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Keywords:

  • biocontrol;
  • elevated CO2;
  • fungal ecology;
  • global warming;
  • soilborne pathogens;
  • soil food web

Abstract

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

Climate changes will influence soil organisms both directly (warming) and indirectly (warming and elevated CO2) via changes in quantity and quality of plant-mediated soil C inputs. Elevated atmospheric CO2 commonly stimulates flow of organic C into the soil system, increases root production and exudation, but decreases litter quality. There is little evidence that atmospheric CO2 enrichment will increase total soil organic matter content because greater C flow into soil stimulates the soil food web, often leading to equivalent increases in soil CO2 efflux. Effects of warming on C allocation belowground, on the other hand, will depend largely on the temperature optima of different plant species. Warming is likely to increase the rate of soil organic matter decomposition by stimulating soil heterotrophic respiration, although some degree of acclimatization to warming is likely. Mycorrhizal and N2-fixing relationships are generally enhanced by CO2 enrichment, but effects of warming are highly variable. Data suggest that energy flow through fungal pathways may be enhanced relative to bacterial pathways by both warming and atmospheric CO2 enrichment. Whether the shift toward fungal domination of soils will increase soilborne fungal disease occurrence in the future is still an open question. Plant heat and drought tolerance, along with resistance to pathogens in warmer and wetter soils, may be achieved, to some unknown extent, by exploitation and management of beneficial soil organisms. Further study is needed to develop a more holistic understanding of the effects of climate change on belowground processes.


Introduction

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

Large volumes of literature on the influence of global environmental changes on natural and managed terrestrial ecosystems exist. Most of this literature pertains to aboveground ecosystem components, however, whilst soil processes have received short-shrift by comparison (Balser & Wixon, 2009). Less than 3% of articles published in ecological journals are focused on belowground organisms or processes (Wardle, 2002) and only a small fraction of those address the effects of environmental changes. Soil biology is inherently difficult to study. When one considers the importance of the belowground system to global biodiversity, biogeochemistry and human welfare, however, it is difficult to understand why so few resources and so little effort has been given to understand how rising atmospheric CO2, warming, and altered precipitation patterns will influence the soil food web.

Understanding how climate change will influence soil community composition and biodiversity, organic matter decomposition dynamics and resultant patterns of nutrient cycling will be needed before it will be possible to predict the fate of global C and N cycles, as well as the functioning of natural and managed ecosystems. Development of a more thorough understanding of the effects of atmospheric CO2 enrichment and warming on the quantity, quality, and timing of plant-mediated soil C inputs may prove to be the key for answering many important questions about soil and ecosystem processes. Unravelling the environmental controls over molecular signalling in the rhizosphere between different groups of soil organisms may prove particularly important because these signals regulate both the input of organic carbon to soil and the uptake of mineral nutrients and water by plants (Phillips et al., 2003).

The effects of global environmental changes on population dynamics of pathogenic soil organisms, along with the ability of plants to resist these organisms, require study (Scherm & Coakley, 2003; Garrett et al., 2006; Chakraborty et al., 2008). Soil pathogens already play important roles in ecosystem processes and account for significant crop losses, and global changes have the potential to influence the outcome of these interactions. Diseases caused by soilborne pathogens can often be more difficult to control than diseases that afflict aboveground structures because they are difficult to detect before serious damage occurs, and control measures are often impractical.

The prospect of controlling soil diseases in the future is complicated by the looming imperative of converting industrial approaches to agriculture and forestry to something more sustainable, whilst maintaining or increasing the global harvest. This is an important issue that will require a better grasp of soil ecology to resolve. The objective of this review is to highlight the function and importance of the soil food web and to discuss how soil organisms might be influenced by environmental changes, including rising atmospheric CO2 and global warming. Finally, I will comment on how soil management practices might be exploited to counteract potential deleterious effects of climate change on belowground processes in managed plant systems.

Structure and function of the soil food web

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

Soil biodiversity exceeds biodiversity of aboveground systems (Wardle, 2002). For simplicity, soil organisms are often subdivided into groups based on their size or functional roles within soil food webs (Fig. 1). Groups commonly include the macrofauna or soil engineers (e.g. earthworms and termites), mesofauna (microarthropods such as mites and springtails), microfauna (nematodes and protozoans) and microflora (bacteria and fungi). Microflora are the primary consumers in the soil food web and they play a key role in decomposing and metabolizing plant-derived organic substrates and mineralizing nutrients. Plant-derived energy flows from microflora to the microfauna or mesofauna that feed on them, such as amoebae, which may in turn be consumed by predaceous nematodes, which are often eaten by mites. Soil engineers play key roles in shaping soil structure at larger spatial scales. Some microflora, such as mycorrhizal fungi, rhizobia, phosphorus solubilizing bacteria and free-living N2-fixing organisms, probably exert the most proximal influence over plant physiology and growth and, therefore, total ecosystem productivity.

image

Figure 1.  Simplified representation of the soil food web, highlighting the major groups of organisms. Black arrows indicate flow of energy (to and from all blue boxes) or indicate an effect of one process or property on another (to and from all aboveground components and orange boxes). Warming of both air and soil will affect virtually all above- and belowground organisms to some extent; some organisms will acclimatize, others will not. The most important effects of environmental changes on the function of the soil food web will follow from changes in the quality and quantity of the plant-derived resource base. These changes could cascade through multiple trophic levels in soil with mostly unknown consequences. Effects of rising CO2 concentration and warming on soil organisms will also be mediated by other atmospheric inputs to the soil food web, including water and N deposition. Figure modified from Lavelle (1995) and Wardle (2002).

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Obviously, lumping soil organisms into size categories or into distinct functional groups, as in Figure 1, is somewhat arbitrary, and significant overlap in function can exist among these groupings. Furthermore, a great diversity of feeding habits is often observed within taxonomic groups and, therefore, closely related species may be playing a significant role at multiple trophic levels (Lavelle, 1995; Briones et al., 2009). For example, among 36 taxa of orbatid mites, several different modes of nutrition were observed, including carnivory or scavenging, fungivory and phycophagy, and primary and secondary decomposition (Schneider et al., 2004). A similar diversity of feeding guilds within a taxonomic group has been reported for other soil organisms, such as collembolans (Berg et al., 2004) and nematodes (Yeates & Newton, 2009).

Relationships of plants with soil endosymbionts, such as rhizobia and mycorrhizal fungi, are particularly sensitive to changes in plant physiology and belowground allocation. Mycorrhizae provide many benefits to plants, including acquisition of nutrients from outside the root zone, uptake of nutrients in soil pores too small for roots to penetrate, uptake of nutrients in forms unavailable for direct plant absorption, and improvements to plant disease resistance (Lambers et al., 2009). There are roughly 20 000 legume species that form symbiotic relationships with rhizobia and this group of bacteria is responsible for approximately a third to a half of plant available N that enters managed forests and agroecosystems (Rogers et al., 2009).

It is important to recognize that symbiotic relationships between plants and many soil microbes, including mycorrhizae and rhizobia, exist along a continuum that spans from mutualism to parasitism. The distinction between friend and foe in the soil milieu is not always clear. This balance hinges upon the relative benefits provided to a plant by a microbe, in terms of enhanced resource uptake or disease resistance, versus the energy costs incurred as a result of exudation of organic carbon substrates into the soil or through direct transfer of carbohydrates from plant to symbiont (Lambers et al., 2009). Mycorrhizal relationships, for example, can benefit plants under some environmental conditions, but decrease plant growth and fitness under a different suite of conditions (Hoeksema et al., 2010). For example, it might be predicted that ongoing atmospheric CO2 enrichment will tip the balance of the mycorrhizal relationships in favour of the plant partner because of the increase in fixed C relative to soil N and P pools. On the other hand, N deposition is likely to shift the benefit in favour of the fungal partner (Hoeksema et al., 2010). Effects of environmental changes on the status of these partnerships will also vary significantly according to plant functional group and taxonomy, so differential effects of environmental changes are likely to alter plant community composition.

It has been suggested that the soil food web should be quite resistant to environmental changes because of redundancy present within functional groups of organisms (Laakso & Setälä, 1999; Briones et al., 2009). Loss of many species, or perhaps even entire functional groups of soil organisms, can be compensated for in some soils by adjustments by other groups (Hunt & Wall, 2002; Wardle, 2002; Briones et al., 2009). The extent to which environmental changes impact a given soil organism is likely to be related to a species’ trophic position (Laakso & Setälä, 1999) or to a soil organism’s life history strategy (i.e. fast-growing r or slow-growing K strategist; Drigo et al., 2008). In some systems, effects could be greatest on organisms occupying the lower trophic levels in soils, such as bacteria and fungi, and these impacts could prove more important for large-scale function of plant systems than direct impacts of environmental changes on top predators (Fig. 1). However, in ecosystems that are regulated more strongly from the top down instead of from the bottom up (i.e. by predation rather than by soil input of plant-derived organic matter, for instance), this may not be the case (Bonkowski, 2004). Variability in how soil food webs are regulated probably explains why responses of both microflora and higher-level consumers to increasing net primary productivity (NPP) can be positive, negative or neutral (reviewed by Wardle, 2002).

Molecular signalling in the rhizosphere controls the nature of relationships between plants and other soil organisms. For instance, legume-derived signals, such as betaines and isoflavonoids, function by chemo-attracting rhizobia and trigger the events that lead to endosymbiosis. Some plants subjected to phosphate limitations produce a group of sesquiterpene lactones called strigolactones that induce growth and branching of arbuscular mycorrhiza (AM) fungal hyphae (Lambers et al., 2009). On the other hand, many rhizosphere bacteria are capable of altering plant root growth patterns by synthesizing the plant hormones auxins and cytokinins. By identifying these sorts of molecular control points, i.e. keystone signalling pathways that have the greatest influence on the soil food web, the probable effects of environmental changes on belowground processes can begin to be tested empirically (Phillips et al., 2003).

Soil carbon storage

The world’s soils contain 3·3 times more C than is contained in the atmosphere and 4·5 times more than the biotic pool (Lal, 2004). Understanding how environmental changes will influence the transfer of C into soil and the biogeochemical fate of plant-derived soil C is key to understanding whether soil will function as a C sink or source in the future.

Carbon is transferred into soil largely through deposition of canopy litter and via rhizodeposition. Rhizodeposition is the movement of organic matter or molecules from roots into the soil and includes sloughing of root cap and cortical cells, turnover of fine roots, and exudation of simple sugars, organic acids, amino acids, polysaccharides and proteins. Periodic replacement of the smallest-diameter roots, particularly in perennial plants, represents an important component of rhizodeposition, and is therefore a key aspect of soil C cycling, because the most distal root orders (1–3) have high N contents and short lifespans (Guo et al., 2008a; Pritchard & Strand, 2008; Strand et al., 2008). Root turnover is metabolically expensive and may account for 30% of global terrestrial NPP (Jackson et al., 1997). Exudation of organic compounds by fine roots may consume an additional 0·5–20% of net ecosystem C assimilation (Farrar et al., 2003; Frank & Groffman, 2009).

Fine roots are also the site of mycorrhizal colonization; in some cases up to 10–20% of the total yearly carbohydrate budget of plants is allocated to support maintenance and growth of fungal symbionts (Johnson & Gehring, 2007). Direct or indirect transfer of organic C from plant to extraradical mycelia is also an important component of soil C processes because mycelia account for up to 20–30% of soil microbial biomass and perhaps 15% of total soil C in some ecosystems (Leake et al., 2004). Godbold et al. (2006) reported that turnover of the external mycelia of mycorrhizae accounted for 62% of C transferred into pools of soil organic matter, significantly more than derived from leaf litter or fine-root turnover combined.

Fine roots, along with their symbionts, are therefore responsible for a significant input of organic C and N into soil, where much of it is eventually made available to other soil organisms. In spite of their importance, little is known about basic fine-root biology, including how long they live and how their rates of production, mortality and turnover are influenced by environmental conditions (Pendall et al., 2004; Pritchard & Strand, 2008; Strand et al., 2008). In fact, published estimates of fine-root longevity vary from a few months to many years, depending on the methodology used to quantify turnover rate (Hendricks et al., 2006; Guo et al., 2008b). Clearly, more reliable estimates of fine-root turnover rates of different species and rates of turnover under different environmental conditions will be required to better understand the potential for environmental changes to affect the soil food web along with the fate of soil C pools (Zak et al., 2000).

Rising atmospheric CO2 and soil biology

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

Plants supply the majority of energy that flows into soil food webs in the form of canopy litter, root necromass and the organic compounds that leak from leaves or are exuded from living roots. Environmental changes, particularly the increase in fixed C availability accompanying increasing atmospheric CO2 concentration, will fundamentally impact the downward flow of energy and the economy of relationships between plant roots and other soil inhabitants (Drigo et al., 2010) (Fig. 1). These changes could cascade through multiple trophic levels in soil with unpredictable consequences for plant metabolism, plant competition, direction and rate of succession, above- and belowground community composition, ecosystem NPP, and agricultural yields.

Root growth and organic carbon flow into soil

The increasing concentration of CO2 in the atmosphere is unlikely to directly affect soil organisms, because the magnitude of projected atmospheric increases are dwarfed by the concentration of CO2 already present in pore spaces of most soils. Plants exposed to CO2 enrichment do, however, realize a significant increase in photosynthesis and growth and, often, C allocation to belowground processes is stimulated to a greater extent than that to aboveground processes. Unequal carbon distribution patterns often result in an increase in both root growth and the root to shoot ratio (Rogers et al., 1996; Pritchard & Rogers, 2000). For instance, growth in CO2-enriched atmospheres resulted in large increases in root production in conifers (Crookshanks et al., 1998; Janssens et al., 1998; Pritchard et al., 2008a), deciduous trees (Crookshanks et al., 1998; Lukac et al., 2003; Norby et al., 2004), grasses (Fitter et al., 1997; Rillig et al., 1999; Milchunas et al., 2005) and field crops (Rogers et al., 1992b; Wechsung et al., 1995; Pritchard et al., 2006). Reports of unchanged or decreased root production in CO2-enriched environments also exist, but are uncommon (Arnone et al., 2000; Johnson et al., 2006; Brown et al., 2007). Greater root production indicates a larger plant-derived resource base available in soil to support a potentially larger belowground community (Fig. 1).

The influence of CO2 enrichment on root turnover rates (i.e. longevity), also critical for controlling the rate of organic C movement into the soil, is either highly variable, or poorly characterized. The prevailing view is that CO2 enrichment will increase fine-root longevity (Eissenstat et al., 2000). Increased longevity might be expected to follow from the increase in carbohydrates available to support root maintenance and protection, for instance, or from the reduction in tissue N concentration in fine roots grown under CO2-enriched conditions. Furthermore, exposure to atmospheric CO2 enrichment has also been shown to increase fine-root diameters (Rogers et al., 1992a; Milchunas et al., 2005) and to preferentially stimulate fine-root proliferation in deeper soil (Norby et al., 2004); root diameter, soil depth and carbohydrate content all correlate positively with fine-root longevity (e.g. Guo et al., 2008b). Empirical data do not support this hypothesis, however, since CO2 enrichment has been shown to increase (Arnone et al., 2000; Johnson et al., 2000; Milchunas et al., 2005) or decrease (Fitter et al., 1997; Thomas et al., 1999; Pritchard et al., 2008a) fine-root longevity with about equal frequency. It is unknown if this variability is an artefact of the methods used to measure fine-root lifespans, or if effects of CO2 on root lifespans simply vary among species or with biotic or abiotic soil conditions. This is a topic that deserves further study.

Changes in root architecture are also found in plants growing with CO2 enrichment. Herbaceous species growing with fertilization often increase root branching and exploration of shallow soil layers, whilst large trees have been shown to preferentially construct roots in deeper soil horizons (Pritchard & Rogers, 2000; Norby et al., 2004; Pritchard et al., 2008a; Iversen, 2010). The shift in rooting depths that is likely to accompany rising atmospheric CO2 could prove important for soil organisms because it will affect spatial patterns of plant-mediated organic C flow into soil. This is noteworthy because of the dependence of the soil food web on plant-derived C coupled with vertical niche separation found among members of the belowground community (Wallander et al., 2004).

In addition to changing root developmental patterns, plants grown in CO2-enriched atmospheres generally release more labile organic carbon compounds into the rhizosphere and construct tissues with higher C:N ratios, increased lignification and higher concentrations of total phenolic compounds (Drigo et al., 2008). The increase in labile exudates has the potential to stimulate (i.e. ‘prime’) the soil food web, resulting in greater microbial biomass or activity, higher rates of organic matter decomposition, increased N mineralization rates and loss of soil C (Carney et al., 2007). Furthermore, there is also evidence that soil macrofauna (detritivores) will need to consume more litter produced by CO2-enriched plants in order to meet their dietary nutrient requirements. Faster breakdown of litter by macrofauna could also contribute to faster loss of soil C (Coûteaux et al., 1991; Hättenschwiler et al., 1999; Wardle, 2002).

On the other hand, increased recalcitrance of litter and root detritus, as a consequence of higher C:N and increased lignification, can also lead to immobilization of soil N in CO2-enriched environments thereby inducing nutritional limitations to microbial metabolism which might slow organic matter decomposition rates (Hu et al., 2001; Drigo et al., 2008). Although there is support for both of these scenarios in the extant literature (Zak et al., 2000; Pendall et al., 2004; Lukac et al., 2010), it appears that the first is more likely, since soil respiration often increases in CO2-enriched plant systems, sometimes to a greater extent than soil autotrophic respiration is stimulated. Zak et al. (2000) summarized the results of 47 separate reports on the effects of atmospheric CO2 enrichment (typically ambient + 360 p.p.m. CO2) and concluded that total soil respiration increased by an average of 51% for grasses, 49% for herbaceous dicots, and 42% for woody plants. These findings largely explain the observation that most long-term experiments have been unable to detect significant changes in soil C accumulation in CO2-enriched plant systems (Pendall et al., 2004).

Implications of altered soil C inputs under elevated CO2 for soil organisms

Whilst it is clear that CO2 enrichment will increase flow of organic substrates into the soil, rising atmospheric CO2 concentrations are unlikely to stimulate all groups of soil organisms equally, and therefore changes in soil community composition are likely to occur. Higher C:N of litter and exudation of more organic C compounds, for instance, could favour soil fungi over bacteria, since bacteria generally have higher N requirements than fungi. A shift from bacterial- to fungal-dominated soil food webs in CO2-enriched plant systems has, in fact, been observed a number of times, and may prove to be a general response (Rillig et al., 1998; Drigo et al., 2007, 2008). A shift in energy flow from bacterial- to fungal-mediated pathways has important implications for belowground processes, although the outcome of such a change for long-term ecosystem function or crop yields has not been elucidated.

Effects of atmospheric CO2 enrichment on mycorrhizal fungi are often profound. A meta-analysis recently indicated that exposure to elevated CO2 (typically double ambient concentrations) elicited a 47% average increase in mycorrhizal abundance and that mycorrhizae were stimulated disproportionately more than roots (percentage colonization increased by more than 30%; Treseder, 2004). Another meta-analysis reported that the mass of ectomycorrhizal (EM) fungi increased by 34% in CO2-enriched environments, whilst that of AM (endomycorrhizal) fungi increased by 21% (Alberton et al., 2005). However, the extent to which mycorrhizae benefit from atmospheric CO2 enrichment will probably depend heavily on the P and N status of both the plant and the symbiont (Treseder & Allen, 2002; Treseder, 2004). In general, plant carbon investment to support symbionts is inversely related to soil nutrient availability. Therefore, the increase in proportional allocation to soil fungi in plants growing in CO2-enriched atmospheres is attributed to the combination of higher photosynthesis along with progressive nutrient limitations that sometimes accompany greater plant biomass production.

It has also been suggested that CO2 enrichment could select for species or genotypes of fungi and bacteria that are more efficient in acquiring mineral nutrients over species that are less beneficial. This was observed, for example, by Gamper et al. (2005), who showed that 8 years of atmospheric CO2 enrichment improved the N nutrition of white clover (Trifolium repens) as a result of selection for the most beneficial AM fungal strains combined with stimulation of N2 fixation in root nodules. Exposure of tree seedlings to elevated atmospheric CO2 led to changes in species composition of ectomycorrhizae in Betula papyrifera (Godbold & Berntson, 1997) and changed the abundance of specific mycorrhizal taxa in large Pinus taeda trees grown with FACE (free air CO2 enrichment) (Parent et al., 2006). Other reports also indicate significant changes in species composition of both EM (Fransson et al., 2001) and AM (Wolf et al., 2003; Drigo et al., 2010) communities. Also intriguing is the suggestion that turnover rate of roots colonized by less beneficial fungal structures may be accelerated relative to turnover rates of roots associated with more beneficial structures (Hoeksema & Kummel, 2003).

Growth, percentage colonization and turnover rates of mycorrhizae will probably influence how long a positive effect of elevated CO2 will be realized by natural plant systems before disequilibrium between availability of mineral nutrients and fixed C negates the CO2-fertilization effect. For example, faster turnover of mycorrhizae, should it occur, might compensate for limitations in size of ecosystem nutrient pools by increasing the rate of nutrient recycling between soil organic matter and plants (Treseder & Allen, 2002; Godbold et al., 2006). In a loblolly pine forest exposed to FACE for 8 years, CO2 enrichment reduced the lifespan of ectomycorrhizal root tips in CO2-enriched plots in deep soil (15–30 cm deep), but increased survivorship in shallower soil (0–15 cm) (Pritchard et al., 2008b). Rhizomorph turnover was accelerated in shallow soil, but effects on survivorship in deep soil varied according to the diameter of the fungal structure. Authors attributed these depth- and diameter-dependent effects to vertical changes in C allocation patterns and altered fungal community structure in CO2-enriched plots (Parent et al., 2006). Little more is known about how environmental changes will influence the turnover rate of fungi.

Unfortunately, most studies on mycorrhizal responses to atmospheric CO2 enrichment were conducted following relatively short exposure durations (months to a few years) on young containerized trees grown in monoculture (Rillig & Allen, 1999; Staddon et al., 2002), primarily in glasshouses or open-top field chambers (OTCs). In many of these studies, data were collected at only a single point in time, thereby ignoring potential for seasonal and yearly variation (Rillig & Allen, 1999). Whilst such an approach has undoubtedly yielded useful generalities regarding the influence of plant C supply on fungal symbionts, it probably falls short of capturing long-term responses of soil biological processes, particularly within intact forest ecosystems dominated by large trees and subject to seasonal and interannual variation in climate (Cairney & Meharg, 1999; Rillig & Allen, 1999; Rillig et al., 2002; Rillig, 2004; Pritchard et al., 2008b). This approach also fails to capture effects of CO2 enrichment on plant–mycorrhizal relationships that may accrue in agricultural fields over several growing cycles.

It is currently unclear how rising atmospheric CO2 will affect fungal and bacterial soilborne diseases because there are simply too few data on this topic. Both increases and decreases in soilborne diseases might be predicted based on documented effects of elevated CO2 on plant and canopy processes. For instance, increased mycorrhizal symbioses, as discussed above, might improve resistance to soilborne, as well as other diseases (van der Putten, 2009). This idea was recently supported by Drigo et al. (2009), who found that the density of Fusarium spp., a genus that includes common root pathogens, was reduced in the rhizosphere of a mycorrhizal plant exposed to elevated CO2. The authors suggested that this effect may have been caused by increased competitive ability of the mycorrhizal fungus relative to Fusarium under CO2-enriched conditions. Alternatively, decreased Fusarium density may have also been caused by the positive effects of CO2 enrichment on Trichoderma, a beneficial fungal genus well known for its ability to suppress plant-disease-causing fungi (Hagn et al., 2007), or to the significant effect of CO2 enrichment on bacterial-derived antibiotics also observed. In a recent study on FACE-grown soybean, there was no effect of CO2 or O3 treatments on sudden death syndrome caused by the soil pathogen F. virguliforme (Eastburn et al., 2010); it should be noted that foliar, not root, symptoms were used for diagnostic purposes in that study. Furthermore, CO2 enrichment could alleviate soil disease pressure through greater biosynthesis of carbon-based secondary metabolites, such as phenolics (Bezemer & Jones, 1998), or as a consequence of reduced root N concentrations (Thompson et al., 1993). Tomato plants grown at high CO2 concentrations (700 p.p.m.) were more tolerant of root rot caused by Phytophthora parasitica, an effect the authors attributed to increased transcription or post-translational turnover of pathogenesis-related proteins in conjunction with higher photosynthesis and water use efficiency (Jwa & Walling, 2001). On the other hand, root exudation might stimulate the growth of pathogens in the rhizosphere. Such an effect might explain the increase in cotton root infestation by Rhizoctonia solani when grown under FACE (Runion et al., 1994). Clearly, more work is needed to elucidate the likely impacts of elevated CO2 on soilborne pathogens.

Although rising atmospheric CO2 is likely to affect soil fungi to a greater extent than bacteria, a number of experiments indicate there could be increases in total soil bacterial biomass in addition to functionally significant shifts in community composition (Drigo et al., 2008). For example, the rhizosphere colonizers Burkholdia and Pseudomonas were both stimulated by atmospheric CO2 enrichment, whilst Bacillus species found primarily in bulk soil were not (Drigo et al., 2009). Endosymbiotic N2-fixing bacteria, such as rhizobia, also benefit from plant growth in high CO2, and often, N2-fixing plant species respond more strongly to atmospheric CO2 enrichment than non-N-fixers (Rogers et al., 2009). Increases in rhizosphere rhizobial populations (Schortemeyer et al., 1996), population size of N-fixing organisms in bulk soil (Allen, 1990), plus increases in the size and number of root nodules (Serraj et al., 1998), are all common effects of atmospheric CO2 enrichment. Whilst stimulation of N fixation by CO2 enrichment almost always occurs in managed plant systems that receive supplemental fertilization, the response of rhizobia in natural systems are sometimes dampened by multiple resource limitations (van Groenigen et al., 2006; Rogers et al., 2009).

Changes in root deployment and C allocation patterns caused by CO2 enrichment may lead to different outcomes over different time scales. For instance, effects of CO2 enrichment (600 p.p.m.) on microbial biomass and community structure associated with perennial ryegrass (Lolium perenne) changed during the course of an 8-year FACE experiment. Exposure to CO2 enrichment increased the population of Pseudomonas spp. after 3 years (Marilley et al., 1999), but no effect of CO2 enrichment was noted after 8 years (Tarnawski et al., 2006). The latter study did, however, find that populations of Pseudomonas spp. that secrete Fe-chelating siderophores were stimulated by exposure to long-term CO2 enrichment. This is significant because the disease-suppressive nature of many rhizobacteria is sometimes attributed to their ability to secrete siderophores. These molecules apparently complex with Fe3+, leading to inhibition of competing rhizosphere organisms that have a low affinity for iron. For example, siderophore production by rhizobacteria was linked to decreases in plant susceptibility to the fungal pathogens that cause rice blast disease (Naureen et al., 2009) and fusarium wilt (de Boer et al., 2003).

Significant effects of CO2 enrichment on other soil organisms, such as nematodes and collembolans, have also been observed. Atmospheric CO2 enrichment (475 p.p.m.) increased root-feeding nematode (Longidorus elongatus) populations in a sheep-grazed pasture under FACE by over 400% after 4 years (Yeates et al., 2003) and more than 3·5-fold after 9 years (Yeates & Newton, 2009). Similarly, at the 9-year harvest, predatory nematodes were doubled and a slight increase in microbial-feeding nematodes was also observed. Interestingly, a strong increase in root-feeding nematodes apparently reduced root biomass by decreasing root longevity from 58 to 38 days (Allard et al., 2005). Other studies have found significant, but more modest effects of CO2 enrichment on nematode populations (Neher et al., 2004; Drigo et al., 2007; Ayres et al., 2008; Li et al., 2009). Exposure of sugar beet and wheat to FACE (550 p.p.m.) for 5 years increased total diversity of collembolans and changed community composition according to the food preference of collembolan groups, reflecting an alteration in the quality of organic C inputs to soil under elevated CO2 (Sticht et al., 2008).

Strong interactive effects of CO2 enrichment with soil N availability on plant growth and ecosystem NPP are often noted (Luo et al., 2004). For example, CO2 enrichment increased the abundance of total soil nematodes, along with omnivores/carnivores, in a wheat field exposed to FACE for 3 years (ambient + 200 p.p.m. CO2), but only in the low-N fertility plots (Li et al., 2009). In general, CO2 enrichment tends to exacerbate plant nutrient limitations by accelerating depletion of available N, which often results in a higher tissue C:N ratio and increased flow of organic carbon through roots into the soil. The stimulation of belowground allocation in CO2-enriched environments may, therefore, be greatest in soils with suboptimal N availability. Although enhanced root growth and stimulation of the soil food web may alleviate N deficiencies in the short term, a increase in the soil C:N ratio over time as enhanced organic C input is unmatched by equivalent increases in mineral nutrient availability is projected to lead to progressive nutrient limitations (Luo et al., 2004). Faster weathering of parent rock material, along with increased rates of biological N fixation in soil, could delay or eliminate progressive nutrient limitations, although this idea is not yet well supported by data (Finzi et al., 2007).

Global warming, altered precipitation and soil biology

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

It can be predicted with some confidence that soils will warm in the future if CO2 emissions are allowed to continue unabated. However, projecting how global and regional precipitation patterns, and soil moisture content, will change in the future is far more tenuous. Some soils will get wetter, others drier. On average, it is thought that soil water content will increase in the future as a result of changes in global hydrology that cause more intense rainfall events. Unfortunately, a general understanding of the effects of soil moisture on carbohydrate allocation belowground to roots and associated rhizosphere organisms has eluded researchers. For example, studies conducted across moisture gradients have shown increases (Santantonio & Herman, 1985; Mitchell et al., 1999), decreases (Comeau & Kimmon, 1989) or no change (Gower et al., 1992; Hendricks et al., 2006) in the amount of C allocated belowground with increasing moisture availability. These results question the popular view that allocation of C belowground is inversely related to soil moisture availability. Predicting the effects of altered soil moisture on the soil food web is, therefore, confounded by the unpredictability of future precipitation, equivocal reports on the influence of soil moisture on belowground plant allocation, and a lack of knowledge about how changes in evapotranspiration caused by environmental changes will influence average soil moisture content in addition to the frequency of wetting–drying cycles. In general, wetter soils will favour a high microbial diversity, rapid microbial turnover rates, high rates of decomposition, and more robust populations of soil meso- and macrofauna compared to dry soils (Wardle, 2002). Many soilborne diseases also tend to be more problematic in wet than in dry soil (Bhatti & Kraft, 1992).

The effects of CO2 enrichment on plant systems follow primarily from changes in the function of a single enzyme, rubisco. By contrast, virtually all enzymatic and physiological processes of plants, ectothermic animals, and microorganisms are sensitive to ambient temperatures, and therefore climate change will probably influence nearly all belowground processes to some unknown extent, with unpredictable long-term consequences. Effects of warming on plant C assimilation, C allocation patterns, growth and development, and biotic interactions are all well documented (Pritchard & Amthor, 2005). The indirect effects of warming on belowground function that follow from changes in canopy processes are likely to be even greater than direct effects of warming on individual soil organisms. Warming, for instance, tends to favour some plant functional groups over others (for example C4 over C3 plants) and changes in species composition in natural communities will have substantial effects on the quality and quantity of plant-derived soil organic matter inputs. Furthermore, as discussed below, warming in the order of 3°C will result in poleward dispersal of many plant species, or even entire communities, and these changes in species assemblages could also prove more important for soil organisms than direct effects of warming (Wardle, 2002). To further complicate matters, warming will interact strongly with other environmental changes, including atmospheric CO2 enrichment, N deposition and changes in precipitation (Pritchard & Amthor, 2005).

Nevertheless, it is clear that rising temperatures will affect roots and their belowground associates, often in unpredictable ways. The effects of soil warming on root growth and turnover, and therefore on interdependent soil organisms, will vary according to temperature optima of individual plant species. Plants that are currently growing in soils with suboptimal temperatures will probably exhibit increases in root growth with warming, whilst root development of plants that are currently growing in soils at optimal or supraoptimal temperatures will be adversely affected (McMichael & Burke, 1998). As soil temperatures increase, maintenance respiration of roots will probably increase in the majority of plant species, and this could decrease the optimal longevity of fine roots, causing them to turn over more rapidly (Eissenstat et al., 2000). By this mechanism, warming could accelerate fine-root-derived C input to the soil (Eissenstat & Yanai, 1997). A global meta-analysis suggested that root turnover increases from boreal to tropical regions (Gill & Jackson, 2000), an observation the authors attributed to increases in maintenance respiration, more rapid mineralization rates, and higher soil pathogen activity in warmer soils. A number of other studies support the hypothesis that root longevity is inversely related to soil temperature (e.g. Forbes et al., 1997; King et al., 1999; Graefe et al., 2008).

Soil warming usually stimulates soil microbial activity, net nitrification rates, P and N mineralization rates, and total respiration in soil (Andresen et al., 2010), although microbial responses to warming are sometimes transient (Balser et al., 2006). Acclimatization of the soil microbial community to experimental warming may result from a wholesale change in the composition of the soil food web, which may in turn reduce temperature sensitivity of the soil community as a whole (Balser et al., 2006). Other possible mechanisms to explain acclimatization have been proposed (French et al., 2009), including changes in substrate availability that occur over longer time periods (Fierer et al., 2005; Davidson & Janssens, 2006), altered litter quality resulting from plant growth under warmer conditions, or physiological adjustments within the soil organisms themselves (Heinemeyer et al., 2006). Understanding how heterotrophic respiration will respond to soil warming is key to understanding whether the soil system will function as a C sink or source in the future.

Ultimately, the future of the soil C pool will be determined by the balance of carbon influx and efflux (autotrophic plus heterotrophic respiration rates). If, for example, warming fails to stimulate heterotrophic respiration, then the soil could serve as a net sink for C in the future, thereby slowing warming rates (assuming some increase in NPP as a result of higher atmospheric CO2 concentrations and/or warming). On the other hand, if soil organic matter decomposition is accelerated by environmental changes relative to the rate of plant-mediated soil C inputs, then significant quantities of soil C could be transferred into the atmosphere, causing a positive feedback to further climatic change (van der Putten, 2009). Alternatively, increased rates of respiration over long time periods could increase N mineralization rates, thereby stimulating plant nutrient uptake and ecosystem productivity and leading to more storage of C in plant canopies. The issue of soil heterotrophic respiratory responses to environmental changes, and the implications of faster or slower respiration for soil C storage, need to be resolved (Billings et al., 2010).

Warming can increase colonization of plants by mycorrhizae directly by stimulating fungal growth and indirectly by stimulating the growth of the host plant (Heinemeyer et al., 2006). For instance, Gavito et al. (2005) found that warmer temperatures led to more extraradical hyphal growth and increased uptake of labile C substrates, and that different species responded similarly but with different temperature optima (see also Heinemeyer & Fitter, 2004). These investigators suggested that extraradical fungal structures may be more sensitive to soil warming than fungal structures directly associated with the host plant. Soil warming also increased the extraradical mycelium of the AM fungus Glomus mosseae associated with Plantago lanceolata (Heinemeyer et al., 2006). Similarly, warming accelerated the rate of transfer of photosynthate from host plants to AM fungi associated with Plantago lanceolata and also caused a shift from storage structures (vesicles) to hyphal growth (Hawkes et al., 2008). Notably, in this latter study, effects of warming on AM growth and respiration were independent of plant size and photosynthetic rate.

Soil organisms other than bacteria and fungi may also prove sensitive to warming, although few studies have been done on larger organisms occupying higher trophic levels, and those that have show highly variable results. Briones et al. (2009) conducted a 2-year mesocosm experiment to evaluate the effects of a 3·5°C increase on a grassland soil community. Although temperature did not affect aboveground plant growth or soil respiration, warming stimulated root growth and significantly affected soil organisms. A total disappearance of epigeic worm species (i.e. surface-dwelling earthworms), a decrease in the larger oligochaetes and Prostigmata mites, and movement of enchytraeids (small worms that feed on bacteria and fungi) to deeper soil were observed, as well as an increase in fungivorous mites, perhaps reflecting the fact that warming apparently caused a shift from a bacterial- toward a fungal-driven food web in this experiment, a result that has been suggested by others (Wardle, 2002). Briones et al. (2009) concluded that warming is likely to have a profound effect on the heterotrophic soil community, thereby increasing C turnover rates, decreasing soil organic matter (SOM) content, and increasing N mineralization rates. Harte et al. (1996) reported increased biomass of meso- and macrofauna in moist soils subjected to warming in a subalpine meadow, whilst warming decreased the biomass of these same groups in dry soils. A meta-analysis found that enchytraeid worm populations will decline precipitously if soils rise above the threshold temperature of 16°C (Briones et al., 2007). These results clearly indicate significant effects of warming on larger soil organisms, but also highlight the unpredictable nature of these responses.

As discussed in more detail by Eastburn et al. (2011) an increase in ambient temperatures will affect pathogen physiology and development rate, in addition to the growth rate and physiology of resistance mechanisms in plant hosts. Temperature optima exist for dispersal and germination of pathogen spores, development of infections in/on host plants, development and maintenance of dormancy, and sporulation (Garrett et al., 2006; Pariaud et al., 2009). For example, the root rot pathogen Monosporascus cannonballus was shown to reproduce more rapidly at higher temperatures (Waugh et al., 2003). Pathogen latency periods are also controlled largely by degree-days and are therefore expected to change with warming (Pariaud et al., 2009). Temperature changes will also influence quantitative aspects of soil disease progression, such as pathogen aggressiveness (i.e. spore release rate, lesion area, percentage of plant tissue infected), which, along with efficacy of resistance mechanisms on the part of the host plant, will combine to control the outcome of plant diseases in terms of reduced plant fitness or decreased agricultural yield. For example, pathogen-isolate × environment interaction accounted for 29% of the variance in disease severity and 19% of the variance for mycotoxin production by Fusarium graminearum (Cumagun & Miedaner, 2004). Rapid rates of environmental change might result in environmental effects having even greater importance in determining disease severity.

The effects of temperature on plant–pathogen interactions will be influenced by the previously described effects of elevated atmospheric CO2 enrichment on plant C and N dynamics. Reductions in tissue N concentrations, increased lignification and increased availability of C skeletons for construction of defence mechanisms may improve disease resistance in CO2-enriched environments and could cancel out potential increases in diseases that might be expected to accompany warmer and wetter conditions in a changed climate. Currently, these sorts of interactive effects are difficult to predict and, thus far, have not been studied for soil pathogens.

It is important to note that warming will influence soil food webs through an increase in annual mean temperatures, via warmer winter temperatures, or by increasing the frequency of temperature extremes in the soil environment. Whilst an average increase in soil temperatures of the magnitude predicted may have only modest effects on soil biology, temperature extremes could have much greater effects by causing periods of acute stress for both plants and soil organisms. However, slightly warmer soil conditions during winter periods may allow novel pathogens to overwinter and colonize new habitats. Furthermore, effects of higher temperatures on soil food webs will probably depend upon whether the new climate is outside the range of the conditions that a given soil system has evolved under (Waldrop & Firestone, 2006). Soil organisms and communities that are adapted to withstand large fluctuations in temperature and moisture availability under the current climate are unlikely to be impacted significantly by climate changes compared to soil systems that have evolved under more stable environmental conditions.

The potential impact of warming on horizontal gene transfer (HGT) between different rhizosphere bacterial species needs to be considered. HGT in the soil can occur through conjugation, transduction or transformation. Conjugation involves transfer of plasmid DNA from one bacterium to another, transduction is transfer of DNA via a bacteriophage, and transformation is direct uptake of extracellular DNA by bacteria. Genes that commonly move between bacteria by these mechanisms are involved in conferring antibiotic resistance, adherence, nutrient uptake and production of toxins, and may influence microbial pathogenicity or mycorrhizal relationships (Frey-Klett et al., 2007; French et al., 2009). These mechanisms of bacterial evolution are influenced by soil temperatures. For example, Lafuente et al. (1996) studied gene transfer between Escherichia coli and Rhizobium meliloti and found that conjugation frequency between these two bacteria increased by an order of magnitude when temperature was increased from 20 to 30°C, in spite of the fact that bacterial biomass remained unaffected. They concluded that above a certain temperature threshold (which probably varies among bacterial species and strains), HGT events would double for every 1°C increase in temperature. Similarly, the frequency of HGT between E. coli and Pseudomonas putida strains increased linearly as temperature was increased in a laboratory study from 5 to 29°C and then declined at temperatures above 35°C (Johnsen & Kroer, 2007). Extreme high and low temperatures tend to decrease the frequency of HGT. HGT events involving movement of antibiotic resistance marker genes from transgenic plants to soil bacteria have been reported (Heinemann & Traavik, 2004), although the transfer of genes from plants to soil fungi is considered to be rare (Richards et al., 2009). Soil warming is likely to increase the rate of evolution of soil microorganisms, especially those with shortest regeneration times, relatively to longer-lived plant species and soil macrofauna.

Invasive organisms and the soil food web

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

The prevailing view is that environmental changes including rising atmospheric CO2 concentration, warming, and altered precipitation patterns, will increase the frequency of plant invasions and exacerbate the negative effects of these events on the environment. Poleward dispersal of plant species, including invasive weeds, for example, is likely to be widespread (Kriticos et al., 2003). As discussed by Wolfe & Klironomos (2005), immigration of plant species into novel habitats may alter the function of native soil food webs by (i) changing the quality, quantity and timing of litter input and rhizodeposition; (ii) causing direct release of novel antimicrobial compounds; (iii) by significantly altering nutrient relations by introducing alternate modes of nutrient acquisition, such as N fixation; or (iv) by altering soil structure or physical properties as a result of novel or dominant rooting habits.

Recent reports indicate that exotic plants are already altering the distribution and success of various beneficial and plant-pathogenic fungi by releasing toxins into the soil or by changing patterns of decomposition as a result of altered quality of detritus (van der Putten et al., 2009). For example, diffuse knapweed (Centaurea diffusa), a plant that has invaded areas of western North America, releases the antimicrobial compound 8-hydroxyquinoline and garlic mustard (Alliaria petiolata) releases glucosinilates from roots which suppress AM fungi (Roberts & Anderson, 2001). Another important example of an invasive plant and soil microbe invading a novel habitat and fundamentally changing ecosystem function involves the firetree (Myrica faya) in Hawaii. This N-fixing plant has invaded N-limited ecosystems, altering soil nutrient dynamics, plant species composition, and presumably function of the soil food web as well (Vitousek & Walker, 1989).

The effects of altered geographic dispersal of organisms caused by environmental changes could prove especially important because immigrants are often favoured over established native species (Chejara et al., 2010). The success of invasive species is sometimes attributed to alleviation of disease pressure from root pathogens in the new habitat (the ‘enemy release hypothesis’; van Grunsven et al., 2007; MacKay & Kotanen, 2008) or to greater benefits conferred by symbionts such as AM fungi in newly colonized habitats (van Grunsven et al., 2010). Evidently, local soil pathogens decrease the performance and success of native species relatively more than they decrease performance of invasive plants (Eppinga et al., 2006; Engelkes et al., 2008).

Dispersal of organisms into new habitats, as discussed above, could result in shifts in the abundance of certain taxonomic groups in soil, or elimination of groups altogether. Such effects could influence ecosystem diversity and function and could make both natural and managed plant systems more susceptible to invasive pathogens or may decrease belowground competition, allowing pathogenic soil organisms to flourish. However, the functional redundancy of soil taxonomic groups common to many soil communities may buffer against significant shifts in the productivity of natural and managed plant systems. In other words, significant changes to different groups of soil organisms may cancel each other out such that soil community shifts will have no net effect on ecosystem productivity (Bradford et al., 2002).

In addition to indirect effects on soil organisms caused by invasive plant species, environmental changes may also directly increase the potential for dispersal of soil organisms (Coleman, 2008; Pariaud et al., 2009). The oomycete Phytophthora cinnamomi, an invasive root pathogen endemic to New Guinea-Sulawesi, for example, has already invaded at least 76 countries, as well as the majority of global biodiversity hotspots (Myers et al., 2000). Phytophthora cinnamomi has caused massive die-off of Quercus spp. and Castanea in the USA and Eucalyptus spp. in Australia over significant geographic areas (Davidson & Shearer, 1989; Dunstan et al., 2010). Models indicate that soil warming associated with climate change is likely to increase the range of P. cinnamomi in both Europe (Bergot et al., 2004) and Australia (Podger et al., 1990). Warming of soil is also likely to cause similar directional dispersal of plant parasitic nematodes, earthworms and many other soil organisms (Boag et al., 1991; Ghini et al., 2008).

Will climate change impact the synchrony of leaf, root and soil microbial physiology?

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

Reports of accelerated leaf and flower phenology in response to climate change, and speculation about how plant and herbivore phenology may become uncoupled under climate-change scenarios, are numerous, but no data are currently available on the effects of climatic variability on the synchrony of plants and soil organisms. In fact, recent reviews on the effects of climate change on plant phenology do not even mention roots or other soil inhabitants (Cleland et al., 2007; Steinaker & Wilson, 2008). This is a potentially significant omission, as carbohydrate production by plant canopies, C allocation belowground, root production and physiology, and composition and function of soil microbial communities, are linked, both in time and space. It is likely that vegetation has coevolved with the soil food web such that some level of synchrony occurs between plant phenology and soil microbial processes. Developmental mismatches between roots and soil microorganisms could fundamentally change ecosystem processes via shifts in biogeochemical cycling and timing of availability of mineral nutrients to plants (Körner & Basler, 2010).

Synchronicity of plants and soil organisms may be of particular importance for disease cycles where pathogenic organisms have evolved to require the presence of a host plant at a developmental stage favourable for infection at a particular time. Plant belowground allocation and rhizodeposition are closely tied to canopy processes such as shoot developmental rate. And the increase in developmental rate of plants that accompanies warming will have corresponding effects on temporal patterns of belowground C allocation and root dynamics. To further complicate matters, developmental rates of soilborne root parasites are also sensitive to temperature and soil moisture. Developmental or phenological mismatches caused by environmental changes may shift the disease cycle in favour of either the pathogen or the plant, or may enable some harmful soil organisms to complete additional life cycles during the growing season, as observed for multivoltine insect species (Ghini et al., 2008).

Soil management

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

It is argued that current industrial agricultural practices that rely on large tracts of monocultures, frequent tillage and excessive and misapplied inorganic fertilizers, cannot be sustained. Accordingly, large areas of land are being managed with more sustainable practices such as no-till (i.e. direct sowing), incorporation of cover crops, planting more genetically diverse assemblages of crops, frequent crop rotation, organic fertility programmes and integrated pest management. Tillage and residue management practices affect soil characteristics such as water holding capacity, soil physical properties, temperature and microbial activity. For example, the presence of soil-surface litter residue coupled with no-till management has repeatedly been shown to enhance soil moisture absorption and retention compared to conventional tillage with no residue. Roots growing under no-till conditions are often found in shallower soil than in tilled soil (Cheng et al., 1990; Rasse & Smucker, 1998).

Although there are many documented benefits of no-till agriculture, including decreased erosion, improved soil infiltration and retention of rainfall, and increased C storage, direct sowing may exacerbate crop losses to some soilborne diseases. Lack of soil disturbance and retention of crop residue on the soil surface may serve as an inoculum source and may create temperature and moisture conditions suitable for pathogen survival. This appears to the be case in the US Pacific Northwest and in parts of Australia where the soil diseases rhizoctonia root rot and bare patch (Rhizoctonia solani and R. oryzae), pythium damping-off and root rot (Pythium spp.), take-all (Gaeumannomyces tritici) and fusarium foot rot (Fusarium pseudograminearum and Fusarium culmorum) are more problematic in fields managed with no-till than in those with conventional tillage (Paulitz et al., 2002). In many instances, the differences in soil biological and physical properties that result from soil management practices may be of a greater magnitude than effects of climate changes (Canadell et al., 1996; Paustian et al., 1996). This was the case in a 3-year study in which incorporation of plant residue, had a greater effect on nematode abundance and diversity than exposure to high atmospheric CO2 concentration in FACE-grown wheat and sugar beet, for instance (Li et al., 2009). Clearly, research on the potential influence of global climate changes on soil ecology must take interactive effects of soil management practices into consideration.

Use of beneficial soil microorganisms

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

As understanding of plant–microbe interactions in the soil has improved, it has become possible to exploit specific beneficial organisms in order to induce faster plant growth, higher crop and timber yields, and improved disease resistance (Gentili & Jumpponen, 2006). In most cases, beneficial microorganisms improve plant growth by (i) secreting plant hormones such as auxins and cytokinins that directly affect plant development; (ii) inhibiting the growth and success of disease-causing microorganisms in the soil; (iii) inducing systemic acquired resistance in plants; (iv) converting N into forms that are available for plant growth; or (v) solubilizing phosphorus and other nutrients, thereby increasing their availability for plant uptake. In a number of cases, the mode of action of beneficial soil microorganisms remains unknown.

Not only can they be used to improve plant growth, nutrition and disease resistance, but soil microorganisms have more recently been shown to increase plant drought resistance and thermotolerance. In a dramatic example, the fungus Curvularia protuberata conferred tolerance to high root-zone temperatures in the tropical panic grass Dichanthelium lanuginosum found in geothermal soils, an effect attributed to the presence of a thermotolerant virus present within the fungus (Redman et al., 2002; Márquez et al., 2007). Interestingly, investigators were also able to improve heat tolerance of tomato plants by inoculating them with C. protuberata in an experimental system (Márquez et al., 2007). Similarly, the plant growth promoting rhizobacterium (PGPR) Pseudomonas sp. strain AKM-P6 significantly enhanced thermotolerance of sorghum seedlings exposed to high temperatures by improving plant physiological resistance mechanisms (Ali et al., 2009). The presence of heat- and drought-resistant species of PGPR, such as Pseudomonas and Paenibacillus species, holds great promise for the application of beneficial soil organisms to improve thermal and drought tolerance of field crops in the future (Timmusk & Wagner, 1999; Srivastava et al., 2008; Ali et al., 2009). Yang et al. (2010) recently reported that the beneficial fungus Trichoderma harzianum, transformed with the gene encoding the antioxidant superoxide dismutase (SOD), was significantly more resistant to both heat and salinity stress. These data suggest that beneficial soil organisms currently sensitive to high temperatures can be manipulated genetically to improve their efficacy under stressful conditions.

Aside from inoculating field crops with beneficial microorganisms to mitigate climate-change effects, another option is to improve the efficacy of plant/microbe symbioses such as mycorrhizae and N-fixing endophytic bacteria through genetic modification. In many cases, however, this approach will require more detailed understanding of the genetic mechanisms involved in the development of the symbiosis, in both the host plant and the symbiont. Some progress has already been made in this area. For instance, inserting additional copies of nodulation genes in Sinorhizobium meliloti enhanced nodulation and increased N fixation in alfalfa (Castillo et al., 1999) and also improved nodulation by Rhizobium tropici on its host legume Macroptilium atropurpureum (Mavingui et al., 1997). Similarly, inserting additional copies of the ribBA gene (involved in riboflavin biosynthesis) into S. meliloti accelerated the nodulation process in alfalfa (Yang et al., 2002). Biocontrol activity of free-living soil organisms such as Pseudomonas fluorescens has also been improved through genetic modification (Hongyou et al., 2005).

Capacity for movement of DNA between soil bacteria should be considered when developing strategies involving beneficial microorganisms, particularly those strategies that involve genetic modifications to improve performance of known beneficial microorganisms. Transfer of genes that increase bacterial fitness from introduced to pathogenic species could have significant implications for soilborne plant diseases, including the increase in pathogen virulence and aggressiveness. HGT could also enable pathogens to exploit additional niches, both within the heterogeneous soil environment and also with respect to the range of compatible host genotypes.

Understanding and managing the trade-offs between frequent applications of chemical fertilizers and biocides, and the efficacy of symbiotic soil organisms for improving plant performance, particularly in the context of environmental change, could be key to maximizing plant performance and minimizing losses to soilborne diseases in the future (van der Putten et al., 2009). Clearly, approaches to developing crop genotypes that focus solely on maximizing yield potential under idealized conditions and that fail to value traits that confer resistance to soil pathogens, as well as relationships with beneficial microbes, need to be rethought. It is now obvious that solving fertility and disease problems with excessive application of chemical fertilizers and antimicrobial compounds cannot be continued indefinitely. Crop development through genetic improvements and conventional breeding should therefore merge with integrated crop management practices involving soil organisms, such as integrated pest management (IPM), and efforts to develop crop varieties and crop management strategies should take into account the changes in temperature regimes and atmospheric CO2 concentrations projected for the future.

Conclusions

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References

Characterizing the processes that govern soil biology, and then understanding how soil organisms are likely to be influenced by global changes, is a daunting prospect. Any environmental change that alters soil nutrient availability or that significantly changes the trophic structure of the belowground food web could influence ecosystem function or crop productivity. Changes in precipitation and warming will affect the soil food web directly and will also influence soil organisms indirectly by altering plant growth, structure and physiology. Although atmospheric CO2 enrichment will not directly affect soil organisms, it will stimulate canopy photosynthesis and increase the amount and change the quality of aboveground litter inputs and rhizodeposition.

The influence of climate changes on natural or managed ecosystem soil processes, including disease pressure, is likely to be inversely related to species diversity present within successive trophic levels. Effects of environmental changes will also vary depending upon whether a given soil food web is regulated mainly from the top down or from the bottom up (Tylianakis et al., 2008).

Applying emerging genetic techniques holds great promise for understanding how soil organisms will respond to global changes (Roesch et al., 2007; Chakraborty et al., 2008). Sequencing entire rhizosphere communities in conjunction with the transcriptome of the associated root holds great promise for linking plant physiological status to rhizosphere activities and community dynamics. This will be a complicated and expensive undertaking because the physiology of roots varies substantially by order (Guo et al., 2008a, b; Pritchard & Strand, 2008) and because quantity and quality of rhizodeposits, and resulting effects on other soil organisms, may vary significantly during the lifespan of a root (Pritchard & Rogers, 2000). One might imagine that the rhizosphere community may undergo succession on a microscopic scale as a root differentially alters the chemistry of the rhizosphere during its ontogeny. Effects of the environment on such processes are therefore a moving target. It seems clear that understanding these processes will require collaboration among plant physiologists, soil ecologists, geneticists and bioinformaticists.

Utilization of beneficial soil organisms, along with improving plant disease resistance through direct genetic modification, are fast-moving fields that hold great promise for ameliorating negative effects of soilborne pathogens on plants in the future, and also for decreasing dependence upon chemical fertilizers and pesticides. Introducing novel genes to crops that boost resistance to microbial root pathogens is one potential way forward. Biocontrol and biofertilization strategies, such as inoculation of crops with beneficial rhizosphere organisms or selecting crop varieties that form beneficial relationships with endosymbionts such as N-fixing bacteria and mycorrhizal fungi, are another. There is significant scope for improving plant mutualisms through genetic transformations, provided the genetic basis for these relationships can be untangled. It should be stressed that development and field application of genetically modified plant-beneficial rhizosphere microorganisms should proceed with great care in light of the frequent HGT that can occur belowground and the potential for soil warming and increased C input to accelerate these rates of transfer (Snow et al., 2005; Johnsen & Kroer, 2007).

It is unclear on what time-scales changes in soil organism functioning must be studied to get a satisfactory view of what future environmental changes portend for interrelated soil and plant processes. Variable effects of environmental changes on different members of soil food webs may be manifested over different time-scales and so short-term effects of environmental changes (as are observed in most simulation experiments) on a given trophic level may change in direction and magnitude over longer time periods. Studies at the Duke FACTS-1 FACE site in North Carolina (USA), where minirhizotrons have been used to monitor dynamics of fine roots, mycorrhizae and rhizomorphs for over a decade, suggests that effects of drought on monthly production and mortality of both fungi and roots during a given year can result in legacy effects on standing crop that linger for several years thereafter (unpublished data; see also Arnone et al., 2008). Inherently slow processes that are mediated by lower trophic levels, such as transfer of C from plant litter into recalcitrant soil organic matter pools (i.e. those that persist for decades or longer), will be particularly difficult to predict.

Obviously, experimental approaches to understanding the effects of warming and associated environmental changes on soil organisms must become more realistic in terms of complexity and duration (Klironomos et al., 2005; Tylianakis et al., 2008). Most of the results discussed in this paper were obtained through controlled laboratory experiments or simplified mesocosm approaches. In the future, it is the interactive effects of multiple simultaneous changes in the field that matter, not the effects of a single environmental change in isolation, but no experimental apparatus has yet been designed that is capable of studying the interactive effects of variation in precipitation, warming, elevated atmospheric CO2, and soil N availability on the soil system, especially within intact forests characterized by large trees (Kimball et al., 2008; Kimball & Conley, 2009). Although attempts to heat vegetation or soil without heating the air have been made in field settings (Ineson et al., 1998; Shen & Harte, 2000), these approaches effectively uncouple atmospheric and soil conditions (Kimball et al., 2008). These sorts of manipulative experiments raise questions about the validity of extrapolating to field conditions. I advocate the initiation of a new generation of climate-change experiments designed to study interactive effects of multiple climate-change factors over timescales of decades or longer.

References

  1. Top of page
  2. Abstract
  3. Introduction
  4. Structure and function of the soil food web
  5. Rising atmospheric CO2 and soil biology
  6. Global warming, altered precipitation and soil biology
  7. Invasive organisms and the soil food web
  8. Will climate change impact the synchrony of leaf, root and soil microbial physiology?
  9. Soil management
  10. Use of beneficial soil microorganisms
  11. Conclusions
  12. Acknowledgements
  13. References