Tetrachloroethene in Drinking-water · surveys in the United States have estimated the average PCE...

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WHO/SDE/WSH/………….

Tetrachloroethene in Drinking-water

Draft background document for development of

WHO Guidelines for Drinking-water Quality

05 June 2018

Version for public review

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Preface

(to be completed by WHO)

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Acknowledgements


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Abbreviations used in the text

BMD benchmark dose

BMDL lower 95% confidence limit of the benchmark dose

BMDLx lower 95% confidence limit estimate of dose corresponding to an x% level of

risk over background levels

bw body weight

CAS Chemical Abstracts Service

CH chloral hydrate

CI confidence interval

CYP cytochrome P450

DCA dichloroacetic acid

DNA deoxyribonucleic acid

EPA Environmental Protection Agency (USA)

FAO Food and Agriculture Organization of the United Nations

GAC granular activated carbon

GDWQ Guidelines for Drinking-water Quality

GSH glutathione

GST glutathione-S-transferase

HBV health-based value

i.p. intra peritoneal

LD50 median lethal dose

LMS linearized multistage

LOAEL lowest-observed-adverse-effect level

LOEL lowest-observed-effect level

MDL method detection limit

NAcTCVC N-acetyl-S-trichlorovinyl-L-cysteine

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NOAEL no-observed-adverse-effect level

NOEL no-observed-effect level

OR odds ratio

PAC powdered activated carbon

PCE perchloroethylene (tetrachloroethene)

PPAR peroxisome proliferator activated receptor

RR relative risk

SCE sister chromatid exchange

SIR standardized incidence ratio

SMR standardized mortality ratio

SSCP single-stranded conformation polymorphism

TCA trichloroacetic acid

TCE trichloroethene

TCVC S-trichlorovinyl-L-cysteine

TCVG S-(1,1,2-trichlorovinyl) glutathione

TCOG trichloroethanol glucuronide

TCOH trichloroethanol

TDI tolerable daily intake

USA United States of America

US EPA United States Environmental Protection Agency

UV ultraviolet

WHO World Health Organization

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Contents 1 EXECUTIVE SUMMARY ............................................................................................................ 1 2 GENERAL DESCRIPTION .......................................................................................................... 2

2.1 Identity .................................................................................................................................... 2 2.2 Physicochemical properties, including speciation .................................................................. 2 2.3 Organoleptic properties ........................................................................................................... 3 2.4 Major uses and sources ........................................................................................................... 3

3 ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE ...................................................... 3 3.1 Water ....................................................................................................................................... 3 3.2 Food ........................................................................................................................................ 4 3.3 Air ........................................................................................................................................... 4 3.4 Bioaccumulation ..................................................................................................................... 5 3.5 Occupational exposure ............................................................................................................ 5 3.6 Estimated total exposure, biomonitoring studies and relative contribution of drinking water 5

4 TOXICOKINETICS AND METABOLISM IN ANIMALS AND HUMANS ............................. 6 4.1 Absorption............................................................................................................................... 6 4.2 Distribution ............................................................................................................................. 6 4.3 Metabolism ............................................................................................................................. 6 4.4 Elimination .............................................................................................................................. 8 4.5 PBPK modelling ..................................................................................................................... 8

5 EFFECTS ON HUMANS .............................................................................................................. 9 5.1 Acute exposure ........................................................................................................................ 9 5.2 Short-term exposure ................................................................................................................ 9 5.3 Long-term exposure ................................................................................................................ 9

6 EFFECTS ON EXPERIMENTAL ANIMALS AND IN VITRO TEST SYSTEMS .................. 11 6.1 Acute exposure ...................................................................................................................... 11 6.2 Short-term exposure .............................................................................................................. 11 6.3 Long-term exposure .............................................................................................................. 12 6.4 Mode of Action ..................................................................................................................... 14

7 OVERALL DATABASE AND QUALITY OF EVIDENCE...................................................... 15 7.1 Summary of health effects .................................................................................................... 15 7.2 Quality of evidence ............................................................................................................... 16

8 PRACTICAL CONSIDERATIONS ............................................................................................ 16 8.1 Analytical methods and achievability ................................................................................... 16 8.2 Source control ....................................................................................................................... 17 8.3 Treatment methods and performance .................................................................................... 17

9 CONCLUSIONS .......................................................................................................................... 17 9.1 Derivation of the health-based value and/or final guideline value ........................................ 17 9.2 Considerations in applying the guideline value .......................................................................... 20

10 APPENDICES .............................................................................................................................. 20 10.1 References ................................................................................................................................ 20

TETRACHLOROETHENE IN DRINKING-WATER

DRAFT Background document for the WHO GDWQ, June 2018

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1 EXECUTIVE SUMMARY




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2 GENERAL DESCRIPTION

2.1 Identity

Tetrachloroethene is also known as perchloroethylene and is abbreviated as PCE.

Other synonyms are: Ethylene tetrachloride; PERC; perchlorethylene; perchloroethene;

perchloroethylene; tetrachlorethylene; 1,1,2,2-tetrachloroethylene; 1,1,2,2-tetrachloroethylene

CAS No.: 127-18-4

Molecular formula: C2Cl4

Chemical structure:

2.2 Physicochemical properties, including speciation

Property Value Reference

Molecular Weight 165.83 IARC, 2014

Boiling point 121 °C Health Canada,

2015

Melting point −22°C Health Canada,

2015

Density at 25°C 1.623 g/mL WHO, 2003

Vapour pressure 1.9 kPa at 20°C Health Canada,

2015

2.53 kPa at 25 °C WHO, 2003

Solubility:

Water (at 25°C) 150 mg/L WHO, 2003

Organic solvents Soluble in ethanol, diethyl ether,

acetone, benzene and chloroform

Partition coefficients:

Log Kow 2.53- 3.40 Health Canada,

2015; ATSDR,

2014

Log Koc 2.2–2.54 ATSDR, 2014

Henry’s law constant 1.8x10-2 atm-m3/mol ATSDR, 2014

Conversion factors 1 mg/L = 147.4 ppm

1 ppm = 6.78 mg/m3

ATSDR, 2014



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2.3 Organoleptic properties

Tetrachloroethene (PCE) is a clear, colourless, non-flammable liquid with an ether-like odour

(Health Canada, 2015). The odour thresholds for PCE in water and air are 0.3 mg/litre and 7

mg/m3 respectively (ATSDR, 1993).

2.4 Major uses and sources

PCE has been used primarily as a solvent in the dry-cleaning industry and as a degreasing

solvent in metal industries, as a heat transfer medium, and in the manufacture of

fluorohydrocarbons as a chemical intermediate (IPCS, 1984, Condie 1985). PCE production

reached its peak during the 1980s (Linak et al., 1992), then over the years its use, especially in

dry cleaning and textile processing, has substantially decreased due to regulation in place in

North America and Europe (IARC, 2014). The major use currently is as a feedstock for

producing fluorocarbons (Glauser & Ishikawa, 2008).

3 ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

Being a volatile organic compound, most PCE released into the environment is found in the

atmosphere, where photochemically produced hydroxyl radicals degrade it to phosgene and

chloroacetyl chlorides with a half-life of 96–251 days (ATSDR, 1997). The partitioning

tendency of PCE in various environmental media has been estimated as follows: air, 99.7%;

water, 0.3%; soil, < 0.01%; sediment, < 0.01% (Boutonnet et al., 1998). In water, it does not

readily undergo hydrolysis or photolysis but is biodegraded by microorganisms to

dichloroethene, vinyl chloride, and ethene. PCE can persist in waters where volatilization

cannot occur. It volatilizes less readily from soil than from water and, with a soil adsorption

coefficient of 72–534, is expected to be fairly mobile in soils. Degradation may occur in

anaerobic soils.

3.1 Water

PCE frequently occurs as a contaminant of groundwater. Due to leaking underground storage

tanks and a nearby dry cleaner, significant levels have been reported in drinking water in

specific US sites, as it was described at the Marine Corps Base in North Carolina where PCE

in drinking water ranged from 2-400 µg/L (mean of 153 g/L), and groundwater levels as high

as 1580 µg/L were observed (IARC, 2014). In provincial Canadian water bodies between the

years 1994-2007, average PCE levels have ranged from 0 to 6.1 g/L and maximum detections

in these waters have ranged from 0 to 54 g/L (Health Canada, 2015).

A survey of drinking-water in the USA in 1976–77 detected PCE in nine of 105 samples at

levels ranging from 0.2 to 3.1 µg/litre (mean 0.81 µg/litre) (US EPA, 2012). In other surveys

of drinking-water supplies in the USA, it was found that 3% of all public water-supply systems

that used well-water contained PCE at concentrations of 0.5 µg/L or higher, whereas those that

used surface water contained lower levels (US EPA, 1987). In the United Kingdom, it has been

detected at levels of 0.4 µg/L in municipal waters (IPCS, 1984, Pearson et al, 1975) and, in

Japan, in approximately 30% of all wells, at concentrations ranging from 0.2 to 23000 µg/L

(ATSDR, 1993). More recently PCE was detected at unspecified levels in 11% (130 of 1,179)



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of drinking water samples from domestic wells in the United States; the reporting limit was 2

µg/L (Rowe et al., 2007). PCE in anaerobic groundwater may degrade to more toxic

compounds, including vinyl chloride (ATSDR, 2014).

3.2 Food

Food is not considered to be a major PCE exposure pathway (Health Canada, 2015). Food

surveys in the United States have estimated the average PCE concentrations in dairy, meat,

cereal, fruit, vegetable, fats and oil, and sugar composite food groups to be 6.6, 12.3, 14.7, 0.8,

0.4, 12.9 and 2.9 μg/kg, respectively. Because of the lipophilic nature of PCE, it may bind to

lipid molecules in foods such as margarine and butter stored in areas where there is PCE in the

air (Health Canada, 2015; US EPA, 2012). PCE concentrations in seafood in the United

Kingdom ranged from 0.5 to 30 µg/kg (Pearson et al., 1975, Dickson et al, 1976). Those in

other foodstuffs ranged from almost undetectable (0.01 µg/kg) in orange juice to 13 µg/kg in

butter (McConnell et al., 1975). Some foods (particularly those with a high fat content) stored

or sold near dry-cleaning facilities may contain considerably higher concentrations (Chutsch

et al., 1990). As part of the Total Diet Study in the USA during 1996-2000, PCE was found in

29 out of 70 (41%) food items from different supermarkets or restaurants (Fleming-Jones &

Smith, 2003). Food was sampled four times a year on a regional basis over a five-year period.

Thus, based on 20 samples of each food item, butter had the most frequent (n=8 of 20 samples)

and highest PCE detection (102 μg/kg). Beef frankfurters (2-60 μg/kg, n=3) had the second

highest detection and margarine (3-42 μg/kg, n=6) had the second most frequent detection.

Potential sources of the contamination were not investigated (Fleming-Jones & Smith, 2003).

3.3 Air

Concentrations of PCE in city air in the United Kingdom range from less than 0.7 to 70 µg/m3

(Pearson et al., 1975). In Munich, suburban and urban air concentrations were 4 and 6 µg/m3

respectively (Lachner, 1976). Surveys in the USA indicated concentrations of less than 0.01

µg/m3 in rural areas and up to 6.7 µg/m3 in urban areas (Lillian et al., 1975). Measurement data

from remote North American sites indicate that background concentrations of PCE have

decreased since 1995 by more than 5% per year (McCarthy et al., 2006). In more recent

surveys, levels of 0.032-0.075 µg/m3 were measured in Antarctica (Zoccolillo et al., 2009);

differences are still measured between rural and urban areas in Italy (0.109–0.719 and 0.461–

4.314 µg/m3, respectively). Levels up to 24 µg/m3 were measured in the USA above

contaminated soil (Forand et al., 2012).

Relatively high levels have been measured in indoor air. PCE was measured in two inner-city

schools in Minneapolis, Minnesota USA with concentrations ranging from 0.1 to 1.3 µg/m3:

the indoor air in homes contained the highest levels of PCE, followed by the personal samples,

outdoors, and indoors in schools (Adgate et al., 2004). In New York, NY USA, residential

indoor air levels of up to peak value of about 5000 µg/m3 (mean value being 34 µg/m3) have

been detected: the sampling period was 2001-2003, where dry cleaners used PCE onsite (New

York State Department of Health, 2010; McDermott et al., 2005). In France concentrations up

to 2900 µg/m3 were recorded (Billionnet et al., 2011; Chiappini et al., 2009) with annual levels

ranging from 0.6 to 124.2 µg/m3

(0.09–18.3 ppb) in residential homes in Paris that were in



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close proximity to dry cleaning facilities, do-it-yourself activities (e.g., photographic

development, silverware), or homes lacking air vents that were built prior to 1945 (Roda et al.,

2013).

In areas were PCE contaminates soil as a result of relevant industrial activities and/or improper

handling or disposal, indoor air contamination can occur and is mainly attributable to soil

vapour intrusion: PCE levels ranged from <0.13 to 1.50 µg/m3

in the homes in San Antonio,

Texas that sit atop an extensive, shallow plume of contaminated groundwater (Johnston &

Gibson, 2014) to 0.1 to 24 µg/m3

in indoor air of residents in the Village of Endicott, New

York, the site of an industrial 1,1,1-trichloroethane spill (Forand et al., 2012). PCE can also

enter the indoor air of homes from sewer gas emissions coming up through the bathroom

plumbing (Pennell et al., 2013).

3.4 Bioaccumulation

With bioconcentration factors reported to range from 39 to 49, PCE is not expected to

bioconcentrate in organisms or to biomagnify within terrestrial and aquatic food chains (US

EPA, 2012), also considering the compound is extensively metabolized in animals.

3.5 Occupational exposure

Average personal exposure of 59 ppm for dry-cleaning workers has been reported in the USA

in 1936–2001 with a temporal trend of decreasing levels (Gold et al., 2008); in Nordic countries

in the mid-1970s median personal exposures were about 20 ppm decreasing to about 3 ppm in

2000 (Lynge et al., 2011). However, high exposures (around 100 ppm in air) still occur in dry-

cleaning facilities in Middle Eastern countries such as Egypt and the Islamic Republic of Iran

(Azimi Pirsaraei et al., 2009; Emara et al., 2010; Rastkari et al., 2011).

Among workers degreasing metal and plastics in the USA the average exposure was reported

to be around 95 ppm during 1944–2001 (Gold et al., 2008), but exposure levels have decreased

by two orders of magnitude over a 60-year period (IARC, 2014).

3.6 Estimated total exposure, biomonitoring studies and relative contribution of

drinking water

PCE can be absorbed following inhalation, oral, and dermal exposure. Air, drinking water,

food, and soil contamination are potential routes of exposure of concern, although inhalation

of contaminated air and ingestion of contaminated drinking water seem to be the most relevant.

Based on the high variability of PCE concentration in air (including indoor environments), it

is not easy to estimate the exposure of the general population via inhalation; by taking 4-6

µg/m3 as an example, estimated exposure would be about 80-120 µg/day for an adult with an

air intake of 20 m3. To evaluate the contribution from drinking water, assuming a concentration

of 0.5 µg/L of PCE, the average daily exposure would be 1 µg for an adult consuming 2 L of

water per day (around 0.8-1.25%of the total inhaled amount). There are insufficient data on the

levels of PCE in foods to allow an average exposure to be determined, hence its contribution

to total exposure.



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In Canada in the early 1990s the total daily intake of PCE was estimated to range from 1.22 to

2.67 µg/kg bw for each of the age groups in the general population. Drinking water ingestion

contributed only a minor amount to the overall exposure, estimated at 0.002 to 0.06 µg/kg bw

per day (<3%). Estimated contributions of food were 0.12–0.65 µg/kg bw per day (7 to 28%).

The highest contributor to exposure was air, which provided >1.22 µg/kg (>62%) of daily

exposure, of which the majority (1.21 to 1.88 µg/kg per day) came from indoor air (Health

Canada, 2015).

A number of biomonitoring studies are available mostly performed in the 1990s, measuring

PCE in expired air or in blood, or trichloroacetic acid (TCA), a metabolite of PCE, in urine

especially among workers in dry-cleaning and other industries, in various countries, as

reviewed by IARC (2014) and ATSDR (2014). Data are also available for the general

population (IARC, 2014; ATSDR, 2014), showing that levels are higher in urban compared to

rural areas, especially for individuals living near a dry-cleaning facility (Brugnone et al. 1994;

Schreiber et al., 2002; Storm et al., 2013).

Due to PCE’s volatility and lipid solubility, incidental exposure could occur through bathing,

showering, or swimming. This indirect dermal or inhalation exposure through drinking-water

can be estimated in terms of litre-equivalents per day (Leq/day) (Krishnan & Carrier, 2008).

For example, an inhalation exposure of 1.7 Leq/day means that the daily exposure to PCE via

inhalation is equivalent to a person drinking an extra 1.7 litres of water per day. In Japan, the

indoor-air exposure to TCE attributable to drinking-water was estimated to be 9.1 Leq/day

(median value for Japanese lifestyle; Akiyama et al., 2018). Overall, drinking water is not a

major source of exposure, unless in the vicinity of a contaminated site.

4 TOXICOKINETICS AND METABOLISM IN ANIMALS AND HUMANS

4.1 Absorption

Animal studies indicate that PCE is rapidly and almost completely absorbed following oral and

inhalation exposure. Dermal absorption of PCE vapour by the skin is minimal compared to

inhalation, but topical exposure to liquid PCE can result in significant absorption (ATSDR,

2014; IARC, 2014; Health Canada, 2015). Concentrations of PCE reached near-steady-state

levels in the blood of human volunteers after 2 hours of continuous inhalation (Stewart et al.,

1961).

4.2 Distribution

PCE fat:blood partition coefficient in humans is in the range of 125–159, whereas the

coefficients for other tissues are in the range of 5-6 (IARC, 2014). Therefore, once absorbed,

due to its high lipophilicity, PCE is mainly distributed to the adipose tissue and other fatty

tissues, including the liver, brain and kidney in both humans and experimental animals (Health

Canada, 2015).

PCE readily crosses the blood-brain barrier and is also found in breast milk and can cross the

placenta and distribute to the foetus (US EPA, 2012; ATSDR, 2014).

4.3 Metabolism



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With the exception of solvent effects that occur at extremely high exposures to the parent

compound, toxicity and carcinogenicity induced by PCE are mainly mediated by its

metabolites. The two pathways of PCE metabolism, oxidation by cytochrome P-450 isozymes

and glutathione conjugation via glutathione-S-transferase (GST), have been extensively

reviewed elsewhere (US EPA, 2012; ATSDR, 2014; IARC, 2014; Health Canada, 2015). In

summary, PCE metabolites are qualitatively similar in humans, rats, and mice although the

three species differ quantitatively regarding the extent of metabolism and the predominant

pathway, recognizing that exposure route and dose also contribute (ATSDR, 2014). Mice

metabolize PCE more extensively than rats, and rats metabolize it more extensively than

humans. Further, saturation of oxidative metabolism, mainly via CYP2E1, although other

isoform can be involved, occurs at similar levels in rats and humans (~100 ppm exposure

concentrations), but oxidative metabolism does not become saturated in mice, which might

explain the lack of renal toxicity in mice (Health Canada, 2015). Geometric mean Vmax values

for PCE metabolism of 13, 144, and 710 nmol/(minute/kg) have been reported for humans,

rats, and mice, respectively (ATSDR, 2014). Regardless of the route of exposure, only 1–3%

of the absorbed PCE is metabolized to its major metabolite, trichloroacetic acid (TCA), by

humans. TCA is a putative toxic moiety for PCE-mediated liver toxicity in mice (Health

Canada, 2015),

Variability of PCE metabolism among humans is reflected by a wide range of reported Vmax

and Km values; It has been suggested that these variations can be attributed to genetic

polymorphisms of the active metabolic enzymes in the studied populations (Lash et al., 2007;

US EPA, 2012).

Oxidative metabolism is postulated to occur mainly in the liver and to a lesser extent in the

lung and kidney (Chiu and Ginsberg, 2011). Although renal metabolism is relevant to rats

(Cummings et al., 1999), the role of renal metabolism is less clear in humans.

After saturation, the second metabolic pathway involving conjugation with glutathione

predominates, although it can also operate prior to saturation, and has been proposed to occur

primarily in the liver and kidney (Chiu & Ginsberg, 2011). Hepatic enzymatic glutathione

conjugation of PCE forms S-(1,2,2-trichlorovinyl) glutathione (TCVG), which is stable, travels

to the kidney, and is biotransformed to S-(1,2,2-trichlorovinyl) cysteine (TCVC) by gamma-

glutamyltransferase and dipeptidases (US EPA, 2012) mainly on the brush-border plasma

membrane of the renal proximal tubular cell. This cysteine conjugate can be bioactivated to

reactive metabolites, or detoxified and excreted into urine, depending on the enzyme involved.

The renal N-Acetyl transferase can detoxify TCVC catalyzing its conversion to N-acetyl-S-

(1,2,2-trichlorovinyl)-L-cysteine (NAcTCVC), which has been detected in human urine

(Volkel et al. 1998). On the other hand, ß-lyase can convert TCVC to a highly reactive 1,2,2-

trichlorovinylthiol compound (TCVT) (Anders et al., 1988; Krause et al., 2003) and flavin

monooxygenase-3 and CYP3A enzymes can lead to the production of the reactive metabolites

TCVC sulfoxide. DCA, the major end-product of the GST pathway, can be formed from TCVT

(US EPA, 2012). In addition, both TCVC and the sulphate can rearrange spontaneously to form

a thioketene, which is the ultimate reactive and toxic agent.



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4.4 Elimination

PCE is eliminated from the body primarily via the lungs in exhaled air as unchanged parent

compound accounting for 80–99% of the intake, regardless of the exposure route in humans.

Urinary excretion of metabolites (predominantly trichloroacetic acid) accounts for 1–3% of the

initial intake (IARC, 2014; Health Canada, 2015). In humans, TCA excretion is linearly related

to inhaled PCE concentrations up to about 50 ppm (ATSDR, 2014).

The elimination half-lives of exhalation and urinary excretion are 65 hours and 144 hours

respectively, in humans (IPCS, 1984; Stewart et al., 1970; Health Canada, 2015).

4.5 PBPK modelling

While a number of inhalation studies are available in occupationally exposed humans and in

experimental animals, the database on ingestion of PCE in drinking water is limited. PBPK

modelling predicts that exhalation can account for 90–99% of initial intake via inhalation and

81–99% via ingestion (Chiu & Ginsberg, 2011).

Considering the main features of PCE kinetics, as summarized above, a linear extrapolation

from high-dose studies in rodents to low-dose human exposures seems not be appropriate,

Indeed:

• PCE is rapidly and well absorbed by both the oral and inhalation routes of exposure

(ATSDR, 2014);

• the metabolic pathways and kinetics of excretion with oral exposure are similar to those

of inhalation exposure (ATSDR, 2014);

• data for oral administration indicate a pattern of effects similar to that of inhalation

exposure;

• differences in first-pass effect (affecting systemic bioavailability) between oral and

inhalation exposures can be adequately accounted for by a PBPK model;

• quantitative differences in metabolism between humans and rodents exist; and

• the metabolite production is not linear since the oxidative pathway is saturated at high-

doses at which the GST pathway start to be active.

The use of PBPK modelling allows a route-to-route extrapolation as well as the estimate of the

internal exposure.

Unlike PBPK models developed for PCE before 2000 generally not including some relevant

information about metabolism, more recent models consider generation of TCA via the

oxidative pathway for rats (Chiu & Ginsberg, 2011), mice (Fisher et al., 2004; Sweeney et al.,

2009; Chiu & Ginsberg, 2011) and humans (Covington et al., 2007; Chiu & Ginsberg, 2011)

as well as GST-mediated metabolism (Sweeney et al., 2009; Chiu & Ginsberg, 2011). While

EPA used the Chiu and Ginsberg model (2011), Health Canada developed and validated its

own model (Nong, 2013) based on previously published models. Use of this PBPK model

allows more precise rodent-to-human and inhalation-to-ingestion extrapolation to equate

laboratory data to human oral exposures.



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5 EFFECTS ON HUMANS

The effects of PCE on human health were extensively reviewed by Health Canada (2105), US

EPA (2012), ATSDR (2014) and IARC (2014), the latter being mainly focused on

carcinogenicity induced by PCE.

5.1 Acute exposure

Oral doses of 4.2–6 g of PCE administered to patients to control parasitic worm infections

caused central nervous system effects, such as inebriation, perceptual distortion, and

exhilaration (Haerer & Udelman, 1964).

PCE is both a potent anaesthetic agent and a cardiac epinephrine sensitizer, therefore after

inhalation of high vapour concentrations human deaths have been reported, as a result of these

two effects (excessive depression of the respiratory centre and/or fatal cardiac arrhythmia

induced by epinephrine sensitization) attributable to the parent compound (ATSDR, 2014).

Post-mortem blood concentrations of PCE in decedents have ranged from 44 to 66 mg/L

(ATSDR, 2014).

Respiratory irritation can be observed in workers exposed to PCE vapours at levels beginning

around 216 ppm or 1464 mg/m3 (ATSDR, 2014). In general, acute studies do not seem to

support liver and kidney effects after short-term exposure, except in extreme cases of accidental

overexposure.

5.2 Short-term exposure

It is not easy to distinguish the health effects associated with PCE short-term exposure to those

related to chronic exposure in humans, since most studies have been carried out primarily in

occupational settings with only limited studies of non-occupational exposure through

contamination of drinking water or indoor air.

5.3 Long-term exposure

5.3.1 Systemic effects

A number of targets of toxicity from chronic exposure to PCE largely focused on the inhalation

route of exposure. The non-cancer-related targets include the central nervous system (CNS),

kidney, liver, immune and hematologic systems, and development and reproduction. In

general, neurological effects were judged to be associated with lower PCE concentrations

compared with other non-cancer endpoints of toxicity. Adverse effects on the liver and kidney

functions were reported by some studies but not reproduced by others (ATSDR, 2014; Health

Canada, 2015).

5.3.2 Neurological effects

The various studies investigating subjective neurological symptoms and objective measures of

neurophysiological, neurobehavioural and neuroendocrine effects of PCE exposure in workers

and the general public have been critically reviewed elsewhere (Health Canada, 2015). The

majority of occupationally exposed cohorts were composed of dry cleaners or other workers in dry

clearing facilities, generally exposed via inhalation; general population cohorts includes residents



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in proximity of or in building also containing dry-cleaning facilities. One cohort of adults exposed

both prenatally and until the age of 5 years to PCE in drinking water pipes was also studied (Getz

et al., 2012).

Visual function, with decrements in colour vision and contrast sensitivity were the two

neurophysiological effects consistently observed in a number of cohorts (Cavalleri et al., 1994;

Echeverria et al., 1995; Gobba et al., 1998; Storm et al., 2011), among which also the cohort

of adults exposed in drinking water for estimated exposure concentrations of ≥40 μg/L (Getz et

al., 2012). Dose-response relationship was evidenced in adults occupationally exposed by

inhalation (Cavalleri et al., 1994; Gobba et al., 1998) with visual effects observed at mean PCE

levels of 4.35–7.27 ppm (Cavalleri et al., 1994;Gobba et al., 1998) Although some other studies

failed in evidencing such an effect, the weight of evidence suggests an association between

PCE occupational and environmental exposure and decrements in colour discrimination

(Health Canada, 2015). These effects seem to be sustained over time with poor recovery

(Gobba et al., 1998; Getz et al., 2012).

In the cohort exposed via drinking water, contrast sensitivity was decreased at the highest frequency

in the high-exposure group (estimated concentrations of ≥ 40 μg/L) (Getz et al., 2012), but this

effect was not consistently reported by other studies (Health Canada, 2015).

The weight of evidence from epidemiological studies investigating neurobehavioral effects or

other neurobiological effects suggests that PCE exposure at elevated levels might be associated

with adverse effects on attention, vigilance and executive function, as well as on attention and

short-term memory function, but not on visuomotor or visuospatial function (Health Canada,

2015).

5.3.3 Reproductive and developmental effects

Several developmental effects, such as eye, ear, central nervous system, chromosomal, and oral

cleft anomalies, were associated with exposure to PCE and other solvents in contaminated

drinking-water supplies (Lagakos et al., 1986). Inhalation exposures have been associated in

female dry-cleaning workers with reproductive effects, including menstrual disorders and

spontaneous abortions (Zielhuis et al., 1989, Kyyronen et al., 1989).

5.3.4 Immunological effects

The limited data available suggest that immunological and associated hematological effects are

possible following PCE exposure in humans, such as by alterations in serum immunoglobulin

E, circulating leukocytes, immunosuppression (host resistance), immunostimulation,

autoimmunity, or allergy-hypersensitivity, in addition to other markers (US EPA, 2012; Health

Canada, 2015).

5.3.5 Genotoxicity and carcinogenicity

A limited number of small cross-sectional studies have evaluated genotoxic and cytogenetic

effects associated with exposure to PCE (IARC, 2014). No significant increase in the frequency

of SCE in association with exposure to PCE was observed in a cross-sectional study among 27

dry-cleaning workers and 26 controls in Japan (Seiji et al., 1990), confirming results of a

previous study on 10 PCE-exposed degreasing workers compared to non-exposed workers in



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the same facility (Ikeda et al., 1980). The frequencies of acentric fragments and chromosomal

translocations in a cross-sectional study in the USA including 18 female PCE-exposed dry-

cleaning workers were not statistically significantly higher than in 18 laundry workers not

exposed to PCE (Tucker et al., 2011).

The carcinogenicity of PCE has been extensively studied, especially in laundry and dry-

cleaning workers and other occupational settings and include both cohort or case-control

studies. An increased incidence of cancer was reported in several cohort and proportionate

mortality studies (WHO, 2003; Blair et al., 1979; Katz & Jowett, 1981; Lynge & Thygesen,

1990) and increased risks of cancer in workers exposed to PCE were found in case–control

studies (Lin & Kessler, 1981; Stemhagen et al., 1983). However, study limitations, such as

concomitant exposures to other chemicals and small sample size, make it difficult to reach a

definite conclusion for most of them.

The National Research Council (NRC, 2010) concluded that evidence for an association

between PCE exposure and lymphoma is suggestive, whereas limited but insufficient for other

cancer types including liver, kidney and bladder cancer. EPA concluded that PCE is likely to

be carcinogenic in humans by all routes of exposure based on sufficient evidence in animals

and suggestive evidence of a causal association between PCE exposure in humans and bladder

cancer, multiple myeloma, and non-Hodgkin’s lymphoma. IARC (2014) classified PCE is

probably carcinogenic to humans (Group 2A).

6 EFFECTS ON EXPERIMENTAL ANIMALS AND IN VITRO TEST SYSTEMS

6.1 Acute exposure

There were no major rodent species or gender differences in susceptibility to lethal effects of

PCE following acute-duration exposure. Oral LD50 values of 3835 and 3005 mg/kg bw were

found for male and female rats, respectively. Acute effects are dominated by central nervous

system depression (Hayes et al., 1986).

A 4-hour inhalation LC50 of 5200 ppm for female albino mice has been reported (Friberg et al.

1953); the highest nonlethal concentrations reported for 4-hour exposure to PCE were in the

range 2,000 to 2,450 ppm in mice and rats (NTP, 1986); the lowest lethal concentrations

reported for 4-hour exposures were 2613–3786 ppm in mice and rats (NTP, 1986).

6.2 Short-term exposure

Groups of male Swiss-Cox mice were given gavage doses of PCE in corn oil at 0, 20, 100,

1000, or 2000 mg/kg of body weight, 5 days per week for 6 weeks (equivalent to 0, 14, 70,

700, or 1400 mg/kg of body weight per day). Mice treated with doses as low as 70 mg/kg of

body weight per day exhibited significantly increased liver triglyceride levels and liver-to-

body-weight ratios. At higher doses, hepatotoxic effects included decreased DNA content,

increased serum alanine aminotransferase, decreased glucose-6-phosphatase serum levels, and

hepatocellular necrosis, degeneration, and polyploidy. The NOAEL was 14 mg/kg bw per day

(Buben et al., 1985).



12

Sprague-Dawley rats (20 per sex per dose) were given PCE in drinking-water at doses of 14,

400, or 1400 mg/kg of body weight per day for 90 days. Males in the high-dose group and

females in the mid- and high-dose groups exhibited depressed body weights. Increased liver-

and kidney-to-body-weight ratios (equivocal evidence of hepatotoxicity) were also observed at

the two highest doses (Hayes et al., 1986).

There was moderate fatty degeneration of the liver in mice following a single (4-h) or repeated

exposure to air containing 1340 mg of PCE/m3. Exposure to this same level for 4 h per day, 6

days per week for up to 8 weeks increased the severity of the lesions (Kylin et al., 1965).

6.3 Long-term exposure

6.3.1 Systemic effects

Chronic (i.e. two years) PCE exposure in laboratory animals has been assessed following

inhalation exposure in rats or mice (JISA, 1993; NTP, 1986) and gavage exposure in rats (NCI,

1977). The target organs were the liver and kidney. Central nervous system effects are

discussed in section 6.3.2 and immune system effects are discussed in section 6.3.4.

Hepatic effects were more prevalent in mice than in rats (Health Canada, 2015). Non-neoplastic

hepatic lesions in mice included angiectasis, central degeneration, and/or necrosis in one or

both sexes exposed to 250 ppm (JISA, 1993). No adverse hepatic effects were observed in rats

in two of the studies (NCI, 1977; NTP, 1986). In the third study, the NOAEL for hepatic effects

was 50 ppm, with increases in spongiosis hepatitis in males at ≥ 200 ppm and decreases in

granulation in females at ≥ 200 ppm that were dead or dying (JISA, 1993). The lowest-

observed-adverse-effect level (LOAEL) for non-neoplastic liver effects, based on necrosis in

male mice in the NTP (1986) study, was 100 ppm (Health Canada, 2015). Hepatocellular

tumors observed in all mouse studies are discussed in section 6.3.5.

Non-neoplastic effects were observed in the kidneys of rats and mice, with the majority of

effects observed in the proximal tubule (Health Canada, 2015). Nuclear enlargement in the

proximal tubule in all rat and mouse exposure groups in the NTP (1986) study and in male rats

exposed to ≥ 200 ppm, female rats exposed to 600 ppm and male and female mice exposed to

250 ppm (as well as low-dose males scheduled for dissection) in the JISA (1993) study. Other

proximal tubule effects included atypical dilation in all high-dose groups (JISA, 1993);

degeneration, necrosis and regeneration in epithelium, inflammatory cell infiltration, fibrosis

and focal mineralization in male and female rats (NCI, 1977); and hyperplasia in a small

number of rats (NTP, 1986). Nephropathy was observed in the majority of male and female

rats and mice exposed by gavage (NCI, 1977), and nephrosis occurred in female mice exposed

by inhalation in the NTP (1986) study. Casts were also observed in the kidneys of dosed male

mice and high-dose female mice in the NTP (1986) study. The NOAEL for renal effects in the

JISA (1993) study was 50 ppm, with nuclear enlargement occurring at 250 ppm; the LOAEL

in the NTP (1986) study was 100 ppm for the same effect. Renal tumors in male rats in the

NTP (1986) study are discussed in section 6.3.5.



13

6.3.2 Neurological effects

As reviewed by Health Canada (2015), three chronic bioassays with PCE examined brain

tissues for gross and histological pathology but did not investigate more subtle neurological

effects. Recognizing this, no adverse effects were noted in rats or mice in the NCI (1977)

study. Relative brain weights were increased in mice inhaling 250 ppm and vitreous deposits

in the brain were increased at 50 ppm but not at 250 ppm in the JISA (1993) study. The NTP

(1986) study identified a significantly positive dose-related trend in the increase in gliomas in

male rats exposed to 200 and 400 ppm; however, the incidence of gliomas in each dose group

was not elevated over control (NTP, 1986).

6.3.3 Reproductive and developmental effects

Only three studies are available, addressing reproductive and/or developmental effects of PCE

exposure via the oral route (ATSDR, 2014; Health Canada, 2015). In addition, these three

studies have limitations due to few doses tested and checked endpoints. Animal studies of

reproductive and developmental effects have focused primarily on the inhalation route of

exposure. Although some contrasting results have been obtained, the weight of evidence

suggest that inhalation exposures during gestation resulted in maternal toxicity in mice, rats,

and rabbits can affect survival of offspring and likely produce alterations in ossification.

Inhalation exposure through gestation or early in development has been linked to neurological

effects in adulthood (Health Canada, 2015).

6.3.4 Immunological effects

Enhanced allergic reactions were observed in mice exposed to doses as low as 2 μg/kg bw per

day (Seo et al., 2012). Health Canada (2015) noted that the study involved passively sensitizing

mice prior to exposure, which leads to difficulties in quantitative dose-response assessment.

This factor, combined with some study limitations, led to the exclusion of this study by Health

Canada (2015) from further consideration in the non-cancer risk assessment.

6.3.5 Genotoxicity and carcinogenicity

The genetic toxicology of PCE has been extensively reviewed (US EPA, 2012; IARC, 2014).

Salmonella reverse mutation assays of PCE as well as other bacterial and yeast tests have

yielded largely negative results with or without hepatic metabolic activation. Micronucleus

formation testing gave inconsistent results; no evidence of unscheduled DNA synthesis was

observed in human lymphocytes or fibroblasts or in rat or mouse hepatocytes (IARC, 2014).

In the presence of rat-liver GST, glutathione and kidney microsomes, simulating the multistep

bio-activation pathway by GSH-conjugation, PCE gave positive results, indicating that its

metabolite TCVG is genotoxic agent. Not all the other metabolites TCVC, NAcTCVC, PCE

oxide, and DCA have been adequately tested in standard assays to support clear conclusions

about their genotoxic potential.

In vivo tests of PCE have been equivocal, with at most, modest evidence of genotoxic effects

in rodent tumor tissues examined (including mouse liver and rat kidney) following exposure at

tumorigenic doses (US EPA, 2012). However, no evidence is available regarding the potential

contribution of PCE genotoxicity to other rodent tumor types (particularly, MCL, testes, and

brain).



14

With respect to chronic bioassays, the exposure by inhalation (6 h per day, 5 days per week for

103 weeks) of F344/N rats to 0, 1.36, or 2.72 g/m3 PCE produced a small (but not statistically

significant) increase in the combined incidence of renal tubular-cell adenomas and

adenocarcinomas in males but not in females. In both sexes, there was an increase in the

incidence of mononuclear cell leukemias at both doses, but the incidence was also unusually

high in concurrent as compared with historical controls (NTP, 1986).

The exposure by inhalation (6 h per day, 5 days per week for 103 weeks) of 49–50 female

B6C3F1 mice (age, 8–9 weeks) to PCE vapor (purity, 99.9%) at concentrations of 0, 100, or

200 ppm (0, 680, or 1360 mg/m3) for 6 hours per day on 5 days per week for up to 103 weeks

resulted in an increase in hepatocellular carcinomas in both sexes (NTP, 1986). When PCE was

administered by gavage in corn oil, there was an increase in the incidence of hepatocellular

carcinomas in both male and female mice but not in Osborne-Mendel rats, recognizing that

survival was reduced in both species as a result of pneumonia, and carcinogenic impurities

were present in the PCE administered (NCI, 1977).

Groups of 50 male and 50 female Crj:BDF1 mice (age, 5 weeks) were exposed to PCE vapor

(purity, 99%) at concentrations of 0, 10, 50, or 250 ppm [corresponding to 0, 68, 340, or 1695

mg/m3] for 6 hours per day on 5 days per week for up to 103 weeks. In males and females,

significant positive trends in the incidences of hepatocellular adenoma, hepatocellular

carcinoma, and hepatocellular adenoma or carcinoma (combined) were observed. The

incidences of hepatic neoplasms were significantly increased at the highest dose compared with

controls. The incidence of hepatic degeneration was also increased at the highest dose in males

(1/50, 1/50, 4/50, and 37/50) and in females (0/50, 1/50, 2/50, and 30/50). The incidence of

splenic haemangioendothelioma and karyomegaly of renal tubular cells was also increased in

the highest dose compared with controls (males: 0/50, 0/50, 6/50, 38/50; females: 0/50, 0/50,

1/50, 49/50) (JISA, 1993).

6.4 Mode of Action

The MoA was extensively analysed and reported by other Agencies (IARC, 2014; Health

Canada, 2015). With the exception of some neurotoxic and cardiotoxic effects due to exposure

to very high levels of PCE by inhalation, which are likely associated to the narcoleptic and

epinephrine sensitivity of the parent compound, toxicity and carcinogenicity induced by PCE

are indeed mainly mediated by its metabolites.

Hepatotoxic and related carcinogenic effects of PCE in mice appear to be due to TCA and

DCA, which is formed in greater amounts by mice than by rats or humans (Schumann et al,

1980; WHO, 2003). The weight of evidence for PCE tends to suggest that a non-mutagenic

mode of action is predominant for hepatocellular tumors (Health Canada, 2015).

The mode of action through which TCA and DCA elicit the benign and malignant

hepatocellular tumors induced by oral or inhalation exposure to PCE in multiple strains and

both sexes of mice remains to be fully elucidated. It has been considered likely that key events

from several simultaneous mechanisms operate, including epigenetic effects as DNA

hypomethylation, cytotoxicity alterations in cell proliferation and apoptosis, disruption of gap-



15

junctional intercellular communication oxidative stress and receptor activation, focusing on a

hypothesized PPARα-activation mode of action.

Peroxisome proliferation is indeed, potentially plausible for PCE-induced hepatocellular

tumors. This has two important consequences: i) a threshold approach could be used for risk

assessment; ii) the human relevance of hepatocellular tumors could also be considered to be

limited, since beside the lower TCA production in humans than in rodents, PPARα expression

is ≥ 10-fold lower in humans and its activation in humans is not related to induction of liver

cell proliferation and suppression of apoptosis. However, some data gaps prevent the confident

identification of peroxisome proliferation as a main MoA (Heath Canada, 2015).

It has been suggested that the induction of kidney tumors in male rats is the combined result of

the formation of a highly reactive metabolite and cell damage produced by renal accumulation

of hyaline droplets (Green, 1990; Olson et al., 1990).

PCE-induced carcinogenicity in the male rats kidney has been associated to GSH-conjugation

genotoxic metabolites, TCVG and NAcTCVC (IARC, 2014), but also to α2µ-globulin-

associated nephropathy, cytotoxicity not associated with accumulation of α2µ-globulin, and

peroxisome proliferation mediated by the receptor PPARα: none of them alone can explain the

renal carcinogenicity in rats, and it is likely that several simultaneous mechanisms operate

The mechanism by which PCE induces immunotoxicity, or by which the toxicity may

ultimately lead to carcinogenesis, could not be identified from the available human and animal

studies. Therefore, it was not possible to assess the plausibility and relevance to humans (IARC,

2014; Health Canada, 2015).

The mechanism by which PCE induces neurotoxic effects has not been identified but

considering the number of molecular targets reported in the studies, it is likely that several

mechanisms are concurrently operating.

7 OVERALL DATABASE AND QUALITY OF EVIDENCE

7.1 Summary of health effects

The highest exposure levels of PCE have been associated with occupational environment (dry

cleaning facilities or industries using metal degreasing products); although decreasing in

western countries it still occurs in some countries. For the general population the important

routes of PCE exposure are inhalation of ambient or indoor air and ingestion of contaminated

drinking water.

The nervous system is more sensitive to PCE than other target organs such as the liver, kidney,

or immune system. Human epidemiological studies (Cavalleri et al., 1994; Echeverria et al.,

1995; Gobba et al., 1998; Schneiber et al., 2002; Storm et al., 2011), and human controlled

exposure experiments (Altmann et al., 1990; Hake & Stewart, 1977) have identified CNS

effects following acute-, intermediate-, and chronic-duration exposures to PCE.

Acute inhalational exposure at 232–385 ppm causes respiratory irritation, whereas inhalation

of high PCE vapor concentrations causes central nervous system depression, loss of



16

consciousness, cardiac failure, and death. Chronic inhalation exposure at dose >4.8 ppm has

been associated with neurobehavioral effects and vision changes in humans.

Studies in animals also identify the liver, kidney, immune system as potential targets of toxicity

and carcinogenicity induced by PCE metabolites.

Data on human carcinogenicity are often weak and sometimes inconsistent. The mode of action

for experimental animal tumors was not fully elucidated, therefore the relevance for humans

could not be completely assessed. The US EPA concluded that PCE is likely to be carcinogenic

in humans by all routes of exposure based on sufficient evidence in animals and suggestive

evidence of a causal association between PCE exposure in humans and bladder cancer, multiple

myeloma, and non-Hodgkin’s lymphoma. IARC (2014) classified PCE as probably

carcinogenic to humans (Group 2A), based on limited evidence in humans and sufficient

evidence in experimental animals.

Reproductive and developmental effects on experimental animals were observed in some

studies (e.g. increased pre-and post-implantation losses, decreased litter sizes, delayed skeletal

development). Available epidemiological evidence in occupational settings or in contaminated

drinking water suffer from a number of limitations including lack of measured exposure levels,

co-exposure to other solvents, lack of control for potential confounders, and small numbers of

subjects and do not provide sufficient bases to draw conclusions.

7.2 Quality of evidence

Overall, PCE is a data-rich compound with many good quality studies on kinetics and toxicity

in humans and animals. Although the main focus of the evidence is on inhalation exposure,

the availability of PBPK modelling allows inhalation-to-oral route extrapolation.

Evidence on neurotoxic effects, which were considered critical, are well established in humans

and animals. Human data have been obtained in occupationally exposed individuals and a clear

dose-response relationship was evidenced in two studies. Many studies consisted of a single

exposure group or had poorly matched control groups. No study was available on the general

population.

Mode of action for tumor induction in experimental animals has not been fully elucidated, and

evidence on human carcinogenicity is often weak and sometimes inconsistent.

Available epidemiological studies on reproductive or developmental effects in occupational

settings or in contaminated drinking water suffer from a number of limitations and could not

be used to draw any firm conclusions.

8 PRACTICAL CONSIDERATIONS

8.1 Analytical methods and achievability

PCE can be analyzed together with trichloroethene by gas chromatography with ISO

10301:1997, which has a limit of quantification of 0.1 μg/L (ISO, 1997).

Four methods for measuring PCE in drinking-water have been approved by the US

Environmental Protection Agency, with generally lower detection methods (US EPA, 2002).



17

• Method 502.2 Revision 2.1, which uses purge and trap capillary gas chromatography

(GC) with photoionization detectors and electrolytic conductivity detectors in series,

has a method detection limit (MDL) range of 0.02–0.05 μg/L.

• Method 524.2 Revision 4.1, which uses purge and trap capillary GC with mass

spectrometry (MS) detection, has an MDL range of 0.05–0.14 μg/L.

• Method 524.3 Version 1.0 measures VOCs in drinking water. The method uses the GC-

MS technique and has a detection limit of 0.036 μg/L. The advantages of the method

include an optimization of the purge and trap parameters, an option for use of selected

ion monitoring (SIM) and the use of solid acid preservatives.

• Method 551.1 Revision 1.0, employing liquid–liquid extraction and GC with electron

capture detectors, has MDLs of 0.002 μg/L and 0.008 μg/L when methyl tertiary-butyl

ether or pentane is used as the extraction solvent, respectively.

8.2 Source control

PCE is primarily, if not exclusively, a groundwater contaminant as it is lost to the atmosphere

from surface waters. The primary cause of contamination is poor handling and disposal

practices, which result in soil contamination and, in vulnerable aquifers, subsequent water

contamination. This may occur some distance from the source, but contamination may be

drawn into the source by pumping. As PCE can persist in waters where volatilization cannot

occur, source control should be the priority action. Control at the source should be relatively

cheap and straight forward by improving handling and disposal practices. Source control will

not be feasible where there is historical contamination, which may be the major source of

contamination in many countries.

8.3 Treatment methods and performance

According to Health Canada (2015), conventional treatment is not effective for the removal of

PCE. The treatment for removal of volatile substances in groundwater is normally air stripping,

but sometimes granular activated carbon can be applied but the removal efficiency is dependent

on process design and operation conditions. At the residential scale, certified point-of-use

treatment devices are available that can remove volatile organic chemicals (VOCs) such as

PCE from drinking water based on adsorption (activated carbon) technologies and may be

installed at the faucet (point-of-use) or at the location where water enters the home (point-of-

entry). Point-of-entry systems may be preferred for the removal of VOCs, because they provide

treated water for bathing and laundry as well as for cooking and drinking thereby decreasing

potential VOC exposure through inhalation.

Treatment is not needed on surface water sources as PCE would volatilize to the atmosphere.

9 CONCLUSIONS

9.1 Derivation of the health-based value and/or final guideline value

Cancer effects

Hepatocellular adenomas and carcinomas are the most relevant endpoint for cancer risk

assessment, since they were observed in both sexes of mice in both chronic inhalation studies,

JISA (1993) and NTP (1986), with significant dose-related trends. The weight-of-evidence for

PCE tends to suggest that a non-mutagenic mode of action is predominant for hepatocellular

tumors (Health Canada, 2015).

Key studies:



18

• NTP (1986) observed significant positive trends in male mice for the incidence of

hepatocellular adenoma and in both sexes for hepatocellular carcinoma. Renal tubular

adenomas and adenocarcinomas in male rats were also observed.

• JISA (1993) observed significant positive trends in the incidences of hepatocellular

adenoma, hepatocellular carcinoma, and hepatocellular adenoma or carcinoma

(combined) in male and female mice. The incidence of splenic haemangioendothelioma

and karyomegaly of renal tubular cells was also increased in the highest dose compared

with controls.

The cancer dose-response modelling for PCE exposure considered the generation of oxidative

metabolites in the liver as the most relevant dose metric for hepatocellular tumors, recognizing

that the external dose PoD from NTP (1986) gave slightly more conservative results (Health

Canada, 2015).

To calculate the TDI, the BMDL10 of 1.7 mg/kg bw per day is divided by a UF of 75

comprising:

• 2.5 for the remaining uncertainty associated with potential interspecies toxicodynamic

differences, since toxicokinetic differences between mice and humans were accounted for

within the PBPK model.

• 10 for intraspecies variability; the variability of PCE metabolism among humans is

reflected by a wide range of reported Vmax and Km values ; It has been suggested that

these variations can be attributed to genetic polymorphisms of the active metabolic

enzymes in the studied populations (Lash et al., 2007; US EPA, 2012).

• 3 for gravity of effects (MoA for carcinogenicity potential not fully elucidated).

TDI (carcinogenic endpoints) = 1.7 mg/kg bw per day

75

= 0.022 mg/kg bw per day

Non-cancer effects

Neurotoxicity was observed in humans and experimental animal studies, with effects in humans

observed at lower concentrations than in animals. The availability of studies in humans,

supported by animal data, as well as the availability of PBPK modelling to extrapolate from

the inhalation to the oral route of exposure, distinguish the present PoD from the previous

guideline value for PCE (WHO, 2003). The PoD was at that time 14 mg/kg bw per day from

two oral toxicity studies in laboratory animals: a 6-week gavage study in male mice and a 90-

day drinking-water study in male and female rats (Buben & O’Flaherty 1985, Hayes et al,

1986).

Key studies:

• Cavalleri et al., (1994) showed a significant decrease in colour vision (mainly blue-

yellow range) in exposed workers.

• Echeverria et al., (1995), using a standardized neurobehavioral battery found changes

in cognitive (reaction time measures and cognitive changes) and visuospatial function.



19

It can be used to qualitatively support the key study, since the exposure assessment was

not as robust as the one performed by Cavalleri et al. (1994).

A NOAEL of 4.8 ppm for colour confusion was identified from Cavalleri et al. (1994). A

BMDL10 of 6.6 ppm can be estimated (Health Canada, 2015), to which the PBPK modelling

validated by Health Canada (Nong et al., 2013) can be applied to extrapolate from inhalation

exposures to equivalent oral doses. In the modelling, the peak concentrations of PCE in the

kidney were used as a proxy for brain values, since partitioning coefficients for PCE in brain

are similar to those in kidney (Dallas et al., 1994), and the effect is very likely mediated by the

parent compound. The external oral dose associated with the BMDL10 was 4.7 mg/kg bw per

day.

As noted by US EPA (2012), the RfDs for non-cancer effects other than neurotoxicity are

within about 10-fold of the RfD, therefore multiple effects may occur at about the same

exposures at which PCE begins to induce neurotoxicity. These results also suggest that it is

important to take into account effects from PCE other than neurotoxicity when assessing the

cumulative effects of multiple exposures.

To calculate the TDI, the BMDL10 is divided by an UF of 300 comprising:

• 10 for database deficiency, taking into account the relatively old studies available;

• 10 for intraspecies variability; and

• 3 to extrapolate from an occupational study with intermittent exposure for 8.8 years;

(this is considered as a conservative approach).

TDI for neurological effects = 4.7 mg/kg bw per day

300

= 0.016 mg/kg bw per day

Cancer and non-cancer endpoints were considered in the derivation of the GV for PCE in

drinking water.

The GV for PCE in drinking water derived considering cancer effects (assuming a non-

genotoxic, threshold mode of action) was higher than the GV derived for the non-cancer end-

points. The GV value was derived for non-cancer effects, and it is protective for both cancer

and non-cancer end-points.

GV = 0.016 mg/kg bw per day x 0.2 x 60 kg bw

2 L

= 96 µg/L (rounded to 100 µg/L)

where:

• 0.016 mg/kg bw per day is the TDI for non-cancer end-points;

• 0.2 is the proportion of total daily intake that is allocated to drinking-water, as there is

evidence of widespread presence in at least one of the other media (air, food, soil, or

consumer products) (Health Canada, 2015);

• 60 kg is the average body weight of an adult;



20

• 2 L/day is the daily volume of water consumed by an adult.

A higher allocation than the default factor of 20% to drinking-water could have been

considered, since releases of PCE into the air and water have continued to demonstrate

decreasing trends over the past few decades (US EPA, 2012; Health Canada, 2015) and thus,

human exposures to PCE are anticipated to continually decline by all probable exposure routes.

However, this was not considered necessary as 100 µg/L PCE in drinking water should be

achievable.

9.2 Considerations in applying the guideline value

For certain population groups, such as dry-cleaning workers, the level of inhalation and dermal

exposure is likely to be high compared to exposure through drinking-water.

In some parts of the world, there may be a need to adapt the guideline value by adjusting the

allocation factor to account for local conditions, including inhalational exposure from long-

time showering and bathing and/or in poorly ventilated buildings. Consideration could also be

given to overall exposure through the various environmental media, as described in in section

9.1.

Requirements for monitoring PCE in drinking-water regulations and standards should be

limited to groundwater sources where a catchment risk assessment indicates the possibility of

its presence. In some cases, this may include draw in from contamination elsewhere in the

aquifer. Monitoring can be at the treatment works and if concentrations are shown to be stable

or effective treatment is in place, then the frequency of monitoring can be quite low.

In the event that monitoring data show elevated levels of PCE, it is suggested that a plan be

developed and implemented to address these situations. Monitoring is not needed on surface

water sources since PCE would volatize out.

10 APPENDICES

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