Lead Authors:
Alexander N. Hristov, The Pennsylvania State University
Jane M. E. Johnson, USDA Agricultural Research Service
Contributing Authors:
Charles W. Rice, Kansas State University
Molly E. Brown, University of Maryland
Richard T. Conant, Colorado State University
Stephen J. Del Grosso, USDA Agricultural Research Service
Noel P. Gurwick, U.S. Agency for International Development
C. Alan Rotz, USDA Agricultural Research Service
Upendra M. Sainju, USDA Agricultural Research Service
R. Howard Skinner, USDA Agricultural Research Service
Tristram O. West, DOE Office of Science
Benjamin R. K. Runkle, University of Arkansas
Henry Janzen, Agriculture and Agri-Food Canada
Sasha C. Reed, U.S. Geological Survey
Nancy Cavallaro, USDA National Institute of Food and Agriculture
Gyami Shrestha, U.S. Carbon Cycle Science Program and University Corporation for Atmospheric Research
Science Lead:
Sasha C. Reed, U.S. Geological Survey
Review Editor:
Rachel Melnick, USDA National Institute of Food and Agriculture
Federal Liaisons:
Nancy Cavallaro, USDA National Institute of Food and Agriculture
Carolyn Olson (former), USDA Office of the Chief Economist


The 2018 EPA inventory (U.S. EPA 2018) attributed 567 Tg CO2e to the agricultural sector for 2015 (excluding emissions related to land use, land-use change, and forestry activities), accounting for 8.5% of total U.S. emissions.1 This proportion reflects a small increase since 1990, primarily due to increased CH4 emissions from manure management. Nitrous oxide emissions from agricultural soil management were the largest sources of GHGs at 295 Tg CO2e, and these emissions, largely due to synthetic nitrogen fertilizer applications, accounted for 77.7% of all U.S. N2O emissions. Other sources primarily included enteric fermentation (166.5 Tg CO2e), manure management (66.3 Tg CO2e and 17.7 Tg CO2e as CH4 and N2O, respectively), rice cultivation (12.3 Tg CO2e), field burning (0.4 Tg CO2e), and CO2 emissions from urea fertilization and liming (4.9 and 3.8 Tg CO2e, respectively). Within the enteric fermentation emissions, beef cattle accounted for 70.9% and dairy cattle 25.6%. Worth noting is that these numbers have been relatively stable since 1990 even though production of beef and dairy products has increased. Agricultural croplands remaining as cropland in the United States (i.e., not converted to or from other land uses) represent a small sink sequestering an estimated 0.1% of the CO2e removed from the atmosphere by land use, land-use change, and forestry activities (U.S. EPA 2018). As noted previously, agricultural practices that remove CO2 from the atmosphere include conversion from cropland to permanent pastures or hay production, reduction in acreage managed with summer fallow, adoption of conservation tillage practices, and increased applications of manure or sewage sludge. Overall, SOC increases in croplands remaining cropland and croplands converted to grasslands collectively offset losses caused by recent conversions of long-term grassland to cropland (U.S. EPA 2015, 2016, 2018; see also Ch. 12: Soils, Section 12.5.1).

In Canada, agricultural soils (55.2 million ha) contain about 4.1 petagrams (Pg) C (0- to 30-cm soil depth) and 5.5 Pg C (0- to 100-cm soil depth), as calculated from the Canadian Soil Information Service National Soil Database and reported in Ch. 12: Soils. As of 2013, Canadian agricultural land removed 11 Tg CO2 per year, which would counter about 2% of the total Canadian national GHG emissions (ECCC 2018). The reduction was attributed to decreased summer fallow and increased adoption of no-till practices in Canadian prairies. However, this value is starting to decline (e.g., down from 13 Tg CO2 in 2005) because changes in SOC stocks and fluxes tend to approach equilibrium at some point after a change in conditions.

5.5.1 Impact of Management Practices


Most cropland carbon stocks are in the soil and reflect management history and practices that increase or decrease soil carbon stocks. Integration of practices that can increase soil carbon stocks include 1) maintaining land cover with vegetation (e.g., use of deep-rooted perennials, elimination of summer fallow, and inclusion of cover crops in annual systems); 2) protecting the soil from erosion (e.g., reduced or no tillage and residue cover); and 3) improving nutrient management (Srinivasarao et al., 2015; Swan et al., 2015). The magnitude and longevity of carbon stock changes have strong environmental and regional differences that are subject to subsequent changes in management practices. Conversely, practices that convert lands from perennial systems, such as converting retired or other lands to row crops, consistently show release of stored carbon back to the atmosphere (Gelfand et al., 2011; Huang et al., 2002). Other management practices with the potential to release stored carbon are inadequate return of crop residues (e.g., Blanco-Canqui and Lal 2009), aggressive tillage (Conant et al., 2007), over application of nitrogen fertilizer, and burning of crop residue (Robertson and Grace 2004; Wang et al., 2011).

The timescale for carbon storage in soils is a critical factor for GHG mitigation. Numerous estimates of the rates and potential magnitude of long-term soil carbon accumulation, storage, and sequestration related to management have been reviewed and presented (e.g., Minasny et al., 2017; Paustian et al., 2016; Sperow 2016; Stockmann et al., 2013; Swan et al., 2015). Management practices that increase carbon inputs include planting high-residue crops and returning crop biomass to the soil; minimizing or eliminating summer fallow (particularly bare fallow); adding cover crops to reduce winter fallow; extending and intensifying cropping rotations (e.g., double-cropping or relay cropping and adding forage perennials); retiring marginal lands to perennials; and adding perennials in buffer strips, field borders, filter strips, grassed waterways, vegetative barriers, and herbaceous wind barriers (e.g., Mosier et al., 2006; Paustian et al., 2016; Sainju et al., 2010; Sperow 2016). Swan et al. (2015) estimated carbon storage rates of 0.42 to 0.95 Mg C per hectare per year among conservation practices that shift to perennials (e.g., retiring marginal land or planting perennials as barriers or borders), while inclusion of cover crops was estimated to accrue 0.15 to 0.27 Mg C per hectare per year. Practices that eliminate summer fallow can increase SOC directly by increasing carbon input or modifying microclimate (i.e., temperature and water), a practice that can decrease mineralization rates by reducing temperature and water content (Halvorson et al., 2002; Sainju et al., 2015).

Numerous publications have reported that ­no-tillage practices store more carbon in soil than those using conventional tillage (e.g., Paustian et al., 2016; Sperow 2016; West and Post 2002). Conversely, others have disputed this claim, especially when including soil carbon measurements deeper than 30 cm (e.g., Baker et al., 2007; Luo et al., 2010; Powlson et al., 2014; Ugarte et al., 2014). No-tillage and other conservation practices were developed to control soil erosion, and this co-benefit is well established. Erosion removes soil carbon from farm fields and relocates that carbon to other parts of the landscape; the amount of this transported carbon that is sequestered in sediments compared to the amount converted to CO2 or CH4 is difficult to estimate (Doetterl et al., 2016). In Ch. 12: Soils, the role of soil erosion is discussed in greater detail and suggests that burial of eroded carbon constitutes a small sink. Comparing SOC sequestration rates from a system managed without tillage to a system with tillage results in negative, neutral, and positive rates of SOC sequestration: 1) 27 ± 19 Mg SOC per hectare per year, (n = 49; Liebig et al., 2005), 2) 0.40 ± 61 Mg SOC per hectare per year (n = 44; Johnson et al., 2005), or 0.45 ± 0.04 Mg SOC per hectare per year (n = 147; Franzluebbers 2010). Likewise, studies using eddy covariance techniques report divergent responses to tillage. For example, Bernacchi et al. (2005) demonstrated that no-tillage agriculture on clay-rich soil built SOC, whereas others (Baker and Griffis 2005; Chi et al., 2016; Verma et al., 2005) used gas exchange techniques to suggest conservation or no-tillage systems were near carbon neutral. In another review, Collins et al. (2012) found that carbon sequestration rates varied from no measurable increase (Staben et al., 1997) to 4 Mg C per hectare per year (Lee et al., 2007), varying with depth monitored, study duration, fertilizer formulation, and location. Several rationales have been postulated for this variability. If sampling depth is shallower than the tillage depth, the apparent change in SOC may be an artifact of sampling depth (Baker et al., 2007) or caused by residue redistribution (Staricka et al., 1991) and vertical stratification of soil carbon (Luo et al., 2010). Meta-analyses by Luo et al. (2010) and Ugarte et al. (2014) suggest that other factors contributing to variability in SOC sequestration include climatic and soil properties interacting with management factors (e.g., cropping frequency, crop rotation diversity, nitrogen, and drainage) along with impacts on rooting depth and above- and belowground biomass, as well as soil heterogeneity and the long time frames required to find a definitive increase or decrease in SOC. Collectively, the evidence indicates that adoption of no tillage may store more carbon, especially in the soil surface, compared to storage with conventional tillage. However, conclusively measuring short-term changes is difficult because of soil heterogeneity and slow rates of change (also discussed in Ch. 12: Soils). In particular, increased N2O or CH4 emissions have been shown to occur for as many as 10 years after no-till adoption (Six et al., 2004), though this effect is greater and more consistent in medium to poorly drained soils (Rochette 2008). Thus, quantifying GHG mitigation by management also must account for changes in N2O and CH4, which can occur coincidently with changes in soil carbon storage (VandenBygaart 2016).

From a carbon emissions perspective, biofuels have received a great deal of attention because of their potential to produce a more carbon neutral liquid fuel relative to fossil fuels. Biofuels from annual crops currently supply about 5% of U.S. energy use, mostly from corn grain ethanol (~36% of the corn grain harvest) and soy biodiesel (~25 % of the soybean harvest; USDA 2017b). Although the potential for reduced GHG emissions with biofuels is compelling, some life cycle assessment analyses suggest that corn grain ethanol has marginally lower (or even greater) GHG emissions compared with those from fossil fuels (e.g., Del Grosso et al., 2014; Fargione et al., 2008). However, more recent studies suggest that currently available technologies can achieve greater GHG reductions of 27% to 43% compared to gasoline when assessed on an energy equivalent basis (Canter et al., 2015; Flugge et al., 2017). Reasons for reduced net GHG intensity for grain- and oil-based biofuels include improved crop-management practices and diminished emissions from land-use change because most of the yield gap from diverting food and feed crops to biofuel feedstocks has been met by increasing per-unit area yields, taking into account the benefits of co-products (e.g., using dried distiller grains for livestock feed) and implementing more efficient feedstock conversion technologies (Flugge et al., 2017). Typically, cellulosic biomass conversion technologies are considered too expensive to compete with liquid fuels derived from other sources (Winchester and Reilly 2015), but innovations at all levels are advancing conversion technology. The impact of cellulosic biofuels on the carbon cycle (Fulton et al., 2015) will depend on ensuring that appropriate mitigation strategies are followed during feedstock choice (perennial or annual) and cultivation (e.g., related to soil carbon stock changes [Blanco-Canqui 2013; Johnson et al., 2012, 2014; Qin et al., 2015]), transportation, and conversion to biofuels (U.S. DOE 2016).

Co-Benefits of Conservation Management

Many common conservation practices improve soil aeration, aggregate stability, and nutrient reserves, while modulating temperature and water and increasing microbial activity and diversity. As a result, soil under some conservation-management regimes can be more resilient to climate variability and more productive (Lal 2015; Lehman et al., 2015). For example, adoption of practices that can conserve soil carbon (e.g., perennial crops, cover crops, and no tillage) may reverse the effects of tillage-intense systems associated with environmental and soil degradation (Mazzoncini et al., 2011). Plant material maintained on the soil surface improves soil physical properties (e.g., Johnson et al., 2016), nutrient availability, and microbial biomass and activity (Feng et al., 2003; Weyers et al., 2013). These improvements result in enhanced soil and water quality and soil productivity (Franzluebbers 2008). Cover crops improve soil health by increasing microbial diversity, biomass, and activity (Bronick and Lal 2005; Lehman et al., 2012, 2015; Schutter and Dick 2002); they also improve soil aggregation, water retention, and nutrient cycling (Blanco-Canqui et al., 2013; Drinkwater et al., 1998; Kladivko et al., 2014; Liebig et al., 2005; Sainju et al., 2006). Thus, there are management practices that simultaneously benefit a number of soil health and carbon storage attributes.

5.5.2 Emissions Reduction


Enteric fermentation and manure management represent 44% of the 2015 agricultural GHG emissions in the United States (U.S. EPA 2018) and 36% and 58% of the agricultural emissions in Canada and Mexico, respectively (FAOSTAT 2017). Of the total U.S. GHG emissions in 2015, however, emissions from enteric fermentation and manure management made up only 3.8% (U.S. EPA 2018). Methane mitigation practices for livestock include practices related to reducing emissions from enteric fermentation (i.e., cattle) and manure management (i.e., cattle and swine) as discussed by Hristov et al. (2013b) and Herrero et al. (2016). Increasing forage digestibility and digestible forage intake generally will reduce CH4 emissions from rumen fermentation (and stored manure) when scaled per unit of animal product. Enteric CH4 emissions may be reduced when corn silage replaces grass silage in the diet. Legume silages also may have an advantage over grass silage because of their lower fiber content and the additional benefit of reducing or replacing inorganic nitrogen fertilizer use. Dietary lipids are effective in reducing enteric CH4 emissions, but the applicability of this practice will depend on its cost and effects on feed intake, production, and milk composition in dairy cows. Inclusion of concentrate feeds in the diet of ruminants likely will decrease enteric CH4 emissions per unit of animal product, particularly when the inclusion is above 40% of dry matter intake.

A number of feed additives, such as nitrates, also can effectively decrease enteric CH4 emissions in ruminants. Because these additives can be toxic to the animals, proper adaptation is critical. However, nitrates may slightly increase N2O emissions, which decreases their overall mitigating effect by 10% to 15% (Petersen et al., 2015). Through their effect on feed efficiency, ionophores are likely to have a moderate CH4-mitigating effect in ruminants fed high-grain or grain-forage diets. Some direct-fed microbial products, such as live yeast or yeast culture, might have a moderate CH4-mitigating effect by increasing animal productivity and feed efficiency, but the effect is expected to be inconsistent. Vaccines against rumen methanogens may offer mitigation opportunities in the future, but the extent of CH4 reduction appears small, and adaptation and persistence of the effect are unknown. A recently discovered enteric CH4 inhibitor, 3-nitrooxypropanol, has shown promising results with both beef and dairy cattle. Under ­industry-relevant conditions, the inhibitor persistently decreased enteric CH4 emissions by 30% in dairy cows, without negatively affecting animal productivity (Hristov et al., 2015). Similar or even greater mitigation potential has been reported for beef cattle (Romero-Perez et al., 2015). If its effectiveness is proven in long-term studies, this mitigation practice could lead to a substantial reduction of enteric CH4 emissions from the ruminant livestock sector.

Animal management also can have an impact on the intensity (i.e., emissions per unit of animal product) of CH4 emissions from livestock systems. For example, increasing animal productivity through genetic selection for feed efficiency can be an effective strategy for reducing CH4 emission intensity. Other management practices for significantly decreasing total GHG emissions in beef and other meat production systems include reducing age at slaughter of finished cattle and the number of days that animals consume feed in the feedlot. Improved animal health, reduced mortality and morbidity, and improved reproductive performance also can increase herd productivity and reduce GHG emission intensity in livestock production (Hristov et al., 2013a).

Several practices are known to reduce CH4 emissions from manure but cannot be considered in isolation of other GHG sources and pollutants such as N2O and ammonia (NH3). Practices such as the use of solid manure storage and composting can reduce CH4 emissions, but N2O and NH3 emissions will increase, and the end result may not be a reduction in overall GHG emissions. Mitigation of carbon emissions also may have tradeoffs with other pollutants including other gaseous emissions, nutrient leaching to groundwater, and nutrient runoff to surface waters. For example, eliminating long-term manure storage can greatly reduce CH4 emissions, but daily spreading of manure throughout the year can cause greater nutrient runoff. Mitigation strategies must be considered from a whole-farm perspective to ensure a net environmental benefit (Montes et al., 2013).

Potential CH4 mitigation strategies include manure solids separation, aeration, acidification, biofiltration, composting, and anaerobic digestion (Montes et al., 2013). Removal of solids from liquid manure reduces available carbon for methanogenesis, and composting or storing the solids in a stack under more aerobic conditions reduces total CH4 emissions. For long-term manure storage, covers likely will become mandatory to reduce NH3, CH4, and N2O emissions. Semipermeable covers such as the natural crust on slurry manure or added floating materials such as straw, wood chips, expanded clay pellets, and some types of plastic can reduce CH4 and NH3 emissions from storage by 30% to 80%, but they also may increase N2O emissions. Greater reductions and perhaps near elimination of emissions can be achieved by sealing the cover and using a flare to convert the accumulated CH4 to CO2. Anaerobic digesters also can be used to enhance CH4 production, capturing the produced biogas and using it on the farm to heat water and generate electricity. Extracting the carbon from manure reduces storage emissions, and the reduction in purchased gas and electricity provides other off-farm environmental benefits. Composting solid manure in aerated windrows can greatly reduce CH4 emissions, but this processing will increase NH3 and N2O emissions (Montes et al., 2013).

Experimental processes of acidification and biofiltration show potential for reducing CH4 emissions if practical and economical systems can be developed (Montes et al., 2013). Decreasing the pH of manure reduces NH3 and CH4 emissions, but the cost of the acid, safety in handling, and difficulty in maintaining the low pH all deter its use. Biofiltration can extract CH4 from ventilation air in barns, but the large size and cost preclude adoption. Biofilters also may create N2O emissions, offsetting some of the carbon reduction benefits.

Rice Production

Rice emits four to five times more CH4 and N2O to the atmosphere (Linquist et al., 2012) and uses two to three times more water per kg than other cereals (Bouman et al., 2007; Tuong et al., 2005). Sustainably oriented production practices have been developed with the goal of mitigating the environmental impact of rice and improving the economic benefits through reductions in production costs. These practices include the irrigation management practice of alternate wetting and drying (AWD) or intermittent flooding, whereby the soil surface is allowed to dry for several days to a week before rewetting in midseason. This practice can be repeated up to five times during the growing season without reducing harvest yield. The concurrent re-oxygenation of the soil layer keeps CH4 emissions low, and studies have shown that water-saving irrigation methods such as AWD reduce net CH4 emissions produced under water-saturated conditions (Linquist et al., 2015; Rogers et al., 2013). Even one 6-day, midseason drainage event, temporarily reducing anaerobic soil conditions, can reduce post-drainage CH4 emissions by 64% with no evident effect on yield (Sigren et al., 1997). This practice also has the co-benefit of reducing grain arsenic concentrations because it changes the soil reduction-oxidation (redox) potential (Linquist et al., 2015). Other irrigation techniques that reduce the inundated soil period also will reduce the CH4 emissions from rice paddies. These methods include the use of drill-seeding rather than water-seeding or transplanting rice (Pittelkow et al., 2014) and carry the additional benefit of reducing the pumping requirements of irrigation water; thus, they will reduce GHG production associated with the energy use of burning fossil fuels—whether through diesel or indirectly through electricity generation. The reduced pumping benefits are particularly true in rice production regions of the Midsouth that are distinct from those in California, where irrigation needs are met from gravity-fed reservoirs draining the Sierra Nevada mountains. However, for any CH4-reducing rice production regime, care must be taken to keep N2O emissions low. As indicated, rates of N2O emissions are particularly sensitive to inputs from nitrogen fertilization, fallow-season field conditions, and midseason or season-end drainage events (Pittelkow et al., 2013). In many cases, both CH4 and N2O are released in any drainage event, with end-of-season drainage transferring 10% of seasonal CH4 and 27% of seasonal N2O to the atmosphere as entrapped gases are released from the soil.

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