Introduction
Agriculture is responsible for 99.4% of ammonia (NH3) emissions from the Republic of Ireland (ROI) (EPA, 2022). The national total emissions in the ROI increased by 12.4% from 109.80 kt in 1990 to 123.4 kt in 2020 with livestock production accounting for the majority of national total emissions (EPA, 2022). In the year 2020, application of manure to soil, manure management and deposition of urine and dung by grazing animals together contributed to 90.1% of the ROI’s national total emissions (EPA, 2022). Other factors contributing to the increase in the ROI’s emissions include the increase in fertiliser use, continued use of urea as an inorganic nitrogen (N) fertiliser as well as the increase in dairy cattle and other cattle populations (EPA, 2022).
Similarly, the agricultural sector remains the main source of NH3 emissions in Northern Ireland (NI) with cattle manure management contributing to at least 35% of the emissions (NAEI, 2022). Emissions in NI have increased since 2011 mainly due to increasing dairy herd size and emissions associated with dairy manure management (NAEI, 2022). The trend has remained fairly stable since 2017 even though the slight declines in dairy cattle numbers and in mineral fertiliser are being offset by an increase in poultry production (NAEI, 2022).
Atmospheric NH3 contributes to acid depositions and high concentrations of nutrients in aquatic and terrestrial ecosystems which poses a threat to the sustainability of these ecosystems (Wilkins et al., 2016; Payne et al., 2017). According to DEFRA (2022), most of NI including designated sites and other priority habitats are now receiving levels of nitrogen (N) which are significantly above their critical loads, or the concentration at which significant and irreparable ecological damage occurs. Similarly, the risk of environmental degradation as a result of NH3 deposition on terrestrial ecosystems is quite high in the ROI due to intensive agricultural production systems. Indeed, after monitoring NH3 concentrations in 12 Natura 2000 sites in the ROI, Kelleghan et al. (2021) observed that 11 sites had exceeded either their critical level or load whereas 10 sites had exceeded both. The MARSH model also identified 80.7% and 5.9% of Natura 2000 sites in the ROI may exceed critical levels of 1 μg/m3 (i.e. habitats where lichens and moss are important features) and 3 μg/m3 (all other habitats), respectively (Kelleghan et al., 2019).
The European Union (EU) has set limits for national emissions of NH3 under the National Emission Reduction Commitments Directive (NECD) 2016/2284. According to the NECD, the ROI and the United Kingdom (UK) are required to reduce NH3 emissions relative to 2005 baseline by 1% and 8% by 2020 and 5% and 16% by 2030, respectively (EPA, 2022; NAEI, 2022). The ROI has exceeded its NH3 limit since 2016 and though recent predictions estimate that the ROI can meet its 2030 target, immediate adoption and implementation of emission reduction strategies is required (Buckley et al., 2020).
Strategies for abating NH3 emissions in livestock production include dietary manipulation (e.g. reducing the crude protein content in animal diet), housing strategies (e.g. low-emission flooring types, regular cleaning of housing floors, the application of urease inhibitors on floors), storage strategies (e.g. reducing slurry pH and NH4 + content, covering of slurry stores, inducing crust formation), field strategies (e.g. adopting low-emission spreading strategies [LESS]), rapid incorporation of slurry after application) and the use of protected urea (i.e. urea with urease inhibitor) as a mineral fertiliser (Bittman et al., 2014; Buckley et al., 2020; Bobrowski et al., 2021). Besides Ireland’s agriculture being dominated by livestock production which has bearings on emissions, livestock production systems are mainly grass-based, making additional mitigation difficult. Even though some NH3 abatement techniques are available to Irish agriculture, their adoption and implementation may be limited at the farm level due to their associated cost. An assessment of NH3 mitigation strategies and their associated costs in Ireland (i.e. Teagasc’s NH3 marginal abatement cost curve [MACC]) shows that the most cost-effective (i.e. 80% reduction) strategy for abating NH3 in Ireland is the implementation of LESS and the substitution of urea fertilisers with protected urea (i.e. urea with the urease inhibitor N-(n-butyl) thiophosphoric triamide [NBPT]) (Buckley et al., 2020).
A number of studies have been conducted to evaluate the effectiveness of NH3 abatement strategies available to Irish agriculture. To ascertain whether these NH3 abatement strategies are indeed contributing to NH3 emission reductions in ROI and NI, a compilation of NH3 EFs derived from these studies is required. Furthermore, a compilation of EFs will provide information that can be used to improve national inventories and potentially identify any new abatement strategies that may be relevant to Irish agriculture. This review presents a narrative of NH3 EFs generated from cattle production systems under Irish environmental conditions.
Methodology
Ammonia emissions data reviewed for this study were retrieved from the digital libraries of Teagasc, Ireland and the bibliographical database Scopus. Articles published in scientific journals were retrieved using the following keyword combination: cattle AND slurry OR urine OR dung AND ammonia AND emissions AND grazing OR housing OR concrete OR yards OR storage OR land AND application AND Ireland. Data search from both databases yielded 45 results. Eligibility criteria for the study selection were predefined to eliminate publication bias. That is, the study should report NH3 emissions from cattle production systems (i.e. slurry, dung, urine and fertiliser application to grassland) during grazing, housing, storage and landspreading under Irish (both ROI and NI) environmental conditions. The studies which did not fulfil the above criteria were excluded from this review or analysis. Out of the 45 assessed for eligibility, 16 articles fulfilled our criteria and were selected.
Data reviewed in this paper relates to NH3 measurements in which results were either reported as experimental results or NH3 emission factors (EFs). Emission factors from the deposition of urine and dung during grazing as well as mineral fertiliser application to grasslands are presented as a percentage (%) of total nitrogen (TN). As the TN is more representative of the N present in urine and urea fertiliser, EFs from dung, urine and mineral fertilisers are expressed as %TN to facilitate comparison of EFs between these N sources. In the case of concrete yards, slurry storage pits and slurry landspreading, EFs are reported as % of total ammoniacal nitrogen (TAN) as NH3 emissions from slurry at these stages mainly emanate from the TAN in the slurry. If EFs were already presented as %TAN, EFs were included unchanged. If EFs were presented as TN and the TAN content of the slurry was reported, EFs were converted into %TAN applied. In instances where EFs were reported as TN without any information on the slurry TAN content (i.e. Bourdin et al., 2014; Burchill et al., 2019; McIlroy et al., 2019), EFs were converted into %TAN based on the assumption that 60% of the TN excreted in cattle slurry is TAN (EPA, 2022). Emission factors generated from models (e.g. Cahalan et al., 2015) were not included in the analysis of this review.
Grazing
Urine and dung deposition
Irish agriculture is dominated by pastoral bovine livestock production, with approximately 93% of the utilisable agricultural area composed of permanent grassland (CSO, 2021). Therefore, in these grass-based production systems, cattle and sheep spend most of the year at pasture. Their excretion of urine and dung during grazing leads to emitting approximately 13.63 kt NH3 per annum, or 10.9% of the total agricultural NH3 emissions (EPA, 2022). This is a small proportion of overall emissions considering the extensive length of the grazing season at over 220 d per annum (Läpple et al., 2012), compared to 47.4% of NH3 produced by manure management (EPA, 2022). Indeed, extending grazing length is in itself a category 1 NH3 abatement measure based on a simple principle that quick infiltration of urine in the soil leads to lower NH3 emissions during grazing compared to housing, storage and landspreading of manure (Bittman et al., 2014).
Fischer et al. (2016) reported NH3 EFs in Ireland between 2.8% and 5.3% TN and between 8.7% and 14.9% TN for dung and urine, respectively, depending on the season of application (spring, summer, autumn). Surprisingly, the largest emissions were observed in spring despite lower air temperatures compared to summer, which is typically associated with higher emissions (Clay et al., 1990; Lockyer & Whitehead, 1990; Sommer et al., 1991). This clearly highlights the importance of other factors such as rainfall aiding infiltration of the excreted material. Emissions were lower from dung than urine, reflecting both a lower N application rate and a higher proportion of organic N in this excreta form. Burchill et al. (2017b) investigated the effect of applying urea fertiliser (with and without N stabilisers) after urine patch deposition on NH3 emissions on two occasions (i.e. May and June). Burchill et al. (2017b) found no significant differences in EFs between treatments receiving the two N stabilisers (dicyandiamide [DCD] and NBPT) in May and June relative to the other treatments (i.e. urine, urea and urine plus urea). The EFs generated from urine patches in May and June were 6.3% and 7.1% TN, respectively, representing spring/summer applications. However, both of these studies used wind tunnel methodology (Lockyer, 1984) which is more suitable for replicated treatment comparisons at small scales rather than the provision of reliable data for the development of country-specific emission factors. As a result, the Irish air pollutant inventory still uses a Tier 1 NH3 EF for grazing animals of 6% TAN (EMEP/EEA; 2019; EPA, 2022). Figure 1 shows the mean NH3 EFs from Irish studies for cattle excreta (i.e. urine or dung) and subsequent fertilisation of urine patches with urea (with and without N stabilisers). As expected, mean EF from dung was lower than urine as most of the N in dung is usually in the organic form and needs to be converted to NH4 + in order to generate NH3 emissions. The addition of DCD to urine alone or in combination with urea to urine patches tended to increase EFs relative to the urine or urine plus urea treatments. As DCD acts as a nitrification inhibitor and not a urease inhibitor, it decreases the rate of conversion of NH4 + to nitrate (NO3 −) and therefore it is not surprising that its inclusion led to increased NH3. Such increases in NH3 emissions after application of N stabilisers such as DCD to urea have also been observed in non-Irish studies (Prakasa Rao & Puttanna, 1987; Asing et al., 2008). In contrast, the addition of the urease inhibitor NBPT and NBPT plus DCD to urea after application to urine patches led to mean reductions in EFs by 28% and 21%, respectively, relative to when urea was applied to urine patches alone. The reduction in EF through the application of NBPT can be attributed to the inhibitory effect of NBPT on urea hydrolysis.
Mean EFs from Irish studies for cattle excreta (i.e. urine or dung) and subsequent fertilisation of urine patches with urea (with and without N stabilisers) on grassland. D = dicyandiamide (DCD); EFs = emission factors; N = N-(n-butyl) thiophosphoric triamide (NBPT).
In general, Irish experimental data are comparable with or slightly lower than results from other countries. For example, Zaman et al. (2013) reported NH3 EFs from urine in the range of 4.9% and 12% TN depending on the year and season of application in New Zealand, while in another New Zealand-based study, Laubach et al. (2013) found urine and dung emissions to be 25.5% TN and 11.6% TN, respectively. A study by Misselbrook et al. (2014) found NH3 from urine patches in the UK to be 25.2% of the applied N and in the US, NH3 EFs were between 10% and 35% TN for urine and 5% and 7% TN for faeces from pasture and shortgrass steppe rangeland (Nichols et al., 2018). However, reviews from Selbie et al. (2015) and Cai & Akiyama (2016) recorded wide ranges of NH3 emission factors driven by animal, soil and climatic conditions as well as measurement methodology, further proving difficulty but also a need for providing a reliable basis for country-specific emission factors.
While grazing is already considered an NH3 abatement measure, additional technologies have been suggested to further reduce emissions from this activity such as reducing crude protein intake in animal diet (Smith et al., 2008; Zaman & Blennerhassett, 2010), sodium chloride (NaCl) supplementation (Liu & Zhou, 2014) and use of urease inhibitors (Saggar et al., 2013). However, evidence is still limited regarding overall efficacy and the effects of season, soil temperature and moisture, rainfall and soil organic carbon on the variability in the efficacy and practicability of the implementation of some of these technologies. Table 1 summarises Irish studies on NH3 EFs as %TN from the deposition/application of urine, dung and subsequent fertilisation with urea (with and without N stabilisers) on pasture. The findings from the studies show that NH3 EFs from urine and dung deposited at pasture vary to a considerable extent owing to differences in the characteristics of excreta, environmental conditions at the time of application and limited experimental data specific to Irish conditions.
Ammonia EFs as percentage (%) of TN after the deposition of cattle excreta (urine and dung), urine plus N stabilisers and urine plus urea (with or without N stabilisers) on grassland (perennial ryegrass and white clover mixture) in Ireland
| Treatment | Time of application | NH3 lost as %TN | % change in EF relative to control | Measurement method | Reference |
|---|---|---|---|---|---|
| Urine (control) | May | 6.3 | NA | Wind tunnel | Burchill et al. (2017b) |
| Urine + urea | May | 6.0 | 5 | ||
| Urine + urea + DCD | May | 5.9 | 6 | ||
| Urine + urea + NBPT | May | 4.6 | 30 | ||
| Urine + urea + NBPT + DCD | May | 6.2 | 2 | ||
| Urine (control) | June | 7.1 | NA | ||
| Urine + urea | June | 9.6 | 35 1 | ||
| Urine + urea + DCD | June | 10.5 | 48 1 | ||
| Urine + urea + NBPT | June | 6.6 | 7 | ||
| Urine + urea + NBPT + DCD | June | 6.1 | 14 | ||
| Dung | April | 5.3 | NA | Wind tunnel | Fischer et al. (2016) |
| Urine (control) | April | 14.9 | NA | ||
| Urine + DCD | April | 19.5 | 31 1 | ||
| Dung | July | 2.8 | NA | ||
| Urine (control) | July | 9.8 | NA | ||
| Urine + DCD | July | 9.7 | 1 1 | ||
| Dung | September | 3.5 | NA | ||
| Urine (control) | September | 8.7 | NA | ||
| Urine + DCD | September | 9.5 | 9 1 |
Soil type/classification: Luvic gleysol.
DCD = dicyandiamide; EF = emission factor; NA = not applicable; NBPT = N-(n-butyl) thiophosphoric triamide; TN = total nitrogen.
1Represents an increase in emission factor.
Mineral fertiliser application to grassland
The use of mineral N fertilisers to meet N demand in agricultural soils is a common practice in most parts of the world including Ireland. While the above practice is required to ensure adequate fodder for grazing animals, the practice is considered as an important source of NH3 in Ireland accounting for 9.7% of the total NH3 emissions (EPA, 2022). Figure 2 shows the mean EFs generated from (i) mineral fertiliser (i.e. calcium ammonium nitrate [CAN] and urea) application (with and without N stabilisers) and (ii) different timing of mineral fertiliser application to grassland. The results show that the mean EF from urea (27.9% TN) was higher than CAN (4.2% TN) which is consistent with the high availability of urea in urea fertiliser which hydrolyses rapidly to NH3 after application. While the use of CAN often leads to lower NH3 EFs relative to urea irrespective of the time/year of application, CAN is expected to generate higher emissions of nitrous oxide (N2O), which is a potent greenhouse gas, compared to urea. This is due to CAN’s higher nitrate content which becomes susceptible to denitrification after soil application (Harty et al., 2016). Results also show that besides the application of urea with maleic and itaconic acid polymer (MIP) which led to an increase in EF by 24%, the application of urea with N stabilisers such as NBPT, N-(n-propyl)-thiophosphoric triamide (NPPT) and the combined application of N stabilisers (i.e. NBPT + DCD, NBPT + NPPT) with urea led to a mean reduction in EF by 50% relative to urea application alone. The addition of DCD alone to urea appeared not to have any effect on EFs from the urea fertiliser. As mentioned earlier, DCD is a nitrification inhibitor and could rather increase NH3 than decrease. In the case of the time of application, EFs for urea were generally higher than CAN irrespective of the time of application. The lowest mean EFs after urea application were obtained in the months of June and July. In the case of CAN, EFs produced from the different months or time of application were fairly similar although emissions tended to increase in the months of April and September.
Ammonia emission factors after (A) mineral fertiliser (with or without N stabilisers) application to Irish grasslands and (B) timing of mineral fertiliser application. CAN = calcium ammonium nitrate; D = dicyandiamide (DCD); M = maleic and itaconic acid polymer (MIP); N = N-(n-butyl) thiophosphoric triamide (NBPT); NP = N-(n-propyl)-thiophosphoric triamide (NPPT).
The findings from the studies summarised in this section clearly indicate that the use of N stabilisers such as NBPT together with urea fertilisers can reduce N losses in the form of NH3 and consequently increase N availability in the soil for grass uptake. In contrast, the use of DCD showed variable results relative to its impact on NH3 EFs which agrees well with other studies (Kim et al., 2012). Thus, it may be assumed that the positive impact of the combined application of urea with NBPT and DCD is most likely a result of the action of NBPT in slowing urea hydrolysis through the inhibition of the urease enzyme. Even though the combined application of DCD and NBPT can have multiple benefits on the environment particularly mitigating NH3 and subsequent N2O emissions as well as NO3 − leaching, the use of only NBPT may be economical to farmers if the overall goal is to abate NH3. Table 2 summarises published literature in Ireland on NH3 EF from mineral fertiliser (urea and CAN) application to grassland.
Ammonia EFs as percentage (%) of TN applied from urea and calcium ammonium nitrate (CAN) application to grassland
| Fertiliser type | Soil type/classification | Time of application | NH3 Lost as % TN | % change in EF relative to control | Measurement method | Reference |
|---|---|---|---|---|---|---|
| Urea | Haplic Cambisol | September | 17 | NA | Wind tunnel | Burchill et al. (2016) |
| Urea | Albic Gleyic | September | 21 | NA | ||
| Urea | Haplic Cambisol | September | 14 | NA | ||
| Urea | Cutanic Luvisol | September | 22 | NA | ||
| Urea | Luvic Stagnosol | September | 18 | NA | ||
| Urea | Luvic Gleysol | September | 13 | NA | ||
| Urea | Luvic Gleysol | May | 8 | NA | Wind tunnel | Burchill et al. (2017b) |
| Urea | Luvic Gleysol | June | 17 | NA | ||
| CAN | MDSL | March | 5 | NA | Wind tunnel | Forrestal et al. (2016) |
| Urea (control) | March | 53 | NA | |||
| Urea – high rate | March | 68 | 28 1 | |||
| Urea + NBPT | March | 15 | 72 | |||
| Urea + DCD | March | 77 | 45 1 | |||
| Urea + NBPT + DCD | March | 18 | 66 | |||
| Urea + MIP | April | 55 | 4 1 | |||
| CAN | April | 7 | NA | |||
| Urea (control) | April | 25 | NA | |||
| Urea – high rate | April | 38 | 52 1 | |||
| Urea + NBPT | April | 6 | 76 | |||
| Urea + DCD | April | 42 | 68 1 | |||
| Urea + NBPT + DCD | April | 6 | 76 | |||
| Urea + MIP | April | 34 | 36 1 | |||
| CAN | June | 2 | NA | |||
| Urea (control) | June | 20 | NA | |||
| Urea – high rate | June | 24 | 20 1 | |||
| Urea + NBPT | June | 5 | 75 | |||
| Urea + DCD | June | 22 | 10 1 | |||
| Urea + NBPT + DCD | June | 4 | 80 | |||
| Urea + MIP | June | 16 | 20 | |||
| CAN | July | 4 | NA | |||
| Urea (control) | July | 26 | NA | |||
| Urea – high rate | July | 24 | 8 | |||
| Urea + NBPT | July | 9 | 65 | |||
| Urea + DCD | July | 22 | 15 | |||
| Urea + NBPT + DCD | July | 5 | 81 | |||
| Urea + MIP | July | 21 | 19 | |||
| CAN | August | 2 | NA | |||
| Urea (control) | August | 30 | NA | |||
| Urea – high rate | August | 43 | 43 1 | |||
| Urea + NBPT | August | 6 | 80 | |||
| Urea + DCD | August | 45 | 50 1 | |||
| Urea + NBPT + DCD | August | 8 | 73 | |||
| Urea + MIP | August | 47 | 57 1 | |||
| CAN | IDCL | March | 4 | NA | ||
| Urea (control) | March | 21 | NA | |||
| Urea – high rate | March | 14 | 33 | |||
| Urea + NBPT | March | 2 | 90 | |||
| Urea + DCD | March | 9 | 57 | |||
| Urea + NBPT + DCD | March | 6 | 71 | |||
| CAN | May | 2 | NA | |||
| Urea (control) | May | 8 | NA | |||
| Urea – high rate | May | 19 | 138 1 | |||
| Urea + NBPT | May | 2 | 75 | |||
| Urea + DCD | May | 7 | 13 | |||
| Urea + NBPT + DCD | May | 2 | 75 | |||
| CAN | June | 3 | NA | |||
| Urea (control) | June | 33 | NA | |||
| Urea – high rate | June | 49 | 48 1 | |||
| Urea + NBPT | June | 4 | 88 | |||
| Urea + DCD | June | 20 | 39 | |||
| Urea + NBPT + DCD | June | 7 | 79 | |||
| CAN | July | 4 | NA | |||
| Urea (control) | July | 31 | NA | |||
| Urea – high rate | July | 34 | 10 | |||
| Urea + NBPT | July | 4 | 87 | |||
| Urea + DCD | July | 13 | 58 | |||
| Urea + NBPT + DCD | July | 6 | 81 | |||
| CAN | September | 8 | NA | |||
| Urea (control) | September | 33 | NA | |||
| Urea – high rate | September | 43 | 30 1 | |||
| Urea + NBPT | September | 9 | 73 | |||
| Urea + DCD | September | 19 | 42 | |||
| Urea + NBPT + DCD | September | 11 | 67 | |||
| CAN | MDSL | July | 5 | NA | Wind tunnel | Krol et al. (2020) |
| Urea (control) | July | 43 | NA | |||
| Urea + NBPT | July | 14 | 67 | |||
| Urea + NBPT + NPPT | July | 14 | 67 | |||
| Urea | July | 47 | NA | IHF method |
CAN = calcium ammonium nitrate; DCD = dicyandiamide; EFs = emission factors; IDCL = imperfectly drained clay loam; IHF = integrated horizontal flux; MDSL = moderately-drained sandy loam; MIP = maleic and itaconic acid polymer; NA = not applicable; NBPT = N-(n-butyl) thiophosphoric triamide; NPPT = N-(n-propyl)-thiophosphoric triamide; TN = total nitrogen.
1Values represent an increase in EF.
Cattle buildings and concrete farmyards
Housing is an essential component of animal agriculture. The primary goal for providing housing for animals is to promote good health and welfare in order to maximise productivity. In Ireland, cattle are mainly housed in naturally ventilated buildings during the winter period. In the winter season, cattle are generally housed either on straw bedding, in sheds over slatted tanks or in cubicle/loose sheds with floors scraped regularly into open tanks (Lanigan et al., 2015). In the straw-based systems, straw is added frequently and either allowed to accumulate or removed frequently as farmyard manure. The effluents (i.e. mixture of urine, spilled water, faeces, food particles, etc.) collected below the slatted floor are stored as slurry. The slurry-based systems (i.e. systems where effluents are collected below slatted tanks) predominate the dairy sector with about 94% of dairy housing in Ireland being slurry based (McIlroy et al., 2019). Cattle buildings are important sources of NH3 emissions. Indeed, manure management from cattle (i.e. dairy and non-dairy) generates 50 kt NH3 per annum which represents about 40% of national total emissions in Ireland (EPA, 2022).
Table 3 summarises NH3 EFs generated from cattle buildings and concrete yards used by cattle under Irish conditions. Emissions from cattle buildings can vary to a considerable extent due to variations in factors such as building design/size, number of cattle and wind velocity/direction (Burchill et al., 2017a). Burchill et al. (2017a) quantified NH3 EFs from four cattle buildings which varied in size, floor type as well as the type and number of livestock housed. The mean EF reported from the study was 13% TAN. The overall mean EF of 13% TAN reported in the Irish study is somewhat lower than the applied EF of 28% TAN in liquid or slurry-based housing in the Irish NH3 inventory (EPA, 2022). The above observation suggests an over-estimation of EFs used in the Irish NH3 inventory. There is therefore the need to generate additional datasets on NH3 emissions from cattle buildings to validate the findings from Burchill et al. (2017a). In the case of concrete farmyards or hard standings, mean EFs generated from the analysis of this review was 35% TAN.
Ammonia EFs as percentage (%) of TAN from cattle buildings and hard standings or concrete yards
| Treatment | NH3 lost (%TAN) | % change in EF relative to control | Measurement method | Reference |
|---|---|---|---|---|
| Cattle buildings | 13 | NA | Passive flux samplers (Ferm tubes) | Burchill et al. (2017a) |
| Concrete standings | ||||
| Slurry (control) | 24 | NA | Dynamic flow-through chamber | McIlroy et al. (2019) |
| Slurry + alum | 8 | 67 | ||
| Slurry + calcium chloride | 10 | 58 | ||
| Slurry + Actisan | 11 | 54 | ||
| Slurry + sulphuric acid | 16 | 33 | ||
| Slurry + Agrotain | 22 | 8 | ||
| Slurry + Envirobed | 23 | 4 | ||
| Slurry + double UI | 21 | 13 | ||
| Slurry + clinoptilotite | 21 | 13 | ||
| Slurry + eugenol | 24 | 0 | ||
| Slurry + sawdust | 22 | 8 | ||
| 1 kg dung + 0.67 L urine | 36 | NA | Wind tunnel | Burchill et al. (2019) |
| 1 kg dung + 1 L urine | 42 | NA | ||
| 1 kg dung + 2 L urine | 48 | NA | ||
| Excreta without pressure wash (control) | 36 | NA | ||
| Excreta + pressure wash 1 h post application | 4 | 89 | ||
| Excreta + pressure wash 3 h post application | 7 | 81 | ||
| Excreta without scraping (control) | 46 | NA | ||
| Excreta + scraping 1 h post application | 9 | 80 | ||
| Excreta + scraping 3 h post application | 22 | 8 |
EFs = emission factors; NA = not applicable; TAN = total ammoniacal nitrogen.
A number of strategies have been proposed as being effective at abating NH3 from cattle buildings or sheds and concrete farmyards. These strategies include regular cleaning of sheds or yards, adsorption of urine using materials such as straw and the use of chemical additives such as urease inhibitors (Burchill et al., 2019; Bobrowski et al., 2021). Ammonia abatement techniques such as application of chemical additives and regular cleaning of yards/shed floors by washing with water or scraping have been evaluated under Irish conditions. Burchill et al. (2019) reported a reduction in EFs of up to 89% after cleaning (i.e. either pressure washing with water or scraping) of excreta. Greater reductions in EFs were noted when cleaning was done immediately (i.e. 1 h post deposition/application) after excreta deposition than later (i.e. 3 h post deposition/application). Similar observations were made in other studies conducted in the UK (Misselbrook et al., 1998, 2001) and were attributed to the removal of the N source (urine and dung) from the emitting surface area after cleaning. In the case of chemical additives, McIlroy et al. (2019) reported NH3 reductions of 67%, 58%, 54% and 33% after application of alum, calcium chloride, actisan (disinfectant) and sulphuric acid, respectively, from slurry applied to a concrete surface under typical Northern Irish cattle housing conditions.
Storage pits
Slurry storage is an important step in managing manure as it allows slurry to be kept until conditions are suitable for landspreading. However, a substantial proportion of slurry N is emitted as NH3 during storage, which reduces the amount of readily available N for plant uptake after landspreading. Research findings on EFs from slurry storage are summarised in Table 4. The mean EF for the control or un-amended slurries was 60% TAN and values ranged between 42–68% TAN. The mean EF (60% TAN) obtained in the present analysis can be considered higher than the EF of 5% and 28% TAN used by the Irish EPA in the Informative Inventory Report for covered slurry during storage and slurry-based cattle houses, respectively (EPA, 2022). It is of note that these studies (Table 4) were small-scale incubations measuring NH3 in a dynamic chamber mode so the airflow is expected to be higher than above a typical tank in the shed. Thus, the values reported in the studies in Table 4 likely overestimate emissions and cannot be used to provide reliable EFs from storage, but rather provide an indication of the efficacy of abatement strategies. Future studies should aim to measure emissions at a scale representative of current farming practice.
Ammonia EFs as percentage (%) of TAN from slurry storage pits in cattle houses
| Treatment | Dry matter (%) | Inclusion rate of additive | NH3 lost (%TAN) | % change in EF relative to control | Storage period (days) | Measurement method | Reference |
|---|---|---|---|---|---|---|---|
| Storage | |||||||
| Slurry (control) | 4 | — | 68.0 | NA | 83 | Dynamic chamber | Kavanagh et al. (2019) |
| Slurry + acetic acid | 4 | 35.7 g | 22.0 | 68 | 83 | ||
| Slurry + alum | 4 | 53.1 g | 6.0 | 91 | 83 | ||
| Slurry + ferric chloride | 4 | 41.3 g | 1.4 | 98 | 83 | ||
| Slurry + sulphuric acid | 4 | 20.8 g | 10.0 | 85 | 83 | ||
| Slurry (control) | 7 | — | 63.1 | NA | 83 | ||
| Slurry + acetic acid | 7 | 43.5 g | 12.9 | 80 | 83 | ||
| Slurry + alum | 7 | 88.5 g | 8.7 | 86 | 83 | ||
| Slurry + ferric chloride | 7 | 70.1 g | 1.6 | 97 | 83 | ||
| Slurry + sulphuric acid | 7 | 19.2 g | 8.9 | 86 | 83 | ||
| Slurry (control) | 7 | — | 42.2 | NA | 70 | Dynamic chamber | Kavanagh et al. (2021) |
| Slurry + apple pulp | 7 | 7.0% | 22.7 | 46 | 70 | ||
| Slurry + apple pulp | 7 | 15.0% | 29.2 | 31 | 70 | ||
| Slurry + brewers grain | 7 | 7.0% | 33.3 | 21 | 70 | ||
| Slurry + brewers grain | 7 | 15.0% | 29.8 | 29 | 70 | ||
| Slurry + dairy washings | 7 | 7.0% | 47.2 | 12 | 70 | ||
| Slurry + dairy washings | 7 | 15.0% | 39.5 | 6 | 70 | ||
| Slurry + dairy waste | 7 | 10.0% | 27.4 | 35 | 70 | ||
| Slurry + grass silage | 7 | 15.0% | 23.9 | 43 | 70 | ||
| Slurry + grass silage | 7 | 7.0% | 32.4 | 23 | 70 | ||
| Slurry + maize silage | 7 | 15.0% | 34.6 | 18 | 70 | ||
| Slurry + maize silage | 7 | 7.0% | 61.3 | 45 | 70 | ||
| Slurry + sugarbeet molasses | 7 | 3.0% | 25.1 | 41 | 70 | ||
| Slurry + sugarbeet molasses | 7 | 5.0% | 11.3 | 73 | 70 | ||
| Slurry + sugarbeet molasses | 7 | 7.0% | 40.5 | 4 | 70 | ||
| Slurry (control) | 7 | — | 66.1 | NA | 116 | ||
| Slurry + commercial additive 1 | 7 | RA | 60.7 | 8 | 116 | ||
| Slurry + commercial additive 2 | 7 | RA | 66.2 | 0 | 116 | ||
| Slurry + commercial additive 3 | 7 | RA | 67.0 | 1 1 | 116 | ||
| Slurry + commercial additive 4 | 7 | RA | 61.2 | 7 | 116 | ||
| Slurry + ferric chloride | 7 | 0.4 | 47.2 | 23 | 116 | ||
| Slurry + ferric chloride | 7 | 0.9 | 43.9 | 7 | 116 | ||
| Slurry + ferric chloride | 7 | 1.1 | 20.2 | 54 | 116 |
EFs = emission factors; NA = not applicable; RA = recommended application rate; TAN = total ammoniacal nitrogen.
1Values represent an increase in EF.
Regarding NH3 abatement measures, the application of chemical amendments (i.e. sulphuric acid, acetic acid, alum and ferric chloride) to slurry during storage delivered the greatest emission reduction of up to 98%. The amendments of slurry with industrial and agricultural waste and by-products (e.g. brewers grain, apple pulp, maize silage effluent, sugarbeet molasses) also showed promising NH3 abatement potential. However, greater reductions in EFs were observed through slurry acidification with commercial acids (i.e. acetic acid, sulphuric acid) and other acidifying additives (i.e. alum and ferric chloride) relative to the industrial/agricultural wastes and by-products. This is possibly due to the direct effect of acidification and chemical additives in reducing slurry pH. The impact of industrial waste products appeared to be influenced by their inclusion rate. On the whole, the greater the inclusion rate, the greater the NH3 emission reduction potential.
Treatment of slurry using commercial additives generally tended to increase NH3 emissions. Even though the mode of action and the composition of commercial additives are often not revealed by manufacturers, the proposed benefits of these commercial additives are often attributed to the role they play in stimulating biological processes and altering the chemical composition (i.e. TAN, NH4 + and pH) in slurries/manures. In the case of biological additives (i.e. microbial-based), NH3 mitigation may be achieved either through (i) stimulation of NH4 + immobilisation which reduces NH4 + concentration in slurry or (ii) enhancement of fermentation which reduces slurry pH through the formation of organic acids. While chemical additives, typically acidifiers (e.g. sulphuric acid, hydrochloric acid) mitigate NH3 losses by directly inducing a decline in pH, physical additives (e.g. biochar, clinoptilotite, peat) on the other hand act by adsorbing NH4 + onto their surfaces.
Although important reductions in NH3 emissions have been reported in the literature with these additives during slurry storage, there is also evidence of either no effect (Matulaitis et al., 2013; Provolo et al., 2016; Owusu-Twum et al., 2017) or some increases (Van der Stelt et al., 2007; Wheeler et al., 2010) in NH3 emissions after application of some additives. Studies (Wheeler et al., 2010; Matulaitis et al., 2013; Brennan et al., 2015) also reveal that the use of these additives might lead to an increase in other pollutant gases or pollution swapping during storage or after land application which can undermine the sustainability of their use. Therefore, commercial additives should be used cautiously and based on empirical evidence that their use is not accompanied with any negative consequences on the environment.
In addition to the use of slurry amendments and acidification, emissions from slurry storage can also be reduced through slurry management and engineering solutions. Practices with the potential to abate emissions include covering slurry storage tanks, inducing or allowing natural crust formation on slurry surfaces, reducing disturbances in storage tanks such as aeration or mixing, reducing the length of housing to reduce the surface area-to-volume ratio and spreading earlier in the year to reduce manure storage length (Kupper et al., 2020).
Slurry landspreading
Landspreading of slurry is a practice common in livestock production systems and offers an opportunity to reutilise the nutrients present in the slurry. In areas of intensive livestock production, slurry nutrients can be used to supply all or a substantial amount of the nutrients needed on farms and consequently reduce the share of mineral fertilisers utilised in agriculture. The most common slurry-spreading technique in Ireland is broadcast application via splash plate although other techniques such as band spreading/trailing hose, trailing shoe and injection have been introduced as low-emission spreading techniques. Spreading of cattle slurry generates almost 37 kt of NH3 per annum, representing approximately 30% of NH3 emissions in Ireland (EPA, 2022). Figure 3A shows the mean EFs generated from low-emission spreading strategies (LESS) and the splashplate technique. The mean EFs generated for the LESS and splashplate were 50% TAN and 63% TAN, respectively. Therefore, the splashplate EF obtained in this meta-analysis is similar to that used by the Irish EPA in the NH3 inventory (64% TAN). Even though the use of LESS led to reductions of approximately 21% in EFs relative to the splashplate, greater reductions were expected. Indeed, the NH3 abatement potential used by the Irish Informative Inventory prepared by the EPA for LESS methods such as bandspreading and trailing shoe are 30% and 60%, respectively. The lower reductions observed in the presented meta-analysis may be attributed to (i) low level of data availability, (ii) high variability of results and (iii) methodology used to obtain data, which was based mostly on small-scale plot work with hand-simulated slurry applications. Therefore, there is a need to conduct further studies to generate data from large-scale plots evaluating LESS techniques such as bandspreading and trailing shoe using standard farm machinery in order to validate the findings of this review.
Ammonia EFs for (A) slurry application techniques and (B) slurry application techniques at different seasons of application. EFs = emission factors. Spring = March, April and May. Summer = June, July and August. Autumn = October. LESS = Low-emission spreading strategies. Bars represent standard errors of the mean.
Figure 3B shows EFs generated from the splashplate and LESS in the spring (March, April and May), summer (June, July and August) and autumn (October). As expected, EFs were highest in the summer months relative to the spring and autumn months irrespective of the slurry application method used, which confirms the positive correlation between NH3 emissions and ambient temperature. The EFs generated using the splashplate were 60% TAN, 70% TAN and 52% TAN for spring, summer and autumn seasons, respectively. In the case of LESS, EFs generated in the spring and summer seasons were 48% TAN and 64% TAN, respectively. There is no apparent information regarding EFs generated using the LESS in the autumn season in Ireland and this needs to be considered in future studies. The above observation indicates the potential to reduce EFs to a great extent by applying slurries in cool temperature conditions rather than in warm weather, ideally accompanied by LESS.
Table 5 summarises NH3 EFs after landspreading of slurry under Irish conditions. Studies which assessed NH3 abatement potential of LESS methods such as bandspreading and trailing shoe reported emission reduction of up to 45% relative to the splashplate. While the practice of reducing slurry pH either through acidification or the use of chemical amendments led to a reduction in EFs of up to 96% relative to the untreated slurry, slurry separation with sieves on the other hand led to emission reduction of up to 67% relative to the un-sieved or un-separated slurry. Slurry separation results in a low DM slurry which facilitates slurry infiltration after soil application and consequently reduces NH3 emissions (Owusu-Twum et al., 2017). The reduction in EFs in the sieved or separated slurries can be related to enhanced infiltration of slurry in low DM slurries relative to high DM slurries even though some contradictory evidence, particularly regarding the spreading of low DM digestate, has been observed by Pedersen et al. (2021). In addition, the application of slurry in the evening using the splashplate and trailing shoe led to a mean reduction in EFs of 70% and 67%, respectively, relative to the daytime. This observation can be attributed to the lower solar radiation and less wind at night. In general, NH3 emission factors obtained after slurry landspreading from Irish-based studies vary to a great extent with values ranging from 0.8% to 108% of TAN applied (Table 5). These variations can be attributed to variations in factors such as slurry N application rates, slurry DM content, treatment techniques, season/time of application, slurry-spreading technique, soil type and environmental conditions at the time of application.
Ammonia EFs as percentage (%) of TAN applied from slurry-spreading techniques in Ireland
| Mitigation strategy | Spreading method | Time of application | Slurry type | Dry matter (%) | Soil texture | Sward height (high/short) | NH3 Loss (% TAN) | % change in EF relative to control | Measurement method | Reference | ||
|---|---|---|---|---|---|---|---|---|---|---|---|---|
| 1. Application technique and time of application | SP (C) | Day spreading | Cattle | 7.6 | Loam | Short grass | 54.5 | NA | IHF | Dowling (2012) | ||
| TS | Day spreading | 7.6 | Short grass | 35.0 | 36 | |||||||
| SP (C) | Day spreading | 7.6 | High grass | 49.6 | NA | |||||||
| TS | Day spreading | 7.6 | High grass | 27.5 | 45 | |||||||
| SP (C) | Evening spreading | 7.6 | Short grass | 24.5 | NA | |||||||
| TS | Evening spreading | 7.6 | Short grass | 16.0 | 35 | |||||||
| SP (C) | 12 July 2006 | 8.3 | Loam | Short grass | 61.5 | NA | IHF | Dowling et al. (2008) | ||||
| TS | 12 July 2006 | 8.3 | Short grass | 44.2 | 28 | |||||||
| SP (C) | 14 May 2007 | 8.3 | Short grass | 73.1 | NA | |||||||
| TS | 14 May 2007 | 8.3 | Short grass | 43.7 | 40 | |||||||
| SP (C) | March | Cattle | 6.0 | Loam | 68.0 | NA | ALFAM model | Cahalan et al. (2015) | ||||
| BS | March | 6.0 | 39.3 | 42 | ||||||||
| SP (C) | June | 7.1 | 72.3 | NA | ||||||||
| BS | June | 7.1 | 41.8 | 42 | ||||||||
| SP (C) | October | 5.7 | 55.6 | NA | ||||||||
| BS | October | 5.7 | 32.0 | 42 | ||||||||
| SP (C) | March | 7.6 | 62.0 | NA | ||||||||
| BS | March | 7.6 | 35.7 | 42 | ||||||||
| SP (C) | June | 7.1 | 104.7 | NA | ||||||||
| BS | June | 7.1 | 60.5 | 42 | ||||||||
| SP (C) | October | 6.5 | 63.5 | NA | ||||||||
| BS | October | 6.4 | 36.7 | 42 | ||||||||
| SP | March | 5.6 | Sandy clay loam | 35.9 | NA | |||||||
| SP | June | 5.9 | 51.2 | NA | ||||||||
| SP | October | 3.2 | 37.8 | NA | ||||||||
| SP (C) | April | Grass fed | 7.4 (HDM) | Well-drained coarse loam | NA | 58.0 | NA | IHF | Bourdin et al. (2014) | |||
| SP (C) | April | Grass fed | 3.8 (LDM) | 84.2 | NA | |||||||
| SP (C) | April | Maize fed | 7.3 (HDM) | 41.3 | NA | |||||||
| SP | April | Maize fed | 4.0 (LDM) | 79.2 | NA | |||||||
| TS | April | Maize fed | 4.5 (LDM) | 51.6 | 21 | |||||||
| SP (C) | July | Grass fed | 7.4 (HDM) | 108.2 | NA | |||||||
| SP (C) | July | Maize fed | 7.6 (HDM) | 93.5 | NA | |||||||
| TS | July | Maize fed | 4.8 (LDM) | 88.2 | 13 | |||||||
| SP (C) | August | Grass fed | 5.3 (HDM) | 54.1 | NA | |||||||
| SP (C) | August | Grass fed | 2.1 (LDM) | 103.0 | NA | |||||||
| SP (C) | August | Maize fed | 6.0 (LDM) | 63.7 | NA | |||||||
| SP (C) | August | Maize fed | 3.3 (LDM) | 95.8 | NA | |||||||
| TS | August | Maize fed | 3.5 (LDM) | 59.6 | 25 | |||||||
| 2. Chemical amendment | BS | Sandy loam | NA | DC + AT | Brennan et al. (2015) | |||||||
| Slurry (C) | 10.1 | 84.0 | NA | |||||||||
| Alum | 9.4 | 3.0 | 96 | |||||||||
| FeCl2 | 10.1 | 18.0 | 79 | |||||||||
| Lime | 8.2 | 40.0 | 52 | |||||||||
| Poly aluminium chloride (PAC) | 9.6 | 13.0 | 85 | |||||||||
| 3. Separation and acidification | SP | Cattle | NA | Clay loam | NA | DC + AT | Frost et al. (1990) | |||||
| Slurry (C) | 9 May | 19.1 | NA | |||||||||
| SS sieved 5 mm | 9 May | 13.4 | 30 | |||||||||
| SS sieved 3 mm | 9 May | 11.5 | 40 | |||||||||
| SS sieved 0.25 mm | 9 May | 9.9 | 48 | |||||||||
| SS sieved 0.015 mm | 9 May | 6.7 | 64 | |||||||||
| Acidified slurry (C) | 9 May | 1.7 | NA | |||||||||
| Acidified SS sieved 5 mm | 9 May | 2.3 | 35 1 | |||||||||
| Acidified SS sieved 3 mm | 9 May | 3.5 | 106 1 | |||||||||
| Acidified SS sieved 0.25 mm | 9 May | 1.6 | 6 | |||||||||
| Acidified SS sieved 0.015 mm | 9 May | 1.5 | 12 1 | |||||||||
| Slurry (C) | 27 June | 23.1 | NA | |||||||||
| SS sieved 5 mm | 27 June | 23.7 | 3 1 | |||||||||
| SS sieved 3 mm | 27 June | 21.8 | 6 | |||||||||
| SS sieved 0.25 mm | 27 June | 15.2 | 34 | |||||||||
| SS sieved 0.015 mm | 27 June | 13.7 | 41 | |||||||||
| Acidified slurry (C) | 27 June | 5.5 | NA | |||||||||
| Acidified SS sieved 5 mm | 27 June | 5.3 | 4 | |||||||||
| Acidified SS sieved 3 mm | 27 June | 5.8 | 5 1 | |||||||||
| Acidified SS sieved 0.25 mm | 27 June | 3.0 | 45 | |||||||||
| Acidified SS sieved 0.015 mm | 27 June | 3.7 | 33 | |||||||||
| Slurry (C) | 15 August | 12.9 | NA | |||||||||
| SS sieved 5 mm | 15 August | 6.9 | 47 | |||||||||
| SS sieved 3 mm | 15 August | 7.4 | 43 | |||||||||
| SS sieved 0.25 mm | 15 August | 4.2 | 67 | |||||||||
| SS sieved 0.015 mm | 15 August | 1.1 | 91 | |||||||||
| Acidified slurry (C) | 15 August | 1.9 | NA | |||||||||
| Acidified SS sieved 5 mm | 15 August | 1.8 | 5 | |||||||||
| Acidified SS sieved 3 mm | 15 August | 1.3 | 32 | |||||||||
| Acidified SS sieved 0.25 mm | 15 August | 1.3 | 32 | |||||||||
| Acidified SS sieved 0.015 mm | 15 August | 0.8 | 58 |
Control treatments are assigned the letter C in brackets.
ALFAM = ammonia loss from field-applied manure; AT = acid trap; DC = dynamic chamber; EFs = emission factors; HDM = high dry matter; IHF = integrated horizontal flux method; LDM = low dry matter; NA = not applicable; PA = phosphoric acid; SP = splashplate; SS = separated slurry; TAN = total ammoniacal nitrogen; TS = trailing shoe; TN = total nitrogen.
1Values represent an increase in EF.
Synthesis and challenges
Analysis of the data summarised in this paper indicates that in storage, manure amendments, particularly those that induce a decline in slurry pH, are effective at abating NH3 emissions considerably. Waste products generated from on-farm practices and food processing such as silage effluent and dairy washings may also be used as alternatives to chemical acids to reduce slurry pH (through the formation of lactic acid) and consequently reduce NH3 emissions. While the use of such wastes may support the circular economy concept, their use may have limited benefits in terms of mitigating greenhouse gas emissions (Kavanagh et al., 2021). Furthermore, commercial additives reviewed in this study did not show an ability to reduce NH3 emissions as observed in other studies (Matulaitis et al., 2013). Hence, the selection and use of commercial additives and waste amendments should be done based on scientifically proven evidence that (i) significant reductions in NH3 emissions will be guaranteed and (ii) the additives will not lead to pollution swapping during storage or at a different stage in the manure management continuum. This review also highlights that the practice of keeping concrete yards used by cattle clean from excreta particularly through pressure washing with water at short intervals should be encouraged on farms as the above practice has the potential to reduce emissions greatly from concrete yards used by cattle (Burchill et al., 2019).
Regarding the use of mineral fertiliser, our analysis shows that NH3 emissions from the application of urea fertilisers to grasslands can be reduced appreciably through the use of urease inhibitors. Indeed, Teagasc’s NH3 MACC shows that the use of urease inhibitors such as NBPT can provide a reduction in national NH3 emissions of approximately 20%. Similarly, the addition of urease inhibitor to urea is considered the most feasible solution to abate NH3 from urea in other countries such as Germany (Hu & Schmidhalter, 2021). The UK has proposed an even stronger NH3 abatement solution in the form of a legislative ban on urea fertilisers (DEFRA, 2020). However, a ban on urea fertilisers implies that farmers may have to resort to other forms of mineral fertilisers such as ammonium nitrate. A shift to ammonium nitrate fertilisers will also pose a threat to the environment as ammonium nitrate fertilisers could stimulate the release of N2O and increase the risk of nitrate leaching. Nitrous oxide emissions from mineral fertiliser usage particularly from urea can be curtailed through the addition of urease inhibitors.
The adoption of abatement strategies upstream in manure management, such as covering slurry stores, slurry acidification and treatment of slurry by amendments during slurry storage will result in more N being retained in the slurry. High levels of slurry N imply that more N could be lost as NH3 downstream or during landspreading of slurry. Therefore, low-emission spreading strategies (LESS) will play a vital role in abating NH3 emissions in animal agriculture. Even though EFs summarised in the present study for slurry application techniques varied to some extent, Irish studies which compared EFs from LESS and the conventional splashplate showed that appreciable reduction in EFs can be achieved by substituting the splashplate technique with LESS such as trailing shoe and band spreading. This review highlights the positive impact of LESS in abating NH3 EFs is consistent with Teagasc’s NH3 MACC, which shows that the implementation of LESS can deliver a reduction in national NH3 emissions of approximately 60%. In France, it is estimated that slurry injection and direct incorporation of slurry will contribute to 60% of the total abatement potential (Mathias & Martin, 2013). Similarly, low-emission manure application is expected to form approximately 60% of the total technical abatement potential in Germany (Wulf et al., 2017). Research conducted under Irish conditions shows that LESS does not only have the potential to reduce NH3 but also increases the fertiliser value of slurry, if applied at the right time (Lalor et al., 2011). Thus, the implementation of LESS should be a sustainable practice since its adoption could reduce the demand for mineral fertilisers and consequently reduce the proportion of greenhouse gas emissions associated with the production of mineral fertilisers. The reduction in the use of mineral fertilisers such as urea fertilisers could further reduce NH3 emissions.
This review showed that important reductions in EFs after landspreading can be achieved within application timing management (Table 5). For instance, Dowling (2012) reported appreciable reductions in EFs after evening spreading of slurry relative to daytime spreading under Irish conditions, irrespective of the slurry application method (splashplate and trailing shoe). Similarly, studies conducted in France and Denmark show that slurry application between evening and early morning has the potential to reduce emissions by up to 50% (Moal et al., 1995; Sommer & Olesen, 2000). The lower reductions after evening spreading relative to daytime spreading can be attributed to factors such as higher humidity, lower temperatures and lower wind velocity at night. These findings indicate that additional reductions in EFs can be achieved within application timing management without imposing high technology adoption costs on farmers.
Prediction of NH3 emissions in Ireland using the ammonia loss from field-applied animal manure (ALFAM) model shows higher emissions in the summer months relative to spring, autumn and winter (Lalor & Lanigan, 2010). Irish-based studies summarised in this review show that considerable emissions can also occur in the spring months after slurry landspreading and urea fertiliser application. Indeed, environmental conditions which stimulate NH3 emissions such as high wind speed and solar radiation have been observed in the spring months in Ireland. Thus, besides the restrictions in the time/month of slurry application in Ireland, the management of slurry and mineral fertiliser applications based on weather conditions within each month will also be crucial in reducing NH3 emissions in Irish livestock production systems.
A major challenge associated with NH3 abatement in livestock production is the already mentioned potential increase in NH3 emissions downstream after the implementation of an abatement strategy upstream. Besides the adoption of LESS, the above challenge can be mitigated by reducing the amount of N excreted from livestock through a reduction in the crude protein content in livestock diets. According to Bittman et al. (2014), a percentage point reduction in the protein content of animal feed content results in total NH3 emission (from animal housing, manure storage and the application to land) reduction by 5–15% as a result of the reduced ammoniacal N in the manure produced. The amount of N excreted from livestock can also be reduced through genetic improvements, particularly the selection of livestock with high feed efficiency (De Verdal et al., 2013).
Another challenge is the pollution swapping potential of NH3 abatement strategies. Emmerling et al. (2020) used a meta-analysis to evaluate the impact of major agricultural management practices on NH3 and its pollution swapping effect. The result showed that for treatment, storage and application stages, only slurry acidification was effective at reducing NH3 emissions with no pollution swapping effect with greenhouse gases such as N2O, methane (CH4) and carbon dioxide (CO2). Even though the rest of the management techniques evaluated in the meta-analysis such as slurry separation, different storage types, biological treatment and variable field applications were also effective at abating NH3 at varying degrees, they led to an increase in the emission of at least one other greenhouse gas. These findings indicate that the issue of pollution swapping could be addressed through slurry acidification or by implementing a combination of abatement strategies.
While LESS is the main strategy for abating NH3 in Ireland, slurry acidification is the preferred technology for reducing NH3 from the animal sheds to the landspreading stage in Denmark. Danish regulation of nutrients in agriculture requires high levels of nitrogen-use efficiency from slurry (MEFD, 2017), thus making it important to conserve ammoniacal nitrogen in the slurry before application to land. Slurry acidification is rarely used in Ireland and other EU countries owing to issues such as its associated cost as well as health and safety and soil health concerns.
The adoption of LESS or other landspreading abatement technologies may be limited on Irish farms due to its associated cost. This may be the case, particularly for farmers in low-income categories who have already invested in splashplate systems. The provision of incentives especially in the form of subsidies to farmers in the above category to cover the cost of implementation of the technology can facilitate the adoption of LESS on Irish farms. According to the EPA (2022), the adoption rate of LESS increased from 3% in the year 2015 to 36% in the year 2020. The adoption rate of LESS is expected to further increase in the coming years due to the restrictions imposed by the current Nitrates Action Programme in Ireland which requires all farmers operating above 170 kg N/ha to use LESS on their farms.
For further studies
Although a number of studies have been conducted on NH3 mitigation strategies under Irish climatic conditions, there is the need to investigate novel strategies for abating NH3 emissions to add to the suite of options already available to Irish agriculture.
Most of the Irish studies which quantified NH3 emissions from slurry landspreading were conducted using small-scale plots with hand-simulated slurry applications. Additional studies are required to generate data from large-scale plots evaluating different slurry-spreading techniques available to Irish agriculture using standard farm machinery in order to validate the findings of this review.
The application of urease inhibitors to manure or urea-based mineral fertilisers is regarded as a potent strategy to abate NH3 emissions in agriculture. Although protected urea offers a low NH3 emission alternative to straight urea (Forrestal et al., 2016; Krol et al., 2020), the potential to further utilise urease inhibitors to reduce emissions from livestock wastes has not yet been made commercially feasible (Sigurdarson et al., 2018). The reason for the limited adoption in animal agriculture is attributed to factors such as the absence of efficient and automated application systems, the need for reapplication, limited chemical stability of the inhibitors and the associated cost to the farmer (Sigurdarson et al., 2018). Further studies are required to improve the above aspects to facilitate the adoption of urease inhibitors on a commercial scale.
Slurry separation is used in some countries as a tool to improve manure management on farms. The technique is often employed in situations where there is the need to export slurry nutrients from farms with excess nutrients to those in deficit of slurry nutrients. The technique also facilitates the transport of organic matter in slurry over long distances to large/central biogas plants. Research shows that the application of the low DM liquid fraction obtained after separation to soil can enhance infiltration of slurry and TAN and consequently minimise NH3 losses (Owusu-Twum et al., 2017). Unlike other European countries such as Spain, Italy and Denmark where the technique has been studied to a great extent, there is limited research findings on the impact of the above technique in Irish agricultural production systems. Slurry separation could offer a cost-effective approach to improve manure management and facilitate the production of biogas which is still in its infant stages in Ireland. An assessment of the environmental (i.e. NH3, greenhouse gas emissions and nutrient leaching), economic and agronomic consequences of the above technique under Irish agricultural production systems is necessary in order to recommend the technique to farmers.
While anaerobic digestion (AD) is regarded as an environmentally friendly approach for reutilising the DM present in manure, the resulting effluents obtained after the AD process (i.e. digestate) might lead to increased NH3 emissions either during storage and/or landspreading due to the increase in effluent pH and TAN after AD. A review of the literature shows limited findings regarding the environmental impacts (i.e. NH3 and greenhouse gas emissions) of applying digestates obtained from AD of slurry to Irish soils. Further studies are required to adequately document the relative benefits of AD under Irish environmental conditions.
Recent research findings reveal the potential application of algae as a biofertiliser (Baweja et al., 2019; Al-Myali, 2021; Zou et al., 2021). These organisms could also play a critical role in managing nutrients in slurry prior to landspreading as algae can be cultured using the elements (e.g. nitrogen, carbon and phosphorus) in slurry/manure as substrates. The increasing cost of fertilisers and the need to identify low-emission alternatives to land application of slurry could result in this being a potentially viable fertiliser in the future. Nevertheless, information regarding the environmental benefits and economic viability of the above strategy is still lacking in the literature and needs to be investigated.
In order to facilitate the adoption of slurry acidification in Ireland, research is required to generate ample scientific data, particularly regarding issues such as its NH3 abatement potential, associated cost, impact on other pollutant gases from slurry as well as ways to address its health and safety concerns.
Even though this review highlights the potential to abate NH3 emissions during slurry storage through the use of slurry amendments, there is limited information regarding their impact on other N loss pathways such as N2O and NO3 − leaching under Irish conditions and should be studied. As previously mentioned, the majority of dairy system in Ireland are slurry-based. As a result, the slurry-based system has been the focus of most research on NH3 mitigation. Information regarding NH3 mitigation techniques from the straw-based system where excreta is collected as farmyard or solid manure is rare and needs to be investigated.
Limitations of the study
The main limitations in the study is the low number of samples and large data variability. This resulted in our data having a low statistical power to detect effects. For example, in sections relating to slurry storage, EFs were generated from small-scale incubations which measured NH3 in a dynamic chamber mode which means that the airflow is expected to be higher than above a typical tank in the shed. Thus, EFs generated from the storage stage in the present study can only be used to provide an indication of the efficacy of abatement strategies and cannot be used to provide reliable EFs as they may overestimate NH3 emissions. Further studies are required to generate additional datasets in order to perform a comprehensive assessment of NH3 EFs generated from cattle production systems in Ireland and the efficacy of NH3 abatement strategies available to Irish Agriculture.
Conclusion
This review provides an overview of NH3 EFs in Ireland. EFs generated from the deposition of excreta during grazing, housing/storage and slurry landspreading varied considerably. The urease inhibitor NBPT was effective at reducing EFs after urea application to grassland. Acidifying additives or chemical amendments were effective at abating emissions from concrete yards or hard standings used by cattle and during slurry storage. Emissions from landspreading can be reduced within daily application timing management. Studies which compared LESS with the splashplate generally reported lower NH3 EFs for the LESS compared to the splashplate. Further studies are required to generate additional data to validate the findings from this review. Further research is also needed to evaluate the impact of these NH3 abatement strategies on other forms of N such as N2O and NO3 to ensure that the abatement of one form of N does not lead to pollution swapping either at the same stage or at a different stage in animal agriculture. Finally, research on the efficacy of new abatement strategies and mitigation options for solid manure systems in Irish agriculture is also required.