Ammonia emissions from intensive dairying: a comparison of contrasting systems in the United Kingdom and New Zealand

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1 Agriculture, Ecosystems and Environment 92 (2002) Ammonia emissions from intensive dairying: a comparison of contrasting systems in the United Kingdom and New Zealand S.C. Jarvis a,, S. Ledgard b a Institute of Grassland and Environmental Research, North Wyke Research Station, Okehampton EX20 2SB, UK b AgResearch Ruakura Research Centre, Private Bag 3123, Hamilton, New Zealand Received 5 June 2000; received in revised form 26 February 2001; accepted 25 June 2001 Abstract Much of the increased atmospheric concentration of ammonia (NH 3 ) is derived from dairy farms. In order to better understand the potential for controlling emissions, a comparison was made of NH 3 emissions from dairy farms in the United Kingdom and New Zealand using recently derived emission factors. This desk study demonstrated distinct differences between the two farms: total N input, off-take and surplus in the UK were 1.7, 1.2 and 1.8 times greater than in New Zealand. More sources of NH 3 were identified in the UK which were, in the main, associated with the housing phase, and therefore manure generation. Total NH 3 losses were equivalent to 57 and 24 kg N ha 1 in the UK and New Zealand farms, respectively: these represented similar proportions of the total N inputs to both systems but because of stocking density in New Zealand and its overall greater milk production per ha, losses of NH 3 whether expressed per LU or unit of milk, were at least two times greater in the UK. Options for controlling emissions and their impact included improved slurry storage and application methods, efficient grazing patterns and increased N fertilizer use efficiency in the UK. Options were fewer in New Zealand, the most practical being more effective fertilizer application. In both cases, farming practices which increase N use efficiency, or improve C:N dietary balance for highly productive cows have much potential for reducing volatilization. The approach taken illustrated some important differences between the two farms and provided an easily adopted method to make assessments of NH 3 loss. Additionally, a set of indicators was derived to allow quantifiable comparisons Elsevier Science B.V. All rights reserved. Keywords: Ammonia; Dairy; Farming systems; Indicators; Mitigation options; N managements; New Zealand; UK; Volatilisation 1. Introduction There is much interest in the extent of ammonia (NH 3 ) emissions from animal production systems (Jarvis and Pain, 1990). It is accepted that intensive animal management has been responsible for increases in atmospheric NH 3 /ammonium (NH 4 + ) concentrations that have occurred (Pain et al., 1998). This Corresponding author. Tel.: ; fax: address: steve.jarvis@bbsrc.ac.uk (S.C. Jarvis). is of concern because of the environmental impact that NH 3 has, both in the atmosphere (as an important control over atmospheric chemistry; Buijsman et al., 1998), and after deposition (Duyzer, 1994) where it influences nitrogen (N) cycling in natural ecosystems and changes ecological balance. Because of rapid reactions with other atmospheric species (Singles et al., 1998), some of the effects of NH 3 deposition can be immediate and widespread. As well as having environmental effects, NH 3 volatilization provides a substantial N loss from grassland based farms. One of the most important farming /02/$ see front matter 2002 Elsevier Science B.V. All rights reserved. PII: S (01)

2 84 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) types is dairying which has a number of emission sources, viz. grazing, fertilizer application, housed animals, and stored and applied excreta (Sommer and Hutchings, 1997). Earlier studies (Jarvis, 1993) indicated that ca. 20% of fertilizer N input to a UK dairy farm was volatilized and that, by using existing management technologies, it is possible to reduce emissions by ca. 35% (Jarvis et al., 1996). It is important to consider mitigation options because of developing national and international protocols (Lekkerkerk, 1998) to reduce emissions. Much information on NH 3 has been accrued (see Stevens and Laughlin, 1997) and this has been used to produce, for example, national inventories (Pain et al., 1998) to assist policy makers. There is also a need to provide information at practical levels to test management options and farm system scales provides a good basis for this. Volatilization has well defined physico-chemical controls, but interaction of these with the interactive climatic, managerial and physical components of all phases of production make prediction difficult. Approaches taken in the past have used either a combination of emission factors and model prediction (Jarvis et al., 1996), or a mechanistic modeling basis (Hutchings et al., 1996). In the present study, the pragmatic approach was taken of using the most recently available emission data to quantify net NH 3 loss from two similar sized dairy farms under intensively managed grassland with contrasting management practices and conditions; namely one in south-west (SW) England, UK and the other in North Island, New Zealand. The objectives were: (i) to further develop a simple methodology with quantifiable indicators for NH 3 emissions using improved and additional emission factors (EFs); (ii) to compare two different dairy farms in the UK and New Zealand, and (iii) to investigate the impact of mitigation options on the two farms. 2. Materials and methods 2.1. Farm description Two model farming systems were defined from published survey and statistical information following methods used in previous systems analyses of dairy farm N flows (Jarvis, 1993; Jarvis et al., 1996). Although the information for the UK farm is based on 1991 information and there will have been some changes in the statistics, this was used because it allows comparison with a previous analysis and also provides a direct comparison with a comparable area in New Zealand. Each farm (Table 1) represents a typical, average management appropriate to the local soil, environment and economic conditions to provide the basis for the comparison. The UK farm is situated in SW England (rainfall >1000 mm per year) on moderate to poorly drained soils. The farm follows current good guidelines for fertilizers (e.g. RB209, MAFF, 1994) and produces, with Table 1 Production and other characteristics for model dairy farms for examination of ammonia emissions in the UK and New Zealand a UK New Zealand Soil type/sward 25 ha: poorly drained clay loam (dystric gleysol): long term ryegrass swards 76 ha: well drained silt loam (umbric dystrochept): long term ryegrass/white clover swards 51 ha: moderately drained loam (dystrochept): reseeded ryegrass swards Animals 102 Friesian cows yielding 5554 l per cow 202 Friesian cows yielding 3348 l per cow (all other non-milking cattle grazed off-farm) 110 Other cattle (followers and beef calves) Total LU = 165 Total LU = 202 Purchased feeds t DM (concentrates and other supplements) 28 t DM (grass silage) Forages 393 t DM ensiled grass (26.5% DM) 37 t DM ensiled grass (18.5% DM) 574 t DM grazed grass 776 t DM grazed grass Purchased bedding 45.6 t DM Wastes 1693 m 3 slurry 670 m 3 dirty water 2340 m 3 dirty water a Appropriate values are expressed on an annual basis.

3 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) kg N fertilizer ha 1 per year, sufficient grazed and ensiled grass to maintain 102 milking cows and a total of 164 livestock units (LU) (including followers and beef cattle, where 1 LU = 500 kg). Concentrates are purchased to supplement forage and winter bedding is also purchased. Manures are collected, stored in an open lagoon and spread as slurry; dirty water from milking parlour washings is collected and stored in a circular tank. Amounts of both materials are calculated from published data (MAFF, 1992). The model New Zealand farm is located in the main dairying region of Waikato in the North Island (rainfall 1200 mm per year) on freely draining soil derived from volcanic ash. Annual pasture dry matter (DM) production of approximately 14 t DM ha 1 is achieved with N inputs by N 2 fixation by white clover (Trifolium repens L.) but with strategic application of urea fertilizer (Butler and Johnson, 1997). This pasture supports an average of 2.7 cows ha 1 (202 LU) which graze throughout the year: all non-milking animals (9 22-month-old heifers) are raised off the farm. Some supplementary feed (mainly grass or maize silage) is purchased to meet autumn/winter feed deficits. Cows are milked for 227 days between early spring and late autumn to match feed requirements with the pattern of pasture growth. Excreta are not stored in New Zealand, but washings from dairy sheds and associated yards are collected and applied to 12 15% of the area at <25 mm per occasion to ensure that annual N returns are less than the local required rate of 150 kg ha 1 per year (Environment Waikato, 1997). Total N inputs (Table 2) to the UK farm are ca. 1.7, off-takes 1.2, Table 2 Farm gate nitrogen balances and surpluses on model UK a and New Zealand b dairy farms UK Total (t) kg ha 1 New Zealand Total (t) kg ha 1 Annual N input Fertilizer Atmospheric deposition N 2 fixation by clover Feeds/bedding Annual removals c Annual surplus a Infromation taken from Jarvis (1993). b Ledgard (unpublished). c In milk and meat. and surplus 1.8 greater than in the New Zealand system Farm ammonia sources The major source of NH 3 within any cattle production system is urea, especially that excreted in urine: it may also be applied as fertilizer. The first stage in NH 3 volatilization is conversion from urea to inorganic form by urease which is present universally within all sectors of dairy systems. Fluxes of NH 3 from the source are then dependent upon environmental and soil (especially ph and moisture) conditions, plant growth rate (as an important competitor removing NH 4 + from the system) and farm managerial actions and activities (Table 3). It is clear that the UK farm has more sources. Previous studies (Sommer and Hutchings, 1997) showed that the largest losses from livestock farms are from excreta deposited, stored and applied as a result of winter housing, a major component of the UK farm. Each component of the two farms and the local influences on volatilization are discussed below Grazed swards The UK swards used for grazing are predominantly perennial ryegrass (Lolium perenne L.), although some clover is present (contributing 10 kg N ha 1 ). The swards receive, on average, 250 kg N ha 1 as NH 4 NO 3 per year: direct emissions from fertilizer are included in the EF for the grazed sward (Table 4). A proportion of the swards is reseeded periodically to maintain sward quality and all grazed swards are rotationally grazed. All animals graze for 185 days from April (spring) to September (autumn). Milking cows spend 2 h each day at milking with a proportional decrease in urine and dung deposited in the field. On the New Zealand farm, the swards are permanent pasture with perennial ryegrass and white clover, which constitutes 17% of DM production and fixes ca. 140 kg N ha 1 per year. Pastures are rotationally grazed with urea added at 53 kg N ha 1 per year as split applications in autumn and early spring to meet feed shortages. On average, 15% of the pasture is ensiled and fed to cows in the field in autumn to remove identified feed surpluses and to maintain pasture quality.

4 86 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) Table 3 Key phases in UK and New Zealand dairy farm production systems providing opportunity for ammonia emissions S. no. Source Controlling factors/influences UK New Zealand 1 Grazed swards Total N inputs (kg ha 1 ) Fertilizer N (kg ha 1 ) 250 (NH 4 NO 3 ) 53 (urea) Grazing days Cut swards Fertilizer N (kg ha 1 ) Housing (days) Over winter 180 NA a Milking (2 per day) (2 per day) 4 Farm manures Stored slurry Open lagoon NA Applied slurry Vacuum tanker NA 5 Dirty water/dairy shed effluent Stored Circular tank Small pond Applied Irrigation system Irrigation 6 Collection yards (2 per day: NA (included in 5) all year emission) 7 Laneways Included in 1 (2 per day: all year emission) a NA: not applicable or only minor impact Cut fertilized swards The UK farm depends on the production of large amounts of DM (2385 kg per LU) from silage. Areas of the farm will be managed for this and will be harvested in up to three successive cuts followed by grazing the later in the year. There will be a direct loss of NH 3 associated with the NH 4 NO 3 applied to silage areas as split dressings after each cut. At a requirement of t DM, and an average rate of 8 t DM ha 1, 49.2 ha are required to produce silage which receives N fertilizer at the same rate as the grazed swards. An EF for direct losses from fertilizer applied to silage areas on the UK farm is included in Table 4. In the New Zealand farm, all N fertilizer is applied to grazed swards: because urea has a much higher emission rate than NH 4 NO 3 (Van Der Weerden and Jarvis, 1997), this is estimated separately, using a New Zealand specific factor (Table 4). The cut grass may provide a NH 3 source: Whitehead and Lockyer (1989) found significant losses under controlled conditions and more recently, Sutton et al. (1998) found enhanced field losses in spring after an early grass cut. However, within a whole farm context, the amounts are small and have not been considered here Housing losses Excreta, predominantly urine, dropped whilst animals are indoors, provide an important NH 3 source (Pain et al., 1998). The nature of the building determines the pattern and rates of loss, but although there may be some deposition onto wet surfaces, once volatilized, NH 3 will find a pathway to the exterior and contribute to the overall flux. All cattle on the UK farm are housed indoors over winter for 180 days and straw is provided for bedding. In both farms, 2 h of each milking day is assumed to be spent in the milking unit and the impact of this is taken into account. No other time is spent indoors by the New Zealand cows Farm manures The UK farm generates large volumes of manures (calculated on the basis of 57 l per LU per day; MAFF, 1992) which is stored as slurry in an open top lagoon (2.5 m deep with a surface area of ca. 964 m 2 (taken from average depths for lagoons in England and Wales; Nicholson and Brewer (1997); and a required maximum storage capacity of 1782 m 3 ). The lagoon is emptied twice each year, but it is assumed that there is an emitting surface present throughout since frequent inputs ensure that the surface does not dry out. Because straw is used, the slurry has a relatively high DM content (4 8%) containing 0.5% total N and a total NH 4 + -N content (TAN) of 2.25 kg m 3. The slurry is broadcast onto the pasture according to codes of good agricultural practice (MAFF, 1992) by a conventional vacuum pressurized tanker. This has considerable potential for large NH 3 losses (Pain et al., 1998; Stevens

5 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) Table 4 Calculated ammonia emissions from model UK and New Zealand dairy farms a S. no. Emission source Emission factor (EF) Emission 1 Grazed swards c g N per LU per day = (0.0683x) (where x = kg ha 1 input) (+/ 20%) 2 Cut fertilized swards NH 4 NO d 3 (UK) = 1.6% N applied, urea e (New Zealand) = 12% N applied (+/ 54 and 34%, respectively) UK Total (kg) (kg ha 1 ) b New Zealand Total (kg) Slurry Stored 2.1gNm 2 per day surface area Applied 37% of TAN applied (+/ 39 and 7%, respectively) (kg ha 1 ) b 4 Dirty water Stored 0.4gNm 2 per day stored area Applied 15% of TAN applied (no information) Housing Cows 38.5 g N per LU per day f Non-milking 38.5 g N per LU per day f Calves 10.6 g N per LU per day f (no information) Collection yards 8.31 g N per cow per day (no information) Laneways 0.30 kg per cow (no information) g Total a Values are calculated from emission factors from Misselbrook et al. (2000), except where noted, and the activity information in Table 3. Numbers in parentheses refer to estimates of ranges in emission factors (from Misselbrook, unpublished). LU: livestock unit and calculated on the basis of 500 kg per animal. b Calculated mean over whole farm. c Assumes 2 h per day milking. d For silage production only. e For whole farm (New Zealand emission rates). f Includes 2 h per day for milking during grazing for UK: New Zealand figure is for 2 h milking only. g Estimate only: see text for derivation. and Laughlin, 1997). There are no similar farm wastes on the New Zealand farm to provide a source (Table 3) Dirty water/dairy shed effluent Both farms produce dirty water. In the UK this largely results from cleaning the milking parlour (calculated from an average figure of 18 l per cow per day; MAFF, 1992) and contains 0.4 kg TAN m 3 which can volatilize either during storage or on application to land. It is stored in a 3.7 m high circular tank (from Nicholson and Brewer, 1997) (223 m 2 surface area) and applied to land by sprinkler irrigation. The New Zealand dairy shed effluent/water is collected in a small sump/pond with an assumed area of 380 m 2. After each milking, the water, which equates to ca. 50 l per cow per day with an average total N concentration of 0.05% (includes 25% TAN; Ledgard et al., 1996), is applied by sprinkler irrigation onto pasture Collection yards and laneways Collection yards and laneways are important components of UK and New Zealand farms, respectively. Collection yards, although cleaned periodically, provide an emitting surface with new excretal material being deposited twice each day. Laneways provide access routes between milking sheds and paddocks on

6 88 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) New Zealand farms and act as a similar source to collecting yards, typically being an area equivalent to 1.5% of the grazed pasture area. Detailed measurement of cow movement and excretion patterns indicated that up to 4% of total excreta is deposited on laneways (Zegwaard and Ledgard, unpublished). Laneways are occasionally covered with sandy materials and are frequently disturbed by animal movement so that their surfaces generally remain plant-free. It is assumed in the UK calculations that excreta of cows in transit to milking parlours contribute to the NH 3 losses from grazing or from the collection yardways. 3. Results and discussion 3.1. Calculated emissions For the UK, recent studies (Pain et al., 1998; Misselbrook et al., 2000) have identified and produced EFs; in some cases, estimates of uncertainty have been made (Table 4) for these. In the absence of New Zealand specific EFs, those defined by Misselbrook et al. (2000) have been used to provide the basis of comparing the farms (Table 4): exceptions are described below. There may, of course, be differences between UK and New Zealand which will influence the EFs, but because they are averages obtained from a broad range of conditions, they are likely to be generally applicable within similar climatic zones and, in the absence of other data, provide the best current means of investigating system differences Grazed swards Jarvis and Bussink (1990) established a relationship between N input and NH 3 loss per unit area which was modified firstly by Pain et al. (1998) and subsequently by Misselbrook et al. (2000) to allow a means of estimating NH 3 emissions on a LU basis in relation to N inputs. This has been used here and it is assumed that all cattle graze for 185 and 365 days per year in the UK and New Zealand, respectively. Because emissions are determined largely by excreta, this period has been reduced by 2 h each day for the milking cows. Despite the lower inputs, but because of the longer grazing season and greater stock numbers, overall grazing losses are substantially higher in the New Zealand system (Table 4). The emission rates (16.5 kg N ha 1 ) for New Zealand compare with recently measured rates of 15 kg N ha 1 in a recent study (Ledgard et al., 1999) Fertilized swards The estimated rates for the UK grazed swards also include the net effect of direct losses from applied NH 4 NO 3, but a separate estimate has been made for that applied to the cut swards. The effect of the greater losses from applied urea on the New Zealand farm are accounted for separately from grazing effects. Emissions from urea fertilizer can vary enormously (Jarvis and Pain, 1990; Van Der Weerden and Jarvis, 1997); whilst the UK inventory EF for urea is 23% of the N applied, a value of 12% as determined for local conditions (Ledgard et al., 1999) has been applied to all New Zealand swards. This provides the second biggest emission source on the New Zealand farm (Table 4). The lower loss rate from urea in New Zealand can be attributed to the highly ph-buffered soils and to the time of application (cool, wet conditions in early spring and mid-late autumn) Slurry Direct loss occurs from the slurry lagoon and is calculated as an average emission occurring throughout the year. Slurry spreading provides the biggest opportunity for loss on the UK farm and the estimate assumes a fixed rate of loss from the TAN applied and, because the slurry has relatively high DM (4 8%) (Misselbrook et al., 2000), the rate is high. Overall, emissions from slurry are equivalent to 26.7 kg N ha 1 and represent 47% of the UK farm total. Contributions from excreta deposited during milking on the New Zealand farm are accounted for in losses from the milking parlour and dirty water/dairy shed washings as discussed below Dirty water/dairy shed washings Emissions from the UK dirty water in the storage tank are relatively low, but occur throughout the year; application to land also results in a low loss rate because of low TAN and DM contents and together comprise only a small proportion (<2%) of the total. The New Zealand dirty water (dairy parlour washings and yard housings 2 per day, results in larger volumes of water with lower N contents (0.13% TAN) than the UK equivalent (0.4% TAN). The estimate for loss again assumes an active emission surface throughout

7 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) the year, but because the area of the holding pond is small, total emissions are also small. Storage and application of dirty water to land on the New Zealand farm is responsible for <6% of the total loss Housing The UK animals are housed for 180 days in the present model farm. Information for this component of management is scarce but Pain et al. (1998) and Misselbrook et al. (2000) established annual EFs per LU including time spent indoors for milking during grazing. This loss is equivalent to 14.6 kg N ha 1 over the whole farm and is 25% of the UK total. Only time spent at milking contributes to this loss in New Zealand and is calculated as a proportion of the UK rate Collection yards/laneways The yards used to assemble cows prior to milking in the UK provide an effective emission source (Misselbrook et al., 1998). Although these areas are scraped cleaned, the rough concrete surface provides a good medium for emission throughout the year. The EF used (Table 4) is based on measurements by Misselbrook et al. (1998). The New Zealand laneways provide a comparable source on which excreta are deposited 2 per day. There are no data for emissions from laneways but, because of the frequent addition of excreta (at 2.2 the density of that in the pasture), no plant sink to remove N, and physical disturbance of the surface, losses would be high, estimated to be 2 that from grazed swards. Both these factors have been applied to that for grazed areas to provide an EF for laneways Total losses Total NH 3 losses from both farms represent substantial removals of N and, across each farm, were equivalent to 57 and 24 kg N ha 1 for UK and New Zealand, respectively (Tables 4 and 5). Whilst information on absolute losses is useful, it is important that these are expressed on a uniform basis; Table 5 provides indicators for NH 3 loss which allow an improved basis for comparison. Losses from both farms represent similar proportions of N inputs, whether on the basis of all inputs (i.e. including atmospheric deposition, feeds and bedding) or fertilizers and fixation only (17 and 21%, and 12 and 13% for UK and New Zealand farms, re- Table 5 Indicators for NH 3 emissions from dairy farming systems in the UK and New Zealand NH 3 -N emissions UK New Zealand Farm total (kg) kg ha kg per kg N input For total farm input 0.17 (17.0%) 0.12 (12.1%) For fertilizer/fixation 0.21 (20.8%) 0.13 (12.6%) kg per LU Cows Total kg per 1000 l milk kg per kg protein N removed % Farm N surplus spectively). Because of lower stocking rate and milk production/ha in the UK, losses of NH 3 expressed per LU or per 1000 l milk are twice those in New Zealand. The UK farms rears young stock and it is more correct to express values on the basis of total N conversion into milk +bodyweight protein. Although this (Table 5) reduces the difference, the UK farm produces 1.9 more NH 3 -N per kg protein-n than in New Zealand. This, in large part, results from the housing in the UK, but is also a consequence of higher N input. When NH 3 loss is expressed in relation to the farm N surplus, there is little difference between the two farms with NH 3 -N loss being ca. 20% of the surplus in each case Potential for control There is much discussion on appropriate means to reduce NH 3 volatilization from animal production. The United Nations Economic Commission for Europe (UNECE) has been developing a second protocol to limit European NO x emissions that would also consider controls on NH 3 (UNECE, 1999). Methods to examine net fluxes of NH 3 from farming systems have been based on annual mass balances (Jarvis, 1993; Jarvis et al., 1996), mass flow of NH 3 -N through the farm with volatilization occurring at various stages (Cowell and ApSimon, 1998) or a mechanistic approach to the controls over volatilization (Hutchings et al., 1996). These methods have been developed for W. European conditions and may need some modifi-

8 90 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) cation to be totally applicable to New Zealand farms, but in the absence of other information the simple approach of applying single, simple changes to the EFs used in the present paper provides a basis for initial discussion. A limited number of previously defined mitigations (Cowell and ApSimon, 1998; UNECE, 1999) which are applicable to the two farms and with quoted values for likely reductions in emission have been used to compare the two farms. A first point to make is that effects of any specific action taken to reduce NH 3 fluxes will be interactive with other sources downstream and impacts of different options are therefore not simply additive. There will also be effects on other N loss pathways. Further, any general improvements to N management and efficiency will have the effect of reducing NH 3 emissions. In a previous systems analysis (Jarvis et al., 1996), NH 3 emissions were reduced from 48 to 24 kg N ha 1 when management was changed to reduce NO 3 leaching. This not only included reducing fertilizer N but also reducing N inputs in animal feed. In general, 1 kg reduction in N excretion results in a kg reduction in NH 3 -N emission (UNECE, 1999) The UK farm A number of practical options exists for the UK farm, most of which relate to the management and utilization of slurry. Other components of the farm have potential for abatement but either the overall impact on total emission is relatively small or the costs are great (Cowell and ApSimon, 1998). The range of measures proposed which can be applied to animal houses, e.g. decreasing areas fouled by manures, using straw to absorb urine, flushing/scraping +/ acid to regularly remove slurry are all relatively expensive. Increasing grazing time (losses per day whilst animals are in the field are lower than when they are housed) decreases loss but it can be calculated from Table 4 that increasing grazing by 10 days would reduce NH 3 emission by <4%. There would also be other possible managerial penalties attached to this option. The slurry store is another important source (Tables 3 and 4) and proposals have been made to reduce this flux, the most practical and viable options are to allow natural crusts to develop or to add materials to promote this (e.g. straw, bark, peat, etc.). The effects of these are poorly quantified, especially where natural crusting will already be taking place. UNECE (1999) considers that a 35 40% reduction is achievable and this gives a maximum potential reduction from the present store of ca. 190 kg per annum. However, this would also increase the potential for loss when the slurry was applied to land by increasing TAN content (Table 4). The net effect is that overall loss would be reduced by 124 kg NH 3 -N (<3% of the annual total). In fact, effects may be less than this, since increasing the slurry DM content of the slurry by adding materials would also increase the spreading EF (Misselbrook et al., 1998). Changing the store to a sealed tank would have greater effect, but would be expensive. Other than by reducing application rates and therefore overall farm inputs, changes to the fertilizer management have little direct effect. However, there would be substantial effect of changing the application method for the slurry. Immediate incorporation into the soil is an effective option where there is tillage on the farm: however, for many UK dairy farms this is not a practical option unless maize (Zea mays L.) is being grown for silage. A number of practical solutions for reducing loss from slurry application have been examined with quantitative data on abatement efficiency. These include band spreading (10% reduction), trailing shoe (40%), injection (open slot 60%, closed slot 80%) (UNECE, 1999). Selection of the method will depend upon the soil type, slope and sward. Costs also are variable, injection being nearly three times more expensive than bandspreading. For the present farm, the trailing shoe method has been selected as the most appropriate. The direct effect of applying all slurry this way reduces the total emission by approximately 14% (i.e. from 57 to 49 kg NH 3 -N ha 1 ). In fact, the benefits are likely to be greater than this because the reduced emission will reduce the farm N requirements and hence reduce emissions further, but only by 20 kg N for the whole farm The New Zealand farm Ammonia sources on the New Zealand farm are fewer and hence there are fewer options for abatement. Most sources are relatively small, only fertilizer application and grazing have large enough emissions to offer practical opportunities for reduction. Changing from urea to another form of fertilizer N would reduce emissions from >8 to <1kgha 1 (an overall

9 S.C. Jarvis, S. Ledgard / Agriculture, Ecosystems and Environment 92 (2002) reduction of 26%). However, the higher relative costs of alternative sources of N (i.e. ca. 90 and 130% more expensive for (NH 4 ) 2 SO 4 or calcium ammonium nitrate per kg N than urea, respectively) are not likely to promote this change. Timing applications to coincide with rainfall could reduce the current rate of loss (e.g. by 70% with 4 mm rain within 1 h, Black et al. (1987)). In some regions of New Zealand, particularly with heavy textured soils, there is interest in using feed pads for ca. 90 days over winter to decrease pasture damage from grazing and reduce nitrate leaching. However, indications are that this would raise NH 3 emissions by 27% (using UK factors for housing and slurry), i.e. more than twice the current loss. Lowering dietary protein contents and increasing energy intake would divert proportionately more excreted N into dung rather than urine and thus reduce NH 3 losses. This would require a major change in New Zealand s low-cost pasture system and is therefore unlikely to attract any significant uptake. Studies comparing grass/clover with grass + fertilizer N in New Zealand with similar N inputs showed only small differences in mean N concentration in the herbage (i.e and 3.23% N, respectively; Sprosen et al. (1997)). It is unlikely that this would result in reduced NH 3 emissions. Systems research using very efficient production with no fertilizer N has indicated the potential for a high potential DM production, high pasture utilization with a high stocking rate and therefore high production and removal of N in milk proteins (47% conversion of fixed N into milk N; Ledgard et al. (1999)). This and other strategies to improve production per unit area without increasing N inputs (e.g. through improved genetic resource, development of more efficient dietary utilization, etc.) will reduce NH 3 emission (and other N losses) per capita of production. These effects apply to the UK system as well as to the New Zealand one. 4. Conclusions Emissions of NH 3 from dairy farming systems are inevitable. Patterns of cycling and transformation of N are such that there will always be potential for volatilization to occur. The present approach provides a simple means of understanding impacts of whole farming systems and providing comparisons. In analyses of the present nature, it is important that common standpoints be developed so that valid comparisons can be made. The derivation of key indicators for NH 3 provides the basis of a practical means for both policy makers and practitioners to determine and compare relative effects. In the present case, the apparently substantial difference between the two farming systems is reduced when NH 3 losses are expressed per unit of protein produced and is reduced further if compared with the N surplus of the system. The overall high N inputs to UK dairy systems and, more importantly, the requirements for winter housing for the cows and collection, storage and application of their excreta, contribute to the significantly greater losses from the UK than the New Zealand farm. The UK dairy farm, because it has more potential sources of NH 3, and a greater overall total loss has a greater potential for attempts to mitigate against loss to be successful. In both environments, options to reduce N surplus in the system, increase N use efficiency and improve C:N dietary balance for highly productive cows have considerable potential to reduce volatilization (and will also have benefits for other N losses). Acknowledgements This work was funded by MAFF, London and AgResearch, whilst S.C.J. was on a study tour of New Zealand funded by the British Council and the New Zealand Lotteries Board. IGER is sponsored by the BBSRC, Swindon, UK. References Black, A.S., Sherlock, R.R., Smith, N.P., Effect of timing of simulated rainfall on ammonia volatilization from urea applied to soil of varying moisture content. J. Soil Sci. 38, Buijsman, E., Aben, J.M.M., van Elsakker, B.G., Mennen, M.G., An automatic atmospheric ammonia network in the Netherlands, set up and results. Atmos. Environ. 32, Butler, P.J., Johnson, T.J., Fertiliser use in the Bay of Plenty, Waikato and South Aukland regions. In: Proceedings of the New Zealand Grassland Association, Vol. 59, pp Cowell, D., ApSimon, H., Cost effective strategies for the abatement of ammonia emissions for European agriculture. Atmos. Environ. 32, Duyzer, J.H., Dry deposition of ammonia and ammonia aerosols over heathland. J. Geogr. Res. 99,

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