How do emission rates and emission factors for nitrous oxide and ammonia vary with manure type and time of application in a Scottish farmland?
Introduction
Manures and slurries provide a significant nitrogen (N) input to agricultural land. In 2011 the total N excreted by livestock in the EU was 9.2 Tg which is only 15% less than the N added by synthetic fertilisers (Velthof et al., 2015). The large quantity of nutrients present in manures means that they are commonly applied to agricultural land to recycle N, phosphorus and potassium for plant growth (Defra, 2010). However, considerable amounts of the applied manure N will not be utilised by crops as a result of nitrification and denitrification, and the subsequent emissions of nitrous oxide (N2O) (Chadwick et al., 2011), dinitrogen (N2) (Cardenas et al., 2007), and ammonia (NH3) (Misselbrook et al., 2005a). Leaching of nitrate (NO3−) into groundwater and surface waters leads to further N loss from the soil (Rodhe et al., 2006) and other environmental impacts including eutrophication and soil acidification.
Globally, agricultural soil is responsible for 65% of N2O emissions (Reay et al., 2012), a greenhouse gas (GHG) approximately 300 times more powerful than CO2, that is also responsible for stratospheric ozone layer depletion (Stocker et al., 2013). In the UK it is estimated that 73% of anthropogenic N2O emissions and 92% of NH3 emissions are from agricultural sources, including direct emissions from soils, animal wastes and manure stores (Dore et al., 2008, Skiba et al., 2012). Indirect N2O emissions also result from deposition of volatilised NH3 and NO3− leaching and transport in aquatic and terrestrial environments.
The potential for N2O and NH3 emission after manure applications to agricultural soil is dependent on a combination of manure properties and environmental conditions. High temperatures, high wind speed and low rainfall immediately following manure application promote NH3 emissions from manures containing a high amount of readily available N (Meisinger and Jokela, 2000, Misselbrook et al., 2005a), meaning that the timing of application can be critical if significant losses of N from the soil are to be avoided. Conversely, loss of N via N2O emissions is higher when manure is applied in wet conditions as N2O production via denitrification will occur before the crop is able to utilise the available N. Nitrate leaching will also occur if excess rainfall and drainage take place between manure application and crop N uptake (Defra, 2010, Shepherd and Newell-Price, 2013). It is generally recommended therefore to apply manures when crops are actively growing and removing N from the soil (Granli and Bøckman, 1994, Meisinger and Jokela, 2000, Defra, 2010).
In the UK, manure application in autumn and winter is restricted by Nitrate Vulnerable Zone (NVZ) regulations to decrease NO3− pollution of aquatic environments. Expansion of these measures to other areas could assist in decreasing indirect N2O emissions from NO3− leaching and direct N2O emissions from denitrification if application in wet conditions is avoided. The time of application should aim to provide a balance between the need to apply manure during the period of maximum crop N requirement, and the need to reduce seasonal climate effects on emissions (Meisinger and Jokela, 2000). Reducing losses of N from the soil is also beneficial for crop growth as more N is available for use by the growing crop (Rodhe et al., 2006, Shepherd, 2009).
The magnitude of N2O and NH3 emissions generated from manures is also dependent on their total-N content and the proportion present as readily available N (ammonium-N and uric acid-N), which varies with manure type (Defra, 2010, Shepherd and Newell-Price, 2013). Large quantities of readily available N (35–70% of total N) are typically found in slurries and poultry manures, compared to only 10–25% of total N in farmyard manure (FYM) (Defra, 2010). Manures containing large amounts of readily available N have a higher probability of losing N via NH3 volatilization (Misselbrook et al., 2005a), N2O production (Chadwick et al., 2011), or as a result of NO3− leaching (Chambers et al., 2000, Dampney et al., 2000, Shepherd, 2009). Manure moisture content can also affect N2O emissions, as an increase in soil moisture can enhance the production of N2O, with greatest N2O emissions most likely to occur between 50 and 70% WFPS (Flechard et al., 2007). Slurry typically has a moisture content of > 90%, increasing the risk of high N2O emissions after application (Jørgensen et al., 1998). The moisture content of manures can also affect NH3 emission rate, and slurries with higher moisture contents are generally associated with lower NH3 emissions as they rapidly infiltrate into the soil, with the majority of the emission typically occurring in the 12 h post-application (Sommer and Hutchings, 2001). Poultry litter, in contrast, has a much lower moisture content and a lower initial loss of NH3, but emissions occur over a longer timescale as uric acid is broken down and urea hydrolysed to NH4+ (Meisinger and Jokela, 2000, Jones et al., 2007). It has also been suggested that the C:N ratio of organic manures may affect N losses from soil. Akiyama et al. (2004) argue that higher C:N ratios in manure compared with inorganic chemical fertilisers provide optimum conditions for denitrification. The high C contents of organic manures (typically 35% organic C), can also stimulate microbial activity, thereby creating anaerobic zones in the soil that allow denitrification and N2O production to occur at a lower % WFPS than for chemical fertilisers (Akiyama et al., 2004). Incorporation of manures into the soil immediately after application, and the method of slurry application can also influence the extent of N2O and NH3 emissions (Webb et al., 2010). However, the use of these methods and their degree of success will depend on the presence/stage of crop growth.
The amount of N2O or NH3 emitted from N sources applied to soils is often calculated using an emission factor (EF), which defines the quantity of N2O or NH3 emitted as a proportion of the total N applied. The UK currently uses the IPCC's Tier 1 EF in its national N2O inventory, where N2O emissions from soils receiving organic amendments are equal to 1% of the total N applied (IPCC, 2006), with no accounting for locally variable factors such as soil type or climate, variations in manure type, or the time of application. The IPCC default EF for NH3 emission following manure application to land is 20% of the applied N. However, the EF used to estimate NH3 emissions from manure application in the UK NH3 emissions inventory is derived from an empirical model taking account of manure type and some soil and climatic factors (Nicholson et al., 2013).
The variety of conditions affecting N loss from soils amended with livestock manures mean it is imperative that applications are carefully managed to avoid significant environmental pollution. It is vital to understand how the form and time of application may affect environmental impacts. The results of the research presented in this paper which forms part of a nationwide project, will contribute to reducing uncertainty in the UK's agricultural GHG inventory, and will enhance the sustainability and GHG mitigation potential of farming systems (GHG, 2013). This study aimed to compare soil N2O and NH3 emissions and EFs following autumn and spring manure applications to arable land in Scotland. Nitrous oxide and NH3 emissions were measured for all manure types following application in both seasons, and the suitability of the IPCC Tier 1 EFs to represent N2O and NH3 emissions from different manure types and seasons of application was assessed. Effects of the timing and form of manure application on crop yield and crop N uptake were also investigated, to assess the impact of the type and time of manure application on crop production.
Section snippets
Site description and experimental design
Two 12 month field experiments were undertaken at Boghall farm (NT 248653, 190 m elevation), in East-central Scotland in 2012/2013. Both experiments were located in the same field, on a sandy loam soil (pH 6, 6% OM), with a 30 year (1980–2009) site mean annual precipitation of 979 mm and mean daily temperature in July and January of 14.3 °C and 3.3 °C, respectively. Spring barley (Hordeum vulgare) had been grown in the field for the previous four years. The site was one of a network of UK sites
Weather, soil moisture and daily N2O emissions
As the majority of N2O emissions are thought to occur in the month immediately following application of an N source to soil (Dobbie et al., 1999), rainfall, temperature and soil moisture during this period were assessed, along with conditions throughout the annual experiments. Approximately twice as much rainfall (172 mm) was measured in the first 30 days of the autumn experiment compared to the spring experiment (82.4 mm), with a maximum daily rainfall of 48.6 mm in the first month of the autumn
Timing of application
Wetter conditions observed in the month of autumn applications compared to spring applications in this study reflects 30 year long-term average seasonal differences (78.3 mm: October; 42.5 mm: April), and suggests that livestock manure should be applied in spring if production of N2O is to be minimised. The observed relationship between N2O emissions, large rainfall events and increasing soil WFPS does though emphasise that it is the short-term weather after application that is the strongest
Conclusion
The results of this research demonstrate how manure type and the time of its application can influence N2O and NH3 emissions, and that the trade-off between N2O and NH3 emissions could be crucial in deciding on timing and method of application for different manure types. The variation in the extent of emissions from different types of manure demonstrates the effects of manure properties such as moisture content, total N and available N content on emission generation. Emissions of N2O were
Acknowledgements
We acknowledge assistance from colleagues in fieldwork and sample analysis, in particular the work carried out by Gail Bennett and Milly Bowden (ADAS), Gemma Miller, Maria Borlinghaus, John Parker, Chris Sillick (SRUC) and the SRUC farm and technician staff. Advice from members of the GHG inventory project including Dr. Catherine Watson (AFBI) is gratefully acknowledged. Funding for this work was provided by the UK Department for Environment, Food and Rural Affairs (Defra) AC0116, the
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