The role of plants and land management in sequestering soil carbon in temperate arable and grassland ecosystems

Geoderma 128 (2005) 130 – 154 www.elsevier.com/locate/geoderma

The role of plants and land management in sequestering soil carbon in temperate arable and grassland ecosystems R.M. Reesa,*, I.J. Binghamb, J.A. Baddeleyb, C.A. Watsonb a

Crop and Soil Research, SAC Edinburgh, Penicuick EH26 OPH, UK b SAC Aberdeen, UK Available online 7 January 2005

Abstract Global climate change and concerns about soil quality have led to a widespread interest in the opportunities that are available to sequester carbon in soils. To achieve a better understanding of the changes in C storage, we need to be able to accurately measure and model inputs and losses of C from soils. This in turn requires a thorough understanding of the biological processes involved and the way in which they are influenced by the soil’s physical and chemical environment. The amount of C present in a soil is determined by the difference between C addition and C loss. Because these fluxes are large relative to changes in C storage, net storage can be very difficult to measure, particularly in the short term. Carbon is added to soil from plant and animal materials deposited on the soil surface. It is known that approximately 50% of C assimilated by young plants can be transferred below ground; some is used for root construction and maintenance as well as root respiration; some organic C is lost to the soil through exudation and root turnover. A comparison of eight studies has shown that the input to the soil of root derived organic C during a growing season can range between 0.1 and 2.8 t C ha 1. Quantifying inputs from different processes has proved difficult and the relative importance of exudation and root death under field conditions remains uncertain. The chemical composition of substrates released by exudation and root death is known to be very different. Exudates contain high concentrations of soluble organic substrates and as a consequence are highly labile, whereas additions of C from root death have structural organic substrates with lower potential decomposition rates. Losses of C from soil occur as a consequence of plant and microbial respiration. However, identifying the source of evolved CO2, whether it be from root or microbial respiration, is much more difficult. Some new methods using isotopic labelling and pool dilution have been developed to separate plant and microbial respiration, and despite difficulties, these promise to provide valuable information on the processes of C input and loss from soils. At a field scale measurements and models would suggest that soil and crop management can play a significant role in determining the extent of C sequestration by soils and the proportion of labile C present. A comparison of 11 field studies showed that soil respiration varies between 4 and 26 t C ha 1 year 1, with management such as tillage, drainage, grazing and manure application exerting a strong influence on the magnitude of fluxes. Net ecosystem exchange of C has been shown to be at least an order of magnitude lower than respiratory losses in comparable studies, but land management is important in determining the direction and magnitude of the C flux. Recent studies have suggested that although the overall quantity of C stored in European soils is increasing, this increase is confined largely to forested areas and that many cropped * Corresponding author. E-mail address: [email protected] (R.M. Rees). 0016-7061/$ - see front matter D 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.geoderma.2004.12.020

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soils are losing soil organic matter. It is has been suggested that that the biological potential for C storage in European cropland lies between 9 and 120 Mt C year 1. In order to take advantage of this potential and to develop management systems that promote C storage we need to achieve a better understanding of the processes of C input and loss, and develop improved models using pools that coincide with measurable soil C fractions. D 2004 Elsevier B.V. All rights reserved. Keywords: Carbon sequestration; Rhizodeposition; Roots; Soil respiration; Land management; Soil organic matter; CO2; Rhizosphere; Soil management

1. Introduction There is an urgent requirement to improve our understanding of the processes contributing to C storage in soils. This has arisen because of the need to sequester C to overcome global climate change (Paustian et al., 1998a) and to improve soil quality, as we develop more sustainable and land management practices (Carter, 2002). Protecting and enhancing stocks of soil organic matter will remain a global challenge that needs to involve environmental scientists, socio-economists and policy makers for the next decade and beyond. This exercise will require an improved understanding of the mechanisms underlying C sequestration and involve a reevaluation of modern and traditional farming practices, as the principles of environmental protection and sustainable management of non renewable resources gain a higher priority. Reflecting these concerns, the European Union’s common agricultural policy will place increased importance on soil quality and land management, as financial support is shifted from production towards environmental protection (European Commission, 2002). A range of new techniques has become available in recent years that have advanced our understanding of the processes underlying organic matter transformations in soil. Through the use of isotopic tracers we now have estimates of the quantities of plant derived C that are added to soil. Both continuous labelling and pulse labelling techniques have indicated that the soil environment is a major sink for C derived from exudation and root turnover (Meharg, 1994). This C contributes to root construction and maintenance, but is also added to the soil through plant respiration and rhizodeposition. The addition of C to soil stimulates microbial activity, which is the driving force behind trans-

formations of organic matter. Molecular techniques are allowing us to examine the relationship between microbial activity, biodiversity and function in soils. It has become clear that the diversity of microbial communities is greater than had been previously thought (Tiedje et al., 2001), however, such diversity may not always be a pre-requisite for soils to undertake a wide range of functions. Many soil functions have been shown to be unaffected by experimentally manipulating the biodiversity of microbial communities (Griffiths et al., 2001), indicating that there is widespread functional redundancy within the system. Where soil functions can be linked to specific biological populations an opportunity exists to manipulate soil processes through alterations of the soil microbial community (Atkinson and Watson, 2000). Experimental techniques allowing physical and chemical characterisation of soils are becoming increasingly sophisticated and are allowing us to explore the relationships between living organisms and the environment within which they function. Young et al. (1998) emphasise the importance of the soil’s structure in maintaining biological diversity through the provision of habitat and by influencing patterns of resource storage and transport. Modern tools in analytical chemistry are forcing us to re-examine long held theories of soil organic matter formation. For example Hatcher (2003) points to evidence from recent NMR studies that suggest that soil organic matter forms as a consequence of the transformation of biopolymers by microbial action and not as a result of random condensation reactions. The physical fractionation of soils is also helping us to understand how organic matter pools can become chemically and physically isolated. Hassink et al. (1997) have shown that the amount of organic matter that a soil can store is largely regulated by its silt and clay content although

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management can also influence storage, particularly of larger macro-organic matter fractions (Balesdent et al., 2000; Denef et al., 2001). Spatial variation of soil organic matter content particularly at national and regional scales are also strongly influenced by climate and land use (Joos et al., 2001; Paustian et al., 1998b). The atmospheric concentration of CO2 is currently higher that it has been for at least 420 000 years, and may well double in this century. This is leading to a very complex series of interactions between the biosphere, hydrosphere and atmosphere as the equilibration between soils, plants and the atmosphere breaks down. Human activity has had a major impact on the C cycle. Pre-industrial concentrations of CO2 were below 300 AL L 1, but today’s atmospheric CO2 concentrations are around 370 AL L 1 and rising. Measurements of the isotopic signature of CO2 in the atmosphere have established that fossil fuel consumption is principally responsible for the concentration rises (Quay et al., 2003). Land use change also contributes to an increase in atmospheric CO2 as a consequence of deforestation and cultivation of new arable land (Schimel et al., 2001). Most recent estimates suggest that during the period between 1989 and 1998 land use changes resulted in the addition of 1.6 Pg C per year to the atmosphere (Houghton, 2000). However, not all of the anthropogenically produced CO2 remains each year in the atmosphere. There is a significant removal of atmospheric C by the oceans (2.3 Pg C year 1) and an increasing sink within the terrestrial biosphere (Bremer et al., 1998). Studies of the C cycle during the 1990s suggest that as a consequence of both fossil fuel consumption and land use change the atmospheric pool of C is currently increasing by around 3.2 Pg per year (Frank et al., 2002). The regional distribution of sink strength in the terrestrial biosphere is more difficult to determine but it is suggested that the mid to high latitudes of the northern hemisphere are particularly important (Janssens et al., 2003). The Kyoto protocol commits signatories to reducing C emissions to 0.3 Pg below 1990 levels. Although this could be achieved partly by reduced dependence on reserves of fossil fuels, it is likely that land use change will play a vital role in contributing to atmospheric CO2 removal. It has been estimated that agricultural soils offer the potential to sequester between 0.4 and

0.9 Pg C year 1, through improved management of existing agricultural soils, restoration of degraded land, the more extensive use of setaside, and the restoration of wetlands (Paustian et al., 1998a). These opportunities to sequester C will probably only be available until the middle of the 21st century after which it is anticipated that soils may return to being net source, as a consequence of respiratory losses increasing more rapidly than inputs by photosynthesis (Prentice, 2001). To understand why the soil has a particular organic matter content, it is necessary to quantify both inputs and outputs of C to the system. Unfortunately these are both difficult to measure and the difference between input and output is normally small, further complicating estimates of change in storage. To predict how the organic matter content responds to changing environmental and management pressures at local, regional and global scales, we need robust models that take account of the complex physical, chemical, and biological controls influencing organic matter turnover. Models of soil organic matter turnover allow us to encapsulate our understanding of the mechanisms regulating organic matter dynamics in soil and then to investigate the effects of environmental and management change on rates of turnover. Although short-term (less that 1 year) environmental and management changes remain difficult to predict, a number of models can successfully predict longerterm change in a range of diverse environments (Smith et al., 1997; Parton et al., 1998; Paustian et al., 1997b). The integration of soil organic matter models within whole ecosystem simulations allows a more holistic assessment of ecosystem responses to environmental change and is highlighting management strategies that can be used to optimise C sequestration through targeted management of soils and vegetation (Paustian et al., 2000; Eve et al., 2002). Models have also highlighted the complexity of interactions that exist within the soil environment, and have indicated areas where our understanding is still incomplete. This review aims to summarise our understanding of the mechanisms by which plants influence the quantity of C contained in soils, through both a consideration of the processes of C addition and loss. We then go on to discuss how land management can be used to influence that content in agro-ecosystems.

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2. Carbon addition to soil Most C additions to soils originate from plants. Plant C can be added to soil by deposition of leaf litter on the soil surface and incorporation of crop residues into the soil after harvest. Root systems also make a significant contribution to C inputs (Fig. 1). It has been widely reported that around 50% of the C fixed in net photosynthesis is transferred below ground and partitioned between root growth, rhizosphere (root plus microbial) respiration and addition to soil organic matter (Lynch and Whipps, 1990; Nguyen, 2003). Typically 5–10% of the net fixed C can be recovered in soil. These figures are based on a summary of 95 14C-labelling experiments covering a broad range of plant species (Farrar et al., 2003). However, for the most part the experiments are relatively short-term and conducted on young plants. In an annual crop, net assimilation and transfer of C below ground declines as the crop flowers, sets seed and matures. In a discussion of C sequestration it is, therefore, more relevant to consider C inputs over the entire cropping season. It has been estimated that a typical winter wheat crop grown in temperate W. European conditions may transfer over 1.6 t ha 1 of C below ground in a season; approximately 18% of the net C fixed in photosynthesis (Swinnen et al., 1995a). After accounting for the C lost as CO2 from root respiration, the

Continuous labelling (a) Pulse labelling (b)

CO2

Shoot biomass 57a 63b

Root biomass 22a 19b

Rhizosphere respiration 14a 12b Soil pools 7a 5b % of net fixed C

Fig. 1. Fluxes of net C determined using 14C isotopic tracers. Results for continuous and pulse labelled experiments were compared in 43 studies of 8 arable and 12 perennial grassland species. Data from Nguyen (2003).

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input of organic C was in the order of 1.1 t ha 1. To put this in context, it is equivalent to about 30% of the potential C input from straw if all the straw were to be incorporated, and clearly root-derived C is the major input if the straw is baled and removed. These estimates of below ground C transfer are broadly in line with those reported from other work on cereals (Table 1). By comparison estimates for other crops can differ widely. The variation may result, in part, from genotypic differences in net C assimilation and partitioning, although a significant part of it may be associated with differences in the duration and experimental conditions adopted (Table 1). Animal wastes are a further source of C input to soils, but they are quantitatively much less important than plant additions. The input of C to soil from root systems has been termed rhizodeposition. This is a collective term for several component processes, exudation, secretion, sloughing and lysis of cells and root tissue senescence. Carbon dioxide lost in root respiration is sometimes included as a component of rhizodeposition (Lynch and Whipps, 1990), but often the term is reserved for the deposition of organic material (Swinnen et al., 1995a). Although the above component processes have been recognised for many years, our understanding of the underlying mechanisms and their quantitative contribution to overall rhizodeposition remains poor. The main reason for this is the difficulty in isolating the individual components experimentally, particularly under field conditions. The major challenge ahead, therefore, is to improve our estimates of rhizodeposition of C and our understanding of the mechanisms involved. This is important for at least two reasons. Firstly, the chemical nature of the C material deposited via exudation, secretion, cell sloughing and root tissue senescence will differ in a way that is likely to influence its fate in the soil. For example, in laboratory incubations, the mineralisation rate of glucose and root mucilage-C is greater than that of root tissue-C (Mary et al., 1993). Consequently, accurate estimates of the relative contributions of these different fractions to total rhizodeposition, under realistic soil conditions, are necessary in order to predict their impact on soil C sequestration. Secondly, a greater understanding of the physiological and molecular mechanisms involved in rhizodeposition would help identify opportunities

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Table 1 Examples of estimated root-derived C inputs Species

Organic C input (t ha 1)

Growth period

Comments

Reference

Sequential 14C pulse labelling; field experiment, the Netherlands. Growth period covers tillering to grain maturity Sequential 14C pulse labelling; field experiment, the Netherlands Sequential 14C pulse labelling; field experiment, Australia. *Estimate includes C lost in root respiration Estimated from measurements of root biomass at crop maturity, and assumed values of exudation and root turnover; field experiment, Canada Estimated from measurements of root biomass at crop maturity, and assumed values of exudation and root turnover; field experiment, Canada Estimated from measurements of root biomass at crop maturity, and assumed values of exudation and root turnover; field experiment, Canada Sequential 14C pulse labelling; controlled environment experiment Range represents high and low N fertilizer supply respectively; controlled environment experiment

(Swinnen et al., 1995b)

Winter wheat

1.0 –1.1

167 days

Spring barley

0.8 –1.3

137 days

Wheat

1.3*

175 days

Winter wheat

1.1

Oats

1.2

Barley

2.1

Perennial ryegrass

2.8

111

Lettuce

0.12– 0.16

60

(Swinnen et al., 1995b) (Keith and Oades, 1986)

(Bolinder et al., 1997)

(Bolinder et al., 1997)

(Bolinder et al., 1997)

(Kuzyakov et al., 2001) (Kuzyakov et al., 2001)

Includes C allocated to root biomass, plus rhizodeposition of organic C.

for manipulating C inputs to the soil via the root system. 2.1. Pathways of rhizodeposition Exudation refers to the diffusion of low molecular weight compounds, principally sugars, amino acids and organic acids, across the plasmalemma of root epidermal and cortical cells into the soil solution. Sugars generally form the majority of the exudates, with rates of exudation and final concentrations often 3–10 times higher than that of amino acids and organic acids (Hodge et al., 1998; Jones et al., 2003a). Inhibition experiments have shown (at least for the few species studied) that this efflux is energetically passive, driven by the steep concentration gradient between the cell cytoplasm (mM concentrations) and soil solution (AM concentrations, Jones and Darrah, 1994b; Sacchi et al., 2000; Nguyen, 2003). In the case of organic acids the efflux is also dependent on the

electrochemical potential gradient across the membrane (Jones et al., 2003a). Some species can regulate exudation of organic acids in response to certain soil conditions by opening specific anion channels to facilitate diffusion (Xia and Saglio, 1992; Jones, 1998; Ryan et al., 2001). Carbon fluxes between the root and soil are not uni-directional, with plant roots having the capacity for re-capture of sugars and amino acids (Jones, 1998; Sacchi et al., 2000). Low molecular weight sugars, proteinaceous amino acids and organic acids are not the only material released from roots. Some species also secrete enzymes such as phosphatases, and release non-proteinaceous amino acids (e.g. phytosiderophores) in response to specific nutrient deficiencies (e.g. P and Fe). These help increase nutrient availability through solubilization or chelation of soil minerals (Engels et al., 2000). Moreover, many roots are covered by a layer of C-rich mucilage. Evidence suggests that this mucilage is secreted by the outer

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layers of root cap cells. The process is energy dependent and involves transport by Golgi vesicles and secretion into the cell wall and intercellular spaces (Roy et al., 2002). The mucilage is composed of high molecular weight polysaccharides, some protein and phospholipid (Matsuyama et al., 1999; Osborn et al., 1999; Knee et al., 2001; Read et al., 2003). It has been implicated in a number of important functions including, lubrication of the root tip as it penetrates the soil, adsorption of, and possible protection of the root from, heavy metals, maintenance of root–soil contact during soil drying, and stabilisation of soil structure (Engels et al., 2000). The outer layer of cells of the root cap become detached, and may be embedded in the mucilage, as the root grows through the soil (Iijima et al., 2000, 2003a). They are steadily replaced by new cell production. In the majority of species studied, the cells are alive when detached and remain viable for sometime. These sloughed border cells may contribute to the lubrication of the root and help protect the main root from soilborne pathogens (Bengough and McKenzie, 1997). As a root system develops there is a gradual degeneration and death of specific cells, tissues or whole roots in most species. Root hairs produced a short distance behind the root tip, are often considered short-lived (Nguyen, 2003). In the main root axes of some graminaceous species, including wheat, barley and perennial ryegrass, there is a progressive senescence of most of the cortex. This is a poorly understood process, and its contribution to C deposition is uncertain. The major unknown is the extent to which any mobilised C is retained and re-translocated by the root. Earlier research using a vital staining technique, suggested that cortical senescence was a programmed developmental event (Henry and Deacon, 1981; Lascaris and Deacon, 1991). However, the technique was later discredited because of doubts over the ability of the stain to penetrate the tissue (Wenzel and McCully, 1991). In our laboratory we have followed the progress of cortical senescence along the seminal roots of wheat by mapping the presence and absence of turgor pressure using a single cell pressure probe and found it to be comparable to that revealed by staining (Bingham, unpublished results). We are currently investigating the potential for retranslocation of C from the senescing regions of the

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cortex. Tissue senescence may be less evident in other graminoids such as maize, where only the epidermis tends to be lost. The root cortex may also be sloughed off, as a result of radial expansion, in dicotyledonous species that produce secondary thickening. Loss of whole roots represents the most dramatic form of C transfer from plant to soil and may occur through senescence processes or the activity of herbivores. Median root lifespans of crop species show a large range from several weeks to many months (Pritchard and Rogers, 2000). In general, annuals and perennials show broadly different root life histories. This is probably a function of annuals needing to complete their life cycle in 1 year, whereas perennials are not constrained to one season. Annual species tend to have an early phase of high root production through until a definite stage of the life cycle (e.g. flowering). After that, new root production may do one of two things. It may stop, accompanied by very low mortality, so that most roots are maintained until the end of the season (Fitter et al., 1996). Alternatively root production decreases and mortality increases, resulting in a slow decline in the root population (Huck et al., 1987). The latter is the most commonly observed pattern for annual crops (Pritchard and Rogers, 2000). Our own observations on barley have shown that mortality commences around the start of stem extension, and can exceed new root production during grain filling (Bingham, unpublished results). Perennials often produce and lose roots simultaneously so that the standing crop of roots at any one time represents balance of these processes (Hendrick and Pregitzer, 1993). This replacement of fine roots in perennial systems may account for 50–80% of net production (Caldwell, 1977). One reason for this high turnover is that it may be energetically less expensive to produce a new root than it is to maintain an existing root that is in an area of depleted soil resources (van der Werf et al., 1988). This has been used as explanation for discrepancies between C assimilation and standing biomass estimates (Fitter et al., 1997; Cheng and Johnson, 1998). 2.2. Methodologies Before the relative contributions of these different pathways to total C deposition can be assessed it is important to consider the techniques used for

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determining rhizodeposition and their inherent limitations. Techniques include trapping and collection of border cells, mucilage and root exudates in sterile and non-sterile media (Hodge et al., 1996; GroleauRenaud et al., 1998; Engels et al., 2000), biosenor reporting of exudation (Jaeger et al., 1999; Darwent et al., 2003), isotopic labelling using 14C or 13C (Meharg, 1994), direct observation of roots (Watson et al., 2000) and mucilage production (Iijima et al., 2003b). All have their limitations and results must be interpreted cautiously. Techniques utilising sterile media, microscopic examination of roots and biosensor reporters can realistically only be used in laboratory experiments and usually on relatively young plants. Sterile media are often used to quantify exudation since loss of exudates through microbial activity is avoided. However, this may lead to unreliable estimates because micro-organisms can themselves stimulate exudation, possibly by increasing the concentration gradient for diffusion (Meharg and Killham, 1991; Jones et al., 2003a). They may also influence other components of rhizodeposition through effects on root growth and morphology (Lynch and Whipps, 1990). Solution culture has been widely used in these studies, but it could lead to an overestimation of exudation and mucilage production, if the diffusion gradient is increased compared to that found in soil, by flow of solution over the root surface (Jones et al., 2003a; Nguyen, 2003). Further, the contribution of root tissue senescence to rhizodeposition of C is generally not accounted for in these studies, because of their short duration. Isotopic labelling and observational techniques have been the methods of choice for most field experiments. Labelling has the advantage of allowing C inputs into the soil to be followed directly, but the estimates of C addition may vary depending on the method used. Pulse-labelling is easy to use in the field, but can result in lower estimates of transfer below ground compared to continuous labelling (Meharg, 1994; Nguyen, 2003, Fig. 1). This is because short chase periods can lead to incomplete translocation before the experiment is terminated. Pulse-labelling also raises difficulties in estimating true C fluxes because of isotope dilution during the chase period. Moreover, the method does not lend itself to estimating soil C inputs from root senescence

and decay of structural material. Continuous labelling overcomes these problems, but is technically difficult to achieve over long time periods in the field. A compromise is to use repeat pulse-labelling over the whole season, although this may not uniformly label all plant C pools. Using isotopic techniques it is possible to get reasonable estimates of C transfer below ground (Warembourg and Estelrich, 2000) and we can be confident that most of the C that is transferred will ultimately be added to the soil OM pool or released as CO2, especially, in the case of annual crops, if the post harvest period of root decay is also considered. Typically, only small amounts of organic C are lost by leaching (McCracken et al., 2002). The major uncertainties arise when attempting to estimate the partitioning of below-ground labelled-C in the different root and soil fractions. For example physical separation of fine roots from rhizosphere soil is very difficult and in many cases the so called rhizosphere pools will contain root tissue, or the contents of cells damaged during extraction, rather that just soil C (Jones et al., 2003b). This will tend to overestimate rhizodeposition via exudation and secretion. Another technically complex problem is the separation of root respiration from microbial respiration. Labelled-C collected as CO2 released from the soil may be derived directly from root respiration or after exudation and subsequent metabolism by micro-organisms. Thus the accuracy of estimates of exudation hinge on the ability to measure the relative contributions of root and microbial respiration. A number of techniques have been developed to do this, but none are wholly satisfactory (see below). Other approaches using the natural abundance of C isotopes to quantify plant C addition to soils have proved useful. Plants with C3 and C4 metabolisms have slightly different 13C contents with C3 plants having more negative y13C values than C4 plants. Thus when it can be established that a soil has supported a vegetation consisting of only C3 or C4 plants over a long period (decades), then by switching the plant type, the rate of addition of new plant C can be measured by quantifying the rate at which the soil’s isotopic signal changes (Boutton, 1991). This technique has proved valuable both in terms of quantifying rates of C input to soils and in allowing us to determine the ages of individual C pools within the

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soil (Roscoe et al., 2001). Its application is however limited by the need to have sites with a long-term history of either C3 or C4 plants that can be swapped prior to the commencement of a study. The technique also suffers from similar difficulties to studies using 14 C labelling, in separating fine root material from bulk soil prior to analysis. The contribution of tissue senescence and mortality to rhizodeposition can also be assessed by direct observation. There are two basic approaches. The first is destructive sampling of root and soil to quantify root length and biomass. The second is nondestructive measurement of root number, length and diameter using rhizotrons and minirhizotrons . Both approaches have been reviewed extensively elsewhere (Oliveira et al., 2000; Hooker et al., 2000; Quideau et al., 2001). Minirhizotron observations have the advantage of enabling the dynamics of root production and mortality to be determined in the field. In order to estimate C inputs, results must be scaled up using a calibration with root length densities determined from soil coring (Crocker et al., 2003). An accurate measure of the average C content of the root tissue is also required and this can be difficult to obtain because of losses associated with washing and sampling roots from soil cores. Crop root systems often have many very fine roots. These are difficult to see and harvest by eye, and are prone to mechanical damage during sampling, particularly by snapping off inside small soil aggregates. Although fine roots make a small contribution to total root biomass they may represent a high proportion of root length and often have the highest nutrient contents but lowest C content (Pregitzer et al., 2002), resulting in a low C:N ratio and potentially high decomposition rate relative to thicker roots. The loss of soluble compounds during washing is a further potential source of error but affects older, brown roots more than young roots (Bo¨hm, 1979). Losses by this route may be minimised by processing samples within a day of collection (Van Noordwijk, 1993). Other major difficulties with the minirhizotron technique are in establishing reliable criteria for assessing root death and in estimating the amounts of nutrients, including C compounds, retranslocated prior to senescence (Hooker et al., 2000). Published studies have used a variety of visual (e.g. colour,

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surface texture) characteristics to classify roots as alive or dead. Separations based on these attributes are, however highly subjective and classification is prone to large variation between operators. A further complication is that many of the characteristics relate more to root decomposition than to actual death. Validation of visual characteristics against physiological measurements of root viability such as tetrazolium chloride staining have been used with some success (e.g. Comas et al., 2000), but their application is complicated by problems with penetration of the stain into the tissues (Ruf and Brumner, 2003). In our laboratory we have used the presence or absence of turgor pressure as an indicator of root viability. In clover this relates well to root colour. 2.3. Contribution of pathways to total rhizodeposition Summarising the results from a selection of experiments reported in the literature, Nguyen (2003) has highlighted the typical range observed for the production of exudates, border cells, mucilage and root hairs. Direct comparison between species is difficult because of the wide range of culture conditions, plant ages and units of measurement reported in the different studies. For example rates of exudate production (in solution culture) ranged from 0.1 to 7% root DM day 1, but for comparative purposes this is not the most useful form of expression. Net exudation is tightly coupled to C translocation to the root (Dilkes et al., 2004) and the greatest rate occurs close to the root apex where the phloem is unloaded (Jones and Darrah, 1994a). Consequently, rates expressed on the basis of standing root biomass are likely to change markedly with plant age. Thus a more useful form of expression is in terms of new root growth. Rates of exudation of approximately 196–226 Ag C mg 1 DM root growth have been reported for wheat seedlings grown axenically (Prikryl and Vacura, 1980). Root cap cell production ranged from 1 to 2.8 Ag C mg 1 DM root growth and mucilage production from 4 to 19 Ag C mg 1 DM root growth (assuming a C content of 40%, Nguyen, 2003). Using some of the above values for wheat, and assuming a maximum root biomass of 126 g m 2, which is typical for winter wheat crops in the UK, the relative season-long inputs of C from root

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biomass, exudation, the production of root cap cells and mucilage, and the sloughing of root hairs can be calculated (Table 2). For simplicity the calculations assume that root mortality prior to flowering is negligible. No account has been made of cortical senescence because we have insufficient knowledge of the proportion of the root system that is affected. The calculations suggest that major pathways for rhizodeposition are exudation and the ultimate death and decay of root tissue post-flowering and postharvest. Root cap cells, mucilage and turnover of root hairs contribute relatively little to the total. In spite of the crude nature of these calculations, the estimate of C input by exudation is broadly comparable with those determined by 14C labelling (Swinnen et al., 1995b). The latter will include C from sloughed cells, tissue (cortex) and mucilage. 2.4. Factors influencing rhizodeposition Numerous experiments using exudate trapping, biosensor reporting and isotope labelling techniques, have shown that rhizodeposition of C is influenced by a wide range of soil, climatic and plant management factors (Paterson and Sim, 2000; Sacchi et al., 2000; Darwent et al., 2003; Nguyen, 2003). It is clear from the mechanisms outlined above, that rhizodeposition

Table 2 Estimated cumulative C allocation to different root fractions of winter wheat for the period up to flowering (maximum root biomass) Fraction

kg C ha

Living root biomass

504

Exudates

285

Root cap cells and mucilage

Sloughed root hairs

3

22

1

Notes Assuming root mass at flowering of 126 g m 2 (Barraclough and Leigh, 1984) and C concentration of 40% Based on an exudation rate of 226 Ag C mg 1 DM root growth (Prikryl and Vacura, 1980) Based on production rate of 6.4 Ag C mg 1 DM root growth and assumed C concentration of 40% (Nguyen, 2003) Assuming a root length at flowering of 23.4 km m 2 (Barraclough and Leigh, 1984), root hair density for wheat of 30 mm 1, medium hair length, and C content per hair of 1.12 ng (Nguyen, 2003)

is a function of the growth and ultimately senescence of roots. Root growth and morphology are highly dependent on the supply of photosynthates from the shoot. Thus rates of respiration, root extension and branching are all reduced if the supply is restricted (Bingham and Stevenson, 1993; Bingham et al., 1996, 1997). The supply of photosynthates will be influenced by factors such as light availability, foliar diseases and defoliation. The proportion of net fixed C transferred below ground varies with crop phenology, the greatest transfer occurring during the vegetative phase (Keith and Oades, 1986; Swinnen et al., 1995b). Thus crops with a longer vegetative period may sustain higher rates of net assimilation and transfer below ground over the season. This may account in part for the higher C deposition by grasses compared to cereals (Table 1; Kuzyakov, 2001). In addition, species with similar life histories may also differ in patterns of C partitioning. In a comparison of 12 species, the percentage of fixed 14C transferred below ground ranged from 41% to 76% (Warembourg and Estelrich, 2000). Root growth and, hence rhizodeposition, is also affected directly by soil conditions. These may influence some components of rhizodeposition, but not others. For example in sand culture, compaction has been shown to increase border cell production by young maize roots, but not total C deposition (Iijima et al., 2000) or exudation when expressed per unit root weight (Bingham and Bengough, unpublished data). Modification of root system morphology by a variety of soil factors is likely to affect exudation. The number of root tips is especially important as these are the preferential sites for exudation. For example, low N supply has been found to increase C exudation in several species (Hodge et al., 1996; Paterson and Sim, 2000; Darwent et al., 2003). In barley, the increased exudation was correlated with the greater number of root tips resulting from a greater root length (Darwent et al., 2003). Moreover, low N supply typically increases soluble sugar concentrations in roots (Bingham and Farrar, 1989; Bingham et al., 1997). On the other hand, a number of studies have demonstrated that total rhizodeposition may be decreased by N deficiency, although the effects are species dependent (Liljeroth et al., 1990; van der Krift et al., 2001). Thus N addition to grasses adapted to high soil fertility (Holcus lanatus) doubled rhizodeposition of 14C,

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whilst low soil fertility grass species (e.g. Festuca ovina) showed no response. The amount of rhizodeposition was correlated with root biomass (van der Krift et al., 2001). These examples illustrate the importance of taking a long-term view when considering the potential impact of biotic and abiotic factors on C deposition. Although low N supply might lead to a short-term increase in root length and C exudation in young cereal plants compared to those well supplied with N (Darwent et al., 2003), the longer-term consequences of low N availability are usually a reduced canopy size, restricted light interception and hence restricted biomass production. Thus despite allocation of a greater proportion of total plant biomass below ground in N deficient plants, the net seasonal effect will be a reduced C input to the soil, if the overall mass of the root system is less. In order to maximise C inputs we must therefore maximise crop productivity. Factors that influence the balance between inputs via exudation and root death and decay may help determine the subsequent fate of the organic matter deposited. Exudates are highly labile, whereas additions of C from dying roots include structural material (e.g. cellulose and lignin) with lower potential decomposition rates. The release and fate of organic substrates from roots is influenced by soil microbial communities. The rhizosphere is known to have a larger microbial population and a higher microbial activity than the surrounding bulk soil (Hbjberg et al., 1996). A study of the rhizosphere environment is therefore necessary to understand the responses of plants and soils to global environmental change, (Janssens et al., 1998). Alterations in C flow to the rhizosphere can be induced by increases in atmospheric CO2 concentration, which can cause a significant increase in C input to the rhizosphere, either in the form of increased or chemically altered exudation, through increased litter production, or through increased production and turnover of fine roots (Pregitzer et al., 1995; Paterson et al., 1996; Hungate et al., 1997). However, the effect of this increased C input to the soil is not consistent. The microbial biomass and activity in the rhizosphere have been shown to increase due to elevated CO2 (Dı´az et al., 1993; Schenk et al., 1995; Rees and Parker, 2001), but neutral responses or a decline in activity have also

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been reported (Rice et al., 1994; Ross et al., 1995; Schortemeyer et al., 1997). These conflicting results could be partly explained by the role of plant species or community in mediating the response, with the added effects of differential nutrient availability (O’Neill, 1994). It is also possible that the quality of substrates released as a consequence of alterations of C flow to the root is altered (Rees and Parker, 2001; Paterson et al., 1996) thereby inducing different microbial responses.

3. C loss from soil Carbon can be lost from soils as a consequence of respiration, leaching and soil erosion (Lal, 2003). On a global basis respiration is thought to be more important than erosion, and leaching losses. There is however a difficulty in assigning losses to different processes. Ultimately all C fixed by photosynthesis will be oxidised and returned to the atmosphere, and erosion can be viewed a transport process that influences the location of the environment in which oxidation takes place (Gregorich et al., 1998). This review focuses on the processes of microbial respiration in soils in contributing to soil C turnover. The microbial biomass that plays a major role in transforming inputs of organic matter also controls C loss. All C fixed through photosynthesis is ultimately returned to the atmosphere as CO2. If this were not the case then the biosphere would slowly grind to a halt as C accumulated in a recalcitrant pool, even if the rate of accumulation were very small. On a global basis the production of CO2 by respiration from land surfaces is approximately 55 Pg C year 1 (Prentice, 2001). This flux is slightly less than the uptake of C by net primary production and contributes to a significant sink of approximately 1.9 Pg C year 1 in the terrestrial biomass. There is considerable uncertainty regarding the size of this sink, however, the approximate balance of C exchange between soils and the atmosphere in tropical regions suggests that the northern hemisphere is likely to be accumulating C in terrestrial pools (Schimel et al., 2001). The distribution of C stocks and therefore the potential for C loss from the world’s soils is highly uneven with the largest C densities occurring at high latitudes in the N hemisphere (Schimel et al., 2001).

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Conceptually it is usually considered that C exists in soils in a number of discrete pools with varying rates of turnover and loss. In order to simulate C loss from soils, it is commonly assumed that soil organic matter can be divided into a labile pool and one or more recalcitrant pools (which normally dominate the overall C pool), each decaying according to first order kinetics. A range of models have been developed and tested extensively using this basic approach such as Century (Parton et al., 1998), Roth-C (Coleman et al., 1997), Candy (Franko et al., 1995), and DNDC (Li et al., 1992). These models are largely empirical, (using soil pools that are not verifiable by experimental measurements), but have been shown to provide good predictions of C loss particularly over longer time periods. To obtain a more mechanistic understanding of the processes of C loss from soils, it is helpful for models to simulate the behaviour of measurable C pools. The choice of pools used to simulate C dynamics is constrained by the need to ensure pool homogeneity and uniqueness (Smith et al., 2002), however, recent work by Sohi et al. (2001) and Six et al. (2002) has shown that models can be developed on the basis of fractionation of organic matter pools by density. These new approaches promise to elucidate the role of physico-chemical controls on organic matter breakdown in soils and ultimately to improve our predictions of C balance. Soil management plays an important role in influencing losses of C by respiration. Data in which field measurements (mostly using dynamic chambers) of soil respiration were recorded from arable and grassland soils shows an almost six fold variation in fluxes, ranging between 4 and 26 t C ha 1 year 1 (Table 3). These studies also show that rates of respiration are altered by various management operations. There is strong evidence that cultivation contributes to a significant increase in respiration as a consequence of the accelerated oxidation of labile C (Eriksen and Jensen, 2001; Alvarez et al., 1995; Soussana et al., 2004). There is also good evidence to demonstrate that soil respiration rates are influenced by grazing (Cao et al., 2004), drainage (KasimirKlemedtsson et al., 1997), and manure addition (Jones et al., in press). The losses of C that result from soil respiration will normally be accompanied by an input through photosynthesis. The difference between net ecosystem respiration and net C uptake is defined as

net ecosystem exchange (NEE), and provides an indirect measure the system’s C sequestration potential. Such measurements are most appropriately made by micrometeorological techniques in which exchange of C at the land surface is measured continuously. A range of such studies have been published in the past decade, and all of those presented in Table 3, were conducted over 1 year or longer. Although there is a scarcity of such measurements in arable agriculture, these studies (mainly grassland) again indicate the importance of management in controlling C flows (Soussana et al., 2004). Although it is not possible to make direct comparisons between the data describing soil respiration and NEE, it is interesting to note that the respiration fluxes are generally one or two orders of magnitude larger than NEE. This illustrates the large internal flows of C within ecosystems and highlights the importance of understanding the processes controlling respiration rates if we wish to influence C sequestration. The role of environmental variables in controlling rates of soil respiration have been extensively reviewed and are a consequence principally of temperature and water availability in influencing organic matter decomposition (Pendall et al., 2003). Various studies have shown that different forms of management can contribute to different rates of loss of C from the soil (see Section 4). However, these studies do not provide us with any valuable information about the origin of the C lost, in particular whether it originates from the plant, or from a soil pool. In order to obtain a mechanistic understanding of the processes contributing to C loss from soils and thus ultimately to C storage, we must be able to partition losses from root and microbial respiration. There are four methods that can be used to achieve this: (i) an isotope dilution method uses the dilution of a pool of labelled sugars in the rhizosphere to quantify root C inputs (Cheng et al., 1993), (ii) labelled rhizodeposits to measure C loss directly from roots (Swinnen, 1994), (iii) use of characteristic differences in the time periods of C loss by roots and microorganisms following pulse labelling to discriminate between microbial and root respiration (Kuzyakov et al., 1999), (iv) the elution of 14C labelled exudates from soil before micro-organisms utilise them (Kuzyakov and Siniakina, 2001). A comparison of these methods by Kuzyakov (2002b), showed that the

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Table 3 Soil respiration and net ecosystem exchange of C from field measurements of crop and grassland soils Country

Crop

Management

Measurement period

Soil respiration (t C ha 1 year 1)

Reference

Argentina

Wheat/soya

Ploughed Mintilled

April 1991–Mar 1992

(Alvarez et al., 1995)

Canada Canada China

Barley Fallow Barley Alpine

Denmark

Grassland

11.60 11.50 4.93 10.85 14.19a 4.17 5.56 5.62b 10.44b 15.47

Sweden

UK

Grass

US, Kansas US, N Dakota

Grassland/ Prairie

US, Nebraska

Grass

US, Colorado

Prairie

Country

Crop

Canada

Denmark

Arable

France

Grassland

Hungary Italy Switzerland

Grassland Grassland Grassland

UK

Grassland

US, N Carolina US, Oklahoma US, Arizona

Grassland Grassland Grassland

US, California

Grassland

a b c

1994–1996

Intensive grazing Extensive grazing Uncultivated Ploughed Fallow Drained cereals Drained row crop Unfertilised Ammonium nitrate Poultry manure Ungrazed Simulated grazing Grazed Ungrazed Day Night Heavy grazing Light grazing

May 1998–Apr 1999

1989

April 2002–Dec 2003

April–Oct. 1996–2000 Jul–Aug 1987 1995–1997

Management

Measurement period

Ungrazed and Uncultivated

1998 1999 2000 2002 2003 2002–2003

Barley Barley+Grass High N input Low N input Alpine High N input Low N input Ploughed Cut Cut annually Burning

Low productivity

Extrapolated from a 69 day study. Extrapolated from a 3 month study. Extrapolated from an 18 day study.

2002–2003 2002–2003 2002–2003 2002–2003 April 2001–April 2002 March 1997–March 1988 1997 1998 1999 2000 1991

5.45 19.09 11.13 11.13 17.23 13.47 11.20 25.19 21.17 10.32c 8.60c 26.49 24.14 Net ecosystem exchange (t C ha 1 year 1) 1.10 0.22 0.18 0.11 1.20 0.40 1.40 1.20 4.31 4.02 2.81 4.49 0.29 0.97 0.07 1.31 2.12 1.10 0.58 1.33

(Akinremi et al., 1999) (Rochette et al., 1992) (Cao et al., 2004) (Eriksen and Jensen, 2001) (Kasimir-Klemedtsson et al., 1997)

(Jones et al., 2004)

(Bremer et al., 1998) (Frank et al., 2002)

(Grahammer et al., 1991) (LeCain et al., 2002)

Reference (Flanagan et al., 2002)

(Soussana et al., 2004) (Soussana et al., 2004) (Soussana et al., 2004) (Soussana et al., 2004) (Soussana et al., 2004) (Soussana et al., 2004) (Novick et al., 2004) (Suyker and Verma, 2001) (Emmerich, 2004)

(Valentini et al., 1995)

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estimate of the contribution of root respiration varies between 40% and 80% (Fig. 2). One difficulty is that each method is dependent on a number of assumptions, some of which may be difficult to justify. In addition to the difficulties of measuring the C loss from plant and microbial respiration it is necessary to consider the importance of soil nutrients in influencing the process. The presence of nutrients can fundamentally alter the rate at which C added by plants is cycled in the soil. One of the unusual features of organic-C is that it provides a source of both nutrients and energy, since organic-C is required for both tissue construction and energy generation by organisms. Increased C supply in the rhizosphere can slowdown N cycling and produce a feed back on plant growth which has been termed the soil hypothesis (Loiseau and Soussana, 2000). The soil N supply can both limit C sequestration and interact with CO2 to influence mineralisation and nitrous oxide release. A number of detailed studies have been able to show this interaction between C input to the soil and N supply. In a study by van Ginkel et al. (1997) net C input to the soil was strongly limited by the combined affects of elevated CO2 concentrations and N availability. This effect resulted from increased dry matter production combined with reduced decomposition of

100%

Root derived CO2

80%

60%

40%

20%

0% Isotop Dil Model RD

Root

14-

CO2

Exudates

Microbial

Fig. 2. The partitioning of root and microbial respiration using four contrasting isotopic techniques. Isotop Dil=An isotope dilution method (Cheng et al., 1993), Model RD=Model rhizodeposition (Swinnen, 1994), 14CO2=a method modelling 14CO2 efflux (Kuzyakov et al., 2001) Exudates=an exudate elution procedure (Kuzyakov and Siniakina, 2001) Data from Kuzyakov (2002b).

root material at the high CO2 concentration. In work carried out by Martin-Olmedo et al. (2002), N recovery from soil derived pools in an elevated CO2 environment was significantly higher when there was a high background N concentration. This was probably a consequence of increased immobilisation taking place where C was added to the low N soil. In these circumstances, microbial populations can mineralise SOM in preference to root-derived substrates because the former contains higher relative concentrations of nutrients. These results demonstrate that as C input to the soil changes, below ground processes can be expected to change at the same time, and this can then have an important influence on soil organic matter storage. Models used by Cannell and Thornley (1998) indicate that responses of grasslands to elevated CO2 would vary according to their N status, with N poor grasslands showing little short-term increase in C storage due to N limitation in the plant biomass which would in turn restrict C addition to the soil. Experimental evidence to support this hypothesis is sparse or ambiguous (Soussana and Hartwig, 1996). Van Groeningen et al. (2002) showed in an eight year FACE experiment that elevated CO2 did not result in any differences in C storage in treatments with high and low-N availability. However, rates of C sequestered by the particulate organic matter fractions per unit of N added were higher in the low N treatment. Experiments that have looked at the effects of nutrient availability independently of other factors have generally shown little influence on organic matter accumulation. Glendining et al. (1996) showed that in an experiment maintained at Rothamsted in England for over 150 years, where different rates of N fertiliser were applied to fields planted with winter wheat, there was no significant impact on concentrations of soil organic matter, although the amounts of potentially available N were increased. This suggests that the quality of the organic matter pool was changed in response to fertiliser addition despite the overall pool size remaining constant. Halvorson et al. (2002) also found that N applications to winter wheat had no effect on soil organic matter accumulation despite the return of increased amounts of crop residues. These effects are likely to be explained by increased decomposition rates of residues from the crops receiving higher fertilisation inputs.

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The interaction between C deposition enhanced by elevated CO2 and soil nutrient dynamics has been studied in our laboratory (Rees, unpublished). The response of a range of grassland soils taken across a gradient of N deposition to CO2 enrichment was examined over 18 months. Dry matter production increased in plants exposed to elevated CO2, although the largest responses were at sites with lowest N deposition, contrary to model predictions (Cannell and Thornley, 1998). Although no significant differences in soil respiration between treatments were observed in this study, the consequences of increased C inputs on soil nutrient transformations were apparent. Soils that were maintained under an elevated CO2 atmosphere, showed a small but significant increase in the loss of dissolved organic C. Within the rhizosphere, C fixed by photosynthesis is released in substrates containing N. As C flow from the plant increases (for example as a consequence of CO2 rises) the C:N ratio of substrates increases as additional flow of C is not necessarily accompanied by additional N. The C:N ratio of rhizodeposits (in short-term experiments this is likely to be largely in the form of exuadates) therefore rises. It is this change in C:N ratio that has a large potential effect on processes of mineralisation, immobilisation and microbial activity, with feedback influencing plant growth (Paterson, 2003). A number of contrasting scenarios can result. Lower C:N ratios promote Nmineralisation and more rapid soil N turnover of root derived substrates. Alternatively it is possible that the higher C:N ratios of plant C will slow down microbial activity as limitations of N and other nutrients, slow down decomposition processes (Cheng et al., 1996). In circumstances where plant derived C becomes less available, micro-organism are more likely to utilise soil-derived C pools (Cheng, 1999) The switching between soil derived and plant derived C pools has a significant impact on the turnover of soil organic matter and can be demonstrated by the use of isotopic tracers. It was found by Liljeroth et al. (1994) that CO2 release from native soil organic matter was decreased when winter wheat was grown in soils with a high concentration of available-N. However, Cheng et al. (2003) found that soil organic matter decomposition was significantly enhanced by rhizosphere C additions through what is commonly described as a priming effect. Such effects are thought to operate by

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increased microbial growth in response to the additional C input followed by enhanced utilisation of soil-derived substrates. One of the difficulties in attributing alterations in soil derived C losses to inputs of plant C is that the presence of plants inevitably introduces a range of confounding factors such as altered water availability, and soil structural changes, thereby complicating the interpretation of such effects (Kuzyakov, 2002a). The problems associated with short-term measurements of C loss from soils are very different from those associated with measurements of C input. In broad terms it is possible to obtain reasonably accurate measurements of respiratory C loss from soil surfaces using as range of chamber and micrometeorological techniques (Janssens et al., 2000). The difficulty arises in attempting to partition this loss between the various source processes. Not only is there an uncertainty in determining the contribution of plant and microbial sources, but within the soil it is difficult on the basis of present knowledge to determine the relative contribution of different organic matter fractions and in explaining some of the complex interactions described above. This understanding is developing rapidly through the applications of techniques described in this paper and will assist us in improving our models of soil C dynamics.

4. Land management and C storage It is argued that land management can make a major contribution to C sequestration by the development of systems where biological inputs of C exceed losses (principally by respiration). Farming practice over many centuries has been used to achieve precisely this objective in building soil fertility, prior to the use of synthetic fertilisers. However, the new challenge of quantifying C flows in order to construct budgets of global C flows requires us to accurately quantify changes in C storage that result from alterations in land management. The conventional approach to this has been to use sequential measurements of soil C concentration, but such measurements are very poor at identifying short-term changes in overall soil C storage. This is because the addition or loss of C to a soil is difficult to quantify against a large background pool, particularly given the variability in

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soil organic C that is normally present. Conen et al. (2003) showed that by using repeated sampling, it would take between 26 and 43 years to identify a 10% (i.e. large) drop in soil organic C. However, improved estimates of input and loss discussed in the preceding two sections provide us with an alternative and shortterm method of identifying the changes in organic C storage, that result from land use change, and are likely to become increasingly important in assessing the value of options discussed in this section. The aim of agricultural management practices is generally to increase the production of harvestable plant parts for either human or livestock consumption. However, in a European context a move towards more sustainable agricultural practices and the delivery of environmental goods is resulting in an increasing awareness amongst the farming community of issues such as increasing organic matter and improving soil structure. The basic criteria for increasing soil organic C is that the amount of C added in residues, including plant roots, exceeds the amount of C lost in decomposition. Thus quantifying the effects of individual management practices and their combinations on C sequestration is vital for improving the potential of farming systems to sequester C (Smith et al., 2000). It has been estimated that crop residues in the US alone have the potential to sequester over 22 million tonnes of C (Lal et al., 1999). Increasing primary productivity of crops increases the quantities of residue added into soil in unharvested plant parts, this will include both above and below-ground residues. Since the 1940s there has been a steady increase in crop yield of major food crops (http:// statistics.defra.gov.uk). For example, wheat yields in the UK have risen from 2.82 to 7.08 t ha 1 between 1948 and 2001. This is a result of both agrochemical use and plant breeding. In arable crops the use of both inorganic (Paustian et al., 1992; Gregorich et al., 1996) and organic fertilisers (Sauerbeck, 2001; Fortuna et al., 2003), together with tillage (Paustian et al., 1997a; Smith et al., 1998; Follett, 2001), residue management (Singh et al., 1998; Duiker and Lal, 1999), and the choice of crops (Bolinder et al., 1997; Paustian et al., 1997a) including the use of cover crops (Sainju et al., 2003; Kuo et al., 1997) all have a major influence on both crop yield and C sequestration. The efficient management of both irrigation and pesticide use also play a part in this (Lal et al., 1999).

Weed biomass may also be an important contributor to C sequestration and will be affected by agrochemical use, although few published studies consider this effect. Within Europe (EU-15) it has been estimated that the C storage potential of cropland is around 90–120 Mt C per year (Smith, 2004). The potential of a range of different management practices, including tillage and different cropping strategies as well as more dramatic land use change, together with an estimate of the uncertainty of the values are given in Table 4. The highest potential value is for change from cropping to grassland but this must be set against the possible use of that grassland for grazing and consequent emissions of non-CO2 greenhouse gases. However, it has been suggested that in reality only 20% of the biological potential is actually achievable (Smith, 2004). Achieving this is intimately linked to changes in both legislation and agricultural policy. These estimates of potential C sequestration are considerably lower that the measured values of NEE presented in Table 3. Although the NEE measurements do not directly equate with C sequestration potential (since C Table 4 Carbon sequestration potential of land management options in Europe (from Smith, 2004) Practice

Soil carbon sequestration potential (t C ha 1 year 1)

Estimated uncertainty

Zero tillage Reduced tillage Set-aside Permanent crops Deep-rooting crops Animal manure Cereal straw Sewage sludge Composting Improved rotations Fertilization Irrigation Bioenergy crops Extensification Organic farming Convert cropland to grassland Convert cropland to woodland

0.38 (0.29)* b0.38 b0.38 0.62 0.62 0.38 (1.47)* 0.69 (0.21)* 0.26 0.38 N0 0 0 0.62 0.54 0–0.54 1.2–1.69 (1.92)* 0.62

N50% NN50% NN50% NN50% NN50% NN50% NN50% NN50% NN50% Very high Very high Very high NN50% NN50% NN50% NN50% NN50%

All estimates are based on extrapolation from Smith (2004), except those marked by asterisk (*), where the figure in brackets is derived from Vleeshouwers and Verhagen (2002).

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added to sites could be lost by other processes such as erosion and leaching), it does suggest that the sequestration potential of some grassland and arable sites may be greater that has been previously recognised. Carbon sequestration data from 67 long-term experiments have been recently analysed by West and Post (2002) to determine the impact of tillage practices. On average a change from conventional to no-till can result in sequestration of 57F14 g C m 2 year 1. This is within the range of values (10–60 g C m 2 year 1) given by Follett (2001) in a recent review. The mean residence time (MRT) of C is generally greater in no-till than conventionally tilled soils. For example, for corn in France the MRT was measured using 13C techniques and estimated to be 127 years for no-till compared with 55 years under conventional tillage (Balesdent et al., 1990). Ploughing depth can also influence C sequestration, Borresen and Njos (1994) found that where shallow ploughing (12 cm) took place soil organic C was higher in the surface layer than following deep ploughing (18 or 24 cm). The total soil organic C (SOC) in the upper 40 cm was however unaffected by plough depth. Rotavation (to a depth of 10 cm) as compared to ploughing increased the SOC in the surface layer (Borresen and Njos, 1993). Other work has shown that ploughing can increase soil organic C storage by transferring organic matter into deeper soil layers, where it decomposes more slowly (Nieder and Richter, 2000). Averaging the results of 7 long-term European field experiments Sauerbeck (2001) found that the soil organic matter content compared with an unfertilised control could not be raised by more than about 30%. Using practical application rates of both farmyard manure (FYM) and mineral fertiliser, soil organic matter increased in the order FYM+mineral fertiliserNFYMNmineral fertiliserNunfertilised. Fortuna et al. (2003) showed that the addition of compost made from dairy manure and oak leaves and added at a rate of 4.5 t ha 1 increased soil organic C by 14% over a 4 year period. In most crops there are a number of residue management options which will affect C sequestration. For example, in cereals, straw management options include incorporation or mulching postharvest, burning, or use for livestock bedding and then returning as manure at a later date and in a more

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decomposed form. Straw incorporation and mulching of residues have both been shown to have a positive effect on C sequestration (Singh et al., 1998; Duiker and Lal, 1999; Jacinthe et al., 2002). A linear relationship has usually been shown to exist between soil organic C content and C input from crop residues (Paustian et al., 1997c; Duiker and Lal, 1999). However, residue quality is also important; Paustian et al. (1992) found a positive effect on soil C of residues with higher lignin contents. C:N ratio of residues have also been shown to be important in C sequestration (Paustian et al., 1997a; Drinkwater et al., 1998). Both studies suggest that even low C:N ratio materials can be beneficial for C sequestration. The study by Lal et al. (1999) indicated the beneficial effects of grass cover crops compared with leguminous ones. West and Post (2002) found that increased diversity in crop rotation either through change from monoculture to rotation or by increasing the number of crops in rotation was associated with a change of 20F12 g C m 2 year 1, in an analysis of data from 67 experiments. Choosing species with large deep root systems is desirable for increasing C inputs (Fisher et al., 1994; Ma et al., 2001). It has been argued that C deposited deep in the soil profile is less prone to oxidation and hence subsequent loss (Fisher et al., 1994). Large differences in shoot:root ratio have been found between cereal species, but this does not necessarily lead to wide variation in root biomass. For example, in a study in Quebec shoot:root ratios varied from 4.9 for winter wheat to 2.5 and 2 for oats and barley respectively (Bolinder et al., 1997). However, these were mostly associated with differences in shoot biomass and no significant differences in root biomass were found (Bolinder et al., 1997). Similarly, no difference was found in root biomass between individual cultivars of wheat, oats, barley and triticale (Bolinder et al., 1997). Xu and Juma (1993) and Ma et al. (2000), on the other hand, have reported significant differences in root biomass between cultivars of barley and switchgrass respectively. In grassland there is generally more soil organic C than under cropland (Cole et al., 1993), as a result of several factors including lack of disturbance, greater return of plant residues, high root biomass, manure application and the return of dung during grazing. As

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with arable crops, grazing practices which increase grassland productivity have the potential to increase SOM and C sequestration (Conant et al., 2001). Grazing can result in higher soil C than ungrazed grass due to more rapid annual turnover of shoot material and also due to changes in species composition (Reeder and Schuman, 2002). They found that exclusion of grazing allowed an increase in annual forbs and grasses with less dense and fibrous rooting systems. Over and above management changes to cropping systems, more dramatic land use change has great potential to sequester C. For example, land use change from cropland to forest or grassland has been estimated to have a global sequestration rate of 338 and 332 kg C ha 1 year 1 (Post and Kwon, 2000). Lal et al. (1999) showed that tall fescue and smooth brome grass increased the mean SOC pool by 18.5% compared with a corn/soybean rotation. Fast growing energy crops have also been shown to substantially increase C sequestration compared with either conventional agricultural systems or unmanaged systems (Zan et al., 2001). Amendment with organic wastes incorporation of cereal straw, reforestation of surplus arable land and the extensification of farming through arable rotations all provide opportunities to increase soil organic matter. Smith et al. (2000) have examined each of these options on a Europe-wide scale in order to estimate the relative quantities of C that can be sequestered by adopting these practices on available arable land. Together with other recent studies (e.g. Cole et al., 1997) they show that land management practices can make a significant contribution to reducing CO2 loss from soils. One of the interesting aspects of this work is the comparison between the application of farm wastes such as sludges and manure and practices such as straw incorporation, extensification, zero till and woodland. The latter group of management practices offers by far the most significant opportunity to increase C sequestration. This further emphasises the importance of plant growth in contributing to C inputs to soil and highlights an area in which further research is required. However, measurements of total organic matter are not always the best indicators of change. The results from a long-term rotational experiment in which

arable crops were grown after a grassland phase have shown as predicted by organic matter models, that total organic matter does not change significantly between phases of the rotation (Helfrich, 2003). However, the particulate organic matter increases significantly during the grassland phase. It was found that the inter-aggregate particulate organic matter fraction increases significantly during the grassland phase. This is important because the location of organic matter is likely to influence its longevity in the soil environment (Hassink et al., 1997). Comparison of ley/arable rotations with all arable rotations in Norwegian long-term cropping experiments showed the beneficial effect of a high proportion of ley (4 years out of 6) on soil organic C (Singh et al., 1998). One of the difficulties of predicting long-term change in C sequestration is in understanding the potentially additive effects associated with changes to more than one practice. Grant et al. (2001) demonstrated that combinations of practices which, for example, reduce residue removal and tillage and at the same time increase manure application can have additive effects as a result of providing N and augmenting soil organic C. However, there are also potential negative feedbacks. Thus reduced tillage operations as discussed above can lead to increases in C storage in soils. But at the same time tillage can also influence the production of other greenhouse gases, in particular N2O (Ball et al., 1999). The increased loss of N2O which has been observed under reduced tillage operations has been linked to lower gas diffusivity and a consequent reduction in aeration (Skiba et al., 2002). In developing future farming systems and in evaluating the effects of alternative farming practices such as organic agriculture, these combined effects must be quantified. Few studies have evaluated different farming systems in terms of their ability to sequester C. New land management practices designed to engineer changes to our atmosphere must take into account the importance of environmental and edaphic factors in influencing C storage, such as texture (Hassink et al., 1997), temperature and water availability (Pendall et al., 2003). For example, a review of field trials on the effects of no-till compared with conventional till for Australian environments highlights that a positive effect of C sequestration was only present in the wetter areas

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(Chan et al., 2003). This is in contrast to the studies of tillage reported above where no-till generally had a positive effect on C sequestration. Environmental and edaphic factors are also likely to influence the timescale over which C sequestration rates continue following changes in management practice. There will also be a relationship with the management practice itself. West and Post (2002) estimated that following a change in crop rotation, a new equilibrium was not reached for 40 to 60 years. The same authors estimate changes resulting from a move from conventional to no-till practices has a peak C sequestration after 5 to 10 years and continues for a further 10 to 15 years. There is clearly a difficulty with these estimates given the short time scale of many experiments. An added problem is that once equilibrium has been reached, the new soil C level will drop quickly as soon as the best managed practices are no longer maintained (Sauerbeck, 2001). Most agricultural practices are associated with hidden C costs and it is important to assess net rather than gross rates of C sequestration (Schlesinger, 1999). Some practices such as reduced cultivation therefore become more attractive as they have the added benefit of reducing CO2 emissions from fuel use. Indirect energy costs are usually substantially lower in organic than in conventional farming systems (Shepherd et al., 2003). It is also important to note that some management practices are likely to have a negative impact on C sequestration and these will include practices that encourage over grazing, noncontour ploughing, fallow and other practices which allow soil erosion to take place.

5. Conclusions The overall balance of terrestrial C in Europe presents a very diverse picture. Some of the most recent estimates (Janssens et al., 2003) suggest that although there is a modest increase in C storage within European soils (111 Tg C year 1) this is a consequence of a relatively large accumulation of C in forests (+377 Tg C year 1), accompanied by losses of C from agriculture ( 199 Tg C year 1) and peatlands ( 67 Tg C year 1). There are relatively few good experimental data that can be used to validate these

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model predictions. One of the best sets of experimental data was reported recently in a study in Belgium (Sleutel et al., 2003). Their measurements of organic C decline during the 1990s from agricultural soils in Belgium match very closely to the model’s predictions of C loss from the soils. These studies indicate that despite the potential of agricultural soils to sequester C, a major challenge remains to move land management practices in new directions. The studies reported in this paper describing NEE of C emphasise the importance that agricultural systems can contribute to increased C storage. In developing new strategies to sequester soil C we must take advantage of our existing knowledge of soil processes and use the methods described in this review to obtain accurate estimates of C input and loss. We know that a wide range of substrates are released from plants at different times during their life cycle, but up until now, this knowledge has not been used to design management systems with the intention of building soil organic matter. Our knowledge of rhizodeposition tells us that although exudation contributes a large amount of C to soils, the management of this process is unlikely to have a significant effect on C sequestration. This is because the C compounds released by exudation tend to be relatively labile and are rapidly respired by microbial communities. There is likely to be a greater opportunity to influence C sequestration by manipulation of plant root systems through choice of crop species, and by modifying the soil environment in which it grows (through tillage). There may also be some scope for selecting cereal varieties on the basis of the quality and quantity of root biomass produced in order to help build a soil C pool. The fate of C in soil is strongly influenced by the soils nutrient content and management, and attempts to modify the rates of C storage in soils need to take account of these effects. Using modified management practices there is a potential to halt or even reverse these organic matter declines. There remain areas where our knowledge is incomplete and where further research is required in order to reconcile apparently conflicting evidence. Thus our ability to quantify the amount and timing of below ground inputs is poor. The relative importance of root death and exudation in contributing to soil C inputs are also uncertain. Studies undertaken using

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minirhizotrons and some pulse labelling work would suggest that root death and senescence are major contributors to C addition to soil. However, other pulse labelling studies would indicate that the exudation and C loss that occur before root death are important. There is a further uncertainty with regards to the relative importance of root and microbial processes in contributing to soil respiration. The new techniques highlighted in this review are helping to make progress in these areas, and should reduce the uncertainties that underlie predictions of soil organic matter change in response to management options. We have argued that improved measurements of C input and loss from soils will provide us with more accurate ways of assessing changes in soil C stocks. There are a number of other reasons for making improvements to these measurements. Firstly it will aid our understanding of the mechanisms that control C storage and therefore allow us to design intelligent management systems that optimise the return of C to the soil. Secondly it will help us to develop more robust models of soil C turnover and as a result make better predictions of organic matter dynamics. Finally it will help to improve our understanding of nutrient fluxes in the soil environments much of which are driven the dynamics of the soil C pool.

Acknowledgements The authors are grateful for the invitation to present this work to a conference on Mechanisms and regulation of organic matter stabilisation in soil held in Munich, in 2003. Research funding was received from the Scottish Executive Environment, and Rural Affairs Department.

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