LEAD IN - Secretary of State for Environment, Food and...

CHARACTERISTIC BEHAVIOUR AND POTENTIAL MITIGATION OF SOME DIFFUSE POLLUTANTS IN ENGLAND

AND WALES: A REVIEW OF AMMONIUM, NITRITE, SOME POTENTIAL PATHOGENS AND BIOLOGICAL OXYGEN

DEMAND

A REVIEW PREPARED AS PART FULLFILLMENT OF DEFRA PROJECT ES0121 ‘COST-DP: COST EFFECTIVE DIFFUSE

POLLUTION MANAGEMENT’

PROJECT TEAM:Haygarth, P.1, Chadwick, D.1, Granger, S.1, Chambers, B.2, Anthony, S.3,

Smith, K.3 and Harris, D.3

1Institute of Grassland and Environmental Research (IGER), North Wyke Research Station, Okehampton, Devon, EX20 4LR.

2ADAS, Gleadthorpe, Meden Vale, Mansfield, Notts, NG20 9PF.3ADAS, Wolverhampton, Woodthorne, Wolverhampton, WV6 8TQ.

EXECUTIVE SUMMARY

This review fulfils milestones 1-4 of project ES0121 ‘COST-DP: Cost effective diffuse pollution mitigation’. This project follows on from Defra projects PE0203 and NT2511 which reviewed P and NO3

- respectively. Project ES0121 reviews the remining significant diffuse pollutants; these are Ammonium (NH4

+), Nitrite (NO2-),

Pathogens and Biological oxygen demand. The review of each pollutant has been broken down into:

An introduction to the pollutant Defining the problem in terms of Source, Mobilisation and Delivery Mitigation measures Pollution swapping

Pollutant summary:

AmmoniumAmmonium (NH4

+) is applied to agricultural land to promote plant growth and optimise agricultural yields. It can be applied in inorganic forms as fertilisers or via animal manures and other organic residues and effluents. Ammonium can also be released from organic nitrogen forms through the process of mineralization. Ammonium transfers to watercourses can cause eutrophication while dissolved ammonia (NH3) is directly toxic to fresh water fish, and the nitrification of NH4

+ to NO3

- can cause oxygen depletion.

NitriteNitrite (NO2

-) is seldom measured in aquatic systems and often included in nitrate values. The two dominant processes involved with NO2

- turnover in the environment are nitrification of NH4

+ and the reduction of NO3- during denitrification. Nitrite is

highly toxic and high concentrations can adversely affect plants and soil micro-organisms, while in watercourses NO2

- poses a threat to fish and some species of invertebrates.

PathogensThe rumen and digestive tract in farm livestock is host to a rich diversity of microflora and can act as a reservoir for pathogenic micro-organisms i.e. E. coli. O157, Salmonella, Listeria, Campylobacter, Cryptosporidium and Giardia. Pathogen presence in manures is affected by factors such as animal type, age, diet and management, as well as regional and seasonal influences. Pathogenic micro-organisms may be transmitted from animals to humans either directly or indirectly through water and food chain contamination.

Biological oxygen demandMost agricultural organic wastes contain substantial quantities of biologically degradable material, which means considerable potential for pollution once these effluents gain access to watercourses. The presence or absence of micro-organisms influences the pollution potential of the waste. The main effect of loading surface waters with organic matter is the rapid depletion of available oxygen as a result of the increased microbial activity stimulated. In fast-flowing waters, a rapid restoration of aerobic conditions can quickly follow (as long as the pollution source is blocked), but in stagnant or slow-moving waters, anaerobic conditions quickly develop associated

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with the generation of foul odours and, in the longer term, the reduction in biodiversity of the system.

The mitigation matrix

A mitigation matrix has been developed (Appendix I) to help provide Defra and the project team with a provisional ‘guide’ to key pollutant behaviour properties and mitigation options. This lists the main mitigation measures currently available to agriculture and relevant to the pollutants of interest in this study. Each measure is rated according to how effective it is considered to be at mitigating each pollutant. The cost-effectiveness of each measure within the matrix has also been scored and the ease of implementation is also rated. Other measure characteristics that have also been rated include public and farmer acceptability, potential for pollution swapping, potential for conflicts with other measures, and the uncertainty of effectiveness of the measure. As part of this process, a list of typical farm systems has been created (Appendix II). This has been based upon livestock/cropping system, soil type and drainage system, manure system, and grazing regime.

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CONTENTS

1. INTRODUCTION AND CONTENT.

2. AMMONIUM.1. INTRODUCTION.2. DEFINING THE PROBLEM.

2.1. SOURCES.2.1.1. Managed Manures.2.1.2. Outdoor Livestock.2.1.3. Inorganic Fertilizers.2.1.4. Silage Clamps.2.1.5. Atmospheric Deposition.2.1.6. Mineralisation.

2.2. MOBILISATION.2.3. DELIVERY.

2.3.1. Surface Pathways.2.3.2. Preferential Pathways.2.3.3. Through-flow Pathways.2.4.4. Field Runoff and Leaching of Inorganic Fertilizers.

3. MITIGATION MEASURES.3.1. Hard Standings.3.2. Manure Stores.3.3. Grazing/Outdoor Livestock.3.4. Land Spreading of Manures, Dirty Water and Other Organic …...Residues.3.5. Silage Clamps and Big Bales.3.6. Septic Tanks.3.7. ‘Upstream’ Measures.

4. POLLUTION SWAPPING.

3. NITRITE.1. INTRODUCTION.2. DEFINING THE PROBLEM.

2.1. SOURCES.2.1.1. Direct Sources.2.1.2. Indirect Sources.2.1.3. Processes of NO2

- Formation in Soil.2.1.4. Relative Contributions of Nitrification and …………

Denitrification to Soil NO2-.

2.1.5. Causes of NO2- Accumulation.


3. MITIGATION MEASURES.4. POLLUTION SWAPPING.

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4. PATHOGENS.1. INTRODUCTION.2. DEFINING THE PROBLEM.

2.1. SOURCES.2.1.1. Microbial Pathogens.2.1.2. Protozoa.2.1.3. Viruses.2.1.4. Summary


2.3.1. Field Losses Following Manure Spreading.2.3.2. Field Losses Following Livestock Grazing.2.3.3. Losses During Slurry Storage.2.3.4. Farmstead Runoff.2.3.5. Livestock Access Direct to Water Courses.

2.4. PATHOGEN LOSS RISK MATRIX.3. MITIGATION MEASURES.

3.1. DIETARY AND MICROBIAL MANIPULATION.3.1.1. Cattle.3.1.2. Pigs.3.1.3. Poultry.3.1.4. Summary.

3.2. MANURE STORAGE AND TREATMENT.3.2.1. Storage.3.2.2. Slurry Treatment.3.2.3. Solids Composting.

3.3. MANURE SPREADING AND GRAZING ANIMALS.3.3.1. Spreading Method.3.3.2. Direct Deposition to Water Courses.3.3.3. Die-off in the Soil Environment.

3.4. METHOD EFFECTIVENESS.3.2.1. Storage.3.2.2. Slurry Treatment.3.2.3. Solids Composting.


5. BIOLOGICAL OXYGEN DEMAND.1. INTRODUCTION.2. DEFINING THE PROBLEM.

2.1. SOURCES.2.2. MOBILISATION.2.3. DELIVERY.

2.3.1. Outdoor Grazing.2.3.2. Manure and Dirty Water Land Applications.2.3.3. Summary.

3. MITIGATION MEASURES.3.1. STORAGE.

3.1.1. Management of Manure.3.1.2. Management of Dirty Water.3.1.3. Management Benefits.

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3.2. GOOD MANAGEMENT PRACTICE.3.2.1. Manure Application Practice.3.2.2. Manure Application and Soil Management.3.2.3. Solids or Liquids Manure Management Systems.

3.3. MANURE EXPORTS.3.4. TREATMENT SYSTEMS.

3.4.1. Mechanical Separation.3.4.2. Anaerobic and Aerobic Digestion.3.4.3. Solids Composting.3.4.4. Use of Treatment Additives.3.4.5. Manure Processing.3.4.6. Soil Treatment Processing.3.4.7. Constructed Wetlands.


6. A CONCEPTUAL MODEL FOR DIFFUSE POLLUTANT BEHAVIOR: A MEASURE CENTRIC APPROACH.1. A MEASURE CENTRIC APPROACH.2. THE MITIGATION MATRIX.3. CONCLUSIONS.

7. APPENDIX I. Mitigation Matrix.

8. APPENDIX II. Model Farm Scenarios.

9. REFERENCES.

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1. INTRODUCTION AND CONTENT

Tackling water pollution over the last 40 years has targeted easily controllable point sources within a catchment such as sewage treatment works and industrial outfalls. Significant progress has been made across the UK addressing these sources of pollution but as further point source control becomes less cost effective, attention is now being directed towards the contribution of diffuse pollution from agriculture.

Diffuse pollution from agriculture and urban areas is now recognised as one of the most significant water quality problems facing the world. In 1995, the US Environment Protection Agency (USEPA) reported that 40% of US rivers, lakes and estuaries did not meet water quality requirements (USEPA, 1995) and that diffuse pollution was identified as the primary cause of this problem.

In recognition of this growing problem, the EU Water Framework Directive (EC, 2000) was adopted by the EU parliament in December 2000. This Directive marks a shift away from effluent based controls to water quality based controls and total maximum daily loads. The directive puts the emphasis on river basin management plans similar to those that were developed in the US during the 1990s.

The terms ‘point source’ and ‘non-point source’ pollution are often used to describe the difference between easily identifiable, generally controllable point sources of pollution from the more diffuse pollution within a catchment, such as runoff from fields and roads. Pollution sources within a catchment tend to be closely linked to land uses (i.e. the application of manures to farmland) and the changes to land use that cause the mobilisation of pollutants (i.e. ploughing of fields or road construction).

Agriculture covers 76% of the land area of England and Wales (Defra, 2001) and as such is a key generator of diffuse pollution. An example of this is nitrogen; it has been estimated that agriculture contributes 70% of the diffuse nitrogen inputs to inland surface waters (The Royal Society, 1983). Activities such as ploughing, the spreading of manures and inorganic fertilisers, and the application of agrochemicals can all give rise to the inadvertent contamination of water supplies.

Pollutants of concern originating from diffuse sources are presented in Table 1. However, the most troublesome pollutants from agriculture are sediments, nutrients, faecal pathogens and pesticides. Future policy for managing water quality requires an understanding of the measures that can decrease losses of pollutants and the costs of implementing them. Defra projects PE0203 and NT2511 (P and N cost curve respectively) highlighted mitigation measures and their associated costs for phosphorus and nitrogen. These projects initiated this Defra project; ES0121 (COST-DP: Cost effective diffuse pollution mitigation) which deals with many of the remaining diffuse pollutants. It aims to provide hard scientific information on the processes involved in diffuse, ammonium, nitrite, pathogenic (E. coli. and Cryptosporidium) and biological oxygen demand (BOD), pollution. The project has prioritised options for mitigating each of these pollutants in terms of costs, pollution reduction, practicalities and applicability in England and Wales. This review contributes in part to fulfilment of project ES0121 by meeting the milestones (1-4):

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1. To undertake a ‘bottom up’ literature review of the potential mitigation options available for nitrite, ammonium, pathogens (E. Coli as a bacterial indicator and Cryptosporidium) and BOD.

2. To produce an objective system for classifying the functional behaviour of diffuse pollutants.

3. To prepare a mitigation matrix for diffuse pollutants.4. To construct typical ‘model farm’ scenarios that can be used to explore

mitigation options.

Table 1. Diffuse pollution concerns (modified from D’Arcy et al., 2000).

Pollutant Example Source Environmental Problem

SedimentsRunoff from agricultural land;

upland erosion; forestry; construction sites

Destruction of gravel riffles; sedimentation of natural ponds and pools; carrier of nutrients and toxic

compounds

Nitrogen Agricultural fertilisers; atmospheric deposition

Eutrophication; contamination of potable waters; acidification

PhosphorusSoil erosion; agricultural fertilisers;

urban runoff (detergents, organic material)

Eutrophication of fresh waters: Ecological degradation Blue green algae Increased need for filtrationCosts for potable reservoirs/rivers

Organic WastesAgricultural wastes (slurry, silage

effluent, dirty water); sewage sludge; industrial wastes for land application

Oxygen demand; nutrient enrichment

Faecal PathogensSeptic tank system failures; animal

faeces; application of organic wastes to land

Health risks; non-compliance with recreational water standards

PesticidesGolf course maintenance; municipal

applications; agriculture; private properties

Toxicity; contamination of potable supplies

Oil and HydrocarbonsCar maintenance; disposal of waste

oils; spills from storage and handling, traffic emissions; road

runoff; industrial emissions

Toxicity; contamination of urban stream sediments; groundwater

contamination; nuisance (surface waters); taste (potable supplies)

Trace Metals Urban runoff; application to land of industrial and sewage sludge Toxicity

Iron Water table rebound following mining Toxicity; aesthetic nuisance

Initially, a literature review on pollutant behaviour and its mitigation has been undertaken. Within this, pollutants have been described in terms of three conceptual locations at which mitigation measures may be applied. These are the ‘SOURCE’ of the pollutant, the processes by which the pollutant is ‘MOBILISED’, and how the pollutant is ‘DELIVERED’ to surface waters. Mitigation options for targeting the pollutants within England and Wales have then been highlighted which is followed by the potential problem of pollution swapping when tackling the pollutants.

Using the source, mobilisation and delivery concepts (Haygarth et al., in press), and the information drawn from the literature review, a conceptual model describing diffuse pollutant on the basis of their characteristics has been proposed. Using this model, a mitigation matrix (Appendix I) has been produced. Mitigation measures targeting pollutants in their source, mobilisation and delivery conceptual locations

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have been scored for their effect on pollutants, cost per hectare, public and farmer acceptability, conflicts with other measures, pollution swapping and uncertainty of the effect of the measure. A list of farm systems is presented in Appendix II. The systems that each measure can be applied to are also included within the mitigation matrix.

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2. AMMONIUM

1. INTRODUCTION

Ammonium (NH4+) is applied to agricultural land to promote plant growth and

optimise agricultural yields. It can be applied in inorganic forms as fertilisers or via animal manures and other organic residues and effluents. When applied at times and rates to satisfy crop demands, NH4

+ is a valuable resource. However, if applied when there is little or no crop demand and at rates greater than the crop can utilise, then there is increased risk of transfer to watercourses. An added complication is that NH4

+ is released from organic nitrogen (N) forms (from soil or from added organic residues) by the process of mineralization. Hence, high rates of organic N additions can also result in release of NH4

+ in excess of crop demands.

Ammonium transfers to watercourses can supply the aquatic environment with a limiting nutrient for algal growth, and thus encourage eutrophication problems. Also, dissolved ammonia (NH3) is directly toxic to fresh water fish. The Freshwater Fish Directive (E.C., 1978: 78/659/EEC) was implemented to protect and improve the quality of fresh waters in order to support fish life, particularly Salmonids and Cyprinids. Fresh water quality is assessed according to pH, temperature and concentrations of dissolved oxygen, suspended solids, biological oxygen demand (BOD), total phosphorus, nitrites, phenolic compounds, petroleum hydrocarbons, chlorine, zinc, copper, non-ionised NH3 and ammonium-nitrogen (NH4

+-N).

When one molecule of NH3 dissolves in water it reacts to form ammonium hydroxide, which dissociates completely to give an NH4

+ ion and a hydroxyl ion.

NH3 + H2O NH4+ + OH-

The reaction is reversible and at higher pH values (i.e. more alkaline), the increase in hydroxyl ions alters the equilibrium and increases the proportion in solution of non-ionised (free) NH3 (highly toxic to fish) to ionised NH4

+ (virtually non-toxic to fish). Ammonia may be converted in water courses when nitrified to nitrite (NO2

-) (also very toxic to fish) and thence to nitrate (NO3

-), but both these processes of oxidation also remove oxygen from water.

In order to reduce the risk of direct toxicity of non-ionised NH3, oxygen depletion through nitrification and eutrophication, the Freshwater Fish Directive (FWFD) has set mandatory threshold concentrations for total NH4

+-N of 0.78 mg l-1. Guideline levels of 0.03 and 0.16 mg l-1 of total NH4

+-N have also been set for Salmonid and Cyprinid fish, respectively.

Ammonium transfers from agriculture to watercourses was the subject of a recent Defra review (Chadwick et al., 2003). This review is summarised in this section and amended with more recent information.

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2. DEFINING THE PROBLEM

2.1. SOURCESThere are several potential sources of the NH4

+ and non-ionised NH3 that can be transferred to watercourses, including managed manures, outdoor livestock, inorganic fertilizers, silage clamps, atmospheric deposition and mineralisation of organic forms of N in agricultural soils. Livestock and manure management are of principal concern, although atmospheric deposition and inorganic fertiliser are also sources.

2.1.1. Managed Manures

i. Dirty WaterDirty water consists mainly of parlour and dairy wash water as well as runoff from soiled yard areas (collection and dispersal yards, open feed areas and loafing yards). On some farms, silage effluent, the liquor from midden areas and slurry storage may also be included in the dirty water system. Therefore, dirty water quality is likely to vary considerably. Cumby et al., (1999) collected dirty water samples from 20 UK dairy farms on three occasions during the year (February, June and September) and the average NH4

+-N analyses were 310 mg l-1 (sd 106%), 580 mg l-1 (sd 84%) and 480 mg l-1 (sd 86%), respectively. Concentrations varied from under 100 to 2000 mg l-1, with a median value of approximately 380 mg l-1. There were seasonal variations on individual farms, as well as large differences between farms.

The average hard standings (dirty yard) area in this study was 1070 m2 per farm. Annual volumes of dirty water ranged from 191 to 6080 m3 per farm, so on the basis of the median NH4

+-N analysis, the notional annual “export” of NH4+-N in the dirty

water may be as high as 2310 kg. This is probably unrealistically high, but attributing the high and low ranges in concentration to the low and high range in dirty water export volumes, respectively, suggests an export of NH4

+-N in the range 380-610 kg, which may be feasible. In relation to the typical housing production of slurry N for a 200 cow dairy herd (i.e. 13.5 t total N, containing 6.7 t NH4

+-N), this latter estimate of dirty water NH4

+-N appears reasonable.

Research has, in the past, sought to quantify the extent, pathways and nature of pollution from farmland, buildings and livestock manure management, but has tended not to consider hard standings. A recent study (WA0516) collected a limited number of runoff samples from 4 hard standing types: dairy cow collecting/feeding yards, sheep handling yards, general purpose yard (arable farm) and sugar beet storage area (following removal of the sugar beet). These measurements indicate that concentrations of potential pollutants in runoff from hard standings require that runoff is either collected or adequately treated before discharge to a watercourse (Table 2).

The mean and range in NH4+-N concentrations are not dissimilar to those reported

from the study on dirty water. Moreover, taking the average dirty yard area from the study of Cumby et al. (1999), a tentative estimate of potential dirty water NH4

+-N export can be made based on the analyses, excess winter rainfall and some likely runoff emission factors (Smith et al., 1984).

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On this basis, the potential export from a farm with 1070 m2 hard standings (foul yard) area over an average winter might be:

1070 m2 (yard area) x 0.25m (winter rainfall) x 0.85 (runoff factor) = 227m3

which at 239 mg l-1 NH4+-N would contain 54 kg of NH4

+-N. Another Defra project (WA0523), suggested that 80-90% of drainage from dairy hard standings was to a tank or lagoon, with c.50% from beef hard standings. The remainder is presumed to drain into proximate fields and ditches, although better information is required on this aspect.

Table 2. Concentrations of potential pollutants in runoff from hard standings.

% Dry matter

NH4+-N

mg l-1

NO3--N

mg l-1

Total Nmg l-1

MRPmg l-1

Total Pmg l-1

BODmg l-1

Dairy cow yards

0.16(0-0.92)

239(57-603)

0.0(0-0.2)

1021(160-2960)

28(9-82)

54(12-115)

2955(1-7300)

Sheep handling 0.10 7 0.0 11 12 13 191

General purpose yard

0.01(0-0.01)

54(45-63)

42.0(0-83.9)

39(38-40)

0.1(0-0.2)

3(0-5) ND

Sugar beet storage area

0.21(0.17-0.25)

1(0-2)

22.2(19.6-24.8)

ND ND ND ND

Values in parentheses are ranges. ND = not determined.

ii. Slurry StoresA considerable proportion of manure is stored as slurry in the UK (Chadwick et al., 1999). The main risk of NH4

+ transfers from slurry stores to watercourses is by failure of the store structure. There is also a risk from over filling of stores, where farmers have inadequate storage capacity.

Slurry store type and the farm's slurry storage capacity influence both the risk of overflow and the flexibility the farmer has to spread slurry to land. Slurry storage capacity in dairy units is commonly between 3-4 months (the Code of Good Agricultural Practice for the Protection of Water recommends 4 months storage capacity). Surprisingly, over 15% of the farms in a recent survey (Smith et al., 2001c) had little or no storage. On many of these units there is only a small below ground tank, which is emptied and spread to land on a daily basis. This type of unit will have an increased risk of overflowing and slurry is likely to be applied at times when soil and weather conditions are not conducive to high nutrient utilisation, resulting in a greater risk contamination of watercourses with NH4

+. iii. Septic TanksAs long as the structural integrity and management of septic tanks is sound, then they should not pose a significant threat to water quality.

iv. Solid Manure HeapsManure heaps on concrete represent a reduced risk of contaminating watercourses with nutrients such as NH4

+, if any effluent arising from these heaps is collected and disposed of carefully. It is the field heaps that provide the greatest source of NH4

+

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transfers to watercourses. The majority of solid manure is stacked in field heaps (Table 3) as this is the cheaper option.

Table 3. Percentage of solid manure stacked on a concrete base or on a field in England (Scott et al., 2002).

Concrete Heaps

Field Heaps

Dairy cows 25 75Beef cattle 20 80

Pigs 35 65Poultry 0 100

The Code of Good Agricultural Practice for the protection of water (MAFF, 1998a) advises that field heaps should not be sited over field drains, within 10 m of a watercourse, or 50 m of a potable watercourse. However, many farmers are unaware of the old drainage systems within their fields and hence field heaps sited over, or close to drains could result in transfers of NH4

+ to watercourses.

Concentrations of NH4+-N leaching from pig manure heaps over a 6-month period

ranged from 10 to 1850 mg l-1 (Defra project WA0716), although the mass of NH4+-N

lost from the heap was small (0.23 kg t-1) and represented <1% of the N content of the manure. In the same study, the NH4

+-N concentration in leachates from broiler litter heaps ranged from 38 to 11800 mg l-1. High concentrations (>2000 mg l-1) of NH4

+-N in leachates from broiler litter heaps were also measured in another Defra project (WA0712). Concentrations leaching from layer manure heaps were generally lower (<2000 mg l-1). These observations raise particular concerns about the polluting potential of these sources in relation to the standards specified within the Freshwater Fish Directive.

Studies carried out in the Irvine and Girvan catchments in the West of Scotland (Aitken, 2003), indicated that the majority (58%) of all farms had middens, but nearly half of these had no containment, resulting in possible discharge of effluent (slurry) to a watercourse. Furthermore, some material contained in middens was considered unsuitable for this form of storage, i.e. manures from livestock fed on a silage-based diet with minimal bedding, as opposed to traditional solid and stackable farmyard manure. A detailed survey undertaken on a sub-sample of 20 farms indicated that 60% of them had significant discharges to watercourses.

2.1.2. Outdoor Livestock

Outdoor livestock (primarily cattle, sheep and pigs) deposit urine and faeces directly onto fields, tracks and into streams/rivers. Table 4 summarises the estimated quantity of NH4

+-N from this source. Outdoor poultry are not considered to be a significant source, since such a small percentage of land area is used for this purpose.

Urine is the principal source of NH4+ and is generated after hydrolysis of excreted

urea in urine. This process is controlled by the ubiquitous enzyme, urease. In contrast, faeces have a low NH4

+ content but contain significant quantities of N in the organic form, which can then mineralise in/on the soil and release NH4

+.

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Table 4. Total NH4+-N loading by outdoor animals.

No. of animals in the UK1

(x 1000)

No. of animals

with access to outdoors1

(x 1000)

Typical period of time spent outdoors per year1

(days)

Typical N loading per day2

(kg/d)

Proportion of total N as

NH4-N 3

(%)

Total loading of NH4-N by animals outdoors

(t/yr)Dairy cows

(Cows, heifers, heifers in calf)

2698 150* 0.263 -0.318 50 62,399

Beef cattle(Sucklers,

growers/fatteners, Calf)7648 183 0.039 -

0.159 50 73,720

Sheep and lambs 35834 180-365 0.003 -0.025 50 89,189

Breeding sows 638 185 365 0.053 60 2,147TOTAL 227,455

1 – UK Ammonia Inventory (2003)2 – MAFF (2001). 3 – Chambers et al. (1999) * Time spent grazing by dairy cows takes into account the time spent indoors during milking, time in collection yards as well as time in the building during the over-wintering period

2.1.3. Inorganic Fertilisers

Table 5 shows the typical content of NH4+ and NO3

- in many inorganic-N fertilizers and the fraction that each contributed of the total N fertilizer used within the UK for 2001/02. During this period 1203 x 103 t of N was applied in the form of inorganic fertilizer (International Fertilizer Industry Association, 2004) of which 743 x 103 t was NH4

+-N.

Table 5. Inorganic N fertilizer NH4+ and NO3

- content and use within the UK.

Type of fertilizer§ % N as

# x 103 t artificial N used in UK

2001/02

x 103 t N in the form of

NO3- NH4

+ NO3- NH4

+

Calcium ammonium nitrate 50 50 26 13 13Ammonium nitrate 50 50 550 275 275

Ammonium sulphate 100 6 0 6Urea 100a 136 0 136

Ammonium phosphate 100 20 0 20Nitrogen solutions 25b 75b 92 23 69NPK compound 40b 60b 373 149 224

§ Taken from Bøckman et al. (1990) unless otherwise stated.# International Fertilizer Industry Association (2004). a Assuming all urea is converted to NH4

+ (Bristow et al. 1992).b Chambers, B. ADAS Gleadthorpe (pers com 2005).

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2.1.4. Silage Clamps

Silage clamps and big bale silage can generate considerable quantities of effluent if silage is cut and not wilted sufficiently. Values in the literature indicate a range of 1000-5000 mg l-1 total N in fresh effluent (McDonald et al., 1960, 1991; Purves and McDonald, 1963; Stewart and McCullough, 1974; Haland 1979; Jones et al., 1990). Stewart and McCullough (1974) concluded that the total N of silage effluent comprised mainly of amino acids and amides with about 10% of the total N in the form of simple NH3. The range of NH4

+-N observed by McDonald et al. (1991) were 120 mg l-1 at 3 days after ensiling to 510 mg l-1 after 60 days ensiling.

The dry matter (DM) content of baled silage averages 30% in the UK, some 5% higher than clamp silage. Very little silage effluent would be expected from a crop of 30% DM content. However Haigh (1992) has calculated that the total silage effluent from bales in England and Wales during 1983-87 accounted for 1.8 million litres, compared with 77 million litres from clamp (bunker) silos. Clamp silage is made on concrete with collection facilities. In some instances, silage effluent is collected in dirty water or slurry tanks and is thought to be partly responsible for the increased NH4

+-N concentrations of these effluents in the summer months (Cumby et al., 1999).

2.1.5. Atmospheric Deposition

The livestock sector is responsible for 85% of the NH3 emissions in the UK (Pain et al., 1998). Much of this emitted NH3 is deposited close to the source of emission. Hence, with a regionalised livestock industry (e.g. dairying in the West and pigs and poultry in the East), it would be reasonable to expect that deposition of reduced N forms will be greater around livestock units. However, on a broader scale, other factors will determine the actual deposition rates, e.g. vegetation type, topography, rainfall etc. (Defra, 2002).

2.1.6. Mineralisation

Ammonium-nitrogen is released from forms of organic N as a result of the process of mineralization. Mineralisation and immobilisation control the release of NH4

+ from organic N found in soil organic matter and organic amendments to the soil, e.g. animal manures and crop residues. The rate of N mineralization from these sources is controlled by the decomposer organisms, the physical climate (temperature, moisture etc.) and the chemical composition of the organic material (Swift et al., 1979).

2.1.6.1. Soil organic matterAgricultural soils contain significant amounts of organic matter (SOM) and N. The baseline quantity of native SOM depends on drainage and texture status of the soil, cropping history, climate and topography. In general, SOM accumulates in grassland soils and declines in arable soils. Mineral N supply from SOM can constitute a large proportion of total crop uptake, so it important that farmers take this into account when planning fertiliser and manure applications to prevent accumulation of N in the soil which is then at risk of transfer to water and air. Generally, the greater the SOM content the greater the N supply from the SOM (Barreto and Bell, 1995), although this

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relationship is weakened by the different nature (composition) of the organic matter and differences in climatic and soil conditions. Therefore, the accuracy of predicting N supply from SOM is limited.

The release of NH4+ from SOM is not likely to result in diffuse pollution of

watercourses since release is likely to coincide with either crop uptake or nitrification to NO3

-. However, losses of NO3- and potentially NO2

- to water courses are possible, particularly if a flush of mineralization is produced, e.g. cultivating grassland.

i. Cultivation of grasslandGrass residues contain significant quantities of organic N and the quantity released on ploughing will depend on the age of the sward and its previous fertiliser and manure management. After cultivation large amounts of organic N can be mineralised and large amounts of mineral N are released often in excess of the requirements of the next crop.

2.1.6.2. Organic amendmentsOrganic amendments such as animal manures and crop residues supply organic matter which releases N more rapidly than SOM. Evenness of spread is an important factor controlling the proportion of applied N taken up by the crop.

i. Animal manuresAnimal manures vary in their total N content and form (Chadwick et al., 2000) because of differences in feed, feed conversion by different animal species, age of animal, bedding material and water intake. Typical values of total N and plant available N in manures are provided for farmers (Defra, 2000), but because of the large variability between manures from even the same animal species, published values are subject to large uncertainty. In general, between 40 and 90% of the total N content of solid manures and slurries is present in organic forms, which are more slowly available for plant uptake than the mineral fraction (Chadwick et al., 2000; Defra, 2000).

Chadwick et al., (2000) showed that N mineralization was greater from manures with low C:organic N ratios (e.g. layer manure and pig slurry) than manures with high C:organic N ratios (e.g. beef FYM and dairy slurry). Relationships between manure C:N ratios and N mineralization have been reported elsewhere, Serna and Pomares (1991) demonstrated a significant relationship between C:N ratio and N mineralisation, r = -0.69 and Floate (1970) showed a weak relationship between the C:N ratio of sheep faeces and N mineralised, r = -0.56. Although, Castellanos & Pratt (1981) found no relationship between manure C:N ratios and N mineralization for a range of stored and fresh animal manures.

Generally, organic materials with C:organic N ratios of 15 or more will initially immobilize N, whilst C:organic N ratios of <15 will result in mineralization, and the lower the C:N ratio the greater the mineralization rate (Kirchmann, 1985; Beauchamp, 1986; Mary and Recous, 1994; Jarvis et al., 1996). Clearly, for those manures with low C:organic N ratios (e.g. poultry manures and pig slurries) it would be desirable not to apply these in the autumn as the organic N they contain is likely to mineralize relatively rapidly and may be at risk from loss by NO3

- leaching. The likelihood of losses in the form of NH4

+ is low. Applying these manures in late winter-early spring is likely to make best use of mineral N and mineralized organic N.

16

ii. Crop residuesN pulses from additions of crop residues supplement background SOM N mineralisation. The N in such residues is released more rapidly than from SOM and depends on quantities of residue returned, residue C:N ratio and physical structure of the residue. The quantity of N released depends on the proportion of the crop harvested (the N harvest index) and the growth of the crop during the season. Sometimes a poor harvest due to pests, diseases and drought may lead to high N contents in non-harvested parts (Powlson et al., 1992). Crops that have received generous N fertiliser applications will return larger residues (Glendining et al., 1992).

According to Shepherd et al. (1996) between 20 and 145 kg N ha-1 can be returned to the soil in arable crop residues. This excludes N returned in roots or in root exudates. The ADAS N index system (Defra, 2000) provides advice on what allowances to make for previous crop’s residues. Again, the risk of losses in the form of NH4

+ from crop residues is low since nitrification will also proceed in the soil generating NO3

-.

2.2. MOBILISATION

Direct urination and defecation by grazing livestock into streams poses a direct risk of NH4

+ transfer to watercourses. Often livestock are given free access to streams and rivers which are used as the sole means of watering livestock. However, farmers are generally concerned with the erosion of riverbanks and there has been a move towards fencing off sections of streams to prevent this. It is unknown to what extent this practice is still used.

Accidental applications of manures and dirty waters to streams and ditches can occur if a farmer is unaware of the range the manure is spread from equipment or if a farmer chooses to ignore or is unaware of the COGAP for Protection of Water (MAFF, 1998a) advice that applications of manures should not take place within 10 m of a watercourse and 50 m of a potable water body.

Ammonium is soluble and will form salts in soil water once it has dissolved from either, NH4

+ based fertilisers, urine deposited on the soil, or mineralised from soil organic matter or organic amendments. Thus a proportion of the NH4

+ in the soil is at risk of transfer to watercourses in the dissolved form via routes of high connectivity (lateral, subsurface, and preferential flow).

The NH4+ ion can displace other ions on the soil surfaces and become fixed to clays.

The literature suggests that only a relatively small proportion of NH4+ in the soil is

fixed in this way. Trehan (1994) reported that only 5% of cattle slurry NH4+-N was

fixed by soil following slurry spreading. So, transfers of NH4+ via detachment of soil

and colloids are likely to small.

Some NH4+ may also be immobilised in the soil by the microbial biomass, particularly

if amendments with high C:N ratios (>15) are made to the soil (Powlson et al., 1994; Chadwick et al., 2000). Trehan (1994) reported that 34% of the 15N labelled NH4

+-N content of cattle slurry had become immobilised into organic forms within 9 days of

17

application. This was considerably greater than the 6-8% measured by Chadwick et al. (2001b).

In the soluble phase, NH4+ will leach down through the soil profile but because of its

positive charge will be attracted and adsorbed to soil particles and hence will be retained within the soil. This will provide an opportunity for the NH4

+ to be nitrified to NO3

-, therefore, leaching losses of NH4+ from all of the major sources tend to be

small.

2.3. DELIVERY

Ammonium derived from manures and other sources can be transported to watercourses in a number of ways, for example, in flow over the surface of the land, either in dissolved or particulate forms. The soil surface can be sufficiently unstable and the hydraulic energy (rainfall or application rate) sufficiently great, to cause erosion of soil particles that may also be transferred down-slope in an overland flow. Potential pollutants can also be transferred via vertical flow through the soil profile to ground water or via drains to surface waters. From the data available, we are unable to discriminate between surface runoff and sub-surface flow. Evidence of this route of transfer is presented along with drain flow, storage losses and leaching.

2.3.1. Surface pathways

i. Surface runoff from fieldsOverland flow generally accounts for only a small proportion of applied N, compared with the proportion transferred in percolates through the profile (Burgoa et al., 1993). Blevins et al. (1996) found that after two growing seasons, less than 2% of fertilizer-N was lost to overland flow, whereas 30% had moved below 1 m in the soil.

Nitrogen in any manure solids left on the soil surface, or associated with fine particles that are readily moved during soil erosion, can be lost through surface runoff to a watercourse. The factors that determine N loss by erosion are: the amount of sediment moved, the N content of the soil moved, and the N content of the manure solids. The other material subject to loss to surface water is N dissolved in runoff water. This portion is usually small (Meisinger and Randall, 1991; Blevins et al., 1996). However, it is variable, and depends on a number of factors, such as the degree of soil cover, source of N applied, application rate, timing and duration of the application. Surface conditions are also important, and affected by slope, soil characteristics, and land management. Finally, runoff is highly dependent on the intensity of rainfall after application. The largest losses occur if a soluble N source is applied to a bare soil surface, and a significant rainfall event occurs soon after application (Edwards and Daniel, 1993; Sharpley, 1997). In many cases, dissolved N is transported into the soil with the initial infiltration that precedes runoff (Meisinger and Randall, 1991). Incorporating the N source such as manure dramatically reduces runoff losses. In most cases, runoff N losses are small, 3 kg N ha -1 annually or less (Meisinger and Randall, 1991; Nichols et al., 1994; Blevins et al., 1996).

Sharpley (1997) investigated NH4+-N in overland flow on ten Oklahoma (USA) soils

amended with poultry litter. Ammonium-nitrogen concentrations decreased with 10

18

successive rain events, starting seven days after litter application. Increasing the time between litter application and rainfall from 1 to 35 days reduced NH4

+-N from 5.5 to 0.1 mg l-1.

Concentrations of NH4+-N are consistently reduced as runoff passes over the soil

surface or interflow passes through the upper soil layer (Chadwick and Pain, 1998), probably due to immobilisation on surface exchange sites on the soil surface and dilution by soil water. (Some rapid nitrification may have also been responsible for the decrease in NH4

+-N concentrations).

In the same study (NRA, 1994) it was shown that NH4+-N concentrations were greater

in leachates leaving surface lysimeters on a freely draining medium loam soil (mean reductions of 35% for NH4

+-N) and least in leachates from lysimeters on an underdrained heavy loam soil (mean reductions of 49% for NH4

+-N), a result of the greater number of exchange sites.

An important question is then: Is the 10 m-rule (highlighted in the COGAP for Protection of Water, MAFF, 1998a) sufficient to protect watercourses under all soil and weather conditions? Study WA0503 concluded that in general, that if effluent passed through the soil matrix it had lower concentrations of NH4

+-N than effluent that passed over the soil surface, since there was greater opportunity for adsorption of NH4

+ on a greater number of exchange sites. Also, slope was an important factor in controlling the proportion of NH4

+-N removed from the runoff by the soil.Smith et al. (2001b) demonstrated that slurry solids are an important factor that influenced runoff potential. Increasing slurry application rate and, in particular, slurry solids loading, increased solids and NH4

+-N losses via surface runoff. The threshold, above which the risk of losses via surface runoff appeared to be greatly increased, was c. 2.5-3.0 t ha-1 slurry solids, which approximates to the 50 m3 ha-1 limit suggested for slurry within the Code of Good Agricultural Practice for the protection of water (MAFF, 1998b). Sealing of the soil surface by slurry solids appeared to be the probable mechanism generating surface runoff following slurry application.

Other important factors influencing NH4+ transfers to water following manure and

dirty water applications include the soil moisture status at the time of application (cracks and macropore flow) and drainage.

ii. Track runoffDairy tracks are used twice per day and at an increased stocking rate than at pasture. A proportion of dung and urine are voided on these impermeable surfaces. Thus, sufficient rainfall can transfer any deposited NH4

+ towards proximate ditches and fields. Estimates of the N deposited on tracks will vary according to the distance between the dairy parlour and grazing paddock. Detailed measurements in New Zealand (NZ) (Ledgard, pers. comm.) have shown that on average, dairy cows spend about 3% of their time (annual average) in transit to the milking parlour and that excretal deposition corresponds to this. Thus, N in excreta to farm tracks and lanes (which typically constitute 1.5% of farm area) can be large, e.g. equivalent to about 9 kg N ha-1 yr-1 (total N in dung and urine) transferred from the grazed area, or about 600 kg N ha-1 yr-1 on the lanes for an average NZ dairy farm. In addition to dung and urine directly voided onto tracks from animals extra manures may also arrive as a result of accidental spillages from manure spreading practices.

19

The fate of the NH4+ will depend on soil texture (i.e. runoff) and the proximity of

watercourses. Based on the findings and experience of hard standings, tracks may be a significant source of both pathogens and NH4

+ and therefore, require further investigation.

2.3.2. Preferential pathways

i. Drain flow losses from fields

- CattleAmmonium loss from grazed land appears to be relatively small. From the Rowden Drainage Experiment (Tyson et al., 1993), mean concentrations in all plots (receiving N inputs ranging from zero to 400 kg N ha-1 yr-1) were below 1.5 mg l-1 NH4

+-N and the maximum concentration recorded was under 2.5 mg l-1 NH4

+-N (400 kg N drained plot. Total N lost in surface runoff as NH4

+-N was <6 kg N ha-1 and losses from the plots were increased with field drainage by up to two-fold. During this experiment, all the plots were grazed, but no additional animal manures were applied.

From 1993 to 1996 (inclusive), the Rowden Drainage Experiment was operated as a series of individual small farms (farmlets) with each 1 ha plot representing the whole grassland farming cycle of cutting (silage), grazing (cattle) and with additions of fertilizer and slurry. Despite this more intensive regime, peak concentrations of NH4

+-N remained below 1 mg l-1 and mean concentrations were <0.5 mg l-1, even in the highest N plots (280 kg N ha-1). The total N leached in the form of NH4

+-N was <2 kg N ha-1. In the winter of 1996-7, the peak concentration in these high N plots just exceeded 3 mg l-1 NH4

+-N with overall leaching of about 4 kg N ha-1 in the ammoniacal form.

From 1998-2001, further N studies were conducted on farmlets (Defra project NT1829) and in the four successive over-winter periods (1998-9, 1999-00, 2000-1, 2001-2), mean concentrations remained below 1 mg l-1. There was, however, an increasing tendency for drainage to continue beyond springtime, so that some leaching coincided (or followed closely) with fertilizer applications.

- SheepCuttle and James (1995) measured NH4

+-N in drainage from a reseeded upland pasture and an adjoining 'unimproved' (Molinia dominated) pasture on a cambic stagnohumic gley soil (Wilcocks series) at an altitude of 400 m in mid-Wales. The reseeded area received 103 kg fertilizer-N ha-1 between April and August in the reseeding year and 65 kg N ha-1 in April in each of the following three years. No fertilizer was applied to the Molinia area. Both areas were grazed by sheep between April and October, with average stocking rates of 15 and 5 sheep ha -1 on the reseeded and Molinia areas, respectively.

Concentrations of both NO3--N and NH4

+-N in water from the Molinia area were in the range 0.01 to 1.8 mg l-1. Concentrations from the reseeded area were generally similar and showed similar fluctuations except following fertilizer applications when they increased for short periods, when the NH4

+-N content of the bulk water sample reached a peak of 3.7 mg l-1, two weeks after the fertilizer application. Concentrations

20

fell to background levels within 1-2 weeks. Annual drainage was equivalent to between 1020 and 1768 mm (mean 1522 mm yr-1). Total quantities of mineral-N (NO3

- + NH4+) leached each year were 3.0 - 4.1 kg N ha-1 from the Molinia area and

2.7-10.0 kg ha-1 from the reseeded area, with the greatest loss from the latter in Year 2.

- Outdoor pigsOver 25% of the breeding herd is now raised outdoors. Because of the treading damage and for welfare reasons, the outdoor pig herds are generally sited on freely draining soil. The behaviour of the animals disturbs the soil surface; hence, land supporting outdoor pigs is vulnerable to runoff and soil erosion losses. High surface soil NH4

+-N concentrations, coupled with a vulnerability to erosion, mean that there is a significant risk of NH4

+ transfer to surface watercourses. However, leaching losses of NH4

+ are not likely to be significant on these freely draining soils.

- Slurry StorageEarth banked lagoons represent a particular risk. A Defra-funded BGS/ADAS study (WA0517) used boreholes on 8 sites on sandstone and chalk to assess the extent of contamination emanating from unlined manure and slurry stores. Porewater was extracted and analysed. Ammonium-nitrogen concentrations in porewater at depths >10m were zero on two sites, between 0.2 and 8.6 mg l-1 on three sites and 83 to 622 mg l-1 on the remaining three sites. However, the results from the field studies and contaminant modelling suggested that, in the majority of situations, unlined earth-based slurry stores pose little threat to potable groundwater drinking supplies (Gooddy et al., 2001).

However, earth banked unlined slurry lagoons/pits over chalk have been shown to pose a risk to ground water. Self-sealing of the lagoon floor by slurry solids will normally occur rapidly, thereby preventing groundwater pollution; however disruption of the self-sealing layer is a serious concern unless emptying is carefully managed. Physical evidence of slurry contamination was observed along fissure walls during drilling studies in the chalk beneath the lagoon in 1993 (Gooddy et al., 1998).

Ritter and Chirnside (1990) also reported significant groundwater pollution when the clay lining of an earth-banked lagoon had been allowed to crack. In a survey study of waste-handling facilities on livestock farms in Great Britain in the early 1990’s, 14% of dairy farms used excavated, unlined earth-banked pits or lagoons for slurry storage (Colman et al., 1994). Nicholson and Brewer (1997) estimated that up to 16 t x 106 of slurry are held in earth-banked structures, in England and Wales. Such stores may represent a continuing pollution hazard, if not carefully managed. However, such risks are likely to reduce in the future with implementation of the Control of Pollution (Silage, Slurry and Agricultural Fuel Oil) Regulations (Anon., 1991). These regulations state that the walls and bases of earth-banked slurry stores should be ‘impermeable’, which is defined by Mason (1992) as a permeability of 10-9 m s-1, with minimum thickness of 1 m.

- Dirty Water SpreadingImpermeable cracking clay soils are especially difficult to manage in terms of water movement and contaminant transfers. In the winter they pose a risk of runoff and drain flow transfers of NH4

+ to watercourses and also in the summer, preferential flow

21

of effluent down cracks can rapidly transfer contaminants to drains. A project sponsored by the National Rivers Authority (N.R.A., 1994) was set up to determine under what soil and weather conditions, could dirty water be safely applied to under-drained land. The results show that effluent contamination of drain flow can occur at any time of the year even at low application rates. Some concentrations were extremely high (>500 mg l-1).

It is clear that the application of dirty water to drained land is associated with significant pollution risk, both under dry soil conditions and with the soil at or above field capacity. Drainage systems with permeable fill represents the greatest risk. The results of these studies confirm that great care is needed to ensure careful management of dirty water applications to drained land if significant water pollution is to be avoided.

- Slurry SpreadingThere are in the region of 6.4 million hectares of drained soils in England and Wales (Withers et al., 2000). The majority of schemes (60%) have been installed on medium and heavy textured soils to correct surface wetness problems. On clay soils, manures are commonly applied in the autumn when soils are dry and can carry the weight of heavy application machinery without causing damage to the soil structure. However, autumn application timings are widely recognised as presenting the biggest potential risk of diffuse nutrient pollution, as crop nutrient uptakes are low and over winter drainage means that applied manure nutrients can be lost from the soil via percolation to ground waters in drainflow and surface runoff. Whilst clay soils are generally considered retentive of N and P, there is clearly a risk of nutrient loss via rapid macropore (by-pass) flow to drains, especially in soils with mole drains and piped systems with overlying permeable fill.

Application timings and cultivation can influence NH4+ transfers to drains (Williams

et al., 2002). Losses of manure N (treatment – untreated control) between September 1999 and March 2000 were greatest (9% of total N applied) following autumn slurry application to uncultivated stubble (ploughed the following spring). Losses of N were lowest (2% of total N applied) following autumn application to the autumn cultivated (ploughed or disced) treatments. The observed reduction in N loss on the autumn cultivated treatments was probably due to the slurry being mixed within the soil matrix, thereby reducing the scope for rapid preferential flow (by-pass) of slurry nutrients directly to the drains. Nitrogen losses following the winter timing were intermediate between the autumn ploughed and uncultivated stubble treatments. Around 90% of N losses from all application timings were as NO3

--N, with soluble organic N and NH4

+-N making up c. 9% and 1% of total N losses, respectively.

Although total N loss, as NH4+-N, was relatively small in this study, peak

concentrations in drainage events following slurry (and fertilizer) applications were sufficiently high to be a cause for concern. Ammonium-nitrogen concentrations peaking at c. 3 mg l-1, suggested some “memory effect” following the September slurry application. However, the higher peak concentrations in April and May, following spring fertilizer N applications indicate the importance also of this latter source. The results indicate that cultivation before the start of winter drainage is an effective method of minimising N and P losses via drainflow following autumn slurry applications to drained arable clay soils.

22

- Solid Manure SpreadingThe NH4

+ content of solid manures typically represents <25% of the total N content when fresh, and <10% after storage for 3-6 months. The fact that it contains relatively low concentrations of NH4

+ (in comparison to animal slurries) and that it has a high dry matter content (25-60%), means that solid manure spreading represents little risk of transfers of NH4

+ to watercourses. The exception could be applications of poultry manure (which contain uric acid), where hydrolysis to urea and NH4

+ could result in large NH4

+ concentrations in the soil.

2.3.3. Through-flow pathways

- Cattle and sheep grazingMore-recent studies have been sited on freely-drained soils where water samples have been collected using ceramic cup samplers. Although these samplers are suitable for studies of NO3

- leaching, they are less suitable for measurements of NH4+ because of

the cation exchange properties of the ceramic material. However, concurrent measurements of NO3

- and various cations indicate that measured concentrations of cations do fluctuate in response to concentration changes in the surrounding soil, though response times are likely to be influenced by differences in the cation exchange properties of the soil and ceramic cup. As these relationships have not been examined in detail, measurements of NH4

+ concentrations obtained with ceramic cup samplers should be treated with caution.

Cuttle et al., (1992) conducted a study on a typical brown earth soil (East Keswick series) at IGER, Aberystwyth and compared leaching from recently-sown grass/clover and pure grass pastures. There were 60 samplers in each pasture type. The ryegrass/white clover pasture received no N fertilizer, whereas the ryegrass pasture received 150-200 kg fertilizer-N ha-1 yr-1 (as ammonium nitrate). Pastures were grazed by sheep between March/April and November. Over three years, NO3

--N concentrations in samples from different ceramic cup samplers ranged from <0.01 to 158 mg l-1. The higher concentrations were assumed to be from areas where urine had been deposited. In spite of the very high concentrations of NO3

--N at some sampling positions, NH4

+-N concentrations were generally <0.05 mg l-1 and showed no relationship to the NO3

--N concentrations). Patterns were similar for the grass/clover and fertilised grass pastures. Total drainage was between 454 and 696 mm yr-1. Mean losses of NO3

- were 9-17 kg N ha-1 yr-1 from the grass/clover plots and 2-23 kg ha-1 from the fertilised grass plots. In comparison, losses of NH4

+-N were <0.2 kg N ha-1.

A similar experiment was carried out using ceramic cup samplers at an upland site at IGER, Bronydd Mawr, near Brecon. Soils were brown earths of the Milford series. Unfertilised grass/clover pasture was compared with grass-only pasture receiving 100 kg N ha -1 as ammonium nitrate fertilizer; all plots were grazed by sheep. At the Aberystwyth site, NO3

--N concentrations in water samples ranged from 0.01 to >100 mg N l-1, whereas NH4

+-N concentrations did not exceed 0.05 mg l-1. Total annual drainage during the 3 year investigation ranged from 836 to 1190 mm. Nitrate losses averaged over the whole pasture were equivalent to 13-24 and 10-13 kg N ha-1 yr-1

from the grass/clover and grass treatments, respectively.

23

More-recent studies have measured leaching from two grazed fields on the organic dairy farm at IGER, Trawsgoed (unpublished). In this case, there are only 10 samplers per field. Soils are brown earths over gravel (Rheidol series). Though the fields do not receive inorganic fertilizer, they receive dressings of cattle slurry in most years. Although most water samples contained similar concentrations of NH4

+-N to those in the earlier studies, samples from several samplers have contained much higher levels. This generally only affected one or two samplers per year, with different samplers affected in different years. In 1995-96, the peak NH4

+-N concentration was 0.8 mg l-1, but in 1996-1997 and 1997-98 peak concentrations reached 15.8 and 13.7 mg l-1, respectively. Typically, concentrations would increase over several weeks in December and then decline to background levels by March.

Urine patches are the main source of NO3--N leaching from grazed fields and other

IGER studies at Trawsgoed has specifically examined leaching from urine-affected areas. Ceramic cup samplers were installed in small field plots on the organic farm and on the adjacent conventional farm. Soils were again brown earths (Rheidol series). Cattle urine was applied to the plots on different dates and leaching measured over the following winter. Maximum NO3

--N concentrations from urine-treated areas exceeded 550 mg l-1; however NH4

+-N concentrations were generally similar to those for untreated plots (< 0.01-0.07 mg l-1). There was an exception with two of the 56 plots, where NH4

+-N concentrations increased for 2-3 weeks, reaching peak values of 27 and 14 mg l-1 before returning to background levels.

A previous study (Hatch et al., 1997) on 1 ha grazed grassland plots on poorly drained soil, demonstrated mean NH4

+-N concentrations were <0.5 mg l-1 in soil solutions obtained using ceramic cup samplers (60 cm depth) from land grazed by beef cattle.

- Manure spreadingWithin the Nitrate Sensitive Area (NSA) scheme in England, leachate samples were collected on a regular basis during the period of excess winter rainfall at selected sites within each of the designated NSAs (Lord et al., 1999). These samples were analysed for NO3

--N and NH4+-N on each occasion. Nitrogen leaching losses were calculated

as the sum of the cumulative NO3--N and NH4

+-N losses, but the latter were almost invariably negligible. The leachate NH4

+-N concentrations usually being at, or below the analytical detection limit of 0.05 mg l-1 NH4

+-N. However, there were very high NH4

+-N transfers from one site with a history of repeated and heavy pig manure applications within the Wildmoor NSA, near Bromsgrove, Worcs. In the first year of monitoring leachate, NH4

+-N concentrations reached values >40 mg l-1 in the autumn and early winter and, although NO3

--N concentrations were higher, NH4+-N

represented a significant contribution to total N leaching. This represents a likely “worst case” scenario since the naturally loamy sand topsoil had been rendered “organic” as a result of the high level of past manuring on the site.

Thereafter (i.e. more than one year into the study), whilst leachate NO3--N continued

to remain high (e.g. in 1991/92 and even in 1999/00, as a result of the high residues), the NH4

+-N concentrations were restored close to background levels. These data indicate that, only under exceptional ‘disposal’ practice, is it likely that any significant NH4

+-N leaching will occur and even then, such losses are unlikely to persist beyond the season following the discontinuation of the disposal practice.

24

Smith et al. (2002) measured NO3--N and NH4

+-N concentrations from ceramic cups from experimental plots receiving slurry at target rates of 200 kg N ha-1. Nitrogen leaching ranged from 0 to >50% of applied slurry N, with the highest losses occurring following applications in the September to November period.

Although the average leachate N concentration (and N loss) was significantly affected by application timing, NH4

+ leaching was consistently very low, with no apparent equivalent treatment timing affect. The NH4

+-N concentration represented, on average, c.1% or less, of the total leachate N concentration. These results confirm what would be anticipated intuitively. Under many circumstances it is likely that NH4

+ applied in manures, if not lost from the soil surface by volatilisation as gaseous NH3, will be rapidly nitrified or, at low soil temperature, will be held within the soil matrix by the cation exchange complex.

In a related arable investigation (Beckwith et al., 1998), generally similar results were obtained with very low NH4

+-N recovered in leachate samples compared with NO3--N:

the NH4+-N representing ~0.007% and ~1.8% of the total mineral N in leachate in

1990/1 and 1992/93, respectively. 2.3.4. Field Runoff and Leaching of Inorganic FertilizersThere are few reports of direct runoff or leaching of fertilizers from grassland or arable land. However, heavy summer rainfall falling on dry ground, immediately after fertilizer applications, can result in high concentrations of NH4

+-N in leachate, due to by-pass flow (Barraclough et al., 1983; Haigh and White, 1986). The maximum NH4

+-N concentrations in runoff, following early season applications (viz. March) of ammonium nitrate, or urea (both applied at rates of 86 kg N ha-1) were 1.7 and 4.5 mg l-1, respectively. Both types of fertilizer leached 3.0 kg ha-1as NH4

+-N, before the next fertilizer dressing had been applied (Scholefield and Stone, 1995). In a second experiment receiving 45 kg N ha-1 (applied as urea in February), <10% of the fertilizer N was leached, but nevertheless, the maximum NH4

+-N concentration exceeded 6 mg l-1 (Scholefield and Stone, 1992). Ammonium-nitrogen concentrations in spring drain flows from arable land at ADAS, Boxworth were up to 5 mg l-1.

3. MITIGATION MEASURES

Mitigation measures are indicated below for the principal sources of NH4+ transfers to

water, i.e. hard standings, manure stores, grazing/outdoor livestock, land spreading of manures, dirty water and other organic residues to land.

3.1. Hard Standings

Hard standings can be defined as any non-roofed area on the farm with a firm base that can support loads and would be likely to generate runoff during all but the lightest rainfall events. Hard standings would include areas such as dairy cattle collecting yards, feeding and stock handling areas and machinery washing areas. Such areas are usually constructed from concrete and are therefore, impermeable to deposited dung and urine. Yards are either scraped and/or washed to ensure high

25

levels of hygiene, particularly in the case of dairy collection yards, resulting in large volumes of dirty water.

Routine sampling from the limited number of dairy farm hard standings in Defra project WA0516, demonstrated that concentrations of pollutants found in yard runoff required that all runoff be either collected or adequately treated prior to discharge to a watercourse. The management of the effluent from hard-standings is critical to prevent transfers of pollutants, including NH4

+ to watercourses. Therefore: ALL effluent from hard standings must be collected and stored prior to

spreading

Volumes of dirty water requiring storage can be minimised by ensuring that, the area of dirty yards open to rainfall is limited wash water use is not excessive clean water from roofs is collected in well maintained gutters and directed

away from the dirty water store, preferably to a clean water store which can be used e.g. for cleaning down parlours etc.

3.2. Manure Stores

i. Slurry and dirty water The structural integrity of above ground tin and concrete tanks should be

inspected regularly to prevent catastrophic failure Bunds should be constructed around above ground tanks to ensure that any

spill/collapse is contained Farmers should ensure that slurry and dirty water stores are of sufficient

capacity to cope with the wet winter months when all slurry generated has to be stored and land spreading cannot take place

New earth banked slurry lagoons over chalk should be lined Stores should not be sited within 50 m of a well or borehole OR 10 m of a

watercourse

ii. Solid manure field heaps Heaps should be carefully sited away from drains, 10 m from a water course

and 50 m from a borehole or well Heaps should be constructed in an ‘A’ shape and not an ‘M’ shape to aid

rainfall shedding from the heap and not leaching through the heap Heaps should be covered if possible to prevent leaching of soluble nutrients

including NH4+. This can be achieved by constructing heaps under roofs (in

buildings) or by covering with plastic sheets.

iii. TreatmentTreatment of animal manures and dirty water offers the opportunity to remove potential problem pollutants prior to spreading on the land, i.e. to reduce the source term. Ammonium can be removed from slurries and dirty water by biological and physical processes,

Physical processes

26

Separation allows the liquid portion containing the NH4+ to be spread via

mobile and static irrigators some distance from the farmsteading, i.e. the NH4+

can be spread to a greater land area and hence reduce the risk of over application to one area of land.

Biological Processes Aeration – converts NH4

+ to NO3- and this can be further denitrified to N

gases. Aerators exist for slurries and dirty water, whilst active compositing may also result in more rapid conversion of NH4

+ to NO3- than by static piling

of solid manures. That said, static piling of solid manure in the open (i.e. uncovered) can result in very low NH4

+ content at the time of FYM spreading to land. This will depend on length of storage and rainfall.

Soil filters – dedicated areas of soil can be used to treat manure (Martinez, 1997b) and dirty water (Defra project - WA0518) and NH4

+. The soil acts as a nitrification reactor and converts NH4

+ to NO3-. Such soil systems can

generate high concentrations of NO3-.

3.3. Grazing/Outdoor Livestock

i. Cattle and sheep Prevent direct access of cattle and sheep into streams. Provide alternative

water supply Remove livestock from land when soil conditions become wet/waterlogged

and poaching becomes a risk. This will reduce the potential risk of transfer via erosion. Potential use of wood chip stand off pads where effluent is collected and stored until more suitable time for application.

Reduce stocking density Feeding and drinking troughs can be moved to drier, more vegetated areas

when the soil becomes compacted and/or badly poached Move livestock if mobbing areas under trees and against hedgerows become

waterlogged and compacted. This can be achieved through the use of fencing. Attention should be paid to animal welfare

ii. Outdoor pigs Remove pigs from areas which are badly disturbed and have evidence of

surface losses of sediment onto fresh land. Dp not return until vegetation cover has returned

Maintain a buffer of vegetated land between pig area and vulnerable watercourses to act as filters to reduce the transfer of sediment and pollutants such as NH4

+. This may still be required long after the pigs have been moved on to another area

Reduce stocking density Use nose rings to reduce soil disturbance Site selection is crutial – look for sites which are flat and not directly

connected to rivers

3.4. Land Spreading of Manures, Dirty Water and Other Organic Residues

27

To reduce the accumulation of NH4+ in the soil which is then at risk of transfer to

watercourses, it is essential that the N in manures (and fertilisers) is matched with crop demand. To achieve this, the following rules apply,

Do not apply manures to fallow land Know the plant requirements (Defra, 2000) and use decision support systems

e.g. Planet/MANNER Know (analyse) the total N and NH4

+-N content (and uric acid content for poultry) of manures

Take into account the N content (and availability) of manure when calculating crop inorganic fertiliser N rates

Take into account soil N supply (analyse soil)

To prevent rapid transfer of NH4+ to drains,

Do not apply dirty water to drained soils Do not apply NH4

+ based fertilisers, slurries and dirty water to cracking clays Do not apply NH4

+ based fertilisers, slurries and dirty water to waterlogged or cracked soils (MAFF, 1998a)

To reduce surface transfers to watercourses, Incorporate slurries and manures into arable land Shallow inject across the slope Do not to apply slurries, dirty water and solid manures when heavy rain is

forecast within the next 2-3 days Do not apply manures, slurries and dirty water to deeply frozen or steeply

sloping land (MAFF, 1998a) Use buffer strips or constructed/natural wetlands between agricultural land and

watercourse

3.5. Silage Clamps and Big Bales

Collect all effluent from clamp silage or big bale silage stored on concrete and direct to dirty water or slurry store

Maintain the structural integrity of the store (seals, panels) Wilt silage to >25% moisture content Site big bale stores away from drains, ditches and potable water supplies

(wells and boreholes)

3.6. Septic Tanks

Maintain structural integrity of septic tanks. Inspect regularly.

3.7. ‘Upstream’ StrategiesThe mitigation practices described above are mainly ‘end of the pipe’ management practices, particularly for animal manure and livestock management. There are, however, additional feeding practices that can be used to reduce the NH4

+ voided by the animals that then requires handling. These include:

Use of higher digestability diets Use of feed additives and replacement of crude protein with specific amino

acids, particularly in pigs and poultry (Pfeiffer, 1993; Schutte, 1994;

28

Misselbrook et al., 1998) through reduction of the crude protein content of pig feed and substitution with synthetic amino acids the pig slurry resulting from the modified diet contained 40% less NH4

+-N and 20% less total N than the ‘control’ slurry. There was no impact on pig performance. Smits et al. (1995) showed that by reducing the crude protein content of cattle diets (grass + concentrates) the slurry NH4

+-N content was reduced by 20% and the total N content by 10% compared to the control.

Phase feeding can be used to match nutrient supply to livestock needs, through feeds, more closely. The metabolic requirements of livestock will vary depending on the phase of growth and lactation. This has been used successfully in pig and poultry systems to reduce N in excreta.

4. POLLUTION SWAPPING

Some of the methods to reduce NH4+ transfers to watercourses may enhance the

potential loss of another form of the same pollutant or indeed another pollutant altogether. Some examples are shown below:

Biological treatment to remove NH4+ results in increased NO3

- concentrations (Project WA0518). Also, slurry aeration can result in significant N2O losses (Burton et al., 1993). Ammonia emissions may also be increased via aeration of slurry and turning of solid manure heaps.

Removal of livestock from pasture during wet conditions to reduce poaching and losses of NH4

+ and other contaminants by erosion and surface lateral transfers may result in increased NH4

+ losses direct form the stand off pad if the pad is not drained effectively and leachates directed to storage facilities.

It is also important to note that any management practice designed to conserve NH4+-

N (and reduce NH3 volatilisation) will increase the source strength of NH4+-N once

applied to the land.

29

3. NITRITE

1. INTRODUCTION

While Ammonium (NH4+) and Nitrate (NO3

-) in agricultural soils and in river systems have been widely studied and well documented, Nitrite (NO2

-) has seldom been measured separately and often included in NO3

- values. Generally, NO2- has been

viewed as a transient intermediate in a number of microbial and chemical processes rarely accumulating in terrestrial and aquatic ecosystems. The metabolism of NO2

- in soils has been shown to be extremely rapid and only accumulates when some factor causes a lower rate of consumption than production. In aquatic systems it has been estimated that the world-wide average concentration of NO2

- in unpolluted water is only 1.0 µg N l-1 (Maybeck, 1982), and where pollution is absent, NO2

- rarely accumulates. The two dominant processes involved with NO2

- turnover in the environment are nitrification following mineralization and NO3

- reduction during denitrification (Burns et al., 1996). Nitrification requires an aerobic environment while anaerobic conditions are required for denitrification. These aerobic and anaerobic environments on a micro-scale can exist in close proximity to each other and the mobility of NO3

- and NO2- makes it possible for these ions to move from one

zone to another and specific N transformations occur in each.

With the intensification of farming the inputs of NH4+ and NO3

- to agricultural systems have increased dramatically and inevitably so have losses of both to aquatic systems. Accordingly, as both NH4

+ and NO3- are precursors to NO2

- formation, NO2-

levels have also increased.

Burns et al. (1995b) and Jones and Schwab (1993) have confirmed that NO2- can

occur in soil at relatively high concentrations while Chapman and Liebig (1952) have demonstrated that the NO2

- produced may persist for several months. High concentrations of NO2

- are toxic to germinating seeds and soil micro-organisms. Plants grown in solutions containing NO2

- are small, chlorotic, prone to wilting and

30

55

6

5

3

2

Organic N NH4+ NO2

- NO3-

NH4+ NO2

- NO3-Organic N

N2, N2O

N2, N2O, NO

Aerobic

Anaerobic

2 2

1 1

3, 4 4

1 – Nitrification2 – Assimilatory nitrate reduction3 – Denitrification4 – Dissimilatory nitrate reduction5 – Diffusion and transport6 – Chemodenitrification (Adapted from

Smith et al., 1995)

Figure 1. Nitrogen Transformations

have sparse, lignified root systems. Nitrite toxicity increases with decreasing pH (Bingham et al., 1954) and with deficiencies of iron and of magnesium and with poor root aeration (Phipps and Cornforth, 1970).

Increased levels of soil NO2- may contribute to concentrations of NO2

- in river waters resulting in levels above European Union (EU) guidelines as outlined in the Freshwater Fish Directive (European Economic Community, 1978). The guide values for NO2

- in rivers supporting salmonid and coarse fish are 3.0 and 9.0 µg NO2--N l-1,

respectively. Watson et al. (2000) reported that drainage from grassland plots receiving 100 kg N ha-1 yr-1 and above exceeded these recommended levels for both NO2

- and NH4+. Recent reports from Germany (Werner, 1991) and Northern Ireland

(Smith et al., 1995) have also reported concentrations in the range of 100 to 200 µg NO2

--N l-1 within certain river systems. At high concentrations, NO2- poses a threat to

fish due to the sensitivity of fish haemoglobin to NO2-, which impairs its ability to

take up oxygen. Some species of invertebrates with the blood pigment hemocyanin are also sensitive to NO2

- in aquatic systems. Toxicity is primarily related to the pH of water, and each time pH increases by one unit the concentration of the toxic factor increases ten times (Petit, 1990). The presence of monovalent ions such as chlorides and bicarbonates strongly reduce the toxicity on NO2

- (Tomasso et al., 1979). Therefore the toxicity of NO2

- is raised in fresh water and reduced in brackish and sea water.


2.1. SOURCES

2.1.1. Direct Sources

No direct sources are reported for NO2- applied directly to agricultural land in the

form of artificial fertilizers. Agricultural wastes also appear to have little or no NO2-

at the point of application. Agricultural slurry contains 40-60% NH4+/uric acid (Smith

and Chambers, 1993; MAFF 2000) and is too anoxic for NO2- formation though

nitrification (Chadwick and Chen, 2002), while farmyard manure is generally too biologically active for NO2

- to persist. However, in France, aerobic treatment of slurry is seen as necessary to remove nitrogen (NH4

+ and NO3-) through

nitrification/denitrification as di-nitrogen gas (N2) in order to protect watercourses. Such treatment can result in NO2

- accumulation, depending on aeration level and raw slurry composition (Béline et al., 1999).

2.1.2. Indirect Sources

As Figure 1 shows, there are two precursors to NO2- formation, namely NH4

+ and NO3

-. With either or both of these ions present in soil, NO2- formation can occur.

Table 5 contains the typical content of NH4+ and NO3

- in many agricultural inorganic-N fertilizers and the fraction that each contributed of the total N fertilizer used within the UK for 2001/02.

Chapman and Liebig (1952) have reported that more NO2- was produced from urea

than any other studied fertilizer including ammonium nitrate and calcium nitrate.

31

Bezdicek et al. (1971) also found that urea produced the most NO2- and free NH3

when compared with diammonium phosphate and ammonium sulphate.

Other potential sources of NH4+ and NO3

- are slurry, urine patches, farmyard manure and dung patches. Considerable accumulations of NO2

- were found to occur in soils which received urine of a high-N concentration (Monaghan and Barraclough, 1992) and with the high NH4

+ content of most slurries it would be expected that similar flushes of NO2

- could occur after application in favourable conditions. The application of manure to soil though, has been reported the produce little NO2

-

(Chapman and Liebig, 1952).

A further potential source of NH4+ is the large quantity of soluble organic nitrogen

(SON) present in agricultural soils. The size of the SON pool is variable but can sometimes be equal to the mineral N content of the soil, however only about one-tenth of it is leached in the form of dissolved organic nitrogen (DON) (Murphy et al., 2000). Smith (1987) has shown that DON outside of soil, i.e. after leaching into streams and rivers, is not “exceptionally susceptible” to mineralization. However, it is thought that SON, within soil, it plays an important role in mineralization/immobilization. However, information on the role of SON in N transformations is sparse and often contradictory (Murphy et al., 2000).

2.1.3. Process of NO2- Formation in Soil

2.1.3.1. Autotrophic NitrificationNitrification may be defined as the oxidation of reduced nitrogen compounds. There are two groups of autotrophic bacteria which obtain energy for growth from the oxidation of either NH4

+ to NO2- or NO2

- to NO3-. Five genera of NH4

+-oxidisers known to exist in soils: Nitrosomonas, Nitrosolobus, Nitrosovibrio, Nitrosospira and Nitrosococcus and one NO2

--oxidiser, Nitrobacter (Belser, 1979). Various authors have found different genera of NH4

+-oxidisers to predominate in soil. Belser and Schmidt (1978) found Nitrosomonas and Nitrosospira to be more common than Nitrosolobus, whereas evidence is accumulating that Nitrosolobus may predominate in some soils (Walker, 1978; MacDonald, 1979). However, all oxidise NH3 through the sequence shown in Figure 2.

2.1.3.2. Heterotrophic NitrificationIt had been previously believed that nitrification was only undertaken by a few specialised autotrophs, however, it is now becoming apparent that heterotrophic bacteria and even fungi also contribute. Indeed, in some soils it may be the primary source of nitrification (Killham, 1986). Unlike autotrophic nitrification, a much more heterogeneous group of prokaryotes and eukaryotes are involved in heterotrophic nitrification.

32

NH4+ NH2OH NOH NO2

- NO3-

NitrosomonasNitrosococcusNitrosospiraNitrosolobusNitrosovibrio

Nitrobacter

Figure 2. Nitrification

Nitrification rates obtained from pure isolated cultures were 103 to 104 times lower for heterotrophs than for autotrophs (Focht and Verstraete, 1977) and their contribution therefore may only be of importance where autotrophs do not thrive. However, evidence presented by Kuenen and Robertson (1988) suggests that heterotrophic nitrifiers denitrify simultaneously and therefore accumulate little or no NO2

-. Therefore their nitrification rates maybe higher than they would appear if based on NO2

- accumulation, although still lower than many of the nitrification rates published for the autotrophs.

There are still doubts as to the role of heterotrophic nitrification in natural ecosystems. The fact that nitrification occurs in situations where autotrophic organisms would not be expected to be active, or where they can not be isolated does not alone prove that nitrification is heterotrophic.

2.1.3.3. NO3- Reduction

Three main routes of NO3- reduction are available to bacteria (Robertson and Kuenen,

1984) and these are shown in Fig 1. Nitrate assimilation and respiration (dissimilatory pathway) may proceed simultaneously within the bacterial cell, although the two processes are enzymatically separated at the NO3

- reduction level. Nitrate assimilation is a tightly regulated process which normally proceeds slowly at the rate that NH3 is required for growth, so NO2

- rarely accumulates. However, dissimilatory NO3

- reduction is a more rapid process which can lead to correspondingly rapid accumulation of NO2

-.

i. Assimilatory PathwayDuring aerobic growth, many plants, eukaryotic and prokaryotic micro-organisms reduce NO3

- to NH4+ which is then assimilated into the organic pool as glutamate or

glutamine. Assimilatory reduction is not energy yielding and in principle, any NO3--N

which is processed via this pathway remains in the biomass.

ii. Dissimilatory Pathways Dissimilatory NO3

- reductionBy contrast, dissimilatory NO3

- reduction is a more rapid process and involves the conversion of NO3

- to NH3. This can therefore be considered N-conserving and can result in a correspondingly rapid and massive accumulation of NO2

-. Many obligative anaerobes or aerotolerant fermentative bacteria exploit the high mid-point redox potential of the NO3

-/NO2- couple by using NO3

- as an acceptor for electron transfer reactions. Nitrobacter has also been shown to grow under anaerobic conditions via this pathway but that under these conditions accumulating NO2

- concentrations inhibit its growth (Freitag et al. 1987).

DenitrificationDenitrification maybe defined as the reduction of NO3

- via NO2- to N2 or nitrous oxide

(N2O):

This is a two stage process with the initial step of NO3- to NO2

- being solely biotic, however, the conversion of NO2

- to nitrogen gasses can occur both biotically and abiotically. Most denitrifiers are facultative anaerobes, and it is now generally

33

accepted that when oxygen and NO3- are present and available to the cell, oxygen will

be reduced instead of NO3-. As a rule NO3

- reduction begins when oxygen levels are depleted, however, some organisms have been found to be capable of simultaneously utilizing NO3

- and oxygen (O2) as terminal electron receptors in respiration (i.e. Thiosphaera pantotropha). This phenomenon has been termed aerobic denitrification (Robertson and Kuenen, 1984).

The conversion of NO2- to N gasses can occur abiotically through a number of

processes; these are known collectively as chemodenitrification and are all dependent on pH and concentration. They have been extensively described by Chalk and Smith (1983), however, in summary:

Decomposition of nitrous acid (HNO2):The chemical stability of NO2

- in soils solutions is linked to the equilibrium of HNO2

and NO2-. Under acidic conditions HNO2 self-decomposes via the following reaction

(Allison, 1963):

However, Nelson and Bremner (1970) believe that it is better represented by:

where nitrogen dioxide (NO2) dissolves and dismutates in the soil (Smith and Chalk, 1979). Thermodynamically, above pH 5.47 spontaneous decomposition of HNO2 into NO2 does not occur (Van Cleemput and Baert, 1978).

α-amino acids and HNO2:Under suitable conditions, HNO2 will react with compounds containing free amino groups (i.e. amino acids, urea and amines) to produce N2:

This is commonly referred to as the Van Slyke reaction, although the true Van Slyke reaction only involves α-amino acids (Allison, 1963).

Organic matter and HNO2:Organic compounds other than amino groups may react with HNO2

resulting in gaseous nitrogen evolution (Bremner, 1957). Clarke and Beard (1960) and Nelson and Bremner (1970b) showed that acid soils with low organic matter retain NO2

-

better than those with high organic matter contents. Smith and Clarke (1960) described greater N2 losses from a soil with high organic matter compared to a soil with low organic matter.

Metallic cations and HNO2:Under acid conditions Nelson and Bremner (1970b) have shown that only Cu+, Fe2+

and Sn2+ ions promoted NO2- decomposition, although they point out that many

aerobic soils do not normally contain sufficient amount of these cations to promote NO2

- decomposition. However, they point out that significant amounts of ferrous iron have been detected in waterlogged soils where NO3

- reduction may also be occurring.

Ammonium and HNO2:

34

Formation of the unstable compound ammonium nitrite (NH4NO2) occurs from the reaction between NH4

+/NH3 and HNO2- under acid conditions, and follows the

following reaction described by Allison (1963):

However, Smith and Clark (1960) suggest that loss of N2 from an acid aerobic soil treated with NH4

+ and NO2- was not due to NH4

+/NO2- decomposition but from

reduction of NO2- by some other agent.

2.1.4. Relative Contributions of Nitrification and Denitrification to Soil NO2-

The processes of nitrification and denitrification are important in converting excess N in agricultural systems to diffuse N pollution. This point is well made when it is considered that only about 25% of the 1.25 x 106 t N fertiliser used in the UK during 1997 was actually applied in the form on NO3

- and that reduced forms of N dominate organic wastes (Edwards et al., 2000). The formation of NO2

- in soil is the result of a combination of these processes and its accumulation occurs as a result of differences in the rates of its production and consumption.

The occurrence of NO2- within soil is highly variable spatially and with depth. Jones

and Schwab (1993) found no particular pattern in the appearance of NO2- in soil

solution at two depths, and Burns et al. (1995b) found that although NH4+ and NO3

-

decreased with depth NO2- did not correlate with depth. This is not altogether

unexpected as grazed grasslands receive an uneven return of excretal N as well as, in some cases, applications of artificial fertilizer enhancing the natural heterogeneity of NO2

- precursors and other soil properties. Many authors have noted flushes of NO2-

under certain conditions. Chapman and Liebig (1952), Wetselaar et al. (1972), Chalk et al. (1975), Burns et al. (1995b) and others have all shown that NO2

- can accumulate in soil as a result of nitrification of NH4

+. Lloyd (1993) and Burns et al. (1995b) have also shown that there is some evidence for NO3

- reduction contributing to soil NO2-.

Burns et al. (1995a) found that >50% of NO2- consumed in soil incubations was

recovered in the NO3- pool indicating that nitrification was the main NO2

- consuming process. However, it was also indicated that recycling of NO3

--N back to the NO2-

pool occurred later, after the initial consumption of NO2-. The relative contributions

of nitrification and denitrification to soil NO2- have been studied by Burns et al.

(1996). In soils with moisture contents ranging between 40-60% both the nitrification and denitrification of contributed to soil NO2

- simultaneously. During soil incubations it was found that initially reduction of NO3

- was the main source of NO2- although

total soil NO2- remained low. Over time, as soil NO2

- increased, it was found to be nitrification of NH4

+ that became dominant. After this pulse of nitrification had passed denitrification again became the main source of NO2

-. Overall it was shown that nitrification was responsible for most of the NO2

- produced. However, denitrification occurred simultaneously and produced smaller, though still significant, amounts of NO2

-. Smith et al. (1997) employed modelling procedures to data obtained from soil incubation and also concluded that nitrification was the main source from NO2

- in soil and that the rate of NH4+ oxidation only needed to be slightly

greater than that for NO2- oxidation for NO2

- to accumulate. Similar conclusions were also drawn from NO2

- accumulations in river sediment (Smith et al., 1997b).

2.1.5. Causes of NO2- Accumulation

35

The processes that cause the formation, accumulation and decomposition are summarised in Figure 3. The build up of NO2

- occurs when the rate of production exceeds the rate of consumption. This is caused by many factors during the nitrification process.

i. TemperatureAs expected with biological pathways in general, temperature has a strong effect on microbes. A rise in temperature results in an increase in microbial activity up to an optimum point beyond which any further increases cause a decline. Nitrification is a two stage process involving several groups of bacteria, and the effects of temperature have different effects on each the stage.

Belser (1979) reported that Nitrosomonas activity increased more than Nitrobacter activity with increased temperature. Quinlan (1980) also demonstrated that at super-optimal temperatures NO2

- consumption is inhibited more than NO2- production per

unit rise in temperature, leading to NO2- accumulation. Quinlan (1980) also showed

that NO2- accumulation would also occur at sub-optimal temperatures when NO2

-

production is stimulated more than NO2- consumption. Chapman and Liebig (1952)

and Tyler and Broadbent (1960) also reported that NO2- oxidisers were very sensitive

to low soil temperatures and that NO2- accumulated in soils varying in temperature

between 50-60°F and persisted for several months.

ii. pHUnder conditions of low pH, NO2

- does not accumulate (Chapman and Liebig, 1952) and it has been suggested this is due to chemical rather than biological pathways (Tyler and Broadbent, 1960; Reuss and Smith, 1965; Bulla et al., 1970; Nelson and Bremner, 1969; 1970; Bollag et al., 1973). Other authors (Chapman and Liebig, 1952; Bezdicek et al., 1971; Chalk et al., 1975; Burns et al., 1995b; Van Cleemput and Samater, 1996) have reported that a large flush of NO2

- can occur in soil with raised pH. This has been attributed to the inhibition of Nitrobacter under these conditions. However, higher pH may also be causing NO2

- accumulation in conjunction with increased free ammonia (NH3).

iii. Elevated NH3 LevelsMonaghan and Barraclough (1992) reported that high urine-N concentrations resulted in NO2

- accumulation and reduced nitrification activity due to the presence of free NH3. In soils with a high pH and NH4

+ concentration, NO2- oxidisers may be

selectively inhibited (Prosser and Cox, 1982). Free NH3 dissociates from NH4+ at

elevated pH and inhibits NO2- oxidoreductase in Nitrobacter (Yang and Alleman,

1992).

The percentage of undissociated NH3 can be calculated from the formula given by Smith et al. (1997):

where T = temperature (°C).

36

Smith et al. (1997) presented evidence that Nitrobacter is inhibited by free NH3 to a greater extent than Nitrosomonas. Alleman (1984) reported that, based on results published by Anthonisen et al. (1976) and Prakasam and Loehr (1972), Nitrobacter retardation will initially develop at NH3 concentrations as low as 0.05 mg l-1. This is approximately two orders of magnitude below the level for an equivalent effect on Nitrosomonas.

iv. Reduced Oxygen LevelsLaanbroek and Gerards (1993) have shown that under conditions of limited O2 supply, some NO2

- oxidisers are repressed while the NH4+ oxidisers are unaffected, causing

NO2- accumulation. They also have shown, along with Freitag et al. (1987) and Bock

et al. (1988), that under limited O2 conditions, partial denitrification by Nitrobacter also occurs, recycling NO3

- to the NO2- pool as well as producing some NH3 and N2O.

v. Reduced Carbon Dioxide Levels It is generally assumed that autotrophs dominate the majority of nitrifying systems. Jones and Paskins (1982) observed that low CO2 tensions lead to elevated NO2

-

formation. It was suggested that a metabolic shortcoming in Nitrobacter’s competition for anabolic carbon was the probable cause for this NO2

- build-up.

vi. Light IntensityThe susceptibility of nitrifying bacteria to inhibition by sunlight has been described Olsen (1981). It was reported that 50% inhibition of NH4

+-oxidisers and NO2--

oxidisers occurred at light intensities three orders of magnitude less than the intensity of full sunlight and that Nitrobacter was more sensitive to sunlight than Nitrosomonas.

vii. Uneven Distribution of Bacterial PopulationsMonreal et al. (1986) found that NH3-oxidisers outnumbered NO2

--oxidisers and NO2-

flushes occurred, in soil incubated with nested urea-N. Where urea had been mixed thoroughly with the soil NO2

--oxidisers out numbered NH3-oxidisers and no NO2-

flush occurred. So NO2- flushes could be due to an imbalance in the numbers of NO2

-

and NH3-oxidising bacteria which, in turn, could be due changes in soil conditions (i.e. pH, N and C supply, moisture) vertically and horizontally leading to differing rates of NO2

- production and consumption.

2.2. MOBILISATION

As with NO3-, NO2

- is highly soluble and only weakly bonded by soil particles. As such, the majority of NO2

- is lost from soils through leaching because of the low capacity of most soils to retain these anions. It is therefore very mobile and in general, any downward movement of water through the soil profile will cause leaching of NO2

- if present.

Figure 3. Summary of the process of NO2- formation, accumulation and

decomposition.

37

2.3. DELIVERY

Due to the necessity for NO2- to form and accumulate within the soil profile little is

lost in rapid ‘overland’ movements of water. If conditions are suitable for NO2-

accumulation then the majority is lost in solution via the slow through flow of water through the soil column and drains. Some NO2

- is removed during high through soil flows (macro-pores and drainage etc) however, the majority of this drainage initially bypasses the soil NO2

- pool.


Smith et al. (1995) estimated that only 40% of the NO2- load within the Lough Neagh

river system in Northern Ireland had come from land drainage directly. This NO2-

was being formed within the soil profile and subsequently leached downwards (Burns et al., 1995b). Formation of NO2

- occurred though both nitrification and denitrification (Burns et al., 1995a; 1996) with, depending on soil conditions, nitrification of NH4

+ being the dominant process. The remaining 60% of the NO2-

river load was believed to have originated from nitrification of NH4+ at the sediment-

water interface, although denitrification of NO3- may also be contributing, even under

aerobic conditions. Smith et al. (1995) suggested that NO2- build up was due to

inhibition of Nitrobacter by free NH3 which was predicted to be in the range of 65 to

38

Soil Organi

c Matter

NO2-

Chemodenitrification:Low pH, high organic matter, α-amino acids, Fe2+ Sn2+ Cu+

Biodenitrification:Neisseria & Flavobacterium

High pHHigh free NH3

Low/High TemperatureLow oxygen

Low CO2

NH4+ NO3

-

Build-up and leaching

NO2-

N2, NO, N2O

NH4+ NO3

-

Leaching

Soil Detachm

ent

Artificial FertilisersSlurry, FYM, Dung and Urine Patches

Agricultural Inputs

Agricultural Inputs

-O2

+O2+O2

+O2

-O2

+O2

SoilR

iver

76 µg N l-1. This would indicate that control of NO2- is best dealt with through the

mitigation of its precursors, particularly NH4+, within both soil and watercourses.

Within the Lough Neagh system it is estimated that 50% of the NO3- river loadings

was derived from the leaching of agricultural fertilizers through land drainage (Smith and Stewart, 1989) while NH4

+ originated from agricultural effluents as a result of farm pollution (Foy and Kirk, 1995). However, as N-transformation within the river system was occurring at the sediment-water interface perhaps soil erosion is a further source of NH4

+ leading to NO2-.


Nitrite is the intermediate of many biotic and abiotic pathways as shown in Figure 1. Nitrite is generally highly reactive and its accumulation in soils occurs as being a result of differences in the rates of its production and consumption.

The main process involved in NO2- formation is the nitrification of soil NH4

+ to NO3-.

Therefore if NO2- is accumulating via this pathway, it maybe removed by

manipulating soil conditions to allow efficient oxidation to NO3-. This would,

however, cause an increase in the soil NO3- pool and make greater amounts of NO3

-

available for leaching. This is possibly a moot point however, as NO2- is also highly

soluble and therefore soil N is just as likely to be lost via NO2- leaching. The other

important pathway for NO2- formation is NO3

- is reduced to NO2- (Figure 3). If the

pathway were to be completed, removing soil NO2-, then many gaseous forms of N

are formed including some greenhouse gases. However, there is evidence that a great deal of N2, the ideal end product for the removal of excess N, could be formed through this process.

39

4. PATHOGENS

1. INTRODUCTION

In England and Wales, around 67 million tonnes of animal manure are estimated to be collected annually from farm buildings and yards, requiring handling, storage and subsequent land application. Of these, approximately 45% are solid based manures and the remainder liquid slurries (cattle and pig), Table 6. Additionally, around 45 million tonnes of excreta are deposited directly in the field by grazing cattle, sheep and pigs (Chambers et al., 2000).

Table 6. Estimated quantities of animal manure handled annually in England and Wales (Chambers et al., 2000).

Manure type Total fresh manure(million tonnes)

As solid manure(million tonnes)

As liquid manure(million tonnes)

CattlePig

PoultrySheepTotal

53.08.93.51.967.3

21.04.33.51.930.7

32.04.6--

36.6

Manures are applied annually to around 16% of tillage land (0.6 million hectares) and 48% of grassland (2.3 million hectares) in England and Wales (Chambers et al., 2000). Applications are made throughout the year depending on crop type, soil conditions and manure storage capacity (Smith et al., 2000a; Smith et al., 2001a). The well managed application of animal manures to land enables their nutrient and organic matter contents to be used to help meet crop nutrient requirements and to improve soil quality. However, where poorly managed, manures can also be a major source of diffuse pollution of water by pathogens and nutrients. Pathogen losses from agriculture, particularly to surface water systems, can impact on compliance with the Bathing Water Directive (EC, 1976) and Shellfish Waters Directive (EC, 1979), and food safety, where surface water is used to irrigate ready-to-eat crops such as lettuce, carrots and cabbage.


2.1. SOURCES

The rumen and digestive tract in farm livestock is host to a rich diversity of microflora and can act as a reservoir for pathogenic micro-organisms (Rasmussen et al., 1993). Manures can contain micro-organisms such as E. coli. O157, Salmonella, Listeria, Campylobacter, Cryptosporidium and Giardia. Pathogen presence in manures is affected by factors such as animal type, age, diet and management, as well as regional and seasonal influences (Nicholson et al., 2000). Pathogenic micro-organisms may be transmitted from animals to humans either directly or indirectly through water and food chain contamination.

40

2.1.1. Microbial pathogens

i. Escherichia coliMost strains of E. coli are harmless; however, several strains are potential pathogens (e.g. O157, O111, O26, O103 and O145). Verotoxin producing E. coli O157 is considered to be the most severe strain. Dairy and beef cattle are generally thought to be the most important reservoirs of E. coli O157 (Chapman et al., 1997). Faecal swab samples taken from an abattoir in North Yorkshire showed that 15.7% of 4,800 cattle samples contained E. coli O157, 2.2% of 1000 sheep and 0.4% of 1000 pig samples, with no occurrence in poultry faeces. A major prevalence study in Scottish beef cattle (aged between 12 and 30 months) collected 14,849 samples from 952 groups of finishing cattle. Animal level prevalence was 7.9% and group level prevalence 22.8%; 95% of samples were positive for verocytotoxigenic (VTEC) E. coli O157 (Synge et al., 2001). Seasonal patterns in the shedding of E. coli O157 have been reported, with excretion rates highest in spring and autumn. Additionally, increased shedding rates have been associated with the movement of animals to housing, increases in herd numbers, the introduction of new animals to the herd, infection at an early age, the onset of lactation and the spreading of slurry on grazing land. Prevalence is further complicated by the observation that VTEC infections of livestock appear subject to regional variation (Mechie et al., 1997; Synge, 2000; Synge et al., 2001).

Recent survey work funded by the Food Standards Agency – FSA (Hutchison et al., 2002) has provided up to date information on E. coli O157 prevalence (Table 7) and levels (Table 8) in fresh animal excreta and stored manure samples collected throughout England, Wales and Scotland. E. coli O157 was present in the fresh excreta of 15% of cattle, 16% of pigs and 22% of sheep samples, although the sample numbers for pig and sheep were limited (93 and 23, respectively). This recent data suggests greater E. coli O157 prevalence in pigs and sheep than the data previously reported by Chapman et al. (1997) of 0.4% for pigs and 2.2% for sheep. E. coli O157 mean levels for each livestock manure type were 467 colony forming units – CFU g-1

of fresh excreta for cattle, 782 g-1 of excreta for sheep and 3908 g-1 of excreta for pigs.

Table 7. Percentage of British livestock manures that contained zoonotic micro-organisms.

PathogenCattle Pig Poultry Sheep

Fresh Stored Fresh Stored Fresh Stored Fresh Stored

E. coli O157SalmonellaListeriaCampylobacterC. parvumG. intestinalis

(n=522) (n=249)

15.4% 13.2%4.5% 4.8%28.8% 34.5%15.9% 15.6%

7.1% 4%3.8% 1.6%

(n=93) (n=35)

16.1% 26%8.6% 2.8%20.4% 22.8%18.3% 17.1%18.3% 5.7%3.2% 2.9%

(n=48) (n=17)

ND ND12.5% 5.8%18.75% 11.8%22.9% 1.8%ND NDND ND

(n=23) (n=6)

21.7% 16.7%8.6% 0%

30.4% 33.3%21.7% 16.7%

30.4% 0%21.7% 0%

ND = Not determined because it was not appropriate to test manures from these species for all pathogen typesn = Number of samples.

41

Table 8. Mean levels of zoonotic pathogens observed in British livestock manures (data are geometric means for positive isolations only).

Levels of pathogens found in positive samples for each animal species (CFU g-1)

Pathogen Cattle Pig Poultry SheepFresh Stored Fresh Stored Fresh Stored Fresh Stored

E. coli O157 467 306(n=85) (n=33)

3908 1296 (n=15) (n=9)

ND ND 782 5000(n=5) (n=1)

Salmonella 670 296(n=25) (n=12)

738 2000(n=8) (n=1)

12 1900(n=6) (n=1)

707 -(n=2) -

Listeria 213 554(n=159) (n=86)

113 414(n=19) (n=9)

420 110(n=9) (n=2)

198 159(n=7) (n=2)

Campylobacter 529 397(n=88) (n=38)

624 351(n=17) (n=6)

447 589(n=11) (n=2)

386 100(n=5) (n=1)

C. parvum 19 8.4(n=39) (n=10)

58 33(n=17) (n=3)

ND ND 20 - (n=5) -

G. intestinalis 20 1(n=12) (n=4)

68 12(n=3) (n=1)

ND ND 10 -(n=7) -

ND = Not determined because it was not appropriate to test manures from these species for all pathogen typesn = Number of samples.

ii. SalmonellaSalmonella species are capable of prolonged survival in harsh environments outside their host (Winfield and Grossman, 2003). Although historically the main reservoir has been considered to be poultry, cattle manures can also harbour significant numbers of these pathogenic bacteria. Recently most laying hens have started to be vaccinated against Salmonella enteritidis before going into lay, in response to pressures from the major retailers rather than any legal requirement (Nicholson et al., 2000).

In 1998, a total of 1375 isolations of Salmonella from cattle were reported in the UK (Anon., 1998c), representing a considerable reduction in the number of cattle infected compared with the early 1990s. Jones and Matthews (1975) examined 187 cattle slurries and found Salmonella present in 11% of samples, although numbers were exceptionally low, typically less than one organism g-1 of slurry. However, Jones (1976) reported data which showed that heifers exhibiting no clinical signs of Salmonellosis could excrete as many of 108 Salmonella dublin g-1 of faeces. The recent FSA survey data (Hutchison et al., 2002) indicated a prevalence of Salmonella of between 4.5% and 8.6% for fresh cattle, pig and sheep excreta, and 12.5% for poultry excreta (Table 7), with mean levels ranging between 12 and 707 CFU g -1 of fresh excreta for the four livestock types.

iii. CampylobacterCampylobacter are common to the intestinal track of humans and animals, and the strain associated with most reported human infections, Campylobacter rejuni, causes over 90% of Camplyobacter enteritis in the UK (Stanley and Jones, 2003). During 1997, 62% of all reported cases of gastrointestinal infection in the UK were attributed to Campylobacter (Jones, 1999). Campylobacter spps. are widely found in the intestinal tract of many animals, especially poultry (Jones et al., 1999; Stanley et al., 1998). Indeed, Stanley et al. (1998) reported an intestinal carriage at slaughter of

42

89.4% in beef cattle in north-west England between 1993 and 1995. The recent FSA survey data (Hutchison et al., 2002) indicated a prevalence of Campylobacter of between 15.9% and 22.9% in fresh cattle, pig, poultry and sheep excreta (Table 7), with mean levels for each livestock type ranging between 386 and 624 CFU g -1 of fresh excreta.

iv. Listeria monocytogenesListeria is most often associated with dairy product contamination, with excretion of this pathogen reported in cattle faeces (Pell, 1997). Listeria is able to survive and proliferate under hostile environmental conditions (Sleator et al., 2003). The recent FSA survey data (Hutchison et al., 2002) indicated a prevalence of Listeria of between 18.8% and 30.4% for the four fresh livestock excreta types (Table 7), with mean levels for each livestock type ranging between 113 and 420 CFU g-1 of fresh excreta.

2.1.2. Protozoa

i. CryptosporidiumCryptosporidium parvum is a zoonotic parasite that infects the gastrointestinal tract of warm-blooded animals, including humans, causing the disease Cryptosporidiosis. Cryptosporidium oocysts can remain viable for about 18 months in a cool damp environment (Nicholson et al., 2000). The oocysts are remarkably resistant to many common disinfectants, including chlorine based compounds. Sturdee (1998) in a study on the incidence of Cryptosporidium in farm mammals in the English Midlands concluded that Cryptosporidium was now ubiquitous amongst mammals in the UK, and that there was now an irreducible minimum background level in wildlife that acted as a reservoir for continual reinfection of domestic livestock.

The recent FSA survey data (Hutchison et al., 2002) indicated a prevalence of Cryptosporidium of 7.1% in fresh cattle excreta, 18.3% in pig excreta and 30.4% in sheep excreta (Table 7), with mean levels for each livestock type of 19, 58 and 20 CFU g-1 of fresh excreta.

ii. GiardiaGiardia is a single celled parasite that can cause the disease Giardiasis through the consumption of faecally contaminated materials. The recent FSA survey data (Hutchison et al., 2002) indicated a prevalence of Giardia of 3.8% in fresh cattle excreta, 3.2% in pig excreta and 21.7% in sheep excreta (Table 7), with mean levels for each livestock type of 21, 68 and 10 CFU g-1 of fresh excreta.

2.1.3. Viruses

Viral pathogen sources from livestock manures represent a much lower risk to human health than bacterial and protozoan pathogens (Nicholson et al., 2000), because few zoonotis viruses infect cattle (Pell, 1997).

43

2.1.4. Summary

Nicholson et al. (2000) provided a useful summary of the presence of pathogens in UK animal manures drawing upon data collated in an extensive review of the literature (Table 9).

Table 9. Pathogens found in UK animal manures.

Pathogen Cattle Pigs Sheep PoultryE. coli O157 √ √ √ XSalmonella √ √ √ √

Listeria √ √ ND √Campylobacter √ √ √ √

Cryptosporidium √ √ √ XGiardia √ √ ND X

√ Found in UK animal manuresX Not found in UK animal manuresND = No data

2.2. MOBILISATION

A review of the processes involved in the mobilisation of pathogens has been described by McHugh et al. (2004). Rainfall is important in the mobilisation and transport of pathogens (Wyer et al., 1994) and it is generally assumed that micro-organisms behave like soil particles with mobilisation occurring by raindrop impact and flow (Tyrrel and Quinton, 2003). It has been proposed that pathogens may exist in three states within a soil-manure mixture; attached to soil particles, attached to organic matter or free and unattached (Tyrrel and Quinton, 2003). The form in which the pathogens exist is important as it directly affects the quantity of pathogens mobilised and the distance over which they are transported. This is due to the velocity of water required to mobilise a particle with mass and to keep it in suspension (Tyrrel and Quinton, 2003). The physical properties of pathogens also affect their relative rates of mobilisation as a consequence of interactions with the soil. These include the presence of cellular appendages, cell size, hydrophobicity and electrical charge (Mawdsley et al., 1995). As soil water content increases, the rate of movement of water and pathogens increases and is a function of water content, pore size and hydraulic gradient (Worrall and Roughly, 1991).

2.3. DELIVERY

There are a variety of transfer routes through which pathogens may be transported from soils to receiving waters, but the factors that control these transfers are not well understood (Hornberger et al., 1992). Pathogenic micro-organisms may enter surface waters via overland flow, by means of sub-surface transfer through highly permeable soils or through by-pass flow in soil profile cracks or artificial field drainage systems. The main driving force for dispersion from the soil environment to surface waters is rainfall events, particularly where these generate high runoff volumes and ‘turbid’ flows.

44

The transport mechanisms of micro-organisms within soils can be divided into physical, geochemical and biological processes (Tim et al., 1988). Physical processes include advection (transport as a result of heat energy) and dispersion (transport over a large area, largely as a result of diffusion). The geochemical processes act to delay pathogen transfer through the soil matrix by filtration, adsorption and sedimentation mechanisms. In addition, biological processes such as growth and chemotactic migration (movement in response to a nutrient gradient) may influence pathogen transfer through the soil environment. Figure 4 (Oliver et al., in press) shows a conceptual model of microbial transmission through the agricultural environment. It highlights two important components: (i) the routes of transfer available from source to receptor, and (ii) the continuum of micro-organism die-off from source to receptor.

Figure 4. Conceptual model of micro-organism transmission from surface applied faecal materials to surface waters (adapted from Oliver et al in press).

2.3.1. Field Losses Following Manure Spreading

45

DIE

-OFF

RECEIVING WATERS

DIE

OFF

OFF

DELIVEREDTO

SOIL SYSTEM:(Matrix flow/Subsurface lateral

flow/Preferential flow/Tile drainage)

AND RETAINEDOR

SOURCE: (Excreta/Slurry/FYM)

MICROBE TRANSFER:(freely suspended/

soil particle associated/waste particle associated)

THROUGH

DIRECTLY VOIDED BY CATTLE

OVERLAND FLOW

DIE

-OFF

DIE

-OFF

i. Overland flow pathways.Abu-Ashour and Lee (2000) showed that rainfall was the major factor driving both the vertical and horizontal movement of bacterial micro-organisms in soil. Similarly, recent work by Vinten et al. (2002) highlighted the importance of hydrological processes at the soil surface and the promotion of microbial transfers via the mobilisation of slurry colloids following raindrop impact. Fenlon et al. (2000) and Cook and Baker (2001) also noted that periods of heavy rainfall can cause significant losses of E. coli by both surface runoff, and by-pass flow through the soil profile. Quinton et al. (2003) in a laboratory study showed that the soil incorporation of slurry following land spreading, compared with surface application, reduced faecal coliform losses in surface runoff, although pathogen survival in the soil may have been increased through protection from UV light. Culley and Phillips (1982) reported that spring runoff gave the highest microbial concentrations regardless of the timing of manure applications. In summary, overland flows can provide an efficient microbial transfer route, but the impact of such losses is likely to be in the short-term following land spreading due to die-off in the soil environment.

ii. Soil matrix flows. A large body of research has focused on the vertical transfer of pathogens in leachate and the similarities with colloid filtration theory. Vertical displacement of micro-organisms through the soil profile has been demonstrated in a variety of soil column experiments (Aislabie et al., 2001; Gagliardi and Karns, 2000; Brush et al., 1999, Woolum and Cassel, 1978). The initial moisture content of the soil was shown to be important in facilitating bacterial movement with continuous water films required to enable bacterial transfer. Gagliardi and Karns (2000) concluded that if soil pores were not clogged, E. coli O157 was able to travel below the soil surface layers for periods in excess of 2 months following the initial manure application. Field-scale work examining the effects of manure application showed that whilst leachate collected at 90 cm could contain faecal coliform levels soon after application, they declined to undetectable levels within 30-60 days (Stoddard et al., 1998). Tan et al. (1991) concluded that microbial travel times were much more rapid in coarser textured soils with larger pore spaces compared with finer textured soils, a conclusion also reached by other authors (e.g. Huysman and Verstraete, 1993). More recently, soil adsorption characteristics were shown by Schijven et al. (2002) to hinder the transfer of microbes. The major soil components affecting the sorption of bacteria are clay and organic matter (Aislabie et al., 2001), with greater E. coli sorption being reported on higher clay content soils due to the greater specific surface area (Ling et al., 2002). Such adsorption, coupled with filtration, can clearly have an important influence on pathogen transfer processes and rates.

The downward translocation of micro-organisms relies very much on pore size openings and the soil matrix system. Instances of pore clogging by bacteria will undoubtedly affect microbial transfer, which will depend primarily on the grain size of the porous medium. Johnson et al. (1995), among others, have suggested that colloid filtration theory provides a conceptual framework for modeling microbial movement. However, travel distances of micro-organisms based on filtration theory have been demonstrated to underestimate true translocation lengths within the soil (Simoni et al., 1998). Microbial transport can be modeled incorporating detachment functions that are associated with microbial residence time within the soil. However,

46

the soil ecosystem accommodates much complexity and makes the study of microbial movement a difficult task.

Colloids can act as vehicles for pathogen transport, with large voids facilitating relatively easy colloid movement. However, to date, only a few studies have been published that detail direct evidence for colloid facilitated transport of contaminants (Kretzschmar et al., 1999).

iii. Macropore (by-pass) flow. A range of literature has indicated that macropore (or by-pass) flow is an important mechanism for pathogen transport in soils (e.g. Fontes et al., 1991; Gannon et al., 1991; Mawdsley et al., 1996; Harvey, 1997). Such preferential pathways serve as routes of relatively rapid water flow and allow cells, along with other colloids and contaminants, to successfully by-pass the sieving and constraining matrix of soils. Although macropores often make up only a small volume of the soil body they serve as important routes for both the lateral and vertical transfer of cells entrained in carrying water. Macropores may be formed naturally or through soil faunal activity, plant root presence or soil shrinkage. In particular, the role of field drainage systems on poorly drained soils and earthworms in providing large pore networks have been highlighted.

Drainflow losses. Joy et al. (1998) showed that the contamination of tile drainflows after the application of liquid manure was strongly associated with the presence of flow in the tiles prior to application, with elevated concentrations recorded 5 days following slurry application. Concentrations exceeded 1000 CFU/100 ml, when 8.6mm of rain fell in 24 hours after application. However, concentrations were as low as 1 CFU/100 ml where rainfall did not create drainflow until 40 days after application. Culley and Phillips (1982) demonstrated that faecal streptococci were capable of movement through the soil profile into field drains located at a depth of 75cm, provided that water was present to facilitate this downward transfer.

Work on drained soils in Scotland showed that heavy rainfall following cattle slurry application to a clay loam soil resulted in up to 7% of the applied E. coli being transported in the drain flow (Fenlon et al., 2000). Vinten et al. (2002) showed that the transport of E. coli to drains was mainly associated with rainfall and drainflow between 3 to 7 days after cattle slurry application, with the first drainflow events after slurry application containing 105 to 106 CFU/100 ml.

Work involving the application of simulated rainfall to isolated silty clay loam soil blocks treated with slurry containing Cryptosporidium oocysts showed that numbers collected in the runoff were initially high and stayed at a plateau of 103 ml-1 for up to 70 days after application (Mawdsley et al., 1996).

Soil mesofauna. Opperman et al. (1987) showed that the movement of cattle slurry and coliforms through soil was enhanced by the presence of earthworm created physical tunnels for transport. Moreover, earthworms along with mites and millipedes, were also suggested by Bowen and Rovira (1999) to act as vectors for microbial transport through the attachment of micro-organisms to such mesofauna. Soils, and in

47

particularly grassland soils, can contain considerable numbers of earthworms whose channels can facilitate pathogen transport.

2.3.2. Field Losses Following Livestock Grazing

In addition to pathogen losses following manure spreading, runoff during the livestock grazing season has also been demonstrated to transfer pathogens to surface waters (Fernandez-Alvarez et al., 1991; Jawson et al., 1982; Howell et al., 1995), with increased microbial counts in runoff water dependent on stocking density (Gary et al., 1983). The effects of grazing on the microbial quality of runoff may persist for some time after animals have been removed from the grazing pasture (Jawson et al., 1982). E. coli O157 was shown to persist on pasture for 105 days following excretion in sheep faeces (Ogden et al., 2002). Stephenson and Street (1978) showed that faecal coliform counts in stream runoff reached concentrations of up to 2500/100 ml shortly after cattle were introduced and remained at high concentrations for up to 3 months after cattle were removed.

Figure 5 (Oliver et al., in press) illustrates naturally occurring flow pathways associated with the soil ecosystem and shows transfer modes operating within them.

2.3.3. Losses During Slurry Storage

A study on one unlined slurry store located on the Upper Chalk in Hampshire measured nutrient and microbial contamination of porewaters at a depth of up to 76m, as a result of fissure flow through the unsaturated zone beneath the structure (Withers et al., 1998). In a more recent UK study, covering eight earth-based (unlined) structures, although Cryptosporidium and E. coli O157 were present in many of the cattle slurry lagoons, neither organism was found in the aquifer material beneath (Gooddy et al., 2001).

The failure or mismanagement of slurry storage structures can cause water pollution. The numbers of substantiated water pollution incidents in England and Wales (1987-1999) from agriculture and from slurry/solid manure stores is summarised in Table 10. Slurry storage accounted for 75 to 85% of all incidents from manure storage facilities.

2.3.4. Farmstead Runoff

Studies carried out in the Irvine and Girvan catchments in the West of Scotland (Aitken et al., 2001) indicated that the majority (58%) of all farms had middens, but nearly half of these had no containment facilities to prevent the discharge of effluent to watercourses. A more detailed survey undertaken on a sub-sample of 20 farms indicated that 60% had significant discharges to watercourses. The other main potential sources of microbiological contamination of watercourses were runoff from poorly contained byres, self-feed silage aprons and cow tracks.

Table 10. Water pollution incidents in England and Wales (1987-1999) from agriculture and from slurry/solid manure stores.

48

Year ’87 ’88 ’89 ’90 ’91 ’92 ’93 ’94 ’95 ’96 ’97 ’98 ‘99Total

agriculture 3,890 4,141 2,889 3,147 2,954 2,770 3,051 3,338 2,731 2,111 1,884 2,050 4,254

Slurry stores 922 1,032 758 632 681 619 583 820 466 332 278 353 NASolid manure

stores 148 194 121 118 133 155 146 193 155 68 37 63 NA

Category 1 Incidents - - - 99 67 63 36 32 28 35 22 29(Source: Environment Agency, 2000)Notes:1998 is the last year for which data were provided for slurry and solid manure stores separately.Includes 2,012 Category 4 incidents (i.e. incidents assessed to have no environmental damage),usually caused by catastrophic failure or mismanagement of slurry storage structures

N/A – not available.

Figure 5. Influential controls governing potential pathogen transfers from agricultural settings (Adapted from Oliver et al in press).

49

Recent survey data from England and Wales suggest that 80-90% of runoff from farmstead hardstandings on dairy farms and c.50% on beef farms is collected, with the remainder most likely seeping into proximate fields and ditches (Chadwick et al., 2003). Additionally, woodchip corrals are becoming increasingly common on livestock farms, and unless the leachate is collected and recycled with the farm slurry/dirty water, these facilities can also act as a source for pathogen losses.

The transport of pathogens present in animal manures can also occur through spillage from manure spreaders travelling on farm roads, public roads and through streams. Rodents and birds may also act as vectors (Cowan, 1998; Sturdee, 1998; Hooda et al., 2000; Böhm and Hartung, 1994).

2.3.5. Livestock Direct Access to Water Courses

The direct deposition of excreta into watercourses by animals has long been a cause for concern and controls have been proposed to reduce the risks associated with such practices (National Rivers Authority, 1992). The main reason for allowing animals to have access to watercourses is for drinking water supply. Additionally, dairy cattle can also contribute if they need to cross a stream or ford on the way to the milking parlour. The defecation of livestock excreta directly into a water course or in close proximity, can contaminate surface water systems with large numbers of viable micro-organisms (Tiedemann et al., 1987). Moreover, Nagels et al. (2002) who investigated microbial transfers from livestock to surface water systems concluded that direct deposition of faecal matter was likely to be the most significant loss pathway, alongside ‘wash-in’ from surface runoff. In a West of Scotland study (Aitken et al., 2001), 50% of the farms with grazing animals had access to watercourses, with 13% of dairy herds in the River Irvine catchment and 60% in the Water of Girvan catchment regularly crossing streams on a daily basis.

2.4. PATHOGEN LOSS RISK MATRIX

Information was drawn together on potential pathogen loss routes and the relative magnitude of losses of viable pathogens to water systems as part of a Defra project WA0804 (Investigation of the Routes by which Pathogens Associated with Livestock Slurries and Manure may be Transferred from the Farm to the Wider Environment). A ‘star’ rating system was used to represent the likely risks of losses from different sources under contrasting conditions. Where there was little data available from the scientific literature, estimates were made based on practical knowledge and expert judgement of the likely risks of microbiological transfer (Table 11).

The studies carried out in the West of Scotland (Aitken et al., 2001) indicated that the main pathways through which the microbial contamination of surface water systems occurred were: livestock excretion while drinking or crossing watercourses; contaminated runoff from slurry stores, middens and farmsteadings etc. running directly into watercourses; runoff or sub-surface drainage from fields where manures had been applied or where livestock were grazing. A detailed survey undertaken in the Irvine and Girvan catchments in the West of Scotland indicated that 60% of the farms had significant discharges to watercourses. Insufficient slurry storage capacity, leading to land application in inappropriate circumstances (e.g. when soils were at or

50

close to field capacity) where surface runoff or drainflow were likely to occur soon after application, was also seen to constitute a problem.

There is little data for England and Wales on the number of farms with significant discharges from uncontained farmstead runoff. However, it was possible to obtain an indication of discharges from dairy and beef hardstandings from Defra project WA0523 (Webb, 1999), which indicated that 80-90% of runoff from farmstead hardstandings on dairy farms and c.50% on beef farms was collected, with the remainder most likely seeping into proximate fields and ditches (Chadwick et al., 2003).

As livestock typically graze out in the fields for 180 days a year, this activity presents a significant risk of microbial pollution where surface runoff or drainflow occur during or shortly after the cessation of grazing. Also, grazing emissions largely occur during the late spring - early autumn when they can impact directly on bathing water quality (the bathing water season in England and Wales runs from 15 May-30 September). Similarly, the land spreading of manures (particularly fresh slurry) can present a high risk of microbial pollution where surface runoff and drainflow occur shortly after land application.

Table 11. Estimated risks of pathogen transmission to the water environment (Defra project WA0804).

* = low risk ***** = high riskOperation Risk conditions Manure type

Land

app

licat

ion

of m

anur

es

Direct application of manure into a watercourse ***** Slurry (fresh) ***** Underdrained soils at or close to a field capacity, with

rainfall within 7 days causing drainflow ****Dirty water/slurry(stored) ****

Land with moderate/steep slopes at or close to field capacity, with rainfall within 7 days causing runoff / interflow ***; and where adjacent to a watercourse ****

FYM (fresh) ***

‘Dry’ soils, where drainflow/surface runoff/interflow is unlikely within 1 month *

FYM (stored and heated)*

Gra

zing

Direct deposition of faeces in a watercourse ***** Sheep ***** Grazing on soils with underdrainage, where rainfall

causes drainflow ****Outdoor pigs *****Cattle ****

Grazing on land with moderate/steep slopes, where rainfall causes runoff/interflow ***; and where adjacent to a watercourse ****

Grazing on ‘dry’ soils, where drainflow/surface runoff/interflow is unlikely to occur during or soon after the end of grazing*

Farm

stea

d ru

noff

etc

Uncontained hard standings runoff **** Uncontained runoff from ‘wet’ (and fresh) farmstead

middens/woodchip corrals ***

Slurry *****Dirty water ****FYM (fresh) ***

Field manure heaps ** FYM (stored and heated)*

Runoff from tracks used by cattle ** Slurry stores * (to *****)

N.B. the risks represented by asterisks in adjacent columns are not necessarily related i.e. they are relative risks within each category

The overall risks of microbiological transfer were estimated for the different routes (Defra project WA0804). Based on the loads likely to be transferred to the water

51

environment and their likely frequency of occurrence, the greatest risks of transfer were assessed to be from:

Pollution from farmsteadings (hardstandings, midden and woodchip corral runoff)

Runoff and drainflow during and after livestock grazing in the field

Runoff and drainflow following the landspreading of liquid manures (e.g. slurry and dirty water)


Pathogen control measures in place or under consideration in the UK and in other countries to minimise the risks from livestock manure management to the food chain were reviewed by Nicholson et al. (2000). The measures identified included the manipulation of diet and dietary additions, minimum storage periods (typically 90 days) for slurry and solid manures, avoidance of recontamination of stored manure, slurry treatment, solids composting, land spreading methods, minimum harvest intervals, minimum grazing intervals and general good practice.

3.1. DIETARY AND MICROBIAL MANIPULATION

3.1.1. Cattle

There has been little work focussed on pathogen control within the gut of the animal. Faeces, in the form of pats or pellets from grazing animals, slurry or solid manure, can act as a point source for infection of animals and man either directly or indirectly, through an array of cyclical events. Many factors have been shown to increase faecal pathogen shedding by herbivores and include, fasting, dietary changes, seasonality, age and physiological state of the animals, and the use of antiprotozoal feed additives.

Significant reductions in gut pathogen numbers could potentially be achieved at relatively little cost to the farmer, although at this stage, without further research, it is difficult to give estimates of log10 reductions in pathogen loads. Potential measures include (Chambers, 2003):(a) The development of calf milk replacement feeds that improve the host's immunity

to infection by unwanted pathogenic micro-organisms. The calf represents a high risk pathogen host.

(b) The development of novel silage inoculants that could not only enhance the nutritive value of silage, but also act as a vehicle for supplying lactic acid bacteria that produce products antagonistic to pathogenic bacteria. Silage inoculants are widely used by farmers and could also include beneficial microorganisms capable of competing with pathogens for gut attachment sites and thus eliminating pathogens.

(c) The introduction of novel forages through conventional breeding programmes, into agricultural swards that either reduce the survival of pathogens on the sward with time after their deposition, or release antimicrobial products on ingestion to

52

reduce or eliminate pathogens in the gut, as has been suggested by Davies et al. (2001).

3.1.2. Pigs

Any feed additives or dietary manipulation carried out to increase feed efficiency and growth, by means of suppressing the presence of harmful bacteria in the gut, are likely to have the potential to reduce the level of pathogenic bacteria in pig faeces.

The relevant types of feed additives registered under the Feedingstuffs Regulations, are antibiotics, probiotics, organic acids, prebiotics and the metals zinc and copper. The majority of antibiotics, which were accepted for use as feed additives, have now been banned and it is thought that those remaining will also be withdrawn soon due to concerns over residues in tissues and the evolution of resistant bacterial strains. The dietary addition of high levels of zinc and copper are also now coming under scrutiny, and there are proposals within the EU to limit their addition to recommended nutritional requirement levels only. There are few published data in the scientific literature on the effect of other feed additives on the bacterial load of the gut and/or the faeces. The use of liquid feeding regimes may also have the potential to reduce pathogen loads in faeces. Studies in the UK and Denmark have shown that fermentation of liquid diets by naturally occurring lactic acid bacteria can reduce the pH of the diet to around 4.0, which reduced the number of bacteria in the gastrointestinal tract by a similar amount to an in-feed antibiotic

3.1.3. Poultry

Antibiotics The routine treatment of poultry with in-feed medicinal products has conferred robust protection against a number of bacterial pathogens and disease syndromes. However, concern over the possibility that resistance among farm animals to commonly used antibiotics might be transferable to humans, has led to the withdrawal of a number of products throughout the EU.

ProbioticsAlternatives to antibiotic feed additives are available and their use is increasing. Products work by manipulating the gut microflora to lessen the impact of pathogenic challenges.

VaccinationPoultry are currently vaccinated against a wide range of bacterial and viral diseases. A recent innovation has been the use of vaccines against Salmonella infection. This vaccine is used in laying hens and breeding birds. In 2001, a new combined vaccine was put on the market to protect against both S. enteritidis and S. typhimurium.

GeneralThe majority of laying hen flocks are vaccinated against S. enteritidis, but antibiotics are not generally added to layer feeds. In contrast, broiler chickens are not vaccinated, but may receive in-feed antibiotics. Some broiler producers use probiotics in the feed when chicks are young in order to help them to establish a healthy gut flora; this is not normal practice for laying hens.

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It is unlikely that antibiotics (or probiotics) would be an acceptable way forward for laying hens. The industry has worked hard to shed the public perception that birds are fed on chemically enhanced diets. For broiler producers, the decision to include antibiotics in feed is taken usually in conjunction with veterinary advisers on the basis of the site disease history, and the policies of supermarket buyers.

3.4.1. Summary

In general, manipulation of gut micro-organism levels using feeds containing synthetic or naturally occurring antimicrobials, or using feeds to change the gut pH, are areas which are complex and at present poorly understood. Whilst initially favourable research has been undertaken it is possible that such approaches to pathogen control may be useful only as short-term solutions to reducing bacteria or pathogen levels in manures. More research is required in this area to determine the long-term implications of individual dietary treatments.

3.2. MANURE STORAGE AND TREATMENT

3.2.1. Storage

Results from work funded by the Food Standards Agency were used to provide information on pathogen die-off during storage. Generally, numbers of Salmonella, E. coli O157, Listeria and Campylobacter in batch-stored solid manures (dairy cattle FYM, pig FYM and broiler litter) declined rapidly to below levels of detection after around one week. In batch stored dairy cattle slurry and dirty water, numbers of Salmonella, E. coli O157 and Campylobacter generally declined to below levels of detection after 3 months (Nicholson et al., 2002b).

3.2.2. Slurry Treatment

i. Anaerobic digestion.There is a substantial scientific literature on the development of and research into anaerobic digestion. Detailed review material is available from a number of sources (Monnet, 2003; Svoboda, 2003; Burton and Turner, 2003) and thus only an outline is included within this report. Anaerobic digestion is achieved by utilising micro-organisms to break down complex organic substances, in the absence of oxygen, in a heated enclosed digester vessel, at temperatures between the extremes of 25 and 700C. Anaerobic digestion is commonly carried out on pumpable slurries, although more recently, high solids content (20-40% DM range) plug-flow reactors have been developed (Monnet, 2003). For slurries, the optimum dry matter content is 6-8% and it is likely that the majority of cattle and pig slurries could be successfully digested, provided that excess bedding was excluded. One of the products of the process is biogas, a mixture of methane (60-70%) and carbon dioxide (30-40%).

The process can be either mesophilic (25-450C) or thermophilic (55-700C); although the latter process gives higher gas yields, equipment is more costly to install and is normally used only in large centralised digesters. All digesters that are in commercial use in the UK operate on a continuous process, with a nominal retention time of 12-20

54

days; the lower figure for pig slurries and the higher for cattle slurries (MAFF, 1998c). The digestion process does not significantly reduce the volume of the slurry, nor its nutrient content.

Typical residence times and temperatures adopted in a farm scale mesophilic digester would be 15 days retention at 350C. In a thermophilic process, typical times and temperatures would be a minimum of 10 days at 550C. Some centralised anaerobic digestion plants in Denmark have an additional 700C pasteurisation process built in, which adds significantly to capital costs.

Effects on pathogens. A study of pathogen reduction in 10 large-scale Danish Biogas plants indicated that, in mesophilic systems, the Pathogen Reducing Effect (PRE) was modest and corresponded to a log10 reduction of 1-2 units. In contrast, thermophilic plants were capable of achieving a PRE log10 reduction of 4 units (Bendixen, 1999). Similar investigations in Germany confirm that either a thermophilic process or pasteurisation at 700C for one hour is necessary to achieve the inactivation of pathogens (Böhm et al., 1999). These findings generally concur with UK surveys undertaken on mesophilic digestion of sewage sludge where an average log10

reduction of 2 units was observed (UKWIR, 1999a and b). A study undertaken specifically on inactivation of viruses in animal slurries concluded that fermentation temperatures at or above 550C were the most important virucidal factor, and that thermophilic processes were generally likely to kill the majority of viruses (Pesaro et al., 1999).

ii. Aerobic treatment.The aerobic treatment of slurry is normally carried out only for odour control purposes which are achieved via the microbial breakdown of the many compounds (organic and inorganic) that contribute to manure odour. This also results in a reduction of pathogen numbers. Improved physical and chemical characteristics are other significant benefits. Aerobic treatment is generally only suitable for separated slurry or dilute effluents (<3% dry matter) containing no bedding (Defra, 1998). Unseparated pig slurry can be aerated, but cattle slurry may require both dilution and mechanical separation for the process to be trouble-free and effective. A number of approaches are used to achieve aeration, either in-situ in the slurry store, or in a purpose-designed aeration vessel. These range from blowing compressed air through porous diffusers with very small outlets, or by entraining air in a fast moving stream of liquid in submerged nozzles, or floating devices with discs or rotating impellers (Cumby, 1987a). Temperatures of the aerated slurry will rise by 5-250C depending on the slurry analysis, degree of aeration, tank insulation and ambient temperature. The process can create foam and its control can be a problem (Cumby, 1987b). Continuous flow systems can reduce slurry odours with a retention time of 1-2 days provided that a reasonably constant and well-mixed flow of slurry is maintained.

Effects on pathogens. A 90% reduction of Salmonella occurred within 2 days in aerated slurry (Jones and Matthews, 1975). Aeration also increased the reduction of Campylobacter in dairy slurry (Stanley et al., 1998). Aeration of farm-scale slurry tanks stored under winter ambient conditions increased temperatures to between 190C and 400C and after 2-5 weeks reduced Salmonella levels by over 99% for both cattle slurry contaminated with S. infantis and pig slurry contaminated with S. typhimurium.

55

In this study similar effects were found for Yersinia, Listeria, faecal coliforms, enterococci and coliphages (Heinonen-Tanski et al., 1998).

iii. Slurry additives. Lime.

The addition of lime has been commonly used to treat slurry following disease outbreaks, such as foot and mouth. A pH of 12 for at least 2 hours is generally regarded as sufficient to produce good pathogen kill (DOE, 1996).

Acid.Lowering slurry pH by the addition of nitric or sulphuric acid has been used as a method of reducing ammonia emissions (Stevens et al., 1997). Reduction to pH 4.0 for a period of two days is generally regarded as effective in killing pathogens, except E. coli O157, which is extremely acid tolerant (Russell and Jarvis, 2001) and would probably require a reduction to pH 3.0. However, direct addition of corrosive mineral acids is unlikely to be taken up by farmers on both health and safety and economic grounds, as already illustrated by the decline in use of acids as silage additives by farmers.

Carbohydrates.Stimulating the growth of indigenous microbial populations using additives that are carbon sources, such as starch or sugar, and reducing the pH by lactic acid and volatile fatty acid production (Subair, 1995) have potential to reduce pathogen numbers. Non-proprietary carbon based additives such as molasses and soluble starch used at a rate of 1 to 2% w/w in the slurry were sufficient to lower the pH, which considerably reduced the emission of ammonia and odour (Hendricks and Vrielink, 1997). Additives that are carbohydrate substrates would need to be used at the 1% w/w addition level, which is likely to be impractical and expensive at the farm scale, although more research is needed.

Effects on pathogens. Enterobacteria declined to less than 103 colony forming units (cfu) ml-1 in an untreated control after 16 days and had completely disappeared where 5% glucose was added and the pH had fallen to <4.5 (Hobbs, 2000). A smaller and less dramatic decline was seen at the 1% and 2% levels, where the final pH was between 6.2 and 6.5.

3.2.3. Solids Composting

Composting is a relatively simple technology and is defined as the controlled biological oxidation of organic matter to produce a stable and humified product (Groenhof, 1998). The process has many benefits including a reduction in pathogen content.

The composting process, from start to finish, takes approximately 3-12 months. This is shown in Figure 6 and comprises the initial thermophilic stage (Phase 1), lasting up to 3 months, followed by a maturation phase lasting up to 9 months. During this time the material will have reduced by about 50% of its original volume and become ‘stabilised’. In addition to this the temperature within the heap will have reached between 60-700C, and will have killed any pathogens associated with the feedstock materials.

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There are three main types of composting systems: 1. Windrow 2. Static pile with forced ventilation 3. In-vessel

Control over the composting process increases from windrow to in-vessel composting, as does the capital cost. The labour cost decreases in the same succession, with the overall running cost mainly depending on the costs of labour and energy.

Figure 6. Typical temperature profile of a compost heap with its associated phases of microbial activity (Groenhof, 1998).

i. Windrow composting. The feed stock is piled in long rows (windrows) and turned at intervals using mobile equipment like tractors with front loaders or compost-turners. The most common method, the conventional windrow, is aerated through natural ventilation (convection and diffusion) and also during turning which is required to ensure homogenous composting. This process requires an extensive area of ground which can be compacted soil, but more ideally a concrete base with the facility for containing the leachate. In regions with high rainfall, leachate production can be reduced and improved control of composting achieved by roofing the composting area. Virtually all of the farm manure composting operations use windrow methods.

ii. Static pile composting. This process uses an active aeration system. Perforated pipes are laid on the floor or in the floor channels and are covered with porous material like straw, wood chips etc.,

Temp (C)

55

70

Mesophilic MesophilicThermophilic

Pathogen Kill

Phase 1 Maturation

57

which aids the efficient distribution of air. The feed stock is then piled on the base and covered with a layer of matured compost to provide thermal insulation and partial odour removal. Aeration, controlled by temperature feedback, is used to sustain the pile in an aerobic state, to maintain the temperature of the pile and to control the moisture content of the pile. The latter helps mainly in the final stage of composting, when the increased aeration rate contributes to compost drying.

iii. In-vessel composting. To ensure homogeneous composting processes with temperature control and therefore inactivation of pathogens and odour reduction, in-vessel composting systems are used. In-vessel composting is usually a multistage process. Pre-composting or full composting is achieved in the first stage in a bioreactor, with the final composting and maturing in windrows. The most common types of reactors are horizontal and vertical plug-flow and an agitated bin reactor. This type of composting system, which is well controlled and thoroughly mixed, is faster than the other two systems, but is more complicated to control and the processing mechanisms are expensive.

3.3. MANURE SPREADING AND GRAZING ANIMALS

3.3.1. Spreading Method

There is potential for pathogens to be transmitted in air during the land spreading of manures, particularly by slurry spreading methods which generate aerosol-sized droplets. At present, broadcast spreading is the most widely used slurry application technique in the UK, with >90% of slurry spread this way (Smith et al., 2000a; 2001c). Broadcast slurry spreading techniques carry a high risk of aerosol generation and have been shown to disseminate pathogens over distances of up to 650 m, especially under windy conditions (Hahesy, 1995). It is possible that adjacent crops, grazing land and surface waters could become contaminated by aerosol-borne pathogens. Band spreaders and shallow injectors give little risk of aerosol generation, but reduce the slurry surface area left on the soil compared with broadcast spreading, so the slurry will dry less quickly and be exposed to less UV radiation, thus increasing the potential for pathogen survival. Similar comments apply to deep slurry injection, although pathogens will be removed from the soil surface and are less likely to contaminate growing crops or be ingested by animals.

In order to reduce ammonia and odour emissions, it is recommended (MAFF, 1998c) that manures are incorporated into the soil soon after spreading; this will protect manures from UV radiation, reduce drying and may therefore lead to slower pathogen die-off than where manures are left on the soil surface. However, it will lower the risks of water pollution from surface runoff and drainflow soon after manure application.

3.3.2. Direct Deposition to Water Courses

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In terms of the amount of manure applied to land and deposited while grazing, the amount directly voided into streams by livestock and that entering accidentally during manure application are likely to be relatively small. However, the risk posed, especially by fresh excreta could be of great importance in terms of peak concentration effects. Manure directly voided into water courses represents a highly concentrated source of pathogens which is instantly delivered avoiding the process of mobilisation. To avoid such contamination, water courses should be fenced off from livestock and an independent water supply used for the watering of animals. Fords and crossing points should, where possible, be removed from use and alternative routes found. Accidental contamination during manure spreading activities can be completely avoided with farmer vigilance and adhearance to the Code of Good Agricultural Practice for the Protection of Water (MAFF, 1998a). This clearly advises that applications of manures should not take place within 10m of a water course and 50m of a spring, well or borehole.

3.3.3. Die-off in the Soil Environment

After the introduction of micro-organisms into soil, either through the direct deposition of excreta or following manure spreading, most bacteria and protozoa have difficulty in surviving. Common inhabitants of the gastro-intestinal tract are not adapted to survive in soils, their preferred habitats for optimal growth are warm (c.37ºC), moist and nutrient rich environments – these conditions are not generally present in soils. A large number of factors affect pathogen survival in soils, including nutrient availability, moisture status, oxygen supply, temperature, organic matter content, pH, soil type and microbial predation. A typical survival curve is demonstrated in Figure 7 (Oliver et al., in press). Nicholson et al. (2000) concluded that temperature was the single most important factor determining pathogen survival in handled manures, with high temperatures (>55ºC) and freezing conditions being very effective in killing pathogens. However, within normal soil temperature ranges (5-25ºC) pathogen survival is generally longer under lower temperature conditions.

Grazing livestock, as well as spread manures, provide a potential source of pathogen losses. Particularly in the case of cattle excreta, where a crust can form on the surface, providing a protective environment for microbial survival (particularly from UV light) until the next rainfall event provides a potential means of transfer. Hence, faecal deposits may provide a long-term continuous source of microbial pollution.

In a comprehensive review of the scientific literature, Nicholson et al. (2000) reported maximum soil survival time for bacterial pathogens as up to 3 years for Salmonella and up to 3 months for Cryptosporidium (Table12).

Table12. Pathogen survival times in soil.

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Pathogen Maximum reported survival timeE. coli O157 Up to 6 monthsSalmonella Up to 3 years

Listeria Up to 2 yearsCampylobacter Up to 20 days

Cryptosporidium Up to 3 monthsGiardia No data

Recent FSA funded research (Hutchison et al., 2002; Nicholson et al., 2002b) recorded typical decimal (i.e. one log10) reduction times of 1-2 days (range 0.84-2.75) during the initial steep part of the pathogen die-off curve following the land application and soil incorporation of a range of manure types (cattle/pig slurry, cattle/pig FYM and poultry manure), Table 13. The maximum survival time of E. coli O157, Salmonella and Campylobacter was generally no more than 1 month after application to both sandy arable and clay loam grassland soils, with Listeria showing a tendency for survival at low levels for longer than 1 month, particularly on the clay loam grassland soil. Over all, the patterns of die-off and survival longevity were similar for the four microbial pathogens. Hence, we are confident that generic E. coli survival can be used as a robust indicator of the survival of microbial pathogens (Nicholson et al., 2002b; Vinten et al., 2002), including intestinal streptococci. Indeed, the studies of Vinten et al. (2002) indicated that the die-off rate of E. coli O157 was the same or quicker than generic E. coli.

Table 13. Decimal reduction times of pathogens in animal manures spread to land calculated from the initial linear part of the die-off curve.

Manuretype

Treatment Salmonella E. coli.O157

Listeria Campylobacter Mean

Solid: D values1 (days)Dairy cattle Immediate 2.18 2.31 2.94 2.91 2.59Dairy cattle Delayed 1.04 5.06 2.59 0.98 2.42Dairy cattle Unincorporated 1.12 1.04 3.44 1.13 1.68

Pig Immediate 2.90 1.00 1.32 1.58 1.70Pig Delayed 1.28 0.95 0.78 0.74 0.94Pig Unincorporated 1.33 0.67 0.93 0.79 0.93

Poultry Immediate 1.28 1.05 1.00 1.11 1.11Poultry Delayed 0.76 0.66 0.81 1.15 0.84Poultry Unincorporated 1.66 0.75 0.97 0.79 1.04

Slurry:Dairy cattle Immediate 3.98 1.62 1.37 4.03 2.75Dairy cattle Delayed 1.11 0.89 0.88 0.86 0.94Dairy cattle Unincorporated 0.89 0.99 0.88 0.91 0.92

Pig Immediate 2.44 4.89 1.13 1.03 2.37Pig Delayed 1.89 0.97 0.66 0.63 1.04Pig Unincorporated 0.79 1.76 0.84 2.26 1.41

1D values = die off time in days.Figure 7. Factors contributing to survival of micro-organisms introduced to soils (Adapted from Oliver et al. in press)

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3.4. METHOD EFFECTIVENESS

Chambers (2003) assessed the pathway reduction potential of a large number of potential control measures by reference to existing literature and research work funded by the FSA (Nicholson et al., 2002a;b; Hutchison et al., 2002). The

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approaches were combined into a series of measures grouped under diet/housing, manure storage and treatment, manure spreading and livestock grazing and farm runoff (Table 14). The effectiveness and reliability of the suggested measures to reduce microbial pathogen loads was assessed, both in terms of maximum potential reduction (from one to six log10 reduction) and reliability (i.e. is the process likely to be easily controllable under typical farm conditions). The results are summarised in columns three and four of Table 14. Measures associated with extended storage periods or the treatment of solid manures and slurries generally gave the best reduction in pathogen loads, ranging from two to six log10 reductions.


Potential measures to control pathogens associated with animal manure management may influence the magnitude of emissions of ammonia (NH3), nitrous oxide (N2O) and methane (CH4) to air, and nitrate (NO3) and phosphorus (P) losses to water (so called ‘pollution swapping’). A detailed review of this area was undertaken by Chambers (2003). The main conclusions for the most effective slurry storage and solid manure treatment options to control pathogens were:

Longer storage periods for slurry (e.g. 3 months storage in two separate tanks) were estimated to increase ammonia emissions by 3.4 kt N/annum and methane emissions by 3.8 kt CH4/annum. Recent Inventory estimates of losses from these sources under present management practices were c.8.3 kt N/annum (1999) and c.4.6 kt CH4/annum, respectively. Nitrous oxide emissions from slurry stores were considered to be unaffected.

Storing all solid manures (for 3 months) before land spreading was estimated to increase ammonia losses by c.1 kt N/annum (Inventory estimate of present losses 1.4 kt N/annum), nitrous oxide emissions by c.1 kt N2O/annum (Inventory estimate of present losses c.4 kt N2O/annum) and methane emissions by c.1.3 CH4/annum (Inventory estimate of present losses c. 25 kt CH4/annum). However, following solid manure storage, ammonia emissions at land spreading were estimated to be reduced by 20 kt N (38 kt N down to 18 kt N), as a result of manure total and ammonium-N content decreases during storage. Thus, the net balance from the storage and land spreading of solid manures would be a decrease of c.19 kt N/annum.

Increasing slurry storage capacity to >3 months on all livestock farms and storing all solid manures before land spreading was estimated to increase nitrate leaching losses following slurry applications by 3 kt N/annum (largely as a result of changes in the timing of applications), and to decrease losses following solid manure spreading by 5 kt N/annum (largely as a result of manure total and ammonium-N content decreases during storage).

Table 14. A summary of measures suggested to control pathogens in manures and their effectiveness and reliability in reducing loadings (adapted from Chambers, 2003).

Phase Possible measures to be considered Log 10 Reliability

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1: D

IET

/H

OU

SIN

G

1. Dietary manipulation/ microbial manipulation of pathogen load in guta) Cattle: fresh and conserved forage, use of silage additives+ b) Poultry: vaccination/probiotics/antibioticsc) Pigs: use of copper and antibiotics

+

+

2: S

TO

RA

GE

/TR

EA

TM

EN

T

2.1 Slurrya) Store slurry for 30 daysb) Store slurry for 90 daysc) Provide additional slurry store(s) to avoid recontamination of stored slurry with fresh material i.e. batch storage for 90 days.d) If no storage, consider treatment:

(1) anaerobic digestion(2) anaerobic digestion plus pasteurisation

(3) aeration (4) additives – LIME – ACID(5) additives – carbohydrate substrates

2.2 Solidsa) Batch store solids for at least 90 days prior to application.b) Compost with thorough mixing : twice within first week and > 55OC for 3 days.c) Avoid recontamination with fresh manure (provide additional manure pad(s) if necessary).

GOOD PRACTICE

GOODPRACTICE

3:SP

RE

AD

ING

OF

MA

NU

RE

S A

ND

G

RA

ZIN

G A

NIM

AL

S a) Incorporate solid manure or slurry prior to drilling. b) Avoid direct application of manures or direct deposition of faeces into a watercoursec) Allow 7 days between manure application/ livestock grazing and onset of runoff or drainaged) Allow one month between manure application/livestock grazing and onset of runoff or drainage

()

GOOD PRACTICE

()

GOOD PRACTICE

4:FA

RM

R

UN

OFF

E

TC

a) Contain hardstandings runoffb) Contain runoff from ‘wet’ (and fresh) farmstead middens/woodchip corralsc) Minimise runoff from field manure heapsd) Minimise runoff from tracks used by cattlee) Maintain secure slurry storage

GOOD PRACTICE “ “ “ “

GOOD PRACTICE “ “ “ “

Pathogen reduction: one star for each order of magnitude: maximum score =e.g. Reliability: = very variable, = totally consistent, with intermediate scores+ Although only one assigned, this approach has some potential, but more research is needed to assess this.() may increase survival within the soil, but the pathogens are less likely to contaminate water systems as they are ‘protected’ from loss within the soil matrix.

5. BIOLOGICAL OXYGEN DEMAND

1. INTRODUCTION

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Most agricultural organic wastes contain substantial quantities of chemically and biologically degradable material, which means considerable potential for pollution once these wastes or effluents gain access to watercourses. Biochemical oxygen demand (BOD). Together with the chemical oxygen demand (COD), BOD is used as a measure of the polluting potential of organic wastes in water bodies. A laboratory test measures the amount of dissolved oxygen (in mg l-1) consumed by biochemical action when a sample is incubated at 20oC for a given number (usually 5) of days (BOD5).

Chemical oxygen demand (COD) is a measure of the amount of oxygen consumed in the combined microbial oxidation of decomposable and inert organic matter and the oxidation of reduced substances in water. The laboratory test involves refluxing the sample with sulphuric acid and potassium dichromate. The COD is always higher than the BOD, but measurements can be made in a few hours while BOD5

measurements take five days.

There would probably be a reasonable relationship between BOD and COD for specific effluent types, but probably not accorss a range of effluents. BOD is more relevant for agricultural and natural water systems since BOD relies on biological activity while COD reflects the stronger oxidising effect of chemicals.

Examples of the pollution potential of a range of these agricultural effluents are shown in Table 15.

Table 15. Typical BOD5 and COD of livestock manures and related effluents.

Pollutant BOD5 (mg l-1) COD (mg l-1) SOURCECattle Slurry 10,000-20,000 110,000 Anon, 1986, Robertson, 1977

Pig Slurry 20,000-30,000 90,000 Anon, 1986, Robertson, 1977Poultry Slurry 25,000-35,000 130,000 Anon, 1986, Robertson, 1977

Dilute Parlour andYard Washings 1,000-2,000 5000-11,000 Anon, 1986, Vanderholm, 1984

Vegetable Washings 100-3,000 Anon, 1986, Vanderholm, 1984Milk 140,000 Anon, 1986, Vanderholm, 1984

Silage Effluent 12,000-83,000 Anon, 1986Effluent from

brewer’s grains 12,000-43,000 Anon, 1986

The values quoted here are from general reference sources where the definition of yard water/washings are more strictly adopted.

The microbiological properties of livestock wastes and effluents are dependent on the growth of micro-organisms contained in the wastes. The growth of micro-organisms is under continuous flux and is influenced by environmental conditions such as temperature, effluent pH, availability of essential nutrients and presence or absence of oxygen. The presence or absence of micro-organisms influences the pollution potential of the waste, its behaviour during storage and its chemical composition. Most of the micro-organisms found in wastes require oxygen in order to support growth. Thus they are in direct competition with other forms of life when discharged into a watercourse or spread on land. It is this requirement for oxygen that is measured in terms of BOD. It must be borne in mind however that not all the organic matter contained in waste is capable of microbial degradation.

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Establishment of the dominant species of micro-organism and overall microbial metabolism is heavily influenced by a number of environmental factors such as temperature, nutrient and pH status. Oxygen, which is an important environmental factor, can only be supplied by external sources, be it by simple diffusion from the atmosphere or by mechanical means. When oxygen is freely available, aerobic conditions exist. Under these conditions a state of equilibrium is eventually reached whereby stable end products are produced. In contrast, a complete lack of oxygen gives rise to anaerobic conditions yielding end products which are not stable and which are capable of further breakdown by aerobic organisms. In addition to aerobic and anaerobic organisms, wastes may contain large numbers of what are know as facultative anaerobic organisms, which can grow both in the absence or presence of oxygen.

In practice, storage of waste in channels and tanks tends to favour the establishment of anaerobic conditions. Because of environmental conditions, incomplete anaerobic breakdown of organic matter may occur, giving rise to wastes with a high oxygen demand and the production of noxious gases. The ratio of COD to BOD5 can give a useful guide to the actual biodegradability of an effluent. Wastes containing a high proportion of biodegradable substrate will have a low COD to BOD5 ratio whilst those containing a high proportion of non-biodegradable solids will have a high COD to BOD5 ratio. For example, pig slurry (often containing little or no bedding material) will have a lower COD to BOD5 ratio than slurry from cattle fed a ration containing a high proportion of roughage, giving large quantities of food residues in the slurry and which may also contain significant amounts of bedding. In the context of waste treatment it can be seen that the problems associated with treating cattle slurry are likely to be greater than those encountered with pig slurry because the former contains larger concentrations of non-biodegradable material. However, pig slurry will usually be associated with greater pollution risk because of its relatively high BOD (Table 15) and nutrient content (Anon, 2000b).

Impact of BOD in waterThe main effect of loading surface waters with organic matter is the rapid depletion of available oxygen as a result of the increased microbial activity stimulated. In stagnant or slow-moving waters, anaerobic conditions quickly develop with the associated generation of foul odours and, in the longer term, the reduction in biodiversity of the system. The introduction of effluents with a high BOD will often be in association with high ammonium (NH4

+-N) concentrations and the nitrification of NH4+-N will

also remove some available oxygen from the system. Where pollution of streams occurs, the impact can be exacerbated by the apparently limited level of dilution that occurs in moving water (Burton and Turner, 2003). Rather, a plug-flow effect is invoked, in which fish and higher life forms are swept downstream by an anaerobic “front”. Pollution incidents involving slurry or other high BOD sources, have usually been very pronounced, with a high and very visible fish kill. Of course, in the fast-flowing stream, a rapid restoration of aerobic conditions can quickly follow (as long as the pollution source is blocked), but for ponds or other poorly aerated (slow moving) waters, the damage is long-lasting.

65

Trends in biological and chemical water quality have been observed in systematic studies of the type reported by Foy and Kirk (1996). Water quality, measured on a fisheries ecosystem scale of 1 (good/salmonid) to 6 (bad/fish absent), of forty-two lowland streams in two Northern Ireland river catchments was inversely correlated with the stocking rate of grazing animals. A decrease in water quality of one class was associated with an increase in the combined grazing/stocking rate of cattle and sheep of 0.6 dairy cow equivalents/ha. This dairy cow equivalent (DCE) stocking rate was significantly correlated with maximum BOD and total NH4

+-N concentrations and minimum dissolved-oxygen levels. The worst pollution events, with BOD5

concentrations in excess of 100 mg l-1, occurred at the end of May and were caused by discharges of silage effluent. Smaller BOD5 peaks, which occurred in late winter and early spring, were related to the land spreading of animal slurries. It was concluded that poultry and pig farms did not appear to have a major impact on water quality. Similarly, Berka et al., (2001) found significant negative relationships between surplus N applications and dissolved oxygen, while NH4

+-N and nitrate (NO3--N)

concentrations in the wet season were positively correlated with the N surplus.

In another catchment study, water quality was measured in 42 streams in the Colebrooke and Upper Bann catchments in Northern Ireland over the period 1990–1998 (Foy et al., 2001). Despite ongoing pollution control measures, biological water quality, as determined by the invertebrate average score per taxon (ASPT) index, did not improve and there was no appreciable decline in recorded farm pollution incidents. However, a greater proportion of incidents were recorded from less polluting discharges such as farmyard runoff. In contrast, there was an improvement during 1997 and 1998 in annual, chemical water quality classification based on excedence values (90th percentiles) for dissolved oxygen, BOD5 and NH4

+-N concentrations. In 1998, 11.9% of streams were severely polluted compared to 26.2% in 1990, while the proportion classed as of salmonid water quality, increased from 40.5% in 1990 to 59.6% in 1998. Although water quality in 1996 did not improve relative to 1990 values, there was a notable increasing trend from 1990 in the numbers of samples taken during the summer which had good water quality with low NH 4

+-N (<0.6 mg N l-1) and high dissolved oxygen (>70% sat). The trend for samples with low BOD (<4 mg l-1) was more erratic, but an improvement was apparent from 1994. These improvements in chemical water quality suggest that farm pollution declined after 1990. The fact that this was not reflected in stream biology may reflect the limited time scale for biological recovery; an important factor preventing biological recovery may be the high pollution capacity of manures and silage effluent, so that even reduced numbers of farm pollution incidents can severely perturb stream ecosystems.


2.1. SOURCES

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Several major sources of BOD are readily identified (Table 15), to which can be added likely significant sources of NH4

+-N highlighted in section 2.1 in the ammonium section. It seems highly likely that a number of major sources are common to both of these pollutants (and to some extent pathogens).

Table 16. Mean analyses of dirty water samples from 20 dairy farms (1991/92) (after Cumby et al., 1999)

Analyte Sampling timeFebruary June September

Mean sd (%)1 Mean sd (%)1 Mean sd (%)1

Total solids (%) 0.57 89 1.65 82 1.03 77BOD5 (mg l-1) 2660 68 9670 100 7450 88COD (mg l-1) 6550 84 17300 84 16300 86

NH4+-N (mg l-1) 310 106 580 84 480 86

Total N (mg l-1) - - 950 81 700 86pH 7.60 7 6.38 20 6.65 12

Phosphate2 (mg l-1) - - 490 70 340 65Potassium (mg l-1) - - 1500 70 850 48

Notes: - no analysis available1 Standard deviation as % of mean2 Phosphate expressed as PO4

3-

Uncontained washings from animal collection/feeding yards are significant sources of potential pollutants including BOD. Dirty water quality, including from yard runoff and wash-water is likely to vary considerably, not just according to source, but with time of year (Table 16). Recent survey data (Defra project WA0523) suggest that ~80% of these yard washings are collected, the remainder presumably seeping into proximate fields and ditches. Leachates from field heaps of solid manure may also have a high BOD and therefore the position of such heaps in respect to drains and ditches is critical if transfers to water are to be avoided. Dairy tracks have received little attention, but their impermeable nature and high concentration of animals twice a day, would suggest that they could be sources of BOD runoff with potential transfer to proximate fields and ditches. Samples of yard runoff collected from a range of yards indicated that concentrations of potential pollutants in runoff from hard standings require that such runoff is either collected or adequately treated before discharge to a watercourse (Table 17). Silage clamp effluent should be collected and stored before spreading to land as it has a high BOD5 (up to 60,000 mg l-1). Big bale silage should pose less risk as it generally has a higher dry matter (DM) and produces less or no effluent.

Table 17. Concentrations of potential pollutants in runoff from hard standings (results from Defra contract WA0516).

Dry matter%

NH4+-N

mg l-1NO3

--Nmg l-1

Total Nmg l-1

MRPmg l-1

Total Pmg l-1

BOD5

mg l-1

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Dairy cowyards

0.16(0-0.92)

239(57-603)

0.0(0-0.2)

1021(160-2960)

28(9-82)

54(12-115)

2955(1-7300)

Sheep handling 0.10 7 0.0 11 12 13 191General

purpose yard0.01

(0-0.01)54

(45-63)42.0

(0-83.9)39

(38-40)0.1

(0-0.2)3

(0-5) NDSugar beet

storage area0.21

(0.17-0.25)1

(0-2)22.2

(19.6-24.8) ND ND ND NDValues in parentheses are ranges. ND = not determined.

Data on vegetable washing and processing wastes are also presented in Table 18; factors likely to impact on the quantity and analysis of the effluent include crop, weather and time of year, and, especially, soil type. Washing carrots from peat soils requires around twice the volume of wash water than carrots from mineral soils.

Table 18. Waste characteristics from washing and processing of vegetable crops.

Vegetable Production Process Effluent

Period Volume (m3 t-1) BOD mg l-1

Suspended solids mg l-1

Beetroot Sept-Feb Canning 4.5 4,000 1,250

Broad Beans July-Aug CanningWashing 6.7 300-700 -

Carrots Aug-Mar

(Mineral Soil)Washing

(Peat Soil)Canning

Dehydration

2.7

5.55.617.8

240

801,4001,220

4,120

2,1902,000700

Parsnips Oct-Feb Washing 2.5 - -

Peas June-AugWashingFreezingCanning

1.618.06.7

8,000-9,0001,000

1,100-4,000

---

Potatoes May-Mar

WashingCanning

Lye PeelingDehydration

5.616.530.0

1,4002,570300

2,0001,0001,200

Spinach April-May Canning 32.0 300 600Source: Wheatland and Bourne, 1970; Holdsworth, 1968.

2.2. MOBILISATION

Annual statistics on water pollution in England and Wales (Anon, 1999) provide information relevant to incidental BOD transfers to surface waters. Of 17863 substantiated incidents in 1998, 2050 (11%) were caused by agriculture. Of these, in 1998 the major sources were dairy and beef cattle, together accounting for 50% of the total. Almost the same numbers (2026 incidents) were classified as caused by “organic wastes”, of which the largest numbers were attributed to cattle slurry (23%). Yard washings (13%), silage liquor (6%) and dairy washings (1%), together accounted for a further 20% of pollution incidents. Altogether, livestock manures, either directly or indirectly, accounted for 55% of these agricultural pollution incidents.

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Failure or mismanagement of slurry storage structures has often caused water pollution. Numbers of substantiated water pollution incidents in England and Wales (1987-1996) from agriculture and from slurry/manure stores are recorded by the Environment Agency (Anon, 2000a). Slurry storage accounted for between 75-85% of all incidents associated with manure storage each year. However, major (category 1) incidents reduced from 99 in 1991, to 22 in 1998 (Anon, 1999). Types of slurry store causing incidents are not categorised by number and a detailed breakdown of figures is not available for 2000 and 2001. Pollution incidents of this nature will normally be associated with significant transfer of BOD and NH4

+-N to water, because of the source materials. The risk of both BOD and NH4

+-N enrichment of receiving waters following a catastrophic failure or mismanagement of slurry storage must be particularly acute.

Taking another of the major incidental sources, runoff from farm steadings, the mean and range of runoff BOD5 concentrations found in the hardstandings study (WA0516; Table 17) were not dissimilar to those in dirty water reported by Cumby et al, (1999). The typical dirty yard area for dairy farms is estimated at 7.8m2/cow, from “weighted” estimates using the data of Brewer et al., (1999) and of Defra contract NT2402. An estimate of potential loss of BOD from hardstandings runoff over a typical winter period can be made. This requires assumptions about the proportion of runoff (Smith et al., 1984), as well as typical overwinter rainfall. Thus, for a 100 cow herd:

100 cows x 7.8 m2 (yard area) x 0.25m (winter rainfall) x 0.85 (runoff factor) = 165.7m3

with an average BOD5 at 2955 mgl-1 (Table 17), the loss to proximate fields or ditches might be as high as c. 491 kg BOD for a 100 cow herd.

Sealing of the soil surface by slurry solids seems likely to be an important mechanism for mobilisation of particulate material and nutrients (both associated with solids and in solution), following slurry application on susceptible soils. Research was carried out on N and P losses via surface and subsurface flow following manure application to arable land at ADAS Rosemaund (Smith et al., 2001b). The results showed an increased loss of solids and NH4

+-N in surface water flow during runoff events, from slurry treatments, compared to control plots receiving inorganic fertilizer only, or no treatment, but had little effect on NO3

--N losses by this route. Increasing slurry application rate and, in particular, slurry solids loading, increased solids and NH4

+-N losses via surface runoff. Although runoff BOD was not measured, it is very likely that a similar pattern would have been observed.

Although the categorisation of pollution incidents no longer identifies land application of slurries as a specific source, it is likely that a significant proportion of the c. 470 cattle slurry and the c. 100 pig slurry, substantiated water pollution incidents in 1998 (Anon, 1999), were the result of excessive rates of slurry application. High rates of slurry, when followed soon afterwards by a rainfall event, on susceptible soils (which include grassland with surface compaction), are likely to result in mobilisation of BOD and nutrients, via surface runoff.

2.3. DELIVERY

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2.3.1. Outdoor Grazing

Grazing and outdoor animals clearly represent potential sources of BOD. However, risk of transfer to watercourses is likely to be small for cattle, unless there are summer storms or direct access of cattle to streams. Extended grazing may increase the risk of transfer from grazing cattle, primarily via lateral flow or preferential flow to drains. Leaching losses are likely to be small, as the BOD of liquids will be broken down as it passes through the soil profile. Outdoor pigs pose more of a threat as the land is usually highly disturbed and unstable and is more likely to be subject to erosion. Transfers of effluents with a high BOD from sheep grazing land could be significant via lateral flow, as sheep generally graze all year round and are present on hillsides in wet conditions. Concentrations of NH4

+-N from porous cups under sheep grazing land have been shown to be low (Cuttle et al., 1992) and are likely to be low also in BOD.

2.3.2. Manure and Dirty Water Land Applications

Manure and dirty water applications to land could significantly increase BOD and NH4

+-N transfers to water. There is evidence of rapid transfer of NH4+-N and, hence

BOD, to drains following dirty water applications at a range of application rates (2 – 25 mm) and soil moisture deficit (Anon, 1994). It is clear that application of dirty water to drained land is associated with significant pollution risk, both under dry soil conditions and with the soil at or above field capacity.

The application of dirty water has also been studied on drained clay soils at ADAS Terrington and at IGER Trawscoed (Williams and Nicholson, 1995). Dirty water characteristics from the study were similar to those reported from other studies with BOD5 from 40 to 3155 mg l-1. Dirty water was applied at a range of soil moisture status. At Trawscoed, dirty water applications of 110 m³ ha-1 were made to silty clay loam soils, drained with 75 mm diameter tiles at 20 cm spacing and 75 cm depth with permeable fill to 30 cm depth. There was a calculated soil moisture deficit of 120 mm. Drain flow was initiated within 30 minutes of the irrigator crossing the drain line; the drain flow was discoloured and foul smelling. At the same site, applications of between 60 m³ ha-1 and 110 m³ ha-1 to saturated soil, (with surface ponding and with the drains carrying water), caused flows to increase and become contaminated, also within 30 minutes of the irrigator crossing the drain lines. Drain flows had a BOD5 of between 44 and 90 mg l-1, with NH4

+-N concentrations ranging between 7 and 81 mg l-1 and total P between 9 and 25 mg l-1 (Figure 8). However, applications of 40 m³ ha-1 did not contaminate drain flow.

Figure 8. Impact of low rate dirty water irrigation on drainage water quality, following a 110 m³ ha-1 application to soils with moisture content greater than field capacity (IGER Trawscoed, 1990).

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At ADAS Terrington, in April 1991, on clay loam soils (Wallasea series, Hodge et al., 1984), but without permeable fill over 80 mm plastic pipes, two 130 m³ ha-1

applications to soil with a moisture deficit of 20 mm initiated drain flow that had a BOD5 ranging between 72 and 292 mg l-1, with NH4

+-N and total P concentrations between 8 to13 mg l-1 and 25 to 33 mg l-1, respectively (Figure 9a). In October 1991, applications totalling 240 m³ ha-1 to soil with a calculated moisture deficit of 100 mm did not initiate drain flow. However, in September 1992, one application of 250 m³ ha-1 to soil with a moisture deficit of 20 mm initiated contaminated flow, with BOD 5

values of 118 to 310 mg l-1, (149-179 mg l-1 NH4+-N, and 24-29 mg l-1 total P) (Figure

9b). However, 5 consecutive applications of 50 m³ ha -1 (observing “good practice”, Anon, 1998a), under similar moisture conditions did not initiate drain flow. In January 1993, when soils were at field capacity, one application of 40 m³ ha-1 initiated a trace of contaminated flow.

Similar risks are associated with the application of slurries to drained clay soils. Results confirm the high risk of N and P losses via drain flow (up to 9% of slurry total N applied), following slurry applications to arable clay soils, although the monitoring to date has not included assessment of BOD (Williams et al., 2002). Pollution risk is reduced if slurry is applied to land that is cultivated before the start of winter drainage. The reduction in losses associated with cultivation may reflect the disruption of soil macro-pores and increased mixing of the slurry within the soil matrix, thereby inhibiting the rate of water and slurry nutrient movement through the soil to the drains.

It is clear that the application of either slurry or dirty water to drained land is associated with significant pollution risk, both under dry soil conditions and with the soil at or above field capacity. This is reflected in the results obtained at Trawscoed and Terrington and also from more recent results at Boxworth (Williams et al., 2002). Drainage systems with permeable fill represent the greatest risk; great care is needed to ensure careful management of slurry and dirty water applications to drained land if significant water pollution is to be avoided.

Figure 9. Impact of low rate dirty water irrigation on drainage water quality (ADAS, Terrington, 1991-92).

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(a) Following applications totalling 260 m³ ha-1 to a drained clay soil with 20 mm moisture deficit.

(b) Following single application of 250 m³ ha-1 to a drained clay soil with moisture deficit of 20 mm.

2.3.3. SUMMARY

It seems that the most significant sources of BOD transfers from agricultural systems to watercourses are likely to be associated with;

Catastrophic failure or mismanagement of slurry storage structures Uncontained runoff from hard standings and tracks used by livestock Manure applications to land, particularly slurry applications to drained land Dirty water applications to drained land Field heaps of solid manure High rates of application of slurry or dirty water Soil type, soil moisture status, and surface and sub-surface conditions are all

likely to influence rates of BOD transfer following land applications 3. MITIGATION MEASURES

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Mitigation of polluting effects of organic materials in water, as expressed by BOD can be achieved through a range of measures, many of which will be effective also against NH4-N and, to some extent, pathogens. It can also be seen that some of these measures apply at source, others to the mobilisation and transport phase and some will impact on all stages.

3.1. STORAGE

3.1.1. Management of Manure: slurry and solid manure systems

In order to gain maximum benefit from the nutrient content of manures and slurries in meeting the fertiliser requirement of crops, particularly nitrogen supply, manure applications should generally coincide with, or be just in advance of, the period of maximum crop growth. This will also be the period of maximum nutrient uptake and will generally be in late spring. This implies the need for adequate storage to contain manures generated during times when spreading is undesirable/impossible due to adverse ground conditions (usually excessive wetness), or prohibited, for example as a result of the timing restrictions imposed by the Nitrate Vulnerable Zones (NVZ) action programme for slurries and poultry manures (Anon, 1998b). The length of the required closed period for spreading varies across the UK, as a result of the distribution of severe winter weather conditions, soil type and underlying geology, restricted area designation (Anon, 1998b), or according to convenience and farmer preference.

Total store size will depend upon the required storage period, the number and type of livestock on the farm and the associated slurry or manure production (Smith and Frost, 2000; Smith et al., 2000b). Water addition, through the use of wash water or rainwater collected on yards draining towards the store, will often result in a doubling of the slurry volume produced on farm. The addition of litter used for bedding also needs to be considered, particularly in solid manure systems (Smith et al., 2001a). In view of the potential for serious pollution incidents as a result of catastrophic failure or mismanagement of slurry storage outlined earlier, it is clearly of crucial importance that storage requirements are adequately planned and that the stores themselves are well designed and managed.

3.1.2. Management of Dirty Water

It is important to promote the understanding that dirty water is a powerful potential pollutant (Table 16 and 17) that should not simply be allowed to drain away to the nearest ditch or ‘soakaway’, or disposed of in the cheapest and most efficient way. Dirty water has been defined (Pain & Menzi, 2003):

“Water derived from washing of equipment and floors in milking parlours, rainfall RUN-OFF from concrete areas or HARDSTANDINGS used by LIVESTOCK and contaminated with FAECES, URINE, waste animal feed etc. Sometimes referred to as BROWN WATER. Contains organic matter and so poses a risk of water pollution but has very low FERTILISER VALUE”.

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Dirty water does NOT include the liquid draining from a manure midden, or through the strainer walls of a slurry storage compound or a weeping wall store.

Dirty water needs to be collected and controlled and can accumulate very quickly where rainfall is incident over open yards and roofed areas. Therefore, measures should be taken to separate the drainage from clean areas (yards and roofed areas), from those contaminated with manures or slurry, or from open feed areas (Chadwick et al., 2003). Moreover, it is important that the difference between dirty water and slurry is understood, since the rules affecting the management and application of dirty water to land sometimes differ. In particular, in NVZs, the no-spreading time periods, for high available N content manures and slurries, do not apply to dirty water (Anon, 1998b).

3.1.3. Management Benefits

With regard to the storage of slurries and dirty water, it is clear that costs can never be justified on the basis of the improved N conservation and the resultant potential for increased savings on fertiliser costs. However, farmers continue to have a poor perception of the value of manures and annual statistics on fertiliser use show that the average fertiliser use on fields receiving a manure application prior to the crop is almost the same as on those fields not receiving a manure application. It is not unrealistic that improved storage facilities and better manure management will provide a considerable economic and environmental benefit. Additional benefits arising from a correctly planned, well-designed storage system include; savings in labour, efficiency (in terms of spreading operations), convenience, inproved working conditions for staff, and the protection of soil structure and fertility. Furthermore, an element of ‘treatment’ may be apparent in terms of BOD levels during the storage period, although this benefit is likely to be limited because most slurry storage systems are filled and emptied concurrently. An example is provided by current experiments on the treatment of dirty water stored before application (Figure 10). A reduction in BOD of 90%, from the initial level of c. 2500 to 270 mg l -1 was observed in the dirty water stored in tanks under ambient conditions (Chadwick, personal communication).

Storage can thus act both to reduce the BOD load at source and can reduce the risk of mobilisation and transport, by increasing flexability and improving the timeliness of slurry and dirty water applications. This will greatly reduce the risk of slurry/dirty water applications to land at or near field capacity, to frozen ground, or to drained clay soils with permeable fill over the drains and well developed cracks connecting the surface to the drainage system.

Figure 10. Decline in BOD in dirty water from a commercial dairy unit during storage at ambient conditions (April – June, 2004).

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3.2. GOOD MANAGEMENT PRACTICE

Research during the last 10-15 years has contributed to a greatly improved understanding of the fate of manure nutrients following land application and has highlighted reasons for the poor perception of the value of manures by farmers. The lack of good technical information and advice has also been progressively addressed in recent years, so that the improvements provided by research can be promoted, allowing better understanding amongst farmers and consultants. It is recognised that much can be achieved by the adoption of best practice, much of which is outlined in the Code of practice for water (Anon, 1998a). Where this guidance is lacking in necessary detail, supporting information is often available, for example on the integration of manures within crop production systems and associated nutrient management plans (Anon, 2000b; Chambers et al., 1999a; 2001).

More recent initiatives have included the Nutrient Demonstration Farms project (Defra contract NT2001), in which work with four commercial farms covering a range of livestock and cropping enterprises has aimed to encourage the adoption of improved management practices. This work has successfully demonstrated to farmers that manures and (reduced) inorganic fertiliser inputs can be readily integrated within commercial farming systems, without compromising crop yields and quality (Williams et al., 2000; Williams & Chambers, 2002).

3.2.1. Manure Application Practice

A major focus on improved manure application practice is necessary, with controlled application of manures and slurries of particular importance. Application rate is of particular significance to the risk of nutrient emissions to water (Smith et al., 2001b), to good utilisation of manure nutrients and to potential impacts on crop quality (Laws et al., 2002). The results of the latter studies indicate that silage quality is unlikely to be compromised by slurry applications made in early spring (February to April), where these are at agronomically sensible rates. To ensure that intended rates of nutrients in slurry/manure have been applied to the crop, the machinery used should

75

be carefully calibrated and operated to give accurate application and that it should be capable of being set up to apply an intended rate with an even spread pattern.

In simple on-farm tests of spreading rate and lateral spread variability involving a total of 41 spreaders or tankers (Smith & Baldwin, 1999), performance was found to be generally poor in terms of observed lateral spread pattern (data for solids spreaders shown in Figure 11). Moreover, in view of a lack of awareness amongst users about either the capacity of the spreader or the nutrient content of the manure, there was little chance on these farms of correctly allowing for the contribution from manure nutrients. However, it can also be seen that a simple adjustment of the spreading bout width enabled a much reduced coefficient of variance (CV) in lateral spread pattern (“optimised” CV in Figure 11), though it must be said that this may sometimes result in an unacceptable increase in application rate.

For mineral fertiliser spreaders, CV <15% is considered acceptable and for manure spreaders the aim should be to achieve CV <25% (Chambers et al., 1999b). Some simple agronomic experiments on silage grass plots, in which a range of CVs, based on the CEN assessment procedure (Anon, 1996), were established at a single slurry application rate of 40 m3 ha-1, to test the sensitivity of crop response to spreading imprecision. Even though there was a significant response to fertiliser N, imprecision in the lateral spread pattern of manures, up to moderately high CV had no significant effect on either grass DM yield or N offtake. This result is not surprising, given that the target slurry application rate supplied 39 and 46 kg ha-1 NH4

+-N, respectively, in the experiments in 1996 and 1997, i.e. well within the linear or steep part of the N response curve. Variation in N supply as a result of variable slurry application, is therefore unlikely to have resulted in any net observable difference between the different spread patterns. These findings are relevant to common spreading practice on UK farms, given that application rates and slurry N content were well within the range normally found.

3.2.2. Manure Application and Soil Management

As outlined in section 2.2., the results of Smith et al., (2001b) showed that increasing slurry application rate and in particular slurry solids loading increased both solids and nutrient losses via surface runoff. Limiting slurry application rate to within a limit of c. 2.5-3.0 t ha-1 solids, which approximates to the 50 m3 ha-1 limit for typical slurry suggested within the Water Code (Anon, 1998a) is an effective way of minimising risk of losses via this route. This application rate is also approaching the limit of what would be considered agronomically sensible practice. Other approaches likely to reduce the risk of runoff from susceptible soils would include:

Avoidance of a fine surface tilth through gentler secondary cultivations (leaving a rougher surface).

Slurry band application, injection or incorporation. Neither of these latter approaches has been tested experimentally however.

i. Secondary cultivations.There are more serious risks of pollution associated with the rapid movement of nutrients to drains through soil macro-pores following slurry or dirty water application to structured clay soils with under drainage (Figs 8, 9a & b). In the case of arable land, these risks can be greatly reduced by appropriate cultivations, either

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before, or soon after slurry application (Williams et al., 2002). It is clear that the disruption of the soil macro-pores and increased mixing of slurry within the soil matrix effectively inhibit the movement of slurry nutrients through the soil. However, this strategy cannot be applied to grassland. It is possible that a combination of limiting slurry application rate and carefully controlled application timing (in relation to soil moisture deficit and likely drainage events) may be effective in reducing this pollution risk pathway. However, further research is necessary to evaluate these and other options.

ii. Slurry band application, injection or incorporation. A range of reduced emission slurry application techniques have been developed (Huijsmans et al., 1997), which have been shown to be effective in reducing gaseous ammonia emissions, but which are known, also, to provide a number of other benefits. These include improved application precision and better control of application rate. As a result of the improved control and partly as a result of slurry placement, potential for negative impacts (including nutrient loss in run-off) is reduced (Prins and Snijders, 1987; Laws et al., 2002). The costs of different techniques can vary considerably and, based on machinery costs and farm data from 8 European countries, Huijsmans et al., (2004) were able to estimate relative costs, taking into account, also, the effect of farm size. Their estimates suggest an increase in the costs of slurry spreading over conventional surface broadcasting, of approximately 30-50%, 40-50% and 60-70%, for trailing hose, trailing shoe and shallow injection techniques, respectively.

Figure 11. Comparison of %CV of lateral spread pattern at average spread width with estimated optimum CV after adjustment of bout width, in on-farm observations (Smith & Baldwin, 1999).

3.2.3. Solids or Liquids Manure Management Systems Contrasting manure management systems with potential to reduce environmental emissions in one flux pathway can have a negative impact on other emissions. It is necessary to consider the possible impacts of contrasting strategies. For example, in the ten-year period up to 1998, there were typically five times more recorded point-

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source pollution incidents per year from slurry stores than from solid manure stores (Anon, 1999). However, stored solid manures can present a high risk to surface water pollution if field heaps are poorly sited (e.g. close to ditches/streams, directly over field drains) (Hatch et al., 2004).

It is well known that nitrate leaching losses are greatest in the period immediately following manure applications containing ‘high’ concentrations of readily available N (e.g. slurries and poultry manures) compared with ‘low’ available N manures (e.g. farmyard manure), Figure 12. When ‘high’ available N manures are applied during the autumn period, all of the manure readily available N can be lost from sandy soil profiles, where there is sufficient over winter drainage. However, N release following all animal manure applications will continue over the long-term, due to the continued release of organic N through mineralisation (Anon., 2000b).

Although less relevant to the risks associated with BOD, there are analogies with N losses via leaching and by-pass flow through structured clay soils. It seems likely that the risks of water pollution from BOD associated with solid manure systems are likely to be significantly less than from systems based predominantly on slurry or including large volumes of dirty water.

Figure 12. Nitrate leaching losses following manure applications to free draining sandy and shallow soils over chalk under arable tillage (Chambers et al., 2000).

3.2.4. Livestock grazing management

As described earlier (2.3.1) the risks of BOD losses to ground and surface water are likely to be small, as the BOD associated with the grazing deposition of dung and urine will largely be broken down as it passes through the soil profile. However, the risk of exceeding the capacity of the soil for BOD treatment in this way will be reduced by observing N-based stocking rate limits (Anon, 1998b). Transfers of nutrients and BOD via lateral flow, particularly on sloping or compacted grassland soils, can be minimised by excluding grazing of susceptible land under excessively wet conditions. The strategic use of stand-off pads for cattle and sheep accommodation overwinter provides a number of potential benefits, including sward protection (French and Hickey, 2001). Experience with outdoor pigs, which are perceived to pose a pollution threat (due to surface disturbance and instability),

78

suggests that the risks can be greatly reduced by controlling stocking rates and nose ringing, which together, allow grass cover to be maintained throughout the breeding cycle.

3.3. MANURE EXPORTS

An estimated 90 million tonnes of farm manures are annually applied to agricultural land in the UK (Williams et al., 2000), but with over 12 million ha of agricultural land, there is no overall problem with manure surpluses. However, in some areas of the country there are high densities of animals (e.g. pig production in Humberside) where there may be local surpluses of manure. Quite commonly these very large units have insufficient associated land area to safely utilise the manure produced. If excessive applications of manures are applied to land then the risk of surface runoff or through drain flow of liquids with high BOD is greatly increased. Over-application of manures can be avoided through the export of manure to other farms. Evidence of this can be seen in the recent Farm Practices Survey (Scott et al., 2002). This showed that almost no cattle FYM or slurry was exported from the producing farm, but that 29% of the FYM and 25% of slurry was exported from pig units, and 69% of the manure was exported from poultry units. This would seem a practical and sensible arrangement between cooperating farmers and can be a successful way of dealing with surplus manures and, hence, of reducing risks of environmental pollution.

Manure bank schemesFormal manure bank or farm waste brokerage schemes have been in operation in some countries, e.g. Netherlands, Denmark, Belgium and USA. An example of which was the Manure Agency East Netherlands (MBO), which was reviewed by Peirson (1997). The MBO was a foundation set up in 1994 and run by a board of 7 appointees (including 4 farmers). In 1993/94 they secured 4,000 contracts with farmers and 250 with contractors. Farmers and contractors joined on a voluntary basis but had to sign up for a minimum of 10 years. The original cost of setting up the MBO was 14 million guilders (c. £4m).

The MBO administered the disposal of manures to land and looked for new disposal options. In addition to a joining fee, farmers paid a development or disposal fee. Pricing for disposal was based on the quality premium scheme previously operated by the government manure banks. The MBO aimed to remove surplus manure and slurry and control the local market in order to keep prices stable and demand from arable farmers high. In total the MBO was responsible for 2.2 million tonnes of manure and had storage for 360,000 tonnes in arable areas. Within the catchment area of the scheme the MBO controlled 35% of the total manure/slurry produced. In the middle of the area, which was not adjacent to arable land, 80-90% of the manure/slurry produced was controlled by the MBO. On the boundaries, membership was only 15 - 20%. This was mainly due to the fact that the disposal fee was based on a flat rate, not taking account of distance from disposal areas. It was therefore more cost effective for farmers on the outskirts to make their own arrangements. Farms located more than 12-15 miles from ‘disposal’ areas were considered likely to have problems with disposal. In areas where farmers needed to export manure, cooperating farmers (importing as well as exporting) inevitably found it cheaper and more convenient to make a private arrangement, with money exchanged only between the donor and

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recipient farms and no fee to the manure bureau. Such schemes have ultimately failed in the Netherlands, largely on a cost basis.

In the UK a number of factors seem likely to inhibit the development of a similar scheme, including, importantly, the relatively low cost of purchased inorganic fertilisers, the costs of manure applications and, amongst farmers, the perceived difficulties and imprecision of quantifying the fertiliser value of organic manures.

3.4. TREATMENT SYSTEMS

A wide range of treatment technologies is available, which can contribute to mitigation of potential pollution, some of the options being particularly effective for reduction of BOD (Table 19).

Table 19. Summary of manure treatment options available on farms (after Burton, 1997).

Option Benefits Drawbacks Comments

No Treatment(Direct spreading)

Routine Task.Least cost.Avoids need for intensive spreading campaign

Poor utilisation of nutrient.Risk of land damage.Pollution risk.

Common option

Storage

Better nutrient utilisation by targeted spreading.Flexibility.Enables treatment options.

Crusting and sedimentation problems.Capital costs.Increased odour potential.

Integral component of treatment processes.

MechanicalSeparation

Reduces liquid volume.Reduces crusting and sedimentation in storage.Improved homogeneity.Easier pumping.Composting of fibre

Cost of pit, pump, gantry and separator.Operational costs.Reliability.

Important process for store management and crop utilisation.Used with biological treatment processes.

Aerobictreatment

Reduces odour and BOD.Provides mixing.Generates heat, which could be utilised.

Capital and operational costs.Separation necessary for most slurries.Selection of optimum system difficult.

Best option where environmental pollution is a risk, particularly odour.

Anaerobicdigestion

Reduces odour and BOD.Biogas production.Easier handling of liquid.

High costs.Management critical.Continuous gas’ production requires use if benefits are not to be lost.

Continuous process.Attractive option where energy supply an issue.

Solidscomposting

Reduces odours.Saleable product.Can include other by-products.

Volatile emissions Capital and operational costs.Marketing skills required.

Very important to establish markets before following this route.

The chosen process should be an integral part of the manure management system at the farm and should be capable of meeting the clearly defined objectives of the treatment system, which need to be identified before considering any treatment option. The best solutions will be within the technical and financial resources available for the unit concerned. Technologies are available that can turn slurry into potable water, capable of discharge into a watercourse, but only at enormous cost.

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What are more certainly required are practical options that meet the needs of the problem without greatly exceeding these.

The benefits and drawbacks of the broad range of treatment options are summarised in Table 19. Storage can be considered a treatment option and will usually form an integral part of a treatment package, although would scarcely be considered as “treatment” by farmers in their rationale for building a manure store.

3.4.1. Mechanical Separation

This relatively simple process can offer advantages both in terms of the improved handling and management characteristics of the two products. There are two basic methods of solids liquid separation. One uses the difference in density between the solid particulate matter and the liquid (settlement or centrifuging) and the second uses the shape and size of the particles to cause separation (screening and filtration).

There can be several reasons and advantages for undertaking separation: improved infiltration of the liquid into the soil, for reduced odour and NH3

emissions; reduction in herbage contamination with slurry solids and, hence, reduced risk

of negative impact on silage quality or pathogen transfer to grazing animals; easier handling of liquids, facilitating improved accuracy of spreading; reduction of nutrient loading via slurry application (may be significant in cases

of nutrient surplus); improved homogeneity of the liquid phase (with reduced sediment and

generally no crusting); reduced storage volume for slurries; reduced energy requirement for mixing and pumping and reduced risk of

blockages; reduced risk of blockages during subsequent operations; useful pre-treatment for biological treatment.

Some disadvantages also have to be considered: storage, handling and spreading of two separate materials; necessary investment in machinery; farm labour and technical input requirement.

After solids are removed, they can be applied to land, dried, composted, or used elsewhere, e.g., in Japan have been used for bedding after composting.

i. SettlementSolids with a density greater than that of water can be settled out by holding the effluent in a tank or allowing passage at low velocity. Fast-moving liquids pick up and transport solids; when velocity slows, the solids settle. Settlement is most effective in dilute wastewaters, e.g. flushing water, yard runoff (Miner and Smith, 1975). Settlement in these dilute effluents occurs fairly rapidly with most occurring within the first 10-20 minutes of retention. This can be seen in the results for potato and carrot washing (Table 20), with settlement particularly effective for solids reduction, though with relatively little effect on BOD (Jones, 1945).

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Table 20. Impact of settlement of effluent from washing and processing of vegetable crops (Jones, 1945).

Period of Settlement

(min)

Potatoes Carrots

Pre-Washing Peeling Pre-Washing Peeling

SS BOD SS BOD SS BOD SS BODmg l-1 mg l-1

01020304060120

2500330

-230

-180100

5741-

41-

3649

990033002100

-1400940810

275021502130

-208019901950

3330520370

-220250200

473128-

292928

3100450440

-430420390

265022502220

-222021502050

% removed in 2 h

96 14 92 29 95 40 87 23

SS = suspended soils

ii. Solids-liquid separationQuicker separation can be achieved using a mechanical screening process. A wide range of equipment is available, usually involving sieves or screens in various configurations. These include run-down screen, vibrating screen, belt press, drum press, press screw/auger separator, sieve centrifuge, decanter centrifuge. Costs vary widely reflecting sophistication and performance. At the low end are basic screening packages, e.g. sieves with pump and mixing equipment, costing c. £15,000 and, at the high end are the centrifuges at >c. £60,000. Despite the range of equipment available, relatively few have been taken up by farmers in the UK; recent estimates suggest that 8% of pig farms and 7% of dairy farmers use some form of mechanical separation to assist in slurry management (Smith et al., 2000a; Smith et al., 2001a).

Performance characteristics can vary substantially and depend on several factors, including:

separator type sieve mesh size (also centrifugal force) slurry type additives (e.g. water; flocculent) solids content of the slurry

Performance is usually assessed in terms of slurry flow rate and relative output of solids and liquid, with separation % of solids and of the major nutrients, N, P and K. Data from a range of the more common separators are shown in Table 21.

Best nutrient separation results appear to be possible with the decanter centrifuge. The range in performance is the result of variable slurry solids content as input to the centrifuge and machine setting. To some extent this allows some adjustment in performance of the technique, depending upon objectives, e.g. maximum nutrient removal in the solids, or production of high DM solids which will compost easily and quickly. Separation can be expected to remove a significant proportion of the organic load from slurry, in terms of COD (with removal of the coarser solids) but might be expected to impact rather less on BOD. Few data are available on these aspects of performance. However, Shutt et al., (1975) reported the removal of 35% solids, 62% BOD and 69% COD from pig slurry, by a simple run-down screen, even with only

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3% of the volume removed. Performance varied with the screen slot width (0.1 cm better than 0.15 cm) and slurry inflow rate. Operation of a vibrating screen with the same pig waste waters indicated rather lower removal efficiencies of BOD and COD (only c. 4% and 10-15%, respectively) than the run-down screen, but with optimum application rate varying according to screen opening size.

Whilst it seems likely that some reduction in slurry/effluent BOD can be expected from a mechanical separation process, performance will vary with slurry type, screen size and flow rate set-up. It should be noted that assessment of the removal efficiency can only be reliably made on the optimised set-up.

Table 21. Separation efficiency and technical data for common separators (Burton & Turner, 2003).

Belt press Sieve drum Screw press Sieve centrifuge

Decantercentrifuge

Flow rate (m3/h) 3.3 8-20 4-18 1.9-5.5 5-15Separation efficiency %

Dry matter 56 20-62 20-65 13-52 54-68N 32 10-25 5-28 6-30 20-40P 29 10-26 7-33 6-24 52-78K 27 17 5-18 6-36 5-20

Volume reduction % 29 10-25 5-25 7-26 13-29Specific energy (kWh/m3) 0.7 1 0.5-2.0 2.2-6.7 2.0-5.3 3.4.2. Anaerobic and Aerobic Digestion

The processes of anaerobic and aerobic digestion have been described within the Pathogens section of this document. As well as having significant effects on pathogen levels within slurries they have also been shown to reduce BOD.

i. Anaerobic digestion. Anaerobic digestion can result in a considerable reduction of odour (Pain et al., 1984) and significant reduction in the pollution potential of slurries as assessed by BOD and COD (Table 22) (Hobson and Robertson, 1977).

Table 22. Reduction of BOD5, COD, total solids (TS) and volatile fatty acids (VFAs) in slurries as a result of mesophilic anaerobic digestion.

Parameter Pig Cattle PoultryBOD5 75 55 80COD 50 35 50TS 40 30 60

VFA 73 70 80 The quantity of total N in the waste was not changed, except for a small proportion of NH3 being transferred to the biogas. It appeared also that some organic N was reduced to NH4

+-N, thus increasing the NH4+-N to a relatively high value (4-6 g l-1

NH4+-N), particularly in digested poultry wastes (Dohanyos et al., 2000). Phosphorus

was also found to be partially released into the liquid phase, by digestion. Overall, therefore, anaerobic digestion decreases the C:N ratio and increases the concentration of immediately accessible plant nutrients.

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Whilst anaerobic digestion is ‘proven technology’ which has been available for over twenty years, uptake of the process has been minimal and restricted to enthusiastic farmers or those sites with specific factors, such as the need for odour control or a direct use for the biogas produced, which favour the process.

ii. Aerobic treatmentAerobic treatment of slurry is normally carried out only for odour control purposes and this is achieved via the microbial breakdown of the many compounds (organic and inorganic) that contribute to manure odour. This results in the stabilisation of organic compounds and, hence, the reduction of COD and BOD.

Livestock slurries can be aerated for various a) times, b) dissolved oxygen target concentrations and c) range of temperatures. Also the aeration can be run as a batch or a continuous process. All of these parameters affect the characteristics of the treated slurries. The most efficient is a continuous culture process, and therefore most research has been devoted to it. As a continuous process, aeration will generate heat and can be performed at mesophilic temperatures (25 to 45ºC) and thermophilic temperatures (>50ºC) (Evans et al., 1983). Laboratory experiments provided data allowing the generation of mathematical equations describing the characteristics of treated pig slurry at those temperatures. Thus:

BOD5 = 1.568/R + 0.152BODf

whereR is the mean treatment time (days)BODf is BOD of the fresh slurry (g l-1)

Thus, pig slurry of 10% DM and BOD5, typically of 35 g l-1, after 5 days of mesophilic aeration, would have a BOD5 of 5.6 g l-1, or c. 16% of the original BOD. Such continuous aerobic treatment is known, also, to result in strongly offensive pig slurries becoming inoffensive (by odour panel assessments) within 2-3 days at mesophilic treatment temperatures (Evans et al., 1983).

A number of agitation systems have been installed which use relatively small amounts of air to mix slurry at intervals. However, these are unlikely to achieve significant oxygenation and therefore there is no guarantee that existing systems would always achieve conditions conducive to a significant reduction of BOD or odour or pathogen control.

3.4.3. Solids Composting

The three main types of composting system have been described within the Pathogen section of this review and in terms of BOD, the composting of solid manures is not strictly of relevance. Nevertheless, composting of solids may involve the use of slurries as a source of nitrogen and does sometimes produce leachates with high BOD. However, composted manures will pose a much reduced risk of pollutant run-off during storage or following land application, as a result of the metabolism and stabilisation of organic compounds that otherwise may contribute to such pollution.

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During treatment the moisture content decreases from about 70% to less than 30%, and the organic content from about 75% to 50%. By oxidising the biodegradable carbonaceous compounds to CO2 the compost is biologically stabilised, i.e. when stored without access to air and re-wetted, it does not generate any odorous compounds and its biological activity is minimal. This also means that the potential BOD emissions, e.g. in leachate from stored material is greatly reduced.

3.4.4. Use of Treatment Additives

Numerous additives have been investigated over the last three decades, including various commercial products, which are intended to help prevent or to alleviate the main problems associated with the manures arising from intensive livestock production.

The most common additives include; bacterial/enzymic preparations; plant extracts; chemicals including acids, oxidising agents, disinfectants, urease inhibitors, masking agents or products with physical properties such as adsorbents. However, the effectiveness of these additives, particularly the commercially available products, remains to be conclusively proven (Pain et al., 1987; Ritter, 1989).

The role of manure additives in minimising the impact of accidental discharges of slurry or dirty water is yet to be investigated. There would however, appear to be potential for additives to reduce both the BOD and nutrient loading of livestock slurry. Reductions in BOD loading would appear feasible through enhancing the degradation rate of organic matter. Williams (1983) found that the volatile fatty acid (VFA) fraction of slurry accounted for up to 70% of its BOD5; additives that degrade the VFA fraction may well, therefore, lower BOD. Commercial digestive additives claim to lower BOD but no experimental evidence of this effect can be found. The mechanism by which the activity of these products might be achieved is not specified, so it can only be assumed that they may enhance the degradation of VFA fraction in the slurries. McCrory & Hobbs (2001) postulated three alternative mechanisms for these effects. Firstly, the additives may add micro-organisms that degrade the organic compounds found in livestock slurries more readily than the natural population. Secondly, they may add enzymes which catalyse the degradation of more recalcitrant organic compounds in livestock slurries, rendering them easier to degrade by the natural or added micro-organisms. Thirdly, they may add an additional carbon source. It would be assumed that the chosen carbon source would have a high C: N ratio and be readily available to micro-organisms.

A series of laboratory experiments were undertaken on pig and cattle slurries, with a range of commercial additives and some single component agents (Defra contract WA0522). Limited effectiveness of most commercial slurry and solid manure additives was clearly demonstrated. Most products claimed microbial activity and often had low counts of viable organisms coupled with poor growth characteristics in slurry, largely due to limited availability of appropriate substrate. It was concluded that slurry additives are in their infancy and while there are some successes, rapid development and research would be the key to benefit the livestock producer and the environment. Investment in the livestock waste additives industry has been limited to date.

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Compared to other options, the use of manure additives can appear attractive, as generally there is no need for capital expenditure. The initial cost, quantity required, frequency of application, and hazard potential of an additive are the most important considerations. As slurry will be continually accumulating in a store, it is likely that all additives will need to be added at some frequency. Microbiological additives though, may benefit from continuous accumulation of slurry. To effectively immobilise N, micro-organisms need to maintain a high population. Frequent application of fresh slurry may provide the readily available organic matter needed for maintenance. The need for frequent addition of additives to slurry could be eliminated if additives are placed in the feed of livestock. Although several examples were presented in the literature, little work has been done in this area, and further research is required. There appears also to be the need to assess the influence of some other additives with regard to possible impact on BOD capacity. It would seem likely that additives that inhibit microbial degradation may well also inhibit any ‘natural’ reduction in effluent BOD, as a ‘side-effect’ of their main activity.

In the light of the very limited supporting research data indicating benefit, additives are unlikely to find wide/popular applicability. It still seems likely that much of the commercial uptake of these products is on a “first encounter” basis, with little repeat business.

3.4.5. Manure Processing

In addition to the movement of manures from areas of surplus to areas of low manure availability and attempts to limit livestock numbers, large scale manure processing might be a useful strategy for abatement of environmental problems associated with livestock manure surpluses. It was, for some time, regarded as the best solution for the surplus of pig manure in the Netherlands (Ten Have, 1993). Processing of liquid manures, whether on a single farm, or regional scale, comprises a series of basic treatment steps, logically arranged. The detail of the steps and optimal arrangement depend on the slurry composition, on local circumstances and on the required products and product quality. The basic steps considered feasible at the present time comprise 8 groups of options (may be combined or used as alternatives), which are summarised in Table 23 overleaf.

The Promest, central processing system developed in the Netherlands is shown in outline in Figure 13 (Rulkens and Ten Have, 1994). Raw pig slurry is digested anaerobically and the digested slurry separated into solid and liquid fractions. Aeration of the liquid fraction, under pH-controlled conditions promotes nitrification. Effluent from the aeration process is evaporated, generating clean water and a concentrate. The concentrate, dried with the slurry solids and sludge from the aeration tank, is used to generate the organic fertiliser product (lower RH corner of Fig 13). Nitrification during the aeration stage enriches the N fertiliser value of the product and other minerals can be added. In fact, the introduction of central pig slurry processing failed, mainly because of high costs and the lack of sound organisational and financial arrangements, as well as the lack of a well-organised network for the distribution and marketing of the products (Rulkens et al., 1998). The anticipated effect of farm-scale measures on (the reduction of) manure volume and composition also had a negative effect on the introduction of central processing. The cost of processing, at c. €25-30 .t-1 was too high and could not compete with local

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arrangements for disposal on neighbouring farms (Hans Willers, personal communication). The Promest system is no longer in operation.

Figure 13. Simplified process diagram of the Promest system (after Rulkens and Ten Have, 1994).

Table 23. Basic treatment steps in animal slurry processing (after Rulkens et al., 1998).

1. Separation of colloidal and suspended particles:

- Filtration (in combination with coagulation/flocculation)

- Centrifuging (in combination with coagulation/flocculation)

- Sedimentation- Separate collection of urine and faeces

in pig houses

5. Removal or concentration of minerals:

- Evaporation (production of concentrate and condensate)

- Reverse osmosis (production of concentrate and permeate)

- Freeze concentration (selective removal of water)

- Catalytic incineration 2. Removal, concentration or conversion

of organic compounds (soluble and insoluble):

- Mechanical separation of colloidal and suspended particles (filtration or centrifuging resulting in a sludge (cake) and a liquid phase)

- Anaerobic treatment

6. Treatment of the manure cake and concentrate from an evaporator or reverse osmosis process:

- Drying (production of a dry cake and a gas phase containing water and volatile compounds)

- Incineration of the wet or dry cake aimed at energy production

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ScreeningMANURE

Anaerobic Treatment Biogas

Centrifuging

Drying

Aerobic Treatment

(Nitrification)

Lime

Dust Removal

Condensation

Sludge Cake

Liquid

Vapour

Evaporation Concentrate

Condensation

Vapour

Clean Water

PelletisationMinerals

Product

DryProduct

Condensate

Liquid

- Aerobic treatment- Wet Oxidation (subcritical or

supercritical)- Hydrothermolysis (supercritical)

- Pyrolysis- Gasification

3. Removal, concentration or conversion or immobilisation of N compounds (including NH3):

- Stripping and absorption in mineral acid- Nitrification (immobilisation)- Denitrification- Acidification with mineral acids

(immobilisation)- Ion exchange- Precipitation as ammonium magnesium

phosphate- Membrane separation

7. Treatment of exhaust gases from a drying process:

- Dust removal - Condensation- Bioscrubbing- Biofiltration- Incineration

4. Removal of phosphorus:- Filtration/separation of colloidal and

suspended particles- Precipitation (soluble compounds)- Biological phosphate removal

8. (Micro)biological conversion of nitrogen:

- Production of fungi- Production of algae- Production of amino acids- Production of bacteria- Production of yeast- Production of duckweed

3.4.6. Soil Treatment Processing

The movement of effluent through soil results in a high degree of purification as long as the ‘treatment’ capacity of the soil is not exceeded. The purification capacity of the soil relates to a combination of physical filtration, chemical reactions (e.g. with respect to phosphate and toxic metals) and biological/microbiological activity in which degradation of substrate compounds and utilisation of the nutrients occurs. One treatment system based on these features of natural soil fertility is the ‘Solepur’ process developed in Brittany, France (Martinez, 1997). The pilot-scale ‘Solepur’ system was used to treat pig slurry and is based on three main components:

managed field (grass or arable) which is drained and hydrologically isolated and to which the slurry is applied. Drainage is collected and passed to:

a reactor for promoting denitrification, via intermediate storage; a non-managed field to receive the denitrified drainage water.

Large volumes of slurry (c. 1000 m3 ha-1 yr-1) were applied to the managed field, with an average annual nutrient load of 5000 kg ha-1 N, 1600 kg ha-1 P and 1700 kg ha-1 K. The process removed 99.9% COD, 99.9% P and c. 90% N from the slurry. The final leachate contained a very low concentration of organic matter, but high nitrate levels, resulting from the oxidation of slurry N in the soil. Whilst the ‘Solepur’ system retained its capacity for the removal of nutrients and organic matter, over 5 years it appeared that gaseous emissions were in some cases encouraged (Chadwick et al., 1998). Ammonia losses were typical of those from surface applications of slurry at agronomic rates; CH4 emissions varied considerably according to soil conditions and slurry application rate. Emissions of N2O were very high following slurry applied in October – at 23% of N applied, in contrast to only 0.17% loss following slurry application to dry soil in June.

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Other options, of more immediate interest to the UK, include the scope for soil-based treatment of dirty water (Defra contract WA0518). In this work, percolation systems were constructed on a permeable soil, with intact and disturbed soil and the overland flow system was constructed using an impermeable soil.

The percolation systems worked on a continuous basis and were capable of reducing BOD, MRP and NH4

+-N by >90%, at dirty water application rates of 2mm or 8mm. Disturbed soils were more effective (P<0.05) at reducing BOD than intact soil blocks, presumably because the soil disturbance reduced the possibility of by-pass flow by disruption of any deep macro-pores or fissures. Leachates from both intact and disturbed soils sometimes contained high NO3

--N concentrations (max. 90 mg l-1). Overland flow systems, working on a batch-flow basis, were capable of significantly removing BOD (>85% removal over 10 days), NH4

+-N (>90% removal over 10 days) and MRP (>90% removal over 10 days).

Delivery rate would be expected to affect treatment efficiency via the impact on residence time in the system. However, with dirty water applied at 25 mm, over 1, 6 or 12 hours, there appeared to be no significant impact on BOD, MRP, NH4

+-N and NO3

--N concentrations. Similarly, across three weekly dirty water application rates, treatment efficiencies showed only a marginal increase (P>0.05) at the lowest rate of 60 mm per week (Table 24). Nitrate leaching, however, was much reduced at the higher loading rates, presumably as a result of reduced nitrification under the anoxic conditions at the highest application rates.

Table 24. Reduction (%) in the total loss of BOD, MRP, NH4+-N and NO3

--N in leachates after varying rates of dirty water application to lysimeters at IGER, North Wyke. (WA0518).

Application rate mm/week BOD MRP NH4

+-N NO3--N

Disturbed 60 80 73 62 -14395125 75 68 59 -322250 69 61 49 -5

Intact 60 80 69 63 -11757125 75 69 55 -3759250 75 68 57 -72

3.4.7. Constructed Wetlands

Constructed wetlands have gained popularity over the last 30 years, particularly for the treatment of municipal wastewater as the complex microbial communities present will deplete both particulate and dissolved organic matter reducing COD and BOD. However, more recently there has been interest in the application of these systems for the treatment of farm wastewaters. The conventional (low rate irrigation) systems for land application are known to result in sufficient numbers of pollution incidents (see

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section 2.3.2, on sources of BOD), even when apparently well designed. In these circumstances, treatment of dirty water and subsequent discharge to a watercourse appears to offer an attractive alternative approach to land spreading.

From a review of published literature (Defra contract WA0658), it was estimated that there might be over one thousand constructed or natural wetlands, world-wide, receiving and treating a variety of wastewaters. Mostly, these receive only dilute effluents, for example, secondary treatment of domestic sewage with a typical BOD5

of 100-250 mg l-1. The BOD5 of dirty water associated with livestock production is considerably higher (Table 16 and 17) on UK farms. Otte (2004) suggests that about 75 constructed wetlands have been set up in Ireland, with most of these, less than 10 years old. Also, many were built on an ad-hoc basis, without proper design considerations. In the US, the majority of wetlands for treatment of animal waste are said to have been installed since 1989 (Stone et al., 2002).

Design parameters are evolving with time and vary somewhat according to the location and situation, the influent requiring treatment and the agreed objectives. The USDA Natural Resources Conservation Service (USDA, 1991) design guidelines were based on BOD5 loading (presumptive method). The guidelines stated the maximum levels of BOD5 discharging from the wetland, with a recommended residence time of at least 12 days. Since then more physically based models have been used. Two (Reed et al., 1995; Kadlec & Knight, 1996) are based on a first-order kinetics, area-based uptake model. Reed et al., (1995) incorporated flow rate, wetland depth, wetland porosity, a temperature-based rate constant and inflow and outflow concentrations. The Kadlec & Knight (1996) model incorporates hydraulic loading rate, inflow and outflow concentrations, a temperature-based rate constant and a background concentration parameter. Burton and Turner (2003) reported results from a subsurface flow wetland used to treat the washwater from a dairy farm in the Emilia-Romagna region in Italy. Mean analyses of the inflow and outflow water from the system, suggested very effective removal of solids and organic matter (reductions of 91.4% total suspended solids; 92.4% COD; 93.6% BOD) but rather lower reductions of total N, NH4

+-N and P (46.7%, nil and 64.5%, respectively).

It is clear that constructed wetlands have a potentially important role in the options available for mitigation of both point and diffuse pollution. It is likely that they would be seen as a component of an overall strategy, to be used in conjunction with other measures. It is important to remember that wetlands provide an important buffering capacity, even if subjected to shock loadings. Such loadings may exceed their capacity for treatment, resulting in system ‘failure’, but they would still provide important protection to water courses, reducing the damage which may result from an otherwise catastrophic pollution incident.


Mitigation of polluting effects of organic materials in water, as expressed by BOD, can be achieved through a range of measures, many of which may be effective also against other pollutants. It can also be seen that some of these measures apply at source, others to the mobilisation and transport phase and some will impact on all stages.

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There will rarely, if at all, be any conflict arising from the BOD mitigation options identified within this review. Perhaps the exception would be aeration which, while effective in reducing BOD5 (by typically 80-90% under mesophilic conditions) and, also certain pathogens, is likely to exacerbate ammonia losses, particularly in the case of intermittent, batch treatment processing. A similar conflict is likely to arise as a result of aerobic composting of solid manures, which is very effective for pathogen removal and the removal of VFAs, odours and moisture content and, hence, potential for BOD release via leachate. However, during the composting and storage phase, NH3 and N2O emissions are likely to increase compared with conventional storage (Defra contract WA0519).

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6. A CONCEPTUAL MODEL FOR DIFFUSE POLLUTANT BEHAVIOR: A MEASURE CENTRIC APPROACH

1. THE MEASURE CENTRIC APPROACH

Traditionally, mitigation measures for diffuse pollution have focused on pollutants in isolation, often without considering the interacting effects that the measures may have on other pollutants (pollution swapping). To address the broader problem of diffuse pollution in a more sustainable way, there is a need to move away from consideration of individual pollutants and towards a ‘measure centric’ rather than ‘pollutant centric’ approach. Conceptually, there are three ways that pollutant transfer to surface waters can be reduced. These involve controls applied at each of the SOURCE, MOBILISATION and DELIVERY stages of the pollutant flux pathways. This approach is taken from that already published elsewhere (Heathwaite et al., 2003; Haygarth et al., in press).

1.1 SOURCE CONTROL

Source controls are targeted at reducing the quantity of potential pollutant vulnerable to mobilisation and delivery to surface waters. The sources of pollutant can be both natural and anthropogenic but the magnitude of a pollutant source should be measured at the source area, i.e. the site of production or mobilisation. From the perspective of a farm system, pollutant sources can be separated into external, internal and recycled sources.

ExternalExternal sources are the primary sources of potential pollutant that are notionally imported across the farm gate. This includes mineral fertilisers, some feedstuffs, and natural atmospheric deposition.

RecycledRecycled sources are the secondary sources that are generated internally to a farm as part of the production system. This includes farm manures, slurries and dirty water, and also excreta from grazing animals.

InternalInternal is used to describe the tertiary sources that are generated internal to the soil profile. This includes the decomposition of organic materials, weathering of bedrock and soil formation.

1.2 MOBILISATION CONTROL

The term mobilisation covers the process by which pollutants become mobile. Mobilisation controls can be targeted at reducing the efficiencies of the different processes by which pollutants are entrained into hydrological pathways from agricultural land. These processes can be separated into solubilisation, detachment and contingent transfers.

92

SolubilisationSolubilisation is used to describe the chemical or biological release of potential pollutants from soil sources into soil water. The processes includes de-sorption of potential pollutants from soil particles, and the release of organic molecules by the death and decay of plants and microorganisms. Solubilisation is a low-energy environment process. Solutes are defined as less than 0.45m diameter (Haygarth and Jarvis, 1999) and this definition spans bio-colloids (bacteria and viruses) and humic macro-molecules.

DetachmentDetachment is used to describe the removal of soil fragments, and any accompanying adsorped potential pollutants, away from the bulk mass by the hydraulic action of flowing water or by raindrop impact. Detachment is predominatly a high-energy environment process and soil fragments are defined as more than 0.45m diameter (Haygarth and Jarvis, 1999). Detachment is often referred to as soil erosion, however, this infers a subsequent transport processes (saltation or movement in suspension) that carries the soil fragments elsewhere. Therefore, soil erosion cannot strictly be a mobilisation process.

ContingentContingent is used to describe pollutant mobilisation that encompasses both ‘accidental’ and ‘incidental’ mobilisation (Preedy et al., 2001). Accidental is used to describe mobilisation that occurs when there is a physical or management system failure, such as occurs when a store leaks pollutant directly to a watercourse; or when guidelines for application of a potential pollutant are incorrectly followed and results, for example, in spray drift over a watercourse. Accidental mobilisation is entirely preventable. Incidental is used to describe mobilisation that occurs when rain and runoff interact directly with fresh applications of pollutant to the soil surface, including those from fertiliser, manure and excreta. In reality, incidental losses will include solubilisation and detachment, however Haygarth and Jarvis (1999) have argued that, conceptually, incidental loss should be kept separate because of the unique circumstances leading up to its occurrence and control. Incidental mobilisation cannot be totally prevented, because there are agronomic restrictions on changes in the timing of potential pollutant inputs and because we cannot achieve perfect prediction of future weather conditions to support tactical changes.

1.3 DELIVERY CONTROL

Moving water provides the energy and is the carrier for the vast majority of pollutant movement off agricultural land. Transport controls can be targeted at changing the relative importance or reducing the efficiencies of the different pathways water can take between the land and receiving watercourse. These can be separated into surface, preferential and through-flow pathways.

Surface FlowSurface flow is used to describe the process of water movement exclusively over the soil surface, due to either infiltration or saturation excess, whereby land drainage makes a rapid and direct connection to a surface watercourse. Physical and chemical interaction with the bulk soil is restricted to a shallow layer, not more than a few

93

centimetres deep. Surface runoff is a high-energy process, able to maintain a large suspended particle load, and with a discharge timescale measured in minutes to hours.

Preferential FlowPreferential flow is used to describe the rapid vertical and lateral movements of water within the soil profile, and occurs via cracks, macro-pores and artificial drainage channels. It is spatially heterogeneous and characterised by incomplete equilibration with the bulk of the soil water. Preferential flow is an intermediate energy process, with a discharge timescale measured in minutes to hours.

Through FlowThrough flow is used to describe the slow, vertical and lateral movements of water through the soil profile. It includes lateral interflow beneath the soil surface, and vertical drainage to the groundwater table. It is spatially homogeneous and is characterised by interaction with the bulk of the soil mass and with water movement through the finer soil pores. This results in a high potential for chemical exchange and physical filtration of the pollutant load. Through flow is a low-energy process, with a discharge time scale measured in days or weeks.

Using these definitions it is possible to group pollutants in terms of their generic characteristics described within the literature (Figure 14)

2. MITIGATION MATRIX

A mitigation matrix has been developed (Appendix I) to help provide Defra and the project team with a provisional ‘guide’ to key pollutant behaviour properties and mitigation options. This lists the main mitigation measures currently available to agriculture and relevant to the pollutants of interest in this study. The matrix categorises whether the measures apply to source, mobilisation or delivery mitigation of the pollutants, as discussed above. Each measure is rated according to how effective it is considered to be at mitigating each of the diffuse pollutants. Effective mitigation of a pollutant is allocated a positive score; however, if that measure is considered to exacerbate the impact of a pollutant, then it is scored negatively.

As part of this process, a list of typical farm systems has been created (Appendix II). This has been based upon livestock/cropping system, soil type and drainage system, manure system, and grazing regime. While this does not attempt to include every possible farm system in England and Wales, it is considered to contain the predominant systems and provide a systematic means of defining model systems and identifying a way to provide comparisons. In the mitigation matrix (Appendix I), the farm systems to which each measure applies has been included.

The cost effectiveness of each measure within the matrix has also been scored on a cost (per hectare) basis for its implementation; ease of implementation is also rated. Other measure characteristics that have also been rated include public and farmer acceptability, potential for pollution swapping, potential for conflicts with other measures, and the uncertainty of effectiveness of the measure. Where measures overlap with those within the draft list for ‘Environmental Stewardship Options’, it has been indicated using the proposed option codes (Martin, 2004, personall comms).

94

The advantage of the matrix is that it is intended to be ‘measure centric’. This means that it can be used as a provisional guide to assess potential cross pollutant benefits of individual measures. For example, by mitigating against the detachment process during mobilisation, we will be mitigating against not one but several pollutants, including NH4

+, faecal pathogens and BOD (Figure 14) to varying degrees but not against others, in this case, NO2

-. Of course, implicit in this approach is the fact that different pollutants, having different properties, will have varying responses to the measures – these needs should be taken into consideration when consulting the matrix.

Some key outputs from the matrix are: Removing/reducing inorganic fertiliser through better awareness of the

nutrient properties of the organic manures being applied is an extremely effective way of reducing NH4

+ sources. This measure is very cost effective and potentially may have cost benefits to farmers in reducing inorganic N imports to the farm.

Treatment of manures and slurries at source provides excellent opportunities to reduce pathogen numbers. The composting of solid manure (>55C for 3 days), batch storage of solid manure (90 days) and the anaerobic digestion and pasteurisation of slurry are considered to be the most effective measures.

The addition of lime and the aeration of slurry will also reduce pathogen numbers, albeit less effectively. These treatments also reduce the source strength of NH4

+, NO2- and BOD. This would result in increased atmospheric

N emissions however. Changing from a slurry based system to a solid manure system offers an

effective way of reducing the mobilisation of most diffuse pollutants although the costs of the change are high.

Application timing of organic and inorganic fertiliser would appear to be the cheapest and most effective way of reducing diffuse pollutant mobilisation.

Injection of slurry would increase losses of NH4+, NO2

- and pathogens. This is due to the increase in source size resulting from reduced NH3 atmospheric emissions and also potentially increased hydrological connectivity to drains. Placing pathogens directly into the soil may also prolong their survival when compared to survival times on the soil surface.

The establishment of artifical wetlands provides the best means of reducing delivery of all pollutants to watercourses. This measure would be very acceptable to the public, however, costs and ease of implementation may be prohibitive.

3. CONCLUSIONSIn the past, apparent mitigation of a pollutant has merely passed the problem to somewhere else in the farm system. It is now clear that a holistic approach is needed to deal with diffuse pollutants and that they should be viewed together as a single problem and not a series of separate pollutants to be dealt with individually. Furthermore an awareness of pollutant interactions has to be considered for an effective control strategy.

95

The characterisation of diffuse pollutants by behaviour rather than chemistry allows for a substantial degree of integration. Such characterisation has been proposed in terms of internal, external and recycled sources, mobilisation via detachment, solubilisation and contingent processes and delivery to water bodies by surface, preferential and through flows. Attributing pollutants to these groups and sub-groups and understanding how the mitigation measures affect those characteristics will facilitate a more complete approach to diffuse pollution prevention.

By removing/reducing pollutants at source, any further mitigation is more effective or may even be rendered unnecessary. By understanding pollutant behaviour in terms of mobilisation and by applying measures that mitigate against those processes, whole suites of pollutants may potentially be abated. Similarly, viewing measures in terms of the processes that deliver pollutants to surface waters will allow reductions in many pollutants simultaneously, or allow a prioritisation of strategies to apply the most cost-effective measures or the mitigation of the most damaging pollutants.

The development of the mitigation matrix proposed in Appendix I paves the way ahead to a combined, holistic approach to the problem of diffuse pollution.

96

NO2-

Path

Path

BOD

SOURCE

MOBILISATION

DELIVERY

NH4+

NH4+

NH4+

NO2-

NO2-

Path

BOD

BOD

Figure 14. A conceptual m

odel for the generic characteristics of diffuse pollutants.

Compatibility Pollution Conflicts with Farm Systems toMeasure Description Source Mobilisation Transport with ES scheme NH4+ NO2- BOD E. coli Crypto Cost/ha Implementation Public Farmer swapping other measures Uncertainty Which Applicable

Reduce/remove gut pathogens (antibiotics) RECYCLED ALL ALL 0 0 0 1* 1* 2 2 1 2 2 3 1 A1 to H3Reduce dietary N intake RECYCLED ALL ALL √ 2 2 0 0 0 3 1 2 2 3 3 2 A1 to B6, D1 to H3

Arable Reversion Conversion from arable production to grassland with extensive grazing with low fertilizer N inputs EXTERNAL ALL ALL √ 0 0 -1 -1 -1 1 1 3 1 2 3 2 I1,2,3

Reduce total farm livestock numbers Reduce numbers of livestock by 50% RECYCLED EXTERNAL ALL ALL √ 2 2 2 2 2 1 1 2 1 3 3 3 A1 to H3

Export manure off the farm Take 50% of the manure to farms that require extra nutrients RECYCLED ALL ALL √ 2 2 2 2 2 1 2 2 2 3 3 3 A1 to B6, D1 to H3

Composting of solid manure (>55C for 3 days) RECYCLED ALL ALL √ 1 1 1 3* 2* 3 2 3 2 2 2 2 A2, A4, A6, B2, B4, B6, C1-C4, D2, D4, D6, E2, E4, E6, G1 to H3

Batch storage of solid manure (90 days) RECYCLED ALL ALL √ 1 1 1 3* 2* 3 3 3 3 2 2 2 A2, A4, A6, B2, B4, B6, C1-C4, D2, D4, D6, E2, E4, E6, G1 to H3

'Fill and draw' storage of slurry RECYCLED ALL ALL √ 0 0 0 1* 1* 3 2 2 3 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Batch storage of slurry (90 days) RECYCLED ALL ALL √ 1 1 1 3* 2* 1 A 1 2 1 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Anaerobic digestion of slurry RECYCLED ALL ALL √ -1 -1 2 1* 1* 1 A 1 3 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Anaerobic digestion and pasteurisation of slurry RECYCLED ALL ALL √ 0 0 2 3* 2* 1 A 1 3 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Aeration of slurry RECYCLED ALL ALL √ 2 2 3 1* 1* 3 A 2 2 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Lime added to slurry RECYCLED ALL ALL √ 3 3 0 2* 1* 2 A 2 2 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Acid added to slurry RECYCLED ALL ALL √ 0 0 0 2* 1* 2 2 2 1 1 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Carbohydrates added to slurry RECYCLED ALL ALL √ 0 0 1 1* 1* 1 2 2 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Cover slurry store to reduce volume RECYCLED ALL ALL √ -1 -1 0 0 0 2 A 2 2 2 3 3 3 A1, A3, A5, B1, B3, B5, D1, D3, D5, E1, E3, E5

Integrate fertilizer with manures EXTERNAL ALL ALL √ 2 2 0 0 0 3 3 3 2 3 3 3 A1to E6, G1 to I3Introduce clover into the grassland system EXTERNAL ALL ALL √ 1 1 0 0 0 3 A 2 3 2 2 3 2 A1 to C3-4

Remove artifical fertilizers from farm system EXTERNAL ALL ALL √ 3 3 0 0 0 1 1 3 1 3 3 3 A1 to E6, G1 to I3

Change fertilizer type Change ammonium nitrate to calcium nitrate to reduce ammonium losses EXTERNAL ALL ALL √ 3 3 0 0 0 2 2 3 2 3 3 3 A1 to E6, G1 to I3

Collect and store dirty water RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 2 0 3 1 1 2 A 2 3 2 3 3 3 A1 to E6, G1 to H3

Minimise dirty yard area RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 1 0 2 1 1 3 A 3 3 3 3 3 3 A1 to E6, G1 to H3

Fertilizer application timing Avoid fertilizer spreading to fields at times of high risk EXTERNAL CONTINGENT SURFACE PREFERENTIAL √ 2 0 0 0 0 3 2 3 2 3 3 3 A1 to E6, G1 to I3

Avoid 'fresh' - solid manure spreading to fields at times of high risk RECYCLED CONTINGENT SURFACE

PREFERENTIAL √ 1 0 1 1 1 3 3 3 3 2 2 2 A2, A4, A6, B2, B4, B6, C1-C4, D2, D4, D6, E2, E4, E6, G1 to H3

Avoid 'fresh' - slurry spreading to fields at times of high risk RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 2 0 2 2 2 3 A 2 3 2 3 3 2 A1, A3, A5, B1, B3, B5, D1, D3, D5,

E1, E3, E5

Manure application location Do not apply slurry/manure to well connected hydrological areas (Defra water code) RECYCLED CONTINGENT SURFACE

PREFERENTIAL √ 2 0 2 2 2 1 3 3 3 3 3 3 A1 to E6, G1 to H3

Fertilizer application location Do not apply fertilizer to well connect hydrological areas (Defra water code) EXTERNAL CONTINGENT SURFACE

PREFERENTIAL √ 2 0 0 0 0 1 3 3 3 3 3 3 A1 to E6, G1 to I3

Incorporate fertilizer into soil EXTERNAL CONTINGENT SURFACE PREFERENTIAL √ 2 -1 0 0 0 3 2 3 2 2 3 2 A1 to E6, G1 to I3

Incorporate 'fresh' slurry into soil RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 2 -1 2 1 1 2 2 3 2 2 3 2 A1, A3, A5, B1, B3, B5, D1, D3, D5,

E1, E3, E5

Incorporate 'fresh' solid manure into soil RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 2 -1 1 1 1 2 2 3 2 2 3 2 A2, A4, A6, B2, B4, B6, C1-C4, D2,

D4, D6, E2, E4, E6, G1 to H3

Bandspread RECYCLED CONTINGENT SURFACE PREFERENTIAL √ -1 0 0 -1 -1 1 2 3 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5,

E1, E3, E5

Shallow injection RECYCLED CONTINGENT SURFACE PREFERENTIAL √ -1 -1 1 -1 -1 2 2 3 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5,

E1, E3, E5

Deep injection RECYCLED CONTINGENT SURFACE PREFERENTIAL √ -1 -1 -1 -1 -1 2 2 3 2 2 2 2 A1, A3, A5, B1, B3, B5, D1, D3, D5,

E1, E3, E5Site manure heaps away from water courses and drains (Defra

water code) RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 3 0 3 2 2 3 3 3 3 2 3 3 A2, A4, A6, B2, B4, B6, C1-C4, D2,

D4, D6, E2, E4, E6, G1 to H3

Site manure heaps on concrete and collect effluent RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 3 0 3 3 3 2 A 2 3 2 2 3 3 A2, A4, A6, B2, B4, B6,C1-C4, D2, D4,

D6, E2, E4, E6, G1 to H3

Change manure management system Change from slurry to solid manure handling systems RECYCLED CONTINGENT SURFACE PREFERENTIAL √ 2 1 2 2 2 1 A 1 3 1 2 3 2 A1, A3, A5, B1, B3, B5, D1, D3, D5,

E1, E3, E5

Fence off rivers from livestock RECYCLED SOIL

CONTINGENT DETACHMENT SURFACE √ 2 0 2 3 3 2 A 3 3 2 3 3 3 A1 to C4, F1

Provide bridges to avoid fording RECYCLED SOIL


Move livestock feeders and troughs regularly RECYCLED SOIL


Woodchip application to tracks and standings RECYCLED SOIL

CONTINGENT DETACHMENT SURFACE √ 1 0 1 1 1 2 3 3 2 3 3 2 A1 to C4, F1

Re-site gateways in high risk areas RECYCLED SOIL

CONTINGENT DETACHMENT SURFACE √ 1 0 1 1 1 2 A 2 3 1 3 3 3 A1 to I3

Provide crop cover during critical periods Undersow cover crops SOIL DETACHMENT SOLUBILISATION

PREFERENTIAL THROUGHFLOW √ 1 1 0 0 0 2 2 2 2 3 3 2 A5, A6, B5, B6, D1 to I3

Reduce field stocking rates to limit grazing pressures when soils are wet

RECYCLED SOIL


Avoid grazing high risk fields under wet conditions RECYCLED SOIL

CONTINGENT DETACHMENT SURFACE √ 1 0 0 1 1 2 2 3 2 3 3 2 A1 to C4, F1

Cultivate across the slope ALL DETACHMENT SURFACE √ 1 0 0 0 0 3 2 2 2 3 3 2 A5, A6, B5, B6, D1 to I3Avoid tramlines over the winter SOIL DETACHMENT SURFACE √ 1 0 0 0 0 3 2 3 2 3 3 3 A5, A6, B5, B6, D1 to E6, G1 to I3Maintain organic matter levels SOIL DETACHMENT SURFACE √ -1 -1 0 0 0 3 3 2 2 3 2 2 A5, A6, B5, B6, D1 to I3

Fertilizer Form Use slowly available nitrogen fertilizers EXTERNAL SOLUBILISATION ALL √ 1 1 0 0 0 1 3 2 2 2 3 2 E6, G1 to I3

Establish riparian strip ALL CONTINGENT DETACHMENT SURFACE √ 1 1 1 1 1 1 A 3 3 2 3 3 2 A1 to I3

Establish artificial wetlands ALLCONTINGENT DETACHMENT

SOLUBILISATION

SURFACE PREFERENTIAL √ 1 1 2 1 1 1 A 1 3 2 2 3 2 A1 to I3

Establish in field buffers ALL CONTINGENT DETACHMENT SURFACE √ 1 0 1 1 1 1 A 3 3 2 3 3 2 A1 to I3

Install hedges and make fields smaller ALL DETACHMENT SURFACE √ 1 0 1 1 1 1 A 1 3 1 3 3 2 A1 to I3Cultivate land ALL ALL SURFACE √ 1 0 1 1 1 2 2 2 2 3 3 2 A1 to I3

Allow drainage to deteriorate ALL ALL PREFERENTIAL THROUGHFLOW √ 1 1 1 1 1 2 2 3 1 2 2 2 A1 to I3

Environmental 1 = >£125 1 = Very difficult 1 = A lot 1 = many 1 = Very Uncertain See Appendix IIStewardship 2 = £50-£125 3 = Very easy 3 = None 3 = none 2 = Some Uncertaintyoption codes 1*= 1-2 log 10 reduction 3 = <£50 3 = Certain

(Martin, 2004. 2*= 3-4 log 10 reduction A = Amortised Costpers coms) 3*= 5-6 log 10 reduction

Reduce hydrological connectivity

SOU

RC

E

Establish buffer zones

Storage and treatment of manures

Manure application timing

DE

LIV

ER

YM

OB

ILIS

AT

ION

Siting of manure heaps

Limit livestock access to water courses

Dietary manipulation

Acceptability

Soil incorporation

Management of dirty water

Control Livestock Movement

Measure Target Area Importance of measure to control pollutants

Soil Cultivation

Manage livestock tracks and congregating areas

Reduce fertilizer inputs

Change slurry application technique from Broadcast

3 = >75%, 2 = 25-75%, 1 = <25% reduction3 = Acceptable1 = Unacceptable

-3 = >75%, -2 = 25-75%, -1 = <25% increase

98

8. Appendix II

MODEL FARM SCENARIOS

Livestock type/cropping

Soil type/drainage status Cropping Manure type*/

grazingFarm Category

NumberDairy Heavy – drained

Heavy – drainedHeavy – undrainedHeavy – undrained

Light – free drainingLight – free draining

GrassGrassGrassGrass

Grass/maizeGrass/maize

Slurry +grazingFYM + grazingSlurry +grazingFYM + grazingSlurry + grazingFYM + grazing

A1A2A3A4A5A6

Beef Heavy – drainedHeavy – drained

Heavy – undrainedHeavy – undrained


GrassGrassGrassGrass

Grass/maizeGrass/maize

Slurry +grazingFYM + grazingSlurry + grazingFYM + grazingSlurry + grazingFYM + grazing

B1B2B3B4B5B6

Sheep – lowlands

– uplands

Heavy – drainedHeavy – undrained

Light – free draining

All soils

GrassGrassGrass

Grass

FYM + grazingFYM + grazingFYM + grazing

FYM + grazing

C1C2C3

C4Pigs – indoor

(finishing unit)Heavy – drainedHeavy – drained

Light – free drainingLight – free drainingLight – free drainingLight – free draining

Arable – combinableArable – combinableArable – combinableArable – combinable

Arable – rootsArable – roots

SlurryFYMSlurryFYMSlurryFYM

D1D2D3D4D5D6

Pigs – indoor(breeding unit)

Heavy – drainedHeavy – drained

Light – free drainingLight – free drainingLight – free drainingLight – free draining

Arable – combinableArable – combinableArable – combinableArable – combinable

Arable – rootsArable – roots

SlurryFYMSlurryFYMSlurryFYM

E1E2E3E4E5E6

Pigs – outdoor Light – free draining Arable – combinable ‘Grazing’ F1Laying hens Heavy – drained


Arable – combinableArable – combinable

Arable – roots

Layer manureLayer manureLayer manure

G1G2G3

Broilers Heavy – drainedLight – free drainingLight – free draining


Arable – roots

Broiler litterBroiler litterBroiler litter

H1H2H3

Arable (no livestock)

Heavy – drainedLight – free drainingLight – free draining


Arable – roots

–––

I1I2I3

*This represents the ‘main’ manure type on the farm, as most farms will have both solid and liquid manures. Also, many farms will have dirty water to manage as part of their manure management system, particularly Farm Categories A1 – B6.

100

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DEFRA projects

NT1829 – Further N cycling studies on farmlets

NT2001 - Integration of animal manures in crop and livestock farming systems: Nutrient demonstration farms.

NT2402 - The Impact of nutrition and management on N and P excretion by dairy cows.

NT2511 – The cost curve of nitrate mitigation options.

PE0203 – Cost surve assessment of phosphorus mitigation options relevant to UK agriculture.

WA0503 - The fate of aqueous pollutants in runoff following slurry and dirty water applications to land.

WA0516 – Run-off and emissions from hardstandings.

WA0517 – Impacts of farm waste stores on groundwater quality.

WA0518 – Development of soil based treatment systems for dirty water.

WA0519 – Enhancing the effective utilisation of animal manures on-farm through compost technology.

WA0522 – Treatment of livestock wastes through the use of additives.

WA0523 – Survey element of WA0516 Run-off and emissions from hardstandings. (Completion - March, 1999).

WA0658 – Feasibility study – constructed wetlands for the treatment of dirty water and the creation of new wildlife habitats on farms.

WA0712 – Management techniques to minimise ammonia emissions during storage and land spreading of poultry manure.

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WA0716 – Management techniques to reduce ammonia emissions from solid manure (‘REAMS’).

Ongoing DEFRA-funded projects cited

WA0804 - Investigation of the routes by which pathogens associated with slurries and manure may be transferred from the farm to the wider environment.

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