A review of water quality concerns in livestock farming areas

The Science of the Total Environment 250 Ž2000. 143᎐167

A review of water quality concerns in livestock farming areas P.S. Hoodaa,U , A.C. Edwardsb, H.A. Andersonb, A. Miller a a

School of Biological and Molecular Sciences, Oxford Brookes Uni¨ ersity, Oxford OX3 0BP, UK b Macaulay Land Use Research Institute, Craigiebuckler, Aberdeen AB15 8QH, UK Received 1 July 1999; accepted 7 January 2000

Abstract Post-war changes in farming systems and especially the move from mixed arable᎐livestock farming towards greater specialisation, together with the general intensification of food production have had adverse affects on the environment. Livestock systems have largely become separated into pasture-based Žcattle and sheep. and indoor systems Žpigs and poultry.. This paper reviews water quality issues in livestock farming areas of the UK. The increased losses of nutrients, farm effluents Žparticularly livestock wastes., pesticides such as sheep-dipping chemicals, bacterial and protozoan contamination of soil and water are some of the main concerns regarding water quality degradation. There has been a general uncoupling of nutrient cycles, and problems relating to nutrient loss are either short-term direct losses or long-term, related to accumulated nutrient surpluses. Results from several field studies indicate that a rational use of manure and mineral fertilisers can help reduce the pollution problems arising from livestock farming practices. Several best management practices are suggested for the control of nutrient loss and minimising release of pathogen and sheep-dip chemicals into agricultural runoff. 䊚 2000 Elsevier Science B.V. All rights reserved. Keywords: Water contamination; Nitrate; Phosphate; Organic waste; Sheep-dipping chemicals; Bacterial and protozoan organisms

U

Corresponding author. Tel.: q44-1865-483267; fax: q44-1865-483242. E-mail address: [email protected] ŽP.S. Hooda. 0048-9697r00r$ - see front matter 䊚 2000 Elsevier Science B.V. All rights reserved. PII: S 0 0 4 8 - 9 6 9 7 Ž 0 0 . 0 0 3 7 3 - 9

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1. Background The effects of modern agriculture on the wider environment are causing concern for a number of reasons. Pollution arising from agricultural activities has increased mainly as a result of the intensification of food production systems. The demand for food production has been met by a combination of high yielding crop varieties and greater reliance on pesticides, fertilisers, and imported animal feedstuffs; since 1950 the consumption of N fertiliser in the UK has increased more than sixfold. Increasing use of fertilisers and imported feeds results in larger soil nutrient pools, potentially increasing the risk of nutrient losses we.g. nitrogen ŽN. and phosphorous ŽP.x through leaching and runoff. Similarly, the total number of livestock Žcattle, pigs, sheep and poultry. in the UK increased from 108 000 000 in 1940 to 188 000 000 in 1987. This has led to intensified farming practices where large numbers of animals are reared on relatively small areas, with the waste production Že.g. farmyard manure, slurry, dirty water, silage effluents and poultry litter. from these farms being large, and its disposal locally being aggravated by the limited land area available. Other European countries Že.g. Germany, France, The Netherlands, Ireland, Denmark, and Belgium. have also seen similar increases in fertiliser usage and livestock numbers ŽFAO, 1999.. The intensification of livestock production with

its associated increased demand for fodder has encouraged farmers to rely more heavily on chemical fertilisers and imported feeds, and very often the waste is considered as a disposal problem rather than a useful source of plant nutrients. Consequently, quantities of farmyard manure and slurry far in excess of crop requirements are frequently applied to soils, with storage and weather considerations often determining the timing and rates of application, rather than agronomic interests. The livestock wastes contain valuable quantities of N, P, K ŽTable 1. and other micronutrients and their methods of use requires rationalisation in order to complement the benefits derived from fertilisers. Practices including manurerslurry applications at times when their beneficial effects cannot be fully realised also have detrimental implications for the wider environment, including water quality. In livestock farming areas, excessive loss of nutrients Žprincipally N and P. and farm effluents Že.g. silage, slurry. in surface runoff andror through leaching are the principal causes of degradation in surfaceand ground water quality. Although not widely recognised, the chemicals used in sheep-dipping operations and bacterial and protozoan pathogens also pose a threat of pollution to watercourses. In this paper, we review major water quality concerns in livestock farming areas with particular reference to studies in north-west European and North American countries relevant to the situation pertaining to the UK.

Table 1 Major nutrients in typical livestock wastes a Total nutrients Žnutrients available in parentheses . N

P

K

Solids (kg t ) Cattle FYM Ž25% DM. Pig FYM Ž25% DM. Broiler litter Ž60% DM.

6 Ž1.5. 6 Ž1.5. 29 Ž10.0.

3.1 Ž0.78. 2.62 Ž1.53. 9.60 Ž5.67.

5.80 Ž3.48. 3.31 Ž2.90. 13.27 Ž9.95.

Slurries (kg my 3) Cattle slurry Ž6% DM. Pig slurry Ž6% DM.

3 Ž1.0. 5 Ž1.8.

0.52 Ž0.26. 1.31 Ž0.65.

2.98 Ž2.49. 1.99 Ž1.65.

y1

a

Source: Webb and Archer, 1994. Abbre¨ iations: DM, dry matter; FYM, farmyard manure.

P.S. Hooda et al. r The Science of the Total En¨ ironment 250 (2000) 143᎐167

2. Nutrients Increased loss of nutrients in agricultural runoff has potentially serious ecological and public health implications. In this regard nitrogen and phosphorus are particularly important as both are implicated in aquatic eutrophication ŽLevine and Schindler, 1989.. However, because of the free air᎐water exchange of N and the fixation of atmospheric N by some blue᎐green algae, P is generally regarded as the eutrophication-limiting nutrient in most aquatic ecosystems ŽSharpley and Menzel, 1987.. Eutrophication and the associated ecological effects result in a general decline in overall water quality, restricting its use for general and drinking purposes ŽUSEPA, 1990; Sharpley and Withers, 1994.. The loss of nitrate in agricultural runoff has potentially serious implications for the quality of potable water, since concentrations in potable water above 50 mg ly1 have been implicated in some human health problems, including infant methemoglobinemia Ž‘blue baby’ syndrome ᎏ Royal Commission on Environmental Pollution, 1979; WHO, 1985.. In 1970, WHO recommended a drinking water nitrate upper limit of 50 mg ly1 Ž11.3 mg ly1 NO 3 ᎐N., and although that remains

145

in most European countries ŽCEC, 1980., the current American standard and WHO ‘guide value’ is 45 mg nitrate ly1 Ž10 mg ly1 NO 3 ᎐N.. Nutrients in agricultural runoff can arise from point or diffuse sources of pollution ŽFig. 1., with major point-sourced pollution incidents occurring due to poor containment of slurry or silage effluents. Such point sources of pollution are easy to identify and control, and the 1989 Water Act in the UK provided necessary legislation for controlling these incidents. However, the diffuse sources of pollution, such as losses of nutrients through leaching and in surface runoff ŽFig. 1. are more difficult to assess and control, and it is these sources which will form the focus of this review. 2.1. Nitrogen loss in surface runoff The combined use of livestock manure and mineral fertiliser results in considerable enrichment of surface soils with nutrients ŽCulley et al., 1981; Daniel et al., 1993., and overland flow from such areas during or after a rainfall event contains relatively high levels of nutrients Že.g. N, P,

Fig. 1. Pathways of diffuse and point sources of nutrients and farm effluent inputs to catchment waters in livestock farming areas.

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K., organic matter and suspended particles ŽKhaleel et al., 1980.. The runoff water may have a relatively large concentration of organic-N associated with particulate matter, but generally contains little nitrate ŽHallberg, 1989.. For example, surface runoff during a 6-week period, from a 102-km2 catchment area, under rotational grazing in Florida, exported 1.3 kg N hay1 as inorganic-N Žincluding up to 0.3 kg NO 3 ᎐N hay1 . to the drainage canal, compared with 4.4 kg N hay1 as dissolved organic-N ŽDierberg, 1991.. Large concentrations of nitrate found in streams and other watercourses draining grassland and arable areas come predominantly from ground water discharges and sub-surface flow including tile or pipe drainage into the streams ŽBaker and Laflen, 1983; Hallberg, 1987; Gangbazo et al., 1995.. Several studies of nitrogen loads andror concentrations in catchment waters are summarised in Table 3; others are discussed below to illustrate key points about nitrogen loss in surface runoff. The amount of N in surface runoff is strongly influenced by a combination of land use and management practices, soil types and climatic conditions. Dierberg Ž1991. noted that the total N loss in surface runoff during a 6-week period from a relatively undeveloped sub-catchment was much smaller Ž0.7 kg N hay1 . than that from a second sub-catchment composed of cattle grazing pastures Ž2.4 kg N hay1 .. The nitrate losses from the two sub-catchments of the Indian River Lagoon ŽFlorida, USA. were only 0.1 and 0.3 kg NO 3 ᎐N hay1 , respectively, with the remainder being largely organic-N. The author concluded that agricultural land use within the catchment Ž256 km2 . was the largest exporter of dissolved inorganic-N. The transport of N in runoff from sites where livestock manure has been applied is dependent on the timing and rate of manure application, together with site Že.g. soil type, slope. and climate Že.g. rainfall amount and intensity . factors. For instance, total N lost in runoff from alfalfa plots which received livestock manure at the rate of 120 kg N hay1 yeary1 was 18.7 kg hay1 yeary1 when applied in the winter; compared to the loss of 9.1 kg N hay1 yeary1 when applied in the spring. Similarly, cattle slurry applied at the rates

of 224 and 560 kg N hay1 yeary1 resulted in a total N runoff losses of 16 and 54 kg N hay1 yeary1 , respectively Žcited in Khaleel et al., 1980.. Ammonia ŽNH 3 . concentrations in surface waters are generally very small due to the average pH of most surface waters being sufficiently low . to convert all ammonia to ammonium ion ŽNHq 4 ; however, as little as 0.02 mg ly1 NH 3 may be toxic to fish and other forms of aquatic life, particularly at high pH ŽCooper, 1993.. In some localised livestock farming areas, watercourses have been found to show elevated concentrations of NH 3 ; continuous monitoring of Clarbeston Stream in south-west Wales draining exclusively livestock farming areas showed that the stream was grossly polluted with background level of 3᎐5 mg ly1 NH 3 ᎐N, with peaks as high as 20 mg ly1 . The results also showed that rainfall events were generally followed by NH 3 ᎐N peaks 3᎐4 h later in the river ŽSchofield et al., 1990.. This deterioration in water quality following rainfall events was probably due both to waste washing from the farmyards and to runoff from slurry-treated fields. Similar incidental rises in NH 3 ᎐N concentration were observed by Foy and Kirk Ž1995. while evaluating the effect of land use on stream water quality in 21 agricultural catchments in Northern Ireland. Although nitrate in surface runoff is generally small, large fluxes of N in other forms Žespecially dissolved organic or particulate mineralrorganic. in the runoff eventually contribute nitrate into the surface water through mineralisation and nitrification. Together with nitrate in sub-surface flow Žtile drains. and ground water discharge, this may raise the loads in rivers and other receiving waters even in the absence of a runoff event. Nitrate concentrations in surface waters draining from arable farming areas are generally considered to be more significant than from livestock farming areas. Houston and Brooker Ž1981. found that the NO 3 ᎐N concentration in the predominantly grassland sub-catchments Ž81:14 grasslandrarable. of the River Wye in Wales was 2.93 compared with 4.99 mg ly1 in other sub-catchments with less grassland Ž59:35 grasslandrarable . ŽTable 3.. These findings are supported by other similar studies from America ŽKeeney and

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DeLuca, 1993., Ireland ŽNeill, 1989. and Scotland ŽHooda et al., 1997a; Domburg et al., 1998..

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The predominant form of N leaching loss was clearly demonstrated in recent studies where 97᎐98% of the N in sub-surface flow was in the form of NO 3 ᎐N ŽLowrance, 1992; Gangbazo et al., 1995.. However, significant amounts of organic-N have also been measured in subsurfaceflow under specific conditions, e.g. cattle slurry applied to grassland soils when they were at or around field capacity moisture content ŽJordon and Smith 1985., or when rain or snow followed a slurry application ŽHooda et al., 1998a.. The extent of nitrate leaching is strongly influenced by the various land use and management

2.2. Nitrogen loss through leaching . and ammonium ŽNHq . are the Nitrate ŽNOy 3 4 two forms of inorganic nitrogen in soils; nitrate ion is freely mobile in the soil solution and is, therefore, potentially vulnerable to leaching below the rooting zone as water moves through the soil, whereas ammonium ion tends to be retained by the soil through cation exchange mechanisms.

Table 2 Summary of selected field studies reporting nitrate leaching through field drains as influenced by land use and management practices Land use and management

Fertiliser N Žkg N hay1 .

Mean nitrate-N Žmg N ly1 .

Nitrate-N leached Žkg N hay1 .

Reference

Grass plots, fertilised, cut for silage, no grazing, no manurerslurry

250 500 900

2.06 13.2 66᎐98

3.8 27 150

Barraclough et al. Ž1983.

Grass plots, fertilised, cut for silage, no grazing, no manurerslurry

0 300 450

᎐ ᎐ ᎐

10 44 51

Thomsen et al. Ž1993.

Permanent fertilised pasture, dairy cattle grazed, no slurryrmanure

51 408

2.3 4.3

15.3 28.5

Sherwood and Ryan Ž1990.

Permanent fertilised pasture, beef cattle grazed but no slurryrmanure

200 400

Permanent fertilised pasture, cattle grazed, no manurerslurry

100 200 300 400

2.98 5.35 8.45 17.5

16 24 36 71

Watson et al. Ž1992.

Permanent fertilised pasture, cut for silage, cattle and sheep grazed, intensive cattle slurry inputs Permanent grass᎐clover pasture, no mineral-N, cut for silage, cattle and sheep grazed, intensive cattle slurry inputs Arable, fertilised, cereal᎐sugarbeet rotation Vegetable crops, fertilised and manured Arable, maize, fertilised and manured Moderately fertilised, sugarbeet᎐wheat rotation Coniferous woodland

245

3.96᎐10.2

30᎐45

Hooda et al. Ž1998a.

3.1᎐8.5

24᎐38

Hooda et al. Ž1998a.

0

120

7᎐19 20᎐60

25᎐30

38.5 133.8

Scholefield et al. Ž1993.

᎐

Strebel et al. Ž1989.

300᎐600 300 150

34᎐70 23.7 36.5

᎐ 136.6 50

Strebel et al. Ž1989. Theocharopoulos et al. Ž1993. Rossi et al. Ž1991.

᎐

-1

᎐

Hallberg Ž1989.

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practices. For a given production system, leaching of nitrate tends to increase with N inputs ŽTable 2.. Nitrate leaching from non-grazed grasslands is generally less than that from arable land use, with or without manure inputs ŽTable 2.. Two factors that cause this difference are, firstly, fertiliser-N input in non-grazed grassland systems is relatively smaller than in most arable systems, and secondly, ploughing operations after each crop andror manure application in an arable system favour soil mineralisation and nitrification processes. At least in the short-term, this results in larger nitrate pools in arable soils than in soils under perennial grass cover. In intensively managed grassland systems where fertiliser-N inputs alone can vary between 300 and 420 kg N hay1 ŽADAS, 1990., nitrate leaching can, however, be greater than both non-grazed grassland and arable systems ŽTable 2.. Manure produced on intensive livestock farms is often far in excess of agronomic requirements. This leads to repeated manure applications at rates that are greater than crop requirements, with a consequent large amount of unutilised Žsurplus. N ŽChang and Entz, 1996.. For example, a study of 177 Dutch dairy farms cited in Korevaar and Den Boer Ž1990., showed an average N surplus of 486 kg N hay1 . Such large N surpluses at dairy, pig and poultry farms have also been common in other western European countries ŽPolitiek and Bakker, 1982; Jarvis, 1993.. These N surpluses are clearly the main source of excessive N losses to the environment, including those in the form of nitrate leaching. For example, a study of over 200 groundwater wells, used for private water supplies in Delaware ŽUSA., showed a mean NO 3 ᎐N concentration of 21.9 mg ly1 in poultry production areas compared with 6.2 for corn and soybean production, and 0.58 for forested areas ŽRitter and Chirnside, 1987.. Climate Žrainfall and temperature. and soil types can also strongly influence nitrate leaching. In a Northern Ireland study, the annual leaching losses of NO 3 ᎐N varied between 7.8 and 30.3 kg hay1 from an intensively managed grassland receiving 300 kg N hay1 solely through inorganic fertilisers ŽJordon and Smith, 1985., with the greatest nitrate leaching occurring after heavy

rainfall in the autumn of 1982 and winter of 1983᎐1984. A similar range in nitrate leaching losses from south-west England ŽScholefield et al., 1993. and south-west Scotland ŽHooda et al., 1998a. have been reported. Generally, for individual sites there is an inverse relationship between the concentration of nitrate leached and rainfall amount. Soil texture is also an important factor that can result in contrasting nitrate leaching between sites. Bergstrom and Johansson Ž1991. monitoring nitrate leaching losses from different types of soils receiving 100 kg N hay1 found that the largest losses Ž65 kg NO 3 ᎐N hay1 yeary1 . occurred on sandy and peat soils. Two loamy soils lost between 25 and 40 kg NO 3 ᎐N hay1 yeary1 , with the smallest losses of 20 kg NO 3 ᎐N hay1 yeary1 occurring in a clay soil. Similarly, in an Irish study, while 15% of the applied N in slurry was lost by leaching from a well-drained loam, only 2% was leached from an impermeable gley soil ŽJarvis et al., 1987.. However, overall smaller leaching losses from less permeable soils Že.g. clay. may result in increased losses in surface runoff and through gaseous emissions due to increased denitrification in such soils. Nitrate concentrations in surface waters generally reflect the intensity of agricultural production, but not in every instance, particularly when comparing catchments with different soil types, soil-N supply and N use efficiency. Hooda et al. Ž1997a. monitored nitrate concentrations in six predominantly agricultural catchments, three each in the north-east and west of Scotland. The western catchments are in one of the most intensive dairy farming areas in the UK, receiving 200᎐300 kg mineral-N hay1 and regular cattle slurry inputs. The north-eastern catchments, on the other hand, represent a mixed pattern of land use Žbeef, sheep, arable and forestry., receiving relatively small mineral-N Ž80᎐110 kg N hay1 . and manureN. Despite fertiliser- and manure-N inputs being more than double in the western catchments, stream nitrate concentrations were approximately four times larger in the north-eastern catchments. The larger nitrate concentrationsrloads in the north-eastern catchments ŽTable 3. were attributed to a combination of enhanced leakage from the mineralisation sector of the N-cycle, and

Table 3 Summary of selected catchment studies on nitrogen and phosphorus in streamsrrivers draining predominantly farming areas Catchment land use and management practices

Loads andror concentrations in catchment waters

Reference

Phosphorus

Wales: two large catchments, 142 km2 Frome Ž59% grassland and 35% arable. and 144 km2 Trothy Ž81% grassland and 14% arable., grazed manured and fertilised pastures, no manurerslurry in arable areasa .

4.99 mg NO3 ᎐N ly1 ŽFrome. 2.93 mg NO3 ᎐N ly1 ŽTrothy. 0.05᎐0.06 mg NH4 ᎐N ly1

0.42 mg MRP ly1 ŽFrome. 0.19 mg MRP ly1 ŽTrothy.

Houston and Brooker Ž1981.

Nebraska, USA: a small pastureland Ž43 ha., no mineral-P, rotationally grazed by cattle with moderate stocking density.

0.78 kg NH4 q NO3 ᎐N hay1 2.8 kg TKN hay1

0.69 kg TP hay1 0.38 kg MRP hay1

Doran et al. Ž1981.

Ontario, Canada: 137 ha catchment, cropped with maize and alfalfa, manured and fertilised, ditch drained. Northern Ireland: six large catchments Ž93% grass and 7% arable., fertilised and manured pastures, grazed as well as cut for silagea .

0.03᎐0.39 mg NH4 ᎐N ly1 4.9᎐13.5 mg NO3 ᎐N ly1 4.8᎐32.5 kg NO3 ᎐N hay1 3.1᎐10.7 kg KN hay1

0.01᎐0.12 mg MRP ly1

Phillips et al. Ž1982.

0.23᎐0.84 kg SRP hay1 0.09᎐0.22 kg SOP hay1 0.07᎐0.48 kg PP hay1

Foy et al. Ž1982.

Georgia, USA: eight catchments, forestry Ž15᎐70%. and pastureland Ž23᎐39%. as the main land uses.

3.2᎐36.9 kg NO3 ᎐N hay1 3.6᎐35.7 kg TN hay1

0.012᎐0.745 kg SRP hay1 0.168᎐1.41 kg TP hay1

Nearing et al. Ž1993.

Lac Leman, France: a small agricultural catchment Ž14 ha., grassland as the main land use.

14.6 kg TN hay1 Ž73% as NO3 ᎐N.

0.60 kg TP hay1 Ž; 50% as DP.

Dorioz and Ferhi Ž1994.

Scotland: The River Ythan catchment Ž689 km2 ., 95% agricultural Žarable, beef, pig, sheep and poultry., 5% forestryruplandrurban, intensively fertilised and manured.

19 kg NO3 ᎐N hay1 5.2᎐7.5 mg NO3 ᎐N ly1

0.26 kg MRP hay1 0.026᎐0.065 mg MRP ly1

MacDonald et al. Ž1995.

Scotland: six small catchments Ž98᎐1128 ha.; three predominantly dairy farming ŽDF. and three had mixed land use ŽMLU. patterns Žarable, beef, forestry., fertilised and manured.

3.5᎐8.8 kg NO3 ᎐N hay1 ŽDF. 0.54᎐1.04 mg NO3 ᎐N ly1 ŽDF. 18.5᎐22.9 kg NO3 ᎐N hay1 ŽMLU. 2.86᎐4.25 mg NO3 ᎐N ly1 ŽMLU.

0.66᎐1.19 kg MRP hay1 ŽDF. 3.94᎐5.59 kg TP hay1 ŽDF. 0.09᎐0.14 kg MRP hay1 ŽMLU. - 0.20 kg TP hay1 ŽMLU.

Hooda et al. Ž1997a. Hooda et al. Ž1997b.

P.S. Hooda et al. r The Science of the Total En¨ ironment 250 (2000) 143᎐167

Nitrogen

a Catchments with significant point-sourced P inputs. TP, total phosphorus; MRP, molybdate reactive phosphorus; SRP, soluble reactive phosphorus; DP, dissolved phosphorus; TKN, total Kjeldahl nitrogen; SOP, soluble organic phosphorus; KN, Kjeldahl nitrogen; TN, total nitrogen.

149

150

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of a shorter growing season, together with permeable soils in these catchments. In the western catchments, a long growing season coupled with relatively smaller soil N-cycle source and finertextured soils, probably resulted in more-efficient N utilisation and hence less nitrate leaching, even allowing for greater fertiliserrslurry inputs ŽHooda et al., 1997a.. Both manure type Že.g. cattle or pig slurry, poultry manure. and time of its application can also have significant impact on nitrate leaching. Manure is generally composed of NH 4 ᎐N and organic-N, often in equal ratios. While manureborne NH 4 ᎐N is readily available to plants, organic-N must undergo mineralisation to become available, and NH 4 ᎐N has to be biologically oxidised to NO 3 ᎐N before manure-N can contribute to nitrate leaching. Therefore, the concurrence Žor not. of plant uptake and the factors that control the processes of N mineralisation and nitrification, will largely determine the effect of manure-N on nitrate leaching. Poultry manure application to an arable loamy sandy soil made in October᎐November resulted in an average loss of 131 kg NO 3 ᎐N hay1 while the loss from January᎐February application averaged 35 kg NO 3 ᎐N hay1 ŽUnwin et al., 1991.. In the same experiment, cattle slurry supplying similar amounts of N applied in October and November resulted in leaching losses of 41 and 59 kg NO 3 ᎐N hay1 , respectively. Similar experiments conducted on grassland showed that cattle slurry application in September resulted in smaller nitrate losses compared to that applied in October or November ŽSmith and Chambers, 1993.. This confirms that N applied at a time when it cannot be utilised by the plants is likely to result in greater leaching losses, and N is generally utilised much more effectively by grass and field crops following spring applications of slurryrmanure, with least benefits from autumnrearly winter applications ŽBailey, 1993.. Ruminants excrete more than 75% of their N intake in urine and faeces ŽBall and Ryden, 1984.. In grazed grassland systems, this additional source of N input into the soil in the form of urine and dung patches results in N losses through nitrate leaching and gaseous emissions ŽGarwood and

Ryden, 1986; Jarvis et al., 1987.. The importance of this effect on nitrate leaching depends on the stocking density of grazing animals and on the length of the grazing period ŽStrebel et al., 1989.. According to a recent estimate, N deposited by grazing livestock in the form of urine and faeces ranges between 400 and 1200 kg N hay1 ŽGarrett et al., 1992., but a substantial amount of this N is likely to be lost through NH 3 volatilisation and as N2 , N2 O and NO x by Žde.nitrification. Nevertheless, for a given level of fertiliser application, N input in a grazed sward remains far greater than a non-grazed sward, and thus gives rise to larger nitrate leaching losses ŽTable 2.. Nitrate leaching from grassland ŽJordon and Smith, 1985; Sherwood and Ryan, 1990; Watson et al., 1992; Scholefield et al., 1993; Hooda et al., 1998a. and arable soils receiving manure inputs ŽAngle et al., 1993; Jemison and Fox, 1994. increases with N input, and intensifies above a certain level of N input. Under north-west European conditions, soils under perennial grass cover will generally loose less than 50 kg NO 3 ᎐N hay1 yeary1 , provided fertiliser N input does not exceed 250 kg N hay1 ŽTable 2.. However, far greater nitrate losses have occurred in situations prone to leaching or with greater N inputs ŽRyden et al., 1984; Macduff et al., 1989; Barraclough et al., 1992; Watson et al., 1992.. 2.3. Phosphorus loss in surface runoff Phosphorus in surface waters arises from various sources, which include runoff from agricultural soils, domestic, farm and industrial effluents Že.g. farm and municipal sewage, silage effluents ., groundwater discharge and atmospheric deposition ŽRyding et al., 1990.. However, P in agricultural drainage is the main source, particularly in areas with no inputs from sewage sludge or other point sources of pollution ŽIsermann, 1990; Nearing et al., 1993; Withers, 1993; Sharpley and Withers, 1994; Sims et al., 1998.. A number of important studies on P outputs andror concentrations in catchment waters are summarised in Table 3; others are discussed below to illustrate key factors that influence P transport in surface and subsurface runoff.

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Phosphorus transport in surface runoff occurs both in soluble and in particulate forms, generally dominated by the latter ŽSharpley and Menzel, 1987., and, like N, is influenced by several factors, including land use and management practices, soil type, time and rate of fertiliser and manurerslurry application. Nearing et al. Ž1993., monitoring water quality of eight catchments that feed Lake Lanier in Northern Georgia, observed that export of both soluble and total P from predominantly forested catchments was much less than that from catchments with a substantial area under fertilised pastures or arable crops. Similarly, a long-term study of 16 catchments in Oklahoma by Sharpley et al. Ž1989. showed that the concentrations of soluble and total P in runoff from native grass catchments Ž0.13 and 0.34 mg P ly1 , respectively. were considerably smaller than from cropped catchments Ž0.21 and 2.42 mg P ly1 , respectively.. The importance of soil type in controlling P loss in surface runoff has also been demonstrated, e.g. by Sharpley Ž1995., who investigated the relationship between the concentration of P in runoff and in soil from 10 Oklahoma soils amended with poultry litter, supplying 0᎐500 mg P kgy1 soil. Two soils with 200 mg kgy1 Mehlich-3 P supported a dissolved P ŽDP. concentration of 0.28 Žmontrorillonitic clay, thermic Pellustert . and 1.36 mg ly1 ŽStigler silt loam, thermic Aquic Paleudalf.. The DP in runoff was related to soil P sorption capacity, with the release of P to runoff increasing as P sorption decreased, regardless of the levels of extractable soil-P. Hooda et al. Ž1997b., studying the transport of P in streams draining predominantly livestock farming areas in Scotland, found that not only soil type affected the amount, but also the form of P. The P outputs in streams draining the well-drained coarse-textured podzols were more than four-times less than those draining the poorly-drained fine-textured gleysols ŽTable 3., despite the fact that the former group of soils had four-times larger acetic acid extractable-P. Furthermore, while the losses from catchments with well-drained soils were essentially entirely in the form of molybdate reactive phosphorus ŽMRP., particulate phosphorus ŽPP. formed a major proportion Ž45᎐86%. of total phosphorus ŽTP. that

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was lost from catchments with poorly-drained soils ŽHooda et al., 1997b.. These findings generally support the earlier observations of P in surface runoff being transported predominantly in the particulate phase Žcited in Sharpley and Menzel, 1987; Sharpley and Withers, 1994; Haygarth and Jarvis, 1999.. However, the study also has highlighted the importance of soil type in determining the distribution of P among its various forms in catchment waters, with P occurring almost exclusively in the form of DP ŽMRP. in certain catchments Že.g. well-drained soil with little or no erosion.. Nevertheless, P loss in any form, including PP, may provide a long-term source of P to the aquatic biota of a water body ŽSharpley and Menzel, 1987.. In intensively managed grassland systems in the UK, fertiliser P inputs may vary between 20 and 50 kg P hay1 ŽChadwick, 1991., and together with continual long-term application of livestock slurries and manures, particularly at rates greater than crop requirements, this results in excessive accumulation of P and other nutrients in the soils. In the Southern Plains of America, longterm pig and poultry manure applications increased P and N levels in the surface Ž50 cm. soils by four- and five-times, respectively ŽDaniel et al., 1993., and similar Canadian work involving land spreading of cattle slurry also showed a steady increase in soil NPK levels in the 0᎐15-cm layer ŽCulley et al., 1981.. These reports suggest that long-term excessive manure or slurry application can result in saturation of soils with these nutrients, especially P which is relatively less mobile in soils. Recently-conducted surveys have shown that a total area of 270 000 ha in the sandy areas in the middle, eastern and southern parts of the Netherlands are P-saturated due to intensive application of livestock wastes ŽUunk, 1991.. According to similar surveys, 30% of grassland soils in the UK have moderate Ž26᎐45 mg kgy1 . to high Ž) 45 mg kgy1 . Olsen extractable soil P contents ŽWithers, 1993.. Phosphorus in such sites with long-term repeated manure applications is vulnerable to losses in runoff, and Sharpley et al . Ž1991. clearly demonstrated this relationship while monitoring the effects of long-term applications Ž9᎐15 years. of

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manure on three sites. Between two- and ninefold increases in TP occurred in the upper 50 cm of the soils, leading to a 10-mm event runoff averaging 0.09 and 1.41 mg ly1 from untreated- and Žpig and poultry. manured soils, respectively. Likewise, grazing activities generally increase P loss in runoff from pastures, tending to increase with stocking density, DP and TP in runoff events from a grazed pasture averaged 1.38 and 2.14 mg ly1 , respectively, compared to 0.83 and 1.28 mg ly1 from a non-grazed pasture ŽSchepers and Francis, 1982.. The increased P losses from grazed pastures are at least partly due to additional P inputs associated with animal excretion in such areas. Where fields are intensively fertilised, manured and grazed, such intensive livestock management can also have a major impact on surface runoff P, with up to 20% of the applied P being lost under some extreme conditions. Phosphorus concentrations as large as 30 mg ly1 were recorded in initial surface runoff from areas receiving regular livestock slurry applications, and the concentrations averaged ) 1 mg ly1 , even several weeks after the waste application ŽSherwood and Fanning, 1981.. However, the largest losses of P have been due mainly to incidental losses of manures being applied to sloping, poorly drained andror frozen soils ŽKlausner et al., 1980; Uhlen, 1981.. Additional evidence of large diffuse-sourced P losses was provided in a recent study by Hooda et al. Ž1997b. who measured P outputs in six agricultural catchments in Scotland. They measured P outputs as high as 5.59 kg TP hay1 yeary1 in streams draining intensive cattle farming catchments, the highest losses being from a catchment with predominantly poorly-drained gleysols of inherently low P sorption capacities. Similar studies from the Netherlands reported that a concentration of 1 mg P ly1 was not uncommon in streams draining intensive livestock farms on sandy soils ŽUunk, 1991., a P concentration 10 times higher than that normally present in the runoff water from agricultural land. 2.4. Phosphorus loss through leaching Phosphorus leaching, including that via subsur-

face runoff Že.g. tile and mole drainage. is rarely considered as an important pathway for the transport of agricultural P to waters, a view stemming from the Žgenerally-. strong retention of P in soils, and earlier work which showed P leaching being an insignificant fraction of P losses via surface runoff ŽHanway and Laflen, 1974.. That still might be the case where water percolates through well-structured mineral soils, which have been fertilised in accordance with crop requirements, but combinations of certain farming practices Že.g. excessive fertiliser and manure application., soil properties Že.g. drainage by preferential-flow, low P retention due either to P-saturation or inherently low sorption capacity. and artificial drainage Že.g. tile or mole. may result in significant P leaching ŽSims et al., 1998.. The issue of P leaching through field drains Žsubsurface runoff. was first highlighted more than 25 years ago ŽRyden et al., 1973., and was recently comprehensively reviewed by Sims et al. Ž1998., who noted that P losses through artificial drainage systems could be as important as in surface runoff. Artificial drainage has become part of modern agriculture both in North America and northwestern Europe, and a number of studies reporting P losses through such drainage systems in livestock farming areas are summarised in Table 4. The perception in intensive livestock farming that manure Žincluding poultry litter and livestock slurry. is a waste disposal problem leads to overlapping application of both manure and mineral fertilisers, and as a result, a net accumulation rate of P in excess of 20 kg P hay1 yeary1 has been observed in many European Union countries ŽBrouwer et al., 1995; Haygarth et al., 1998; Hooda et al., 1999.. Similar large P accumulations in soils receiving long-term poultry litter, beef feedlot excreta, and cattle slurry have been reported in North America ŽDaniel et al., 1993; Mozaffari and Sims, 1994; Simard et al., 1995.. Such large increases in P accumulation and resultant saturation of P sorption capacities of surface soils will increase downward P migration, with a potentially high risk of P leaching through field drains Žsubsurface runoff.. Mozaffari and Sims Ž1994., investigating P accumulation in 48 Southern Delaware

Table 4 Summary of selected field studies of phosphorus leaching through tilermole drains as influenced by land use and management practices a P concentration Žmeanrrange. in subsurface drainage Žmg ly1 .

P losses in subsurface Žleaching. drainage Žkg hay1 yeary1 .

Reference

Fertilised Ž50 kg P hay1 yeary1 and ufertilised pastures

0.06 SRP Žunfertilised . 0.19 SRP Žfertilised.

0.17 TP Žunfertilised . 0.71 TP Žfertilised.

Sharpley and Syers Ž1979.

Cropped to maize, wheat and soybean, manured with dairy cattle manure

0.011 MRP Žno manure. 0.022 MRP Žmanured, 200 t hay1 .

NR

Hergert et al. Ž1981.

Cropped to continuous silage maize, manured with livestock manure Ž134 kg P hay1 yeary1 .

SRP Žcontrol plots. 0.02᎐0.17 SRP Žmanured plots.

NR

Phillips et al. Ž1981.

Permanent pasture, fertilised with mineral-P and livestock slurry, cut for silage, grazed by dairy cattle

0.05 SRP

0.1᎐0.2 g SRP hay1 dayy1

Smith et al. Ž1995.

Permanent fertilised pasture Ž25 or 50 kg P hay1 yeary1 , no slurryrmanure, beef cattle grazed

0.02 MRP

0.2 MRP

Hawkins and Scholefield Ž1996.

Permanent grassland in intensive dairy and pig farming areas, receiving both mineral and manurerslurry inputs

SRP ranged up to 4.8

1.29 SRP

Stamm et al. Ž1998.

Permanent grass pasture, no mineral-P, cut for silage, dairy cattle and sheep grazed, intensive cattle slurry inputs,

0.16 and 0.30 MRP Žmean values.

1.27᎐1.34 MRP 2.97᎐3.58 TP

Hooda et al. Ž1999.

Permanent grass᎐clover pasture, mineral-P Ž25 kg P hay1 yeary1 , cut for silage, dairy cattle and sheep grazed, intensive cattle slurry inputs

0.26 and 0.38 MRP 0.27 Žmean values.

1.68᎐2.03 MRP 3.47᎐5.03 TP

Hooda et al. Ž1999.

a

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Land use and management

Abbre¨ iations: NR, not reported; TP, total phosphorus; MRP, molybdate reactive phosphorus; SRP, soluble reactive phosphorus.

153

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soils with a history of long-term poultry litter application found evidence of downward P movement up to a depth of 40 cm, and Hooda et al. Ž1999. reported significant downward movement of cattle slurry-derived P in a non-calcareous gleysol to a similar depth. Whether or not such downward movement of P in soils leads to increased P loss to surface waters via subsurface runoff or to groundwater will be determined by the site hydrology, the depth of field drains, the subsoil chemistry and the groundwater table. The depth of field drains Žtilermole. often varies between 50 and 100 cm; the shallower the depth, the greater the risk of P being transferred to surface waters through this route ŽStamm et al., 1998.. Nevertheless, many studies have reported greater P leaching Žvia artificial drainage systems. from manured andror fertilised soils compared to that from unfertilisedrunmanured soils ŽTable 4.. However, it is the antecedent history of manuring and fertiliser application at a site rather than P inputs in a single season or year that determines P leaching losses. Continuous longterm excess P inputs will gradually increase the Degree of Soil Saturation with P ŽDSSP ᎏ where 100% s total P sorption capacity., simultaneously increasing the risk of P leaching. Hooda et al. Ž1999. assessed the effect of long-term P inputs on P accumulation and leaching from two pastures established on a non-calcareous gleysol, where after 9 years, NaHCO3-soluble P in the topsoil Ž0᎐10 cm. averaged 38 mg P kgy1 for the grass pasture Žonly slurry-P inputs., and 47 mg P kgy1 for the grass᎐clover pasture Žslurry- and fertiliser-P inputs.. The P leaching losses from the grass᎐clover pasture were significantly greater than those from the grass pasture, reflecting the enhancement in DSSP associated with the larger inputs. In the Netherlands, a DSSP value of 25% is considered critical, above which significant P losses in subsurface runoff are expected to occur ŽUunk, 1991; Breeuwsma et al., 1995.. Phosphorus leaching is more prevalent in certain soils than in others. Considerably greater P leaching losses occur from organic and sandy soils than fine textured mineral soils, possibly because of the more-significant contents of constituents responsible for P sorption Žclays, carbonates and

sesquioxides. in the latter soils ŽSims et al., 1998.. Similarly, large P leaching losses have also been reported to occur in soils where a significant proportion of subsurface runoff occurs through macro-pores ŽStamm et al., 1998; Hooda et al., 1999.. As with P loss in surface runoff, large concentrationsrquantities of P have also been measured in the subsurface-flow following cattle slurry applications to soils when they were at or around field capacity moisture content ŽJordon and Smith 1985.. Hooda et al. Ž1999. measured P leaching losses as high as 5.05 kg TP hay1 from intensively managed pastures ŽTable 4., but concluded that ) 30% of these are incidental losses, which occurred when rain andror snowmelt followed cattle slurry spreading. In contrast to surface runoff, P leaching occurs predominantly in the form of dissolved P ŽHeckrath et al., 1995; Chardon et al., 1997; Beauchemin et al., 1998; Hooda et al., 1999.. It appears that P losses in runoff Žboth surface and subsurface . from livestock farming areas will generally not exceed 2 kg TP hay1 yeary1 ŽTables 3 and 4., but could be much greater from specific fieldsrcatchments or with conditions such as incidental losses. These may be small losses from an agronomy point of view and not of economic importance to farmers, generally accounting for - 5% of applied P, but they contribute to the degradation of surface waters by accelerated eutrophication ŽSharpley and Withers, 1994. and, therefore, are of environmental importance.

3. Organic effluents Organic wastes generally contain a large proportion of solids, which can rapidly blanket benthic habitats, with consequential changes in faunal species composition ŽCooper, 1993.. Organic waste may find its way into water courses either by direct discharge from slurry or silage effluent storage or runoff from slurryrmanure applied fields ŽFig. 1.. Organic contamination causes rapid growth of micro-organisms in water, resulting in a high biochemical oxygen demand ŽBOD., and as a consequence, the concentration of dissolved oxygen in water and sediments falls well below the

P.S. Hooda et al. r The Science of the Total En¨ ironment 250 (2000) 143᎐167

155

Table 5 Ranges of BOD concentrations for various wastes a Source

BOD Žmg ly1 .

Silage effluents Pig slurry Cattle slurry Liquid effluents draining from slurry stores Dilute dairy parlour and yard washing Ždirty water. Milk Untreated domestic sewage Treated domestic sewage Clean river water

30 000᎐80 000 20 000᎐30 000 10 000᎐20 000 1000᎐12 000 1000᎐5000 140 000 300᎐00 20᎐60 -5

a

Source: MAFF Ž1998..

levels that can support all forms of aquatic life. In different effluents ŽTable 5., BOD varies depending on how readily they can be metabolised by micro-organisms, e.g. dairy wastes and silage effluents, with large contents of readily-metabolised components, have much higher BOD values than livestock slurries. Thus, livestock waste runoff or direct discharge of farm effluents with large BOD values are capable of causing fish kills or severe disruptions in aquatic ecosystems by rapidly depleting the dissolved oxygen in water bodies. Khaleel et al. Ž1980. reviewed the results of a number of farm waste studies, and found that the BOD of runoff water is strongly influenced by livestock farming activities. The BOD of runoff water from pastures and rangelands varied between 2 and 265 mg ly1 whereas that of runoff from feedlots varied between 9 and 3450 mg ly1 . The review showed that large BOD values were associated with very high stocking densities or the direct discharge of farm effluents into the streams. In Northern Ireland, Foy and Kirk Ž1995. found that in 21 agricultural catchments, dairy cow stocking density was significantly correlated with streamwater BOD Ž1.8᎐73.5 mg ly1 .. They further noted that larger BOD peaks were caused by discharges of silage effluents whereas smaller BOD peaks occurred due to the loss of animal slurry in runoff from slurry-applied fields. Recently Hooda et al. Ž1998b. provided additional evidence of the impact of land use and farm management on streamwater BOD, where mixed land use Žarable, poultry, beef, sheep, and forestry. catchments gave values - 5 mg ly1 Ž0.9᎐4.3. while

predominantly dairy farming catchments varied from - 1 to ) 42 mg ly1 . They concluded that while smaller BOD values were a characteristic of the less intensive farming systems in the mixed land use catchments, larger BOD values were attributed to direct discharges of farm effluents Že.g. silage effluents, milk parlour washings, midden drainage. in the more intensive dairy farming catchments. Rainfall events characteristically govern the BOD of streams draining livestock areas. Generally, when there is no rainfall, the BOD remains within the ranges for natural waters unless farm effluents are discharged directly into streams, and here, the streams may have a high BOD even in the absence of any rainfall event ŽHooda et al., 2000.. Intensive monitoring of two farm streams in south-west Wales, showed that discharges were discrete events with BOD levels rising fivefold, and the timing of these BOD peaks coincided with direct discharges of yard and parlour washings into the watercourses ŽSchofield et al., 1990.. In this study and other similar work Že.g. Hooda et al., 2000., water pollution was also reflected in the status of benthic macrovertebrate communities in the streams, which was poor, as organic pollution-sensitive species were significantly less abundant down-stream from the input. The numbers of reported farm waste pollution incidents in the UK have been increasing. The pollution incidents reported in Scotland rose from 310 in 1984 to 539 in 1993 ŽScottish Farm Waste Liaison Group, 1994.. Similarly, those reported in England and Wales have increased from 2367 in

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1981 to 4141 in 1988 ŽWAA, 1989.. The majority of these incidents were caused by accidental discharge of livestock slurry andror silage effluents to the receiving water ŽWAA, 1989; SOAFD, 1992; MAFF, 1998.. Both slurry and silage effluents have a very high polluting potential, with BOD ranging from 10 000 to 80 000 mg ly1 ŽTable 5., and with ammonia ŽNH 3 ᎐N. ranging from 200 to 1200 mg ly1 . Typical NH 3 ᎐N concentrations recorded in streams following typical farm waste pollution incidents varied from 2 to 700 mg ly1 , and the dissolved oxygen concentration decreased to as low as 0.7 mg ly1 , the detrimental effects of these events on aquatic fauna being demonstrated in four-field simulations, carried out in two streams in south Wales ŽMcCahon et al., 1991.. During each simulation Ž6᎐24 h in duration. the response of several invertebrate species in the downstream-polluted zoneŽs. was compared with that of the species maintained in an upstream-unpolluted reference zone. The results showed that significant invertebrate mortalities were observed only under conditions of reduced dissolved oxygen. However, dosing with NH 3 caused significant reductions in the feeding rate of Gammarus pulex but it recovered during the post-exposure period.

4. Sheep-dipping chemicals Sheep suffer from a number of external insect parasites and immersing the animal in a bath of pesticide solution Ž‘sheep-dip’. controls these parasites; such pesticides generally contain either an organophosphorus or synthetic pyrethroid compound as the active ingredient. Whilst no one could argue the benefits of sheep dipping in terms of animal health and productivity, the chemicals are known to have deleterious effects on water quality ŽCoddington, 1992.. In dipping practice, these chemicals may find their way into the water courses primarily through the disposal of spent dip baths, the draining off of the animal, or dip leaking directly into watercourses ŽFig. 1.. A water quality survey of the River Tweed catchments in Scotland showed the presence of sheep dip chemicals in 17 out of 20 catchments sampled

during the 1989 dipping season ŽVirtue, 1992.. A follow-up survey involving 1302 sheep farms carried out during the 1990 dipping season confirmed the presence of diazinon Ž10᎐150 ng ly1 . and propetamphos Ž70᎐400 ng ly1 . in surface waters throughout the area as a whole ŽCurrie, 1992.. Eight serious pollution incidents were also observed, with dip active ingredient concentrations in stream water recorded above 1 ␮g ly1 , and in one case the concentration of propetamphos in the water was more than 1 mg ly1 . Similarly, in the Grampian region of Scotland, organochloride insecticide analysis of samples taken during 1984᎐1985 from the River Ugie and its tributaries showed, by the timing and the nature of the compounds found, that sheep-dips were likely sources of pollution of freshwaters in the region ŽLittlejohn and Melvin, 1991.. However, the analysis showed no evidence of serious contamination of the waters by phenolic compounds that are present at high concentrations in sheep-dip fluids. The contamination of watercourses with sheepdip chemicals can have adverse effects on the aquatic ecosystem. It has been shown that these chemicals cause pollution as evidenced by fish kills and reduction in stream biota ŽCoddington, 1992; Currie, 1992; Virtue and Clayton, 1997.. A majority of farmers in the UK dispose of the spent dip either to a soakaway or direct disposal on the land close to the dipper without spreading. This practice, however, is not considered environmentally safe. For example, Littlejohn Ž1992. monitoring the disposal of dip solution from a dip tank to a soakaway, situated 300᎐400 m from the nearest stream, found that dip-chemicals appeared in the stream within 2 h of the dip solution being released. As stipulated in the Codes of Good Practice, the soakaway disposal method for spent sheep-dip is no longer an acceptable practice ŽSOAFD, 1992; MAFF, 1998..

5. Bacterial and protozoan pathogens The microbiological quality of surface waters helps to determine their acceptability for both drinking and recreational purposes. Livestock

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157

Fig. 2. Pathways of catchment water contamination with microbial and protozoan micro-organisms.

waste, particularly untreated slurry and faeces of grazing animals can carry a variety of bacterial and protozoan pathogens. Contamination of water with slurry andror faeces, which can occur through a variety of pathways ŽFig. 2., therefore represents a potential health risk to human and animals; such a risk potential will depend on a number of factors, such as animal health Žseverity of infection., slurry treatment and storage period, time and method of slurry application, stocking density of grazing animals and survival time of the organisms. Infected animals can excrete pathogens, which may survive prolonged anaerobic storage of slurry ŽRankin and Taylor, 1971; Kearney et al., 1993.. Land spreading of such slurry can cause water pollution either through direct contamination with slurry aerosols or through runoff from slurry applied fields ŽFig. 2.. Although not widely recognised, the use of raw sewage in farming systems in some developing countries Že.g. China, India, Thailand, Indonesia. can also contaminate water with bacterial and protozoan pathogens ŽAdhikari et al., 1997; Wang, 1997.. Faecal contamination has been reported in streams draining dairy farms ŽJanzen et al., 1974., subsurface runoff from manure-applied fields ŽCulley and Philips, 1982., sur-

face runoff from grazed pastures ŽDoran and Linn, 1979. and poultry litter treated grass fields ŽGiddens and Barnett, 1980.. Bacterial contamination of runoff water has traditionally been assessed using counts of selected bacterial indicators, such as total coliforms ŽTC., faecal coliforms ŽFC., faecal streptococci ŽFS. or enterococci. Livestock grazing activities have been found to increase bacterial counts in runoff water. For example, indicator bacterial densities in streamwater were significantly higher when at least 150 cattle were grazing, but bacterial counts dropped to levels similar to those adjacent to an ungrazed pasture following the removal of cattle or when only 40 head of cattle were grazing ŽGary et al., 1983.. Similarly cattle grazing increased FC counts which exceeded several-fold the USEPA standard for bacterial contamination of primary contact water Ž200 faecal coliforms 100 mly1 . ŽDoran et al., 1981; Howell et al., 1995.. The effect of grazing on bacteriological quality of runoff persisted for more than 1 year after animals were removed from a pasture in the Pacific Northwest ŽJawson et al., 1982.. Niemi and Niemi Ž1991. studied the occurrence of Escherichia coli and faecal streptococci in ditches, brooks and natural ponds in six agricultural and 22 uninhab-

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ited pristine areas in southern Finland. While their findings were consistent with the previously-discussed studies, faecal indicators in approximately half of the samples of pristine areas were also detected, presumably due to contamination by wild animals living in the area ŽFig. 2.. There are a few studies that have information on water contamination by specific pathogens and only a few important ones are briefly discussed here to illustrate the major problems that can arise and to suggest appropriate control measures. 5.1. Cryptosporidium This protozoan parasite, found in surface waters is thought to originate from both sewage sludge and livestock waste. In extensive surveys of surface waters in the USA, Cryptosporidium oocysts have been detected in approximately 75% of lakes, streams and rivers examined ŽRose, 1989.. Cryptosporidium oocysts are resistant to many disinfectants used in routine chemical treatment ŽAngus, 1983., and they may remain viable in water for at least 140 days ŽCurrent, 1986.. A number of different species have been identified from different animals but it is Cryptosporidium par¨ um that is generally believed to be one of the most concern to health as the causative agent of cryptosporidiosis. In the 1980s, cryptosporidiosis was recognised as a common cause of acute selflimiting gastroenteritis, and has been reported in humans from all over the world, with prevalence rates of 0.6᎐20% in developed countries and 4᎐32% in underdeveloped countries ŽSoave, 1989.. Recent studies show that cryptosporidial infections are prevalent in cattle, pig, and sheep ŽQuilez et al., 1996; Olson et al., 1997.. Outbreaks of cryptosporidiosis in Sheffield, England ŽBarbara et al., 1990. and Ayrshire in Scotland ŽAnon, 1990. were attributed to slurry spreading or other cattle activities. Svoboda et al. Ž1997. conducted a major study of C. par¨ um oocysts on cattle farms in England and Scotland and found that - 8% of oocysts applied on the bedding leached out, with their viability decreasing rapidly during composting of the bedding material, and that their sur-

vival in the field was limited either because of desiccation or freezing. Land spreading of wellcomposted manure, therefore, presents only a small risk of runoff water being contaminated with C. par¨ um. However, not only has the organism survived for more than 21 days when Cryptosporidium-contaminated slurry was applied but also the oocysts showed downward migration through the soil profile, with potential transfer to water with subsurface runoff ŽMawdsley et al., 1996.. Sturdee et al. Ž1998. carrying out extensive studies of C. par¨ um over a 17-month period on a well-managed farm estate in Warwickshire, showed that over 70% of all surface water sampled during that period contained C. par¨ um oocysts, although they were not able to test their viability. Their study did not show any consistent correlation between the numbers of oocysts in the streams on the estate and slurry spreading or grazing in adjacent fields, but there was some correlation with heavy rainfall and elevated numbers of oocysts in the water courses. Since this was a well-managed model farm using current best practices, it is unsurprising that slurry spreading and grazing were not found to result in elevated levels in adjacent water, but one of the principal findings to emerge was that accidental leakage into drainage ditches, from farmyard runoff or damaged drains, led to an important route of contamination of streams with fresh viable oocysts. All such runoff should be directed to a slurry lagoon, but even on a well-managed farm it is possible for apparently small but significant amounts of material to get into drainage ditches instead. This is a particular problem where young livestock are involved, such as young calves, which show elevated levels of infection and shed more oocysts than older livestock groups Ž10 4 oocysts gy1 faeces.. The implication here is that particular care is needed in maintaining the integrity of drains around intensive calving units, to ensure that all slurry and washing are taken to the slurry lagoon. Analysis of slurry prior to spreading showed that many oocysts were being spread into the fields ŽSturdee et al., 1998.. However, work by Ruxton Ž1995. demonstrated that a combination of the high concentration of ammonia and preda-

P.S. Hooda et al. r The Science of the Total En¨ ironment 250 (2000) 143᎐167

tion of pathogens by other organisms in the slurry should destroy most of the oocysts or ensure that they are non-viable by the time of application to the fields. The other major issue to emerge is the significance of wild animals as reservoirs of the parasite. Studies by Chalmers et al. Ž1995, 1997., Webster and MacDonald Ž1995. and Sturdee et al. Ž1998. showed that C. par¨ um is ubiquitous in UK wild mammal populations, and highly prevalent in rodents, and up to one-fifth of the oocysts in agricultural drainage could come from wildlife sources. The implications for water quality and for cross-infection potential between other species are such that the UK government are now introducing a new drinking water standard, limiting the concentrations of Cryptosporidium par¨ um to 0.1 oocysts ly1 . Because it is difficult to destroy the oocysts by conventional chlorination treatment, there is likely to be an increased emphasis on measures to limit the number of oocysts getting into watercourses in the future. Overall the risk of water contamination with C. par¨ um remains high, particularly in situations where rains follows Žcontaminated. slurry spreading or where infected animals have direct access to streams or other water bodies. 5.2. Giardia This is another highly potential protozoan parasite that can cause water-borne diarrhoreal infections to both man and animals Žlivestock and wild populations alike.. It is now considered as the leading water-borne parasitic disease in the USA, and thought to be one of the most common intestinal diseases world-wide ŽBemrick and Erlandsen, 1988.. The organism infects the small intestine and is excreted in large numbers, as small cysts, during an infection. The infection is more prevalent in children and young animals than in adults. As with Cryptosporidium, Giardia infection is also prevalent among young farmed animals Že.g. cattle, pigs, sheep, horses. and very often the infections are concurrent ŽXiao et al., 1993; Quilez et al., 1996; Olson et al., 1997.. Giardia labbila cysts have survived up to 33 days in animal waste and 47 days in water ŽSnowdon et al., 1987.. Runoff from infected-waste applied

159

fields can therefore contaminate fresh waters with this organism. In a survey of sheep and cattle in Canada, Buret et al. Ž1990. found that 35.7% of suckling lambs and 22.7% of calves faecal samples showed the presence of cysts. No incidence of cysts was found in adult cattle, but 4.1% of the adult sheep did excrete cysts. Similarly, Giardia spp. has been reported in cattle in Switzerland ŽGasser et al., 1987.. Like C. par¨ um, Giardia cysts are hard to destroy with conventional water disinfection treatments, and prevention of their spread is therefore essential. 5.3. Salmonellae These have long been recognised as a cause of enteric infections in man and many animals and can survive for up to 1 year in anaerobically stored slurry. In a study of 187 samples of cattle slurry, 11% were found to contain Salmonella spp. ŽJones and Mathews, 1975.. Similar studies of poultry houses show a high incidence of Salmonellae infection in poultry, with a high degree of litter contamination ŽOpara et al., 1992; Carr et al., 1995.. After land application of infected excreta, organisms have survived for less than 3 weeks on grass leaves ŽTaylor and Burrows, 1971.. Other reports, however, suggest that this organism can survive from 3 days to 36 months in faeces, and from 5 to 968 days in soil ŽJones, 1986.. This means water pollution can occur months after spreading contaminated slurryr manure. Jack and Hepper Ž1969., studying an outbreak of Salmonella typhimurium infection in cattle at a farm in South Devon isolated the organisms from a slurry spraying system and from four carrier cows. They concluded that the spread of the disease in the herd was caused by irrigation with infected slurry. Salmonella infection among other livestock species is also frequently reported. According to one report, more than half of the broilers and more than one-third of pigs sent for slaughter are infected with Salmonella spp. ŽStrauch, 1986.. High infection incidence among livestock and its ability to survive in the field over longer periods of time greatly increases the risk of water contamination with Salmonella spp. All of these examples of microbial pathogens

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demonstrate that either some form of slurry treatment similar to that undertaken in the biological treatment phase of sewage sludge prior to land spreading andror a longer storage period will minimise the risk of spreading of slurry-borne pathogens both directly to water courses and to potential ‘reservoirs’ of wild animal populations.

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6. Measures to reduce water pollution Point sources of pollution are easy to identify and can effectively be controlled by following Codes of Good Agricultural Practices, laid down for the protection of water ŽSOAFD, 1992; MAFF, 1998.. However, the diffuse sources of pollution, such as loss of nutrients through leaching and in runoff, are more important and serious in that they are difficult to assess and control. Solutions to the problem of diffuse sources of pollution require a sound agricultural policy of sustainable farming systems, and centre around best management practices ŽBMPs.. The policy should outline that land be used only to support agricultural production without abuse ŽCooper, 1993.. Some of the established BMPs are discussed below.

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6.1. Control of fertiliser and li¨ estock waste in surface runoff In addition to inorganic fertilisers, regular land application of manures and slurries in livestock farming has increased soil N and P to levels exceeding plant requirements ŽKing et al., 1990; Sharpley et al., 1993.. Surface runoff from such areas then becomes enriched in nutrients and organic matter. However, by adopting the following BMPs, loss of nutrients and other pollutants in surface runoff can be minimised.

6.2. Minimising leaching potential

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The nutrient content of manures and slurries is highly variable, and should be determined particularly for N and P, prior to any land application, and then used for deciding their rate of application. Commercially available kits or computer packages Že.g. Agros meter, MANNER. can also be used for on-farm nutrient estimations in manures ŽChambers et

al., 1999.. Rates of manure or slurry application should not supply nutrients Že.g. N, P. in excess of the crop requirements and any additional inorganic fertiliser should be applied at rates, which allow for the nutrients supplied through the use of manures. Livestock waste should not be applied to saturated or water-logged, snow-covered, frozen and steep sloping grounds because this may cause high losses of N and P in surface runoff. For example, losses of up to 20 and 17%, respectively, of applied N and P in early spring runoff have been reported when manure was applied to frozen ground ŽYoung and Mutchlen, 1976.. Surface-applied livestock slurry has greater potential for nutrient and organic matter loss in runoff compared with that incorporated into the soil, and farmers should be encouraged to use modern slurry application techniques, including direct injection. Diluting the slurry with water prior to the conventional surface application can also help in minimising its loss in runoff by facilitating its penetration into the dry soil. Vegetative filter strips or ‘buffer strips’ are an effective BMP for removal of pollutants, particularly particulate N, P, organic matter and other sediments, from runoff water. Field research has proven that, the use of buffer strips helps to minimise agricultural pollutant inputs into water bodies adjacent to grasslands and other animal confinements, such as feedlots ŽDillaha et al., 1988; Clausen and Meals, 1989..

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As for reducing surface runoff losses, a rational use of fertilisers and livestock waste is necessary for controlling excessive nutrient accumulation in soils, which, thereby will help in reducing transport of N and P by leaching. Because P has low solubility in soils, measures to control nutrient leaching are often directed toward N fertiliser management. Most drainage from agricultural land in most

P.S. Hooda et al. r The Science of the Total En¨ ironment 250 (2000) 143᎐167

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north-western European and North American regions occurs during autumn and winter. Any nitrate remaining in the soil profile after the end of growing season Žwhich coincides with the beginning of the leaching season. is susceptible to leaching because of little opportunity for nitrate removal by plant uptake. Solutions to nitrate leaching problems, therefore, need to be based on minimising the amount of nitrate available for leaching. Applying fertilisers based on soil testing and strictly according to the plant requirement, and evolving strategies that minimise the residence time of nitrate in soil Ži.e. timing the application with plant uptake. can help mitigate the nitrate leaching problem. For instance, application of N based on soil testing compared with a conventional application of 210 kg N hay1 was found to decrease the potentially leachable N by over 30% ŽTitchen and Scholefield, 1992.. An alternative to using inorganic N fertiliser in grassland is the use of mixed swards, i.e. a mixture of grass and perennial legumes, such as clover Ž Trifolium spp. and lucerne Ž Medicago sati¨ a L... These mixed pastures, where the principal source of N is through biological fixation, offer a means of reducing nitrate-leaching losses. This was demonstrated in a recent Ohio study where beef cattle grazed orchard grass Ž Dactylis glomerata L.. and tall fescue Ž Festuca arundinacea Scherb.. pastures, which were fertilised with 224 kg N hay1 yeary1 for 5 years. In the beginning of the 6th year, lucerne was interseeded into the pastures and fertiliser application discontinued ŽOwens et al., 1994.. Nitrate concentrations in groundwater collected from field drains decreased, on average, by 58% during the following 2-year period before reaching prefertilisation levels over the next 8 years. This environmentally-benign nature of legumebased pastures was further supported by recent findings of significantly less nitrate leaching from grass᎐clover mix pastures than from mineral fertilised pastures ŽWatson et al., 1992; Kristensen et al., 1995; Hooda et al., 1998a., and in addition provides a safeguard against

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incidental N waste such as fertiliser loss in rainwater. As with inorganic fertiliser, the timing of slurry or manure applications is very important for most efficient nutrient use, hence minimising their leaching potential. Application in August᎐September resulted in smaller nitrate losses compared with October᎐November treatments ŽFroment et al., 1992; Smith and Chambers, 1993., therefore, particularly under north-west European and North American climatic conditions, applications of manures and slurries to land after September should be avoided. It is also important that slurry should not be applied in winter months, as large losses of slurry effluents through field drains haven been recorded from such applications ŽJordon and Smith 1985; Hooda et al., 1999.. The significance of the time of slurry application was clearly demonstrated by the latter authors, who noted that up to 30% of P leaching losses through field drains occurred following a slurry application in winter when the soils were close to field capacity moisture content. Ideally slurryrmanure applications should be made in spring and summer months for effective N use efficiency and hence much reduced leaching potential ŽBailey, 1993.. The use of nitrification inhibitors Že.g. dicyandiamide., in conjunction with manure and fertiliser, is likely to be limited by the cost factor, but the potential benefit of their use is considerable because they reduce nitrate leaching by controlling the process of nitrification ŽWadman et al., 1989; Froment et al., 1992.. Treating manures with P-binding materialr chemicals before their land spreading offers a means of reducing P solubility and hence P loss to water. It has been shown that treating poultry manure with alum and ferrous sulfate reduced the P concentration in runoff by 87 and 63%, respectively ŽShreve et al., 1995., while other chemicals Že.g. slaked lime, quick lime, ferrous chloride and ferric chloride. have produced similar effects ŽMoore and Miller, 1994.. However, while similar strategies can be adopted with other type of livestock waste

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Že.g. slurry., the effects of their long-term use on P availability and soil quality are not yet known. 䢇

6.3. Control of sheep-dipping chemicals, bacterialr protozoan organisms in runoff

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Soakaways, commonly used for the disposal of used sheep-dip, are no longer considered safe means of disposal as they represent a high risk to ground and surface water. To prevent or mitigate contamination of ground and surface waters with sheep-dip chemicals, spent-dip should be diluted with three parts of slurry or water prior to land spreading away from drains and farm streams ŽMAFF, 1998.. To minimise the risk of parasite organisms escaping into surface water all farm building drains need to be regularly maintained and drainage from calving units should not be allowed to mix with other farmyard drainage. Many examples in the literature of outbreaks of disease are traced to such accidental leakage ŽBetts et al., 1995.. The control of wild rodents’ population, particularly mice, in the vicinity of intensive animal rearing facilities, especially calving pens, during autumn and winter will reduce the possibility of cross-infection and also reduce transport of parasitic organisms such as Cryptosporidium and Giardia into watercourses. Slurry and manure spreading should only be undertaken when the material has been stored for a sufficient time for potential pathogens in it to be destroyed or reduced to a non-viable state by composting. This may mean holding material for several months before spreading and then applying it only when there is no likelihood of precipitation. Similarly the application of dirty water to land presents a high risk of spreading bacterialrparasitic organisms. Therefore, dirty water should be applied only to fields with the least risk of causing water pollution ŽMAFF, 1998.. Similarly limiting moisture content in the litter base of poultry houses may provide a less favourable envi-

ronment for the multiplication of Salmonella and thus should help reduce the risk of water contamination following land spreading of poultry litter ŽCarr et al., 1995.. The use of slurry injectors for subsurface application and the avoidance of slurry spreading on saturated, frozen and steeply sloping soils recommended in Section 6.2 above will also help to limit the escape of any remaining pathogenic organisms into surface water bodies. Combined with the use of buffer strips adjacent to streams and rivers to provide a barrierrfilter in which pathogens are likely to be retained and then degraded by prolonged exposure to the external environment. The important control measure is to prevent them getting into the water directly or other organisms where they can again proliferate and become source reservoirs for infection.

Acknowledgements One of us ŽPSH. would like to thank Dr I. Svoboda of the Scottish Agricultural College, Ayr for several useful discussions on some of the issues covered in this paper. The Scottish Office Agriculture, Environment and Fisheries Department funded the initial part of this work while PSH was based at the Macaulay Land Use Research Institute, Aberdeen. Oxford Brookes University provided further funds for the concluding part of the work. References ADAS. Nitrogen for grassland. London: Agricultural Development Advisory Service, Ministry of Agriculture Fisheries & Food, 1990. Adhikari S, Gupta SK, Banerjee SK. Long-term effect of raw sewage application on the chemical composition of ground water. J Indian Soc Soil Sci 1997;45:392᎐394. Angle JS, Gross CM, Hill RL, McIntosh MS. Soil nitrate concentrations under corn as affected by tillage, manure and fertiliser applications. J Environ Qual 1993;22:141᎐147. Angus KW. Cryptosporidiosis in man, domestic animals and birds: a review. J R Soc Med 1983;76:62᎐70. Anon. Cryptosporidium in water supplies: report of the group of experts. London: HMSO, 1990.

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