Potential mechanisms for achieving agricultural benefits

Plant Soil (2010) 337:1–18 DOI 10.1007/s11104-010-0464-5

MARSCHNER REVIEW

Potential mechanisms for achieving agricultural benefits from biochar application to temperate soils: a review Christopher J. Atkinson & Jean D. Fitzgerald & Neil A. Hipps

Received: 17 December 2009 / Accepted: 7 June 2010 / Published online: 30 June 2010 # Springer Science+Business Media B.V. 2010

Abstract Natural organic biomass burning creates black carbon which forms a considerable proportion of the soil’s organic carbon. Due to black carbon’s aromatic structure it is recalcitrant and has the potential for long-term carbon sequestration in soil.

Soils within the Amazon-basin contain numerous sites where the ‘dark earth of the Indians’ (Terra preta de Indio, or Amazonian Dark Earths (ADE)) exist and are composed of variable quantities of highly stable organic black carbon waste (‘biochar’). The apparent high agronomic fertility of these sites, relative to tropical soils in general, has attracted interest.

Biochars can be produced by ‘baking’ organic matter under low oxygen (‘pyrolysis’). The quantities of key mineral elements within these biochars can be directly related to the levels of these components in the feedstock prior to burning. Their incorporation in soils influences soil structure, texture, porosity, particle size distribution and density. The molecular structure of biochars shows a high degree of chemical and microbial stability. A key physical feature of most biochars is their highly porous structure and large Responsible Editor: Yongguan Zhu. C. J. Atkinson (*) : J. D. Fitzgerald : N. A. Hipps East Malling Research, New Road, East Malling, Kent ME19 6BJ, UK e-mail: chris.atkinson@emr.ac.uk

surface area. This structure can provide refugia for beneficial soil micro-organisms such as mycorrhizae and bacteria, and influences the binding of important nutritive cations and anions. This binding can enhance the availability of macro-nutrients such as N and P. Other biochar soil changes include alkalisation of soil pH and increases in electrical conductivity (EC) and cation exchange capacity (CEC). Ammonium leaching has been shown to be reduced, along with N2O soil emissions. There may also be reductions in soil mechanical impedance. Terra preta soils contain a higher number of ‘operational taxonomic units’ and have highly distinctive microbial communities relative to neighbouring soils. The potential importance of biochar soil incorporation on mycorrhizal fungi has also been noted with biochar providing a physical niche devoid of fungal grazers. Improvements in soil field capacity have been recorded upon biochar additions. Evidence shows that bioavailability and plant uptake of key nutrients increases in response to biochar application, particularly when in the presence of added nutrients. Depending on the quantity of biochar added to soil significant improvements in plant productivity have been achieved, but these reports derive predominantly from studies in the tropics. As yet there is limited critical analysis of possible agricultural impacts of biochar application in temperate regions, nor on the likelihood of utilising such soils as long-term sites for carbon

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sequestration. This review aims to determine the extent to which inferences of experience mostly from tropical regions could be extrapolated to temperate soils and to suggest areas requiring study.

Keywords Biochar . Black carbon . Biochar . Carbon sequestration . Charcoal . Climate change

Introduction The burning of plant-derived biomass, whether maninduced or from natural fires, has considerable impact on the ecology of many landscapes (Alexis et al.

2007; Rovira et al. 2009). Around 40% of the earth’s land surface is modified by fire (Alexis et al. 2007), and these events leave soil surface residues which are generally low in organic carbon, as natural combustion is an oxygen rich process. The residues from these burning events are frequently described as ‘black carbon’ (Preston and Schmidt 2006). This is a generic term used to describe highly condensed material generated from the incomplete combustion of fossil fuels and organic matter (Schmidt et al. 2001; Rovira et al. 2009). It is often released into the atmosphere as particulates where it remains for a short time (at least relative to carbon dioxide) (Ramanathan and Carmichael 2008). The term ‘black carbon’ has also been used to describe this material within soil when either deposited from the atmosphere or directly from the combustion of vegetative matter. It is this ‘black carbon’ within soils that is the primary focus of this review. It is a term also used to describe the considerable quantities of carbon within urban soil (‘char’) which is derived from local sources following incomplete combustion of fossil fuel (Rawlins et al.

2008). Here however we are interested in the ‘black carbon’ produced from natural burning (biogenic carbon) that can provide a considerable proportion of the soil organic carbon (SOC); in some cases, i.e.

Australian soils, Brazilian oxisols and German chernozemic soils it can range from 30% to 45% of the total SOC (Skjemstad et al. 1996; Schmidt et al. 1999; Glaser et al. 2000).

The occurrence of natural and man-induced burning can have considerable influence on the pace and direction of ecosystem/habitat development (Tyron 1948; Ritchie 1995; Bowman 1998; Bond and Keeley 2005; Preston and Schmidt 2006; Ansley et al. 2006;

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Alexis et al. 2007). The ecological literature contains many vegetational and soil physiochemical descriptions of how both natural burning and charcoal production can induce habitat changes (Oguntunde et al. 2004; Hart et al. 2007; Oguntunde et al. 2008), including increases in plant primary production (Santos et al. 2003). Slash and burn agricultural land conversion is traditional and extensive world-wide (Lehmann et al. 2003b; Steiner et al. 2007), but is generally inefficient (only <2–3% of the original preburn biomass remains as charcoal) and unsustainable due to nutrient losses, particularly N and S (Glaser et al. 2002; Alexis et al. 2007), while after 5–10 years of biological decomposition <10–20% remains (Jenkinson and Ayanaba 1977). When practiced, as is frequent within the tropics, slash and burn relies on extended fallow periods (which can be around 20 years) after only 1 to 3 years of agricultural usage. Soil organic matter and nutrient depletion and leaching are explained as the primary causes for the degradation of tropical soils within these agricultural systems (Zech et al. 1997).

Black carbon, due to its aromatic structure, is resistant to decay and can therefore occupy a significant proportion of the soil carbon fraction aided by its potential for a prolonged life within the soil (Kuhlbusch et al. 1996; Dai et al. 2005). Estimates suggest that around 0.05–0.2 Pg (1015 g) C yr−1 of black carbon or charcoal are stored annually in soil (Kuhlbusch 1998; Lehmann et al. 2005; Nguyen et al.

2009). There are examples from chernozemic soils in Germany, Canada and the USA where charred matter accounted for >45% of the total soil organic matter (Schmidt et al. 1999; Ponomarenko and Anderson 2001; Skjemstad et al. 2002). It is, however, debateable as to whether repeated wildfires increase soil black carbon content due to surface carbon deposits being consumed in subsequent fires (Dai et al. 2005; Rovira et al. 2009; see ref. within Ansley et al. 2006).

Soil organic matter varies considerably depending on soil type (sandy soils may contain <5% organic matter while it can be close to 100% in wetland soils), but is it the largest carbon reservoir in the biosphereatmosphere system and therefore changes in size can have a large influence on global carbon balance (Lal 2008). Despite the slow rates of production of SOC compared with other flows in the carbon cycle, it is the relative stability (shows resistance to ‘i.e. weathering or recalcitrant’) with respect to microbial

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decomposition that facilitates SOC accumulation.

SOC is, however, on the decline due to soil degradation which has been calculated to have contributed around 80 Pg of carbon (∼60% of the carbon released due to fossil fuel consumption) into the atmosphere (Lal 2004). Climate, particularly temperature and water, have been proposed as important factors in determining the rate of organic carbon decay (Jenkinson and Ayanaba 1977; Sorensen 1974). Some evidence shows that both mean annual temperature and rainfall influence organic carbon in soils. However, it remains unclear if this relationship comes about due to effects on soil carbon production or on decomposition or both (Glaser and Amelung 2003). SOC can also decline radically due to an increase in the rate of decay and its structure in response to some human land induced changes (e.g. tropical forest to grassland) (Solomon et al. 2007a).

Such declines in SOC also take place in response to differences in agricultural practices, such as soil tillage (Maia et al. 2010).

Traditional charcoal production utilises carbon dioxide sequestered into woody biomass tissue via the process of pyrolysis (‘conventional or slow pyrolysis’). Pyrolysis occurs when tissue of biological origin is burnt (or charred) in the absence of, or at low levels of, oxygen to produce ‘biochar’ (Mohan et al.

2006; Preston and Schmidt 2006). Upon pyrolysis approximately 50% of the carbon contained in the original source of biomass can be retained within the biochar; however recovery rates are highly dependent of the pyrolysis process (FAO 1985; Daud et al. 2001; Demirbas 2001; Baldock and Smernik 2002; Lehmann et al. 2002; Laird 2008).

The recalcitrant nature of biochar does not, however, mean that it remains unchanged after application to soil. The stability of carbon derived from biochar has been analysed. Taking soil samples from both Anthrosols and adjacent Amazonian soils has revealed that over an extended period (around 500 days) of incubation, biochar rich soils had a higher proportion of carbon within organic-mineral fractions compared with adjacent biochar-poor soils (Liang et al. 2008). Despite these data from laboratory-based studies on tropical soils, we still have limited direct in situ evidence to begin to model likely effects of long-term biochar application in temperate soils. Long-term studies within the temperate zone are an absolute requirement not only to determine a

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cost benefit of carbon sequestration, but also the extent of an opportunity for agronomic manipulation benefits.

The extent and time over which soil-incorporated biochar carbon remains stable is important, but as yet remains unclear, particularly with respect to different climatic factors (Lehmann and Rondon 2005; Wardle et al. 2008; Cheng et al. 2008b; Hammes et al. 2008; Cheng and Lehmann 2009; see reviews by Hammes and Schmidt 2009 and Lehmann et al. 2009; Nguyen et al. 2009; Nguyen and Lehmann 2009). The processes of ‘erosion’ and ‘transportation’ outside the soil measurement zone are occurring and, if not included in an estimation of carbon retention, would account for errors in estimated losses via degradation (Rumpel et al. 2006; Major et al. 2010a). Stable biochar residue appears to provide a means by which carbon can be sequestered and stored over much longer periods relative to that achievable either within living organisms, or landfill of untreated biowaste; biochar production could potentially mitigate the release of greenhouse gases and further climate change by pyrolyzing such waste biomass. Analysis is being undertaken to assess the potential of pyrolysis of organic biomass as an alternative to the use of non-renewable fossil fuels (Graber and Hadas 2009).

Global biochar production varies around 0.05 to 0.3 Gt (109) C yr−1, while annual net global plant primary production is around 60 Gt C yr−1 (Field et al. 1995, 1998). Diverting more of this global primary plant production, particularly waste material from agricultural production, into biochar and its subsequent soil incorporation could provide a significant and extensive location for storing sequestered carbon.

Lehmann (2009) provides a broad analysis of the associated carbon costs of various biological sequestration approaches. This analysis for biochar, derived from Gaunt and Lehmann (2008), shows biochar sequestration with the lowest carbon costs, compared for example to forest and geological sequestration, and similar to simply burying logs (Lehmann 2009).

Further benefits (environmental and economic) could be envisaged from biochar, such as improvements in soil water quality and income derived from soil carbon sequestration, even in the absence of agronomic benefits (Lal 2004). The conditions under which these benefits will occur requires extensive research before it can provide part of an adaptationary and economic solution to rising GHG emissions.

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Reports suggest that soil incorporated biochars can enhance plant growth (Glaser et al. 2002; Lehmann et al. 2003a; Woods 2003; Yamato et al. 2006; Steiner et al. 2007; Rondon et al. 2007; Kimetu et al. 2008; Chan et al. 2008; Blackwell et al. 2009; Asai et al. 2009).

Regions within the Amazon-basin contain numerous localised sites where the ‘dark earth of the Indians’ (Terra preta de Indio, Portuguese), Anthrosols or Amazonian Dark Earths (ADE) exist (Sombroek et al. 2003; Kern et al. 2003). Their distribution within the Amazon Region is considerable. These sites contain variable quantities of highly stable organic waste (biochar or black carbon within ‘middens’ of domestic waste) which increases the carbon sink capacity of soil (Mann 2002; Kern et al. 2003; Marris 2006; see review of Lehmann et al. 2003b; Myers et al.

2003; Woods 2003; Teixeira and Martins 2003; Steiner 2007). Evidence from carbon dating shows the deposits to be from the ‘pre-Columbian period’ (between 500 and 6,000 years old; Lehmann and Rondon 2005). These deposits vary in their nature but have been shown to be anthropogenic in origin, having been created by Amerindians through various activities which include waste from inhabitation (Terra preta soils) to possible ‘slash and char’ agriculture (Terra mulata soils) (Denevan 1996; Mann 2002; see reviews

by Myers et al. 2003; Kern et al. 2003 and Erickson 2009) and show enhanced nutrient concentrations compared to surrounding soils (Lehmann et al.

2003a). The extent to which human activities in the pre-Columbian era have impacted on ecosystems beyond the known Terra preta sites is debatable, but recent evidence shows the extensive and complex nature of other anthropogenic sites of similarly ancient agricultural activity and landscape change (McKay et al. 2010). Discrete patches of Terra preta are known to contain 70% more black carbon than surrounding soils, i.e. 9% total carbon compared with 0.5% for untreated soil within the same region (Woods 2003). They have also been characterised, relative to other soil types in the region, as supporting a wider genetic agrobiodiversity (Clement et al. 2003).

The higher fertility of Terra preta soils originally attracted the interests of geographers, soil scientists and archaeologists, where in general tropical (Amazonian) soils, can be low in fertility (Glaser et al. 2003). Within many weathered tropical soils, organic matter is the primary reservoir for plant nutrients (N, S [95% of total] and P [75–20% of total]) (Tiessen et al. 1994;

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Steiner et al. 2007). This organic matter, being predominantly confined to the soil surface, is easily leached of available and mobile nutrients. In some cases with Brazilian anthrosols, which have received large amounts of biogenic Ca-P (as fish bones), it has been shown that, compared to inorganic P sources (geogenic Ca-P), biogenic P persisted longer with enhanced bioavailability over long periods (Sato et al.

2009).

Water percolation in well-aggregated soil can be a rapid process in the humid tropics, but nutrient leaching in Terra preta has been shown to be unexpectedly low due to nutrient retention (Lehmann et al. 2003a, b).

Understanding and developing novel ways to ensure that the processes which cause nutrient loss, by leaching, are restricted will become very much more critical in the future as population drives the need for more efficient use of land and because of potentially dwindling finite supplies of growth limiting elements in agricultural production, such as phosphorus.

Biochar physical and chemical properties The physical properties of biochar are key to understanding the way biochar functions within soil and its potential to act as a route to sequester atmospheric carbon dioxide (Downie et al. 2009).

Incorporation of biochar can influence soil structure, texture, porosity, particle size distribution and density, thereby potentially altering air oxygen content, water storage capacity and microbial and nutritional status of the soil within the plant rooting zone (see Amonette and Joseph 2009). It is also apparent that the soil water regime can itself modify biochar stability depending on the initial properties of the feedstock used; biochars produced at lower temperatures and from more labile feedstock are more easily altered (Nguyen and Lehmann 2009). Differences in biochar particle size per se, however over the range of 2 mm to 20 mm do not, at least in some studies, appear to have a significant influence on crop yields (Lehmann et al. 2003b).

Biochar is a carbonaceous material which contains polycyclic aromatic hydrocarbons with an array of other functional groups (Schmidt and Noack 2000; Preston and Schmidt 2006; Krull et al. 2009). Their highly porous structure can, with some biochars, contain significant amounts of extractable humic and

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experiments using biochar applications to soil.

Descriptive information of a biochar’s key characteristics is vital in developing our understanding of its agricultural impacts. Information regarding ‘availability’ of minerals is particularly relevant to our understanding of potential biochar benefits to plant growth (e.g. see Gundale and DeLuca 2007).

Trompowsky et al. (2005) showed that the final pyrolysis temperature influenced the yield of biochar humic and fulvic acids produced from two Eucalyptus species. Similarly differences in pyrolysis temperature using the same feedstock (chicken manure) produced biochars with very different properties including, EC, pH and P and N concentrations (Chan et al. 2008).

Temperature is primarily responsible for the level of carbon lost during pyrolysis and the physical and structural changes apparent (Downie et al. 2009).

However, temperature differences during pyrolysis have been linked to the creation of micro-pores (Angstrom dimensions) formed during the loss of water molecules during dehydroxylation. This increase in porosity resulted in a 3-fold increase in surface area

fluvic acids (Trompowsky et al. 2005), with a molecular structure showing a high degree of chemical/ microbial stability, and an environment dependent turnover on a timescale of millennia (Cheng et al.

2008a and see below). Solomon et al. (2007b) show evidence that the anthropogenic induced accumulation, via pyrolysis, of highly refractory aryl-C structures was the most likely cause for the observed stability of SOC, at least within Amazonian dark earths.

The heterogeneous composition of biochars means that their surfaces can exhibit hydrophilic, hydrophobic, acidic and basic properties, all of which contribute to their ability to react with soil solution substances. The variability in biochar physical and chemical properties depends on the material used to produce it (feedstock), the availability of oxygen and the temperatures achieved during pyrolysis (Lua and Yang 2004; Gundale and DeLuca 2006; Amonette and Joseph 2009). Table 1 shows a somewhat restricted list of the elemental and some chemical constituents of several different feedstocks used to produce biochar. It is critical that such information is made available from all

Table 1 Chemical constituents of biochars produced from various feedstock sources under different production temperatures Biochar feedstock

pH

C g kg−1

N g kg−1

Bark: Acacia mangium

7.4

C/N

P g kg−1

K g kg−1

Production (°C)

Information source

398

10.4

38

260–360

Yamato et al. 2006

Coconut: Cocos nucifera, shell

690

9.4

73

500

Tsai et al. 2006

Corn: Zea mays residue

675

9.3

73

10.4

350

Nguyen and Lehmann 2009

Corn: Zea mays, residue

790

9.2

86

6.7

600

Nguyen and Lehmann 2009

680

1.7

400

0.2

1

450

Chan et al. 2007

45

0.6

6.2

400

Magrini-Bair et al. 2009

700

Novak et al. 2009

Green waste

6.2

Peanut: Arachis hypogaea, shell Pecan: Carya illinoinensis, shell

499 7.6

Pecan: Carya illinoinensis, shell Poultry: litter

834 880

9.9

380

11 3.4

245

4.0

220

20

700

Busscher et al. 2010

19

25

22

450

Chan et al. 2007

Poultry: Broiler litter

258

7.5

34

48

30

700

Lima and Marshall 2005

Poultry: Broiler cake

172

6.0

29

73

58

700

Lima and Marshall 2005

Rice: Oryza sativa, straw

490

13.2

37

500

Tsai et al. 2006

Sewage sludge

470

64

Sugar cane: Saccharum spp. bagasses Wood: unknown

710

17.7

7

56

40

450

Bridle and Pritchard 2004

500

Tsai et al. 2006

350

Rondon et al. 2007

wildfire

DeLuca et al. 2006

708

10.9

65

6.8

Wood: Eucalyptus deglupta

7.0

824

5.7

144

0.6

Wood: Pinus ponderosa Pseudotsuga menziesii Wood: Quercus spp.

6.7

740

16.6

45

13.6

759

1.0

759

1.1

350

Nguyen and Lehmann 2009

884

1.2

737

2.2

600

Nguyen and Lehmann 2009

Wood: Quercus spp.

0.9

Lehmann et al. 2003a

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(Bagreev et al. 2001). Natural oxidation can also take place at the surface of particles and this can be related to not only exposure time, but also mean annual temperatures (Cheng et al. 2006; Cheng and Lehmann 2009). This oxidation can impact greatly on changes in soil biogeochemistry and nutrient release and retention (Cheng et al. 2008a).

Biochar porosity, which determines its surface area, shows pore size distribution that is highly variable and encompasses nano- (<0.9 nm), micro- (<2 nm) to macro-pores (>50 nm) (Downie et al. 2009). Larger macro-pores are a key to its function in soil, i.e. aeration and hydrology; they also provide a habitat niche for microbes. Smaller pores are involved with molecule adsorption and transport. Soil structure varies with soil type and is closely linked to particle size distribution. For example, sandy soils have limited specific surface area (sand 0.01 to 0.1 m2 g−1), and can only store relatively small quantities of water or nutrients compared with the greater specific surface area of clay particles (5 to 750 m2 g−1) (Troeh and Thompson 2005). The inclusion of biochar in one study with a sandy soil showed an enhancement in specific surface area (× 4.8) relative to adjacent soils (Liang et al. 2006). Downie et al. (2009) quoted biochar sources that had specific surface areas significantly greater than those of clay (>1,500 m2 g−1). The porous nature of biochar is also important as it can provide refugia for some beneficial organisms such as mycorrhizae and bacteria (see below) (Saito 1990; Pietikäinen et al. 2000). The porosity and surface area of a biochar will have very important effects on its nutrient retention capacity by surface binding of both cations and anions to its surfaces (Liang et al. 2006; Chan and Xu 2009). However depending on pyrolysis conditions, low temperatures have been shown to be particularly important, when the recondensing of volatile organic compounds on the char can block these pores and their adsorption potential (Kwon and Pignatello 2005; Pignatello et al. 2006). These authors suggest that biochar porosity declines due to the presence of substances, such as lipids, fulvic and humic acids, which themselves show biochar production temperature differences in molecular rigidity which can restrict pore access.

It is very apparent that the physical and chemical properties of biochar feedstock along with the conditions under which it is produced can have a large impact on the effects of biochar has when applied to

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soil, e.g. the level of humic acid etc. We already have some notion of how the pyrolysis process and the feedstock type can be used to create specific purpose or ‘designer biochars’. This area clearly requires more collaboration and development between pyrolysis engineers and agricultural scientists.

Nutrients Organic biomass derived from manures and compost contains large amounts of carbon and macro- and micro-nutrients (see Chan and Xu 2009). It appears that utilisation of these sources of organic matter, as biochar feedstocks will alter the availability of key macro-nutrients such as N and P, and some metal ions (e.g. Ca and Mg), when incorporated into soil (Lehmann et al. 2003a; Gundale and DeLuca 2006; Major et al. 2010a, b). Feedstock and pyrolysis conditions influence the carryover of minerals to biochars (see Amonette and Joseph 2009), with the presence of key biochar elements linearly dependent on the levels within the initial feedstock (Alexis et al.

2007). There are, however, likely to be different mineral pyrolysis temperature optima (Bridle and Pritchard 2004). Relative differences in feedstock nutrient value are suggested to be conserved even when prepared under different pyrolysis conditions (DeLuca et al. 2009).

The Terra preta soils, despite being known for their high fertility, are linked with below optimum availability of key plant nutrients for crop production (Lehmann et al. 2003a, b). Nitrogen is cited as one such element where content is high, but ‘availability’ is low, while the opposite was true for P and Ca. The largest proportion of N occurs in the organic form, or is organically bound in these soils while the opposite in true for P (Lehmann et al. 2003b). Similarly, with biochars, nutrient levels and availability can be very low, e.g. N and P (Bridle and Pritchard 2004). The ratio of C to N within biochars can be very variable depending on feedstock and pyrolysis conditions (see data in Krull et al. 2009, table 4.1). This ratio not only influences the recalcitrant properties of the biochar, but may also affect the types of C and N released during mineralisation (Krull et al. 2009). Given the high C/N ratios for biochar there is an expectation that N immobilisation occurs, inducing plant N deficiency.

Again, the recalcitrant nature of the carbon restricts N immobilisation (Chan and Xu 2009). Somewhat

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indirectly biochar incorporation has been shown, in many studies, to induce soil alkalisation which can increase soil nitrification (Lehmann et al. 2003b; Oguntunde et al. 2004; Topoliantz and Ponge 2005; Yamato et al. 2006; Chan et al. 2007; DeLuca et al.

2009; Hua et al. 2009; Major et al. 2010a, b), but increases in soil nitrification rates do not rely solely on soil pH changes (Berglund et al. 2004). Increases in soil pH are likely to influence P availability which is highly pH-dependent; in acid soils (<pH 4) insoluble iron and aluminium phosphates form, while in alkaline soils (>8.5 pH) insoluble Ca phosphates dominate.

Chemical properties Biochar soil additions cause changes in pH, electrical conductivity (EC), cation exchange capacity (CEC) and nutrient levels (Liang et al. 2006; Gundale and DeLuca 2007; Warnock et al. 2007; Amonette and Joseph 2009). The increases in soil pH induced by biochar application are not surprising given the well documented use of material such as wood ash in modifying pH and nutrient availability, particularly P and K (Clarholm 1994; Mahmood et al. 2003).

Elevated CECs are due to increases in charge density per unit surface of organic matter, which equates with a greater degree of oxidation, or increases in surface area for cation adsorption, or a combination of both.

Liang et al. (2006) reported that, in anthrosols, as a consequence of black carbon particle surface oxidation the adsorption of organic matter and its charge density (CEC per unit surface area) increased.

Ammonium leaching, albeit from greenhouse biochar experiments, was reduced (by 60%) (Lehmann et al.

2003a; Major et al. 2009), while in some cases N2O emissions can be reduced (Spokas and Reicosky 2009). Other studies, using soils in Amazonian field trials, have confirmed that biochar can act as an adsorber reducing N leaching and increasing N use efficiency (Steiner et al. 2008c). There would appear to be huge area of research required to ensure that the prolysis process and the feedstock used have the potential to optimise soil N for plant availability while minimising leaching. Efficiency of N use will be an absolute requirement to sustain future population growth. To achieve this much more needs to be understood regarding the mechanistic influence of biochar (direct and indirect) on nitrification and Navailability.

Some biochars also appear to reduce the mobility of heavy metals (Cu and Zn) (Hua et al. 2009) and other organic soil contaminants (i.e. insecticides, Hilber et al. 2009). Positive surface charges will, however, decrease as the biochar oxidises, changing its absorption properties (Cheng et al. 2008a). Biochar application can also reduce the concentration of soluble compounds such as phenols in the soil solution (Gundale and DeLuca 2007). The availability of some elements toxic to plant growth, particularly at low pH, such as Al, Cu and Mn, can be reduced by biochar incorporation (Sierra et al. 2003; Steiner 2007; Steiner et al. 2008a), while the availability of other elements can increase, with biochar induced increases in soil pH enhancing solubility e.g. N, P, Ca, Mg and Mo. Physical properties Little is published on the effects of biochar incorporation on soil physical properties (see review by Hammes and Schmidt 2009). Factors such as biochar mobility within the soil profile are important, particularly with respect to benefits to plant production and potential movement into the ground and surface waters (Leifeld et al. 2007). The limited evidence, from very few studies (over 35 years) shows with sugar cane production and waste burning long-term, that cane biochar moves down the subsoil (Skjemstad et al.

1999; Leifeld et al. 2007; Major et al. 2010a, b). Other evidence, at certain sites, shows the movement of biochar into the subsoil over 100 years, (Hammes et al.

2008). This has been linked to a decline in biochar particle size with time in the soil (Lehmann et al.

2009). However, relative to biochar, black carbon losses from agricultural residues and in particular soil organic matter, were considerably greater (Rumpel et al. 2006; Steiner et al. 2007; Major et al. 2010a, b).

It has been shown that soil bulk density (BD) within some Amazonian dark earths, but not all, was lower in the upper relative to the lower horizons (Teixeira and Martins 2003). These authors have collated information about increases in BD with depth of profile from a number of soils, sites and authors (see Table 2 in Teixeira and Martins 2003). They indicate total porosity was high in the upper aerated horizons, but declined as the organic matter content fell with depth. However, these properties were deemed to be advantageous to agriculture. They also

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indicate that these changes may cause mechanical resistance to root penetration to increase, but in general the hydraulic properties of these dark earths compare favourably with much ‘lighter’ soils. This was suggested to be contrary to what would be apparent from only field observations of dark earths (Teixeira and Martins 2003). It may therefore be necessary to consider the potential of critical evaluation of the benefits of soil biochar incorporation to improve mechanical impedance. Some soils with a high level of impedance/compactness, and low infiltration can become waterlogged, and show restricted root growth and plant development. In some situations, e.g. upland rice growing, biochar applications can improve soil water permeability (Asai et al. 2009). The ease with which plant roots penetrate soil is known to impact on growth and yields (see e. g. Whalley et al. 2006). Certain soils, particularly when dry, limit root penetration at depth and can be improved through biochar application (Chan et al.

2007).

Considerable knowledge is required to develop an improved understanding of the way in which biochar incorporation into soil changes its physical structure on application and subsequently how the physical nature of the soil/biochar alters long-term. Experimentally, this is a challenge if done in situ which appears the most appropriate. The careful design and constructions of soil biome regimes will likely be required in the short-term to develop longer-term modelled predictions prior to field data becoming available.

Soil biota The structure and function of biological communities within soils is complex, with its varied inhabitants grouped into algae, archaea, arthropods, bacteria, fungi, nematodes, protozoa and other invertebrates.

The presence and variable abundance of these groups has a profound effect on soil function and health and productivity, as does the application of organic matter and biochar. Amazon soils contain a diverse range of micro-organisms adapted to the soil’s biochemistry and ecology (Thies and Suzuki 2003; Kim et al. 2007; O’Neill et al. 2009; Thies and Rillig 2009). Taxonomic studies using molecular approaches (ribosomal gene fingerprinting) show that Terra preta soils contain higher numbers of operational taxonomic units (OTU)

compared with pristine forest soils, i.e. 396 OTU, compared with 291 OTU (Kim et al. 2007). A similar comparison of bacterial species richness showed Terra preta soils were about 25% richer than forest soil, with 14 phylogenetic groups compared with only 9 in forest soils (Kim et al. 2007).

Bacteria Evidence from the use of wood ash, in pot experiments, has shown that bacterial activity, measured through isotopic incorporation of labelled thymidine and leucine, was enhanced along with bacterial community structure (Mahmood et al. 2003), while suggestions have been made that the application of charcoal to soil can have significant impacts on carbon utilisation profiles and population structures (O’Neill et al.

2009), along with increases in basal soil respiration and respiration rate per micro-organism (Pietikäinen et al. 2000: Steiner et al. 2008a). More recent work has indicated, contradictory, reductions and stimulations of measured respiratory products when different biochars were applied to different soils (Spokas and Reicosky 2009). A number of explanations for this respirational increase have been suggested and include increases in pH, and the availability of minerals. For example, the presence of heterotrophic phosphate-solubilising micro-organisms was enhanced after charcoal soil additions (Kimura and Nishio 1989). Along with increases in microbial community biomass, there were linear increases in microbial efficiency (CO2 released per unit of soil carbon) over a range of charcoal applications from 50 to 150 g kg−1 of soil (Steiner et al. 2008a). There have also been suggestions that biochar could promote the growth of micro-organisms causing the decay of the more labile compounds within biochar (Hamer et al. 2004).

Diazotrophs, a specialised group of bacteria (and the Archaea), share a common function: they possess the enzyme nitrogenase and the ability to reduce atmospheric N2 into NH3, which can be nitrified (NO3−) prior to plant uptake. These diazotrophs act as either free-living N-fixing soil bacteria (e.g. Azospirillum sp.; Azobacter sp.) or as mutualists within plants (e.g. the rhizobia that form legume nodules or the actinorrhizal associations of Franksia sp.) (DeLuca et al. 2009). The free-living bacteria are less effective at N fixation than symbiont rhizobia, i.e. 5 compared to 3–206 kg N ha−1 y−1 respectively. Despite the

Plant Soil (2010) 337:1–18

ubiquitous presence of these free-living N-fixing soil bacteria, there have been few studies that show that applications of activated carbon (Berglund et al. 2004) and biochar (Gundale and DeLuca 2006) can increase nitrification. The biochar micro-environment may also provide a favourable niche (fine structural pores) in which oxygen concentration declines; for nitrogenase to function effectively, low oxygen tensions are required with Fe and Mo ions (Thies and Rillig 2009). Biochars are generally low in inorganic-N and this can provide diazotrophs with a competitive advantage for colonisation of the biochars large surface area. This factor combined with biochar’s potential for NH4+ exchange with the soil solution could modify soil-N availability to the plant and stimulate nodulation and fixation.

Little is known about the impacts of biochar on N immobilisation and denitrification (DeLuca et al. 2009).

The reducing of NO3− to N2 in the absence of oxygen is achieved via several intermediates (NO2−, NO, N2O) which can be released to the atmosphere. Biochar may have the potential to catalyse the reduction of N2O to N2, reducing the emission of a key greenhouse gas.

However, supporting evidence is limited (Van Zwieten et al. 2009). These authors also discuss how biochar application might reduce CH4 emissions. The concept that biochar applications might reduce denitrification requires much more investigation. The impacts of potential reductions in agricultural production of such a potent GHG would be considerable given the current emission levels and the likely increases in food production with global population growth.

A role has been suggested for biochar in adsorbing and protecting chemical signalling molecules derived from plants such as the nod factors which enhance root nodulation via rhizobia (Thies and Rillig 2009).

Evidence exists to show that increasing biochar application rates to soil can increase the proportion of N derived from fixation by Phaseolus vulgaris, and this increased yield (Rondon et al. 2007). These beneficial effects were linked to increased availability of Mo and B (source not determined), with an increase in soil pH. Rhizobia show increased function in neutral pH soils, so increasing alkalinity in acidic soil enhances nodulation and fixation.

Fungi The different functional groups within ‘soil fungi’, i.e. saprophytes, pathogens and mycorrhizae, respond

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differently to biochar application (Thies and Rillig 2009). Saprophytes potentially modify the persistence of soil biochar, through decomposition. Their invasive hyphal growth and extracellular enzymatic capacity enables them to colonise biochar pores. Soil pathogenic fungi are extensive and particularly important with respect to plant disease management. However, the effect of biochar on soil pathogens (population structure and function) appears limited. Matsubara et al. (2002) has shown that asparagus seedlings tolerance to Fusarium oxysporum was enhanced with biochar application.

Mycorrhizal fungi are ubiquitous key components in nearly all terrestrial vegetation and crop systems, showing a very high degree of specificity and mutualism, enhancing plant survival, establishment and growth. There are a number of types, the most common being arbuscular mycorrhizae (AM) and ecotomycorrhizae (EM) which are distinctive in their morphologies. Soil biochar incorporation appears to have a positive impact on mycorrhizal fungi (MF). Positive impacts have been described for soil applied wood ash (Mahmood et al. 2003).

Some arbuscular mycorrhizae increase root colonisation sites in the presence of biochar (Ishii and Kadoya 1994; Warnock et al. 2007). This increased colonisation has yet to be clearly linked with specific biochar characteristics, just with the presence of additional SOM. Soil biochars can also increase endomycorrhizal plant associations, enhancing P availability (Garcia-Montiel et al. 2000). Several mechanisms explain how mycorrhizal colonisation, with biochar applications, may enhance agricultural performance (see review Warnock et al. 2007). These potential beneficial mechanisms are: nutrient availability is enhanced, or there are changes in soil physiochemical properties; alterations in other beneficial or detrimental soil microbes (e.g. mycorrhizal helper bacteria (MHB) or phosphate solubilising bacteria (PBS); enhancement of the ability of MF to resist plant fungal pathogen infection, through enhanced root colonisation (Matsubara et al. 2002).

Alterations in plant and fungi rhizosphere chemical signalling impacting on mycorrhizal colonisation rates are also suggested as mechanisms. This may involve alterations in the production and intensity with which root-derived signal molecules are produced, as well as the detoxification of rhizospheric allelochemicals which hinder mycorrhizal establish-

Plant Soil (2010) 337:1â&#x20AC;&#x201C;18

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ment. The presence of biochar in the soil can provide a physical niche for growing hyphae and bacteria, devoid of fungal grazers (Saito 1990; Pietikäinen et al.

2000; Warnock et al. 2007). Soil grazers in the size range of collembola and protozoans (>1.6 mm) would be excluded given the measured biochar pores size.

Warnock et al. (2007) propose that biochar additions promote crop production through a combination of factors. While it is unclear which is the most important, published data support an increased nutrient availability hypothesis. Despite the fact that biochars themselves may contain only small amounts of plant nutrients to benefit mycorrhizal fungi (Lehmann et al.

2003b; Gundale and DeLuca 2006), it is proposed that biochar increases soil nutrient availability through alterations in soil physico-chemical properties. These changes in themselves alter the availability of nutrients and potentially mycorrhizal abundance, changing the local nutritional balance, e.g. N/P ratios, and thus affecting root colonisation (Miller et al. 2002). The concept that nutrient availability or increased efficiency of use, due to the presence of biochar, requires critical examination to determine the role played by soil micro-organisms.

Invertebrates As yet there is little, if any, evidence that the soil fauna in general is affected either directly or indirectly by biochar application (Chan et al. 2008; Thies and Rillig 2009). It is, however, well known that processes that influence the flow of energy and organic matter within the soil will impinge on bacterial and fungal-based energy chains, which impact at higher trophic levels. This potentially could be important for soil fauna that ingest biochar, as would be the case with geophagous organisms such as earthworms (Topoliantz and Ponge 2005). These authors suggest that a specific earthworm Pontoscolex corethruus was a significant factor in the development of Terra preta soils in Brazil, particularly with respect to improving soil quality. Such activity would also relocate biochar and aid soil incorporation (Hammes and Schmidt 2009). Chan et al. (2008) suggest that the biochar production process with respect to pyrolysis temperature can influence earthworm soil/ biochar type selection preferences. The mechanisms behind the wormâ&#x20AC;&#x2122;s selection processes are as yet unknown.

Biochar induced changes in crop production It is frequently suggested that biochar applications to soil can increase agricultural productivity (reviewed

by Lehmann et al. 2003b; Blackwell et al. 2009).

These authors show that in a high proportion of the studies (>90%), biochar-induced increases in crop yield were apparent, while Lehmann and Rondon (2005) report that, depending on the amount of biochar added, significant improvements in plant productivity were achieved ranging from 20% to 220%. Blackwell et al. (2009) indicates that the crop list investigated is restricted and does not include work on grasslands, shrubs and trees, or even perennial tropical crops. In the latter case, tropical soils are usually highly weathered, acidic (elemental toxicity), show a high degree of leaching and some have a high clay content (Oxisols and Ultisols), so many soil productivity enhancing management approaches will be likely to yield positive benefits here as opposed to many temperate agricultural soils.

The importance attached to the extent with which biochar application might increase agricultural production is an important driver in any attempt to develop systems that economically incorporate pyrolysis products within the soil. It is not the only consideration (carbon sequestration is also very important), but it requires long-term investment in agricultural experimentation.

Water relations Soil moisture content is primarily determined by soil texture and precipitation rate. Soil water availability is a key factor in determining agricultural productivity world-wide and will be exacerbated by climate change (Bates et al. 2008). Increases in soil organic matter will likely increase water availability. Terra preta soils have been shown to have higher moisture contents (Lehmann et al. 2003b and quoted within from Hartt 1885; Teixeira and Martins 2003) than neighbouring soils. Improvements in soil field capacity have been recorded upon biochar addition (Chan et al.

2007). An increase in the soil water content or water holding capacity of a soil amended with biochar will likely have greater benefits in sandy soils. Compared to loamy and clayey textures from a podzolic forest soil, it was the sandy textured soil that showed increased moisture levels on charcoal application

Plant Soil (2010) 337:1â&#x20AC;&#x201C;18

(Tyron 1948). This suggests that soils with high clay fractions are less likely to benefit from biochar applications than sandy soils (Woolf 2008). This notion is, however, somewhat inconsistent with potential agricultural benefits, seen with Terra preta soils being derived generally, or solely, from improvements in the soil/plant water balance. Increasing the retention of water, particularly in the rooting zone, can also influence, increases as well as decreases in nutrient movement and leaching (Major et al. 2009).

By their nature charcoal and biochar are particularly porous and once their initial hydrophobicity (Bornemann et al. 2007) is overcome they have the potential to oxidise and absorb and retain water (Cheng et al. 2006). In some cases biochar use can enhance soil water permeability, but this would be more of a challenge with soils with high clay content (Asai et al. 2009). These authors also suggest that soil water holding capacity is increased through biochar application. Glaser et al. (2002) have also demonstrated an 18% higher water retention value for Amazonian anthrosol relative to nearby soil with no biochar. This is supported by the notion that irrigation requirements should potentially decline in the presence of biochar. As yet there is little clear experimental proof that biochar does impact on soil or plant water relations. Major et al. (2009) suggest that due to the physical characteristics of biochar there will be changes in soil pore-size distribution and this could alter percolation patterns, residence time and flow paths of the soil solution. It has also been suggested that if biochar contains sufficient amounts of humic substances, which they can, then this can increase soil water holding capacity (Piccolo et al. 1996). If water holding capacity is increased then expectations of nutritional benefits could also be gained for soil solution mobile elements.

Less direct alterations in soil water content induced by biochar include changes in soil oxygen tensions which influence soil micro-organisms and their capacity to oxidise SOM (Van Gestel et al.

1993). Soil wetting and drying cycles alter the level of soil saturation which influences organic matter decomposition (Sorensen 1974), nutrient availability (Nguyen and Marschner 2005) and microbial activity (Van Gestel et al. 1993). The soil water regime has also been shown to influence the stability of biochar, with changes in soil saturation levels being important in determining oxidative breakdown (Nguyen and

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Lehmann 2009). The potential opportunity that incorporation of biochar into soil may offer with respect to the conservation of soil moisture is considerable and warrants more appropriate experimental evaluation. This is particularly true if the likely impact of further climate change influences extensive agricultural regions as fresh water becomes scarcer. Opportunities to conserve soil moisture content and enhance the fraction available for plant extraction will be a critical factor in sustaining agricultural production. It is also likely, should biochar provide a route by which soil moisture content can be conserved, that it could be a useful tool in landscape amelioration particularly in more sandy soils (arid regions) where soil moisture retention is critical in the initial reestablishment of vegetation and crops.

Crop nutrition The importance of soil mineral nutrition as a limiting factor in maximising crop yield, with N the most limiting factor, is well known. Nitrogen primarily exists in the soil in organic complexes, which are subsequently ammonified (NH 4+ ) then nitrified (NO3â&#x2C6;&#x2019;), before plant uptake. It is well established that post-fire soils show changes in soil nutrient dynamics, for example, elevated N cycling and increased N availability (Covington and Sackett 1992; Gundale and DeLuca 2007). Studies from savanna-type grasslands show a trend for N accumulation with burning which may occur indirectly due to P stimulated N fixation by cyanobacteria or the stimulation of N2 fixation if root nodule forming species are present (Ansley et al. 2006), while biochar application has been shown to stimulate N fixation by beans in association with Rhizobia spp. symbionts (Lehmann and Rondon 2005; Rondon et al. 2007).

There is as yet no evidence to support the idea that free-living N-fixing bacteria are influenced by biochar application. It is, however, well known that excessive soluble forms of N in the soil solution reduce N fixation (Dazzo and Brill 1978), while available P can stimulate it, and therefore its presence in a soluble form, could increase bacterial N2-fixation (Lehmann et al. 2003a, b).

Biochar studies suggest that despite significant loss of labile N volatilised on burning (70â&#x20AC;&#x201C;90%), charcoal residues can contain considerable amounts of the

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element (Raison 1979; Rovira et al. 2009). These authors showed that the amounts of N in the biochar ranged from 21 to 370 mg kg dw−1. When plant tissues are pyrolysed at relatively low temperatures (100°C) carbon volatilizes along with oxygen is lost, while P volatilisation requires 700°C. The process of pyrolysis therefore enhances the availability of P relative to biochar carbon (DeLuca et al. 2006). This enhancement of P content in biochar is apparent when poultry litter is pyrolysised at 450°C compared to 550°C and this may be the cause for observed increases in radish yields (Chan et al. 2008).

However, there is also conflicting evidence that biochar application does little to contribute directly to the soil nutrient status (a source of available mineral elements), again it is combinations of biochar and applied fertiliser which cause yield enhancements via reduced fertilizer (nutrient) leaching (Lehmann et al. 2003b).

Terra preta soils contain significantly greater amounts of N as well as other elements such as Ca and P (Glaser et al. 2001; Lehmann et al. 2003b; Steiner 2007). These are the key elements influencing the production potential and fertility of most soils (Russell 1988). Phosphorus has long been considered a primary limiting nutrient in soils that are highly weathered in the humid tropics (Williams and Joseph 1976).

Biochar applications to forest soils in different geographic regions show stimulated N transformation (DeLuca et al. 2006), but soil N dynamics are by no means fully understood. As has already been mentioned the capacity of a biochar and its potential to both adsorb N (NH4) and subsequently increase Navailability to the plant may explain varied yield responses; this is a key question that warrants further experimentation. Pot experiments with radish (Raphanus sativus) suggest that charred green waste did not directly enhance yields, while some evidence indicated that when biochar incorporation takes place with the addition of N fertilisers growth stimulation can be synergistic (Steiner et al. 2007; Chan et al. 2007; Chan et al. 2008; Asai et al. 2009), Nutrient bioavailability and plant uptake of P, as well as K, Ca, Zn and Cu have in some cases increased in response to charcoal application, while N leaching declined (Lehmann et al. 2003a; DeLuca et al. 2009; Major et al. 2010a, b). Rice yields have been shown to increase with biochar application when soils

Plant Soil (2010) 337:1–18

are known to be low in available P (Asai et al. 2009).

Abiotic events such as drought are known to decrease P availability. While numerous complex reactions with clay and organic matter are apparent, it is unclear how changes in fluxes between insoluble and solubleP are linked with biochar applications. Mechanistic explanations include biochar acting as a source of soluble P-salts and exchangeable P, as a modifier of soil pH (which ameliorates P-complexing metals) and as an enhancer of microbial activity and therefore indirectly subsequent mineralization of P (DeLuca et al. 2009). For example, Steiner et al. (2008b) have shown that biochar application to Amazonian upland soils promoted microbial biomass, expressed as an increase in metabolic quotient, which they suggest increased the capacity to solubilise soil phosphate.

Biochar generally, when using soil incubation assays, irrespective of temperature of production, has the potential to increase extractable P (PO43−) within the soil solution (Gundale and DeLuca 2006, 2007).

Biochar can alter P availability directly through its anion exchange capacity or by influencing activity/ availability of the cations that interact with P. This can lead to soil oxides of elements such as aluminium and iron being unable to bind with soluble P which has linked with these biochar exchange sites. Phosphorus precipitation also influences the solubility of P and therefore the amount available to the plant. The effectiveness (strength of the ionic bond) with which P combines and forms insoluble compounds with various cations (Ca2+, Al3+ and Fe2+3+), and subsequently precipitates, depends on pH. Biochar can alter soil solution pH and therefore bonding as well as sorbing these metal cations, thus avoiding precipitation with P. An increase in pH can increase alkaline metal (Mg2+, Ca2+ and K+) oxides. This reduces soluble forms of aluminium which is suggested as the most significant biochar factor affecting P solubility (DeLuca et al. 2009). The situation is different in soils that are already alkaline or neutral; adding alkaline metals would potentially enhance Ca bonding with P. Biochar may also provide indirect effects on P availability and uptake through changes in the soil environment for micro-organisms (see above). Symbiotic soil fungi are well documented as enhancers of the efficiency of plant P uptake, particularly in soils low in the element. Under these conditions biochar has been shown to increase the yields of maize and peanut by changes in P availability (Yamato et al.

Plant Soil (2010) 337:1–18

2006). The presence of these beneficial symbiotic fungi, such as mycorrhizae, and their enhancement due to the presence of biochar, may explain why low rates of nutrient application can be effective yield enhancers. The mechanism by which this has been achieved involves mycorrhizal hyphae enhancing the interception of minerals that would potentially be lost as leachates in their absence (Allen 2007).

Recently published work (Novak et al. 2010) highlights the need for considerably more critical understanding of the nutritional role of different biochars in different soils and climates; this has been apparent for some time, but now the possible confounding role played by crop residues is becoming clearer. These authors show that NO3–N immobilisation occurred when crop residues were incorporated with soil applied biochar and this resulted in, albeit temporary, reduced plant available NO3–N. Despite the fact that these were in vitro experiments they highlight the very important issue of our limitations in being able to understand and predict the impact of biochar application on soil nutrition, as well as the added complexity of real short-term interactions between the carbon (organic matter) and nitrogen cycles.

Processes which increase the fertility of agricultural soils may also increase unwanted weed pressure (Major et al. 2003). Some limited research has shown that despite substantial increases in weed ground cover induced by the use of manures and composts, biochar additions did not influence weed abundance or species richness (Major et al. 2005).

Concluding remarks Production of biochar and its incorporation into temperate soils is a relatively novel concept for establishing a long-term sink for atmospheric CO2 storage. Biochar provides a stable and inert form of carbon sequestration which is potentially long-term and substantial, with a low risk of return into the atmosphere. The use of biochar as a source of soil carbon sequestration appears to have important adaptational/mitigation potential, providing economical solutions to its production and incorporation can be found. Clearly, more information is required about the stability of biochar in temperate soils, particularly over the long term. For example,

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markedly lower mean annual soil temperatures in temperate regions may be particularly important with respect the promotion of long-term biochar stability.

There is also considerable potential to explore the added value benefits of biochar, beyond utilisation of key waste sources and avoiding land fill and environmental contamination. These benefits are suggested as improvements in soil nutrient availability by reducing leaching and promotion of soil health. If biochar applications can also be utilised to improve the water holding capacity of freely draining soils then there will be considerable opportunities to improve landscapes and soil that have degraded. The latter are key components in the development of improved agricultural efficiency which will rely at minimum on sustaining, if not increasing, land use efficiency to meet changes in climate and securing food for an increasing global population.

While the potential agricultural influences of biochar use have been identified in tropical regions, its use in the northern hemisphere has not been studied in sufficient critical detail. Studies elsewhere show that incorporation of biochar influences soil physical structure, chemistry and biology. The physical structure of biochars determines their porosity and provides refugia for beneficial soil organisms such as fungi (mycorrhiza) and bacteria. These features can improve ‘soil health’ by enhancing processes like soil nitrification, with the added benefits of catalysing N2O reduction reducing GHG emissions (O’Neill et al. 2009). It appears, at least in some soils, that soil biochar can improve plant productivity. Other benefits reported from biochar incorporation into soils, include reductions in environmental pollution (e.g. metal contamination); reductions in fertiliser applications and increasing efficacy of water usage. Opportunities appear to exist in that the feedstock and the biochar process (pyrolysis conditions) used have the potential to be managed to derive purposeful or ‘designed biochars’ developed for specific end uses.

Acknowledgements Financial support for this study was provided by The East Malling Trust and South East England Development Agency (SEEDA). We are also extremely grateful to the editor and anonymous reviewers for their constructive comments on earlier drafts and to all the many authors who made available copies of their recently published work.

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