SPECIAL TOPICS — Mitigation of methane and nitrous oxide emissions from animal operations: II. A review of manure management mitigation options¹

Abstract

This review analyzes published data on manure management practices used to mitigate methane (CH₄) and nitrous oxide (N₂O) emissions from animal operations. Reducing excreted nitrogen (N) and degradable organic carbon (C) by diet manipulation to improve the balance of nutrient inputs with production is an effective practice to reduce CH₄ and N₂O emissions. Most CH₄ is produced during manure storage; therefore, reducing storage time, lowering manure temperature by storing it outside during colder seasons, and capturing and combusting the CH₄ produced during storage are effective practices to reduce CH₄ emission. Anaerobic digestion with combustion of the gas produced is effective in reducing CH₄ emission and organic C content of manure; this increases readily available C and N for microbial processes creating little CH₄ and increased N₂O emissions following land application. Nitrous oxide emission occurs following land application as a byproduct of nitrification and dentrification processes in the soil, but these processes may also occur in compost, biofilter materials, and permeable storage covers. These microbial processes depend on temperature, moisture content, availability of easily degradable organic C, and oxidation status of the environment, which make N₂O emissions and mitigation results highly variable. Managing the fate of ammoniacal N is essential to the success of N₂O and CH₄ mitigation because ammonia is an important component in the cycling of N through manure, soil, crops, and animal feeds. Manure application techniques such as subsurface injection reduce ammonia and CH₄ emissions but can result in increased N₂O emissions. Injection works well when combined with anaerobic digestion and solids separation by improving infiltration. Additives such as urease and nitrification inhibitors that inhibit microbial processes have mixed results but are generally effective in controlling N₂O emission from intensive grazing systems. Matching plant nutrient requirements with manure fertilization, managing grazing intensity, and using cover crops are effective practices to increase plant N uptake and reduce N₂O emissions. Due to system interactions, mitigation practices that reduce emissions in one stage of the manure management process may increase emissions elsewhere, so mitigation practices must be evaluated at the whole farm level.

manure, methane, mitigation, nitrous oxide

INTRODUCTION

Climate change associated with emissions of greenhouse gases (GHG) resulting from human activities is noted as the defining human development issue currently (UNDP, 2008). Through the production of feed, growth of productive animals and the supporting herd, and the disposal of animal waste, livestock production contributes 8 to 18% (depending on the estimation method used) of global GHG emissions (O'Mara, 2011) and emissions from animal agriculture are expected to increase 17% by the year 2020 (USEPA, 2006). Significant reduction of GHG emissions can be achieved through intensification of animal production and improved production technologies in less developed countries and the use of novel mitigation practices in modern production systems (Smith et al., 2007; O'Mara, 2011).

This article reviews the direct effect of animal manure mitigation practices on emissions of the major non–carbon dioxide (non-CO₂) GHG, methane (CH₄) and nitrous oxide (N₂O), focusing on experimental and field evaluation data and excluding for the most part modeling and simulation studies. The goal was to gather experimental information to support the development of whole farm analyses and life cycle assessments (LCA) necessary to identify opportunities for synergies and consequences of interactions derived from the use of these practices. Interactions among mitigation practices for individual components of livestock production systems are discussed in Gerber et al. (2013).

LIVESTOCK MANURE AND EMISSIONS

Animal manure is a nutrient resource containing most of the essential elements required for plant growth and can be a significant source of N in both intensive and subsistence production systems. Application of manure to cropland increases soil OM, microbial biomass, and mineralization rate (Spiehs et al., 2010; Langmeier et al., 2002) and improves a number of soil properties including soil tilth, water-holding capacity, oxygen content, and fertility; it also reduces soil erosion, restores eroded croplands, reduces nutrient leaching, and increases crop yields (Khaleel et al., 1981; Araji et al., 2001). Animal manure is an alternative to energy intensive and high-cost synthetic fertilizer and can be a very effective fertilizer source when the available nutrient content and mineralization rate are synchronized with crop nutrient uptake.

Unmanaged accumulation of organic wastes, however, presents environmental and health concerns for humans and animals. Concerns include leaching of nitrate (NO₃^–) and pathogens to ground water, unbalanced algal growth and eutrophication of surface water and deterioration of sensitive ecosystems, degradation of soil production potential through accumulation of nutrients, salts, and metals, and emission of gases considered a health and environmental risk (Gerber et al., 2005; Steinfeld et al., 2006; USEPA, 2011).

Manure management refers to all activities, decisions, and components used to handle, store, and dispose of feces and urine from livestock with the goal of preserving and recycling the nutrients in the livestock production system (Brandjes et al., 1996; IPCC, 2006a). This includes manure accumulation and collection in buildings, storage, processing, and application to cropland as well as deposition in pastures and rangelands in grazing systems.

Methane and Nitrous Oxide Emissions.

In ruminant production systems, enteric CH₄ production is the largest contributor to GHG emissions followed by CH₄ from manure and in beef feedlot systems, N₂O from pen surface, and N₂O emissions from soils. Emissions from nonruminant livestock systems are less than that of ruminants and are mostly CH₄ and N₂O from manure storage and land application (Hristov et al., 2013b).

The contribution of manure management to global GHG emission was estimated by Steinfeld et al. (2006) to be 2.2 Gt of CO₂ warming potential equivalents (CO₂e) per year. Methane emissions from manure storage were estimated to be 470 Mt CO₂e/yr in 2010 with an expected 11% increase by 2020 and N₂O emissions from fertilizer use, manure application, and deposition by grazing livestock were estimated at 2,482 Mt CO₂e/yr in 2010 with an expected 18% increase by 2020 (USEPA, 2006). Nitrous oxide emissions from soil application of manure are a major contributor to total GHG emissions from agriculture (Davidson, 2009) with animal waste representing 30 to 50% of the global agricultural N₂O emissions (Oenema et al., 2005). Both CH₄ and N₂O are powerful GHG with global warming potentials (GWP) of 25 and 298 kg CO₂e/kg, respectively (Solomon et al., 2007).

Most of the CH₄ emission from manure is produced under anaerobic conditions during storage with very little following land application. Manure produces less CH₄ when handled as a solid (e.g., in stacks or pits) or when deposited on pasture or rangelands (USEPA, 2005). Therefore, opportunities to reduce CH₄ emission are centered on preventing anaerobic conditions during storage or capturing and transforming the CH₄ that is produced, if anaerobic conditions are present. Data summarized by Chianese et al. (2009) indicate average CH₄ emissions from covered slurry, uncovered slurry, and stacked manure to be 6.5, 5.4, and 2.3 kg/m² per yr although rates vary with temperature and time in storage.

Direct emissions of N₂O from manure storage are small when compared with CH₄ emissions. For N₂O emissions to occur, manure must first be handled aerobically where ammonium (NH₄⁺) or organic N is converted to NO₃^– and nitrite (NO₂^–) during nitrification and then handled anaerobically where the NO₃^– and NO₂^– are reduced to elemental N (N₂), with intermediate production of N₂O and nitric oxide (NO) through denitrification (USEPA, 2010). Most of the N₂O resulting from manure is produced in manure-amended soils through microbial nitrification under aerobic conditions and partial denitrification under anaerobic conditions, with denitrification generally producing the larger quantity of N₂O (Tisdale et al., 1993; USEPA, 2010).

Losses of N₂O from the pen surface of open-lot dairy or beef feedlot facilities, however, can be significant. The fact that a large amount, up to 50%, of the excreted N by beef cattle is not recovered in manure has been well documented for various geographic locations (Loh et al., 2008; Cole and Todd, 2009). Most of these losses are as ammonia (NH₃), but N₂O emissions are also significant (Leytem et al., 2011; Rahman et al., 2013) and depend on a variety of factors, including surface conditions (Aguilar et al., 2011). Manure management practices in beef feedlots vary, but usually pens are cleaned when animals are marketed (i.e., several times a year) or once a year (Eghball and Power, 1994), which creates conditions for NH₃ and GHG emissions off the pen surface. Recently, Rahman et al., (2013) reported CH₄, CO₂, and N₂O emission rates from the pen surface of a North Dakota beef feedlot of 38 g, 17 kg, and 26 g/head per d, respectively. Somewhat lower N₂O emissions were reported for an open-lot dairy in southern Idaho: 0.13, 0.49, 28.1, and 0.01 kg/cow per d (NH₃, CH₄, CO₂, and N₂O, respectively; Leytem et al., 2011). Considering the much greater GWP of N₂O, compared with CH₄ (Solomon et al., 2007), the emissions of these 2 GHG were comparable in that and other studies (Boadi et al., 2004). Comparatively lower N₂O emissions from a free-stall dairy in the Texas panhandle were reported by Borhan et al. (2011): 836, 5,573, and 3.4 g/head per d (CH₄, CO₂, and N₂O, respectively). In the same study, emissions from a beef feedlot were reported as 3.8, 1,399, and 0.68 g/head per d, respectively. The pen surface was estimated to contribute about 84% of the aggregate N₂O emission in this study.

Manure contains most elements necessary for stimulating soil nitrification and denitrification processes that form N₂O. These processes are transient, depending on the amount and form of available N (NH₄⁺ or NO₃^–), soil oxidation–reduction potential, degradable C sources, soil temperature, water content, and microbial population (Cavigelli and Parkin, 2012). Denitrifying organisms can further reduce N₂O to N₂ at rates dependent on soil conditions, with multiple factors controlling the ratio of N₂O to N₂ produced. The fraction of N completely reduced to N₂ also increases as soil water content approaches saturation. Nitrous oxide can also be produced indirectly when manure N is lost through volatilization as NH₃, NO, and nitrogen dioxide (NO₂) and is nitrified and denitrified in soil following redeposition (USEPA, 2010).

Being a result of microbial processes, the emission of N₂O is highly variable as influenced by environmental and metabolic factors, which makes measurement of mitigation effects difficult. Nonetheless, results of adopting mitigation practices can be estimated using the potential N₂O emission reductions obtained when optimal conditions for nitrification and denitrification are assumed. This approach makes it possible to gauge the effect of mitigation practices and their interactions within the livestock production system.

Due to the nature of the antagonistic processes resulting in CH₄ and N₂O emissions (CH₄ is produced under anaerobic conditions whereas production of N₂O requires sufficient levels of oxygen), some practices that result in the reduction of CH₄ production increase N₂O emissions. An example is the aeration of manure during storage to reduce CH₄ emissions. This process may increase N₂O emissions when aeration rate is sufficient to create an aerobic environment.

Opportunities to reduce N₂O and CH₄ emissions from livestock manure are illustrated in Fig. 1 and 2 according to the process of formation and emission of CH₄ and N₂O and the flow of organic C and N during the animal production cycle. The review of manure CH₄ and N₂O mitigation practices based on experimental and commercial scale applications is summarized in Table 1. Data related to the mitigation practices presented in this review may not account for losses, conversions or other differences caused by preceding processes in the system, study duration, and climatic impacts such as winter and summer differences.

Figure 1.

The flow of organic carbon through the livestock production system including opportunities to mitigate methane from livestock manure.

Figure 2.

The flow of nitrogen and organic carbon through the livestock production system including opportunities to mitigate nitrous oxide from livestock manure.

Table 1.

Manure handling strategies offering non-CO₂ greenhouse gas mitigation opportunities

Category	Species1	Potential CH4 mitigating effect2	Potential N2O mitigating effect2	Potential NH3 mitigating effect2	Effective3	Recommended4
Dietary strategies
Reduced dietary protein	AS	?5,6	Medium	High	Yes (N2O and NH3)	Yes (N2O and NH3)
High fiber diets	SW	Low	High	?	Yes (N2O)	Yes (N2O)
Grazing management
Grazing intensity7	AR	?	High?7	?7	Yes (N2O)	Yes (N2O)
Housing
Biofiltration	AS	Low?	?	High	Yes (NH3 and CH4?)	Yes (NH3 and CH4?)
Manure system8	DC, BC and SW	High	?	High	Yes (CH4 and NH3)	Yes (CH4 and NH3)
Manure treatment
Anaerobic digestion	DC, BC and SW	High	High9	Increase?10	Yes (CH4 and N2O)	Yes (CH4 and N2O)
Solids separation	DC and BC	High	Low?9	?11	Yes (CH4)	Yes (CH4)
Aeration	DC and BC	High	Increase?12	?12	Yes (CH4)	Yes (CH4)
Manure acidification	DC, BC and SW	High	?13	High13	Yes (CH4 and NH3)	Yes (CH4 and NH3)
Manure storage
Decreased storage time	DC, BC and SW	High14	High14	High14	Yes (all)	Yes (all)
Storage cover with straw	DC, BC and SW	High	Increase?15	High	Yes (CH4 and NH3)	Yes (CH4)
Natural or induced crust	DC and BC	High	Increase?15	High	Yes (CH4 and NH3)	Yes (CH4)
Aeration during liquid manure storage	DC, BC and SW	Medium to High	Increase?12	?12	Yes (CH4)	Yes (CH4)
Composting	DC, BC and SW	High	?12	Increase12	Yes (CH4)	Yes (CH4)
Litter stacking	PO	Medium	N/A16	?	Yes (CH4)	Yes (CH4)
Storage temperature	DC and BC	High	?	High	Yes (CH4 and NH3)	Yes (CH4 and NH3)
Sealed storage with flare	DC, BC and SW	High	High	?17	Yes (CH4 and N2O)	Yes (CH4 and N2O)
Manure application method
Injection vs. surface application	DC, BC and SW	No effect to increase?	No effect to increase18	High	Yes (NH3)	Yes (NH3)
Timing of application	AS	Low	High19	High	Yes (N2O and NH3)	Yes (N2O and NH3)
Soil cover, cover cropping	AS	?	No effect to High20	Increase?21	Yes (N2O?)	Yes (N2O?)
Soil nutrient balance	AS	N/A	High	High	Yes (N2O and NH3)	Yes (N2O and NH3)
Nitrification inhibitor22
Applied to manure or after urine deposition in pastures	DC, BC and SH	N/A	High	N/A	Yes (N2O)	Yes (N2O)?
Urease inhibitor23
Applied with or before urine	DC,BC and SH	N/A	Medium?	High	Yes (N2O and NH3)23	Yes (N2O and NH3)?
Soil management
Soil condition, texture, water filled pore space, drainage	AR	?	Decrease or Increase24	?	Yes (N2O)?	Yes (N2O)?

¹DC = dairy cattle; BC = beef cattle (cattle include Bos taurus and Bos indicus); SH = sheep; GO = goats; AR = all ruminants; SW = swine; PO = poultry; AS = all species.

²High = ≥30% mitigating effect; Medium = 10 to 30% mitigating effect; Low = ≤10% mitigating effect. Mitigating effects refer to percent change over a “standard practice,” that is, study control that was used for comparison and are based on combination of study data and judgment by the authors of this document.

³Effectiveness is determined on the basis of greenhouse gas or NH₃ mitigation potential (in some cases, effects on feed intake and/or animal productivity were also considered).

⁴Based on available research or lack of sufficient research.

⁵? = uncertainty due to limited research or lack of data, inconsistent or variable results, or lack or insufficient data on persistency of the effect.

⁶Insufficient research. Modeling suggests that enteric CH₄ may increase. If rumen function is impaired, manure CH₄ emissions may increase.

⁷Reduced grazing intensity can reduce up to 50% N₂O emissions by lowering urinary N input to the soil but can also increase N₂O emissions by increasing residual OM during soil freeze-thaw cycles. Ammonia emissions may increase; magnitude is unclear.

⁸Generally, a manure system that reduces the time between feces and urine excretion and removing manure from the animal house will reduce NH₃ and CH₄ emissions from the building (data on N₂O are not conclusive and there is little N₂O emission from buildings and manure storage). Type of manure system also has an effect of greenhouse gas emissions; depending on type of storage, emissions can be greater from manure that is removed daily vs. manure that resides in the animal building for longer periods of time.

⁹Anaerobic digestion and solids separation reduce the source of degradable C in the manure applied to the soil and as a consequence reduce the potential for nitrification and N₂O emissions.

¹⁰Ammonia emissions will increase as more organic N is decomposed to NH₃; magnitude is unclear.

¹¹Reducing the solids content of manure facilitates infiltration and prevents NH₃ losses from land application.

¹²(See also Aeration during liquid manure storage and Composting.) In some cases, aeration of manure may stimulate the transformation of NH₃–N into NO₃–N, which increases the potential for N₂O emissions and reduces the potential for NH₃ volatilization. In most cases, composting increases NH₃ losses.

¹³Manure acidification decreases the fraction of volatile NH₃ in manure but may increase N₂O emissions after land application.

¹⁴Increasing the time manure is stored increases potential emissions by increasing the concentration of NH₃ and CH₄ through degradation of nutrients in manure and the time manure emissions occur, creating a compound effect varying widely in magnitude. This strategy is listed with High mitigation potential because not storing manure would be a total avoidance of storage greenhouse gases.

¹⁵Covering the manure with permeable porous covers promotes nitrification and N₂O emissions.

¹⁶N/A = not applicable.

¹⁷Covering the manure storage with impermeable covers prevents NH₃ volatilization during storage and increases the content of NH₃ in the manure, which generally results in greater of NH₃ emissions after land application.

¹⁸Injecting manure increases the soil organic C and N pools, inducing reduction conditions and promoting N₂O emissions; the magnitude of increase depends on soil moisture, texture, temperature and many other factors.

¹⁹Timing land application of manure to temperature and soil moisture conditions that are not conducive to nitrification and N₂O emissions can have a large but variable impact on emissions.

²⁰Cover crops can increase plant N uptake and decrease accumulation of NO₃^– and thus reduce N₂O production through denitrification, but the results on overall greenhouse gas emissions have not been consistent. Interactions with other soil conservation practices are significant (tillage system, for example) and must be considered when the goal of cover cropping is reducing whole-farm greenhouse gases emissions.

²¹Soil vegetation cover intercepts land applied manure, reducing infiltration and contact with soil, which results in greater NH₃ volatilization losses whose magnitude depends on the application method and manure characteristics.

²²Nitrification inhibitors have reduced N₂O emissions and N leaching in intensive dairy and sheep production systems of Australia and New Zealand.

²³Urease inhibitors work well before urea N in the urine is transformed to NH₃–N in manure; therefore, the opportunity for application may be limited to intensive feedlot production or waste handling systems, in which feces and urine are separated.

²⁴Soil moisture content affects the nitrification and denitrification potential, resulting in large variation in N₂O emissions.

There is still much to learn about the benefits of particular mitigation practices, the effect of combining mitigation practices, the response of environmental indicators such as nutrient conversion, volatilization, leaching, erosion, etc., and the effect on the environmental and financial performance of the production system (farm) as a whole. It is common for one environmental benefit of a technology to negatively impact processes in another area (Gerber et al., 2013).

Ammonia: An Important Component in Greenhouse Gas Mitigation.

Although not a GHG, NH₃ (and its ionized form, NH₄⁺) is an important component of the manure N cycle. Ammonium (a large fraction of the N in manure) is the first product of decomposition of urea through the action of the microbial enzyme urease after urine is deposited on barn floors and pastures. Urease is abundant in fecal matter and in soil, and thus urea excreted in urine is rapidly converted to NH₄⁺ when the environmental conditions (temperature, pH) are favorable. Ammonium N can be converted under aerobic conditions to NO₃^– and both forms of N are readily available to plants whereas organic forms of manure N are generally not readily available (Beegle et al., 2008).

Ammonium N is also the carrier of rapidly available N in the soil and a necessary precursor in the process that leads to N₂O emissions from application of manure and fertilizers and urine deposition in pastures (Tisdale et al., 1993; de Klein et al., 2001). Ammonia is a volatile gas that escapes to the atmosphere reducing the amount of N transported to the soil, which may offset the benefits of manure storage and land application mitigation practices. Simultaneously, NH₃ that escapes to the atmosphere can form particulate matter that may return to soil through dry or wet deposition. This deposition can harm sensitive ecosystems and contribute to N runoff and groundwater pollution and be converted into N₂O through denitrification (Galloway et al., 2004; USEPA, 2010; Hristov et al., 2013b).

Ammonia volatilization is generally the largest pathway of loss for manure N (Harper et al., 2004; Lee et al., 2011a), with losses typically accounting for 30 to 70% of the NH₄⁺ content of cattle manure (Thompson and Meisinger, 2002). Nitrogen emissions can also be in the form of N₂ (Harper et al., 2004) but losses as N₂ have not been well quantified. Thompson et al. (1987) estimated that 2 to 12% of applied manure in English grassland was lost to denitrification (N₂ +N₂0) with surface application and 7 to 21% was lost with injection of manure. The portion of manure N lost as N₂O is relatively low, generally below 2 to 3%, and only in a few reports has it reached 10% (de Klein et al., 2001). Based on a N mass balance approach, Hristov et al. (2011b) estimated that more than 25% of the feed N input on a dairy farm was not accounted for in milk and manure after 24 h following excretion and this was mostly attributed to NH₃ volatilization; losses from beef cattle feedlots with long term exposure were even greater, reaching 50% (Cole and Todd, 2009).

The relationship between manure NH₃ volatilization and N₂O emission is also complex because 1) emissions of both may be reduced by diet manipulation or manure management and 2) if a mitigation technology reduces NH₃ losses, the preserved NH₄⁺ may later increase soil N₂O emissions (Petersen and Sommer, 2011). On the other hand, gaseous losses of N will reduce the availability of N for nitrification and denitrification processes and, consequently, N₂O formation (USEPA, 2010). Therefore, NH₃ emission is considered an important component of the analysis of N₂O mitigation practices presented in this review.

ANIMAL MANAGEMENT AND HOUSING

Animal Housing.

Structures used to house livestock do not directly affect the processes resulting in N₂O and CH₄ emissions; however, the type of structure used determines the manure management methods used to handle, store, process, and use the manure. Housing systems with solid floors that use hay or straw for bedding accumulate manure with higher DM, which is commonly stored in piles creating conditions conducive for nitrification and denitrification and thus greater N₂O emission. Külling et al. (2001, 2003) compared liquid manure with stacked manure handling systems, and their results indicate that farm yard manure and deep litter manure handling systems tend to produce greater N₂O emissions than slurry-based systems. In these studies, quantitative differences in N₂O emissions from the manure handling systems were difficult to determine because protein content in the diet and NH₃ emission from manure also varied. Greater CH₄ emissions were reported from farm yard manure followed by liquid slurry and deep litter manure. Amon et al. (2001) compared composted, anaerobically stacked, and slurry-based manure and found higher NH₃ emissions in composted manure with most of the losses occurring after manure was turned during aeration. These authors found much greater N₂O and CH₄ emissions from anaerobically stacked manure with no significant difference between slurry-based and straw-based manure systems.

Housing systems with slatted floors accumulate manure in liquid or slurry form with that manure stored for longer periods of time increasing the production of CH₄ and reducing the production of N₂O. Hassouna et al. (2010) studied gaseous emissions from cattle housing in France and found higher N₂O emissions in buildings with straw-based bedding and solid manure handling systems when compared with liquid manure handling systems. Nitrous oxide emissions were detected in only 2 of the 14 liquid manure systems studied. The same study found smaller differences between CH₄ emissions from buildings using straw-based solid manure or liquid manure systems and attributed this result to the difficulty in discerning enteric CH₄ emissions from manure emissions because the former produced most of the CH₄ emitted from the buildings.

Hristov et al. (2012) investigated the effect of manure management on barn floor NH₃, CH₄, N₂O, and CO₂ emissions from 12 commercial dairy farms in Pennsylvania. Dairies participating in the study had flush (manure was flushed twice daily), 2 types of scrape (manure was scraped daily), and gravity-flow (manure was accumulated under the building and removed several times during the year) manure systems. Barn floor NH₃ emissions were considerably lower for the flush manure systems (average of 167 mg/m² per h) and highest for the gravity-flow system (426 mg/m² per h). Methane emissions were also lowest for the flush (37 mg/m² per h) and much higher for the gravity-flow system (1,216 mg/m² per h). Carbon dioxide emissions were not different among manure systems (ranging from about 2,000 to 7,000 mg/m² per h), and N₂O emissions were negligible in all systems. This study showed that NH₃, and particularly CH₄, emissions from manure are much greater from dairy barns where manure is stored for prolonged periods of time compared with barns where manure is removed daily.

In contrast to ruminants, housing plays a more important role on GHG mitigation in nonruminant livestock production systems because most of the emission in these systems comes from the manure. Philippe et al. (2007) compared GHG emissions from fattening swine raised on a concrete slatted floor or straw-based deep litter. Swine fattened on deep litter released nearly 20% more GHG than those on slatted floors (6.2 and 13.1 g/swine per day for NH₃, 0.54 and 1.11 g/pig per day for N₂O, and 16.3 and 16.0 g/swine per day for CH₄, respectively).

The type of housing system also determines the feasibility of using anaerobic digestion or composting to treat the manure with its associated effects on GHG emissions. As described by the International Atomic Energy Agency (IAEA, 2008), the types of housing used in Asia facilitate different strategies for manure treatment depending on the capital resources of the producer. A popular alternative for smallholders is anaerobic digestion of animal and household waste, usually funded by government programs. Medium to large producers have better access to capital investment, have specialized production determined by surrounding markets, and generally use raised slatted floors that allow collection of manure for further treatment through solids separation and anaerobic digestion.

Mechanically ventilated structures provide opportunity to treat emitted GHG through filtration and scrubbing as gases are exhausted from facilities. An interesting mitigation technology for animal housing uses titanium dioxide (TiO₂) paint on the interior walls. Industrial uses of TiO₂ show that stimulation of its photocatalytic properties by UV light leads to oxidation of NH₃ and NO_x (e.g., Lee et al., 2002; Allen et al., 2005). Studies by Guarino et al. (2008) and Costa et al. (2012) in swine houses showed that GHG mitigation with TiO₂ paint holds promise. Alkali and alkaline earth metal oxides, hydroxides, and carbonates or bicarbonates have been shown to have high CO₂ absorption capacity and are being investigated for CO₂ sorbent applications (Duan et al., 2012).

Dietary Effects on Manure Emissions.

Diet manipulation to reduce nutrient excretion has been studied for many years, especially related to N and P reduction, but its consideration as a mitigation practice to reduce CH₄ and N₂O emissions from manure is relatively new. Diet can have a profound effect on N losses and particularly the route of N excretion (i.e., feces vs. urine) in most farm animals (Hristov et al., 2013b). Studies with ¹⁵N-labeled urine or feces have demonstrated that urinary N is the primary source of NH₄⁺ in cattle manure, contributing from 88 to 97% of the NH₃ emitted within the first 10 d of manure storage (Fig. 3; Lee et al., 2011a).

Figure 3.

Proportion of ammonia N emitted from manure originating from fecal N and urinary N. In this experiment feces or urine were labeled with ¹⁵N through a continuous infusion of ¹⁵NH₄Cl into the rumen of lactating dairy cows (from Lee et al., 2011a).

With urine being the main source of volatile N emissions, manipulating the route of N excretion becomes an important N₂O and NH₃ mitigation tool. Urea is the main nitrogenous constituent of ruminant urine. In the urine of high-producing dairy cows, urea N represents 60 to 80% or more of total urinary N (Reynal and Broderick, 2005; Vander Pol et al., 2008) and proportionally decreases as dietary CP and intake decrease (Colmenero and Broderick, 2006). On low-protein diets, urinary urea N can be as low as 46 to 53% of the total urinary N (Hristov et al., 2011a; Lee et al., 2012a) with urinary N close or even below 19% of total excreta N (Lee et al., 2011b, 2012b). Therefore, reducing dietary CP concentration is perhaps the most effective method for mitigating NH₃ emissions from stored manure (Hristov et al., 2011b). Emissions from land-applied manure are further reduced because low-CP diets produce manure with a slower mineralization rate of N (Powell and Broderick, 2011). Sauvant et al. (2011) showed, however, that CH₄ production per kilogram digested OM decreased in a linear fashion with increasing dietary CP, that is, decreasing dietary protein concentration likely results in increased concentration of fermentable carbohydrates in the diet, which in turn likely increases CH₄ production. In support of this, Dijkstra et al. (2011) concluded, from a simulation study, that dietary-N mitigation options at the animal level aimed at reducing urinary N excretion may result in elevated enteric CH₄ emission (per kg of fat- and protein-corrected milk). These potential relationships must be considered when manipulating dietary N to reduce manure NH₃ and N₂O emissions.

Several studies have investigated the effect of dietary protein on N₂O (and CH₄) emissions from manure and manure-amended soil. Külling et al. (2001) reported decreased N₂O emissions during simulated storage of manure from dairy cows fed low-protein diets, but the total GHG emissions were not affected by the dietary protein content (due to increased CH₄ emissions from the low-protein manure). Velthof et al. (2005) concluded that decreasing the protein content of swine diets had the largest potential to simultaneously decrease NH₃ and CH₄ emissions during manure storage and N₂O emission from soil. Data on the effect of dietary protein on manure N₂O emissions, however, are not consistent and often no effect or even increased N₂O emissions (from housing) have been reported when lowering dietary protein for swine (Clark et al., 2005; Philippe et al., 2006) and dairy cattle (Arriaga et al., 2010). Manure CH₄ and CO₂ emissions per unit of land may increase immediately following soil application (or during storage) due to the use of a greater application rate for low- vs. high-protein manure to meet the crops N requirements (Lee et al., 2012b).

Feeding and management can significantly affect N excretions and volatilization losses from beef feedlots. Phase feeding is one example of an effective mitigation practice for these types of production systems. Reducing dietary protein concentration during the production cycle to better meet the requirements of the animal can significantly lower N excretions (Cole et al., 2005, 2006; Vasconcelos et al., 2007) and consequently losses from the pen surface. Erickson and Klopfenstein (2010) reported 12 to 21% lower N excretion and 15 to 33% lower N volatilization losses of phase-fed cattle. Manure management can also have a significant impact on N losses. Pen cleaning frequency, for example, decreased N volatilization losses by 19 to 44% and increased manure N by 26 to 41% (Erickson and Klopfenstein, 2010).

Reduced protein N in the animal diet produces manure with a slower N mineralization rate that releases less plant-available N (Powell and Broderick, 2011). Therefore, changes in manure application rate recommendations are needed to reflect N cycling from modified diets. At equal N application rates, whole-crop barley yield was not different between manures from dairy cows fed high- (16.8) or low-crude protein (14.8%) diets (Lee et al., 2013). To minimize N₂O production in all cases, manure application rates should be coordinated with the amount of mineral fertilizer applied, and consideration should be given to application timing and method to prevent N application in excess of plant requirements.

Low-protein diets must be formulated to meet or exceed the animal's energy, metabolizable protein, and AA requirements if feed intake and animal performance are to be maintained (Lee et al., 2011b). Diets severely deficient in RDP will reduce total tract fiber digestibility in ruminants, which may negatively affect DMI and animal performance (Mertens, 1994; Lee et al., 2011b; Aschemann et al., 2012). A meta-analysis by Nousiainen et al. (2009) and Huhtanen et al. (2009) showed that diet CP was the only dietary factor (of the factors studied in that analysis) that was positively related to NDF digestibility in dairy cows. Decreased ruminal degradability of fiber will increase excretion of fermentable OM in manure, which might increase manure CH₄ emissions although the latter effect has not been consistently reported (Hindrichsen et al., 2005). On the other hand, these effects may be counteracted by reduced enteric CH₄ production because fiber degradability in the rumen will decrease. Diets severely deficient in RDP will have a negative impact on microbial protein synthesis and animal productivity and therefore must not be recommended as a mitigation practice.

Feed intake depression with protein- and AA-deficient diets has also been demonstrated with pigs and poultry (Henry, 1985; Picard et al., 1993) and must be avoided to maintain efficient animal production. Supplementation of low-protein diets with synthetic AA may alleviate undesirable effects on feed intake. Growing swine fed a 14% CP diet containing supplemental lysine (0.73% Lys) had intake and growth performance similar to swine fed a 16% protein diet (0.77% Lys; Baker et al., 1975). Analogous results were reported by Yen and Veum (1982) who observed feed intake and ADG for growing swine fed a protein deficient (13% CP) diet supplemented with Lys and Trp to be similar to swine fed a 16% CP diet. Similar to monogastric animals, supplementation of the diet with rumen-protected AA (Lys, Met and His) increased DMI and milk production in high-producing dairy cows (Lee et al., 2012a). Overall, feeding protein close to animal requirements, including varying protein concentration with the productive stage of the animals (phase feeding), is recommended as an effective manure NH₃ and N₂O emission mitigation practice.

De Klein and Eckard (2008) concluded that abatement of N₂O should be considered as part of an integrated approach to improve the efficiency of N cycling in animal production systems. Particular attention should be given to improving animal N utilization, thus reducing urinary N output to the soil–plant system. According to these authors, current technologies could deliver up to 50% reduction in N₂O emissions from an animal housing system but only up to 15% from a grazing system. Dietary CP reduction can reduce both CH₄ and N₂O emissions from stored manure (Atakora et al., 2011a,b; Osada et al., 2011) as well as following land application (Velthof et al., 2005). Land application reductions are not supported by all data because of the large variation in soil conditions. In intensively managed pastoral systems, supplementation of the pasture with low-N feeds such as corn or small grain silage, which generally reduce dietary N concentration, can reduce urinary N losses and, consequently, manure and soil NH₃ and N₂O emissions (by 8 to 36%; de Klein and Monaghan, 2011). In some systems, however, this reduction may be of a smaller magnitude (Velthof et al., 2009), or total GHG emissions may be even increased (Beukes et al., 2010), perhaps due to increased synthetic fertilizer use to grow the cereal silage.

Shifting N losses from urine to feces is expected to reduce N₂O emissions from manure-amended soil due to the lower concentration of NH₄⁺ in manure (this will depend on storage conditions, if manure is stored before application). Feed additives, such as tannins, have redirected excreted N from urine to feces. Carulla et al. (2005), for example, reported a 9.3% reduction in urinary N as proportion of total N losses, and Misselbrook et al. (2005) reported a reduction in manure NH₃ emissions from cows fed tanniferous forage (although the effect in the latter study was confounded by the CP content of the diets). Grainger et al. (2009) observed a 45 to 59% reduction in urinary N excretion (as percent of N intake) with condensed tannins but also a 22 to 30% drop in milk N secretion. Similarly, Aguerre et al. (2010) observed a linear decrease in urinary N excretion (vs. a linear increase in fecal N excretion) in high-producing dairy cows fed diets supplemented with 0 to 1.8% (DM basis) of a quebracho tannin extract. Ammonia emission from slurry from cows receiving the tannin-supplemented diets was 8 to 49% lower than emissions from the control slurry. Tannins also reduced NH₃ emission by 20% when directly applied to the barn floor and 27% after a tannin extract was applied to soil (Powell et al., 2011a,b).

Studies directly investigating the effect of tannins on manure or soil N₂O emissions are scarce. Hao et al. (2011) supplemented cattle diets with condensed tannins from Acacia mearnsii at 25 g/kg DM and followed GHG emissions from composted manure for up to 217 d. Nitrous oxide emissions that occurred during the first 56 d of composting were generally low (up to 0.1 kg N/t compost DM) and not affected by tannin supplementation. Methane (and CO₂) emissions were also not affected by tannin supplementation. The authors speculated that tannin application level was too low, tannins were complexed with protein on excretion in the feces, or that microbes in compost were capable of altering the biological activity of tannins.

Decreased N release rate from manure from animals fed tanniferous forages has been reported (Powell et al., 1994; Cadisch and Giller, 2001) although other reports indicated no effect of condensed tannins on the agronomic value of cattle manure (Hao et al., 2011) and decreased manure N availability may be a concern in agricultural systems relying exclusively on manure as a source of N for crop growth. Indeed, some reports have indicated a significant drop in yield when high-tannin manure (equivalent to 2.2 t tannins/ha per yr) was added to sweet corn and radish plots (27 to 32% reduction for sweet corn and 42 to 46% for radish; Ingold et al., 2012). However, others reported no effect of condensed tannins on the N fertilizer value of feces from sheep receiving a diet supplemented with tanniferous legumes (Tiemann et al., 2009). Clearly, more studies are needed to relate tannin application, through the diet or directly to manure, to GHG emissions from manure during storage or after land application.

Grazing Practices.

Improving pasture quality in terms of forage digestibility is an efficient way of decreasing GHG emissions from the animal and the amount of manure produced. However, in pasture-based production systems, improving forage quality often means increasing N fertilizer application rates, which can have a negative impact on urinary N excretion and thus NH₃ and N₂O emissions.

Nitrous oxide emissions can be particularly high in intensive pasture systems due to high N concentration in urine as a result of the high CP content of pasture (22 to 28% CP, DM basis, in New Zealand, for example), and there are many reports on the relationship between the placement and chemical composition of urine and soil nitrification and denitrification processes. Eckard et al. (2010) pointed out that the effective N application rate within a urine patch from a dairy cow on pasture is between 800 and 1,300 kg N/ha, and N is deposited at concentrations that are orders of magnitude greater than the utilization capacity of the soil–plant system. These authors suggested that a more uniform distribution of urine throughout the paddock would reduce the effective N application rate, which should translate into a reduction in N₂O emissions.

These effects are often compounded by high fertilizer N application rates to stimulate grass growth, which further increases urinary N concentration. De Klein et al. (2001) showed a 40 to 57% reduction in N₂O emissions when grazing was restricted to 3 h/d in the humid late autumn in New Zealand. This reduction was attributed to diminished N input during conditions most conducive to N₂O emissions in New Zealand. When de Klein et al. (2001) included N₂O emissions resulting from application of the effluent collected during restricted grazing periods, N₂O emissions were reduced by only 7 to 11%.

Nevertheless, keeping the animals off the paddocks, in “stand-off” or “feed pads” for most of the day during the wet months of the year (autumn–winter), has been shown to be an effective N₂O mitigation practice in intensive grazing systems (de Klein, 2001; de Klein et al., 2002; Luo et al., 2008a). Not allowing grazing during wet weather also increases pasture productivity due to reduced sward damage and soil compaction (de Klein, 2001; de Klein et al., 2006). One must keep in mind, however, that this practice results in much greater NH₃ emissions (Luo et al., 2010) due to urine and feces being excreted and allowed to mix in the stand-off or feed pad area. According to Luo et al. (2010), reduction of N₂O emissions from intensive grazing systems can be achieved by several strategies: 1) improving N use efficiency through reducing the amount of N excreted by grazing animals, 2) optimizing soil management and N inputs, 3) optimizing pasture renovation, 4) manipulating soil N cycling processes through soil additives, 5) selecting for plants and animals that maximize N utilization, and 6) altering grazing and feeding management.

Biofiltration.

Biofiltration can treat ventilated air from animal buildings using biological scrubbers to control odor, to absorb NH₃, and to convert NH₃ into NO₃. Preventing NH₃ losses may indirectly reduce N₂O emissions by reducing NH₄⁺ deposition and consequent conversion to N₂O (see earlier discussion).

Ammonia removal efficiency in swine and poultry houses from acid scrubbers and biotrickling filters (based on biofilms that degrade the odorous compounds) averaged 96 and 70%, respectively (Melse and Ogink, 2005). Shah et al. (2011) investigated the effectiveness of a coupled biofilter–heat exchanger in reducing NH₃ emissions (and recover heat) in a broiler house. The biofilter was effective in treating very high inlet NH₃ concentrations (>96 mg/kg) with removal efficiencies greater than 79% for empty bed residence times ranging from 4.3 to 29.1 s. The biofilter was apparently also able to trap some sulfurous gases emitted from the broiler house.

Recent reports (Maia et al., 2012a,b) have shown that biofilters used to scrub NH₃ from exhaust streams in animal houses generate N₂O as a result of nitrification and denitrification processes in the biofiltration media. In their first study, Maia et al. (2012a) showed a high correlation between biofilter NH₃ removal and N₂O generation associated with conversion of NH₃ to NO₂^– and NO₃^– in the biofilter. In their second study, Maia et al. (2012b) reported that moisture content between 48 and 52% in the biofilter media was an important factor in obtaining significant NH₃ reduction with reduced N₂O production. Therefore, N₂O production in biofilter scrubbers should be taken into account when implementing biofiltration systems for GHG and NH₃ mitigation.

A few studies have investigated CH₄ mitigation by passing contaminated air from a swine manure storage or from swine housing through a biofiltration system. A Canadian Pork Council (2006) study reported reductions of 50 to 60%, and Girard et al. (2011) reported a maximum reduction of up to 40%. They described their biofilter as “packed with inorganic material,” but the packing material was not disclosed.

Melse and van der Werf (2005) reported up to 85% CH₄ removal from the exhaust stream of a covered swine liquid manure storage using a biofiltration system composed of a mixture of compost and perlite inoculated with CH₄ oxidizing bacteria collected from activated sludge. The CH₄ removal capacity of the biofilter system depended on the concentration of CH₄ in the filtered stream. Therefore, the authors extrapolated that an equivalent system for animal housing with low CH₄ concentration in the filtered stream would require very large biofilter systems to achieve 50% reduction or more, pointing to this as a limitation in the applicability of this technology. In addition, they reported N₂O production in the biofiltration system contributing 4 to 64% of the outlet stream GHG CO₂e, which, as mentioned above, needs to be included in the design and promotion of biofiltration strategies for GHG mitigation. The high residence time necessary in these systems due the low solubility and biodegradability of CH₄ hinders effectiveness (Melse and Verdoes, 2005). Melse and Timmerman (2009) reported on the potential use of multipollutant scrubbers, combining acid scrubbers, biofilters, and water curtains to reduce not only NH₃, odors, and GHG but also particulate matter from animal housing exhaust.

MANURE STORAGE AND TREATMENT

Greenhouse gas emissions from stored manure are primarily in the form of CH₄ (due to anaerobic conditions) although N₂O emissions can occur and NH₃ volatilization losses are often large. A direct way to avoid cumulative GHG emissions is to reduce the time manure is stored (Philippe et al., 2007; Costa et al., 2012). Increasing the time of manure storage increases the period during which CH₄ (and potentially N₂O) is emitted as well as the emission rate, creating a compound increase (Philippe et al., 2007).

Temperature is a critical factor regulating processes leading to NH₃ (Sommer et al., 2006) and CH₄ (Steed and Hashimoto, 1994) emissions from stored manure. Decreasing manure temperature to <10°C by removing the manure from the building and storing it outside in cold climates can reduce CH₄ emissions (Monteny et al., 2006).

Storage treatments with proper aeration and moisture management have reduced CH₄ generation from poultry manure (Li and Xin, 2010). Ventilated belt removal of laying hen manure can reduce CH₄ emissions compared to deep-pit storage (Fabbri et al., 2007). Separation of swine slurry into solid and liquid portions and then treating the solids through aerated composting reduced CH₄ emissions by 99% and N₂O emissions by 75% compared with untreated manure (Vanotti et al., 2008). However, due to the often negative relationship between NH₃ and N₂O emissions (Petersen and Sommer, 2011), this process is likely to increase NH₃ emissions substantially and perhaps total N losses from manure. Amon et al. (2001) reported greater NH₃ losses from an actively turned composting pile of solid cattle manure than from an undisturbed anaerobically stored pile, with the opposite effect observed for N₂O emissions.

Prapaspongsa et al. (2010a,b) compared 14 swine manure management practices based on combinations of thermal pretreatment, anaerobic digestion, anaerobic co-digestion, liquid and solid separation, drying, incineration, and thermal gasification with respect to their energy, nutrient, and GHG balances. The anaerobic digestion-based scenario with a natural crust during storage had the highest GWP reduction through high efficiencies in energy and nutrient recovery with restricted emissions of GHG and NO₃^–. The incineration and thermal gasification-based scenarios and a scenario using only deep injection yielded the greatest reduction in respiratory inorganics and terrestrial eutrophication categories because they had the lowest NH₃ emissions. Manure incineration combined with liquid and solid separation and drying of the solids was a promising management option yielding a high potential energy utilization rate and GHG reduction.

Storage Covers.

Several types of manure storage covers have been reported in the literature including natural crusts on slurry manure stored with a high solids content, straw, wood chips, oil layers, expanded clay pellets, wood, and semipermeable and sealed plastic covers. The effectiveness of the cover depends on many factors, including permeability, cover thickness, degradability, porosity, and management. Semipermeable covers tend to increase N₂O emission because they provide optimal aerobic conditions for nitrification at the cover surface and at the same time create a low oxygen environment just below the cover favorable for denitrification and the production of N₂O (Hansen et al., 2009; Nielsen et al., 2010). Semipermeable covers are valuable for reducing NH₃, CH₄, and odor emissions, but they often increase N₂O emissions (Sommer et al., 2000; Guarino et al., 2006; VanderZaag et al., 2008).

Capturing the gases produced using impermeable membranes, such as oil layers and sealed plastic covers, can reduce NH₃, N₂O, and CH₄ emissions. The results from Guarino et al. (2006) and VanderZaag et al. (2008) suggest that using a vegetable oil layer as a manure storage cover, although very effective, is not practical because of degradability, generation of foul odors, and difficulty in preventing the oil film from becoming mixed or “broken” over the manure surface.

Covering manure storages with impermeable covers is an effective mitigation practice if the CH₄ captured under the cover is burned using a flare or an engine generator set to produce electricity; otherwise, the captured CH₄ builds pressure inside the storage creating an explosion hazard and/or rupture of the cover. Increased air pressure inside the storage structure reduces the fraction of compounds in the gas phase and increases that trapped in liquid manure. The increased gas trapped in the manure liquid is then released when the pressure in the manure storage is reduced to remove the manure. Retaining the CH₄ produced is not beneficial if it escapes at a later stage; therefore, burning or combusting the collected CH₄ to produce electricity or heat is the most desirable option. The effectiveness of impermeable covers depends on transforming the collected gases to less potent GHG such as NO_x and CO₂ (Nicolai and Pohl, 2004; Rotz and Hafner, 2011).

Anaerobic Digestion.

Anaerobic digestion is the process of degradation of organic materials by archaea in the absence of oxygen, producing CH₄, CO₂, and other gases as by-products. This provides a promising practice for mitigating GHG emissions from collected manure. In addition, when correctly operated, anaerobic digesters are a source of renewable energy in the form of biogas, which is 60 to 80% CH₄, depending on the substrate and operation conditions (Roos et al., 2004). Anaerobic digesters also provide opportunities to reduce pathogens and manure odor.

Three practical temperature ranges are generally considered for anaerobic biogas systems: psychrophilic (15 to 25°C), mesophilic (30 to 38°C), and thermophilic (50 to 60°C). These temperature ranges facilitate the growth of specific microbes. Thermophilic systems are more sensitive to environmental changes, such as temperature fluctuations and chemical concentrations produced during the digestion process (Kim et al., 2002; Ahn and Forster, 2002; El-Mashad et al., 2003), because the number of functional microorganism species that thrive at this temperature is considerably less than those that survive at lower temperatures (Ziekus, 1977; Wolfe, 1979; Smith, 1980). Below 15°C, the production of biogas is greatly reduced and CO₂ becomes the dominant product of anaerobic digestion; therefore, anaerobic digestion systems are not recommended for geographic locations with average temperatures below this threshold without supplemental heat and temperature control (Sommer et al., 2007).

The effluent of the digester, commonly called digestate, contains most of the soluble plant nutrients found in the feedstock and the more resilient, difficult to degrade organic material. Digestate is commonly applied directly to crops whereas the sludge, formed by precipitated minerals and undigested OM, may be composted before field application.

Digester designs vary widely in size, function, and operational parameters. Smaller digesters (6 to 10 m³), designed to improve sanitary conditions in developing countries and to provide energy for single family dwellings, were promoted in the 1970s and 1980s throughout Asia and Latin America (Bond and Templeton, 2011; Jiang et al., 2011b). These were designed to function with the waste originating from a few animals (2 to 5 swine, 5 to 10 cows, 100 chickens, or a combination of these) together with the family dwelling waste. According to Dhingra et al. (2011), these types of digesters reduced GHG emissions from 23 to 53% when compared with households without biogas, depending on the condition of the digester, technical assistance, and operator ability. The effectiveness of these types of digesters for mitigating GHG depends mostly on the amount of CH₄ leakage that occurs through digester walls and piping delivering the biogas to the family dwelling (Dhingra et al., 2011). These small digesters have been used by farmers in developing countries only when government subsidies and economic incentives have been available (Bond and Templeton, 2011).

Commercial farm digesters are typically designed to treat liquid manures. There are 4 basic commercial farm-level anaerobic digestion vessel designs (Roos et al., 2004). The most common and simple covered lagoon digestion systems are ambient (psychrophilic) temperature systems that require manure with a solid content of 3% or less and a storage cover to maintain anaerobic conditions. These systems typically create the largest type of digester with the longest hydraulic retention time. Plug-flow digesters and fixed domes use a vessel that receives manure at one end and discharges from the opposite end with no mixing or agitation. These systems may be heated to a mesophilic temperature and require slurry with a solids content of 11 to 13%. Small-scale digesters are often of this design. The more sophisticated complete mix digesters consist of an engineered digestion vessel designed to handle manure slurries with a solid content from 3 to 10%. A mixing system enhances bacterial contact with OM. Supplemental heat is often added to these systems to operate at mesophilic temperatures, which promotes bacterial growth and a shorter hydraulic retention time. Fixed-film digesters use a medium, such as rope, plastic mesh, or beads, placed in the vessel on which bacteria can grow. Dilute manures with a solids content of 3% or less are passed across (or through) the medium in these systems. Whereas other systems rely solely on suspended microbial growth, these also feature attached microbial growth. Widespread commercial farm digester adoption has not occurred because of variable economic return (Hill et al., 1985; Safley and Westerman, 1994; Braber, 1995) and the limited competitiveness of biogas with other fuels used for heat and power (Lantz et al., 2007).

Industrial biogas digesters are used to produce renewable energy for towns and municipalities. These digesters, prevalent mostly in Europe, use biomass collected from several farms to feed the anaerobic digesters. Co-digestion of agricultural biomass, industrial organic waste, and animal manures is common in industrial biogas plants because it allows better optimization of C to N ratio and CH₄ production while reducing the impact of NH₄⁺ on gas production (Ward et al., 2008).

When CH₄ is collected and used as an energy source, it can substitute for combusted fossil fuels reducing the emissions of GHG, NO_x, hydrocarbons, and particulate matter (Börjesson and Berglund, 2006). These authors compared the emissions from the life cycle of raw materials used for anaerobic digestion (6 different feedstocks, including swine manure) and the emission from systems that the anaerobic digestion process replaced. One of the serious concerns identified was uncontrolled losses of CH₄ from biogas plants, including losses from stored digestate. Typical losses from systems storing digested manure were reported to range from 5 to 20% of the total biogas produced (Bjurling and Svärd, 1998; Sommer et al., 2001).

In a follow-up report, Börjesson and Berglund (2007) further explored overall environmental impact when biogas systems replaced various energy producing reference systems. The investigation was based on Swedish conditions using an LCA approach that considered both direct and indirect emissions. Greenhouse gas emissions per unit of heat were reduced 10 to 25% when biogas-based heat replaced fossil fuel-based heat. Emissions from biogas systems contributed 60 to 75% and 25 to 40% of the life cycle emissions of CO₂ and CH₄ in the reference and fuel-based systems, respectively.

During the anaerobic digestion process, N-containing compounds found in substrates, such as proteins, AA, and urea, are reduced to NH₃ (Bernet et al., 2000). Ammonia remaining in the aqueous solution is then transferred to the soil when the digestate is land applied (Bernet et al., 2000; Hafner et al., 2006). Anaerobic digestion stabilizes the organic C in the feedstock (reducing the fraction of easily degradable C in manures) increases plant availability of N and provides less energy to support the growth of N₂O-forming microorganisms, reducing the potential of N₂O emissions when applied to soil. Mineralization of organic N and VFA during anaerobic digestion increases manure pH and available N, which may result in increased NH₃ volatilization (Petersen and Sommer, 2011).

In general, reduction of manure OM content is expected to reduce N₂O emissions from manure-amended soils (Petersen, 1999; Bertora et al., 2008) although Thomsen et al. (2010) reported higher N₂O emissions when treated manure was applied in a wet spring season. These contradictory results led Petersen and Sommer (2011) to conclude that there is not a simple relationship between removal of manure OM and the risk of N₂O emission. To address this controversy, Thomsen et al. (2010) proposed linking the balance between N₂O and N₂ to soil water-filled pore space and oxygen supply. This relationship has been discussed in detail by Petersen and Sommer (2011); the authors concluded that prediction of N₂O emissions from manure-amended soil depends on manure composition and soil conditions. Masse et al. (2011) noted high variability between N₂O emissions, referring to 6 studies that found similar differences in emissions of the gas when comparing digested and nondigested manures.

Data on the anaerobic digestion of poultry waste as a GHG mitigation practice are limited. Several studies show successful biogas production using poultry waste as a component of co-digestion (digesting poultry waste with other manures—beneficial due to the complementary composition of the different manures); however, the impact on GHG mitigation was not reported. The ability to use anaerobic digestion to create, capture, and destroy CH₄ derived from swine manure is well documented (Safley and Westerman, 1994; Masse et al., 2003a,b). Although it is possible to reduce CH₄ emissions by over 60% from swine manure using anaerobic digestion, the amount of CH₄ produced and collected does not directly translate into an equal amount of reduced CH₄ emissions because the untreated manure would not yield the same amount of CH₄ gas. Most literature reviewed focused on research that compared digested manure with manures that received no treatment or a different treatment. In this manner, the biogas removed was not considered in the emission comparisons of nondigested versus digested manure by many authors who used the assumption that biogas produced during digestion is destroyed through controlled combustion. A number of the studies referenced here considered emissions from digested manure after it was land applied. Reductions of N₂O emissions reported in these papers were as high as 70% compared with untreated manure applications. A commonly stated reason for this decrease was that digested manure contains less OM (degradable C) providing less energy for nitrite-forming microorganisms, which subsequently limits N₂O production.

Although most anaerobic digestion systems significantly reduce GHG emissions when compared with traditional manure handling systems, incorrect operation, lack of maintenance, and CH₄ leaks can make them a net contributor to GHG. For this reason, it is imperative that system designs and components ensure containment of nearly all biogas. The potential for anaerobic digestion to mitigate N₂O emissions after the digested manure is land applied is promising, but many parameters involved with field application contribute to conflicting reports.

Anaerobic digestion systems require large initial capital investments during construction along with ongoing maintenance and supervision costs. Historically, the adoption of this technology occurs only when economic incentives are offered as price advantages for biogas (biofuels and renewable policy incentives), when the costs of construction and maintenance are subsidized, or when no competitive alternative energy source is available. Furthermore, instruction and technical assistance to users are necessary in implementing successful anaerobic digestion systems because the correct operation of anaerobic digesters is not trivial and 50% failure rates are common (Bond and Templeton, 2011; Jiang et al., 2011a).

Acidification.

An important factor affecting GHG emissions, in particular NH₃, from stored manure is pH. According to Petersen and Sommer (2011), manure acidification is an effective mitigation option for NH₃ emissions, but the effect on N₂O is not well studied. The relationships between NH₃ volatilization and factors such as air velocity and turbulence, manure temperature, and manure pH have been well documented (Ndegwa et al., 2011).

Ndegwa et al. (2011) listed 15 studies in which cattle, swine, or poultry manure NH₃ emissions were successfully mitigated (from 14 to 100% reduction in emissions) by lowering manure pH with sulfuric, hydrochloric, or phosphoric acids, calcium chloride, alum, or monocalcium phosphate monohydrate. These authors concluded that strong acids are more cost effective at reducing manure pH than weaker acids or acidifying salts. However, strong acids are more hazardous and, therefore, acidifying salts and weaker acids may be more suitable for on-farm use. Acidification of urine and, consequently, manure from cattle or monogastric farm animals has also been attempted using anionic salts, high dietary levels of fermentable carbohydrates, organic (benzoic) acids, or Ca and P salts (Ndegwa et al., 2008).

A commercial system used on several farms in Denmark acidifies a portion of the manure with concentrated sulfuric acid to a pH of 5.5, removes a portion of the acidified manure equivalent to daily manure production, and returns the remaining manure to the storage facility (Sørensen and Eriksen, 2009). These authors concluded that NH₃ volatilization from acidified cattle and swine manure was low after both soil incorporation and surface application. Petersen et al. (2012) studied the effect of acidification on CH₄ (and NH₃) emission from fresh and aged cattle manure during 3 mo of storage using the equipment described above. Manure pH was adjusted to 5.5 with sulfuric acid, samples of manure were stored for 95 d, and NH₃ and CH₄ emissions were monitored. Manure pH increased gradually to 6.5 to 7 during storage. Acidification had a dramatic effect on emissions, reducing CH₄ by 67 to 87% (more pronounced with aged manure) and almost completely eliminating NH₃ emissions. The authors concluded that manure acidification may be a cost-effective GHG mitigation practice.

Application of acidified manure is not expected to greatly impact crop production; the pH range of acidified manure is within the optimal range for corn and many cereal crops (5.5 to 6.5; Tisdale et al., 1993). Approximately 30% of soils worldwide and about 60% in Asia are acidic (<pH 5.5) and already require periodic lime applications to maintain optimal pH (von Uexküll and Mutert, 1995). Smaller quantities of acidified manure would be needed to provide crop N requirements because the reduction in NH₃ emissions provides manure with a greater plant-available N content. However, long-term impacts of land application of acidified manures on soil pH have not been reported, and more frequent application of lime to maintain optimal pH in some soils may be required.

Composting.

Composting is an exothermic, aerobic process of microbial decomposition of OM that has several benefits related to manure handling, odor control, pathogen control, OM stabilization, additional farm income, etc. Composted manure solids (following manure separation into solids and liquid) are also used as bedding in some dairy production systems to reduce the cost of production and provide cow comfort, assuming udder health is not compromised (Husfeldt et al., 2012).

Due to the nature of the composting process, N losses can be high and are influenced by a number of factors including temperature, C to N ratio, pH, moisture, and material consistency (Zeman et al., 2002). Compost can be a source of N₂O emissions with both nitrification and denitrification processes occurring during composting. Bacillus species are the main players in the degradation of OM and betaproteobacterial NH₃–oxidizing bacteria involved in the nitrification process (Maeda et al., 2011). Depending on composting intensity, NH₃ losses can be particularly high, reaching 50% of the total manure N (Peigné and Girardin, 2004). Aeration of the composting heap reduces CH₄ emissions (Thompson et al., 2004; Jiang et al., 2011a; Park et al., 2011) but can increase NH₃ and N₂O losses (Jiang et al., 2011a). Hao et al. (2004) reported up to 30% DM, 53% C, and 42% of initial N being lost during composting of straw-bedded manure. Methane losses accounted for 6% of the C losses. Nitrous oxide losses represented 1 to 6% of the total N losses.

Addition of mature compost with nitrite-oxidizing bacteria to actively composting swine manure was shown to reduce N₂O emission by 70% (Fukumoto and Inubushi, 2009). These authors reported that up to 19% of the total manure N was lost as NH₃ and N₂O. Brown et al. (2008) reviewed the impact of composting of a range of feedstocks (including animal manure) on GHG emissions and pointed out that the primary benefit of composting is the reduction of CH₄ emissions compared with manure stored under anaerobic conditions. These authors estimated, for example, that a facility that composts an equal mixture of manure, newsprint, and food waste could conserve the equivalent of 3.1 Mg CO₂ per Mg of dry feedstock composted if feedstocks were diverted from anaerobic storage lagoons and landfills without gas collection mechanisms.

According to Clemens et al. (2006), raw cattle manure can release from 160 (winter) to 3,600 (summer) g/m³ of CH₄ and 38 to 57 g/m³ of N₂O. For digested manure, the release rates are from 80 (winter) to 1,200 g/m³ (summer) CH₄ and 40 to 76 g/m³ N₂O, respectively. A recent study by Kariyapperuma et al. (2012) reported a 57% decrease in soil N₂O emissions with composted vs. liquid swine manure. Remarkably, emissions during the same period of the following year were not different between composted and noncomposted manure; the authors attributed the lack of difference to a significant reduction in emissions in the second year due to frozen soil. In spite of significant GHG emissions from composting, the review by Brown et al. (2008) concluded that even in a worst-case scenario, these emissions are minimal in comparison to the benefits associated with the CH₄ reduction credits from composting. The authors also stated that it is possible to significantly reduce emissions from compost piles by increasing the solids content of the feedstocks and the C to N ratio. Overall, Brown et al. (2008) concluded that composting can be an effective method for reducing GHG emissions from a range of waste materials, including animal manure. It must be noted, however, that NH₃ losses during manure composting are significant.

LAND APPLICATION

Manure is a valuable resource that is best used as fertilizer. However, increased animal density accompanied by continuous inflow of nutrients from imported feeds can lead to nutrient imbalance at the farm and watershed scales, which can lead to greater water and air pollution. This nutrient imbalance is more likely to occur in intensive animal production systems. When input of recoverable manure nutrients (the quantity of manure nutrients available for land application) grossly exceeds the assimilative capacity of soil and crop utilization, nutrient buildups occur (Saam et al., 2005). Lander et al. (1998) categorized the ratio of recoverable manure nutrients to the assimilative capacity of crop and pasture land at a county level from 1.0 (indicating that the county had county-level excess nutrients) to <0.25 (less than 25% of the nutrients taken up and removed by crops or applied to pasture can be supplied by manure generated within the county). According to the USDA (Kellogg et al., 2000), the number of counties in the United States in 1997 with ratios of 0.5 or greater (i.e., surplus of nutrients) totaled 165 for N and 374 for P. Maguire et al. (2007) estimated that in 89% of counties in the United States, there was a deficit of manure P relative to crop P removal and there was a manure P surplus in the remaining 11%. Therefore, in some regions, manure application can be limited by soil accumulation of nutrients.

Surplus nutrients can be a significant environmental problem for large livestock operations. For example, Hristov et al. (2006) reported average efficiency of use of imported N and P (total exports/total imports × 100) on commercial dairy farms in Idaho (average size of 2,100 cows and 186 ha arable land) of 41 and 66%, respectively. Accumulation of nutrients was occurring on these dairies and as a result, soil P levels in the top 30-cm layer were well above state threshold standards. Soil NO₃^– N was >40 mg/kg for 5 of the 8 dairies, and 2 were over 80 mg/kg. Such high N concentrations exceed the crop needs for optimal growth and thus represent a high potential for N loss to the environment.

When nutrient surplus is not an issue, manure is a valuable source of available nutrients for crops, successfully replacing inorganic fertilizer. In an LCA analysis, Adom et al. (2012) found that N fertilizer input was the largest contributor to GHG emissions for feeds used by the dairy industry in the United States: about 65% due to N₂O release on application and 35% from fertilizer manufacture. These authors recommended farmer education in fertilizer best management practices to effectively reduce GHG emissions on farms.

Application Method and Emissions.

Components of the microbial biomass in the soil use CH₄ as a C source and, with the exception of rice paddies, the soil is often a CH₄ sink. Only when CH₄ concentrations exceed the metabolic capacity of the soil or when the aerobic metabolism of the soil biota is inhibited throughout the soil column are CH₄ emissions significant after land application of manure. Therefore, promoting the aerobic metabolic path and reducing CH₄ load are other approaches used to reduce CH₄ emissions after manure injection (Rodhe et al., 2006). Agricultural soils absorb on average of 1.5 kg CH₄/ha per yr (Chianese et al., 2009), but CH₄ from land-applied manures can be a source of CH₄ emission, diminishing within a few days following application. Sherlock et al. (2002) measured CH₄, NH₃, and N₂O immediately after land application of swine slurry to pasture and up to 90 d afterward. They reported high NH₃ emission rates immediately after application, decaying quickly and totaling 57 kg N/ha or 22.5% of the applied N. Methane emissions were highest immediately after manure application, coming from CH₄ dissolved in the manure. Very low emissions continued for the following week, attributed to anaerobic degradation of fatty acids in the manure. Total CH₄ emissions were slightly higher than 1 kg C/ha, which accounted for 0.08% of the C applied. In contrast, N₂O emission was initially low and dropped to background concentration levels after 90 d, but high emission peaks were observed following rainfall events. Although total N₂O emission accounted for only 2.1% of the N applied, or 7.6 kg N/ha, the authors considered N₂O to be the most important pollutant due to its greater GWP (Sherlock et al., 2002).

An important difference between mineral fertilizer and manure is that manure contains organic C, which, depending on soil conditions, may affect N₂O emissions. Manure C may increase microbial respiration rates in soil, thus depleting oxygen and providing the anaerobic conditions required for denitrification (Pelster et al., 2012). As a result, organic amendments containing large amounts of labile C and available N (cattle, swine, or poultry manure) have lead to increased soil N₂O emission compared with mineral fertilizers. An Intergovernmental Panel on Climate Change (IPCC, 2006b) report assumed N₂O emission factors (Tier 1) for mineral fertilizers and cattle, poultry, and swine manure at 1 and 2% of N input, respectively. Pelster et al. (2012) reported an N₂O emission factor for plots amended with poultry manure as 1.8% of applied N, more than double that of the other treatments, including mineral fertilizer (0.3 to 0.9%), a result attributed to the high C content of poultry manure. These authors concluded that, compared with mineral N sources, manure application increases soil N₂O flux in soils with low C content. Soil N₂O emissions can vary greatly, and emission factors of up to 12% of N input (for nitrate-based fertilizer) and 5% for manure have been reported (de Klein et al., 2001). Nyakatawa et al. (2011) investigated CH₄ and N₂O emissions from soil receiving poultry litter or ammonium nitrate using surface soil incorporation and subsurface band application methods in conventional and no-tillage systems on a Decatur silt loam soil in north Alabama. Plots receiving ammonium nitrate were net emitters of CH₄ and N₂O whereas plots receiving poultry manure were net sinks of CH₄. Nitrous oxide emissions from manure-amended soil depended on application method; surface or soil incorporation resulted in net emission of N₂O whereas manure-amended plots under subsurface band application were net sinks of N₂O.

Incorporating manures can greatly reduce NH₃ emissions, leaving more N susceptible to emission as N₂O through nitrification and denitrification. However, reduction in NH₃ losses with incorporation means that a smaller quantity of manure is required to provide the crop N requirements, and therefore the potential for N₂O production is reduced. Subsurface injection of manure slurries into soil can result in localized anaerobic conditions surrounding the buried liquid manure, which, together with an increased degradable C pool, may result in higher CH₄ emissions than with surface applied manure. Diluting the manure or reducing the degradable C flux through solid separation or anaerobic degradation pretreatments are options to reduce CH₄ emissions from injected manure (Amon et al., 2006; Clemens et al., 2006).

A note of caution is necessary because CH₄ emissions from manure injected into soil are relatively low when compared with the reduction in NH₃ volatilization obtained through subsurface injection. Powell et al. (2011c) investigated the NH₃ volatilization mitigating potential of 3 methods of stored dairy slurry application: surface broadcast, surface broadcast followed by partial incorporation using an aerator implement, and injection. Slurry total N loss was 27.1 (20.5% as NH₃ and 6.6% as NO₃^–), 23.3 (12.0% as NH₃ and 11.3% as NO₃^–), and 9.1% (4.4% as NH₃ and 4.7% as NO₃^–), respectively. The authors reported that although slurry incorporation decreased total N loss, the conserved N did not significantly impact crop yield, crop N uptake, or soil properties at the end of the trial. They explained the lack of response to conserved N by the relatively small differences in slurry N remaining after N loss and the relatively large amount of soil N mineralization rate in the high fertility soil at the study site.

Controlling the amount of N available for nitrification and denitrification in soil as well as the availability of degradable C and soil oxidation–reduction potential are options to reduce N₂O emissions that can be achieved through the manure application method. In the first few weeks after application, manure injection often increases N₂O emission compared with surface applied manure (Dell et al., 2011). Dilution, solid separation, and anaerobic digestion pretreatments of manure before injection reduce the availability of degradable C and as a result tend to decrease N₂O emission. A number of authors have noted that wet soils tend to promote N₂O emissions and that application timing can be important. On many soils, simply avoiding application before a rain event can avoid spikes in emission rates. Maintaining soil pH above 6.5 was shown by Mkhabela et al. (2006) to help reduce N₂O emissions. Nitrous oxide emissions resulting from manure injection into soil are generally low and therefore should be weighed against the benefits of reducing NH₃ volatilization when manure is surface applied. More work is needed to further investigate overall benefits of manure application mitigation strategies.

Practices that result in increased NH₃ emissions in general will reduce the overall efficiency of the production system, reduce the amount of N being recycled on the farm, and increase the demand for N fertilizer, which could increase GHG emissions. For example, lower N₂O emissions are expected when manure is left on the soil surface compared to that incorporated into the soil, largely because a significant portion of the manure N is lost as NH₃ before undergoing nitrification and denitrification. The trade-off between reduced NH₃ volatilization and higher N₂O production may be even greater for incorporation by injection because the concentration of manure in belowground bands leads to conditions that can be more conducive to denitrification than with mixing by tillage (Dell et al., 2011). Manure incorporated in pockets in the soil through injection or shallow ditches greatly reduced NH₃ emission and resulted in reduced N losses and no CH₄ emission. Nonetheless, the increased OM in manure accelerated soil metabolism, depleting oxygen in the soil porous space and thus triggering denitrification and N₂O emissions. By using anaerobic digestion or separating manure solids, the organic content of manure is reduced, which generally results in lower emissions of N₂O after manure injection (Clemens et al., 2006; Velthof and Mosquera, 2011).

Urease and Nitrification Inhibitors.

Microbial processes that result in N₂O production can be manipulated through the use of chemical additives. Urease inhibitors are effective when applied to urine before it is mixed with soil or feces. In open lot feedlots, urease inhibitors have been reported to effectively decrease NH₃ losses. For example, Varel et al. (1999) treated feedlot pens with urease inhibitors, cyclohexylphosphoric triamide, and N-(n-butyl) thiophosphoric triamide (nBTPT). Whereas no urea was found in the control pens, the treated pens retained significant amounts of urea for up to 14 d following treatment. Treating the pens weekly for 6 wk further increased urea conservation, reducing NH₃ volatilization losses.

Nitrification inhibitors [the most widely used are dicyandiamide (DCD) and nitrapyrine] were found to reduce the amount of N₂O produced under controlled experimental or field conditions. Applied over urine and feces deposited under intensive pasture-based systems in New Zealand, nitrification inhibitors were effective in reducing N₂O emissions (de Klein et al., 1996, 2001, 2011; Di and Cameron, 2002, 2003, 2012). Luo et al. (2008b) reported up to 45% reduction in N₂O emissions from dairy cow urine applied to various soils in New Zealand by the nitrification inhibitor DCD and pointed out that the effectiveness of these compounds may be reduced under heavy rainfall. Recent national trials in New Zealand reported an average N₂O reduction by DCD of 50% (Gillingham et al., 2012).

Application of DCD has also resulted in a dramatic 68% reduction in NO₃^– leaching losses from a deep sandy soil pasture of perennial ryegrass and white clover (Di and Cameron, 2002, 2005). In addition, Ca₂⁺ and Mg₂⁺ leaching were reduced by 51 and 31%, respectively, and herbage DM yield in the urine patch areas was increased by 33% (Di and Cameron, 2005). A review by de Klein and Monaghan (2011) suggested potential reductions in NO₃^– leaching of up to 60% and N₂O emissions by up to 55% with DCD application.

It has also been pointed out that the effectiveness of nitrification inhibitors (specifically DCD) depends largely on temperature, moisture, and soil type. For example, the longevity of DCD decreases with increasing soil temperature (Kelliher et al., 2008; de Klein and Monaghan, 2011). Some studies have suggested potential increases in NH₃ volatilization and NH₄⁺ leaching due to increased NH₄⁺ accumulation in soil. It has been shown that DCD may not be effective in reducing NO₃^– leaching in soils that leach substantial amounts of NH₄⁺, which is also influenced by rainfall (de Klein and Monaghan, 2011).

Results of the combined use of nitrification and urease inhibitors have been inconclusive (Khalil et al., 2009; Zaman and Blennerhassett, 2010). Urease inhibitors inhibit urea hydrolysis to NH₄⁺ and thus directly affect substrate availability for NH₃ volatilization. A recent review of the literature using New Zealand as a case study indicated that a urease inhibitor—nBTPT—was effective in inhibiting urea hydrolysis with an average NH₃ emission reduction of 53% ranging from 11 to 93% (Saggar et al., 2012). Because NH₄⁺ is also a source of NO₃^– leaching and N₂O emission, it is expected that inhibition of urea hydrolysis will affect all 3 pathways of N loss in soil, but this has not been consistently observed (Khalil et al., 2009; Zaman and Blennerhassett, 2010).

As stated earlier, nitrification inhibitors can increase soil NH₄⁺ and thus potentially increase NH₃ losses whereas urease inhibitors prolong the stability of urea. If, however, nitrification inhibitor activity is decreased, preservation of N as urea may not decrease consequent losses of N as NO₃^– or N₂O. This scenario is also questioned on the basis of different half-lives of urease and nitrification inhibitors (de Klein and Monaghan, 2011).

Advances in plant biotechnology and microbial enzymology may offer new opportunities for reducing manure-amended soil N₂O emissions. Richardson et al. (2009), based on the idea that soil N₂O emissions from bacterial denitrification processes result from incomplete reduction of N₂O to N₂, suggested potential ways of enhancing this final step in the denitrification process: 1) increasing soil Cu availability to provide sufficient CuA and CuZ, cofactors needed for biosynthesis or assembly of nitrous oxide reductase (N₂OR), 2) improving the understanding of the regulation of N₂OR activities (enzyme repair and de novo synthesis), and 3) use of plants to “scrub” N₂O emissions by expressing bacterial N₂OR in plants.

Cover Crops.

Cover cropping can reduce soil erosion, improve soil quality and fertility, improve water, weed, disease, and pest management, and enhance plant and wildlife diversity on the farm (Lu et al., 2000; Haramoto and Gallandt, 2004). In some production systems, cover cropping can also increase crop yields (Miguez and Bollero, 2005), reduce input costs, and increase farm profitability by reducing N fertilizer use, improving P availability, and reducing weed control costs (Lu et al., 2000; Stockwell and Bitan, 2012; Kassam et al., 2012).

Reduction of N fertilizer use by growing leguminous cover crops has a direct mitigation effect on soil N₂O emissions by reducing soil NO₃^– availability and potential leaching (Christopher and Lal, 2007). Through their symbiotic relationship with Rhizobium (root nodule bacteria), legumes fix atmospheric N, converting it to NH₄⁺, which is consequently incorporated into plant AA and proteins. Thus, inclusion of legumes in plant rotation and consequent incorporation of legume residues into agricultural soils enhances plant-available inorganic N and organic soil N (Heichel, 1987). Cover crops can increase plant N uptake and decrease NO₃^– accumulation and thus reduce N₂O production through denitrification, but the results on overall GHG emissions have not been consistent. Interactions with other soil conservation practices exist (tillage system, for example) and must be considered when the goal of cover cropping is to reduce whole-farm GHG emissions. Interactions among soil conservation and management practices, however, are complex and may easily shift the balance of GHG fluxes. A study from Denmark reported a strong correlation among soil conservation practices, cover cropping, and tillage (Petersen et al., 2011). These authors concluded that reduced tillage may be an N₂O mitigation option in rotations with cover crops but that there was inconclusive evidence that the overall balance of N₂O emissions was positively affected. Another example of these interactions is the study by Garland et al. (2011). These authors demonstrated that differences in cover crop management had the capacity to affect GHG emissions; for example, mowing the cover crop produced larger peak emissions (14.1 g N₂O N/ha per d; no-till system) compared with cover crop incorporation by disking (1.6 g N₂O N/ha per d; conventional tillage system).

A review of soil organic C sequestration and GHG emissions from agricultural activities in the southeastern United States found that combining cover cropping with no tillage enhanced soil organic C sequestration compared with no tillage and no cover cropping (0.53 vs. 0.28 Mg ha/yr; Franzluebbers, 2005). Similar results have been reported for cotton (Causarano et al., 2006), but the C sequestration benefits were minimal and the effect on N₂O emissions were inconsistent in a corn–soybean rotation (Bavin et al., 2009).

Liebig et al. (2010) reported no net GHG mitigation benefit from incorporating a rye cover crop during the fallow phase of a dry land wheat cropping system under no-till management, and similar inconclusive results were reported by a Canadian study (VanderZaag et al., 2011). The simulation analysis of soil conservation practices for several crops (beans, corn, cotton, safflower, sunflower, tomato, and wheat) by De Gryze et al. (2010) in California's Central Valley found that, compared with conventional agricultural management, cover cropping had the largest potential to mitigate soil GHG fluxes resulting in a net reduction of 752 to 2,201 kg CO₂–e/ha per yr (with conservation tillage having the smallest mitigation potential). Similar conclusions were drawn by the authors for alfalfa, melon, and sunflower (De Gryze et al., 2011).

CONCLUSIONS

Conclusions for manure mitigation practices discussed in this review are summarized in Table 1. There are a number of animal and manure management practices that are feasible and can effectively reduce CH₄ and N₂O emissions from manure storage and/or land application. It is important to remember, however, that some of these practices may result in “pollution swapping” or increase NH₃ emissions. Therefore, due to numerous interactions at the animal, storage, and land application phases of the manure management process, GHG mitigation practices should not be evaluated individually in isolation but as a component of the livestock production system (farm) as a whole.

Optimizing the animal diet to improve N use efficiency, balancing N input with production level, and maintaining fiber digestibility while reducing enteric CH₄ fermentation, are important steps in reducing N₂O and CH₄ emissions from manure. Due to the complex interaction between nutrition, production, animal health, and economic performance, diet modification to reduce N inputs should be done carefully to prevent reduced fiber digestibility and maintain animal productivity.

The type of animal housing used establishes the way feces and urine are handled as well as the storage period of the manure, setting the conditions that would determine the magnitude of GHG emissions and providing opportunities for complementary mitigation practices such as composting, anaerobic digestion, biofiltration, and photocatalytic degradation. Therefore, integral animal housing and manure management system design is an important component in the implementation of GHG mitigation practices.

The choice of manure management technology has a strong influence on energy, nutrient, and GHG balances, and to obtain reliable results, the most representative and up-to-date management technology combined with data representative of the specific area or region must be considered. Overall, lowering the concentration of N in manure, preventing anaerobic conditions, or reducing the concentration of degradable manure C are successful strategies for reducing GHG emissions from manure applied to soil. Semipermeable covers are valuable for reducing NH₃, CH₄, and odor emissions but likely increase N₂O emissions; therefore, their effectiveness is not clear and results may vary widely. Impermeable membranes, such as oil layers and sealed plastic covers, are effective in reducing gaseous emissions but are not very practical.

Biofilters can be an effective tool to reduce CH₄ and NH₃ from mechanically ventilated animal housing facilities, but management of nitrification and dentrification processes in the biofilter are essential to control N₂O emissions and the overall GHG mitigation efficiency of the system. Decreasing manure pH through acidification is a promising practice to reduce NH₃ and CH₄ emissions during manure storage but might increase N₂O emissions following land application. The adoption of manure acidification as a GHG mitigation tool may be limited by the practical and legal regulatory constraints that safe handling of strong acids in a farm setting imposes. Application of acidified manure is not expected to greatly impact crop production; however, long-term impacts of land application of acidified manure on soil pH have not been reported. Composting can be an effective method for reducing GHG emissions from a range of waste materials, including animal manure, but NH₃ losses are significant.

Use of anaerobic manure digesters is a GHG mitigation strategy that has a significant potential to capture and destroy most CH₄ from manure, generate renewable energy, and provide sanitation opportunities for developing countries. Proper design and management of digestion systems is important so that they do not become net emitters of CH₄. There may also be potential for mitigating N₂O emissions following land application of the digested manure although results are contradictory. On larger farms, these systems may require large initial capital investments. The adoption of this technology on farms will heavily depend on climatic conditions and the availability of alternative sources of energy. Instruction and technical assistance are also necessary in implementing successful anaerobic digestion mitigation practices. Anaerobic digestion systems are not recommended for geographic locations with average temperatures below 15°C without supplemental heat and temperature control.

Urease inhibitors are effective in reducing urea degradation when applied to animal urine before it is mixed with feces, to drylot pen surfaces (for at least 5 to 10 d), or before urine is deposited in the soil; therefore, excluding manure managements systems that separate feces and urine or continuous treatment of the pen surface, urease inhibitors have limited opportunity to be used effectively in most animal production systems. The use of nitrification inhibitors has been demonstrated as an effective practice to reduce N₂O emission from intensive grazing systems but provide minimal if any economical benefit to the producer and therefore it is limited as an attractive mitigation practice.

Manure application techniques such as subsurface injection reduce NH₃ and CH₄ emissions but can result in increased N₂O emissions. Injection works well when combined with anaerobic digestion and solids separation by improving infiltration. Separation of manure solids and anaerobic degradation pretreatments can mitigate CH₄ emission from subsurface-applied manure, which may otherwise be higher than from surface-applied manure. Timing of the manure application (e.g., avoiding application before a rain) and maintaining soil pH above 6.5 may decrease N₂O emissions. Use of cover crops to reduce the use of commercial fertilizers by increasing N fixation, increase plant N uptake, and reduce the amount of soil N available for NH₃ volatilization and N₂O production through nitrification is very effective GHG mitigation tool, but when applied to farm systems mixed results have been obtained.

Author notes

1

This article is part of a series of articles examining methane and nitrous oxide mitigation practices for livestock operations. The article is derived in part from a published review of mitigation options for the livestock sector funded by the Food and Agriculture Organization (FAO) of the United Nations (Hristov et al., 2013b), with the consent of FAO. The views expressed in this information product are those of the authors and do not necessarily reflect the views or policies of FAO. The authors would like to thank the experts who reviewed the original FAO report (in alphabetic order): M. Doreau, R. Eckard, D. Hongmin, T. McAllister, H. Montgomery, M. Powell, S. Sommer, and M. Tibbo.

© American Society of Animal Science 2013