Corresponding author: Present address: S. McCormack, Centre for Ecology and Hydrology, Bush Estate, Penicuik, Midlothian, EH26 0QB, UK, tel. +44 (0) 131 445 4343, fax +44 (0) 131 445 3943, e-mail: firstname.lastname@example.org
Biochar amendment of soil and bioenergy cropping are two eco-engineering strategies at the forefront of attempts to offset anthropogenic carbon dioxide (CO2) emissions. Both utilize the ability of plants to assimilate atmospheric CO2, and are thus intrinsically linked with soil processes. Research to date has shown that biochar and bioenergy cropping change both aboveground and belowground carbon cycling and soil fertility. Little is known, however, about the form and function of soil food webs in these altered ecosystems, or of the consequences of biodiversity changes at higher trophic levels for soil carbon sequestration. Hitherto studies on this topic have been chiefly observational, and often report contrasting results, thus adding little mechanistic understanding of biochar and bioenergy cropping impacts on soil organisms and linked ecosystem processes. This means it is difficult to predict, or control for, changes in biotic carbon cycling arising from biochar and bioenergy cropping. In this study we explore the potential mechanisms by which soil communities might be affected by biochar, particularly in soils which support bioenergy cropping. We outline the abiotic (soil quality-mediated) and biotic (plant- and microbe-mediated) shifts in the soil environment, and implications for the abundance, diversity, and composition of soil faunal communities. We offer recommendations for promoting biologically diverse, fertile soil via biochar use in bioenergy crop systems, accompanied by specific future research priorities.
Rising atmospheric carbon dioxide (CO2) concentrations have created the need to reduce fossil fuel reliance and enhance natural carbon stocks (IPCC, 2007). Various geo- and eco-engineering strategies have been proposed to augment terrestrial carbon, some of which focus on carbon uptake by plants and soil (Galaz, 2012). Globally, these biological carbon reservoirs contain approximately 2.7 times as much carbon as the atmosphere (Batjes, 1996), and there is great potential to enhance these stores via targeted land management (Fowles, 2006). Biochar and bioenergy production are two such management methods at the forefront of scientific efforts to offset anthropogenic carbon emissions. Both methods capitalize on the ability of plants to assimilate atmospheric CO2 and aim, directly or indirectly, to augment soil carbon stocks (Lehmann, 2007; Sutherland et al., 2009).
Biochar is the relatively recalcitrant, carbon-rich residue produced by pyrolysis, where organic matter is heated in an oxygen-depleted environment (Lehmann, 2007). This drastically increases its resistance to oxidation, creating a stock of slow-cycling organic carbon which may be incorporated into soil for long-term storage (Sombroek et al., 2003). Biochar can potentially improve soil nutrient retention and crop yield in highly leached tropical soils (Lehmann & Rondon, 2006), and possibly reduce turnover of existing soil carbon (Liang et al., 2010). However, biochar addition can also stimulate the mineralization, and hence loss of humus in organic forest soils (Wardle et al., 2008), and reduce the productivity of certain plant species, such as perennial ryegrass (Jeffery et al., 2011). In addition to ecosystem-specific effects, biochar impacts depend on its physical and chemical characteristics, which are largely determined by feedstock type (e.g. agricultural or silvicultural residues, livestock excreta or sewage) and production conditions (heating temperature, atmosphere and duration) (Sohi et al., 2009). Biochar application rates typically range from 1 to >100 t ha−1, which can influence the magnitude of biochar impacts on crop productivity (Jeffery et al., 2011). Biochar effects on soils and soil fauna are therefore largely dependent on biochar production and land management.
There is an intrinsic link between biochar and bioenergy production: biochar can be produced from bioenergy crop residues, bioenergy can be generated during biochar production and biochar can be applied to fields in which bioenergy crops are grown (Gaunt & Lehmann, 2008; Sohi et al., 2009). Scientific interest in their combined use is currently increasing due to the potential for gains in soil carbon sequestration (Tallis, 2010). Consequently, bioenergy cropping and biochar soil amendments represent, individually and in combination, a land-use change with direct and indirect consequences for the functioning of soil.
Bioenergy cropping aims to reduce fossil fuel carbon release by producing a biomass-based fuel generated by photosynthetic carbon capture. Bioenergy cropping may also improve soil carbon storage, particularly in carbon-depleted ex-arable soils, where conversion from previously intensive land uses leads to increased primary productivity and reduced rates of carbon turnover (Blanco-Canqui, 2010). Specifically, bioenergy cropped soils may experience increases in organic carbon content due to minimum tillage practices, multiyear harvest cycles, high litter inputs, and establishment of more extensive rooting systems (Lemus & Lal, 2005).
Although some research has been undertaken on the response of the microbes (primarily bacteria and fungi) to biochar and bioenergy cropping (e.g. Hamer et al., 2004; Steinbeiss et al., 2009), the effects on eukaryotic soil biota (e.g. protozoa, nematode, and enchytraeid worms, microarthropods and earthworms) are yet to be systematically studied (Lehmann et al., 2011; Rowe et al., 2009). These higher trophic levels are integral to the function of soil food webs, which facilitate decomposition of organic matter and belowground nutrient cycling (Bardgett, 2005; Nielsen et al., 2011). Trophic interactions of soil fauna regulate microbial community structure and activity, thereby influencing decomposition rates, nutrient cycling and ultimately plant productivity (Nielsen et al., 2011). The feeding and burrowing of larger soil animals (e.g. earthworms, arthropods) further alters the soil physical environment and modulates resource availability (Bardgett, 2005). Thus, trophic interactions between microbes and soil fauna, together with direct soil engineering by animals, are vital to the stabilization and sequestration of soil organic carbon (Bardgett, 2005; Seastedt, 1984).
Soil fauna is sensitive to environmental changes such as land management and pollution, and thus are an important determinant of soil functional responses to perturbation (Yeates & Bongers, 1999). Biochar and bioenergy cropping aimed at increased soil carbon sequestration may have perverse outcomes by causing unanticipated and adverse changes to soil communities and linked ecosystem functions. Such changes include priming of soil carbon, reduced plant productivity and lower soil bioturbation, aggregation and nutrient cycling. This review focuses on the impact of biochar deployed in those perennial bioenergy crop systems, namely short-rotation coppice (SRC) and perennial grasses (Miscanthus and switchgrass), which represent the greatest divergence from current agricultural management. Specifically, we synthesize the literature to identify the main mechanisms by which the cascading effects of biochar in bioenergy cropping scenarios can affect decomposition pathways in soil via changes in soil communities. Due to the limited research on soil faunal responses to these land uses, we focus on the known effects of biochar and bioenergy cropping on abiotic and biotic soil properties, and extend this to predict the consequences for soil fauna.
Impacts of biochar and bioenergy crops on abiotic soil properties
Structure, porosity and moisture
The addition of highly porous biochar to soil is likely to alter the soil's porosity and structure. Soil structure supports invertebrate diversity and abundance by provision of fine-scale habitat heterogeneity and retention of nutrients and moisture (Bardgett, 2005). Soil aggregates create pore spaces that retain air and water and protect soil organic carbon from physical decomposition (Jones & Donnelly, 2004). Pore spaces thereby provide niches for soil microfauna (protozoa, tardigrades and nematodes) and mesofauna (mites, collembola, enchytraeid worms), offering refugia from desiccation and access to resources (Bardgett, 2005; Neher et al., 1999). Although larger soil organisms such as earthworms and termites can create channels within the soil (Jouquet et al., 2006), the movement of most micro- and mesofauna is limited by pore neck diameter. For example, microbes can occupy pores that are only a few microns wide, whereas nematodes can be up to 100 microns in diameter and accordingly occupy larger pores (Neher, 2010). Changes to soil porosity may thus affect soil trophic interactions; for example, by creating spatial refuges for microbes (Ogawa, 1994).
Both biochar addition and bioenergy crop establishment require tillage, the former for incorporation into the soil, and the latter for planting and weed suppression. Tillage reduces the diversity and abundance of microarthropods and earthworms by disrupting the soil physical structure and resource availability (House & Parmalee, 1985; Lal, 1988; Siepel, 1996). Invertebrate regulation of decomposition is lessened by tilling, as microbial-feeders benefit from the presence of a dense microbial population within a continuous litter layer (Stinner & House, 1990), and because meso- and macrofauna do not survive well under high levels of physical disturbance (Beare, 1997). Microarthropod abundance is commonly observed to decline under traditional tillage practices (Siepel, 1996; Stinner et al., 1988), which could reduce the rate of soil nutrient cycling and decomposition (Cole et al., 2004). Earthworm abundance also declines under high levels of physical disturbance, due to their large size and susceptibility to desiccation (Lal, 1988). This can result in further losses of soil carbon as aggregate formation is reduced (Lal, 1988). Therefore, the initial effect of biochar addition and bioenergy crop establishment on soil fauna may be detrimental, particularly in previously untilled soil. However, subsequent to tillage, biochar can increase aggregate formation by increasing plant productivity, microbial activity and by binding to other soil constituents (Rondon et al., 2004; Steiner et al., 2008).
Perennial biofuel crops (e.g. Miscanthus, willow and poplar) only require tillage prior to crop establishment, after which point the soil may be left untilled for up to 30 years, depending on crop type (Hilton, 2002). This can significantly improve soils that have previously undergone annual tillage for agricultural purposes. The undisturbed, year-round litter layer created by perennial crops in reduced-tillage sites creates a heterogeneous, resource-rich habitat for soil fauna. Reduced tillage promotes extensive root systems and fungal mycelial networks, which support both plant-feeding and fungivorous soil fauna (Bardgett, 2005; Neher, 1999), promoting soil aggregation and physical protection of soil carbon (Jones & Donnelly, 2004).
However, minimum-till systems can become compacted over time, and management practices requiring heavy machinery (e.g. harvesting and fertilizer or pesticide application) can increase soil compaction (Kort et al., 1998; Makeschin, 1994). Soil compaction is detrimental to most soil invertebrates, as it reduces habitable pore spaces and access to plant roots, fungal hyphae and water (Whalley et al., 1995). The frequency of heavy machinery use may be reduced by biochar, as its adsorption of nutrients may lessen fertilizer requirements (Lehmann, 2007). The low bulk density and high porosity of biochar could further diminish the extent of soil compaction (Sohi et al., 2009). Soil fauna, particularly meso- and macrofauna, may therefore benefit from the combined implementation of biochar and perennial bioenergy cropping in a reduced-tillage system. However, biochar is gradually broken down into fine fractions in the soil (Sohi et al., 2009), such that its effects on soil bulk density may lessen over time. The overall response of fauna to soil structural changes resulting from bioenergy cropping and biochar additions are therefore likely to depend on time since crop establishment.
Biochar may further alter the physical nature of the soil by reducing soil albedo (Genesio et al., 2012). A decrease in soil albedo will increase soil temperatures, particularly in equatorial latitudes, leading to decreased soil moisture content (Genesio et al., 2012; Oguntunde et al., 2008). In contrast, perennial bioenergy cropping can increase surface albedo via the provision of a dense canopy, understorey and litter layer (Georgescu et al., 2011). The effect of biochar addition on soil albedo and soil organisms will, therefore, depend on vegetation, soil type and climate. The soil's ability to retain moisture both prevents faunal desiccation and provides a nutrient-rich soil solution supporting microbial and plant productivity (Bardgett, 2005). Albedo-driven changes to soil moisture are likely to affect soil fauna differentially due to species- and taxon-specific moisture requirements. For example, nematodes rely on soil water films for movement, access to microbial prey and as a refuge from desiccation (Bardgett, 2005; Ettema, 1998), whereas microarthropods with chitinous exoskeletons are more effective at modulating water loss (Killham, 1994). Albedo-induced reduction in nematode and microarthropod numbers may decrease the regulation of microbial activity by their grazers, enhancing microbial abundance while reducing nutrient turnover rates (Bardgett, 2005). Similarly, earthworm abundance and activity declines with aridity (Lee, 1985), whereas enchytraeid worms, less affected by water stress, can thrive where earthworm abundance is reduced (Topoliantz et al., 2000). Drought is, however, detrimental to both earthworms and enchytraeids (Lavelle, 1988; Lindberg et al., 2002), and the reduced activity of these soil organisms is likely to decrease decomposition, soil mixing and physical stabilization of soil organic matter (Bardgett, 2005). To promote an abundant and active soil fauna, biochar-induced albedo effects should be modulated, particularly in equatorial or drought-prone regions. In these areas, bioenergy crop sites may be ideal for biochar use, as the dense vegetation canopy and litter layer would greatly reduce insolation of the soil surface.
Biochar generally has a neutral to alkaline pH, depending on feedstock, and hence may have a liming effect on most soils (Jeffery et al., 2011). However, the extent of this effect is dependent on biochar application rate and soil type; a higher application rate will produce a stronger effect on the soil, whereas crops grown on highly acidic soil respond less strongly to biochar addition than on neutral or moderately acidic soil (Jeffery et al., 2011). In contrast, bioenergy cropping can increase soil acidity (Jug et al., 1999; Makeschin, 1994), due to reduced liming inputs and nitrification-induced loss of base cations (van Miegroet & Cole, 1985). Acidity is deleterious to many soil organisms as it can increase solubility of aluminium ions to toxic concentrations, and can also be harmful to plant productivity by reducing nutrient availability (Bardgett, 2005). Liming of acid soils generally (though not always, see Fornara et al., 2011) favours bacterial, rather than fungal, decomposition, thereby often increasing nutrient mineralization and decreasing soil carbon sequestration (Fig. 1) (Bardgett, 2005). Liming has also been shown to reduce collembolan species diversity (Chagnon et al., 2001), which may be a result of declining fungal biomass. Liming can also reduce the abundance of mites (Hågvar & Amundsen, 1981), which could cause a resultant rise in their prey – microbial grazers – thereby increasing microbial activity and decomposition rate (Kajak, 1995). Neutral and alkaline soils tend to be more diverse in terms of both microbes and fauna (Edwards & Bohlen, 1996; Fierer & Jackson, 2006), but extreme liming can also be harmful to certain organisms; for instance, earthworms are generally not found in soils with a pH greater than eight (Liesch et al., 2010). Enchytraeidae thrive in acidic soils (Cole et al., 2002; Edwards & Bohlen, 1996), and microarthropods exhibit a wide range of species-specific pH optima (van Straalen & Verhoef, 1997). Biochar with a high pH can be used to mitigate the acidifying effects of bioenergy cropping, or acidic legacy of previous agricultural land use, with potential benefits for crop productivity (Jeffery et al., 2011). Biochar dose, soil and biochar pH, crop pH tolerance and bacterial mineralization of soil carbon are important considerations in this form of management. Biochar type and dose selection should aim to minimize adverse effects on soil communities by considering the expected response of soil pH in relation to natural soil conditions and other ongoing management practices that may affect pH.
Soil nutrient cycling
Both bioenergy crops and biochar can increase soil organic carbon content and soil carbon-to-nitrogen ratio (Baum et al., 2009; Sohi et al., 2009). Except for a small labile portion, biochar-derived organic carbon is mostly recalcitrant and slow cycling (Sombroek et al., 2003). In contrast, carbon inputs from bioenergy cropping (i.e. leaf litter and root exudates via the fungal mycelium) are readily accessible to microbes and detritivorous invertebrates (Godbold et al., 2006). In ex-arable soils, the year-round, lignin-rich litter layer of perennial bioenergy crops creates a continuous habitat for invertebrates (House et al., 1984; Neher, 1999). This litter layer and the dense vegetation canopy reduce belowground temperatures, thereby decreasing microbial activity in deeper soil strata (Mann & Tolbert, 2000), potentially leading to greater soil carbon stabilization (Clifton-Brown et al., 2007). Changes to the net soil carbon balance with bioenergy cropping will, however, depend in part on the previous land use. Overall carbon gains would be expected in degraded or ex-arable land, whereas carbon losses are anticipated where pasture or seminatural land is converted (Lemus & Lal, 2005). A meta-analysis of land-use change found that conversion of pasture and forest to agricultural crops reduced soil carbon by 59 and 42 per cent respectively (Guo & Gifford, 2002). In the same study, the conversion of cropped land to plantation caused an average gain in soil carbon of 18 per cent. Bioenergy cropping is therefore most beneficial in degraded, nutrient-poor soils such as ex-arable land and brownfield sites, where there is potentially the most to be gained in terms of soil carbon content and soil habitability (Lemus & Lal, 2005).
In agricultural land, the use of inorganic fertilizers is common and can reduce the abundance of collembola, enchytraeid worms, earthworms and cryptostigmatid mites, as well as overall invertebrate species richness (Siepel, 1996; Yeates et al., 1997). Species loss upon inorganic fertilizer application may be partly due to a decline in fungivores and their predators, as bacterial decomposition dominates where nutrients are readily available (Bardgett, 2005). Declines in faunal abundance and diversity (in terms of functional dissimilarity) cause a general reduction in decomposition and nutrient cycling rates, and potentially reduce plant nutrient acquisition (Heemsbergen et al., 2004; Wardle et al., 2004). Bioenergy crops are generally supplemented with inorganic fertilizer inputs, but the high nutrient uptake efficiency of most bioenergy crops means that a lower proportion of fertilizer remains in the soil or is lost via leaching in comparison with arable crops. More efficient use of inorganic fertilizers in bioenergy cropping could amplify invertebrate abundance and diversity in ex-arable land, although it is unlikely that community composition would reach seminatural levels. Biochar can improve the efficiency of soil nutrient retention via its high cation exchange capacity (CEC) (Sohi et al., 2009). The effect of biochar-mediated nutrient retention on soil fauna will likely depend on land-management practices and the amount and type of nutrient inputs applied. Adsorption of toxic ammonium ions to biochar may increase their soil residence time (Ding et al., 2010), with detrimental consequences for soil fauna at high concentrations. In contrast, biochar surface adsorption of organic fertilizers such as manure, compost and mulch may benefit certain invertebrate taxa. For instance, organic fertilizers increase the abundance of earthworms (Edwards & Lofty, 1982), collembola (Culik et al., 2002) and mites (Minor & Norton, 2004). In contrast, enchytraeid worms show a mixed response to such amendments (de Goede et al., 2003; Keeling et al., 1995).
Bioenergy cropping in contaminated ex-industrial or degraded land can be an efficient use of natural resources. While such land may not be suitable for food production, bioenergy cropping can replenish soil organic carbon stocks and reduce concentrations of soil contaminants (Hartley et al., 2009). Willow and poplar, both commonly used in SRC, are effective cover crops for phytoremediation due to their fast growth rates and high uptake of heavy metals (Baum et al., 2009), although uptake is species- and clone dependent (Granel et al., 2002; Meers et al., 2007). Inoculation of willows with mycorrhizal fungi or bacteria has been suggested as a means of further improving their metal uptake capacity (Sell et al., 2005). Heavy metals can increase earthworm mortality (Langdon et al., 1999), reduce microbial biomass and abundance of plant-feeding nematodes (Bardgett et al., 1994), and modify nematode and microarthropod communities (Russell & Alberti, 1998; Yeates & Bongers, 1999). Reductions in pollutant concentrations are therefore likely to promote more species-rich soil communities.
Due to its high CEC, biochar can also be useful for remediation of contaminated soil (Cao et al., 2009) via the immobilization of heavy metals and persistent organic pollutants (Beesley et al., 2010; Gomez-Eyles et al., 2011; Uchimiya et al., 2010). The ability of biochar to immobilize toxins depends on its surface characteristics (i.e. oxygen and volatile concentration and CEC), thus biochar production conditions may be selected to augment its capacity for soil remediation (Uchimiya et al., 2011). However, depending on feedstock, pretreatment and production, biochar may itself be a source of organic contaminants, such as polyaromatic hydrocarbons (PAHs) (Hale et al., 2012), which is an important consideration if applying biochar to vulnerable or degraded systems.
Impacts on soil food webs
Although perennial bioenergy crops are planted as a monoculture on ploughed and herbicide-treated land, their subsequent extensive management facilitates the establishment of diverse vegetation (Dimitriou et al., 2011). The high structural heterogeneity, low disturbance and fertilization, winter harvesting and multiyear harvest cycles make SRC crops particularly conducive to promoting plant diversity (Dimitriou et al., 2011). This can augment the amount and diversity of plant-derived inputs to soil, thereby potentially elevating the abundance, biomass and diversity of microbes (Matlack, 2001; Topoliantz & Ponge, 2005), soil microarthropods (Borges & Brown, 2001; Siemann et al., 1998), earthworms (Cesarz et al., 2007) and certain nematodes (Wardle et al., 2006) (Fig. 2).
In addition to high litter inputs and reduced tillage, soil carbon stocks are also augmented by the transfer of plant-assimilated carbon via the external mycorrhizal mycelium, which enhances microbial biomass and nutrient cycling (Bardgett, 2005). Extensive rooting systems favour plant-associated invertebrates and mycorrhiza provides a high-quality resource for fungal feeders while protecting plants from parasites and pathogens (Fig. 2) (Bardgett, 2005). Initial research indicates that, in some situations, biochar can enhance mycorrhizal colonization and sporulation (Saito & Marumoto, 2002; Warnock et al., 2007); hence the use of biochar in bioenergy cropping sites could further amplify root-derived carbon transfer to the soil, potentially increasing soil carbon sequestration. The effects of this enhancement on soil fauna may depend on the nature of mycorrhizal growth; grazing by fungivores may be restricted by pore neck size if fungal hyphae colonize the pore space of biochar.
Biochar has a range of indirect effects on plant productivity, depending on soil and crop characteristics, as well as biochar type (Jeffery et al., 2011). Although research on biochar application to temperate bioenergy crop sites is scarce, biochar has been found to improve the productivity of crops with potential bioenergy applications (e.g. wheat, sorghum and maize), particularly in equatorial latitudes (Jeffery et al., 2011). In more stable seminatural systems, higher productivity may be associated with amplified litter and root inputs to the soil, potentially benefiting root feeders, detritivores and fungivores in the rhizosphere (Sarathchandra et al., 2001; Wardle et al., 2004; Yeates, 1987). No effect of charcoal has been found on the productivity of SRC willow (Dimitriou et al., 2006; Park et al., 2005). The adverse effects of intensive agricultural management practices, such as frequent fertilization, tillage and harvesting, may offset any potential benefits of higher plant productivity (Bardgett, 2005). Therefore, the overall effect of biochar and bioenergy cropping on soil fauna will depend on concomitant management practices, with greatest benefits in extensively managed perennial systems.
Biochar may affect plant competitive interactions, particularly in bioenergy cropping systems with diverse understorey vegetation, by altering plant species composition with knock-on effects for plant-associated invertebrates. Biochar-induced changes to plant interspecific competition may arise not only from alterations to edaphic properties (e.g. pH, moisture) but also potentially by interfering with plant chemical signals. Although such effects of biochar have not yet been documented, activated carbon is a well-known inhibitor of plant chemical signals (Kulmatiski & Beard, 2006; Ridenour & Callaway, 2001). These secondary metabolites are exuded from roots for many purposes, including: initiating plant–microbe symbioses, acquiring nutrients and deterring pathogens or competitors (Bais et al., 2004). Immobilization of plant signals by activated carbon has been found to change plant species composition, reducing the establishment of certain invasive species which rely on allelochemicals to gain a competitive advantage over native plants (Kulmatiski & Beard, 2006). Reduced effectiveness of plant secondary metabolites may also lessen the plant's ability to establish mycorrhizal symbioses, thereby reducing plant nutrient uptake (Bais et al., 2004). Sequestration of plant defence chemicals by biochar may enhance the abundance of pathogens and plant parasites such as root-feeding nematodes, thereby increasing plant susceptibility to disease and reducing primary productivity, crop yield and litter-derived carbon inputs to the soil (Bais et al., 2006). However, the effects of biochar on belowground plant signalling and rhizosphere interactions remain untested hypotheses requiring empirical validation.
Biochar and charcoal have consistently been found to increase microbial biomass in soil across ecosystems, ranging from boreal forests (Wardle et al., 2008; Zackrisson et al., 1996) to Amazonian uplands (Liang et al., 2010; Steiner et al., 2008). Evidence from microcosm experiments suggests that an increase in microbial biomass should precipitate an increase in microbial grazer populations (e.g. nematodes, protozoans, collembola) and higher trophic levels (e.g. predatory nematodes, mites) (Cole et al., 2004). Such a shift in the soil food web will be a crucial determinant of soil carbon fluxes. Moderate grazer abundance can promote fungal growth and microbial enzyme activity, stimulating decomposition and carbon release, whereas high grazer abundance reduces microbial activity (Bardgett, 2005), potentially allowing accumulation of soil organic carbon. This increase in microbial biomass may be caused by improved soil habitability (e.g. increased pH, soil moisture, plant productivity) or retention of microbes in the soil via adsorption to biochar (Thies & Rillig, 2009). In addition, the small, but biologically significant, labile fraction of biochar (generally less than one per cent of the total biochar mass (Hamer et al., 2004)) can increase the substrate for microbial growth, although its composition and relative quantity varies widely with biochar type (Sohi et al., 2009). Labile content of biochar is affected by pyrolysis type and temperature; fast pyrolysis biochar is less aromatized and contains relatively more labile carbon than biochar created by gasification or slow pyrolysis (Brewer et al., 2011). Higher pyrolysis temperatures also correspond to lower labile content and increased aromaticity (Bruun et al., 2011). As a result, CO2 emissions from biochar-treated soil are higher in soil amended with biochar produced at a lower temperature (Bruun et al., 2011). This implies greater stimulation of microbial activity by low-temperature biochars and a greater potential for priming of preexisting soil carbon. However, the labile portion is a finite resource that will be metabolized within one or two growing seasons of biochar addition to the soil; thus, biochar is likely to exert different short- and long-term effects on the soil carbon balance (Sohi et al., 2009).
In general, bioenergy crop production (e.g. North American perennial grass and European SRC systems) seems to increase soil microbial biomass relative to intensive agricultural use (Lipps & Balser, 2011; Makeschin, 1994). This is probably the result of a combination of high crop productivity, year-round litter inputs and reduced tillage. In contrast to biochar, the positive effects of bioenergy on microbial biomass are likely to be amplified with time since crop establishment, as the soil community recovers after tillage. This longer term stability in microbial biomass would be expected to ultimately support larger populations of microbial-feeding fauna.
Microbial activity is a good indicator of grazing pressure on microbes, with higher activity denoting abundant microbial feeders and more rapid carbon turnover (Hanlon & Anderson, 1979). In a study of Amazonian Terra Preta soils (centuries-old anthrosols which received periodic inputs of charred biomass), carbon stability and microbial metabolic quotient (ratio of carbon mineralized to microbial biomass carbon) were lower in comparison with adjacent unamended soils, despite not having received fresh inputs of charred biomass for decades (Liang et al., 2010; O'Neill, 2007). This may signify that despite larger microbial biomass, microbial feeders are affected negatively by char-induced alterations to the microbial community. One possible explanation of this is the protection of microbes from larger microbivores by the fine-scale pore spaces of charred organic matter. This could have the effect of reducing the rate of soil nutrient cycling, as microbial grazers effectively release nutrients that may otherwise be immobilized in microbial biomass (Bardgett, 2005). In particular, this could result in a reduction in the availability of mineralized nitrogen, which is excreted by nematodes, protozoans and collembola (Neher, 1999) and taken up by plants.
Little research has explored the effects of bioenergy crops on soil microbial activity. However, it can be inferred that bioenergy cropping may reduce microbial respiration, based on findings of slower organic carbon turnover in Miscanthus plantation soils (Siemann et al., 1998) and accumulation of soil organic carbon under SRC crops (Jug et al., 1999; Vesterdal et al., 2002). Slower carbon turnover is associated with fungal decomposition pathways, which are common where soil physical disruption is low (Bardgett, 2005). Changes in microbial community composition will therefore have consequences for both soil carbon sequestration and higher trophic levels, as many microbial-feeding fauna specialize in either fungal or bacterial ingestion (Figs 1 and 2). Protozoans, which feed primarily on bacteria, are unlikely to benefit from fungal dominance. As protozoa are important producers of mineralized nitrogen (Hoorman, 2011), this may create a feedback of reduced nitrogen availability, further favouring fungal decomposition, which predominates where nitrogen is limiting (Bardgett, 2005). Alterations to litter quality by fungal decomposition can also change the community composition of other decomposers, such as earthworms and saprophagous microarthropods (Langley & Hungate, 2003). Whereas soil carbon sequestration and fungivore abundance could therefore increase under reduced tillage, abundance of obligate bacterial feeders may decline.
Current understanding of biochar and bioenergy cropping on soil fauna
Empirical results of soil faunal responses to biochar (and other char-like substances) and bioenergy cropping are highly variable (Tables 1and 2). This is likely due to the great complexity of, and differences among, soil ecosystems (e.g. climate, soil properties, vegetation, land management, faunal community composition and biochar type). Furthermore, much of this work has hitherto been observational and short term, and mechanistic understanding of biotic responses to these forms of land management is limited.
Table 1. A summary of studies reporting the effects of pyrolysis products on soil fauna
Soil fauna group
Location/soil type/duration (if stated)
Wood waste biochar, fast pyrolysis, 3.9 Mg ha−1
Arable clay loam, Québec, Canada, field trial, 18 months
Due to the paucity of research that has addressed biochar effects on soil invertebrates, Table 1 also contains the effects on soil fauna of other pyrolysis products, including charcoal, charred biomass, ash and kiln smoke. Pyrolysis products seem to have a mixed, but mostly negative, effect on soil fauna that directly ingest soil organic matter (e.g. earthworms and enchytraeids), whereas microbial feeders and higher trophic levels may indirectly benefit from the input of organic substrate. Biochar ingestion will be limited by particle size, with larger biochar particles, such as those produced from wood chippings, less likely to be consumed and transported or aggregated by bioturbators. With time, mean particle size will diminish as biochar undergoes physical degradation (Sohi et al., 2009), and this may increase the ingestible proportion. Biochar impacts on soil ingesters could be due to the presence of contaminants or toxins in biochar [such as PAHs, which can be produced during the charring process (Hale et al., 2012)], or due to desiccation from ingesting dry biochar (Li et al., 2011). The negative response of enchytraeids to wood ash addition (Table 1) is likely due to increased soil pH, as enchytraeid worms generally thrive where soils are too acidic to support large numbers of earthworms (Bardgett, 2005).
Most of the research on soil faunal response to agricultural biochar use has focussed on earthworms, which are often considered keystone species due to their function as ecosystem engineers and decomposers (Bardgett, 2005). Currently, studies of other taxa such as nematodes, microarthropods and enchytraeids are scarce (Lehmann et al., 2011). Much of the research involving these taxa comes from studies on wildfire-derived charcoal or ash in forest soils rather than purpose-made biochar in agricultural land (Table 1). Results from these studies are not necessarily applicable to research on synthesized biochar, as differences in char and ecosystem type are likely to be important in determining the nature of the faunal response (Table 1). The duration of exposure to char inputs is also an important factor in determining soil community responses. Although biochar inputs may elicit an initial reduction in abundance or biomass of particular soil taxa, a succession of species may occur over time as organisms adapted to the altered conditions immigrate and become established. Such potentially complex community dynamics require long-term field studies to examine the response of all functional and taxonomic groups of soil fauna to biochar application over time.
The effect of bioenergy cropping on soil fauna (Table 2) depends primarily on the land-use history and current management. Compared with annual crops perennial bioenergy crops support greater soil faunal abundance across many taxa (Bellamy et al., 2009; Felten & Emmerling, 2011; Liang et al., 2012; Makeschin, 1994), although abundances remain lower than in seminatural soils (Brand & Dunn, 1998; Felten & Emmerling, 2011). As one might expect, more intensive management of bioenergy crops (e.g. residue harvesting, herbicide application) will negate the positive impacts of bioenergy cropping on soil fauna (Bengtsson et al., 1997; Minor & Norton, 2008). Also, not all soil fauna are desirable in bioenergy cropping systems; for instance, parasitic root-feeding nematodes can be deleterious to plant productivity (Neher, 2010), and could be difficult to control in the absence of annual crop rotation. Much remains unexplained about the variety of soil faunal responses to bioenergy cropping. Earthworm abundance has, for example, been shown to increase, decrease and show no change in different bioenergy cropping systems (Bellamy et al., 2009; Coates & Say, 1999; Makeschin, 1994). Research on the response of micro- and mesofauna such as protozoa and nematodes to bioenergy cropping is currently scant (Rowe et al., 2009). As with biochar, long-term studies are vital, as the effects of bioenergy cropping on soil fauna and linked functions will likely become more pronounced over time. In ex-arable land, the reduced tillage and perennial crop cover may gradually augment phyto- and soil diversity, soil carbon stocks and soil aggregation. Conversely, in converted pasture or seminatural land, the introduction of tillage and biomass harvesting may reduce carbon stocks over time.
Towards an understanding of biochar effects on soil fauna
As shown in Tables 1 and 2, the reaction of soil fauna to land management varies across taxa and by crop, biochar and soil type. However, the cause of this variance is not well understood. Although biochar is often compared with other organic amendments, such as compost and sewage sludge, there are a number of important distinctions between biochar and uncharred organic matter, such as its long soil residence time, small biologically available fraction and potential PAH content, which necessitate targeted biochar research.
Research on the response of soil fauna to biochar in bioenergy cropping systems should aim to elucidate how biochar may be applied to avoid disruption of biotic interactions and linked processes and promote provision of ecosystem services (Fig. 1). Nutrient cycling and plant productivity in biochar-treated land will depend on the modified soil's capacity to support certain functionally important species and a functionally dissimilar community of invertebrates (Bardgett, 2005; Heemsbergen et al., 2004; Neher, 1999). Biochar avoidance or mortality tests using invertebrate species of interest may provide useful preliminary information (e.g. Liesch et al., 2010; Topoliantz & Ponge, 2003). In these assays, the critical response parameters would include organism survivorship, biomass, activity and reproductive rates. However, the extremely high biological and physical diversity of soils makes it unrealistic to expect such assays to establish a comprehensive understanding of invertebrate responses. Nonetheless, such targeted approaches could be applied in bioenergy cropping sites to ecosystem engineers (e.g. earthworms, termites) or guilds of species known to be differentially sensitive to disturbance, such as nematodes that occupy different trophic niches (e.g. microbial grazers, plant parasites and predators) (Yeates & Bongers, 1999). Avoidance tests should involve biochars with varied physiochemical (e.g. pH, particle size, pore size distribution) and production (e.g. feedstock, pyrolysis method, temperature and duration) characteristics to allow multivariate analysis to identify mechanisms underpinning positive and negative soil faunal responses to different biochar types.
Long-term field monitoring of invertebrates is another necessary, yet largely overlooked, topic of biochar research. Field trials are required to verify the findings of laboratory experiments, as biochar effects could be moderated by environmental factors absent in a controlled environment. In a meta-analysis, Jeffery et al. (2011) found that biochar effects were three times more pronounced in pot experiments compared with field trials. The temporal aspect of biochar research is also an important consideration in this context, as certain biologically relevant properties of biochar, such as labile portion, particle size and CEC, change with time (Baum et al., 2009; Sohi et al., 2009). The labile carbon portion in particular seems to have a strong, but ephemeral influence on biochar-induced changes to microbial activity, as carbon mineralization often peaks within days or weeks of biochar addition (Bruun et al., 2011). The quantity of labile soil organic carbon and the C : N ratio of the soil are markedly different where biochar has been recently applied compared with historically char-amended soils (i.e. Terra Preta soils) (Thies & Rillig, 2009). This implies that the response of soil microbes to biochar addition may change over time. However, biochar is not strictly comparable with Terra Preta soils due to differences in production conditions, (pyrolysis vs. partial burning) feedstock (diversity of feedstocks vs. hardwood), land management and char application practices (Sohi et al., 2009). It is therefore difficult to predict how the effects of biochar on soil communities may change in the long term. Artificial ageing of biochar using induced oxidation is a new, but promising means of predicting long-term effects of biochar in the soil (Sohi et al., 2009). A further complication is the practice of biochar reapplication, which would effectively combine biochar of varying ages and properties. This could have the effect of enhancing resource and habitat heterogeneity, potentially increasing invertebrate functional diversity, but also amplifying biochar-induced modifications to the soil, such as intensifying changes to pH or albedo. Research is needed to determine biochar application rates and frequencies that will optimize plant productivity and the habitability and function of the soil.
Biochar implementation strategies should consider how current land management causes the soil condition to differ from its natural state, with an aim to minimize impacts on biodiversity and soil function. Manipulation of biochar characteristics (e.g. pH, CEC, labile content, porosity, elemental composition) by selection of feedstock and production conditions is currently a topic of growing interest (Brewer et al., 2011). This could allow biochar to be “designed” to match with particular soil types (Novak et al., 2009). For instance, a high production temperature may be chosen to create a biochar low in volatiles for soils rich in organic carbon (Bruun et al., 2011), to reduce any priming effect. In nutrient-depleted or mineral soils, priming of organic carbon may be less of a concern, and lower temperature pyrolysis could be used to maximize biochar yield (Bruun et al., 2011). Nutrient-depleted soils may also benefit from sewage or poultry waste-derived biochar, which tends to be higher in labile nitrogen and phosphorus (Bruun et al., 2011). For highly acidified soils, gasification (a type of biochar production which maximizes the yield of combustible gas by-products) tends to produce the most alkaline biochar (Brewer et al., 2011). Highly porous biochars, produced by high-temperature fast pyrolysis (Zanzi et al., 1996), may be best applied to compacted soils, to produce a reduction in bulk density. Matching of biochar type to soil and crop requirements necessitates an understanding of how biochar effects on soil functions, including faunal abundance and activity. A more comprehensive understanding of the effects of biochar on soil invertebrate populations will help to inform strategic biochar selection.
Due to current rates of population growth, habitat loss and fragmentation, and challenges to food security, efficient and productive land use is of paramount consequence (Campbell et al., 2008). Bioenergy cropping is clearly of both economic and environmental importance, but must be implemented in an efficient manner that preserves or restores the biodiversity and function of soils. Use of biochar in bioenergy crop sites may have certain advantages for both productivity and belowground ecosystems, for instance by reducing soil compaction (Laird et al., 2010), improving soil structure (Lehmann et al., 2009) and promoting stabilization of organic matter (Liang et al., 2010).
In this review, we described how soil fauna contributes to critical ecosystem processes and how this ecology may be altered by biochar and bioenergy cropping impacts on food resources, niche availability and interspecific interactions. However, it is difficult to predict the effects of biochar and bioenergy cropping on soil biotic communities due to the dearth of research and the complexity of biological and physical interactions in soil. A mechanistic understanding of biochar and bioenergy cropping effects on soil communities would enable their more effective implementation, with potential benefits for soil fauna and the ecosystem functions they regulate. Namely, benefits may be gained from “designing” biochar to ameliorate soil properties which limit the productivity of a particular crop species, with minimal alterations to natural or beneficial soil properties. Research in this area could potentially enhance our ability to manage land for the provision of multiple ecosystem services: carbon sequestration, food cultivation and energy production, while supporting diverse aboveground and belowground communities.