Croplands play a key role for the sustainability of world's life. The exponential growth of global population and the need to produce alternative energy to fossil fuel has placed increased demands on agriculture to produce more food and biomass. For these reasons the agriculture poses a huge challenge for the sustainability at global and local environments, because of the different forms of environmental impacts related to agricultural management. The loss of soil organic matter (SOM) through enhanced mineralization and the emissions toward the atmosphere of greenhouse gases (GHGs), two processes strictly connected, are numbered among the environmental impacts of agriculture.
Net primary production (NPP) of the global croplands has been estimated at 15% of global terrestrial NPP (Field et al. 1998). Today, the increasing productivity is allowed by large use of fossil fuels and technologies, giving life to the so-called industrialized agro-ecosystems. These are wide extended, particularly in the developed countries. Cultivated lands cover 14% of the world's vegetated land surface while, across the European Union, agricultural surfaces cover an area of 46%, with 24% of arable and 19% of grasslands (Ramankutty and Foley 1999). The wide input of subsidiary fossil energy toward the agricultural systems improves the primary productivity, but mobilizing matter from other systems and speeding up the cycle matter inside the system, determines pollution and environmental degradation. In fact, one of the main effects of industrialized agro-ecosystem is the alteration of biogeochemical cycles, mainly carbon and nitrogen cycle. Particularly, the agricultural activities contribute considerably to the emissions toward the atmosphere of GHGs, i.e., carbon dioxide (CO2), methane (CH4) and nitrous oxide (N2O).
The main concern of the altered global C cycle is the large imbalance between carbon release to the atmosphere and carbon uptake by other compartments, that leads to a continued increase in atmospheric CO2 to a rate of 4.1 x 109 tons of carbon per year (IPCC 2007a, b). CO2 is considered the main GHG, affecting the phenomenon for more than 50% (IPCC 1996). The terrestrial carbon cycle is supposedly a sink of about 25% of the anthropogenic CO2 emissions (Running 2008). On the other end, over the past few 100 years, the expansion of agricultural land has released substantial carbon to the atmosphere, due to soil carbon depletion by agriculture through removal of photosynthate carbon toward the market system and conventional tillage (CT) practices, which increase SOM mineralization rate (Schlesinger 1984).
Soil CO2 emission is, after photosynthesis, the second largest C flux in terrestrial ecosystems, accounting for 60-90% of total respiration (Longdoz et al. 2000). Global soils, including crop soils, store great amounts of organic carbon (OC) and they are, potentially, one of the sinks of atmospheric CO2 fixed by photosynthesis and released in soil as dead material and/or plant exudates. Global SOM has a basic role in carbon storage as well as in control of biological processes involved in soil-atmosphere exchange of CO2, CH4 and N2O. Despite its long residence times when in steady state, this very large C reservoir, which far exceeds C pools in both aboveground biomass and atmospheric CO2 (Eswaran et al. 1993), represents a very large potential source of CO2, if decomposition exceeds stabilization, i.e., humification.
Globally, plant biomass and SOM store about 500 Pg and 1,100 Pg C, respectively (IPCC 1996). Agricultural soils account for less than one-fourth of the SOC pool (Wood et al. 2000), and SOC levels are usually related to climate, topography and soil texture. Soils in North America, Asia and Europe are considerably richer in SOC (12.2, 12.6, and 14.6 kg C m-2, respectively) than in sub-Saharan Africa (7.7 kg C m-2). However, the effects of human activity on global carbon stocks are inadequately understood, particularly for human-managed terrestrial ecosystems. One of the greatest uncertainties concerns changes in soil C stocks that may occur with different management.
Human activities have also dramatically altered the earth's nitrogen (N) cycle, strictly linked with the C cycle, particularly in the agro-ecosystems, since N immobilized in plant tissues is removed with harvest, thereby reducing the N availability in soil for the next crop cycle. Loss of N soil fertility is recovered by means of mineral fertilizers, whose production needs large amount of fossil fuels (Haber's process). Input of reactive N into the biosphere by man now exceeds the rate of biological N2-fixation in native terrestrial ecosystems (Galloway et al. 2004). This increased reactive N is due not only by N fertilizer production, but also by the fossil fuel combustion used to support food and energy demands. The large input of added nitrogen and SOM management is responsible of a large diffusion of inorganic N in the environment with several impacts. A main impact is represented by the production of N2O, this gas accounts for about 6% of the anthropogenic greenhouse effect (IPCC 1996), and agricultural soils are the major source of atmospheric N2O. This gas is characterized by a warming potential 200 times as large as CO2 and, by reacting with oxygen radicals in the stratosphere to form nitrogen monoxide, is involved in the destruction of stratospheric ozone (Crutzen 1981).
Thus, it is raising a new concern of industrialized agriculture toward a more environmentally sustainable system. Literature in sustainable agriculture identifies two core aspects (1) the over-exploitation of natural resources and (2) the induced pollution to the environment. Within this frame, agricultural soils play key roles, since they are natural resource with long regeneration cycle (centuries or millennia), and source of atmospheric degradation by emitted GHGs. In addition to developing new soil management strategies for sustainable agro-ecosystems, political and social approaches are also needed, based on a common understanding that soil and agro-ecosystems are essential for a sustainable society. For these reasons the sustainable management of soil received a strong support at the Rio Summit in 1992, as well as in Agenda 21 (UNCED 1992), the UN Framework Convention on Climate Change (UNFCCC 1992), Articles 3.3 and 3.4 of the Kyoto Protocol (UNFCCC 1998) and elsewhere. These conventions indicate the recognition by the world community of the strong link between soil degradation and desertification on one hand, and loss of biodiversity, threat to food security, increase in poverty, and risk of accelerated greenhouse effect and climate change, on the other hand. Article 2 of the Kyoto Protocol states that nations adhering to the Protocol have to improve energy use in different sectors of the national economy, including agriculture, and promote policies for sustainable agriculture in order to reduce the impact on climate change. Moreover, article 10 invites nations to draft national or regional programmes to reduce GHGs emissions. In the same article, agriculture is considered as an economic sector liable to such planning. The conference of parties (COP) at Bonn and Marrakech (2001) have included evaluation of carbon sink in forest (article 3.3) and in agriculture (article 3.4), as a result of land use, land use change, and forestry (LULUCF). Improving the capacities of land use and agricultural practices to increase the carbon stocks in soils is one of the policies to be developed to promote SOM accumulation.
Following the results of the European Climate Change Programme (ECCP), a good progress is a policy that allows C sequestration from 5 to 8% of the CO2 emitted by European activities. This topic is strictly associated to other problems of SOM management in southern part of Europe (Spain, Portugal, Italy, and Greece), where desertification is progressively advancing. SOM is often considered a key factor for either soil degradation or soil rehabilitation, and 2% has been suggested as the soil threshold content beyond which degradation occurs. According to UNEP (1991), desertification threatens over 60% of Southern European landscapes, and represents one of the largest environmental threats in the European Union.
SOM decline in cultivated soils has been studied in many long-term experiments (Grace and Oades 1994; Golchin et al. 1995). The restoration of organic matter in cropland is limited, while soil respiration (i.e., CO2 release) tends to increase, thus resulting in a considerable carbon loss as compared with natural terrestrial ecosystems (Buyanovsky et al. 1987). Continuous cropping and inadequate replacement of nutrients, removed by harvested crops or lost through erosion, leaching, or gaseous emissions, deplete fertility and cause SOM levels to decline, often by 50% or more (Matson et al. 1997). When soils are tilled, SOM is more prone to weathering and decomposes fast because favourable conditions in water, aeration and temperature are available for microbial activity. The amount of organic matter lost by either clearing a wooded area or tilling native grassland varies according to soil type, but most of organic matter is lost within the first 10 years (Buyanovsky et al. 1987). Evidence acquired over several years increasingly indicate that certain fractions of SOC are likely to respond more rapidly than total soil C to land use change and management. SOM is divided into labile and non-labile materials. It has been shown that the C and N present in particulate organic matter (POM) can accumulate rapidly under land management systems that minimize soil disturbance and may also provide an early indicator of change in C dynamics and total soil C under different land use and management practices (Cambardella and Elliott 1992; Franzluebbers and Stuedemann 2002). The loss of SOM under cultivation can mainly be attributed to loss of the labile C fraction (Wadman and de Haan 1997).
There is a close relation between SOM degradation and global warming that, in turn, may strongly affect SOM decomposition. Rising temperatures brought about by climate change will cause microorganisms in world's soils to decompose organic matter more rapidly, releasing extra CO2 and accelerating climate change
(Knorr et al. 2005). Over the short term, CO2 increase in atmosphere may enhance plant growth through CO2 fertilization, thus removing some of the excess CO2 (Giardina and Ryan 2000). However, current models predict that, in longer term, rising temperatures will speed up decomposition of SOM and release CO2 in atmosphere to an extent that exceeds any carbon sequestration in soil, and contributes further to climate change (Knorr et al. 2005). Some reports suggest that non-labile SOM is more sensitive to temperature than labile SOM, thus implying that the long-term effects of soil decomposition in a warming world may be even stronger than that predicted by global models (Davidson and Jansens 2006).
Despite the evidence that soil agricultural management is responsible of SOM degradation, the belief that agricultural ecosystems can play an important role in absorbing the surplus of carbon produced by human activities has been recently spread out. Such carbon sequestration would imply the transfer and secure storage of atmospheric CO2 into persistent C pools, thereby preventing its immediate reemission. Hence, a first step of this process is represented by an increased plant photosynthetic CO2 fixation that is stimulated by the larger CO2 concentration in atmosphere (Norby et al. 1992; Paustian et al. 1997). However, the photosynthate carbon should not be totally fixed into stable humus fractions, and variable amounts may be in large part distributed in rapidly cycling plant and soil carbon pools (Hungate et al. 1997; Schlesinger and Licther 2001).
Nevertheless, soil management strategies in croplands can have a great potential for carbon sequestration, since the carbon sink capacity for world's agricultural and degraded soil is still 50-66% of the historic carbon loss of 42-72 Pg (Lal 2004). However, the actual potential carbon storage in cultivated soils may be even smaller if climate change leads to increases in mineralization. Different literature showed that croplands have a great capacity to absorb carbon produced by fossil fuels burning (Schütz et al. 1990; Franzluebbers 2005; Johnson et al. 2005; Su 2007). Lal (2004) estimated that cultivated soil can accumulate 0.4-0.8 Pg C year-1, if recommended farming practices are adopted, such as no-till, crop rotation, cover cropping and manure application. Lal (2001) reported that pasture treatment increased SOC from 9.2 to 55.4 Mg ha-1 after 25 years, and forest management increased SOC from 14 to 48.4 Mg ha-1 after 21 years. This suggests that grassland and pasture treatments would increase SOC stock as much as forest management. SOC sequestration might be increased at the expense of an increase of non-CO2 GHGs emissions (CH4 and N2O), although soil-specific strategic practices, such as synchronized fertilization techniques and optimum water control, among other things, may reduce these emissions.
CO2 from soil is not entirely produced by SOM degradation. In fact, vegetation contributes to total CO2 emission with root and rhizo-microbial respiration. For this reason, the total soil CO2 emission cannot be considered a direct measure of SOM
oxidation, despite some studies continue to interpret it in such a manner (Hanson et al. 2000; references therein). Since plants' contribution does not involve stable organic matter, total soil CO2 flux has no effect on the long-term C balance in soils. Kuzyakov (2006) describes three C pools as sources of CO2 from soil (1) SOM; (2) above and below ground dead plant residues; (3) organic substances released by living roots such as exudates, secretions and sloughed-off root cells. The last group is frequently described as rhizo-deposits.
The soil carbon pools are oxidized by different groups of heterotrophic organisms for their metabolic needs. The most important and active heterotrophs in soil are microorganisms: bacteria, fungi, actinomycetes and protozoans. The contribution of soil fauna (micro-meso and macro fauna) to total CO2 emission from soils usually consists of only a few percent (Konate et al. 2003; Ke et al. 2005). Despite this negligible direct contribution, soil fauna may greatly affect microbial respiration by fragmentation of plant residues (Couteaux et al. 1991; Bonkowski et al. 2000) and by its different roles in the soil trophic webs (Bonkowski 2004). This intensifies their turnover rate and results in increasing CO2 emissions from soil (Mikola and Setala 1998; Wardle et al. 1998).
Contribution of plants' root respiration is most important and highly variable in space (soil characteristics, soil volume containing roots) and time (depending on environmental factors affecting plant activities). In particular, the contribution of vegetation to soil CO2 fluxes is not constant and changes during the year. It depends on plant species (Fu et al. 2002), growth stage (Rochette et al. 1999), soil nutrients status (Bradley andFyles 1996) and environmental factors, such as intensity of light for photosynthesis (Craine et al. 1999; Kuzyakov and Cheng 2001), soil moisture (Flanagan et al. 2002) and temperature (Buchmann 2000). Therefore, it is fundamental to distinguish the microbial decomposition of SOM in roots-free soil, frequently referred to as "basal respiration." Exudates, secretions and sloughed-off root cells from plants are liable to change microbial activity in the so-called "rhizosphere priming effect" (Kuzyakov 2002), and represent the interaction between growing roots and SOM decomposition (Cheng and Kuzyakov 2005). Similar changes in SOM decomposition have often been measured after addition of fresh plant residues to soil (Kuzyakov et al. 2000). The microbial capability in decomposing SOM depends on environmental factors (moisture, temperature, soil nutritional status, etc.) and SOM quality. Typically, SOM is divided into at least two and frequently more pools: inert (or passive), recalcitrant, resistant, decomposable, available, active, etc., each being characterized by different residence times. It is generally accepted that the inert pool is complex and tightly bound to clay minerals. This SOM pool has a very slow degradation rate with a mean residence time of thousands of years (Rethemeyer et al. 2004). This inert pool gives only a minor contribution to CO2 fluxes from soil. The other SOM pools (decomposable, available, active) can be oxidized by microorganisms either through basal respiration or by priming effect.
With regard to the CO2-driven greenhouse effect, only SOM-derived CO2 contributes to change atmospheric CO2 concentration. Due to fast turnover times, microbial decomposition of plant residues and rhizodeposits, as well as root respiration, have no significant effect on C sequestration in the short- or long-term. Since plant C sources frequently amount to more than half of the total soil CO2 flux (Hanson et al. 2000), the flux of plant-derived CO2 masks the contribution of SOM-derived CO2, when measuring CO2 fluxes from planted soils.
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