aRe-calculated in Tg N yr \ anthropogenic total including global warming source of 0.2 Tg N yr-1. Natural sources from Bouwman et al. (1995); stratospheric sink and atmospheric increase from Houghton et al. (1995). bVan Aardenne et al. (2001). cBouwman et al. (2002a) for the 1990s.
dKroeze et al. (2005), Nevison et al. (2004); estimated uncertainty is ±70% from Nevison et al. (2004).
eNevison et al. (2003 , 2004), combining the uncertainties in ocean production and oceanic exchange.
fMosier et al. (1998), Kroeze et al. (1999), Olivier et al. (1998): a single value indicates agreement between the sources and methodologies of the different studies.
Hirsch et al. (2006) inversion results further suggest that N2O source estimates from agriculture and fertilizer may have increased markedly over the last three decades when compared with an earlier inverse model estimate (Prinn et al. 1990). Applications of N fertilizer will further increase in order to feed some 8 billion people over the next 30 years (Zhu et al. 2005). Therefore, agricultural N2O emissions are projected to increase by 35-60% up to 2030 due to increased nitrogen fertilizer use and increased animal manure production (FAO 2003). Similarly, US-EPA (2006) estimates that N2O emissions will increase by about 50% by 2020 relative to 1990.
220.127.116.11 Processes Controlling Production and Approaches to Reduce Uncertainty
Crop production, wherein anthropogenic reactive nitrogen is created and consumed, inevitably causes N2O emissions directly and indirectly (Mosier et al. 1998). For example, up to 71% of the annual total N2O released from anthropogenic reactive nitrogen is presently due to direct emission in Asia (Zhu et al. 2005). N2O is often enhanced where available nitrogen (N) exceeds plant requirements, especially under wet conditions (Smith and Conen 2004; Oenema et al. 2005).
N2O is produced during N transformation in soils, mainly by the microbial processes of nitrification and denitrification (Freney 1997) as well as chemo-denitrification (van Cleemput 1998) and fungal transformations. These microorganisms operate under various optimum conditions, and it is generally assumed that nitrification is an aerobic process, and denitrification is an anaerobic process (Granli and B0ckman 1994; Conrad 1996; Bouwman et al. 2002b). Techniques, such as selective nitrification inhibitor and 15N labeling (Muiller et al. 2004) have been used to identify the N2O production processes. Another way to identify the processes producing N2O is to monitor the changing isotopic composition of N2O based on the findings that the stable isotope ratio of 15N/14N of denitrifier-derived N2O can differ from that of nitrifier-derived N2O (Yoshida 1988; Kim and Craig 1993; Webster and Hopkins 1996).
The microbial basis of N2O production in soil is more complicated than that of CH4. Production of N2O has been demonstrated among bacteria that respire nitrate to nitrite and those that dissimilate nitrate to ammonium (DNRA) (Stevens et al. 1998; Xiong et al. 2007). Since denitrifiers, nitrate respirers, and DNRA cannot be differentiated in situ, and also autotrophic and heterotrophic nitri-fiers are not differentiable, the exact microbial basis of N2O production and consumption is still not clear (Conrad 1996). Moreover, these physiologically defined groups are obviously widespread among the various taxa, bacteria, and fungi.
Though N2O is produced and consumed by defined reactions in individual microorganisms, much of the spatial and temporal variability of N2O fluxes between the soil and the atmosphere can often be explained by only a limited number of factors (Conrad 1996; Bouwman et al. 2002b). Any factor that affects the equilibrium of N transformation in soil, could result in N2O emissions. Nitrification is the predominant N2O-producing process under moderately moist and warm conditions, while denitrification is the predominant process under wet (anaerobic) conditions when NH4+ and NO3- are available in soil. Both empirical models based on various physical and chemical parameters and mechanistic models that differentiate between nitrification and denitrification are successful without considering the structure of the microbial community (Li et al. 1992a, b, 2004; Bouwman et al. 2002a). General denitrification models have shown that N2O production in soil is mainly controlled by the availability of nitrate, labile C compounds, and O2 (Del Grosso et al. 2000). Soil nitrate content and soil water content are the key factors affecting denitrifica-tion and N2O emissions from the agricultural soils (Vallejo et al. 2001; Dobbie and Smith 2003).
The potential source of N2O from crop fields is the volatilization of ammonia, some of which is converted to N2O in the atmosphere (Dentener and Crutzen 1994; Warneck 2000). The N2O channel is 10-60% efficient according to a review by Atkinson et al. (2004). However, there is no laboratory or field observation of this mechanism till now.
Another new source of N2O is emission from plants (Chen et al. 1997; Goshima et al. 1999; Zhang et al. 2000; Xu et al. 2001; Smart and Bloom 2001; Hakata et al. 2003; Zou et al. 2005). Many researches demonstrated the role of growing plants in N2O production and emissions from agricultural systems (Chang et al. 1998; Muiller 2003). Some studies showed a role of plant pathway in ecosystem N2O emissions (Mosier et al. 1990; Yu et al. 1997; Chang et al. 1998; Yan et al. 2000). This source remains yet unquantified.
Croplands are usually fertilized and therefore not likely to be sinks for N2O, though some studies report N2O uptake in fertilized fields (Xu et al. 1997; Kroeze et al. 2007). The crops usually don't efficiently utilize N fertilizers and other N sources (Zhu and Chen 2002; Galloway et al. 2004; Xiong et al. 2008). The surplus N is particularly susceptible to emission as N2O (Velthof et al. 1996; McSwiney and Robertson 2005; Xiong et al. 2006a).
N2O emission from croplands at site scales occurs essentially with great spatial and temporal variability (Veldkamp and Keller 1997; Dobbie and Smith 2003). Spatial and temporal variability is mainly caused by heterogeneity in soil properties and agricultural management (e.g. water, nutrient, crop, tillage, and residue management) (Veldkamp and Keller 1997; Brown et al. 2002; Dobbie and Smith 2003). Agronomic practices, such as tillage and fertilizer applications, can significantly affect the production and consumption of N2O because of alteration in soil physical, chemical, and biochemical activities. Tillage could cause immediate changes in microbial community structure as reported by Jackson et al. (2003). They produce large N2O emissions at the beginning of crop season (Xiong et al. 2006a). It's likely that some emission pulses would have been missed due to the insufficient sampling frequency since N2O emissions mainly occur in pulses (Zheng et al. 2000; Xing et al. 2002; Zheng et al. 2004; Xiong et al. 2002a, b, 2006a).
Improving N use efficiency can reduce N2O emissions (Cole et al. 1997; Paus-tian et al. 1998; Robertson et al. 2000; Dalal et al. 2003; Monteny et al. 2006). By reducing leaching and volatile losses, improved efficiency of N use can also reduce off-site N2O emissions (i.e. indirect emissions) (Xiong et al. 2006b). Adopting reduced or no-till may affect N2O emissions, but the net effects are inconsistent and not well-quantified globally (Smith and Conen 2004; Li et al. 2005). The effect of reduced tillage on N2O emissions may depend on soil and climatic conditions (Helgason et al. 2005). In some areas, reduced tillage promotes N2O emissions, while elsewhere, it may reduce emissions or have no measurable influence. Drainage of croplands in humid regions can 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). The use of rotations with legume crops is an important example to reduce reliance on external N inputs, although legume-derived N can also be a source of N2O (Xiong et al. 2002a, b; Rochette and Janzen 2005).
The fluxes of N2O from paddy soil are small compared to aerated soils, because a larger percentage of the produced N2O is further reduced to N2 by denitrifiers. Submerged rice fields seem to act occasionally even as a sink for atmospheric N2O (Minami andFukushi 1984; Xu et al. 1997). The direct emissions of N2O from rice agriculture are variable and robust global estimates are not yet possible. Since seasonal emission varied from ~ 1 to 150 mg m-2 yr-1, it accounts for a range of potential global emissions from 0.01 to 2Tgyr-1 (Xu et al. 1997; Khalil et al. 1998b; Suratno et al. 1998; Abao et al. 2000).
Several agricultural practices and conditions, that favor reduced emissions of CH4 from rice fields, tend to increase N2O emissions (Lindau et al. 1990; Bronson et al. 1997b; Xu et al. 1997; Cai et al. 1997; Abao et al. 2000; Hou et al. 2000; Xiong et al. 2007). However, the conditions, that affect emissions of these two gases, vary widely. For example, alternate anaerobic and aerobic cycling increases N2O emissions relative to constant aerobic or anaerobic conditions (Granli and B0ckman 1994; Chen et al. 1997; Tsuruta et al. 1997; Zheng et al. 1997; Xing 1998; Xing et al. 2002). In contrast, intermittent irrigation of rice paddies, which causes anaerobic and aerobic cycling, is considered to be one of the options for reducing CH4 emissions (Yagi et al. 1996; Chen et al. 1997; Zheng et al. 1997; Cai et al. 2003; Huang et al. 2004). Antecedent water regime of a soil, independent of the current water status, also affects N2O emissions (Xing et al. 2002) and CH4 emissions (Xu et al. 2000, 2003). After the rice is harvested and the fields are no longer inundated, varying amounts of N2O emissions are observed, but no CH4 emissions occur (Bronson et al. 1997a, b; Xu et al. 1997; Khalil et al. 1998b). In China, 86% of the rice fields are under a rice-upland cropping system and mid-season drainage is widely adopted in rice cultivation.
Mitigation is one aspect of the relationship between CH4 and N2O emissions from rice fields that has been widely studied (Bronson et al. 1997a; Wassmann et al. 2000; Breiling et al. 2005). The question of whether the agricultural practice of intermittent flooding of rice fields causes a net reduction of non- CO2 greenhouse gas emissions has been addressed, and it is argued that even though N2O has a global warming potential some 10 times greater than CH4, the radiative forcing of the additional N2O is only partially off-set by the reduction of CH4 emissions (Bronson et al. 1997a; Breiling et al. 2005). This finding is used to promote the practice of intermittent flooding as a means for mitigating greenhouse gas emissions from rice fields.
6.1.3 Carbon Dioxide (CO2)
CO2 is the most important anthropogenic greenhouse gas. The global atmospheric concentration of CO2 has increased from a pre-industrial value of about 280 to 379 ppm in 2005 (IPCC 2007). CO2 emissions from the agricultural soils are usually included in the land use, land use change and forestry sector. Land use change is the main mechanism by which agriculture can indirectly contribute to the increase of CO2. So there are a few comparable estimates of emissions of this gas from agriculture. CO2 is released largely from microbial decay or burning of plant litter and soil organic matter (Janzen 2004). Agricultural lands generate very large CO2 fluxes both to and from the atmosphere, but the net flux is small or zero depending on the conditions (Paustian et al. 1998). Another source of CO2, that can be attributed to the agriculture, is the energy used to grow and harvest the food produced, but there are no global estimates. This source, however, is included in the estimates of fossil fuel source of CO2.
Soil carbon sequestration has strong synergies with sustainable agriculture and generally reduces vulnerability to climate change. Stored soil carbon may be vulnerable to loss through both land management change and climate change. The conversion of land from forested to agricultural land can have a wide range of effects as far as CO2 emissions are concerned (Hymus and Valentini 2007). Soil disturbance and increased rates of decomposition in converted soils can both lead to emissions of CO2 to the atmosphere, and increased soil erosion and leaching of soil nutrients further reduce the potential of the area to act as a sink for atmospheric carbon. Similarly, land reclamation and changes in land use management can affect an increase in terrestrial carbon uptake. Current estimates suggest that such land-use changes lead to the emission of 1.7PgCyr-1 in the tropics, mainly as a result of deforestation, and to a small amount, of uptake (~0.1 Pg C) in temperate and boreal areas, thus producing a net source of ~1.6PgCyr-1.
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