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Agricultural lands occupy 37% of the earth's land surface. Agriculture accounts for 52 and 84% of global anthropogenic methane and nitrous oxide emissions. Agricultural soils may also act as a sink or source for CO2, but the net flux is small. Many agricultural practices can potentially mitigate greenhouse gas (GHG) emissions, the most prominent of which are improved cropland and grazing land management and restoration of degraded lands and cultivated organic soils. Lower, but still significant mitigation potential is provided by water and rice management, set-aside, land use change and agroforestry, livestock management and manure management. The global technical mitigation potential from agriculture (excluding fossil fuel offsets from biomass) by 2030, considering all gases, is estimated to be approximately 5500–6000Mt CO2-eq.yr−1, with economic potentials of approximately 1500–1600, 2500–2700 and 4000–4300Mt CO2-eq.yr−1 at carbon prices of up to 20, up to 50 and up to 100 US$ t CO2-eq.−1, respectively. In addition, GHG emissions could be reduced by substitution of fossil fuels for energy production by agricultural feedstocks (e.g. crop residues, dung and dedicated energy crops). The economic mitigation potential of biomass energy from agriculture is estimated to be 640, 2240 and 16 000Mt CO2-eq.yr−1 at 0–20, 0–50 and 0–100 US$ t CO2-eq.−1, respectively.
Agriculture releases to the atmosphere significant amounts of CO2, CH4 and N2O (Cole et al. 1997; IPCC 2001; Paustian et al. 2004). Carbon dioxide is released largely from microbial decay or burning of plant litter and soil organic matter (Janzen 2004; Smith 2004b). Methane is produced when organic materials decompose in oxygen-deprived conditions, notably from fermentative digestion by ruminant livestock, stored manures and rice grown under flooded conditions (Mosier et al. 1998). Nitrous oxide is generated by the microbial transformation of nitrogen in soils and manures, and is often enhanced where available N exceeds plant requirements, especially under wet conditions (Smith & Conen 2004; Oenema et al. 2005). Agricultural greenhouse gas (GHG) fluxes are complex and heterogeneous, but the active management of agricultural systems offers possibilities for mitigation. Many of these mitigation opportunities use current technologies and can be implemented immediately. In this paper, we use the latest datasets and techniques to make the first estimates of agricultural GHG mitigation potential for 2030 that include all GHGs with breakdowns for all global regions and all gases.
Opportunities for mitigating GHGs in agriculture fall into three broad categories based on the underlying mechanism:
Many practices have been advocated to mitigate emissions through the mechanisms cited above. Often a practice will affect more than one gas, by more than one mechanism, sometimes in opposite ways, so that the net benefit depends on the combined effects on all gases (Robertson & Grace 2004; Schils et al. 2005). In addition, the temporal pattern of influence may vary among practices or among gases for a given practice; some emissions are reduced indefinitely, other reductions are temporary (Marland et al. 2003a; Six et al. 2004). Where a practice affects radiative forcing through other mechanisms such as aerosols or albedo, those impacts also need to be considered (Marland et al. 2003b; Andreae et al. 2005). The impacts of various mitigation options considered are summarized in table 1. The most important options are discussed below.
Croplands, because they are often intensively managed, offer many opportunities to impose practices that reduce net emissions of GHGs (table 1). Mitigation practices in cropland management include the following partly overlapping categories.
Improved agronomic practices that increase yields and generate higher inputs of residue C can lead to increased soil C storage (Follett 2001). Examples of such practices include: using improved crop varieties; extending crop rotations, notably those with perennial crops which allocate more C below-ground; and avoiding or reducing use of bare (unplanted) fallow (West & Post 2002; Lal 2003, 2004a; Freibauer et al. 2004; Smith 2004a,b). Adding more nutrients, when deficient, can also promote soil C gains (Alvarez 2005), but the benefits from N fertilizer can be offset by higher emissions of N2O from soils and CO2 from fertilizer manufacture (Schlesinger 1999; Robertson & Grace 2004; Gregorich et al. 2005).
Emissions can also be reduced by adopting less intensive cropping systems, which reduce reliance on pesticides and other inputs (and therefore the GHG cost of their production; Paustian et al. 2004). An important example is the use of rotations with legume crops (Izaurralde et al. 2001; West & Post 2002), which reduce reliance on inputs of N, though legume-derived N can also be a source of N2O (Rochette & Janzen 2005)
A third group of agronomic practices are those that provide temporary vegetative cover between agricultural crops. These ‘catch’ or ‘cover’ crops add C to soils (Barthès et al. 2004; Freibauer et al. 2004) and may also extract plant-available N unused by the preceding crop, thereby reducing N2O emissions.
Nitrogen applied in fertilizers and manures is not always used efficiently by crops (Cassman et al. 2003; Galloway et al. 2003). Improving this efficiency can reduce emissions of N2O, generated by soil microbes largely from surplus N and it can indirectly reduce emissions of CO2 from N fertilizer manufacture (Schlesinger 1999). Practices that improve N use efficiency include: adjusting application rates based on precise estimation of crop needs (e.g. precision farming); using slow-release fertilizer forms or nitrification inhibitors (which slow the microbial processes leading to N2O formation); avoiding time delays between N application and plant N uptake (improved timing); placing the N more precisely into the soil to make it more accessible to crops roots; avoiding excess N applications, or eliminating N applications where possible (Cole et al. 1997; Dalal et al. 2003; Paustian et al. 2004; Robertson 2004; Monteny et al. 2006).
Advances in weed control methods and farm machinery now allow many crops to be grown with minimal tillage (reduced tillage) or without tillage (no till). These practices are now increasingly used throughout the world (e.g. Cerri et al. 2004). Since soil disturbance tends to stimulate soil C losses through enhanced decomposition and erosion, reduced- or no-till agriculture often results in soil C gain, though not always (West & Post 2002; Alvarez 2005; Gregorich et al. 2005; Ogle et al. 2005). Adopting reduced or no till may also affect emissions of N2O, but the net effects are inconsistent and not well-quantified globally (Cassman et al. 2003; Smith & Conen 2004; Helgason et al. 2005; Li et al. 2005). The effect of reduced tillage on N2O emissions may depend on soil and climatic conditions: in some areas reduced tillage promotes N2O emissions; elsewhere it may reduce emissions or have no measurable influence (Marland et al. 2001).
Systems that retain crop residues also tend to increase soil C because these residues are the precursors for soil organic matter, the main store of carbon in the soil. Avoiding the burning of residues, for instance mechanizing the harvesting of sugarcane, which eliminates the need for pre-harvest burning (Cerri et al. 2004), also avoids emissions of aerosols and GHGs generated from fire.
About 18% of the world's croplands now receive supplementary water through irrigation (Millennium Ecosystem Assessment 2005). Expanding this area or using more effective irrigation measures can enhance C storage in soils through enhanced yields and residue returns (Follett 2001; Lal 2004a). But some of these gains may be offset by CO2 from energy used to deliver the water (Schlesinger 1999; Mosier et al. 2005) or from N2O emissions from higher moisture and fertilizer N inputs (Liebig et al. 2005), though the latter effect has not been widely measured.
Drainage of agricultural lands in humid regions can promote productivity (and hence soil C) and perhaps also suppress N2O emissions by improving aeration (Monteny et al. 2006). Any nitrogen lost through drainage, however, may be susceptible to loss as N2O (Reay et al. 2003).
Cultivated wetland rice soils emit significant quantities of methane (Yan et al. 2003). Emissions during the growing season can be reduced by many practices (Yagi et al. 1997; Wassmann et al. 2000; Aulakh et al. 2001). For example, draining the wetland rice once or several times during the growing season effectively reduces CH4 emissions (Smith & Conen 2004; Yan et al. 2003), although this benefit may be partly offset by higher N2O emissions, and the practice may be constrained by water supply. Rice cultivars with low exudation rates could offer an important methane mitigation option (Aulakh et al. 2001). In the off-rice season, methane emissions can be reduced by improved water management, especially by keeping the soil as dry as possible and avoiding waterlogging (Cai et al. 2000, 2003; Kang et al. 2002; Xu et al. 2003).
Methane emissions can also be reduced by adjusting the timing of organic residue additions (e.g. incorporating organic materials in the dry period rather than in flooded periods; Xu et al. 2000; Cai & Xu 2004), composting the residues before incorporation or producing biogas for use as fuel for energy production (Wang & Shangguan 1996; Wassmann et al. 2000).
Agroforestry is the production of livestock or food crops on land that also grows trees, either for timber, firewood or other tree products. It includes shelter belts and riparian zones/buffer strips with woody species. The standing stock of carbon above ground is usually higher than the equivalent land use without trees, and planting trees may also increase the soil carbon sequestration (Guo & Gifford 2002; Paul et al. 2003; Oelbermann et al. 2004; Mutuo et al. 2005), though the effects on N2O and CH4 emissions are not well known (Albrecht & Kandji 2003).
One of the most effective methods of reducing emissions is to allow or encourage the reversion of cropland to another land cover, typically one similar to the native vegetation. The conversion can occur over the entire land area (‘set-asides’) or in localized spots such as grassed waterways, field margins or shelterbelts (Follett 2001; Ogle et al. 2003; Falloon et al. 2004; Freibauer et al. 2004; Lal 2004a). Such land cover change often increases storage of C; for example, converting arable cropland to grassland typically results in the accrual of soil C owing to lower soil disturbance and reduced C removal in harvested products. Compared to cultivated lands, grasslands may also have reduced N2O emissions from lower N inputs and higher rates of CH4 oxidation, though recovery of oxidation may be slow (Paustian et al. 2004).
Similarly, converting drained croplands back to wetlands can result in rapid accumulation of soil carbon (removal of atmospheric CO2), although this conversion may stimulate CH4 emissions, because waterlogging creates anaerobic conditions (Paustian et al. 2004). Planting trees can also reduce emissions, but these practices are considered under agroforestry (see §2a(vi)) afforestation or reforestation.
Since land cover (or use) conversion comes at the expense of lost agricultural productivity, it is usually an option only on surplus agricultural land or on croplands of marginal productivity.
Grazing lands occupy much larger areas than croplands (FAOSTAT 2006), but are usually managed less intensively. The following list provides some examples of practices to reduce GHG emissions and enhance removals.
The intensity and timing of grazing can influence the growth, C allocation and flora of grasslands, thereby affecting the amount of C accrual in soils (Conant et al. 2001, 2005; Conant & Paustian 2002; Freibauer et al. 2004; Reeder et al. 2004). Carbon accrual on optimally grazed lands is often greater than on ungrazed or overgrazed lands (Rice & Owensby 2001; Liebig et al. 2005). The effects are inconsistent, however, owing to the many types of grazing practices employed and the diversity of plant species, soils and climates involved (Schuman et al. 2001; Derner et al. 2006). The influence of grazing intensity on emission of non-CO2 gases is not well established, apart from the indirect effects from adjustments in livestock numbers.
As for croplands, C storage in grazing lands can be improved by a variety of measures that promote productivity. For instance, alleviating nutrient deficiencies by fertilizer or organic amendments increases plant litter returns and, hence, soil C storage (Conant et al. 2001; Schnabel et al. 2001). Adding nitrogen, however, may stimulate N2O emissions (Conant et al. 2005) thereby offsetting some of the benefits. Irrigating grasslands, similarly, can promote soil C gains (Conant et al. 2001), though the net effect of this practice depends also on emissions from energy use and other related activities on the irrigated land (Schlesinger 1999).
Practices that tailor nutrient additions to plant uptake, like those described for croplands, can reduce emissions of N2O (Follett et al. 2001; Dalal et al. 2003). Management of nutrients on grazing lands, however, may be complicated by deposition of faeces and urine from livestock, which are neither easily controlled nor as uniformly applied as nutritive amendments in croplands (Oenema et al. 2005).
Biomass burning (not to be confused with bioenergy, where biomass is combusted off-site for energy) contributes to climate change in several ways. Firstly, it releases GHGs, notably CH4, and to a lesser extent, N2O (the CO2 released is of recent origin, is reabsorbed by vegetation and is usually not counted). Secondly, it generates hydrocarbon and reactive nitrogen emissions, which react to form tropospheric ozone. Smoke contains a range of aerosols which can have either warming or cooling effects on the atmosphere, though the net effect is thought to be positive radiant forcing (Andreae 2001; Andreae & Merlet 2001; Menon et al. 2002; Anderson et al. 2003; Jones et al. 2003; Andreae et al. 2005; Venkataraman et al. 2005). Thirdly, fire blackens the land surface, reducing its albedo for several weeks, causing a warming (Beringer et al. 2003). Fourthly, burning can affect the proportions of woody versus grass cover, notably in savannas, which occupy approximately one-eighth of the global land surface. Reducing the frequency or intensity of fires typically leads to increased tree and shrub cover, resulting in higher landscape C density in soil and biomass (Scholes & van der Merwe 1996). This woody plant encroachment mechanism is higher initially, but saturates over 20–50 years, whereas avoided CH4 and N2O emissions are ongoing as long as the fires are suppressed.
Mitigation of radiant forcing involves reducing the frequency or extent of fires through more effective fire suppression (Korontzi et al. 2003); reducing the fuel load by vegetation management; and burning at a time of year when less CH4 and N2O are emitted (Korontzi et al. 2003). Although most agricultural-zone fires are ignited by humans, there is evidence that the area burned is ultimately under climatic control (van Wilgen et al. 2004). In the absence of human ignition, the fire-prone ecosystems would be lit by other agents.
Introducing grass species with higher productivity or C allocation to deeper roots has been shown to increase soil C. For example, establishment of deep-rooted grasses in savannas has been reported to yield very high rates of C accrual (Fisher et al. 1994), although the applicability of these results has not been widely confirmed (Davidson et al. 1995; Conant et al. 2001). Introducing legumes into grazing lands can promote soil C storage (Soussana et al. 2004), through enhanced productivity from the associated N inputs, and perhaps also reduce N2O emissions if the biological N2 fixation displaces the need for fertilizer N.
Lands used for grazing also emit GHGs from the livestock, notably CH4 from ruminants and their manures. Practices for reducing these emissions are considered under §2e below.
Organic soils contain high densities of C, accumulated over many centuries, because decomposition is suppressed by absence of oxygen under flooded conditions. To be used for agriculture, these soils are drained, which aerates the soil, favouring decomposition and therefore high fluxes of CO2 and N2O. Methane emissions are usually suppressed after draining, but this effect is far outweighed by pronounced increases in N2O and CO2 (Kasimir-Klemedtsson et al. 1997). Emissions on drained organic soils can be reduced to some extent by practices such as avoiding row crops and tubers, avoiding deep ploughing and maintaining a more shallow water table, but the most important mitigation practice, probably, is avoiding the drainage of these soils in the first place, or re-establishing a high water table where GHG emissions are still high (Freibauer et al. 2004).
A large fraction of agricultural lands have been degraded by erosion, excessive disturbance, organic matter loss, salinization, acidification or other processes that curtail productivity (Batjes 1999; Lal 2001, 2003, 2004b; Foley et al. 2005). Often the C storage in these soils can be at least partly restored by practices that reclaim productivity including: revegetation (e.g. planting grasses); improving fertility by nutrient amendments; applying organic substrates such as manures, biosolids and composts; reducing tillage and retaining crop residues; and conserving water (Bruce et al. 1999; Lal 2001, 2004b; Olsson & Ardo 2002; Paustian et al. 2004). Where these practices involve higher nitrogen amendments, the benefits of C sequestration may be partly offset by higher N2O emissions.
Livestock, predominantly ruminants such as cattle and sheep, are important sources of CH4, accounting for approximately 18% of global anthropogenic emissions of this gas (US-EPA 2006). The methane is produced primarily by enteric fermentation and voided by eructation (Murray et al. 1976; Kennedy & Milligan 1978; Crutzen 1995). Practices for reducing CH4 emissions from this source fall into three general categories: improved feeding practices, use of specific agents or dietary additives, and longer term management changes and animal breeding.
Methane emissions can be reduced by feeding more concentrates, normally replacing forages (Blaxter & Clapperton 1965; Johnson & Johnson 1995; Lovett et al. 2003; Beauchemin & McGinn 2005). Although concentrates may increase daily methane emissions, emissions per kilogram feed intake and per kilogram product are almost invariably reduced. The net benefit, however, depends on reduced animal numbers or younger age at slaughter for beef animals and on how the practice affects emissions when producing and transporting the concentrates (Phetteplace et al. 2001; Lovett et al. 2006).
Other practices that can reduce CH4 emissions include: adding oils to the diet (e.g. Machmülller et al. 2000; Jordan et al. 2004); improving pasture quality, especially in less developed regions, because it improves animal productivity and reduces the proportion of energy lost as CH4 (Leng 1991; McCrabb et al. 1998; Alcock & Hegarty 2005); and optimizing protein intake to reduce N excretion and N2O emissions (Clark et al. 2005).
A wide range of specific agents, mostly aimed at suppressing methanogenesis, have been proposed as dietary additives to reduce CH4 emissions as follows:
Increasing productivity through breeding and better management practices spreads the energy cost of maintenance across a greater feed intake, often reducing methane output per kilogram of animal product (Boadi et al. 2004). With improved efficiency, meat-producing animals reach slaughter weight at a younger age, with reduced lifetime emissions (Lovett & O'Mara 2002). The whole system effects of such practices are not entirely clear, however; for example, selecting for higher yield might reduce fertility, requiring more replacement animals (Lovett et al. 2006).
Animal manures can release significant amounts of N2O and CH4 during storage, but the magnitude of these emissions varies. Methane emissions from manure stored in lagoons or tanks can be reduced by cooling or covering the sources, or by capturing the CH4 emitted (Clemens & Ahlgrimm 2001; Monteny et al. 2001, 2006; Paustian et al. 2004). The manures can also be digested anaerobically to maximize retrieval of CH4 as an energy source (Clemens & Ahlgrimm 2001; Clemens et al. 2006). Storing and handling the manures in solid rather than liquid form can suppress CH4 emissions, but may increase N2O formation (Paustian et al. 2004). Preliminary evidence suggests that covering manure heaps can reduce N2O emissions (Chadwick 2005). For most animals worldwide, there is limited opportunity for manure management, treatment or storage—excretion happens in the field and handling for fuel or fertility amendment occurs when it is dry and methane emissions are negligible (Gonzalez-Avalos & Ruiz-Suarez 2001). To some extent, emissions from manure might be curtailed by altering feeding practices (Külling et al. 2003) or by composting the manure (Pattey et al. 2005), but these mechanisms and the system-wide influence have not been widely explored. Manures also release GHGs, notably N2O, after application to cropland or deposition on grazing lands, but the practices for reducing these emissions are considered above in §2a,b.
Increasingly, agricultural crops and residues are seen as sources of feedstocks for energy to displace fossil fuels. A wide range of materials have been proposed for use, including grain, crop residue, cellulosic crops (e.g. switchgrass, sugarcane) and various tree species (Cerri et al. 2004; Edmonds 2004; Paustian et al. 2004; Sheehan et al. 2004; Dias de Oliveira et al. 2005; Eidman 2005). These products can be burned directly, but often are processed further to generate liquid fuels such as ethanol or diesel fuel (Richter 2004). These fuels release CO2 when burned, but this CO2 is of recent atmospheric origin (via photosynthesis) and displaces CO2 which otherwise would have come from fossil C. The net benefit to atmospheric CO2, however, depends on energy used in growing and processing the bioenergy feedstock (Spatari et al. 2005).
The interactions of an expanding bioenergy sector with other land uses, and impacts on agro-ecosystem services such as food production, biodiversity, soil and nature conservation, and carbon sequestration have not yet been adequately studied, but bottom up approaches (Smeets et al. 2007) and integrated assessment modelling (Hoogwijk 2004; Hoogwijk et al. 2005) offer opportunities to improve understanding. Latin America, sub-Saharan Africa and Eastern Europe are promising regions for bioenergy, with additional long-term contributions from Oceania and East and NE Asia. The technical potential for biomass production may be developed at low production costs in the range of 2 US$GJ−1 (Rogner et al. 2000; Hoogwijk 2004).
Major transitions are required to exploit the large potential for bioenergy. Improving agricultural efficiency in developing countries is a key factor. It is still uncertain to what extent, and how fast, such transitions can be realized in different regions. Under less favourable conditions the (regional) bioenergy potential(s) could be quite low. In addition, it should be noted that technological developments (in conversion, as well as long-distance biomass supply chains such as those involving intercontinental transport of biomass-derived energy carriers) can dramatically improve competitiveness and efficiency of bioenergy (Hamelinck et al. 2004; Faaij 2006).
It is theoretically possible to increase the storage of carbon in long-lived agricultural products (e.g. strawboards, wool, leather and bioplastics) but with an increase in C held in these products from 37 to 83Mt C yr−1 over the past 40 years and assuming a first-order decay rate of 10–20% per year, this is estimated to be a global net annual removal of 3–7Mt CO2 from the atmosphere which is negligible when compared with other mitigation measures, and the option is not considered further here.
Many mitigation practices (§2) affect more than one GHG and the best available data have been used to estimate the impact on all GHGs of each practice. When assessing the impact of agriculture on changes in GHG emissions, it is important to consider the impacts on all GHGs together (Robertson et al. 2000; Smith et al. 2001; Gregorich et al. 2005). For the non-livestock-based options, ranges for per-area mitigation potentials for each practice are given for each GHG (in t CO2-eq.ha−1 yr−1) for each of four climate regions in table 2. For soil carbon, estimates of soil C storage, CO2 mitigation potential and the low and high values for the 95% confidence interval were derived using mixed-effect modelling on a large dataset of long-term agricultural soil carbon experiments from a variety of countries, though temperate studies were more prevalent in the database (Ogle et al. 2005). Estimates were made using this method for all land-based mitigation options except estimates for soils under bioenergy crops and agroforestry which are assumed to derive their mitigation potential mainly from cessation of soil disturbance; the figures for soils under bioenergy crops and agroforestry are therefore assumed to be the same as for no till within the same climatic region, and for organic soil estimates which are derived using estimated emissions under drained conditions from International Panel on Climate Change (IPCC) guidelines (IPCC 1997, 2003). Soil methane and nitrous oxide emission reduction potentials were derived as follows: (i) for organic soils, the mean of low- and high-nutrient status organic soil N2O emission factors were used from the IPCC good practice guidelines for land use, land-use change and forestry (GPG LULUCF; IPCC 2003), where low and high values correspond to best estimates for low- and high-nutrient status organic soils and for CH4, low, high and median emission values are taken from Le Mer & Roger (2001), (ii) N2O figures for nutrient management were derived from US-EPA (2006) assuming a reduction in N to 80% current N application, (iii) N2O figures for tillage and residue management were derived from US-EPA (2006 using figures for no-till), and (iv) global rice figures were taken directly from US-EPA (2006) so per area figures are not given.
For the livestock-based options, mitigation potentials (dairy cows, beef cattle, sheep, dairy buffalo and other buffalo) for reducing enteric methane emissions through improved feeding practices, specific agents and dietary additives, and longer term structural and management changes/animal breeding are shown in table 3. These estimates were derived using a model similar to that described in US-EPA (2006). The proportional reduction due to the application of each practice was estimated from reports in the scientific literature (see footnotes to table 3 for main references). These were adjusted for (i) the proportion of the animal's life where the practice was applicable, (ii) the technical adoption feasibility in a region, i.e. whether the farmers have the necessary knowledge, equipment, extension services, etc., to apply the practice (average dairy cow milk production in each region over the period 2000–2004 was used as an index of the level of technical efficiency in the region and to score a region's technical adoption feasibility), (iii) the proportion of animals in a region to which the measure can be applied (i.e. if the measure is already being applied to some animals as in the case of bST use in North America, it is considered to be applicable only to the proportion of animals not currently receiving the product), and (iv) non-additivity of simultaneous application of multiple measures. There is evidence in the literature that some measures are not additive when applied simultaneously, such as the use of dietary oils and ionophores, but this is probably not the case with most measures. However, we did account for the fact that once one measure is applied, the emissions base for the second measure is reduced and so on, and we also incorporated a further 20% reduction in mitigation potential to account for unknown non-additivity effects. Only measures considered feasible for a region were applied in that region (e.g. bST was not considered for European regions due to the ban on its use in the EU). It was assumed that total production of milk or meat was not affected by application of the practices, so that if a measure increased animal productivity, animal numbers were reduced in order to keep production constant.
As can be seen from tables 2 and and3,3, some of the mitigation measures operate predominantly on one GHG (e.g. dietary management of ruminants to reduce CH4 emissions) while others have impacts on more than one (e.g. rice management). Some practices benefit more than one gas (e.g. set-aside/headland management) while others involve a trade-off between gases (e.g. restoration of organic soils). Table 2 also shows that the effectiveness of some mitigation practices differs between climate regions and can also differ within a climate region. A practice that is highly effective in reducing emissions at one site may be less effective or even counterproductive elsewhere. This means that there may be no universally applicable list of mitigation practices, but that any proposed practices will need to be tuned to individual agricultural systems present in specific climatic, edaphic and social settings.
The effectiveness of mitigation strategies also changes with time. Some practices, like those which elicit soil C gain, have diminishing effectiveness after several decades; others, such as methods that reduce energy use, may reduce emissions indefinitely. For example, Six et al. (2004) found a strong time dependency of emissions from no-till agriculture, in part owing to changing influence of tillage on N2O emissions.
The per-area/per-animal values for mitigation potential for each climate region, summarized in tables 2 and and3,3, were used to scale-up to regions and to the world by multiplying by the appropriate area under each climate in each region. The regions, climate zones within each region, areas of crop, crop mix and grassland in each climate zone in each region, area of cultivated organic soils within each climate zone in each region, the area of degraded land in each climate zone in each region and the total area of rice cultivation for each region were derived from the FAO Global Agro-Ecological Zones (AEZ; FAO/IIASA 2000), FAO Digital Soils Map of the World (FAO/UNESCO 2002) and FAO statistical (FAOSTAT 2006) databases as follows (figure 1):
All data were converted to real-area projections and the areas in square metre were converted to hectare. Cropland mitigation options were applied to the total crop area (minus those under rice cultivation, irrigation, set-aside or on organic soils or degraded soils, since other mitigation occurred on these lands), total mitigation was taken as the mean of the agronomy, nutrient management and tillage/residue management effects on 95% of the land, plus improved biosolid management on 5% of the land. Grazing land management was applied on all grassland, restoration of organic soils and degraded lands on the croplands occurring on these areas as calculated above, bioenergy on the land projected to be available for bioenergy production in 2030 by the Image v. 2.2 model (Strengers et al. 2004; Hoogwijk et al. 2005). Water management was applied only on the irrigable area identified in the FAO AEZ database, and agroforestry and set-aside only on projected surplus cropland in 2030. The total area of cropland and grassland for each region in 2030 for each SRES scenario is shown in table 4.
For emissions from livestock, total cattle, sheep and buffalo numbers in the various regions were obtained from FAOSTAT (2006). The cattle numbers for each region were broken down into numbers of dairy cattle and other cattle (owing to the different reduction potentials of both types) using FAOSTAT (2006). The biophysical emissions reduction potentials of the various practices were determined as described above. Estimated marginal costs of implementing each mitigation practice are shown in table 5.
In agriculture, there is a relationship between the amount paid for GHGs (i.e. the price of CO2 equivalents) and the level of mitigation realized. The amount of mitigation achieved for a given carbon price can be used to define a marginal abatement curve (MAC) for each practice for each region. We used the MACs from US-EPA (2006) to define the level of implementation (economic potential) for each practice in each region, for carbon prices up to 20, up to 50 and up to 100 US$ t CO2-eq.−1 practices as described below:
At low prices, the dominant strategies are those consistent with existing production such as change in tillage practice, fertilizer application, diet formulation and manure management, while higher prices elicit land use changes that displace existing production, such as biofuels (and afforestation), and allow the use of more costly animal feed-based mitigation options. The portfolio of mitigation strategies also varies over time owing to (i) the limited ecological capacity of the sequestration related strategies (i.e. their approach to a new carbon equilibrium over time) and (ii) the limited market penetration potential of capital intensive strategies like biofuels (which are constrained by the rate of turnover in energy processing plants, prospects and costs of retrofits, and energy product growth; Lee et al. 2005). It is important to note that while the most prevalent cost-mitigation quantity schedules are for single strategies (i.e. the amount of sequestration obtained as prices increase; as in Antle et al. 2001), it is not valid to sum these to gain a total mitigation potential due to resource competition among strategies. For example, Schneider & McCarl (2006) show that at higher prices, adding individual strategies can yield a total mitigation estimate that is as much as five times too large.
The global technical mitigation potential from agriculture by 2030, considering all gases, is estimated to be approximately 5500–6000Mt CO2-eq.yr−1, with cumulative economic potentials of 1500–1600, 2500–2700 and 4000–4300Mt CO2-eq.yr−1 at carbon prices of up to 20, up to 50 and up to 100 US$ t CO2-eq.−1 (table 6). To put these figures in context, annual CO2 emissions during the 1990s were approximately 29000Mt CO2-eq.yr−1, so agriculture could offset, at full biophysical potential, about 20% of total annual CO2 emissions, with offsets of approximately 5, 9 and 14% at CO2-eq. prices of up to 20, up to 50 and up to 100 US$ t CO2-eq.−1.
Of these total mitigation potentials, approximately 89% is from reduced soil emissions of CO2, approximately 9% from mitigation of methane and approximately 2% from mitigation of soil N2O emissions (figure 2). For each region, the biophysical potential is defined by the sum of the potential due to (i) improvements in cropland management (mean of cropland management, tillage practice, nutrient and manure management and water management) for the whole cropland area in 2030, (ii) improved grazing land management for the whole grassland area in 2030, (iii) reduction of soil GHG emissions under bioenergy cropping, (iv) improved rice management of the whole rice area, (v) restoration of native ecosystems on currently cultivated organic soils, (vi) restoration of all degraded lands, (vii) improved livestock management (mean of mitigation due to feeds/inocula/breeding and systems) and (viii) improved manure management. Figure 3 shows the total mitigation potential per region using the mean per-area estimates of potential for all practices and GHGs considered together.
The low, mean and high regional estimates of the biophysical mitigation potential are shown in figure 4. The low and high estimates about the mean (e.g. low and high estimates are approximately 400 and 10600Mt CO2-eq.yr−1, respectively about the mean estimate of 5500Mt CO2-eq.yr−1) are largely determined by uncertainty in the per-area estimate for the mitigation measure. For soil CO2 emission reduction, this arises from the mixed linear effects model used to derive the mitigation potentials, accounting for approximately 89% of the total potential. It is important to note that the most appropriate agricultural mitigation response will vary at the regional level and different portfolios of strategies will be developed in different regions and in countries within a region.
Estimates in the IPCC Second Assessment Report (SAR; IPCC 1996) suggested that 400–800Mt C yr−1 (equivalent to approximately 1400–2900Mt CO2-eq.yr−1) could be sequestered in global agricultural soils with a finite capacity saturating after 50–100 years. In addition, the SAR concluded that 300–1300Mt C (equivalent to approximately 1100–4800Mt CO2-eq.yr−1) from fossil fuels could be offset by using 10–15% of agricultural land to grow energy crops, with crop residues potentially contributing 100–200Mt C (equivalent to approximately 400–700Mt CO2-eq.yr−1) to fossil fuel offsets if recovered and burned. It was noted that burning residues for bioenergy might increase N2O emissions but this effect was not quantified. The SAR concluded that CH4 emissions from agriculture could be reduced by 15–56%, mainly through improved nutrition of ruminants and better management of paddy rice. It was also estimated that improvements in agricultural management could reduce N2O emissions by 9–26%. The SAR noted that GHG mitigation techniques will not be adopted by land managers unless they improve profitability, but that some measures are adopted for reasons other than for climate mitigation. Options that both reduce GHG emissions and increase productivity are more likely to be adopted than those which only reduce emissions.
In the IPCC Third Assessment Report (TAR; IPCC 2001), estimates of agricultural mitigation potential by 2020 were 350–750Mt C yr−1 (approximately 1300–2750Mt CO2 yr−1). It was noted that the range was mainly caused by large uncertainties about CH4, N2O and soil-related emissions of CO2 and that most reductions will cost between US$ 0 and 100tC-eq.−1 (approximately US$ 0–27t CO2-eq.−1) with limited opportunities for negative net direct cost options. The analysis of agriculture in the TAR included only conservation tillage, soil C sequestration, nitrogen fertilizer management, enteric methane reduction and rice paddy irrigation and fertilizers. The estimate for global mitigation potential was not broken down by region or practice.
These estimates, based on the best data currently available, are comparable with previous estimates, but give for the first time, an assessment of the agricultural mitigation potential for all gases, for all regions, at a range of potential carbon costs. The comparison of previous estimates of agricultural mitigation potential with comparable figures from this study is summarized in table 7. Given the differences in areas considered and the different assumptions made in previous studies, the estimates in this study are strikingly similar.
In addition to GHG emission reduction, agricultural land can provide feedstock for bioenergy production. Bioenergy to replace fossil fuels can be generated from agricultural feedstocks including by-products of agricultural production and dedicated energy crops. For residues from agriculture, the energy production and GHG mitigation potentials depend on yield/product ratios and the total agricultural land area, as well as type of production system. Less intensive management systems require reuse of residues for maintaining soil fertility. Intensively managed systems not only allow for higher usage rates of residues but also usually deploy crops with lower crop-to-residue ratios. Estimates of energy production potential from agricultural residues vary between 15 and 70EJyr−1. The latter figure is based on the regional production of food (in 2003) multiplied by harvesting or processing factors and the assumed recoverability factors. These figures do not subtract the potential alternative use for agricultural residues. As indicated by Junginger et al. (2001), competing applications can reduce the net availability of agricultural residues for energy or materials significantly. In addition, the expectations about future availability of residues from agriculture vary widely among the studies. Dried dung can also be used as an energy feedstock. The total estimated contribution could be 5–55EJyr−1 worldwide, with the range defined by current global use at the low end, to technical potential at the high end. Usage in the longer term is uncertain because dung is considered a ‘poor man's fuel’.
Organic wastes and residues together could supply 20–125EJyr−1 by 2050, with organic wastes potentially having an important role. The potential fossil fuel offset for 2050 from agricultural organic wastes and residues when used for energy production, assuming that it replaces gas, its energy content is 20GJ t −1 of dry biomass (IPCC 2001) and 1 t of dry biomass used to generate electricity prevents 0.28t C from gas from being emitted to the atmosphere (Cannell 2003), is 1000–6000 Mt CO2-eq.yr−1. If we assume linear uptake, a rough estimate of the potential by 2030 is 600–4000 Mt CO2-eq.yr−1.
The energy production and GHG mitigation potentials of dedicated energy crops depend on land availability, considering that food demand has to be met, combined with nature protection, sustainable management of soils and water reserves and other sustainability criteria. Since future biomass resource availability for energy and materials depends on these factors, an accurate estimate is difficult to obtain. Berndes et al. (2003) reviewed 17 studies of future biomass availability and showed that no complete integrated assessment and scenario studies were available.
Energy cropping on current agricultural land could, with projected technological progress, deliver over 800EJyr−1 without jeopardizing the world's food supply. Various studies have arrived at differing figures for the potential contribution of biomass to future global energy supplies ranging from below 100EJyr−1 to above 400EJyr−1 in 2050. A recent study (Sims et al. 2006), using lower per-area yield assumptions and bioenergy crop areas projected by the Image v. 2.2 model, suggests more modest potentials by 2025. The differences among studies are largely attributable to uncertainty in land availability and yield levels. The potential fossil fuel offset from dedicated energy crops by 2050, if assumed to supply 100–400EJyr−1 by replacing gas, and assuming 20GJ t−1 of dry biomass (IPCC 2001) and that 1 t of dry biomass used to generate electricity prevents 0.28tC from gas from being emitted to the atmosphere (Cannell 2003), is 5000–20000Mt CO2-eq.yr−1. If we assume linear uptake, a rough estimate of the potential by 2030 is 3000–12000Mt CO2-eq.yr−1.
Total GHG mitigation potential from agricultural bioenergy by 2030, including dedicated energy crops and agricultural wastes and residues is 4000–16000Mt CO2-eq.yr−1. The economic analysis presented above, using figures for bioenergy uptake from Lee et al. (2005), suggests that 4, 14 and 100% of the biophysical potential would be implemented at 0–20, 0–50, 0–100 US$ t CO2-eq., respectively. Assuming that 16000Mt CO2-eq.yr−1 represents the total biophysical potential, economic mitigation potential of biomass energy from agriculture at 0–20, 0–50, 0–100 US$ t CO2-eq. is estimated to be 640, 2240 and 16000Mt CO2-eq.yr−1 accounting for 30, 90–100 and 500% of all other agricultural GHG mitigation measures combined, respectively. The bioenergy mitigation potential is compared to other agricultural GHG mitigation options at a range of prices of CO2-eq. in figure 5.
Like mitigation from bioenergy, where the mitigation effect is usually counted in the user sector, enhanced energy efficiency (i.e. through reduced fossil fuel) is also possible in the agricultural sector. Figure 6 shows the potential for energy savings by 2030 in different world regions, derived by summing estimates from individual countries. These were calculated as emission savings which were calculated as follows:
Improved energy efficiency potentially delivers an additional global GHG mitigation potential of 770Mt CO2-eq.yr−1 by 2030.
Many agricultural mitigation activities show synergy with the goals of sustainability, and many explicitly influence the constituents of sustainable development, including social, economic and environmental indicators. Other mitigation options have more uncertain impact on sustainable development. There are interactions between mitigation and adaptation in the agricultural sector. Mitigation and adaptation may occur simultaneously, but differ in their spatial and geographical characteristics. The main climate change benefits of mitigation actions taken now will emerge only over decades, but where the drivers achieve other policy objectives (e.g. to meet air or water quality standards) there may also be short-term benefits. Conversely, actions to enhance adaptation to climate change impacts, even in the short term, will have consequences at all timescales from short- to long term (Smith et al. 2007).
In many regions, non-climate policies, including macroeconomic, agricultural and environmental policies, have greatest impact on agricultural mitigation options. These are reviewed elsewhere (Smith et al. 2007). Some evidence suggests that, despite significant biophysical potential for GHG mitigation in agriculture, very little progress has been made since 1990 and little is expected by 2010. There are barriers to implementation which may not be overcome without policy/economic incentives (Smith et al. 2005a).
Many agricultural mitigation options have both co-benefits (in terms of improved efficiency, reduced cost and environmental co-benefits) and trade-offs. Balancing the co-benefits with potential adverse effects is necessary for successful implementation. Many agricultural GHG mitigation options could be implemented immediately without further technological development, but a few options are still undergoing technological development. Technological development has been shown to be a key driver in ensuring the efficacy of agricultural mitigation measures (Smith et al. 2005b). Communication and capacity building is also important. In particular, it is important that farm managers understand the issue of climate change or potential opportunities so as to be motivated to act, the technologies and their application, and the costs and benefits of mitigation actions. The long-term outlook for GHG mitigation in agriculture suggests that there is significant potential, but many uncertainties, both price- and non-price related, will determine the level of implementation. These further considerations are discussed in detail elsewhere (Smith et al. 2007).
There are significant opportunities for GHG mitigation in agriculture, but for the potential to be realized numerous barriers need to be overcome. Many recent studies have shown that actual levels of GHG mitigation are far below the technical potential for these measures (e.g. Smith et al. 2005a). The gap between technical potential and realized GHG mitigation occurs due to barriers to implementation, including climate and non-climate policy, and institutional, social, educational and economic constraints. The mix of agricultural mitigation options that are adopted in the future will also depend upon the price of carbon dioxide equivalents. The total biophysical potential of approximately 5500–6000Mt CO2-eq.yr−1 would never be realized due to these constraints, but with appropriate policies, education and incentives, it may be possible for agriculture to make a significant contribution to climate mitigation by 2030. To put the figures calculated here in context, annual CO2 emissions during the 1990s were approximately 29000Mt CO2-eq.yr−1, so agriculture could offset, at full biophysical potential, about 20% of total annual CO2 emissions, with offsets of approximately 5, 9 and 14% at CO2-eq. prices of up to 20, up to 50 and up to 100 US$ t CO2-eq.−1.
One contribution of 15 to a Theme Issue ‘Sustainable agriculture II’.