�6 weeks and 6�8 weeks were without effect on body weight and several biochemical parameters. Tissue damage did not occur. The menstrual cycle and the volume of ejaculation were not affected, although a shortening of the neck of the spermatozoa was observed in the first study. The genotoxic potential of methoxychlor appears to be negligible. The International Agency for Research on Cancer placed methoxychlor in Group 3 in 1979. Although subsequent data suggest a carcinogenic potential for liver and testis in mice, which may be caused by the hormonal activity of proestrogenic metabolites of methoxychlor and thus may have a threshold, the study was inadequate because only one dose was used and because this may have been above the maximum tolerated dose. The genotoxic potential of methoxychlor appears to be negligible. Methoxychlor may be a tumour promoter.

Derivation of maximum acceptable value As the genotoxic potential of methoxychlor appears to be negligible, a tolerable daily intake approach was used for the derivation of the MAV for methoxychlor in drinking-water. The data base for studies on long-term, short-term and reproductive toxicity of methoxychlor is inadequate. Therefore the no observable adverse effect level used for the derivation of the MAV is based on a teratology study in rabbits.


The MAV for methoxychlor in drinking-water was derived as follows:


where:

• no observable adverse effect level = 5 mg/kg body weight per day based a teratology study in rabbits




• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 for concern for threshold carcinogenicity and the limited database).

References

Methyl parathion







WHO. 2003. Methoxychlor in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/105).


New entry August 2005. (Also called parathion-methyl and and metafos.)

Maximum acceptable value There are insufficient data to determine a MAV for methyl parathion in drinking-water. WHO 2004 states that methyl parathion occurs in drinking-water at concentrations well below those at which toxic effects are observed.


1 To source waters

Methyl parathion is an organophosphate insecticide and acaricide, used to control many biting or sucking insect pests of agricultural. 4-Nitrophenol is an insecticide and a break-down product of methyl parathion which does not adsorb to soil particles so may contaminate groundwater. As at August 2005, parathion-methyl is registered for use in New Zealand.


Form and fate in the environment Methyl parathion partitions mainly to air and soil in the environment. There is virtually no movement through soil. By far the most important route for the environmental degradation of methyl parathion is microbial degradation. In most situations, methyl parathion adsorbs to soil particles and degrades rapidly; it is therefore unlikely to contaminate groundwater. Half lives of methyl parathion in water are in the order of weeks to months. Concentrations of methyl parathion in natural waters of agricultural areas in the USA ranged up to 0.0005 mg/L, with highest levels in summer. In water, methyl parathion is subject to photolysis, with a half-life of eight days during the summer and 38 days in winter. When it is applied as an insecticide, methyl parathion breaks down within several months, primarily by photolysis and biodegradation. The rate of degradation increases with temperature and with exposure to sunlight. Its biodegradation half life in soil is 10 days to two months. Degradation was faster in flooded soils than in non-flooded soils. Mineralisation may occur, especially in moist soils. Some volatilisation of applied methyl parathion may occur.

Typical concentrations in drinking-water Methyl parathion has been detected rarely in groundwater outside of areas where it is used. It has been detected in the groundwater of Mississippi at 8 ppb (0.008 mg/L).



Referee method

A referee method cannot be selected for methomyl because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for methomyl for the above reason. However, the following information may be useful: [???]

Health considerations The general population can come into contact with methyl parathion via air, water or food. The organophosphate insecticides are cholinesterase inhibitors. They are highly toxic by all routes of exposure. The oral LD50 for methyl parathion in rats is 18 to 50 mg/kg, in mice is 14.5 to 19.5 mg/kg, in rabbits is 420 mg/kg, in guinea pigs is 1270 mg/kg, and in dogs is 90 mg/kg. Studies with human volunteers found that 1 to 22 mg/person/day have no effect on cholinesterase activity. In a 4-week study of volunteers given 22, 24, 26, 28 or 30 mg/person/day, mild cholinesterase inhibition appeared in some individuals in the 24, 26 and 28 mg dosage groups. In the 30 mg/person/day (about 0.43 mg/kg/day) group, red blood cholinesterase activity was depressed by 37%. When methyl parathion was fed to dogs for 12 weeks, a dietary level of 1.25 mg/kg soon caused a significant depression of red blood cell and plasma cholinesterase. A dietary level of 0.125 mg/kg produced no effects.


Methyl parathion is absorbed rapidly into the bloodstream through all normal routes of exposure. Following administration of a single oral dose, the highest concentration of methyl parathion in body tissues occurred within one to two hours. Metabolism occurs in the liver, eventually to phenols which can be detected in the urine. Methyl parathion does not accumulate in the body. It is almost completely excreted through the kidneys (urine) within 24 hours. Methyl parathion is a possible human teratogen and is not a suspected carcinogen.

Derivation of maximum acceptable value A NOAEL of 0.3 mg/kg of body weight per day was derived from the combined results of several studies conducted in humans, based on the depression of erythrocyte and plasma cholinesterase activities. Methyl parathion decreased cholinesterase activities in long-term studies in mice and rats, but did not induce carcinogenic effects. Methyl parathion was mutagenic in bacteria, but there was no evidence of genotoxicity in a limited range of studies in mammalian systems. A health-based value of 9 mg/L can be calculated for methyl parathion on the basis of an ADI of 0.003 mg/kg of body weight, based on a NOAEL of 0.25 mg/kg of body weight per day in a two-year study in rats for retinal degeneration, sciatic nerve demyelination, reduced body weight, anaemia and decreased brain acetylcholinesterase activity, using an uncertainty factor of 100. Since the toxicological end-points seen in animals were other than acetylcholinesterase inhibition, it was considered more appropriate to use these data rather than the NOAEL derived for cholinesterase inhibition in humans. Intake of methyl parathion from all sources is generally low and well below the ADI. As the health-based value is much higher than methyl parathion concentrations likely to be found in drinking-water, the presence of methyl parathion in drinking-water under usual conditions is unlikely to represent a hazard to human health. For this reason, the establishment of a guideline value for methyl parathion is not deemed necessary. The USEPA (1990) has established a Lifetime Health Advisory (LHA) level of 0.06 mg/L for 4-nitrophenol, a breakdown product of methyl parathion, in drinking water. This means that EPA believes that water containing 4-nitrophenol at or below this level is acceptable for drinking every day over the course of one�s lifetime, and does not pose any health concerns. However, consumption of 4-nitrophenol at high levels well above the LHA level over a long period of time has been shown to cause adverse health effects, including damage to the liver, respiratory stress, and inflammation of the stomach in animal studies.



IPCS. 1992. Methyl Parathion. Geneva: World Health Organization. International Programme on Chemical Safety (Environmental Health Criteria 145).

USEPA. 1990. National Pesticide Survey: 4-nitrophenol. Washington DC: Environmental Protection Agency, Office of Water, Office of Pesticides and Toxic Substances.

WHO. 2003. Methyl Parathion in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/106).


Metolachlor Revised July 2005.


Maximum acceptable value Based on health considerations, the concentration of metolachlor in drinking-water should not exceed 0.01 mg/L.


1 To source waters

Metolachlor may enter source waters as a result of its use as a selective herbicide for pre-emergence and incorporated pre-plant weed control in a variety of crops. The total annual usage of metolachlor in New Zealand in the late 1980s was 19,000 kg with the majority of use in the North Island. As at August 2005, metolachlor is registered for use in New Zealand.

Forms and fate in the environment Metolachlor photodegrades slowly in water exposed to sunlight. Its hydrolysis half life is greater than 200 days in water. It can also be lost from the soil through biodegradation and volatilisation. The half life for metolachlor in soil ranges from 15 to 130 days with the recommended average soil half life being 90 days. Metolachlor can leach beyond the root zone in detectable amounts. The water solubility of metolachlor is 530 mg/L and the sorption coefficient is 200 mL/g. It is fairly mobile and under certain conditions can contaminate groundwater, but it is mostly found in surface water.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of metolachlor (limit of detection = 0.0001 mg/L). Detected in surface water and groundwater at concentrations that can exceed 0.01 mg/L (WHO 2004).

Removal methods Adsorption on to activated carbon appears to be the most promising technique for removal of metolachlor from drinking-waters, but more actual data are still required to determine its effectiveness.


Referee method

Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector (EPA 507).


No alternative methods have been recommended for metolachlor because no methods meet the required criteria.

Health considerations Metolachlor is absorbed and excreted readily in the rat. It is metabolised via dechlorination, O-methylation, N-dealkylation, and side-chain oxidation. No unchanged chemical was isolated. Metolachlor has a low oral acute toxicity. Signs of metolachlor intoxication in humans include abdominal cramps, anaemia, ataxia (loss of coordination), dark urine, methaemoglobinaemia, cyanosis,


hypothermia, collapse, convulsions, diarrhoea jaundice, weakness, nausea, shock, sweating, vomiting, central nervous system depression, dizziness, dyspenea, liver damage, nephritis, cardiovascular failure, dermatitis, sensitisation, eye and mucous membrane irritation, corneal opacity and reproductive effects. In a one-year study in beagle dogs, administration of metolachlor resulted in decreased kidney weight at the two highest dose levels. In two-year studies with rodents fed metolachlor in the diet, the only toxicological effects observed in albino mice were decreased body weight gain and decreased survival in females at the highest dose level, whereas rats showed decreased body weight gain and food consumption at the highest dose level. Metolachlor does not induce gene mutations in bacterial or mammalian cells. There is no evidence from available studies that metolachlor is carcinogenic in mice. In rats, an increase in liver tumours in females and a few nasal tumours in males have been observed.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV of metolachlor in drinking-water. Long-term toxicity data are available in dogs and rodents. The lowest observable adverse effect level used in the derivation is based on an apparent decrease in kidney weight at the two highest dose levels in a one-year dog study. The MAV for metolachlor in drinking-water was derived as follows:


where:

• lowest observable adverse effect level = 3.5 mg/kg body weight per day based on an apparent decrease in kidney weight observed in a chronic dog study




• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 for concern regarding carcinogenicity).




USEPA. 1989. Drinking Water Health Advisory: Pesticides. United States Environmental Protection Agency, Office of Drinking-water Health Advisors. Lewis Publishers.





WHO. 2003. Metolachlor in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/39).


Metribuzin New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of metribuzin in drinking-water should not exceed 0.07 mg/L (70 µg/L). Metribuzin is not mentioned in WHO 2004.


1 To source waters

Metribuzin is a triazine herbicide, registered for use in New Zealand. It is available in four different formulations and is used for pre- and post-emergent weed control. Trade names are Lexone DF, Metriphar 48SC, Python and Sencor DF. It is applied by various methods including aerial, chemigation, and ground application. No information is available on the annual usage of specific active ingredients in New Zealand, although metribuzin is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Metribuzin is very soluble in water: 1200 mg/L (Merck & Co 1996). Microbial degradation is the principal route of removal of metribuzin from the soil (Health Canada 1986). Metribuzin also adsorbs moderately to soil with high clay and/or organic matter content; adsorption decreases as soil pH increases (WSSA 1983, cited in Health Canada 1986). It has a mobility (as Koc) of 60, which indicates that it is weakly adsorbed to organic soil and therefore has the potential to migrate through soil to reach groundwater. Once in groundwater, metribuzin is expected to persist due to its stability to hydrolysis and the lack of light penetration. Conversely, residues of metribuzin are not likely to persist in clear, well-mixed, shallow surface water with good light penetration since parent metribuzin degrades rapidly by aqueous photolysis (USEPA 1998). There is no information available regarding the greatest source of exposure to metribuzin for New Zealanders (ie, dermal contact, inhalation, diet: food, water). The USEPA indicates that dietary exposure to metribuzin residues in food are not of concern. Of greater concern is the inhalation exposure risk posed to metribuzin handlers, particularly mixers/loaders/applicators and field workers.

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included metribuzin. In the New Zealand national pesticides surveys, conducted every four years since 1990, metribuzin has been detected in groundwater twice, at concentrations of 0.00014 and 0.0012 mg/L (Close et al 2001). In the Waikato region, metribuzin has been detected in groundwater at three sites. Concentrations ranged from 0.00002�0.00028 mg/L (Hadfield and Smith 1999). Metribuzin has been detected in groundwater at one location in the Edendale area at concentrations ranging between 0.00003�0.00014 mg/L (Hughes 2000).


Metribuzin was detected in 26 of 1140 samples from municipal and private water supplies in Prince Edward Island (time period not reported) Nova Scotia (1986), Ontario (1979 to 1986), Manitoba (1986) and Alberta (1978 to 1986). Detection limits ranged from 0.00001 to 0.001 mg/L. The maximum concentration of metribuzin, determined in a sample from a well in Ontario, was 0.3 mg/L (Hiebsch 1988, cited in Health Canada 1986).

Removal methods Oxidation of triazines by ozone is reported to be effective (Chiron et al 2000). The water chemistry, in particular the alkalinity and pH, will affect the oxidation rate. Use of activated carbon following ozonization should be considered to adsorb oxidation products. Nanofiltration (membrane technology) in water with a low natural organic matter concentration is reported to remove approximately 50% of atrazine and simazine (Agbekodo, et al.1996). The percentage is increased to 90�100% when 3.6 mg/L of natural organic matter is present. Similar results may be expected for metribuzin as it is from the same chemical family and of comparable size. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of metribuzin, although a guide to the efficiency of the process cannot be provided.


Referee method



Liquid/solid extraction/gas chromatography-mass spectrometer (EPA 525.2) or liquid/solid extraction/gas chromatography-mass spectrometer (EPA 508.1).

Health considerations Rats administered 1 or 200 mg/kg body weight of radioactively labelled metribuzin by stomach tube were reported to eliminate about 80% in the first day following administration, and 95% by the second day. Almost equal amounts were found in the urine and faeces (Health Canada 1986).

Acute poisoning

In studies using laboratory animals, metribuzin generally has been shown to be of low acute toxicity. It is slightly toxic by the oral and inhalation routes and has been placed in Toxicity Category III (the second lowest of four categories) for this effect (USEPA 1998). The acute oral LD50 for rats is 2200 mg/kg, mice 698�711 mg/kg, guinea pigs 250 mg/kg (RSocC 1987). These levels suggest a moderate acute oral toxicity when compared with other pesticides. Effects of high acute exposure in metribuzin poisoned rats included narcosis (stupor) and laboured breathing. Deaths occurred within 24 hours, and survivors recovered slowly without permanent effects (Extoxnet 1996).

Chronic exposure

In two-year feeding studies with rats and dogs, results showed no observable effects at doses of 5 mg/kg/day in rats and 2.5 mg/kg/day in dogs. Reduced weight gain, an increase in the number of deaths, blood chemistry changes, and liver and kidney damage were observed in a two-year study in


which dogs were given 1500 ppm or 37.5 mg/kg/day of metribuzin (Extoxnet 1996). Results of three developmental toxicity studies and one reproduction study suggest that although metribuzin is not considered a developmental toxicant, it is associated with developmental toxicity effects (USEPA 1998). In single high dose studies, metribuzin appears to depress the central nervous system. Other studies indicate that the target organs of metribuzin are the thyroid gland and the liver (Extoxnet 1996). The International Agency for Research on Cancer (IARC) has not classified metribuzin, but the USEPA has classified it as a Group D chemical, not classifiable as to human carcinogenicity.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for metribuzin in drinking-water, as follows:





Close ME, Rosen MR, Smith VR. 2001. Fate and transport of nitrates and pesticides in New Zealand�s aquifers. In: Rosen MR, White PA (eds). Groundwaters of New Zealand, Chapter 8, pp. 185�220. New Zealand: Hydrological Society.

Extoxnet. 1996. Pesticide Information Profiles: Metribuzin. Oregon State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/



Heibsch S. 1988. The occurrence of 35 pesticides in Canadian drinking-water and surface water. Unpublished report prepared for the Environmental Health Directorate, Department of National Health and Welfare, Canada.

Hughes B. 2000. Edendale Pesticide Iinvestigation Report 2000. Southland Regional Council Publication No 2000�14.


Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Metribuzin (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

USEPA. 1998. EPA RED Facts: Metribuzin. Prevention, pesticides and toxic substances (7508W). EPA-738-F-96-006.


Weed Science Society of America. 1983. Herbicide Handbook (5th ed). Champaign, USA.


Molinate Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of molinate in drinking-water should not exceed 0.007 mg/L (7 µg/L).


1 To source waters

Molinate may enter source waters as a result of its use to control germinating broadleaved and grassy weeds in rice. Molinate has never been registered in New Zealand.

Forms and fate in the environment The available data suggest that groundwater pollution by molinate is restricted to some rice-growing regions. Volatilisation is the main route of loss of molinate from soil and rice fields. Photochemical degradation occurs to a lesser extent. Molinate has a low persistence in soil and water with half lives ranging from five to 25 days. The water solubility of molinate is 880 mg/L and the sorption coefficient is 190 mL/g.

Typical concentrations in drinking-water Molinate has never been used in New Zealand. The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of molinate (limit of detection = 0.0001 mg/L). Concentrations in water rarely exceed 0.001 mg/L (WHO 2004).

Removal methods No information on methods of removing molinate from water is available.


Referee method

Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector (EPA 507).


No alternative methods have been recommended for molinate because no methods meet the required criteria.


Health considerations Molinate is not absorbed percutaneously. In rats, it is metabolised primarily to the sulfoxide, then to mercapturic acid. Molinate has not shown evidence of mutagenic activity in a range of bacterial assays. Based on the limited data available, molinate does not seem to be carcinogenic to animals. Evidence suggests that impairment of the reproductive performance of the male rat is the most sensitive indicator of molinate exposure. However, a review of epidemiological data based on the examination of workers involved in molinate production does not indicate an effect on human fertility.

Derivation of maximum acceptable value A tolerable daily approach has been used for the derivation of the MAV for molinate in drinking-water. The no observable adverse effect level used in the derivation is for reproductive toxicity in the rat. The MAV for molinate in drinking-water was derived as follows:

0.2 mg/kg body weight/day x 70 kg x 0.1 = 0.007 mg/L (7 µg/L) 2 L/day x 100

where: • no observable adverse effect level = 0.2 mg/kg body weight per day for reproductive toxicity in the rat • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 0.1 • uncertainty factor = 100 (for inter and intra-species variation)

References USEPA. 1988. Methods for the Determination of Organic Compounds in Drinking Water.


WHO. 2003. Molinate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/40).


Oryzalin New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of oryzalin in drinking-water should not exceed 0.4 mg/L (400 µg/L). Oryzalin is not mentioned in WHO 2004.


1 To source waters

Oryzalin is a selective pre-emergence herbicide. It is registered for use in New Zealand and is available in two different formulations: granules and suspension concentrate. The granule formulation (trade name


Rout Ornamental Herbicide) also contains the active ingredient oxyfluorfen. The suspension formulation has a trade name of Surflan Flo. No information is available on the annual usage of specific active ingredients in New Zealand, although oryzalin is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Oryzalin has low solubility in water: 2.5 mg/L (Merck & Co 1996). Oryzalin biodegrades slowly with a half life of approximately two months. It is not mobile under field conditions and is not volatile. There is no information available regarding the greatest source of exposure to oryzalin for New Zealanders (ie, dermal contact, inhalation, diet: food, water).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included oryzalin, so typical concentrations in New Zealand drinking-waters are unknown. Oryzalin has been detected once during sampling of groundwaters in the Waikato region by Environment Waikato. The concentration was 0.00019 mg/L (Hadfield and Smith 1999). Removal methods No information is available on the removal of oryzalin from water. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of oryzalin, although a guide to the efficiency of the process cannot be provided. Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.


Referee method

None listed for Oryzalin in Table 12.3 of the DWSNZ Nor is it in APHA 1998


None listed for Oryzalin in Table 12.3 of the DWSNZ

Health considerations Oryzalin is moderately well-absorbed from the gastrointestinal tract, metabolized and eliminated following absorption. When oryzalin was administered to male rats, 40% of the dose was excreted in the urine and 40% in the faeces within three days. Similar results were obtained in tests with rabbits, a steer, and with Rhesus monkeys (Extoxnet 1996).


Chronic exposure

In acute toxicity studies using laboratory animals, oryzalin is practically non-toxic by the oral route and the USEPA has placed it in Toxicity Category IV (the lowest of four categories) for this effect. Oryzalin generally is of modest acute toxicity. The acute oral LD50 for rats and mice is greater than 10,000 mg/kg, for cats and dogs it is greater than 1000 mg/kg (RSocC 1987). These levels suggest a low acute oral toxicity when compared with other pesticides. In subchronic toxicity studies, oryzalin caused the accumulation of an iron-containing pigment in the kidneys of rats, an increase in the weights of several organs in mice, and blood, bone marrow and liver effects in beagle dogs. Another chronic toxicity study using beagle dogs showed effects to the blood, liver, kidneys and thyroid gland. In developmental toxicity studies using rats, oryzalin caused reduced maternal body weight gain as well as decreased foetal body weights, an increase in runts and bone development effects. In rabbits, it caused reduced maternal food consumption and weight gain, foetal effects and reduced litter size. Reproduction studies using rats showed increased liver and kidney weights, and decreased food consumption and body weight gain (USEPA 1994). The International Agency for Research on Cancer (IARC) has not classified oryzalin, but USEPA has classified it as a Group C carcinogen, that is, a possible human carcinogen for which there is limited animal evidence (USEPA 1994).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for oryzalin in drinking-water, as follows:



References Extoxnet. 1996. Pesticide Information Profile: Oryzalin. Oregon State University.


Merck & Co. 1996. The Merck Index. An encyclopedia of chemicals, drugs and biologicals (12th ed). Merck Research Laboratories Division of Merck & Co Inc.

Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Oryzalin (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brother Ltd.

USEPA. 1994. EPA RED Facts: Oryzalin. Prevention, pesticides and toxic substances (7508W). EPA-738-F-94-012.


Oxadiazon New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of oxadiazon in drinking-water should not exceed 0.2 mg/L (200 µg/L). Oxadiazon is not mentioned in WHO 2004.


1 To source waters

Oxadiazon is a selective pre-emergence herbicide. It is registered for use in New Zealand and is available as an emulsifiable concentrate (trade name: Foresite 380) and as granules (trade name: Ronstar SG). The granules formulation also contain simazine as an active ingredient. No information is available on the annual usage of specific active ingredients in New Zealand, although oxadiazon is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Oxadiazon has very low solubility in water: 0.7 mg/L (Merck & Co 1996). Its mobility (as Koc) is 3200, which indicates that it is strongly adsorbed to organic soil, and therefore unlikely to be highly mobile in organic soils. There is no information available regarding the greatest source of exposure to oxadiazon for New Zealanders (ie, dermal contact, inhalation, diet: food, water).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included oxadiazon, so typical concentrations in New Zealand drinking-waters are unknown. Monitoring for pesticides in groundwater in the Waikato region has detected oxadiazon at one location, at a concentration of 0.00021 mg/L (Hadfield and Smith 1999).

Removal methods No information is available on the removal of oxadiazon from water. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of oxadiazon, although a guide to the efficiency of the process cannot be provided. Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.



Referee method

None listed for Oxadiazon in Table 12.3 of the DWSNZ Nor is it in APHA 1998


None listed for Oxadiazon in Table 12.3 of the DWSNZ


Acute exposure

The acute oral LD50 for rats and mice is greater than 8,000 mg/kg (RSocC 1987), which suggests a low acute oral toxicity when compared with other pesticides. Oxadiazon may irritate slightly the mucous membranes of the mouth if swallowed.

Chronic exposure

The critical effect of chronic exposure is increased levels of serum proteins and increased liver weights, based on a two-year rat feeding study (USEPA 1987). The International Agency for Research on Cancer has not classified oxadiazon for its ability to cause cancer.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for oxadiazon in drinking-water, as follows:



References Hadfield J, Smith D. 1999. Pesticide Contamination of Groundwater in the Waikato Region. Environment Waikato Technical Report 1999/9.


Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Oxadiazon (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

USEPA. 1987. Integrated Risk Information System (IRIS): Oxadiazon.

Pendimethalin Revised July 2005.


Maximum acceptable value Based on health considerations, the concentration of pendimethalin in drinking-water should not exceed 0.02 mg/L.


1 To source waters

Pendimethalin may enter source waters as a result of its application as a selective herbicide, used in cereals, maize and vegetable crops. It is also used to control suckers in tobacco. The total annual usage of pendimethalin in New Zealand in the late 1980s was 3100 kg with the majority of this being in the North Island. As at August 2005, pendimethalin is registered for use in New Zealand.

Forms and fate in the environment Pendimethalin is a moderately persistent herbicide that can give rise to long lasting metabolites, mainly through photodegradation. Pendimethalin and metabolites bind tightly to soil particles and the leaching potential is negligible. It is lost through photodegradation, biodegradation and volatilization. Half lives in soil range from 30 to 450 days with a recommended average half life of 90 days. Little is known about its more polar degradation products. The water solubility is 0.3 mg/L and the sorption coefficient is 5000 mL/g.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of pendimethalin (limit of detection = 0.0002 mg/L). WHO (2004) states that pendimethalin has been found rarely in drinking-water in the limited studies available (detection limit 0.00001 mg/L).

Removal methods No information on methods of removing pendimethalin from water is available.


Referee method

No referee method has been given for pendimethalin because no method meets the required criteria.


No alternative methods have been recommended for pendimethalin because no methods meet the required criteria. However, the following information may be useful: Pendimethalin can be determined in water samples by extraction with methylene chloride and analysis by gas chromatography with a nitrogen phosphorus detector (eg, Method EPA 507). Confirmation by a second capillary column with different polarity is strongly recommended. No information on a limit of quantitation is available.

Health considerations Pendimethalin appears to be absorbed poorly and excreted rapidly. Following oral administration, about 95% is excreted within 24 hours, principally in the faeces. Maximum tissue concentrations were found in the liver and kidney.


Pendimethalin is of low acute toxicity. In a short-term dietary study in rats, a variety of indications of hepatotoxicity as well as increased kidney weights in males were observed at the highest dose level. In a long-term dietary study, some toxic effects (hyperglycaemia in the mouse and hepatotoxicity in the rat) were present even at the lowest dose level. Pendimethalin does not appear to have significant mutagenic activity. Long-term studies in mice and rats do not provide evidence of carcinogenicity. However, these studies have some important limitations.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for pendimethalin in drinking-water. The lowest observable adverse effect level used in the derivation is based on slight liver toxicity even at the lowest dose tested (5 mg/kg of body weight) in a long-term rat feeding study. The MAV for pendimethalin in drinking-water was derived as follows:


• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 for the use of a LOAEL instead of a NOAEL an for limitations in the database).


WHO. 2003. Pendimethalin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/41).

where:

• lowest observable adverse effect level = 5 mg/kg body weight per day for liver toxicity observed in a two-year rat study




References




Pentachlorophenol Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of pentachlorophenol in drinking-water should not exceed 0.009 mg/L. The guideline value is considered provisional because of the variationsin metabolism between experimental animals and humans. The maximum contaminant level (USEPA 2004) is 0.001 mg/L.



1 To source waters

Pentachlorophenol may enter source waters as the result of its use as a timber preservative. It has been used for this purpose during the past 30�40 years in New Zealand. Until mid 1988 up to 200 tonnes of pentachlorophenol was used annually, however the use of pentachlorophenol has now virtually ceased in response to environmental and occupational health concerns. The presence of dioxins and dibenzofuran impurities in pentachlorphenol has also contributed to the decline in its usage. Domestic use of pentachlorophenol in products such as moss killers which are available from supermarkets and hardware sources may also result in the contamination of source waters.

The P2 Chemical Determinand Identification Programme, sampled from 494 zones, did not find detectable concentrations of pentachlorophenol (limit of detection = 0.0001 mg/L).

1. Liquid/Solid Extraction and Gas Chromatography with an Electron Capture Detector (EPA 515.2).


No known sources.


No known sources.

Forms and fate in the environment Pentachlorophenol is very persistent in the environment, particularly if adsorbed to sediments. Leaching to groundwater has been observed at some timber treatment sites in New Zealand. Oxidation is the main mechanism of pentachlorophenol degradation, and degradation is favoured by light, high pH, and high dissolved oxygen conditions. Half lives of pentachlorophenol in soil range from seven to 120 days.

Typical concentrations in drinking-water Pentachlorophenol was not detected in any of 320 samples from 161 supplies in New Zealand sampled between 1988 and 1992. The detection limit was between 0.00002 and 0.00004 mg/L (0.02 and 0.04 µg/L).

Concentrations in water samples are usually below 0.01 mg/L, although much higher concentrations in groundwater may be measured under certain conditions (WHO 2004).

Removal methods No information on methods of removing pentachlorophenol from water is available. However, isotherm adsorption data indicate that removal by adsorption on to granular activated carbon should be possible.


Referee method



2. Acetylation Liquid/Liquid Extraction Gas Chromatographic/Mass Spectrometric Method (EPA 1653).


Health considerations Food is usually the major source of exposure to PCP unless there is a specific local chlorophenol contamination of drinking-water or exposure from log homes treated with PCP. In short- and long-term animal studies, exposure to relatively high pentachlorophenol concentrations has been shown to reduce growth rates and serum thyroid hormone levels and to increase liver weights and liver enzyme activity. Exposure to much lower concentrations of technical pentachlorophenol formulations has been shown to decrease growth rates, increase weights of liver, lungs, kidneys and adrenal glands, increase liver enzyme activity, interfere with porphyrin metabolism and renal function, and alter haematological and biochemical parameters. Microcontaminants appear to be the principal active moieties in the nonacute toxicity of commercial pentachlorophenol. Pentachlorophenol has been shown to be fetotoxic, delaying the development of rat embryos and reducing litter size, neonatal body weight and survival, and weanling growth.


Pentachlorophenol is not considered to be teratogenic, although birth defects arose as an indirect result of maternal hyperthermia in one study. Pentachlorophenol has been shown to be immunotoxic in several animal species. At least part of this effect is caused by pentachlorophenol itself. Neurotoxic effects have also been reported, but the possibility that these are due to microcontaminants has not been excluded. Pure pentachlorophenol has not been found to be highly mutagenic. The presence of at least one carcinogenic microcontaminant (hexachlorodibenzo-p-dioxin) suggests that the potential for technical pentachlorophenol to cause cancer in laboratory animals cannot be completely ruled out. IARC classified PCP in Group 2B (the agent is possibly carcinogenic to humans) on the basis of inadequate evidence of carcinogenicity in humans but sufficient evidence in experimental animals. There is suggestive, although inconclusive, evidence of the carcinogenicity of PCP from epidemiological studies of populations exposed to mixtures that include PCP. Conclusive evidence of carcinogenicity has been obtained in one animal species (mice). Although there are notable variations in metabolism between experimental animals and humans, it was considered prudent to treat PCP as a potential carcinogen. In addition, pentachlorophenol has been classified as a probable human carcinogen (Group B2) by the USEPA, for exposure via the oral route. However, human studies of high exposure groups, such as timber treatment workers, have not provided sufficient evidence of cancer. Chronic exposure to pentachlorophenol may result in a range of adverse health effects in humans including irritation of the skin and mucous membranes, chloracne, neuraesthesia, depression, headaches and changes in kidney and liver function. The toxicity of pure pentachlorophenol has not been evaluated in humans.

WHO based their 0.009 mg/L guideline value on multistage modelling of tumour incidence in a US NTP bioassay without incorporation of a body surface area correction, recognising that there are interspecies differences in metabolism between animals and humans, with an important metabolite formed in rats being only a minor metabolite in humans. The concentration of PCP associated with a 10-5 upper-bound excess lifetime cancer risk is similar to the guideline value established in the previous edition, so that guideline value is retained.

References Pontius FW (ed). Quality and treatment. AWWA (Water). McGraw-Hill Inc.




Revised July 2005.

New Zealand National Task Group on Site Contamination from the Use of Timber Treatment Chemicals. 1992. Pentachlorophenol Risk Assessment Pilot Study.

Shaw C. 1990. Timber Preservation in the New Zealand Environment: Pentachlorophenol and chlordane. Cawthron Institute Report.


WHO. 2003. Pentachlorophenol in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/62).


Permethrin

Maximum acceptable value (provisional) Based on health considerations, the concentration of permethrin in drinking-water should not exceed 0.02 mg/L. WHO (2004) did not derive a guideline value because �permethrin occurs in drinking-water at concentrations well below those at which toxic effects are observed�.


1 To source waters

Permethrin may enter source waters as a result of its application as a contact insecticide. It is used against a wide range of pests in agriculture, forestry and public health. The total annual usage of permethrin in New Zealand in the late 1980s was 3400 kg with the majority of use being in the North Island. As at August 2005, permethrin is registered for use in New Zealand.


Permethrin is a WHOPES recommended larvicide used to control aquatic invertebrates in water mains.

Forms and fate in the environment Permethrin is photodegraded in water and on soil surfaces. In soil it degrades rapidly by hydrolysis and microbial degradation. The half life in soils ranges from six to 105 days with a recommended average half life of 30 days. The water solubility of permethrin is 0.006 mg/L in some references and 0.2 mg/L in others and the sorption coefficient is about 1000 mL/g.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 346 zones, did not find detectable concentrations of permethrin (limit of detection = 0.0002 mg/L).


Concentrations as high as 0.8 mg/L have been recorded in surface water; levels in drinking-water have not been reported.

Removal methods Conventional coagulation is considered to remove permethrin readily from water. Under normal disinfection conditions, chlorine does not react with either isomer.

Referee method


Permethrin has a low acute oral toxicity in mammals. The cis-isomer is the more toxic form. Oral toxicities of the major metabolites of permethrin are lower than those of the parent compound. The major signs of acute intoxication are effects on the central nervous system, consisting of uncoordinated movements, whole body tremors, and loss of balance. Overt signs of toxicity do not appear until near-lethal doses.




No alternative methods have been recommended for permethrin because no methods meet the required criteria.

Exposure of the general population to permethrin is mainly via the diet.

Permethrin is absorbed readily when given orally. Distribution occurs rapidly in the body, mostly to adipose tissue, followed by liver, kidney and brain. Permethrin administered to mammals is almost completely eliminated from the body within several days.

Forestry workers using permethrin reported symptoms including itching and burning of the skin and itching and irritation of the eyes. Paraesthesia (sensation abnormality) was induced in volunteers about 30 minutes after the application of permethrin solution (total 0.5 mg) to the earlobe. Of 10 volunteers treated with 15�40 mL of a permethrin (1:3) (1%) head louse solution, three developed mild, patchy erythema (skin reddening), which faded 4�7 days later. Dietary permethrin does not appear to affect reproduction adversely in rats or mice. Permethrin has not exhibited mutagenic activity in a range of short-term mutagenicity assays. The International Agency for Research on Cancer has classified permethrin in Group 3 (not classifiable as to its carcinogenicity to humans), as there are no human data and only limited data from animal studies.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of a MAV for permethrin in drinking-water. In 1987, Joint FAO/WHO Meetings on Pesticide Residues (JMPR) recommended an acceptable daily intake for 2:3 and 1:3 cis:trans-permethrin which has been used in the derivation of the MAV given below. It is based on a two-year dietary study in rats that observed clinical signs and changes in body and organ weights and blood chemistry, and from a one-year study in dogs, based on reduced body weight. The MAV for permethrin in drinking-water was derived as follows:



where:


(Also called 2-phenylphenol.)

1 To source waters

• no observable adverse effect level = 5 mg/kg body weight per day for obtained in a rat study

• average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 0.01 • uncertainty factor = 100 (for inter and intra-species variation). JMPR noted that if permethrin is to be used as a larvicide for the control of mosquitos and other insects of health significance in drinking-water sources, the share of the ADI allocated to drinking-water may be increased.

References FAO/WHO. 2000. Pesticide Residues in Food � 1999: Evaluations � 1999. Part II � Toxicology. Geneva: World Health Organization. Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/00.4).

The Royal Society of Chemistry. 1987. The Agrochemicals Handbook (2nd ed). Surrey, England: Unwin Brothers Ltd.





WHO. 2003. Permethrin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/111).


Phenylphenol New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of phenylphenol in drinking-water should not exceed 1.4 mg/L. WHO 2004 states that phenylphenol occurs in drinking-water at concentrations well below those at which toxic effects are observed.


2-phenylphenol is used as a disinfectant, bactericide and virucide. In agriculture, it is used in disinfecting fruits, vegetables and eggs. It is also used as a general surface disinfectant in hospitals, nursing homes, veterinary hospitals, poultry farms, dairy farms, commercial laundries, barbershops and food processing plants. As at August 2005, phenylphenol is not registered for use in New Zealand.


Referee method

Form and fate in the environment 2-phenylphenol is degraded readily in surface waters, with a half-life of about one week in river water.


Removal methods

Analytical methods


Health considerations 2-phenylphenol has been determined to be of low toxicity. Both 2-phenylphenol and its sodium salt are carcinogenic in male rats, and 2-phenylphenol is carcinogenic in male mice. However, urinary bladder tumours observed in male rats and liver tumours observed in male mice exposed to 2-phenylphenol appear to be threshold phenomena that are species- and sex-specific. JMPR has concluded that 2-phenylphenol is unlikely to represent a carcinogenic risk to humans. Although a working group convened by IARC has classified the sodium salt of 2-phenylphenol in Group 2B (possibly carcinogenic to humans) and 2-phenylphenol in Group 3 (not classifiable as to its carcinogenicity to humans), JMPR noted that the IARC classification is based on hazard identification, not risk assessment, and is furthermore limited to published literature, excluding unpublished studies on toxicity and carcinogenicity. JMPR also concluded that there are unresolved questions about the genotoxic potential of 2-phenylphenol.

Derivation of maximum acceptable value The MAV (provisional) for phenylphenol in drinking-water was derived as follows:


where:

• no observable adverse effect level = 39 mg/kg body weight per day in a two-year toxicity study for decreased body weight gain and hyperplasia of the urinary bladder and carcinogenicity of the urinary bladder in male rats




• uncertainty factor = 100 for intra- and interspecies variation. WHO (2004) states that because of its low toxicity, the health-based value derived for 2-phenylphenol is much higher than 2-phenylphenol concentrations likely to be found in drinking-water. Under usual conditions, therefore, the presence of 2-phenylphenol in drinking-water is unlikely to represent a hazard to human health. For this reason, the establishment of a guideline value for 2-phenylphenol is not deemed necessary.


References FAO/WHO. 2000. Pesticide Residues in Food � 1999 evaluations: Part II � Toxicological. Geneva: World Health Organization. Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/00.4).

WHO. 2003. 2-phenylphenol and its Sodium Salt in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/69).


Phorate New entry August 2005. (Also called thimet.)



There are insufficient data to determine a MAV for phorate in drinking-water. WHO 2004 states that phorate is unlikely to occur in drinking-water, so did not develop a guideline value for drinking-water.

1 To source waters

Phorate, an organophosphorus compound, is an insecticide and acaricide that controls pests by systemic, contact, and fumigant action. It is used against sucking and chewing insects, leafhoppers, leafminers, mites, some nematodes and rootworms. Phorate is used in pine forests and on root and field crops, including corn, cotton, coffee, some ornamental and herbaceous plants and bulbs.

As at August 2005, phorate is registered for use in New Zealand.

Forms and fate in the environment In the environment, phorate is degraded by micro-organisms and interaction with water. Phorate itself is not persistent in plants. However, phorate protects plants for a long time because its breakdown product persists in plants and soils. Phorate binds to soil organic matter and clay particles and is almost immobile in soils. Thus, it does not leach easily and is transported mainly with runoff via sediment and water. Phorate has some potential, though minimal, to leach through the soil and contaminate groundwater, particularly where soils are sandy and aquifers are shallow. Phorate is moderately persistent in the soil. Its half life under aerobic laboratory conditions is 82 days, while a field study noted a half life of 7.5 days. It is least persistent in clay soil, while it is slowly released from peat/sand and sandy soils. Phorate disappears almost completely from sand/muck soils within one year. Phorate is unstable in water, especially under alkaline (basic) conditions. As it breaks down in water, non-toxic, water-soluble products are formed.


Removal methods No information is available on processes that can be used to remove phorate from water.


Analytical methods

Referee method

A referee method cannot be selected for phorate because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for phorate for the above reason. However, the following information may be useful: [???]

Health considerations The oral LD50 for rats is 1.0 mg/kg. The oral LD50 for mice ranges from 3.5 to 6.59 mg/kg. Guinea pigs have an oral LD50 of 20 mg/kg. Long-term studies of mice fed high doses of 98.7% pure phorate showed no effects on fertility, gestation, and viability. This suggests that phorate is unlikely to cause reproductive effects in humans. No birth defects were found in two studies on the rat. This suggests that phorate does not cause birth defects. There was some maternal and embryo toxicity at relatively low doses (0.5 mg/kg). Available mutagenicity studies involving microbial and mammalian cells have shown no adverse effects on genes or chromosomes. Thus it appears that phorate does not cause mutations. Valid studies on the carcinogenicity of phorate are not available.

Derivation of maximum acceptable value There are limited and insufficient data on phorate on which to propose a MAV for drinking-water.

References Extoxnet. Pesticide Information Profile. Oregon State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/


USEPA. 1985. Pesticide Fact Sheet for Phorate: Fact Sheet No 34.1. Environmental Protection Agency, Office of Pesticide Programs.

Walker MM, Keith LH. 1992. EPA�s Pesticide Fact Sheet Database. Chelsea, USA: Lewis Publishers.


Picloram New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of picloram in drinking-water should not exceed 0.2 mg/L (200 µg/L). Picloram is not mentioned in WHO 2004. The maximum contaminant level (USEPA 2004) is 0.5 mg/L.



1 To source waters

Picloram is registered for use in New Zealand as a systemic herbicide. It is available in a variety of formulations, some of which contain other active ingredients: clopyralid, 2,4-D or triclopyr. Trade names include: Radiate, Tordon 2G and 50-D, Tordon Brushkiller and Tordon Gold Herbicide, Vigilant. No information is available on the annual usage of specific active ingredients in New Zealand, although picloram is understood to be likely to constitute less than 5% of the �Other Hormone Type� class of pesticides (P Holland, personal communication).

Forms and fate in the environment Picloram is very soluble in water: 430 mg/L (Merck & Co 1996). The principal environmental risks of picloram relate to contamination of surface water and groundwater, and damage to non-target terrestrial plants including crops adjacent to areas of application via runoff or drift. Picloram is among the most mobile of currently (USEPA) registered pesticides and eventual contamination of groundwater is virtually certain in areas where picloram residues persist in the overlying soil (USEPA 1995). Picloram is resistant to biotic and abiotic degradation processes. It is stable to hydrolysis and anaerobic degradation, and degrades very slowly with half lives ranging from 167 to 513 days. Picloram is extremely mobile. Nearly 100% of the chemical leached but none of it degraded over a three-year period in a University of Arkansas study. Given its high persistence, it appears unlikely that picloram will degrade once it reaches groundwater, even over a period of several years (USEPA 1995).

In New Zealand national pesticides surveys conducted every four years since 1990, picloram has been detected in groundwater once, at a concentration of 0.0003 mg/L. Pesticide monitoring by Environment Canterbury has detected picloram twice in groundwater from one location (concentrations 0.00018 and 0.0003 mg/L) in and close to the Level Plain area in South Canterbury (Close et al 2001). Additionally, picloram has been detected at three sites in the Waikato region in pesticide monitoring of groundwater conducted by Environment Waikato. Concentrations ranged from 0.00002�0.0028 mg/L (Hadfield and Smith).

USEPA classified picloram as a Restricted Use pesticide in 1978 as a result of recurring reports of phytotoxicity to economically important crops caused by contamination of water supplies (USEPA 1995).

Studies of mobility and degradation in New Zealand soils have reported mobilities (as Koc) ranging from 19 to 47, which suggest a moderate level of adsorption to organic soils (Close et al 2001). There is no information available regarding the greatest source of exposure to picloram for New Zealanders (ie, dermal contact, inhalation, diet: food, water).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included picloram, so typical concentrations in New Zealand drinking-waters are unknown.

The USEPA Office of Drinking Water�s STORET database indicates that picloram has been reported in 420 of 744 surface water samples (USEPA 1995). However, despite its persistence and mobility in soils, picloram has been detected infrequently in surface water and groundwater in four provinces in Canada (Hiebsch 1988, cited in Health Canada 1988).


Removal methods Picloram is oxidised by ozone (Haag and Yao 1992). This is achieved most effectively at higher pH values which favour hydroxyl radical formation. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of picloram, although a guide to the efficiency of the process cannot be provided. Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.


Referee method

Liquid/liquid extraction/gas chromatography-electron capture detector (EPA 515.2).


High pressure liquid chromatography/photodiode array ultraviolet detector (EPA 555).

Health considerations Absorption of picloram through the gastrointestinal tract is rapid and almost complete. In human volunteers, the absorption half-time was 20 minutes. After oral administration of 5 and 0.5 mg/kg body weight radiolabelled doses, concentrations in the blood were proportional to the dose administered and were highest during the first hour. Picloram was excreted rapidly and unchanged in the urine. Most of the dose (77 to 86%) was excreted within the first six hours, and 94% of the dose was recovered after 72 hours (Nolan et al 1984, cited in Health Canada 1998).

Acute exposure

In studies using laboratory animals, picloram generally has been shown to be of moderate to low acute toxicity with oral LD50 values in the range of 2000 to 8000 mg/kg body weight for the rabbit, mouse, guinea pig and rat (NRC 1974, cited in Health Canada 1988). These levels suggest a low/variable acute oral toxicity when compared with other pesticides. It has been shown to potentially cause the following health effects from acute exposures: damage to central nervous system, weakness, diarrhoea and weight loss. Picloram and its derivatives are only slightly toxic by the oral routes and USEPA have placed it in Toxicity Category III (the second lowest of four categories) for this effect.

Chronic exposure

In a subchronic toxicity study using rats, picloram caused changes in the liver. A dog dietary study resulted in decreases in body weight gain, food consumption, liver weights and several enzymes. Chronic exposure to picloram has the potential to cause liver damage. A chronic toxicity study using dogs resulted in increased liver weight. A chronic/carcinogenicity study using rats resulted in chronic toxicity in males only and no evidence of carcinogenicity. Based on these studies, picloram was classified as a Group E chemical, one showing evidence of non-carcinogenicity for humans (USEPA 1995). Picloram is classified by the International Agency for Research on Cancer (IARC) as Group 3: Unclassifiable as to carcinogenicity to humans.


Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for picloram in drinking-water, as follows:


where: • tolerable daily intake = 0.07 mg /kg body weight per day

Hiebsch S. 1988. The occurrence of 35 pesticides in Canadian drinking-water and surface water. Unpublished report prepared for Environmental Health Directorate, Department of National Health and Welfare.

Nolan R, Freshour N, Kastl P, et al. 1984. Pharmacokinetics in picloram in male volunteers. In: Toxicol Appl

• average weight of adult = 70kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10%.

References Close ME, Rosen MR, Smith VR. 2001. Fate and transport of nitrates and pesticides in New Zealand�s aquifers. In: Rosen MR, White PA (eds). Groundwaters of New Zealand, Chapter 8, pp. 185�220. New Zealand: The Hydrological Society.

Haag WR, Yao CCD. 1992. Rate constants for reaction of hydroxyl radicals with several drinking-water contaminants. In: Environ Sci Technol 26: 1005�13.



Merck & Co. 1996. The Merck Index: An encyclopedia of chemicals, drugs and biologicals (12th ed). Published by Merck Research Laboratories Division of Merck & Co Inc.

National Research Council of Canada. 1974. Picloram: The effects of its use as a herbicide on environmental quality. Ottawa: Associate Committee on Scientific Criteria for Environmental Quality.

Pharmacol 76: 264�[add page number].

USEPA. 1995. EPA RED Facts: Picloram: Prevention, pesticides and toxic substances (7508W). EPA-738-F-95-018.


Pirimiphos methyl New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of pirimiphos methyl in drinking-water should not exceed 0.1 mg/L. Pirimiphos methyl is not mentioned in WHO 2004.



1 To source waters

Pirimiphos methyl may enter source waters as a result of its application as a broad spectrum insecticide for grain storage pests and for industrial and domestic fly control. The total annual usage of pirimiphos methyl in the late 1980s was 19,800 kg. As at August 2005, it is registered for use in New Zealand.

Forms and fate in the environment The water solubility is 9 mg/L and the sorption coefficient is 1000 mL/g (estimate).

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 342 zones, did not find detectable concentrations of pirimiphos methyl (limit of detection = 0.0002 mg/L).



Referee method

Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector or Flame Photometric Detector (Organophosphorus pesticides in river and drinking-water, tentative method 1980; and Organophosphorus pesticides in sewage sludge: organophosphorus pesticides in river and drinking-water: an addition 1985; HMSO 1986).


No alternative methods have been recommended for pirimiphos methyl because no methods meet the required criteria.

Health considerations No specific information available, but as an organo-phosphate, it may be expected to show characteristic effects including inhibition of acetyl cholinesterase and central nervous system depression. Organo-phosphates are absorbed readily through the skin, and through the respiratory and gastrointestinal tracts.

Derivation of maximum acceptable value The MAV for pirimiphos methyl was calculated by the New Zealand Ministry of Health as follows:


where: • acceptable daily intake = 0.03 mg/kg body weight • average weight of adult = 70 kg • proportion of acceptable daily intake allocated to drinking-water = 0.1 • average quantity of water consumed by an adult = 2 L/day.


References HMSO. 1986. Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector or Flame Photometric Detector (organophosphorus pesticides in river and drinking-water, tentative method 1980; and organophosphorus pesticides in sewage sludge: organophosphorus pesticides in river and drinking-water: an addition, 1985).

JMPR. 1993. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues. Geneva, 20�29 September 1993, FAO Plant Production and Protection Paper No 122 (Annex 1).


Pirimisulfuron methyl New entry August 2005.

Based on health considerations, the concentration of pirimiphos methyl in drinking-water should not exceed 0.1 mg/L. Pirimisulfuron is not mentioned in WHO 2004.


1 To source waters

Pirimisulfuron (sometimes spelt primisulfuron) is a urea herbicide registered for use in New Zealand. There is one formulation (Beacon), which is sold as water dispersible granules. The compound is a selective post-emergence systemic herbicide that is absorbed rapidly by plants and transported throughout the plant roots and foliage. No information is available on the annual usage of specific active ingredients in New Zealand, although pirimisulfuron methyl is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).

Forms and fate in the environment Pirimisulfuron methyl is moderately soluble in water: 70mg/l at 20°C (PMEP 2001). Laboratory tests on well-aerated soils indicated a half life of less than two months. In soils without oxygen, breakdown of the compounds took nearly three months (RSocC 1991, cited in PMEP 2001). Sunlight has little effect on the breakdown of the compound in soil or in water. Dissipation of the compound under field conditions was much quicker, with a half life of three to 12 days (RSocC 1991, cited in PMEP 2001). Pirimisulfuron methyl is very mobile in soil in field and laboratory studies. The mobility of the compound in soils indicates that it has the potential to leach to surface water and to groundwater that lies close to the surface in highly vulnerable soils. The breakdown of the compound is much quicker in acidic soils and water rather than in neutral or alkaline soils and water (Walker and Keith 1991). There is no information available regarding the greatest source of exposure to pirimisulfuron methyl for New Zealanders (ie, dermal contact, inhalation, diet: food, water).


Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included pirimisulfuron methyl, so typical concentrations in New Zealand drinking-waters are unknown. Information on typical concentrations in international drinking-waters was unavailable.

Removal methods No information is available on the removal of pirimisulfuron methyl from water. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of pirimisulfuron methyl, although a guide to the efficiency of the process cannot be provided. Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.


Referee method

None listed for pirimisulfuron methyl in Table 12.3 of the DWSNZ Not in APHA 1998 either


None listed for pirimisulfuron methyl in Table 12.3 of the DWSNZ


Acute exposure

Pirimisulfuron methyl is a slightly toxic compound with an acute oral LD50 greater than 5050 mg/kg in the rat, which suggests a low acute oral toxicity when compared with other pesticides. Slight skin irritation was observed in rabbits after dermal application of primisulfuron methyl, but it did not cause skin sensitisation in male guinea pigs (Extoxnet 1996).

Chronic exposure

Rats and mice fed very large amounts of pirimisulfuron methyl for up to ninety days showed no ill effects (RSocC 1991, cited in PMEP 2001). This indicates that even at relatively high levels of exposure (500 mg/kg) there is little toxic effect from the compound over relatively short exposure periods. Male rats fed higher doses of primisulfuron-methyl (up to 1000 mg/kg) for three months showed a decrease in body weight and a decrease in spleen weight. The lowest dose at which adverse effects were noted was about 150 mg/kg (RSocC 1991, cited in PMEP 2001). When dogs were fed moderate doses of pirimisulfuron methyl for a year, a number of effects were noted at the highest doses tested (125 mg/kg) which included changes in the blood such as increased platelets and anaemia. Other changes included decreased cholesterol, pale livers, and thyroid gland changes. This evidence suggests that the chronic risks of human exposure to moderate levels of the herbicide for extended periods of time are slight (PMEP 2001).


In a two-generation study, rats were fed moderate doses (250 mg/kg) of pirimisulfuron methyl. Rats had decreased testicular function and the offspring had decreased body weights. No compound related reproductive effects were noted at doses below 50 mg/kg/day (Walker and Keith 1991, cited in PMEP 2001). Pregnant rabbits fed high doses of pirimisulfuron methyl produced normal offspring. Only at doses of 300 mg/kg and above were there changes in maternal body weight, spontaneous abortion and changes in the maternal stool. In another study, pregnant rats were fed moderate doses of pirimisulfuron methyl. They showed increases in the number of litters having incomplete bone formations, though this study was inconclusive. The USEPA has indicated that it is unlikely, given the results of the three studies noted above, that pirimisulfuron methyl is teratogenic to humans (Walker and Keith 1991, cited in PMEP 2001). No studies have found mutagenic effects. The International Agency for Research on Cancer has not classified pirimisulfuron methyl for its ability to cause cancer.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for pirimisulfuron methyl in drinking-water, as follows:

0.25 mg/kg body weight per day x 70 kg x 0.1 = 0.9 mg/L 2 L

where: • no observable adverse effect level = 0.25 mg /kg body weight per day • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10%.

References Extoxnet. 1996. Pesticide Information Profile: Primisulfuron methyl. Oregon State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

Royal Society of Chemistry. 1991. The Agrochemicals Handbook: Pirimisulfuron methyl (3rd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

PMEP. 2001. Extoxnet: Extension Toxicology Network: Pesticide Information Profile: Primisulfuron methyl. http://pmep.cce.cornell.edu/profiles/extoxnet/

Walker M, Keith L. 1991. EPA�s Pesticide Fact Sheet Database. Chelsea, USA: Lewis Publishers.


http://pmep.cce.cornell.edu/profiles/extoxnet/

Procymidone New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of procymidone in drinking-water should not exceed 0.7 mg/L. Procymidone is not mentioned in WHO (2004).


1 To source waters

No information available. As at August 2005, procymidone is registered for use in New Zealand.


Typical concentrations in drinking-water In the three national surveys on pesticides in groundwater done up until 1999, procymidone was found in a concentration range of 0.0001 to 0.0017 mg/L. It is not widely found, with market gardening in Pukekohe being one of the areas in which it was found on a number of occasions. In other surveys in New Zealand the concentration range found was 0.00001 to 0.0034 mg/L.

Removal methods No information is available on processes that can be used to remove procymidone from water.

Analytical methods

Referee method

[???]


[???]

Health considerations Level causing no toxicological effect: • mouse: 100 ppm in the diet, equal to 15 mg/kg bw/day • rat: 250 ppm in the diet, equivalent to 12.5 mg/kg bw/day • dog: 100 mg/kg bw/day. Estimates of acceptable daily intake of procymidone for humans are 0�0.2 mg/kg bw. Procymidone was negative in various mutagenicity assays, there was no evidence of genotoxicity, and its carcinogenicity has not been classified by IARC.

Derivation of maximum acceptable value The MAV for procymidone in drinking-water was derived, for non-carcinogenic effects, based on an ADI of 0.1 mg/kg, as follows:


0.1 mg/kg body weight per day x 70 kg x 0.1 = 0.35 mg/L 2 L However, because procymidone is unlikely to be found in New Zealand drinking-waters, it was decided to retain the 0.7 mg/L MAV from DWSNZ 1995 and 2000. The 0.7 mg/L MAV had been derived by using the proportion of allowable daily intake allocated to drinking-water = 0.2. The 1989 Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and a WHO Expert Group on Pesticide Residues had decided to adopt a fraction of 0.1 instead (FAO Plant Production and Protection Paper 99, 1989, and Part II � Toxicology. FAO Plant Production and Protection Paper 100/2, 1990, as cited in IPCS INCHEM).

References IPCS INCHEM. International Programme on Chemical Safety (IPCS) and Canadian Centre for Occupational Health and Safety (CCOHS). Information available at: http://www.inchem.org/documents/jmpr/jmpmono/v89pr12.htm. Core site at www.inchem.org

Revised July 2005.



Propanil

Based on health considerations, the concentration of propanil in drinking-water should not exceed 0.02 mg/L. WHO (2004) states that although a health-based value for propanil can be derived, this has not been done, because propanil is transformed readily into metabolites that are more toxic. Therefore, a guideline value for the parent compound is considered inappropriate, and there are inadequate data on the metabolites to allow the derivation of a guideline value for them. National authorities should consider the possible presence in water of more toxic environmental metabolites.

1 To source waters

Propanil may enter source waters as the result of its application as a contact herbicide. It is used post-emergence (mainly in rice) to control broadleaved and grass weeds. It is also used in a mixture with MCPA for wheat. It is not registered currently in New Zealand, although it has been in the past.

Forms and fate in the environment Propanil is degraded rapidly in water by sunlight to phenolic compounds. In soil propanil is biodegraded to various metabolites with half lives ranging from one to 15 days. The recommended average soil half life is one day. The water solubility of propanil ranges from 130 to 225 mg/L and the sorption coefficient is 149 mL/g.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find detectable concentrations of propanil (limit of detection = 0.0001 mg/L).


http://www.inchem.org/documents/jmpr/jmpmono/v89pr12.htm

http://www.inchem.org/

Although used in a number of countries, propanil has only occasionally been detected in groundwater (WHO 2004).

Removal methods No specific information on methods of removing propanil from water is available. However, chlorine has been reported to be effective in the break down of this family of pesticides. Slow sand filtration has no effect on the concentrations of these pesticides.


Referee method

No referee method has been given for propanil because no method meets the required criteria.


No alternative methods have been recommended for propanil because no methods meet the required criteria. However, the following information may be useful: Propanil can be determined in water samples by extraction with methylene chloride and analysis by gas chromatography with a nitrogen phosphorus detector (eg, Method EPA 507). Confirmation by a second capillary column with different polarity is strongly recommended. No information is available for the limit of quantitation.

The probable oral lethal dose for humans is 0.5�5 g/kg body weight. Exposure produces local irritation and central nervous system depression. Ingestion causes irritation with a burning sensation in the mouth, oesophagus, and stomach, with gagging, coughing, nausea and vomiting, followed by headache, dizziness, drowsiness and confusion.

Health considerations Propanil and its metabolites do not appear to accumulate in tissues. Six metabolites have been detected in urine. Propanil has moderate acute toxicity. Two of its environmental metabolites, 3,4-dichloroaniline and 3,3�,4,4�-tetrachlorobenzene are more toxic than the parent compound. Animal studies show that under conditions of long-term exposure, propanil is toxic to red blood cells.

Workers from a pesticide plant who were exposed to the propanil metabolite 3,4-dichloroaniline showed signs of methaemoglobinaemia. Of the 28 workers exposed to 3,4-dichloroaniline and propanil, 17 showed signs of chloracne, which was attributed to the presence of contaminants. Propanil is not considered to be genotoxic. However, at least one of propanil�s environmental metabolites (tetrachloroazobenzene) is genotoxic. Data from a limited study in rats do not provide evidence of carcinogenicity.

Derivation of maximum acceptable value As the limited data available do not provide evidence of carcinogenicity, a tolerable daily intake approach has been used for the derivation of the MAV for propanil in drinking-water. The no observable adverse effect level used in the derivation is from a three-month rat feeding study. The MAV for propanil in drinking-water was derived as follows:

5 mg/kg body weight/day x 70 kg x 0.1 = 0.02 mg/L


2 L/day x 1000

where:

• no observable adverse effect level = 5 mg/kg body weight per day from a three-month rat feeding study




• uncertainty factor = 1000 (100 for inter and intra-species variation and an additional 10 for the short duration of the study and limitations of the database).

References Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: 466�[add page number].



1 To source waters

Propazine has low solubility in water (20°C): 8.6 mg/L (Merck & Co 1996).

WHO. 2003. Propanil in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/112).


Propazine New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of propazine in drinking-water should not exceed 0.07 mg/L (70 µg/L). Propazine is not mentioned in WHO (2004).


Propazine is used as a pre-emergence selective triazine herbicide. It is registered for use as a herbicide in New Zealand, and is available as a suspension concentrate (trade name Agpro Propazine 500) or wettable powder (trade name Gesamil 50WP).

No information is available on the annual usage of specific active ingredients in New Zealand, although propazine is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication).


If applied to an outdoor environment, propazine has a high potential to leach into groundwater or reach surface waters by runoff. Propazine is stable to hydrolysis in sterile aqueous pH 5, 7 and 9 buffered solutions. However, published literature on propazine and related chloro-s-triazines indicate that the chemical may be susceptible to hydrolysis after adsorption on to the surface of soil colloids (a surface catalysis effect). Propazine is moderately persistent to degradation under aerobic soil conditions,


degrading with half lives of 12 to 24 weeks (calculated 15 weeks) in a non-sterile loamy sand and eight to 12 weeks in a sterile loamy sand soil. Batch equilibrium studies suggest that propazine is mobile. There is no information available regarding the greatest source of exposure to propazine for New Zealanders (ie, dermal contact, inhalation, diet: food, water).

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included propazine, so typical concentrations in New Zealand drinking-waters are unknown. No information on typical concentrations in international drinking-waters was available.

Removal methods Oxidation of triazines by ozone is reported to be effective (Chiron et al 2000). The water chemistry, in particular the alkalinity and pH, will affect the oxidation rate. Use of activated carbon following ozonisation should be considered to adsorb oxidation products. Nanofiltration (membrane technology) in water with a low natural organic matter concentration is reported to remove approximately 50% of atrazine and simazine (Agbekodo et al 1996). The percentage is increased to 90�100% when 3.6 mg/L of natural organic matter is present. Similar results may be expected for propazine as it is from the same chemical family and of comparable size. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of propazine, although a guide to the efficiency of the process cannot be provided.


Referee method

None listed for propazine in Table 12.3 of the DWSNZ not in APHA 1998 either



Health considerations Absorption of propazine from the gastrointestinal tract has been found to be rapid and similar for all study groups. Within 48 hours of treatment, 82�95% of the administered dose was recovered from excreta, predominantly the urine. No specific target organs were identified (USEPA 1998). The structural similarity of propazine to other triazine herbicides suggests that propazine may cause endocrine effects similar to those caused by atrazine in female rats (USEPA 1998). No reports on the effects of exposure of humans to propazine have been identified.

Acute exposure

The LD50 for acute oral toxicity in rats is greater than 5050 mg/kg/day (USEPA 1998), which suggests a low toxicity when compared with other pesticides. Administration of lethal or near lethal doses to rats has caused symptoms of lethargy, muscular weakness, runny nose, emaciation, diarrhoea, and laboured breathing. It is mildly irritating to the skin, eyes, and


upper respiratory tract. Contact dermatitis has been reported among workers manufacturing propazine. No cases of poisoning from human ingestion of this herbicide have been recorded (Extoxnet 2001).

Chronic exposure

USEPA has based their NOAEL and LOAEL chronic toxicity concentrations upon decreased body weight (USEPA 1998). Reproductive toxicity is based on decreased ossification and decreased body weight. The International Agency for Research on Cancer has not classified propazine for its ability to cause cancer, but the USEPA has classified it as a Group �C� (possible human carcinogen) chemical based on significant increases in mammary gland adenomas and adenomas/carcinomas in female Sprague-Dawley rats (USEPA 1998).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for propazine in drinking-water, as follows:

0.02 mg/kg body weight per day x 70 kg x 0.1 = 0.07 mg/L 2 L

where: • acceptable daily intake = 0.02 mg /kg body weight per day • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of tolerable daily intake allocated to drinking-water = 10%.

References Agbekodo KM, Legube B, Dard S. 1996. Atrazine and simazine removal mechanisms by nanofiltration: influence of natural organic matter concentration. In: Wat Res 34(11): 2535�42.

Chiron S, Fernandez-Alba A, Rodriguez A, et al. 2000. Pesticide chemical oxidation: state-of-the-art. In: Wat Res 34(2): 366�77.

Extoxnet. 2001. Pesticide Information Profile: Propazine. New York: PMEP, Cornell University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/


USEPA. 1998. EPA Pesticide Fact Sheet: Propazine. Office of Prevention, Pesticides and Toxic Substances (7501C).

Propoxur New entry August 2005. (Also called Baygon.)

Maximum acceptable value There are insufficient data to determine a MAV for propoxur in drinking-water. WHO 2004 states that propoxur is unlikely to occur in drinking-water, so did not develop a guideline value for drinking-water.



1 To source waters

Propoxur is a non-systemic carbamate insecticide which was introduced in 1959(25). [What does �25� refer to?] Propoxur is not used on food crops. It is used against mosquitoes in outdoor areas, for flies in agricultural settings, for fleas and ticks on pets, as an acaricide, on lawns and turf for ants, on flowering plants, and in private dwellings and public buildings. It is also used as a molluscicide, a chemical that kills snails. It is effective against cockroaches, aphids and leafhoppers. Propoxur is one of the chemicals that have, to a large extent, replaced DDT in the control of black flies and mosquitoes. As at August 2005, propoxur is registered for use in New Zealand.

Forms and fate in the environment Because it is both highly soluble in water (2000 mg/L) and has a lengthy soil half-life (28 days), and does not adsorb strongly to soil particles, propoxur has a high potential for groundwater penetration. In one study, there was practically no loss of propoxur from a silt-loam soil to which it was applied during a six-month period, but 25% of applied Baygon was lost from sand in 100 days. In another study, propoxur was very mobile in sandy loam, silt loam and silty clay soils. The rate of biodegradation increases in soils that have been exposed previously to propoxur or other methylcarbamate pesticides.

Typical concentrations in drinking-water No information is available.

Removal methods No information is available on processes that can be used to remove procymidone from water.

Analytical methods

Referee method

A referee method cannot be selected for propoxur because a MAV has not been established and therefore the sensitivity required for the referee method is not known.


No alternative methods can be recommended for propoxur for the above reason. However, the following information may be useful: [???]

Health considerations Carbamates generally are excreted rapidly in urine and do not accumulate in mammalian tissue. If exposure does not continue, cholinesterase inhibition reverses rapidly. The LD50 for propoxur in rats ranges from 83 mg/kg to 150 mg/kg. In rats, propoxur poisoning resulted in brain pattern and learning ability changes at lower concentrations than those which caused cholinesterase-inhibition and/or organ weight changes. The oral LD50 in mice is 23.5 mg/kg, 40 mg/kg in guinea pigs. Twelve-month old male goats have an oral LD50 greater than 800 mg/kg. The oral LD50 for technical propoxur in rats was 50 mg/kg for males and 104 mg/kg for females. Propoxur did not cause mutations in six different types of bacteria. A derivative of propoxur (N-nitroso) is mutagenic. No carcinogenic effects have been reported for propoxur.


Derivation of maximum acceptable value There are limited and insufficient data on propoxur on which to propose a MAV for drinking-water.

USEPA. 1988. (Baygon) Propoxur Health Advisory. Washington DC: Environmental Protection Agency, Office of Drinking Water.

Pyridate may enter source waters as a result of its use as a foliar-acting contact herbicide for dicotyledonous plants and some grassy weeds.

No information on methods of removing pyridate is available.

Referee method

References Extoxnet. Pesticide Information Profile. Oregon State University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/


Pyridate Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of pyridate in drinking-water should not exceed 0.1 mg/L. WHO (2004) did not derive a guideline value because �pyridate is not persistent and is only rarely found in water�. Their 0.1 mg/L GV appeared in their 1998 publication.


1 To source waters

Pyridate is currently registered for use in New Zealand as the emulsifiable concentrate Tough 450 EC, and the wettable powder Lentagran 45 WP.

Forms and fate in the environment Pyridate has a low water solubility (1.5 mg/L) and relatively low mobility. It is not persistent and is hydrolysed, photodegraded and biodegraded rapidly. Under favourable conditions its environmental half life is of the order of a few days.

Typical concentrations in drinking-water No New Zealand data, and very few data from overseas, are available for pyridate in drinking-water supplies.

Removal methods


No referee method has been given for pyridate because no method meets the required criteria.



A tolerable daily intake approach has been used for the derivation of the MAV of pyridate in drinking-water. The no observable adverse effect level used in the derivation is based on increased kidney weight in a two-year rat feeding study.

No alternative methods have been recommended for pyridate because no methods meet the required criteria. However, the following information may be useful: The main pyridate metabolite may be determined by high performance liquid chromatography followed by ultraviolet absorption at 254 or 280 nm (Chemie Linz). Pyridate-D and 2,4-dichlorophenoxypropionic acid isooctyl ester may be determined using a method in which the sample is dissolved in chloroform and extracted with sodium hydroxide, followed by addition of morpholine and measurement of the compounds by UV absorption at 298 nm. No quantitative limits are cited.

Health considerations Following oral administration, pyridate is absorbed rapidly by the gut and distributed to the organs. It is hydrolysed in the blood and in artificial intestinal juices of rats. Pyridate is excreted rapidly, mainly in urine. Pyridate has been tested in long-term feeding studies in rats and mice and symptoms reported included increased liver weight, decreased body growth.

The International Agency for Research on Cancer has not evaluated pyridate. No evidence of carcinogenicity was found in long-term feeding studies in rats and mice. The available evidence indicated that pyridate is not genotoxic.


The MAV for pyridate in drinking-water was derived as follows (WHO 1998):


where:

• no observable adverse effect level = 3.5 mg/kg body weight per day for increased kidney weight in a two-year rat feeding study




Pyriproxifen



References Chemie Linz. 1981. Method of analysis for determination of 6-chloro-4-hydroxy-3-phenyl-pyridazine in leaching water. Unpublished study submitted to the World Health Organization.

World Health Organization. 1998. Guidelines for Drinking-water Quality (2nd ed). Volume 2: Health Critieria and other Supporting Information.



Maximum acceptable value (provisional) Based on health considerations, the concentration of pyriproxifen in drinking-water should not exceed 0.4 mg/L.


1 To source waters

Pyriproxyfen is a broad-spectrum insect growth regulator with insecticidal activity against public health insect pests. It is a WHOPES-recommended insecticide for the control of mosquito larvae. In agriculture and horticulture, pyriproxyfen has registered uses for the control of scale, whitefly, bollworm, jassids, aphids and cutworms. As at August 2005, pyriproxifen is registered for use in New Zealand.

Forms and fate in the environment Pyriproxyfen degrades rapidly in soil under aerobic conditions, with a half life of 6.4�36 days. It disappeared from aerobic lake water sediment systems with half lives of 16 and 21 days. Pyriproxyfen appeared to be degraded much more slowly in anaerobic lake water sediment systems. As pyriproxyfen is a fairly new pesticide, few environmental data have been collected.

Typical concentrations in drinking-water No information is available. No detectable concentrations found in surface water in the USA (WHO 2004).

Removal methods No data available; 0.001 mg/L should be achievable using GAC (WHO 2004).

Analytical methods

Referee method

[???]


[???]

Health considerations Intake of pyriproxyfen from all sources is generally low and below the ADI. JMPR concluded that pyriproxyfen was not carcinogenic or genotoxic. In short- and long-term studies of the effects of pyriproxyfen in mice, rats and dogs, the liver (increases in liver weight and changes in plasma lipid concentrations, particularly cholesterol) was the main toxicological target.

Derivation of maximum acceptable value The MAV for pyriproxyfen in drinking-water was derived as follows:

10 mg/kg body weight/day x 70 kg x 0.1 = 0.35 mg/L (rounded to 0.4 mg/L) 2 L x 100


where:

• no observable adverse effect level = 10 mg/kg body weight per day based on increased relative liver weight and increased total plasma cholesterol concentration in male dogs in two one-year toxicity studies


There are insufficient data to determine a MAV for quintozene in drinking-water. WHO 2004 states that quintozene is unlikely to occur in drinking-water, so did not develop a guideline value for drinking-water.

1 To source waters



• uncertainty factor = 100 for intra- and interspecies variation.

References FAO/WHO. 2000. Pesticide Residues in Food � 1999 Evaluations: Part II � Toxicological. World Health Organization, Geneva. Joint FAO/WHO Meeting on Pesticide Residues (WHO/PCS/00.4).

WHO. 2003. Pyriproxyfen in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/113).


Quintozene New entry August 2005. (Also called pentachloronitrobenzene or PCNB.)



No information available. As at August 2005, quintozene is registered for use in New Zealand.




Analytical methods

Referee method

A referee method cannot be selected for quintozene because a MAV has not been established and therefore the sensitivity required for the referee method is not known.



No alternative methods can be recommended for quintozene for the above reason. However, the following information may be useful: [???]

Health considerations Ex USEPA: a two-year feeding study with dogs (four males and four females/group) given diets containing 0, 30, 180, or 1080 ppm indicated that PCNB (1.4% hexachlorobenzene) caused liver weight increases, increased liver-to-body weight ratios, elevated serum alkaline phosphatase levels, and microscopically observed cholestatic hepatosis with secondary bile nephrosis at 1080 ppm (the highest dose tested). An interim sacrifice at one year occurred with one dog/sex/group; the remaining animals were sacrificed at two years. The cholestatic changes were observed in all animals given diets containing 180 and 1080 ppm PCNB, and one of three male dogs in the 30 ppm dose group exhibited the microscopic changes (no female dogs were affected). The authors noted that these histopathologic changes were moderate in the 1080 ppm group and minimal in the 180 ppm group. Based on these results, 30 ppm was the NOEL and 180 ppm was the LEL in dogs. An uncertainty factor of 100 was used to account for the inter- and intraspecies differences. An additional UF of 3 was used since the database for chronic toxicity is incomplete. Conversion factor: 1 ppm = 0.025 mg/kg/day (assumed dog food consumption).

Derivation of maximum acceptable value There are limited and insufficient data on quintozene on which to propose a MAV for drinking-water.

References NTP (National Toxicology Program). 1986. Technical Report on the Toxicology and Carcinogenesis Studies of Pentachloronitrobenzene in B6C3F1 Mice (feed studies). Report No NIH 86-2581.

USEPA. http://www.epa.gov/IRIS/subst/0254.htm


Simazine Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of simazine in drinking-water should not exceed 0.002 mg/L (2 µg/L). The maximum contaminant level (USEPA 2004) is 0.004 mg/L.


1 To source waters

Simazine may enter source waters as a result of its use pre-emergence to control broadleaved and grass weeds in a wide variety of crop, orchard and non-crop areas.

The total annual usage of simazine in New Zealand in the late 1980s was 71,700 kg with the majority of use being in the North Island. The highest usage was in the Tauranga county (11,200 kg).


Forms and fate in the environment Simazine can be degraded through hydrolysis and N-dealkylation. Its half life in soil ranges from 28 to 170 days with a recommended average half life of 60 days. Even though it has low solubility in water (3.5 to 5 mg/L) it can leach to groundwater. Its sorption coefficient is 130 mL/g.

Typical concentrations in drinking-water Simazine has been detected in tile drainage from an orchard in Canterbury and has also been detected at low concentrations in groundwater in South Canterbury.

The P2 Chemical Determinand Identification Programme, sampled from 343 zones, found simazine concentrations to range from �not detectable� (nd) to 0.0002 mg/L, with the median concentration being �nd� (limit of detection = 0.0001 mg/L). Frequently detected in groundwater and surface water at concentrations of up to a few micrograms per litre (WHO 2004).

Removal methods Simazine can be removed from water by a number of methods: activated carbon adsorption (0.0001 mg/L should be achievable using GAC, WHO, 2004); ion exchange; and oxidation by chlorine, chlorine dioxide, ozone, hydrogen peroxide and potassium permanganate. Coagulation, chlorination, and softening processes are relatively ineffective in removing simazine from water.


Referee method



1 Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector (EPA 507).

Health considerations Simazine is absorbed by the gut of rats and mice and distributed to various tissues, with the highest concentrations in the spleen, liver and kidney. USSR workers manufacturing simazine and propazine reported 124 cases of contact dermatitis. The serious cases lasted 7-10 days and involved erythema (reddening of the skin), oedema and a vesiculopapular reaction that sometimes progresses to the formation of bullae (watery blisters). A study showed an association between ovarian tumours and exposure to triazine herbicides, but the number of subjects included in the study was limited. Simazine does not appear to be genotoxic in mammalian systems. Recent studies have shown an increase in mammary tumours in the female rat, but no effects in the mouse. The International Agency for Research on Cancer has classified simazine in Group 3 (not classifiable as to its carcinogenicity to humans).

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV. The no observable adverse effect level used in the derivation is from a rat study for carcinogenicity and long-term toxicity


study based on weight changes, effects on haematological parameters, and an increase in mammary tumours. The MAV for simazine in drinking-water was derived as follows:


where:

• no observable adverse effect level = 0.52 mg/kg body weight per day established for carcinogenicity and long-term toxicity




• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 for possible carcinogenicity).

References Foster DM, White SL. 1991. New treatment processes for pesticides and chlorinated organics control in drinking-water. J IWEM 5: pp 466�[add page number].


Smith VR. 1993. Groundwater contamination by organic chemicals in Canterbury. Canterbury Regional Council Report 93(20).


USEPA. 1989. Drinking Water Health Advisory: Pesticides. United States Environmental Protection Agency, Office of Drinking Water Health Advisors. Lewis Publishers.





WHO. 2003. Simazine in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/42).


2,4,5-T Revised July 2005. (Also called 2,4,5-trichlorophenoxyacetic acid.)

Maximum acceptable value Based on health considerations, the concentration of 2,4,5-T in drinking-water should not exceed 0.01 mg/L.



1 To source waters

Chlorophenoxy herbicides are not frequently found in drinking-water; when detected, concentrations are usually no greater than a few micrograms per litre (WHO 2004).

Available information indicates that 2,4,5-T can be removed effectively from water by adsorption on to granular, or powdered, activated carbon. Approximately 90% removal has been reported for ion exchange, and 63% for removal by coagulation/clarification/filtration.



2,4,5-T may enter source waters as a result of its use for the control of gorse and other brush weeds. The total annual usage of 2,4,5-T in New Zealand has been above 700,000 kg in the past and was averaging around 500,000 kg in the late 1980s. Most of the usage was in the North Island, with about 250,000 kg being used in the Rangitikei county. In 1990 it was voluntarily withdrawn from use. There was additional concern about 2,4,5-T because of impurities of dioxins in the product.

Forms and fate in the environment 2,4,5-T can be made as either amine salts or butyl esters. The esters hydrolyse rapidly into the acid and the salts dissociate in water. The acid undergoes degradation to produce 2,4,5-trichlorophenol. The half life of 2,4,5-T in soil ranges from 12�59 days with a recommended average half life of 24 days. The water solubility of the acid is 150 mg/L; the ester is practically insoluble and the solubility of the amine salts range from 189,000 to 500,000 mg/L.

Typical concentrations in drinking-water Of 230 source water samples obtained from 212 supplies in New Zealand between 1988 and 1992, four samples contained detectable levels of 2,4,5-T. The concentrations ranged from 0.00005 mg/L (0.05 µg/L) to 0.00019 mg/L (0.19 µg/L). 2,4,5-T was detected on one occasion in a well near Gisborne at a concentration of 0.0001 mg/L (0.1 µg/L). The P2 Chemical Determinand Identification Programme, sampled from 296 zones, found 2,4,5-T concentrations to range from �not detectable� (nd) to 0.0002 mg/L, with the median concentration being �nd� (limit of detection = 0.0001 mg/L).

Removal methods


Referee method



Health considerations In general, chlorophenoxy herbicides are absorbed rapidly from the gastro-intestinal tract and distributed evenly throughout the body. Accumulation in human tissues is not expected, and a steady-state level in


the human body will be achieved within 3�5 days of exposure. Elimination occurs primarily in the urine, mostly in the unchanged form. Biological half lives of chlorophenoxy herbicides in mammals range from 10 to 33 hours. Metabolic conversions occur only at high doses. The salt and ester forms are hydrolysed rapidly and follow the same pharmacokinetic pathways as the free acid forms. 2,4,5-T is considered to be moderately acutely toxic. Symptoms of high oral doses include nausea, vomiting, drowsiness, fever, increases in pulse and respiration, shock, coma and death. Animals subject to long-term exposure to high doses of 2,4,5-T experienced symptoms including reduced body weight gain, elevated urinary excretion of porphyrins, heptacellular swelling, paleness and increased relative liver and kidney weights.

Results of various reproductive studies indicate that 2,4,5-T, without appreciable dioxin contamination, caused teratogenic effects (cleft palate and kidney malformations) in only mice at doses above 20 mg/kg body weight.

Chlorophenoxy herbicides as a group, including 2,4-D and MCPA, have been classified by the International Agency for Research on Cancer in Group 2B (possibly carcinogenic to humans). However, based on the available data from studies on exposed populations and on animals, it is not possible to assess the carcinogenic potential of any specific chlorophenoxy herbicide. Therefore, drinking-water guidelines for these compounds are based on a threshold approach for other toxic effects.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for 2,4,5-T in drinking-water. The calculation was based on reduced body weight gain, increased liver and kidney weights, and renal toxicity in a two-year study in rats. The MAV for 2,4,5-T in drinking-water was derived as follows:


where:

• no observable adverse effect level = 3 mg/kg body weight per day based on reduced body weight gain, increased liver and kidney weights and renal toxicity in a two-year study in rats. The NOAEL was the same for reproductive effects (reduced neonatal survival, decreased fertility, reduced relative liver weights and thymus weights in litters) of dioxin-free (<0.03 mg/kg) 2,4,5-T in a three-generation reproduction study in rats




• uncertainty factor = 1000 (100 for inter and intra-species variation and 10 for the suggested association between 2,4,5-T and soft tissue sarcoma and non-Hodgkin lymphoma in epidemiological studies).

References



Close ME. 1993. Assessment of pesticide contamination of groundwater in New Zealand, 2: results of groundwater sampling. New Zealand Journal of Marine and Freshwater Research 27: 267�73.


USEPA. 1989. Drinking Water Health Advisory: Pesticides. United States Environmental Protection Agency, Office of Drinking Water Health Advisors. Lewis Publishers.








Based on health considerations, the concentration of terbacil in drinking-water should not exceed 0.05 mg/L. WHO (2004) does not mention terbacil.

1 To source waters

Removal methods

Terbacil



Forms and fate in the environment Terbacil is a selective herbicide usually used for control of both annual weeds and perennial grasses. It is sprayed on soil surfaces preferably just before, or otherwise during, the period of active weed growth. Terbacil works in plants by interfering with photosynthesis. It is part of a family of chemicals called substituted uracils. As at August 2005, terbacil is registered for use in New Zealand.

In most soil types, terbacil has a relatively low tendency to be adsorbed to soil particles (Koc = 55 g/mL). It is highly soluble in water and highly persistent in soils. Soil half lives of 120 days or two to five months have been reported. This information indicates that terbacil is likely to be moderately mobile in soil and potentially can pollute groundwater so should not be used on sandy or gravelly soils. In moist soils, terbacil is subject to microbial degradation.

Terbacil is stable to hydrolysis and photo-degrades slowly in water. The water solubility of terbacil is 710 mg/L and the log K is 1.89 (Toxnet). ow




Analytical methods

Referee method

[???]


[???]

The USEPA has established a Lifetime Health Advisory (LHA) level of 90 micrograms per litre (0.09 mg/L) for terbacil in drinking-water. This means that EPA believes that water containing terbacil at or below this level is acceptable for drinking every day over the course of one�s lifetime, and does not pose any health concerns.

Health considerations Terbacil has low acute toxicity in humans and other animals. In general, the uracil herbicides are rapidly excreted in urine by mammals. This may account for their reportedly low toxicity. The oral LD50 of terbacil for rats is 5000 to 7500 mg/kg. Signs of acute terbacil poisoning in rats include weight loss, paleness, lack of movement and rapid respiration. Six out of six male rats survived ten daily doses of 1000 mg/kg. No evidence of toxicity was seen in two-year feeding studies of rats fed doses as high as 12.5 mg/kg or in dogs fed doses as high as 6.25 mg/kg of terbacil. At 125 to 500 mg/kg there was a lower rate of weight gain, liver enlargement and other liver changes in rats. The high dose produced a slight increase in liver weight in dogs. Terbacil is not mutagenic and has given � no evidence of carcinogenicity.

Derivation of maximum acceptable value The MAV for terbacil in drinking-water was derived as follows:


where:

• no observable adverse effect level = 1.25 mg/kg-day based on the absence of increase in thyroid/body weight ratio, a slight increase in liver weights, and an elevated alkaline phosphatase level, in a two-year dog feeding study





References TOXNET. 2004. TOXNET Databases Summary. US National Library of Medicine, National Institutes of Health, Department of Health and Human Services.

USEPA. 1989. Pesticide Fact Sheet Number 206: Terbacil. Washington DC: US Environmental Protection Agency, Office of Pesticides and Toxic Substances, Office of Pesticide Programs.

USEPA. 1989. Health Advisory Summary: Tebacil. Washington DC: US Environmental Protection Agency.

USEPA. 2005. Integrated Risk Information System. US Environmental Protection Agency, the Office of Research and Development, and the National Center for Environmental Assessment.


Terbuthylazine Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of terbuthylazine in drinking-water should not exceed 0.008 mg/L.


1 To source waters

Terbuthylazine, a herbicide that belongs to the chlorotriazine family, may enter source waters as a result of its application as a selective herbicide for the control of grass and broadleaf weeds in forestry and various crops and orchards. The total annual usage of terbuthylazine in the late 1980s was 39,000 kg. Terbuthylazine is registered for use in New Zealand as at August 2005.

Forms and fate in the environment In soil, microbial degradation occurs via dealkylation of the side chain, hydroxylation resulting from hydrolysis of the chlorine atoms and of the dealkylated amino group, and ring clevage.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, found terbuthylazine concentrations to range from �not detectable� (nd) to 0.002 mg/L, with the median concentration being �nd� (limit of detection = 0.0002 mg/L). Concentrations in water seldom exceed 0.0002 mg/L, although higher concentrations have been observed (WHO 2004).



Referee method

No referee method has been given for terbuthylazine because no method meets the required criteria.


1. Liquid/Liquid Extraction and Gas Chromatography with a Nitrogen Phosphorus Detector (Chlorophenoxy acidic herbicides, trichlorobenzoic acid, chlorophenols, triazines and glyphosate in water 1985; HMSO 1986).

Health considerations There is no evidence that terbuthylazine is carcinogenic or mutagenic. In long-term dietary studies in rats, effects on red blood cell parameters in females, an increased incidence of non-neoplastic lesions in the liver, lung, thyroid and testis and a slight decrease in body weight gain, were observed.


Derivation of maximum acceptable value The MAV for terbuthylazine was calculated as follows:

0.22 mg/kg x 70 kg x 0.1 = 0.0077 mg/L (rounded to 0.008 mg/L) 2 L x 100

where:

• no observable adverse effect level = 0.22 mg/kg body weight for decreased body weight gain at the next higher dose in a two-year toxicity/carcinogenicity study in rats


• proportion of acceptable daily intake allocated to drinking-water = 0.1

• average quantity of water consumed by an adult = 2 L/day

• uncertainty factor = 100 (for inter- and intra-species variation).

References JMPR. 1993. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment, and the WHO Expert Group on Pesticide Residues. Geneva, 20�29 September. FAO Plant Production and Protection Paper No 122 (Annex 1).

The Royal Society of Chemistry. 1987. The Agrochemicals Handbook (2nd ed). Surrey, England: Unwin Brothers Ltd.

WHO. 2003. Terbuthylazine in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/63).


Thiabendazole New entry August 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of thiabendazole in drinking-water should not exceed 0.4 mg/L (400 µg/L). WHO (2004) does not mention thiabendazole.


1 To source waters

Thiabendazole is used as a systemic fungicide for spoilage control of citrus fruit; for treatment and prevention of Dutch elm disease in trees; for control of fungal diseases of seed potatoes. It is also used therapeutically for cats as an anthelmintic (nematodes). It is registered for use in New Zealand and is available in a variety of formulations, some of which contain other active ingredients (fosetyl-aluminium, thiram, metalaxyl). Trade names are: Aliette Super, Apron Combi, Apron TZ, Tecto SC, Tecto. No information is available on the annual usage of specific active ingredients in New Zealand, although thiabendazole is understood to be likely to constitute only minor use in the agricultural sector (P Holland, personal communication). Thiabendazole is used for veterinary and human medical purposes. Anthelmintics are medicines used in the treatment of worm infections. In this usage, thiabendazole is listed as a medicine notified as having been supplied under Section 29 of the Medicines Act. Section 29 of the Medicines Act 1981 permits the supply of an unapproved medicine to a medical practitioner for use by a named patient.


Forms and fate in the environment Thiabendazole is practically insoluble in water (RSocC 1987). Thiabendazole binding in soil increases with increasing soil acidity. It is quite persistent. In one study, nine months following application, most of the residues (85�95%) were recovered from soil. It is not expected to leach readily from soil (Extoxnet 2001). There is no information available regarding the greatest source of exposure to thiabendazole for New Zealanders (eg, dermal contact, inhalation, diet: food, water), although dietary intake would be expected to be important because of thiabendazole�s use to protect fruit from fungal damage.

Typical concentrations in drinking-water No Ministry of Health drinking-water surveys have included thiabendazole, so typical concentrations in New Zealand drinking-waters are unknown. No information on concentrations in international drinking-waters are available.

Removal methods No information is available on the removal of thiabendazole from water. However, since it binds strongly to soil, water treatment processes that remove particulate matter should remove a lot of the thiabendazole. Trace organic substances can be expected to adsorb on to activated carbon to some extent, and therefore activated carbon is likely to achieve some removal of thiabendazole, although a guide to the efficiency of the process cannot be provided. Nanofiltration and reverse osmosis may also provide a means of removing this compound from water, but no data are available to support this.


Referee method

High pressure liquid chromatography � fluorescence (EPA 641).


None.

Health considerations Excretion of thiabendazole in the urine and faeces is rapid in most species and is almost complete after 48 hours in rats and 96 hours in sheep. The excretion products are metabolites reaching peak levels in the blood stream one hour after administration to rats, one to two hours in humans, and four hours after administration in cattle. These metabolites are distributed throughout most body tissues in sheep but detectable in only a few tissues at low levels (less than 0.2 ppm) in 16 days and at very low levels (0.06 ppm or less) after 30 days. No evidence of bioaccumulation in animal tissues was found (Extoxnet 2001).

Acute poisoning

The acute oral LD50 for mice is 3810 mg/kg. For rats it is 3330 mg/kg and rabbits 3850 mg/kg (RSocC 1987). These levels suggest a moderate to low acute oral toxicity when compared with other pesticides.


Effects of acute over-exposure to the fungicide include dizziness, anorexia, nausea and vomiting. Other symptoms such as itching, rash, chills and headache occur less frequently. The symptoms are brief and are related to the dose level (Extoxnet 2001).

Chronic exposure

Dogs autopsied after a two-year feeding study showed incomplete development of bone marrow, a wasting away of lymph tissue, and other abnormalities. Most dogs tested at around 100 mg/day for two years developed anemia. The dogs recovered at the end of the study (Extoxnet 2001). The International Agency for Research on Cancer has not classified thiabendazole for its ability to cause cancer.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for thiabendazole in drinking-water, as follows:



References Extoxnet. 2001. Pesticide Information Profile: Thiabendazole. New York: PMEP, Cornell University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

Royal Society of Chemistry. 1987. The Agrichemicals Handbook: Thiabendazole (2nd ed). Surrey, England: Royal Society of Chemistry, Unwin Brothers Ltd.

Triclopyr Revised July 2005.

Maximum acceptable value (provisional) Based on health considerations, the concentration of triclopyr in drinking-water should not exceed 0.1 mg/L. WHO (2004) does not mention triclopyr.


1 To source waters

Triclopyr, a pyridine, may enter source waters as a result of its application as a herbicide, used to control woody plants and many broadleaved weeds in mainly non-crop areas. Unlike a similar product, 2,4,5-T, there is no possibility of dioxin impurities occurring in triclopyr. Some trade names for herbicides containing this product are Garlon, Turflon, Access, Redeem, Crossbow, Grazon and ET. The herbicide may be mixed with picloram or with 2,4-D to extend its utility range.


The total annual usage of triclopyr in New Zealand in the late 1980s was 27,500 kg with the majority of use being in the South Island. As at August 2005, triclopyr is registered for use in New Zealand.


No known sources.


No known sources.

Forms and fate in the environment Triclopyr undergoes fairly rapid biodegradation in soil with an average half life of 46 days. Water solubility is 440 mg/L and the sorption coefficient is 20 mL/g. In natural soil and in aquatic environments, two of the formulations convert rapidly to the acid which in turn is neutralised to a salt. Triclopyr is not adsorbed strongly to soil particles, has the potential to be mobile, and is degraded fairly rapidly by soil microorganisms. Concentrations of 500 ppm had no apparent effects on the growth of common soil microorganisms. The half life in soil is from 30 to 90 days, depending on soil type and environmental conditions, with an average of about 46 days. The half life of one of the breakdown products (trichloro-pyridinol) in 15 soils ranged from 8�279 days with 12 of the tested soils having half-lives of less than 90 days. Longer half lives occur in cold or arid conditions. Breakdown by the action of sunlight is the major means of triclopyr degradation in water. The half life is 10 hours at 25°C. The major metabolite is trichloropyridinol. Triclopyr is translocated readily throughout a plant after being taken up by either roots or the foliage. Cowberries with residues of 2.4 ppm at six days had 0.7�1.1 ppm at 30�36 days, and 0.�-0.3 ppm in 92�98 days. The estimated half life in aboveground drying foliage as in a forest overstory is two to three months.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 296 zones, found triclopyr concentrations to range from �not detectable� (nd) to 0.0006 mg/L, with the median concentration being �nd� (limit of detection = 0.0001 mg/L). Triclopyr was tested but not found in a host of groundwater sites throughout the country (Williams et al 1998).

Removal methods No information on methods of removing triclopyr from water is available.


Referee method

No referee method has been given for triclopyr because no method meets the required criteria.




Health considerations When rats were dosed intravenously at 5 mg/kg, most of the dose was excreted in urine. At 100 mg/kg urinary excretion still predominated. At higher doses, an increasing amount was in the faeces. In dogs, 0.5 mg/kg of triclopyr had a half life of 14 hours for clearance from blood plasma, and a dose of 20 mg/kg had a half life of 95 hours reflecting the unique capacity for excretion of organic acids by the dog. Excreted triclopyr is mostly the parent compound but small quantities of breakdown products are also present. Triclopyr was found in greater quantities in the liver and fatty tissue of the rat when compared with the blood plasma. The dog had higher levels in the kidney than in the blood plasma, and in monkeys, residues in all tissues were the same as in blood plasma. The compound is not expected to concentrate to any significant degree in the tissues of animals.

Acute poisoning

The oral LD50 of triclopyr in rats ranges from 630 to 729 mg/kg and from 2000 to 3000 mg/kg for various formulated products. Similar differences were noted for skin toxicity in the rabbit. The LD50 for the technical material was greater than 2000 mg/kg and greater than 4000 mg/kg for the formulations. Inhalation of triclopyr (technical) did not affect rats but inhalation of some of the formulations did cause nasal irritations. A similar result was seen when rabbit eyes were exposed. The technical material had only a slight effect on rabbit eyes and the undiluted formulated material caused significant eye irritation. Other oral LD50 values for triclopyr are 550 mg/kg in the rabbit and 310 mg/kg in the guinea pig.

Triclopyr is slightly toxic to mallard ducks. When fed the compound, the LD50 was 1698 mg/kg. Bobwhite quail and Japanese quail fed for eight days had LC50s of 2935 ppm and 3278 ppm, respectively.

The compound is practically non-toxic to fish. Triclopyr has a LC50 of 117 ppm for rainbow trout and a 96-hour LC50 of 148 ppm for bluegill sunfish. The compound is practically non-toxic to the aquatic invertebrate Daphnia magna, a water flea (LC50 for the triclopyr salt of 1170 ppm). The compound is non-toxic to bees.

Chronic exposure

Rats fed diets containing between 3 and 30 mg/kg/day of triclopyr experienced no ill effects. Males fed much higher doses (100 mg/kg) had decreased liver and body weight and increased kidney weight. The male mice were also sensitive at moderate doses. They had reduced liver weight at 60 mg/kg/day. Monkeys fed small amounts of triclopyr (30 mg/kg/day) had no adverse effects. Triclopyr fed to rabbits daily at low to moderate doses (25 to 100 mg/kg) caused some maternal toxicity and death but not foetal toxicity or birth defects. The maternal mortality was inconsistent with other studies. There is not enough data to draw any conclusion about the reproductive hazards of triclopyr due to chronic exposure in humans. Pregnant rats given moderate doses (up to 200 mg/kg/day) on days 6�15 of gestation had offspring with mild fetotoxicity, but no birth defects. There were no teratogenic effects in rabbits treated in a similar manner at 10 or 20 mg/kg/day. The evidence suggests that the human risk of birth defects is fairly low due to chronic exposure to triclopyr.


Both bacteria and isolated cells did not mutate in response to the presence of triclopyr. Another mutagen study using rats was weakly positive, but negative in mice, the more sensitive species. There were no chromosome changes noted in rat bone marrow. Triclopyr is not considered to be mutagenic. Rats and mice fed low levels (3 to 30 mg/kg/day) of triclopyr for two years showed no carcinogenic response. Even though the mice did have a high incidence of lymph cancer, this incidence was apparently characteristic of the particular strain of mice and did not represent a dose-related effect.

Derivation of maximum acceptable value The MAV for triclopyr in drinking-water was derived using a tolerable daily intake approach as follows:


where:

• no observable adverse effect level = 3 mg/kg body weight per day from a two year feeding study in rats




• uncertainty factor = 100 (for intra and inter-species variation)


Extoxnet. 2001. Pesticide Information Profile: Thiabendazole. New York: PMEP, Cornell University. Available at: http://pmep.cce.cornell.edu/profiles/extoxnet/

Tsukioka T, Takeshita R, Murakami T. 1986. Gas chromatographic determination of triclopyr in environmental waters. Analyst 111: 145�9.




Williams WM, Holden PW, Parsons DP, et al. 1988. Pesticides in Ground Water Data Base: Interim report. United States Environmental Protection Agency, Office of Pesticide Programs.

World Health Organization. 1993. Guidelines for Drinking-water Quality (2nd ed). Volume 1: Recommendations.


Trifluralin Revised July 2005.

Maximum acceptable value Based on health considerations, the concentration of trifluran in drinking-water should not exceed 0.03 mg/L. WHO (2004) states that some impure technical grades of trifluralin could contain potent carcinogenic compounds and therefore should not be used.


1 To source waters

Trifluralin may enter source waters as a result of its use as a pre-emergence herbicide, for the control of annual grasses and broadleaved weeds in a variety of crops and orchards. It has low water solubility and a high affinity for soil. However, biodegradation and photodegradation processes may give rise to polar metabolites that may contaminate drinking-water sources. The total annual usage of trifluralin in New Zealand in the late 1980s was 7800 kg with the majority of use being in the South Island. The highest usage was in the Ashburton county (3300 kg). As at August 2005, trifluralin is registered for use in New Zealand.

Forms and fate in the environment Trifluralin has a low water solubility (0.3 mg/L) with high soil affinity and is relatively immobile. It undergoes photodecomposition, volatilisation, and biodegradation with soil half lives ranging from 20 to 132 days. The recommended average soil half life is 60 days. The sorption coefficient is 8000 mL/g.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme, sampled from 343 zones, did not find trifluralin at detectable concentrations (limit of detection = 0.0002 mg/L). Not detected in the small number of drinking-water samples analysed; has been detected in surface water at concentrations above 0.0005 mg/L and rarely in groundwater (WHO 2004).

Removal methods Granular activated carbon (0.001 mg/L should be achievable, WHO 2004), conventional treatment (alum coagulation, sedimentation, filtration), and possibly air stripping, can remove trifluralin from water.


Referee method



No alternative methods have been recommended for trifluralin because no methods meet the required criteria.


Health considerations Oral doses of trifluralin are not absorbed readily by the gastro-intestinal tract of the rat. The absorbed fraction of trifluralin is metabolised extensively and trifluralin and its metabolites are excreted principally in faeces. Higher levels of trifluralin were found in fat than in the liver. Toxic effects in animals associated with long-term exposure to trifluralin include slightly increased mean body weight gain, slight change in plasma lipids, and a statistically significant increase in liver weight. In the USA, use of trifluralin was associated with an increased risk for non-Hodgkin�s lymphoma. A study of ovarian cancer in Italy did not suggest an association with trifluralin exposure. Both results were based on small numbers of exposed subjects. A larger US study showed no association with occurrence of cancer. The International Agency for Research on Cancer has evaluated technical-grade trifluralin and classed it in Group 3 (not classifiable as to its carcinogenicity to humans). A number of long-term carcinogenicity/toxicity studies with pure (>99%) test material have not demonstrated evidence of carcinogenicity. Trifluralin of high purity does not possess mutagenic properties. Technical trifluralin of low purity may contain nitroso contaminants and has been found to be mutagenic.

Derivation of maximum acceptable value A tolerable daily intake approach has been used for the derivation of the MAV for trifluralin in drinking-water. The no observable adverse effect level used in the derivation was established for mild hepatic effects from a one-year feeding study in dogs. The MAV for trifluralin in drinking-water was derived as follows:


where:

• no observable adverse effect level = 0.75 mg/kg body weight per day for mild hepatic effects in a one-year feeding study in dogs











WHO. 2003. Trifluralin in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/43).



1080 Revised July 2005. (Also called sodium fluoroacetate.)

Maximum acceptable value (provisional) Based on health considerations, the concentration of 1080 in drinking-water should not exceed 0.0035 mg/L. 1080 is not mentioned in WHO 2004.


1 To source waters

1080 is a highly toxic poison used for the control of possums and rabbits. It may enter sources waters as a result of aerial application. As at August 2005, 1080 is registered for use in New Zealand; it is a Class B, restricted use poison, and a license is required for its use.

Forms and fate in the environment 1080 biodegrades in soil and water and it is translocated from water to plants. It is absorbed by aquatic organisms but it is not bioaccumulated.

Typical concentrations in drinking-water The P2 Chemical Determinand Identification Programme did not find 1080 at detectable concentrations (limit of detection = 0.0001 ng/L).

Removal methods No information is available about methods for removing 1080 from water.


Referee method

Derivatisation with dicyclohexylcarbodiimide and gas chromatography with electron capture detection (Ozawa and Tsukioka 1987).


No alternative methods have been recommended for 1080 because no methods meet the required criteria.

Health considerations Animal studies have shown that 1080 is absorbed rapidly and excreted as unchanged fluoroacetate and a range of metabolites. 1080 poisoning results from the transformation of fluoroacetate into fluorocitrate within cell mitochondria. Acute poisoning is characterised by a symptom-free latent period of half to two hours or longer between ingestion and onset of symptoms (nausea, vomiting, diarrhoea and hyperactive behaviour leading to convulsions, coma and cyanosis). Ventricular fibrillation is noted commonly and is the


primary cause of death. Early symptoms include alteration of heart sounds and premature, weak contractions.

Derivation of maximum acceptable value The MAV for 1080 was calculated by the New Zealand Ministry of Health using an NOAEL derived from a Department of Conservation teratology study of rats (Eason, 1999) as follows:

0.1 mg/kg x 70 kg x 0.5 = 0.0035 mg/L 2 L x 500

where: • no observable effect level = 0.1 mg /kg body weight per day • average weight of adult = 70 kg • average quantity of water consumed by an adult = 2 L per day • proportion of lowest lethal dose allocated to drinking-water = 0.5 • uncertainty factor = 500 (10 for intraspecies variation, 10 for inter-species variation, 5 for the

inadequacy of the studies and database).

References Eason CT, Gooneratne R, Fitzgerald H, et al. 1994. Persistence of sodium monofluoroacetate in livestock animals and risk to humans. Human and Experimental Toxicology 13(2): 119�22.

Eason CT, et al. 1999. A review of recent regulatory and environmental toxicology studies on 1080: results and implications. New Zealand Journal of Ecology 23(2): 129�37.

Ozawa H, Tsukioka T. 1987. Gas chromatographic determination of sodium monofluoroacetate in water by derivatisation with dicyclohexylcardodiimide in water. Analytical Chemistry 59: 2914�17.


2.4 Aesthetic determinands aggressiveness (plumbosolvency) aluminium ammonia calcium chloride chlorine 2-chlorophenol colour copper 1,2-dichlorobenzene 1,4-dichlorobenzene 2,4-dichlorophenol ethylbenzene hardness hydrogen sulphide iron magnesium manganese monochloramine odour pH sodium styrene sulphate taste temperature toluene total dissolved solids 1,2,3-trichlorobenzene 1,2,4-trichlorobenzene 1,2,5-trichlorobenzene 2,4,6-trichlorophenol turbidity xylenes zinc

Aggressiveness Updated August 2005. (Also called plumbosolvency.)

Description and characteristics The aggressiveness of a water is estimated by various empirical indices which should not be considered as absolutes. These indices are guides to the behaviour of calcium carbonate in aqueous systems. They should be supplemented, where possible, with experimentally derived information. Neither the calculations referred to here, nor the most complex computerised calculations, adequately describe all corrosion events that establish �aggressiveness�.

Langelier Saturation Index (LSI) The Langelier Saturation Index is used to evaluate the calcium carbonate (CaCO3) scale-forming and scale-dissolving tendencies of water. It provides no information about the rate, or extent, of precipitation


or dissolution of calcium carbonate. This distinction must be appreciated, because a water may have a tendency to precipitate calcium carbonate, but if the concentration of calcium in the water is insufficient, little solid will form. Nevertheless, assessing these tendencies is useful in corrosion control programmes and in preventing CaCO3 scaling in pipes and equipment such as industrial heat exchangers or domestic water heaters. The LSI is based on the assumption that a scale coating protects the pipe; it is not a direct measure of how a water will react with metal pipework or fittings. The LSI calculation produces a number that may range from a negative to a positive value. A value of zero indicates that the water is in equilibrium with any calcium carbonate solid present, that is, it is saturated and will neither dissolve nor precipitate calcium carbonate. A negative LSI indicates undersaturation, and the tendency for calcium carbonate dissolution. Waters, such as rainwater, containing very little calcium and alkalinity and having very low pH may have an LSI less than -5, but values between -2 and -3 are more common for aggressive reticulated waters. A positive LSI shows oversaturation, and the tendency to precipitate calcium carbonate. Whether the LSI calculated for a water is positive or negative, it must be remembered that the index is only an approximation. The uncertainty in the interpretation of the index increases as the water approaches saturation (zero index value). Near zero, a water with a negative index may be capable of precipitating calcite, and vice versa. The value must be significantly positive or negative before it is possible to be reasonably certain of the precipitating, or dissolving properties of the water. The calcium carbonate saturation index (LSI) is calculated from the calcium, pH, temperature, dissolved solids or conductivity, and alkalinity characteristics of a water.

Guideline value Based on the formation of a calcium carbonate scale acting as a barrier to corrosion, a guideline value of LSI > 0 is assigned.

Typical concentrations in drinking-water Typical values for aggressiveness in New Zealand drinking waters range from LSI +1.5 to -3.0 with most waters being greater than -1.5 and less than 0.

Aggressiveness modification Reduction of scale formation: chemical softening, reverse osmosis, electrodialysis, or ion exchange, will reduce calcium, and thus decrease the LSI and increase the water�s aggressiveness.

Minimisation of aggressiveness Addition of calcium ions; and pH and alkalinity adjustment using combinations of lime, caustic soda, soda ash, sulphuric acid, and carbon dioxide; can increase the LSI and hence decrease the water�s aggressiveness.

Analytical methods The saturation index can be obtained from the following formula: LSI = pH � pHs, where pHs is the pH at which a water is saturated with CaCO3, and is calculated from published nomographs or software using values for temperature, total dissolved solids, calcium and alkalinity.


Health considerations Aggressive waters have the potential to cause significant health effects depending on the nature of the materials used in the distribution system. Corrosion of pipes may lead to heavy metals such as copper, zinc, lead and cadmium being present in the distribution system. Where asbestos-cement pipes are used, corrosion by aggressive water may also release asbestos into the water.



Aluminium (Al3+) Updated August 2005.

Description and characteristics Aluminium is the third most abundant element in the earth�s crust occurring in minerals, rocks and clays. This wide distribution accounts for the presence of aluminium in nearly all natural water as a soluble salt, a colloid, or an insoluble compound. It may be present in water as a fine suspension through natural leaching from soil, clay and rock, or alumino-silicates. Aluminium is used in many industrial and domestic products including antacids, anti-perspirants, food additives and vaccines. It is used commonly by the food industry for food containers and packaging, and in cooking utensils. Aluminium salts are used extensively in water treatment as coagulants for the removal of colour and turbidity. Where aluminium coagulants are used, post-treatment plant flocculation effects can occur precipitating aluminium in the reticulation where it can later be resuspended or redissolved. A whitish gelatinous precipitate of aluminium hydrolysis products may form in the distribution system, which could result in consumer complaints about �milky coloured� water. More commonly though, it attracts iron and manganese, causing dirty (brown) water. It has been estimated that the intake of aluminium from food and beverages is approximately 5�20 mg/day. Drinking-water probably contributes less than 5% of the total dietary intake, although aluminium in water may be more bio-available than aluminium from other sources.

Guideline value Based on aesthetic problems caused by post-treatment plant precipitation, the concentration of aluminium in drinking water should not exceed 0.10 mg/L. However, there are factors other than aluminium concentration that contribute to its solubility properties, such as the presence of natural organic matter, the pH of the water, and the presence of coagulation/flocculation aids. Although data is insufficient to set a guideline value based on health considerations, there has been some concern over the possible health effects of aluminium, particularly on the possible causal link between aluminium concentration and Alzheimer�s disease, and on the bio-availability of aluminium. Water authorities are strongly encouraged to keep aluminium concentrations as low as possible.


Typical concentrations in drinking-water Values of aluminium commonly found in New Zealand drinking-waters range from 0.01 to 0.3 mg/L. However, the 1983�1989 Surveillance Data Review indicated that for all reticulated water samples 28% contained above 0.05 mg/L aluminium, including 10% above 0.2 mg/L. For aluminium-treated supplies, 81% contained greater than 0.05 mg/L aluminium, including 36% above 0.2 mg/L.

Removal methods Naturally-occurring aluminium associated with inorganic and organic particulate matter, and soluble natural organic matter, can be reduced using coagulation, flocculation and filtration. Aluminium carry-over from treatment can be reduced by optimisation of the coagulation process.

Analytical methods Aluminium is measured by atomic absorption spectroscopy, inductively coupled plasma emission spectroscopy, or colorimetric procedures. Some colorimetric techniques have a limit of detection below 0.05 mg/L, and graphite furnace AAS has a detection limit of approximately 0.005 mg/L. Field test kits are available for aluminium, and are almost exclusively based on colorimetric methods. Colorimetric procedures can be carried out in a treatment plant laboratory, by following the manufacturer�s instructions or standard method procedures carefully, and taking the necessary precautions against contamination. Field kit and treatment plant analyses for aluminium are likely to be used for process control.

Aluminium concentrations are expressed in many terms (eg, total, soluble, dissolved, extractable). The term �soluble� means truly soluble, not dissolved which is �filterable through a 0.45 micron filter�. Some finely suspended alumino-silicate clay particles can pass through a 0.45 micron filter but are not truly soluble. Studies have proven that different measurement techniques can be more suited to specific aluminium fractions. Analysis for aluminium fractions is only necessary if the total aluminium concentration is greater than guideline values and remedial action is being investigated.

Health considerations The metabolism of aluminium in humans is poorly understood. It has been suggested that absorption of aluminium is greater in the presence of organic ligands such as citrate. Studies indicate that probably less than 1% of dietary aluminium is absorbed by the gastro-intestinal tract and passes into the blood stream. Some aluminium accumulates in bone tissue but most is removed by the kidneys. In healthy adults the total accumulated body load of aluminium has been estimated at about 35 mg. Investigations have established a correlation between the concentration of aluminium in water used to prepare kidney dialysis fluid and the incidence of dialysis dementia. The condition is treatable using chelation therapy. Aluminium has also been linked to other conditions associated with the use of dialysis units including osteomalacia (a softening of the bones), and anaemia. Reverse osmosis or de-ionisation units are now used to treat dialysis water before use, and aluminium concentrations are kept below 0.01 mg/L. Elevated concentrations of aluminium have been found in the brains of Alzheimer�s patients. Studies indicate that a tentative link may exist, but there is no conclusive evidence that aluminium directly causes Alzheimer�s disease. WHO stated in 1998 and repeated in their 2004 Guidelines:

On the whole, the positive relationship between aluminium in drinking-water and Alzheimer disease, which was demonstrated in several epidemiological studies, cannot be totally dismissed. However, strong reservations about inferring a causal relationship are warranted in view of the failure of these studies to account for demonstrated confounding factors and for total aluminium intake from all sources.

Taken together, the relative risks for Alzheimer disease from exposure to aluminium in drinking-water above 0.1 mg/L, as determined in these studies, are low (less than 2.0). But, because the risk


estimates are imprecise for a variety of methodological reasons, a population-attributable risk cannot be calculated with precision. Such imprecise predictions may, however, be useful in making decisions about the need to control exposures to aluminium in the general population.

Aluminium has been associated with two severe neurodegenerative diseases, namely Parkinsonism dementia (PD) and Amyotrophic lateral sclerosis (ALS), both of which occur with a high incidence in areas where aluminium is naturally present in food and drinking water. There was an appreciable decrease in the incidence of these conditions when the areas became developed with associated changes in dietary habits, importing of food, and improvements to the water supply. The Factsheet in the Australian Drinking Water Guidelines was revised in 2001. It states that water authorities are strongly encouraged to keep acid-soluble aluminium concentrations as low as possible, preferably below 0.1 mg/L. No health-based guideline is set for aluminium at this time, but this issue will be kept under review.



AWWA. 1990. Water Quality and Treatment: A Handbook of Community Water Supplies (4th ed). McGraw Hill Inc.

Garruto RM, Yanagihara R, et al. 1990. Models of environmentally induced neurological disease; epidemiology and etiology of amyotrophic lateral sclerosis and parkinsonism-dementia in the Western Pacific. Environmental Geochemistry and Health 12: 137�51.

Garruto RM, Yase Y. 1986. Neurodegenerative disorders of the Western Pacific: the search for mechanisms of pathogenesis. Trends in Neurosciences 9: 368�74.

Gregor JE. 1992. Aluminium Interlaboratory Comparative Study: Analytical workshop. New Zealand Water Supply and Disposal Association Annual Conference.

Martyn CN, Osmond C, Edwardson JA, et al. 1989. Geographical relation between Alzheimer�s disease and aluminium in drinking water. Lancet (January): 59�62.

Mattingley BI. 1992. New Zealand Drinking Water Surveillance Programme Data Review 1983�89.

Perl DP, Brody AR. 1980. Alzheimer�s disease: X-ray spectrometric evidence of aluminium accumulation in neurofibrillar tangle bearing neurons. Science 208: 297�9.

Waters JD. 1988 . Aluminium: Current toxicity issues. Proceedings New Zealand Trace Elements Group Conference. Lincoln College, Canterbury.

WHO. 2003. Aluminium in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/53).


Ammonia (NH3 and NH4+)

Updated August 2005.

Description and characteristics Ammonia may be found in natural surface waters, but is more frequently found at elevated concentrations in anaerobic groundwaters. At the pH of most natural waters, ammonia (NH3) dissolves rapidly in water to form an equilibrium with the ammonium cation (NH4

+).


Ammonia is used commercially in animal feeds and fertilisers, and in the manufacture of fibres, plastics and explosives. Ammonia products are widely used as cleaning agents and food additives. Ammonia can be an important indicator of pollution as it can be formed as an intermediate product in the breakdown of nitrogen-containing organic compounds, or of urea from human or animal excrement. Ammonia may increase the solubility of some metals such as copper and lead, thus interfering in the formation of passivating films and increasing corrosion rates. It is also a food source for some micro-organisms, and can support nuisance growths of bacteria and algae, often with a subsequent increase in the nitrite and nitrate concentrations. It may be present in unchlorinated drinking-water due to contamination of source water or through microbial metabolism. Ammonia reacts rapidly with chlorine. During breakpoint chlorination most of the ammonia is oxidised to nitrogen and nitrate. However, if the chlorine dose is insufficient, chloramines may be formed.

Analytical methods

Most uncontaminated source waters have ammonia concentrations below 0.01 mg/L. High concentrations (greater than 10 mg/L) have been reported where water is contaminated with animal waste. Many groundwaters contain 0.1�1.0 mg/L of ammonia as N. The odour threshold of ammonia in water is 1.5 mg/L; the threshold for taste is 35 mg/L. Chloramines have an unpleasant biting odour at concentrations in water above 0.2 mg/L. To restrict chloramines forming in the distribution system by the reaction of FAC and ammonia, the ammonia content should be less than 0.3 mg/L. Food can contain substantial amounts of ammonia/ammonium ion, and is the principal source of intake for humans.

Guideline value Based on aesthetic considerations (odour threshold), the concentration of ammonia (measured as ammonium ion) should not exceed 1.5 mg/L. WHO (2004) states that ammonia in drinking-water is not of immediate health relevance, and therefore no health-based guideline value is proposed. However, ammonia can compromise disinfection efficiency, result in nitrite formation in distribution systems, cause the failure of filters for the removal of manganese and cause taste and odour problems.

Typical concentrations in drinking-water Values of ammonia commonly found in New Zealand drinking waters range up to approximately 2 mg/L. However, most waters contain less than 0.05 mg/L of ammonia. Ammonia is generally not present in groundwater unless it is contaminated with sewage or animal excrement. Ammonia in deep, secure groundwaters should not be a concern.

Removal methods Ammonia concentrations in drinking-water supplies can be reduced by chemical or biological oxidation of ammonia to nitrate.

The concentration of ammonia in water can be determined by a number of methods including colorimetric (phenate), titrimetric, ion chromatography, and ion selective electrode techniques. Preliminary distillation may be required for high concentrations and where interferences are present. The limit of detection of ammonia for most methods is 0.01 mg/L.

Health considerations Ammonia is an important metabolite in humans and animals. It is formed in the liver by the de-amination of amino acids, and in the gastro-intestinal tract by the breakdown of food by enzymes and bacterial flora. Almost all ammonia is absorbed, then transported to the liver and used mostly in the urea cycle. Only an


extremely small proportion of the ammonia absorbed in the intestinal tract originates directly from food or water. Ammonia has a toxic effect on humans only if the intake becomes higher than the detoxification capacity of the body. At doses above 32 mg of ammonium ion per kilogram body weight per day (over 1000 mg/L in drinking-water!) ammonium chloride influences the metabolism by shifting acid-base equilibrium, affecting glucose tolerance and reducing tissue sensitivity to insulin. In studies with animals, high doses of ammonia (over 100 mg/kg body weight per day) generally have not produced any significant toxic effects, however there is some evidence that ammonia may act with cancer causing compounds to increase the incidence of tumours. Ammonia and ammonium chloride have shown mutagenicity in some tests with bacteria and animal cells.


Holden WS. 1970. Water Treatment and Examination. London: J&A Churchill.

WHO. 2003. Ammonia in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/1).


Calcium (Ca2+) Updated August 2005.

Description and characteristics The presence of calcium in water sources results from the passage of water through or over deposits of limestone, dolomite, gypsum, or gypsiferous shale. It is often added to drinking-water (usually as hydrated lime) to increase the pH after chemical coagulation or to remove carbon dioxide.

Most calcium compounds are not particularly soluble in pure water. However, their solubility increases in the presence of CO2 (carbon dioxide), and sources of water containing up to 100 mg Ca/L are fairly common. Sources containing over 200 mg Ca/L are rare. Calcium contributes to the total hardness of water. Calcium is present in a great number of industrial products and is a common constituent of food, particularly dairy products. In the building industry, calcium oxide is used in mortar, stucco and plaster. It also finds use in pulp and paper production, sugar refining, petroleum refining, tanning, and as a water and wastewater treatment chemical. Appreciable concentrations of calcium salts precipitate upon heating to form often harmful scale in boilers and pipes, and on cooking utensils. Depending on pH and alkalinity, hardness above about 200 mg/L can result in scale deposition, particularly on heating. Protection and scaling tendencies are estimated by the saturation index of a water (refer to Aggressiveness Data Sheet). Calcium in asbestos-cement and cement lined pipes, and concrete tanks leaches when subjected to aggressive water.

Guideline value A guideline value for calcium in water is incorporated into the guideline value for total hardness.


Typical concentrations in drinking-water Values of hardness found in New Zealand water supplies range from less than 5 mg/L to 475 mg/L as CaCO3. However most waters have hardness values ranging from less than 5 mg/L to 80 mg/L as CaCO3. See hardness datasheet for further information.

Removal methods Chemical softening, reverse osmosis, electrodialysis, or ion exchange treatments will reduce calcium and its associated hardness to acceptable levels.

Analytical methods Calcium concentrations may be determined by atomic absorption spectroscopy, inductively coupled plasma emission spectroscopy, and complexometric (EDTA) titration methods.


The calcium hardness value (in mg CaCO3 per litre) can be calculated by multiplying the calcium concentration (as Ca) by a factor of 2.5.

Health considerations There is no evidence of adverse health effects specifically attributable to high levels of calcium in drinking water. A number of ecological and analytical epidemiological studies have shown a statistically significant inverse relationship between hardness of drinking-water and cardiovascular disease. There is some indication that very soft waters may have an adverse effect on mineral balance, but detailed studies were not available for evaluation.

The taste threshold for the calcium ion in drinking water varies from 100�300 mg/L as CaCO3, depending on the anions present.


WHO. 2003. Hardness in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/6).


Chloride (Cl-)

Description and characteristics Chloride is widely distributed in nature, generally in the forms of sodium, potassium, and calcium salts. It constitutes approximately 0.05% of the earth�s crust, but the greatest amount of chloride in the environment is present in the oceans. The presence of chloride in natural waters can be attributed to dissolution of salt deposits, discharges of effluents from chemical industries, sewage discharges, irrigation drainage, contamination from refuse leachates, seawater intrusion in coastal areas, and in some countries contamination resulting from the salting of roads to control ice and snow and oil well operation. A high chloride content may be damaging to metallic pipes and structures as well as being harmful to growing plants. The chloride content of rainwater can exceed 20 mg/L, especially in windy coastal areas.


The taste threshold of chloride in water is dependant on the associated cation but is in the range of 200�300 mg/L. Chloride is generally present at low concentrations in natural surface waters. Concentrations in unpolluted water are often less than 10 mg/L. Seawater infiltration into groundwater or surface water can often be detected by comparing chloride-to-sulphate ratios.

A high chloride content can accelerate corrosion processes, particularly the pitting of steel, and if the chloride content is higher than the alkalinity, dezincification of duplex brasses can occur.

Guideline value Based on aesthetic considerations (taste and corrosion), the chloride concentration in drinking water should not exceed 250 mg/L.

Typical concentrations in drinking-water New Zealand drinking waters typically contain less than 100 mg/L chloride.

Removal methods Chloride cannot be removed from drinking water by conventional water treatment processes. It can be removed by distillation or reverse osmosis but these require considerable energy and can be expensive to operate. Desalination using cryoscopic techniques (by freezing) is commonly used in the Middle East.

Analytical methods The chloride concentration in drinking water can be determined by titration with silver nitrate or mercuric nitrate, using colorimetric or potentiometric end point detection. The limit of detection is approximately 1 mg/L. Ion chromatography can also be used with a limit of detection of 0.1 mg/L.

Health considerations Chloride is an essential element for humans and animals. It contributes to the osmotic activity of body fluids. The main source of human exposure to chloride is the addition of salt to food, and the intake from this source is usually greatly in excess of that from drinking-water. Healthy individuals can tolerate the intake of large quantities of chloride provided that there is a corresponding intake of fresh water. A normal 70 kg human body contains approximately 80 g of chloride. The chloride content in the body is regulated by the kidneys. Little is known about the prolonged intake of large amounts of chloride by humans. High salt intakes have been reported to increase blood pressure but this is attributed to the sodium content rather than chloride.


WHO. 2003. Chloride in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/3).



Colour Updated August 2005.

Description and characteristics Two terms are used to describe colour.

1 �True colour� is the colour of water from which the turbidity has been removed, usually by filtration through 0.45 micron pore size filter.

2 �Apparent colour� is what one actually sees; it is the colour resulting from the combined effect of true colour and any particulate matter. In turbid waters, the true colour is substantially less than the apparent colour.

Apparent colour may result from atmospheric or microbial oxidation of iron and manganese to insoluble hydroxides that precipitate, giving colour (red, brown and black waters), humus and peat materials, plankton, weeds and industrial wastes (eg, dyeing operations, electroplating, paper manufacturing). The dissolution of metals from pipes and fittings can also discolour drinking water. Badly corroded iron pipes can produce a red/brown colour whereas corrosion of copper pipes can produce a blue-green colour. Colour is measured in Hazen Units (HU) or True Colour Units (TCU) � see Measurement Techniques. A true colour of 15 HU can be detected in a glass of water, and a true colour of 5 HU can be seen in larger volumes of water, for instance in a white bath or basin.

Guideline value For appearance reasons, true colour in drinking water should not exceed 10 Hazen Units (HU) or true colour units (TCU).

Typical concentrations in drinking-water Values of true colour in New Zealand drinking waters range from 0 to 40 HU, however most are between 0 and 5 HU.

Removal methods Constituents of natural colour derived from humic and fulvic acids can be reduced by coagulation and flocculation followed by filtration. Oxidation of organics by chlorine or ozone will also reduce colour but may produce undesirable by-products. Colour derived from natural organic matter should be removed prior to chlorination to prevent the production of disinfection by-products. Consumption of disinfectants by reaction with naturally occurring humic and fulvic material can also cause difficulties in maintaining an adequate level of disinfectant, thus creating the opportunity for microbiological regrowth.

Analytical methods True colour can be measured either spectrophotometrically or using visual comparison. In both cases, the standard unit of measurement is Hazen Units (HU). (True colour is often quoted as True Colour Units; however the numerical values are identical.) Hazen Units are defined in terms of a platinum-cobalt standard. This standard was developed for the analysis of colour in natural waters with a yellow-brown appearance, and is not applicable to waters with �different� colours, such as blue copper-containing waters.


Colour measurements can be made in the field or in a treatment plant laboratory for process control purposes, using platinum-cobalt calibrated visual comparator disks, or using the standard platinum-cobalt solutions directly for comparison. Samples must be passed through a 0.45 micron pore size filter before comparison. Visual colour measurement can be rather subjective and those using the method should be screened for their ability to see the colours in a reasonably quantitative manner. The spectrophotometric method is not really suited to drinking-water. There can be a relationship between natural organic matter measured by UV absorbance and colour measured visually. The relationship tends to be source-specific. UV absorbance is a quick, cheap and reliable test.

Colour is generally related to organic content, and while the colour derived from natural sources such as humic and fulvic acids is not of health significance, disinfection of such a water can produce a variety of toxic or carcinogenic organic compounds as by-products. If the colour is high at the time of disinfection, then the water should be checked for disinfection by-products. Low colour at the time of disinfection does not necessarily mean that the concentration of disinfection by-products will be low. Some natural organic precursors to disinfection by-products are not detected by this colour measurement.

Conductivity is an analytical tool, most often used to approximate total dissolved solids (TDS). Total dissolved solids (mg/L) in a sample can be estimated by multiplying conductivity (in mS/m) by an empirical factor. This factor may vary from 5.5 to 7, but higher values up to 9.6 may be associated with waters high in sulphate. The relationship is also distorted in samples with a high silica content (eg, many groundwaters).


Unattractive coloured water may prompt people to seek other, perhaps less safe, sources of drinking water.


Conductivity Updated August 2005.

Description and characteristics Conductivity is a measure of the ability of an aqueous solution to carry an electric current. This ability depends on the presence of ions, their total concentration, mobility, and valence, and on the temperature of measurement. Solutions of most inorganic compounds are relatively good conductors. Conversely, molecules of organic compounds that do not dissociate in aqueous solution conduct a current very poorly, if at all. Using the International System of Units (SI), conductivity has the units millisiemens per metre (mS/m).

Guideline value There is no guideline value.


Typical concentrations in drinking-water The conductivities of New Zealand water supplies range up to 100 mS/m, but are typically between 3�30 mS/m.

Removal methods Any reduction in conductivity is associated with treatment of water supplies to remove specific cations and anions.

Analytical methods Conductance of a solution is measured between two spatially fixed and chemically inert electrodes. To avoid polarisation at the electrode surfaces the conductance measurement is made with an alternating current signal. The conductance of a solution, G, is directly proportional to the electrode surface area, and inversely proportional to the distance between the electrodes. In the laboratory, conductance of a standard potassium chloride (KCl) solution is measured and the corresponding conductivity is calculated. An instrument capable of measuring conductivity with an error not exceeding 1% or 0.1 mS/m, whichever is greater should be used. The water sample being tested should be warmed to 25°C (usually) or a temperature correction should be made to record the equivalent conductivity at 20oC or 25oC; some instruments have an automatic adjustment or compensator. This temperature must be included with the test result. The conductivity of a sample can be estimated, or the estimation technique can be used as a quality assurance step in the measurement of conductivity. Conductivity can also be included as a step in the ion balance check (APHA1998, section 1030E).

Health considerations Conductivity reflects the presence of ions capable of carrying an electric current. The types of ions present and their concentrations will determine the effect of a high conductivity value. Any health consideration should consider the specific ion(s) rather than the conductivity measurement itself.


Hardness (total) Updated August 2005.

Description and characteristics Strictly speaking, hardness is a measure of the multivalent cations in water that are capable of precipitating (or wasting) soap. The dominant cations that give rise to hardness are calcium and magnesium. For drinking-water, this has now lead to hardness being redefined as the sum of the calcium and magnesium concentrations expressed as calcium carbonate. The principal natural sources of hardness in water are sedimentary rocks or seepage and run-off from soils. Hard water normally originates in areas with thick topsoil and limestone formations. Groundwater rich in carbonic acid and dissolved oxygen usually possesses a high solubilising potential towards soil or rocks that contain appreciable amounts of the minerals calcite, gypsum, and dolomite.


The hardness of drinking-water can be increased by dosing calcium compounds (usually as hydrated lime) to increase the pH after chemical coagulation, or to remove carbon dioxide. Degrees of hardness can be described as follows: •

•

•


Hardness may be calculated from the analytical determination of the individual concentrations of the components of hardness, calcium and magnesium; their sum being expressed in terms of an equivalent quantity of calcium carbonate.

<75 mg/L CaCO3 soft 75�150 mg/L CaCO3 moderately hard 150�300 mg/L CaCO3 hard

• >300 mg/L CaCO3 very hard Public acceptance of hardness can vary considerably among communities and is generally related to the hardness that the consumer has come to expect. Soft water may lead to greater corrosion of pipes, although this will depend on other factors such as pH, alkalinity and dissolved oxygen concentration (refer to Aggressiveness Data Sheet). Total hardness above 200 mg/L may lead to excessive scaling of pipes and fittings, and cause blockage of safety relief valves in hot water systems. Aggressive water can dissolve lime out of concrete lined and fibrolite pipes, to the extent that the water can become very hard, particularly in dead-ends, or in areas of low water usage overnight.

Guideline value Total hardness in drinking water should not exceed 200 mg/L as CaCO3 (calcium carbonate), to reduce effects on taste, soap lathering, and to minimise undesirable build-up of scale. Too little hardness in a drinking-water may result in it being aggressive. An appropriate lower limit of hardness, sufficient to suppress excessive corrosion, has to be determined for each water supply.

Values of hardness found in New Zealand water supplies range from less than 5 mg/L to 475 mg/L as CaCO3. Typical values lie between less than 5 mg/L and 80 mg/L as CaCO3. Calcium tends to appear at a higher concentration than magnesium in most New Zealand surface waters and groundwaters. Waters that have Ca and Mg derived from terrestrial sources will usually have Ca hardness greater than the Mg hardness, but the order is the other way around when the Ca and Mg may arise from a marine influence.

Removal methods Hardness can be reduced readily by treatments such as chemical softening, reverse osmosis, electrodialysis or ion exchange.

Analytical methods

Total hardness (mg equivalent CaCO3/L) = 2.5 [Ca, mg/L] + 4.1 [Mg, mg/L] Hardness can also be measured by the reaction of the polyvalent metallic ions present in a water sample with a chelating agent such as EDTA, and is expressed as an equivalent concentration of calcium carbonate.


Health considerations There is no evidence of adverse health effects specifically attributable to high levels of calcium or magnesium in drinking water. Some epidemiological studies have found that hard water may have a beneficial effect on health, particularly on some types of cardiovascular disease, but the data are inadequate to conclude that the association is causal. The taste threshold for the calcium ion in drinking water varies from 100 to 300 mg/L, depending upon the anions present; for the magnesium ion the taste threshold is about 100 mg/L. There is some indication that water with a hardness of less than about 75 mg/L may adversely affect mineral balance.


WHO. 2003. Hardness in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. World Health Organization, Geneva. (WHO/SDE/WSH/03.04/6).


Hydrogen sulphide (H2S) Updated August 2005.

Description and characteristics Hydrogen sulphde is a gas with an offensive rotten eggs odour that is detectable at very low concentrations, below 0.0008 mg/m3 in air. Hydrogen sulphide is formed in drinking-water by the hydrolysis of soluble sulphides, or through the reduction of sulphate by the action of micro-organisms under anoxic conditions. It may be present in anoxic water drawn from deep wells. When this water is released to the atmosphere, some of the hydrogen sulphide escapes into the air. Hydrogen sulphide may be present in source waters as a result of its discharge from industries in which it is used. These include the production of sulphur, sulphuric acid, inorganic sulphides, thiophenes and other organic compounds, dye manufacturing, petrol refining, coke ovens, paper mills, iron smelters, food processing, cosmetic production, fellmongeries and tanneries. It is also present in sewers and is a major component of sewage odour. Anaerobic conditions in areas of low or no flow in the reticulation (eg, dead ends) may result in the microbial reduction of sulphate present in the water to hydrogen sulphide. In water, hydrogen sulphide will be in equilibrium with the sulphide ions and hydrogen sulphide. The ratio depends on the pH, temperature and salinity. At pH 7.4, about one-third exists as undissociated hydrogen sulphide and the remainder largely as the hydrosulphide. Above pH 10 the sulphide ion will be the dominant form; below pH 5 undissociated hydrogen sulphide will predominate. In well-aerated water, hydrogen sulphide is oxidised readily to sulphates and biologically oxidised to elemental sulphur. In anaerobic water, the microbial reduction of sulphate to sulphide can occur.

Guideline value It is highly unlikely that anyone could consume a harmful quantity of hydrogen sulphide from drinking-water. Based on aesthetic considerations (odour and taste) the concentration of hydrogen sulphide in drinking water should not exceed 0.05 mg/L.


Analytical methods

The Methylene Blue Colorimetric Method (APHA 4500-S0.1 mg/L to be determined if a pathlength greater than 1 cm is used.

An ion selective electrode is available for sulphide, and has a detection limit of approximately 0.01 mg/L.

Soluble sulphides are absorbed from the gastro-intestinal tract, although hydrogen sulphide is absorbed principally by the lung. Animal studies have indicated that after absorption, hydrogen sulphide is distributed to the brain, liver, kidneys, pancreas and small intestine.

Typical concentrations in drinking-water Hydrogen sulphide has not been monitored routinely in New Zealand drinking water supplies. In the USA a maximum concentration of 0.5 mg/L of undissociated hydrogen sulphide has been reported in fresh water.

Removal methods Hydrogen sulphide, and other soluble sulphides, can be stripped from water by aeration, so long as the water is sufficiently acidic to ensure that the sulphide is present as hydrogen sulphide (pH values of approximately 5 and less). Some oxidation of the sulphide occurs during aeration, and oxidation can also be carried out by other oxidising agents, such as chlorine, ozone, and permanganate. The initial product formed when oxidation is carried out by chlorine or oxygen (during aeration) is predominantly colloidal elemental sulphur. However, some polysulphides are also formed, and at very low concentrations these compounds still lead to odours. This problem can be overcome by dechlorination by sulphur dioxide or sodium metabisulphite, after the majority of the colloidal sulphur has been removed by filtration. This converts the remaining colloidal sulphur and polysulphides to thiosulphate, which can then be converted to sulphate by rechlorination.

Samples for sulphide analysis should be collected with as little aeration as possible, and, if total sulphide is to be determined, preserved with four drops of 2N zinc acetate per 100 mL of sample. The sampling bottle should be filled completely.

2- D) will allow sulphide levels of less than


There are no data on the human health effects of ingesting water that contains hydrogen sulphide. Ingestion of sulphides has been known to cause nausea, vomiting and irritation of the mucous membranes. Inhalation of hydrogen sulphide is known to be extremely toxic to humans with exposure to amounts as low as 5 mg/L for 30 minutes producing headaches, dizziness, nausea, gastrointestinal disorders and breathing problems. Inhalation of concentrations above 500 mg/L can cause cardiac failure and death.

Animal data are mainly from short-term inhalation studies. Effects include neurotoxic activity and distortions in cardiac rhythm. No long-term cancer studies have been undertaken on hydrogen sulphide. Sodium sulphide did not induce cancers in experimental animals. Hydrogen sulphide was not found to be mutagenic in tests with different strains of bacteria.



WHO. 2003. Hydrogen Sulphide in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/7).


Description and characteristics

Under reducing conditions such as groundwaters drawn from fine sediment peaty deposits or iron-rich volcanic or metamorphic rocks, iron will be present in the soluble, colourless ferrous (Fethe water is drawn and becomes exposed to the air, the ferrous ion will be oxidised to the rusty/cloudy ferric ion (Fe

Iron (Fe2+ and Fe3+) Updated August 2005.

Iron is the fourth most abundant element by weight in the earth�s crust. Iron occurs commonly in soil and rocks as the oxide, sulphide and carbonate minerals. In water it occurs mainly in the divalent and trivalent (ferrous and ferric) states. The presence of iron in natural waters can be attributed to the dissolution of rocks and minerals, acid mine drainage, landfill leachates, sewage or iron related industries. Iron can be associated with natural organic matter, in soluble, colloidal or particulate forms. Iron is used in many domestic and industrial applications, ranging from iron and steel products and pigments in paints, to food colours and preparations for preventing iron deficiency in humans. Ferric chloride and ferric sulphate are also used as coagulants in water treatment. Iron is a natural constituent in plants and animals, and food is the major source of iron intake. Meats are the richest food sources of available iron (haem-iron). Iron has a taste threshold of about 0.05 to 0.1 mg/L in water, and becomes objectionable above 1 mg/L. High iron concentrations give water an undesirable rust-brown appearance, and can cause staining of laundry and plumbing fittings, fouling of ion-exchange softeners, and blockages in irrigation systems. Growths of iron bacteria, which concentrate iron, may cause taste and odour problems, and lead to pipe restrictions, blockages and corrosion.

2+) state. After

3+) state.

Guideline value Based on aesthetic considerations the concentration of iron in drinking water should not exceed 0.2 mg/L.

Typical concentrations in drinking-water Concentrations of iron found in New Zealand drinking waters range from less than 0.05 to 21 mg/L. However, most waters contain less than 0.4 mg/L.

Removal methods Iron salts can be removed effectively by the standard water treatment processes of coagulation followed by filtration. Groundwater supplies with a high iron content can be treated to form iron precipitates, using aeration, oxidation with chlorine, pH adjustment, or lime softening. Greensand, which is used in some places for iron and manganese removal, is treated with permanganate which oxidises these elements forming insoluble compounds which are filtered out.


Analytical methods The iron concentration in drinking water can be determined using inductively coupled plasma emission spectroscopy or atomic absorption spectroscopy. The limits of detection are less than 0.01 mg/L. Alternatively the phenanthroline colorimetric method, which has a limit of detection of 0.01 mg/L, can be used. In this method, iron is reduced to the ferrous state and treated with the colorimetric reagent to generate an orange-red complex.


WHO. 2003. Iron in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/8).

Magnesium (Mg

Iron is an essential trace element for humans. Daily requirements vary with age and gender. For example, women aged 19�54 years need about 14 mg per day but this requirement doubles for pregnant women, while men require about 7 mg per day. Iron deficiency is common and is of particular concern for infants, women, vegetarians and athletes. The amount of iron absorbed from food by the gastro-intestinal tract varies from 0% to 25% according to individual requirements and the source of iron. Iron is used in the production of haemoglobin, myoglobin and a number of enzymes, and is stored in the spleen, liver, bone marrow and muscle. As a precaution against storage in the body of excessive iron, in 1983 JECFA established a PMTDI of 0.8 mg/kg of body weight, which applies to iron from all sources except for iron oxides used as colouring agents and iron supplements taken during pregnancy and lactation or for specific clinical requirements. An allocation of 10% of this PMTDI to drinking-water gives a value of about 2 mg/L, which does not present a hazard to health. Cases of iron poisoning have been reported, mainly among young children who ingest medicinal iron supplements formulated for adults. Physiological regulation of iron absorption confers a high degree of protection against iron toxicity and there are a number of reports of high doses of iron being taken, particularly by adults, with no adverse effects. Studies with animals over long periods have reported only very mild adverse effects associated with a high iron intake. There is no evidence that iron induces cancer in laboratory animals. Most iron salts have been inactive in tests for mutagenicity and do not induce chromosome aberrations in human cells.


Gillanders L, Lloyd L, Maher K, et al (eds). 1992. Dietitians Clinical Handbook. New Zealand Dietetics Association.


2+) Updated August 2005.

Description and characteristics Magnesium is a common constituent of natural water. Many salts containing magnesium are readily soluble and water sources containing magnesium levels up to 10 mg Mg/L are common; natural waters


rarely contain more than 100 mg Mg/L. Healthy adults require about 350 mg of magnesium per day. Magnesium is a common constituent of food of both animal and plant origin. Magnesium is used in various processes in the textile, tanning and paper industries. Alloys of magnesium find extensive use in moulds and die casting, portable tools, luggage and general household goods. The salts of magnesium are used in the production of magnesium metal, fertilisers, ceramics, explosives and medicines. Important contributors to the hardness of drinking-waters, magnesium salts can form scale in boilers.

Guideline value A guideline value for magnesium in water is incorporated in the guideline value for total hardness based on aesthetic considerations.

Values of total hardness found in New Zealand water supplies range from less than 5 mg/L to 475 mg/L as CaCO ost waters have hardness values ranging from less than 5 mg/L to 80 mg/L as CaCO , surface waters are softer than groundwaters. Most waters contain less magnesium than calcium.



3. However m3. Generally

Removal methods Chemical softening, reverse osmosis, electrodialysis, or ion exchange treatments will reduce the magnesium and associated hardness to acceptable levels. Magnesium forms the insoluble hydroxide at pHs above 10�11.

Analytical methods Magnesium concentrations may be determined by atomic absorption spectroscopy, inductively coupled plasma emission spectroscopy and complexometric (EDTA) titration methods.

Once the magnesium concentration is obtained, the magnesium hardness value in mg CaCO3 per litre can be calculated by multiplying the result by a factor of 4.1.

Health considerations There is no evidence of adverse health effects specifically attributable to high levels of magnesium in drinking water. Concentrations greater than 125 mg/L of Mg can have a slight purgative or diuretic effect when associated with sulphate, particularly for people not accustomed to the water supply. The taste threshold for the magnesium ion in drinking water is less than 100 mg/L as Mg.


WHO. 2003. Hardness in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/6).


Manganese


Refer to the manganese datasheet in inorganic determinands section.

Based on the need to reduce corrosion and encrustation in pipes and fittings and to maintain adequate disinfection, the pH of drinking water should normally be between 6.5 and 8.5. At a pH of 6.5, water can corrode some metals at an unacceptable rate. Values of pH at the lower end of this range should therefore be avoided.

pH Updated August 2005.

Description and characteristics pH is a measure of the hydrogen ion concentration of water. It is measured on a logarithmic scale from 0 to 14, and is described as the negative log of the hydrogen ion concentration (-log [H+]). A pH of 7 is neutral, greater than 7 is alkaline and less than 7 is acidic. A pH of 6 is ten times more acidic than a pH of 7.

The balance amongst carbonate, bicarbonate, dissolved carbon dioxide and/or hydroxide will determine the pH of most waters. Thus, contact of water with limestone (CaCO3) and dissolution or release of dissolved CO2 are significant. The acid-base properties of natural organic matter can influence the pH of the water. The acceptable pH range in drinking-water is primarily based on minimising corrosion and encrustation, and on optimising the disinfection properties of chlorine. High pH can cause scaling and encrustation problems, while lower pH can result in corrosion. For chlorine disinfection efficiency, the pH should be kept below 8; efficiency decreases with increasing pH. Values of pH above 9.5 can cause a bitter taste in drinking water, and can irritate skin if the water is used for ablutions.

Guideline value

Using the Langelier Saturation Index (LSI) as an indicator of the tendency of a water towards corrosion and encrustation, a more appropriate pH range for most New Zealand waters would be pH 7.5 to 8.5. As an example, for a �typical� water with characteristics of alkalinity = 80 mg/L HCO3, calcium = 50 mg/L, conductivity = 25 mS/m at 20oC, and temperature = 15oC, the LSI would range from -0.4 to +0.6 over the pH range 7.5 to 8.5. A pH of 6.5 for this water would result in a LSI of -1.4.

Typical concentrations in drinking-water Values of pH in New Zealand drinking water range from 5.1 to 10.0, and most are between pH 6.0 and 8.5.

Adjustment techniques The pH of water can be adjusted by the addition of acid or alkali. Usually lime, soda-ash, caustic soda, and carbon dioxide are used for this. For waters containing excess dissolved carbon dioxide, the pH can be raised by aeration. Addition of aluminium coagulants or the addition of chlorine gas can lower water pH.

Analytical methods pH can be determined potentiometrically using a standard glass electrode and reference electrode, or combination electrode. Three standards are preferred to calibrate and check the electrodes and meter. Commonly used buffer standards are pH 4, pH 7 and pH 9.2. pH is difficult to measure in unbuffered water.


pH measurements can be performed in the field or the treatment plant laboratory, and are often used on-line for process control. On-line electrodes need to be calibrated regularly (as do all electrodes), especially those situated in areas where they are exposed to high turbidity or attachment of biological growths, such as flashmixers, clarifiers, and immediately after lime slurry addition. Particulate matter can adsorb to the electrode surface, altering its response characteristics. Field tests using pH-sensitive dyes and discs with a comparator or Nessleriser exist but are not recommended because of problems of colour blindness, getting a good colour match under all lighting conditions, and because high levels of FAC can modify some colours. Portable pH meters are reasonably price and reliable today.

Higher pH values may have an indirect adverse effect on bacteriological quality through reduced disinfection efficiency. Lower pH values can increase the solubility of heavy metals from pipes, particularly lead and copper, and may decrease the formation of trihalomethanes.

Health considerations Although pH usually has no direct impact on consumers, it is one of the most important operational water quality parameters (WHO 2004). A direct relationship between pH and human health is difficult to determine as pH is closely associated with other aspects of water quality. Consumption of food and beverages with both low and high pH values is common and does not result in adverse health effects. Some carbonated soft drinks have a pH of 2.5, orange fruit juice has a pH of about 3.8, and the pH of fresh milk is 6.7. The effect of pH on health will depend on the buffering capacity of the water used. This is related to the nature and amount of dissolved inorganic and organic material. Water with a low buffering capacity can change pH rapidly, but water with a high buffering capacity is more resistant to pH change.

In humans, extreme values of pH result in irritation of the eyes, skin and mucous membranes. Eye irritation and exacerbation of skin disorders have been associated with high pH values. Gastrointestinal irritation may occur in sensitive individuals at pH values above 10. Below pH 4, redness and irritation of the eyes have been reported, with the severity increasing with decreasing pH.


WHO. 2003. pH in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/12).


Sodium (Na+) Updated August 2005.

Description and characteristics Sodium is present in a number of minerals, the principal one being rock salt (sodium chloride, NaCl). The sodium ion is widespread in water due to the high solubility of sodium salts and the abundance of mineral deposits. Near coastal areas, wind-borne sea spray can make an important contribution either by fall-out on to land surfaces where it can drain to drinking-water sources, or from wash-out by rain. Seawater intrusion, or old salt waters trapped in sediments, are possible sources of high sodium ion concentrations in groundwaters.


Natural contamination, water treatment chemicals, domestic water softeners and sewage effluent can also contribute to the sodium content of drinking water. Sodium salts are used in the paper, glass, soap, pharmaceutical and general chemical industries, and for a variety of other purposes. Sodium is also used in the food industry and for culinary purposes. Considerable amounts are excreted by humans and it is a common constituent of domestic sewage. Sodium, as sodium salts such as sodium chloride or sodium sulphate, has a taste threshold of about 135 mg/L. The taste becomes appreciable when the sodium concentration exceeds 200 mg/L. Most people are exposed to less than 50 mg of sodium per day by drinking tap water (based on a consumption of two litres per day). As sodium salts are very soluble, virtually all the sodium present in water, whether consumed directly, in the preparation of beverages, or incorporated into food, will be absorbed.

Guideline value Based on aesthetic considerations (taste), the concentration of sodium in drinking-water should not exceed 200 mg/L. There is no health based guideline value for sodium. Medical practitioners treating people with severe hypertension or congestive heart failure are often concerned if the sodium concentration in the patient�s drinking-water exceeds 20 mg/L.

Typical concentrations in drinking-water New Zealand drinking-waters generally contain less than 40 mg/L.

Removal methods Sodium salts are not easily removed from drinking water. Processes such as reverse osmosis, ion exchange or distillation can be used but are expensive to operate; these are used for desalination of seawater.

Analytical methods Sodium concentrations in drinking water can be determined by atomic absorption spectroscopy, inductively coupled plasma emission spectroscopy, flame emission spectroscopy, and ion chromatography. The limits of detection can be as low as 0.001 mg/L depending on the instrumentation used.

Health considerations Sodium is essential to human life and is present in all tissues and fluids. It has been estimated that a total daily intake of approximately 200 mg/person is sufficient to meet the needs of growing infants and children. Sodium intake via the water supply makes only a modest contribution to total intake. Water authorities are encouraged strongly, however, to keep sodium concentrations as low as possible. Whether water is consumed directly or with food or beverages, virtually all of the sodium present in it will be absorbed. Sodium concentration in the body is regulated by the kidneys. In general, sodium salts are not acutely toxic substances because of the efficiency with which mature kidneys excrete sodium. Highly excessive intake of sodium chloride may cause vomiting, thus eliminating much of the salt but the vomit response is variable. Increases in the sodium concentration in plasma may give rise to the sensation of thirst. There is evidence linking excess sodium intake with cardiovascular disease. People with severe hypertension or congestive heart failure may need to restrict their overall dietary intake of sodium further


if the concentration in drinking water exceeds 20 mg/L. Medical practitioners treating people with these conditions should be aware of the sodium concentration in the patient�s drinking-water.


WHO. 2003. Sodium in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/15).


Atmospheric sulphur dioxide (SO ed by the combustion of fossil fuels and emitted by the metallurgical roasting processes, may contribute to the sulphate content of surface water. Sulphur trioxide (SO the photolytic or catalytic oxidation of sulphur dioxide, combines with water vapour to form sulphuric acid, which is precipitated as �acid rain� or snow.

Sulphate (SO42-)


Description and characteristics Sulphate occurs naturally in a number of minerals, and is used commercially in the manufacture of numerous products, including fertilisers, chemicals, dyes, glass, paper, soaps, textiles, fungicides and insecticides. Sulphate, including sulphuric acid, is also used in mining, pulping, and the metal and plating industries. Barium sulphate is used as a lubricant in drilling rigs for groundwater supply. The majority of sulphates are soluble in water, the exceptions being the sulphates of lead, barium and strontium. Dissolved sulphate is considered to be a permanent solute in water. Food is probably the major source of sulphate intake. In general, the average daily intake of sulphate from drinking-water, air and food is approximately 500 mg, food being the major source. In areas where the concentration of sulphate in water is high, drinking water may constitute the principal source. In the water treatment industry, aluminium sulphate (alum) or iron sulphate may be used as coagulants. Copper sulphate is occasionally used for the control of blue-green algae (cyanobacteria) in industrial and public water storage reservoirs.

Under anaerobic conditions, the reduction of sulphate to sulphide by sulphate-reducing bacteria can result in an unpleasant odour due to the release of hydrogen sulphide. Such bacteria can also cause corrosion of pipes.

2), form

3), produced by

The taste threshold for sulphate is in the range 250�500 mg/L.

Guideline value Based on aesthetic considerations (taste), the concentration of sulphate in drinking water should not exceed 250 mg/L. Purgative effects may occur if the concentration of sulphate exceeds 500 mg/L and the anion is associated with magnesium, particularly for people unaccustomed to the water supply.

Typical concentrations in drinking-water Sulphate concentrations in New Zealand waters range from below 5 mg SO4/L to 50 mg SO4/L, however most waters contain less than 25 mg SO4/L, or less than 5 mg/L if the supply is not treated by chemical coagulation.


Removal methods Most sulphate salts are very soluble and cannot be removed from drinking water by conventional water treatment processes. Costly desalination methods such as reverse osmosis or distillation are required for sulphate removal.

Analytical methods The sulphate concentration of drinking-water can be determined by colorimetric, gravimetric, turbidimetric, or ion chromatography methods. Limits of detection are method-dependent, ranging from the gravimetric method being suitable for sulphate concentrations greater than 10 mg/L, to the ion chromatographic method suitable for concentrations above 0.1 mg/L.

Health considerations Sulphate is absorbed rapidly by the gastro-intestinal tract, but a number of factors, such as the accompanying cation, can influence the rate of absorption. Low doses are probably absorbed more effectively than high doses. Sulphate is found in all body tissues but is highest in the metabolically active areas of bone and in tooth formation, and may be important in regulating bone development. Sulphate is one of the least toxic anions. Ingestion of high doses can result in catharsis (loosening of the bowels) with dehydration as a possible side-effect, however, no harmful effects have been reported in studies with animals.


WHO. 2003. Sulfate in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/114).


Taste and odour Updated August 2005.

Description and characteristics Taste and odour in drinking-water can occur naturally, or from the result of chemical and/or biological contamination of water supplies. Generally, taste is indicative of the inorganic constituents and odour of the organic constituents of drinking water. Tastes and odours can also result from the materials used in the distribution system or plumbing. Materials should always be suitable for use in water supplies. Odours and tastes are the primary criteria consumers use to judge the quality and acceptability of drinking water. People�s senses of taste and smell tend to vary, so the acceptability of the same water can vary from person to person, and from day to day for the same person. Similarly, one individual within a group may be more or less sensitive to a particular substance than the group as a whole. The taste perception is generally much less sensitive than that of smell. When �tasting� water, the senses of both gustation (taste) and olfaction (smell) are activated and it is extremely difficult to differentiate between them.


pH, colour, temperature, and turbidity all affect the detection and perception of odour and taste. With increased temperature the more volatile compounds vaporise and stimulate the human sensory organs in the nasal and sinus cavities. Temperature also affects the growth rate of micro-organisms, some of which may produce metabolites with offensive tastes and smells. High concentrations of natural organic matter and turbidity in water are often associated with non-specific taste and odour problems. Where pH controls the equilibrium concentration of the neutral and ionised forms of a substance in solution, it can notably influence its taste and odour. Disinfection chemicals can contribute taste or odour to water. The odour threshold for free chlorine varies with pH, but is between 0.1 mg/L and 0.4 mg/L. Monochloramine and dichloramine have odour thresholds of 0.65 mg/L and 0.15 mg/L respectively. A number of organic compounds can produce taste and odour as by-products of disinfection, particularly chlorination. Some chlorinated phenols, for example, have an antiseptic smell and a very low taste and odour threshold, varying from 1.2 mg/L to 0.0005 mg/L. Contamination of water from spills, discharges or leaks of organic compounds can result in unpleasant taste and odours. Diesel fuel, for example, has a taste and odour threshold of 0.0005 mg/L. A number of health significant organic chemicals give rise to taste and odour at concentrations well below their respective MAVs (refer to respective datasheets). These chemicals include ethylbenzene, styrene, toluene, xylene, monochlorobenzene, 1,2- and 1,4-dichlorobenzenes, trichlorobenzenes, and chlorophenols. Organic metabolites resulting from microbial action in waters or surrounding soils can result in a drinking-water that smells �earthy�, �musty� or �woody�. Compounds most often linked to these tastes and odours are geosmin and methyl isoborneol which have extremely low odour threshold concentrations of less than 0.00001 mg/L. These compounds are produced by a number of micro-organisms including blue-green algae (cyanobacteria). While these two compounds are commonly cited as the causes of tastes and odours, they are just a few of the compounds derived from algae that give rise to tastes and odours. Inorganic compounds generally are present in water at substantially higher concentrations than organic compounds. Taste thresholds for some commonly occurring inorganic ions are about 3 mg/L for manganese, 0.05 to 0.1 mg/L for iron, 3 mg/L for copper, 5 mg/L for zinc, 250 mg/L for chloride, and 250 mg/L for sulphate.

Guideline value There is no guideline value for taste and odour. The taste and odour of drinking water should be acceptable to most consumers. The European Economic Community Standards and the USEPA require that the Threshold Odour Number (TON) not exceed 3. The Threshold Odour Number method is no longer regarded as a reliable method for the determination of odours.

Removal methods Substances producing taste and odour can sometimes be removed by granular activated carbon or less effectively by powdered activated carbon. Oxidation, coagulation/flocculation/filtration, and biological treatments may also reduce tastes and odours. Where micro-organisms are responsible for the chemicals, their physical removal prior to disinfection is recommended.

Analytical methods Measurements of taste and odour are difficult to obtain due to the complexity of the testing programme. The Flavour Profile Measurement method is widely recognised as the appropriate procedure when assessing drinking-water. It provides information on both the strength and characteristics of the odour


and taste of the water. A small panel (five to eight people) can be trained to identify specific odours and tastes associated with common contaminants. These panels are useful for assessing complaints by consumers, identifying the sources of a contaminant, and for the initial assessment of a new or improved purification process. A few analytical procedures are known to be able to detect and quantify the very low concentrations of some odour-causing organic compounds, for example closed loop stripping GCMS.

Health considerations Odour in potable water may indicate pollution of the water or malfunctions during water treatment or distribution. Odours of a biological origin can indicate increased biological activity. Some algae can produce toxins, and the detection of these algae by taste and odour provides a useful early warning of potential problems, although taste and odour do not necessarily indicate the presence of toxins.




Temperature Updated August 2005.

Description and characteristics Temperature is an aesthetic criterion for drinking-water. Generally, cool water is more palatable than warm or cold water. Increased temperature will increase the vapour pressure of volatile compounds in drinking-water and may lead to increased odour. Drinking-water with a high mineral content may be more palatable if it is refrigerated. Turbidity and colour of filtered water may be affected indirectly by temperature, as low water temperatures tend to decrease the efficiency of water treatment processes by, for instance, affecting floc formation rates and sedimentation efficiency. Chemical reaction rates increase with temperature, and this can lead to greater corrosion of pipes and fittings in closed systems. Scale formation in hard waters will also be increased by higher temperatures.

Guideline value No guideline is set due to the impracticality of controlling water temperature. The temperature of the water dictates the minimum C.t value that needs to be achieved when disinfecting for protozoa compliance.

Typical concentrations in drinking-water Temperatures in New Zealand drinking water range from 5oC to 30oC, and may vary seasonally as well as geographically.


Control in drinking-water supplies Control of water temperature in reticulated supplies is seldom practical or effective. Selective withdrawal from deep reservoirs can be used but this may introduce other water quality problems. Aeration causing evaporation can also be used to lower temperature. Some practices can affect water temperatures, such as whether water mains and service pipes are buried underground, and how deep. Long retention times in pipes and storage tanks can increase water temperatures, particularly in the summer.

Analytical methods Temperature measurements, made with a thermometer, must be measured on-site. The intended use of the information will dictate which type of thermometer is most appropriate, and whether the thermometer calibration needs to be traceable to a national standard thermometer.

Health considerations The effectiveness of chlorine as a disinfectant is influenced by the temperature of the water being dosed. Generally higher temperatures result in more effective disinfection at a particular chlorine dose. However, FAC in the distribution system dissipates more rapidly as the temperature increases. Water temperature can directly affect the growth and survival of micro-organisms. Increased temperature can also promote the growth of taste and odour producing organisms in lakes and impoundments, and in the distribution system.


AWWA. 1990. Water Quality and Treatment, A Handbook of Community Water Supplies (4th ed). McGraw Hill Inc.

Total dissolved solids Updated August 2005.

Description and characteristics A high total dissolved solids (TDS) concentration is usually associated with a high ion concentration in the water and therefore a higher conductivity, which is a measure of the waters ability to carry electrical currents. The total dissolved solids in water comprise the inorganic salts and any dissolved organic matter. The principal ions contributing to TDS are carbonate, bicarbonate, chloride, sulphate, nitrate, sodium, potassium, calcium, and magnesium. Total dissolved solids influence other qualities of drinking-water, such as taste, hardness, corrosion properties, and tendency for encrustation. At TDS values greater than 1000 mg/L, scaling may occur through the precipitation of salts from the water. The high level of dissolved material may also enhance corrosion through the improved ability of the water to carry electrical currents necessary for corrosion to occur. Water with low total dissolved solids can be aggressive on reticulation pipework and fittings because protective films are not formed. Total dissolved solids may originate from natural sources including wind blow seaspray, sewage, effluent discharges, urban runoff, or industrial waste discharges.


The palatability of drinking-water has been rated according to TDS as follows (Bruvold and Daniels, 1990):

Total dissolved solids (mg/L) Quality <80 excellent 80�500 good 500�800 fair 800�1000 poor >1000 unacceptable Water with extremely low TDS may taste flat and insipid.

Guideline value Based on taste, and the need to reduce excessive scaling and corrosion, total dissolved solids in drinking-water should not exceed 1000 mg/L.

Typical concentrations in drinking-water New Zealand drinking-waters contain TDS levels, approximated from conductivity measurements, that range from 5 to 2400 mg/L; however, most are less than 300 mg/L.

Removal methods It is difficult to remove dissolved salts from drinking-water. Suitable technologies include reverse osmosis, ion exchange, and distillation, but all of these require considerable energy input and can be expensive to operate. Lime softening may also be effective where total dissolved solids are mainly due to hardness.

Analytical methods The most common and least expensive method of determining TDS values is to convert electrical conductivity measurements by multiplication with a factor that varies with the type of water (refer to Conductivity Data Sheet). The most accurate method of TDS determination entails the summation of the concentration of all the major anions and of cations from a water sample analysis. A gravimetric measurement following filtration is the most common analytical method. Total dissolved solids is defined as the residue remaining after evaporation at 105oC of a 0.45 micron filtered sample. Good quality drinking-water samples shouldn�t need filtering!

Health considerations There is no evidence of deleterious physiological reactions occurring in persons consuming drinking-water supplies that have high total dissolved solid levels. There have been some suggestions that high TDS may reduce the effectiveness of chemical sanitation, but the nature of individual components will give a better correlation with chlorine effectiveness.


Bruvold WH, Daniels JI. 1990. Standards for mineral content in drinking water. Journal of the American Water Works Association 82(2): 59�65.

WHO. 2003. Total Dissolved Solids in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization (WHO/SDE/WSH/03.04/16).



Turbidity Updated August 2005.

Description and characteristics Turbidity in water is caused by the presence of fine suspended matter such as clay, silt, colloidal inorganic and/or organic particles. In addition to being aesthetically displeasing, turbidity can protect pathogenic micro-organisms from the effect of disinfection processes and promote disinfection deficiencies. Heavy metals and pesticides can be adsorbed and so concentrate on to the surface of some suspended particles. Examples of turbidity due to micro-organisms are the summer blooms of blue-green algae in surface water, algal debris, and the detritus from iron bacteria in distribution systems (red water). Turbidity is related to the light scattering properties of a water, and for nephelometric turbidity (required in these guidelines) it is defined such that 40 NTU is the turbidity of a standard formazin suspension. Turbidity is used in the DWSNZ as a surrogate for the removal of protozoan (oo)cysts following coagulation and filtration of a water.

Guideline value Based on aesthetic considerations alone, turbidity should not exceed 2.5 NTU. As a guide, water with a turbidity of 5 NTU would appear slightly muddy or milky in a glass. It would not be possible to see through the glass if the turbidity was over 60 NTU. Crystal clear water usually has a turbidity of less than 1 NTU. Although 2.5 NTU is the aesthetic guideline value, the controlling requirement will invariably be that defined in the DWSNZ for protozoan (oo)cyst removal, for each water treatment process.

Typical concentrations in drinking-water New Zealand drinking waters have a turbidity range up to 180 NTU, however most waters are less than 2 NTU.

Removal methods Coagulation followed by filtration through granular media is used widely to reduce turbidity down to 0.3 NTU.

Analytical methods Turbidity is measured using a commercial turbidimeter, with the detector at right angles to the source. The instrument is calibrated using suitable standards (eg, formazin) and results are expressed in Nephelometric Turbidity Units (NTU). The detection limit for most instruments is 0.1 NTU or better. Turbidity measurements are often made in the field and on-line for process control purposes. Portable turbidimeters and on-line meters are available.


Health considerations Highly turbid water does not necessarily constitute a health hazard but turbidities above 1 NTU can protect harmful micro-organisms from the effects of chlorine disinfection, and the particles may adsorb toxic organic or inorganic compounds. Refer to comments in above sections on protozoan cysts.


Taste problems can occur if the zinc concentration in drinking water exceeds 1.5 mg/L as Zn. Water with a zinc concentration above 3 mg/L tends to be opalescent, develops a greasy film when boiled, and has an undesirable astringent metallic taste.

Zinc (Zn2+) Updated August 2005.

Description and characteristics Zinc is an abundant element in the earth�s crust and is found most commonly as the zinc mineral sphalerite (ZnS). The natural zinc content of soils is estimated to be between 1 and 300 mg/kg. Zinc is present in plant and animal tissues, and food is the major source of zinc intake. Drinking-water usually makes a negligible contribution to total intake. In surface and groundwaters the concentration of zinc from natural leaching is usually less than 0.01 mg/L. Tap water can contain much higher concentrations as a result of corrosion of zinc coated pipes, tanks and fittings. Zinc concentrations in galvanised iron rainwater tanks are typically 2 mg/L to 4 mg/L but have been reported as high as 11 mg/L. Zinc is used as a coating to prevent corrosion of iron and steel products, and in the manufacture of brass. Zinc oxide is an important component in the manufacture of paint and rubber products, including tyres.

Guideline value Based on aesthetic considerations (taste), the concentration of zinc in drinking water should not exceed 1.5 mg/L.

Typical concentrations in drinking-water New Zealand drinking waters contain from 0.02 to 9 mg Zn/L, however most are less than 0.2 mg Zn/L.

Removal methods Zinc concentrations in drinking water can be reduced by alum coagulation at pH 6.5-7.0 (30% removal) or by lime softening at pH 9.5 to pH 10 (60% removal).

Analytical methods The concentration of zinc in drinking water can be determined by atomic absorption spectroscopy or inductively coupled plasma emission spectroscopy. The limits of detection can be as low as 0.001 mg/L.


Health considerations Zinc is an essential element for humans; nutritional zinc deficiency results in retarded growth, infantilism, anorexia, mental lethargy, skin changes and night blindness. The recommended intake for adults is 12 mg per day. Drinking water is not usually a significant source of zinc. Approximately 20-30% of dietary zinc is absorbed by the gastro-intestinal tract. Highest concentrations of zinc are found in the liver, kidney, bone, retina, prostate, and muscle. To function properly various enzyme systems require zinc such as alkaline phosphatase, carbonic anhydrase, and alcohol dehydrogenase. More than 70 zinc metallo-enzymes are known. In humans, consumption of excessive quantities of zinc can result in nausea, vomiting, diarrhoea and abdominal cramps, and subsequent copper deficiency, anaemia and gastric erosion. In animal studies, zinc has been reported to reduce the toxic effects of nickel and cadmium. There is no evidence that occupational exposure to zinc increases the risk of cancer. Zinc has been shown to induce chromosomal aberrations in mammalian cells, but is inactive in bacterial mutation tests.


WHO. 2003. Zinc in Drinking-water. Background document for preparation of WHO Guidelines for Drinking-water Quality. World Health Organization, Geneva. (WHO/SDE/WSH/03.04/17).



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