Synopsis 

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 these practices. Interactions among mitigation practices for individual components of livestock production systems are discussed in Gerber et al. (2013). 

Commentary 

This limited series of the occasional e-letters are comprised of (4) four articles. They will appear fortnightly and are published during November and December, though they will be accessible through our social media pages. 

Analysis 

Land Application 

Manure is a valuable resource that is best used as fertilizer. However, increased animal density accompanied by the continuous inflow of nutrients from imported feeds can lead to nutrient imbalance at the farm and watershed scales, leading to more significant water and air pollution. This nutrient imbalance is more likely to occur in intensive animal production systems. When the input of recoverable manure nutrients (the quantity of manure nutrients available for land application) grossly exceeds the assimilative capacity of the 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., the 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. There was a manure P surplus in the remaining 11%. Therefore, in some regions, manure applications 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 the 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 30cm layer were well above state threshold standards.  

Soil NO – N was >40 mg/kg for 5 of the eight dairies, and 2 were over 80 mg/kg. Such high N concentrations exceed the crop needs for optimal growth and 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 the 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 except 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 mitigate 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 manure 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 crucial pollutant due to its greater GWP (Sherlock et al., 2002). An essential 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 to 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 the 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 the soil are relatively low 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 NO –), 23.3 (12.0% as NH and 11.3% as NO –), 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 the soil and 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. Several authors have noted that wet soils tend to promote N2O emissions and that application timing can be significant.  

On many soils, simply avoiding application before a rain event can prevent spikes in emission rates. Maintaining soil pH above 6.5 was shown by Mkhabela et al. (2006) to reduce N2O emissions. Nitrous oxide emissions resulting from manure injection into the soil are generally low and, therefore, should be weighed against the benefits of lowering NH3 volatilization when manure is surface applied. More work is needed further to investigate the 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, mainly 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 maybe even more significant for incorporation by injection. 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 significantly 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 porous soil space, triggering denitrification, and N2O emissions. 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 decrease NH3 losses effectively. 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 a 45% reduction in N2O emissions from dairy cow urine applied to various soils in New Zealand by the nitrification inhibitor DCD. They 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 NO – leaching losses from a deep sandy soil pasture of perennial ryegrass and white clover (Di and Cameron, 2002, 2005). In addition, Ca + and Mg + leaching was 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 NO – leaching of up to 60% and N O emissions by up to 55% with DCD application. It has also been pointed out that the effectiveness of nitrification inhibitors (specifically DCD) depends mainly 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 effectively reduce 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 NH + 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—effectively inhibited 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 three 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 NH + and thus potentially increase NH3 losses whereas urease inhibitors prolong the stability of urea. If, however, nitrification inhibitor activity is decreased, the preservation of N as urea may not reduce consequent losses of N as NO – or N2O. This scenario is also questioned based on 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 the 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). The reduction of N fertilizer use by growing leguminous cover crops directly impacts soil N2O emissions by reducing soil NO – availability and potential leaching (Christopher and Lal, 2007). Through their symbiotic relationship with Rhizobium (root nodule bacteria), legumes fix atmospheric N, converting it to NH +, which is consequently incorporated into plant AA and proteins.  

Thus, the 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 NO – 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 quickly 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 might be an N2O mitigation option in rotations with cover crops. 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 could 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 (0.53vs. 0.28 Mg ha/yr; Franzluebbers, 2005). Similar results have been reported for cotton (Causarano et al., 2006). Still, the C sequestration benefits were minimal, and the effect on N2O emissions was 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. 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 most considerable 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). The authors drew similar conclusions for alfalfa, melon, and sunflower (De Gryze et al., 2011). 

Conclusion 

There are several 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). 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 essential 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 and 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 essential 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 the soil.  

Semipermeable covers are valuable for reducing NH3, CH4, and odor emissions but likely increase N2O emissions; therefore, their effectiveness is not apparent, and results may vary widely. Impermeable membranes, such as oil layers and sealed plastic covers, effectively reduce gaseous emissions but are not practical. Biofilters can be an effective tool to reduce CH4 and NH3 from mechanically ventilated animal housing facilities. Still, the management of nitrification and denitrification processes in the biofilter is 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. The use of anaerobic manure digesters is a GHG mitigation strategy that has a considerable 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 are essential not to 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 for 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 dry lot pen surfaces (for at least 5 to 10 d), or before the 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 economic benefit to the producer. 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.  

The infusion 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 a very effective GHG mitigation tool, but when applied to farm systems mixed results have been obtained. We want to proceed further, more in-depth on this controversial subject.  

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