An investigation into the reactions of biochar in soil.
Biochar is a carbon-rich solid material produced by heating biomass in an oxygen-limited environment and is intended to be added to soils as a means to sequester carbon (C) and maintain or improve soil functions. Interactions between biochar, soil, microbes, and plant roots are known to occur within a short period of time after application to the soil (Lehmann and Joseph 2009). However, the extent, rates, and implications of these interactions are still far from being understood, and this knowledge is needed for an effective evaluation of the use of biochar as a soil amendment and tool for C sequestration. Recent studies (Steiner et al. 2007; Bruun et al. 2008; Singh and Cowie 2008; Kuzyakov et al. 2009) suggest that the types and rates of interactions (e.g. adsorption-desorption, precipitation-dissolution, redox reactions) that take place in the soil depend on the following factors: (i) feedstock composition, in particular the total percentage and specific composition of the mineral fraction; (ii) pyrolysis process conditions; (iii) biochar particle size and delivery system; and (iv) soil properties and local environmental conditions.
High-temperature pyrolysis (>550[degrees]C) produces biochars that generally have high surface areas (>400 [m.sup.2]/g) (Downie et al. 2009; Keiluweit et al. 2010), are highly aromatic and therefore very recalcitrant to decomposition (Singh and Cowie 2008), and are good adsorbents (Mizuta et al. 2004; Lima and Marshall 2005). Low-temperature pyrolysis (<550[degrees]C), on the other hand, favours greater recovery of C and also of several nutrients (e.g. N, K, and S) that are increasingly lost at higher temperatures (Keiluweit et al. 2010). Low-temperature biochars, which have a less-condensed C structure, are expected to have a greater reactivity in soils than higher temperature biochars and a better contribution to soil fertility (Steinbeiss et al. 2009). In fact, pot and field trials indicate that high mineral-ash biochars produced at temperatures <500[degrees]C have, in some cases, given higher crop yields than more recalcitrant biochars produced at higher temperatures (Chan et al. 2008). Based on this greater reactivity, low-temperature biochars have been blended with minerals and sludges to balance the nutrient content of the amendments, and results of pot and field trials are now entering the scientific literature (Chia et al. 2010). However, optimal heating rates and soaking times must be determined when operating kilns at temperatures of 180-350[degrees]C, and then adopted during production, to avoid the production of compounds that, in sufficiently high concentration, could be toxic to plants, such as acid aldehydes or phenols (Bridgwater and Boocock 2006).
This paper reviews current understanding of the reaction of biochars in soil and reports recent findings from field trials in which biochar has been in the soil (Australian Ferrosol) for 1-2 years. As biochar-soil interactions are expected to be more intense for low-temperature biochars, this work focuses on them; however, some aspects are applicable to other biochars.
Properties of biochars
Biochars produced at temperatures <550[degrees]C, and especially those with high ash content, have intricate surface and internal properties that result in complex interactions with the components of soil (Shinogi et al. 2003; Amonette and Joseph 2009). Low-temperature biochars have a predominantly amorphous C structure, with a lower aromaticity than high-temperature biochars (McBeath and Smemik 2009; Keiluweit et al. 2010). The morphology of the biochar resembles that of the parent material; for example, wood biochar has the exoskeleton of the tracheids, whereas chicken manure biochar has a heterogeneous structure consisting of charred remnants of seeds, hair, proteins, digested food, bedding material, and minerals. The pore structure, size distribution, volume, and total surface area are a function of the properties of the original biomass feedstock, as well as the process conditions (Downie et al. 2009; Keiluweit et al. 2010). Most of the physical, electrical, and chemical properties change as final heat treatment temperature and time increase until a point is reached where most of the carbon is in the form of graphite (Antal and Gronli 2003). As the heat treatment temperature increases, some metals in the carbon lattice can be volatilised (e.g. K initiated at ~400[degrees]C) and the ash phases change their morphology (either transform from a crystalline structure to an amorphous structure or vice versa, e.g. silica) or chemical composition through decomposition, oxidation, or reduction (Womat et al. 1995; Bridgwater and Boocock 2006).
The pH and electrical conductivity of the biochar depend on both the content and composition of the mineral fraction (also referred to as the ash fraction), and this in turn depends on the type of feedstock and process conditions under which the biochar is produced (Chan and Xu 2009; Singh et al. 2010). The nutrient content of biochars is also largely influenced by the type of feedstock and pyrolysis conditions (Singh et al. 2010), whereas the availability of nutrients in biochars is related to the type of bonds associated with the element involved (De Luca et al. 2009; Yao et al. 2010). Phosphorus is mainly found in the ash fraction, with pH-dependent reactions and presence of chelating substances controlling its solubilisation (De Luca et al. 2009). Potassium in biochar is generally available to plants (Amonette and Joseph 2009). Conversely, nitrogen availability from biochars has been shown to vary widely depending on final temperature of pyrolysis, heating rate, time of holding at final temperature, and type of feedstock (Amonette and Joseph 2009). While some researchers have indicated a low N availability (Gaskin et al. 2008; Yao et al. 2010) and suggested that N is mostly present as heterocyclic N (so-called 'black N'; Knicker et al. 1996), others have observed considerable N availability from chicken litter biochars (Chan et al. 2008), where it is mainly found as nitrate on the surface of the biochars.
The mineral components that exist within the C structure of the biochar differ in the range of structural ordering (Yao et al. 2010) and in electrical and magnetic properties. Ishihara (1996) reported that wood charcoal carbonised at <300[degrees]C, 300-800[degrees]C, and >8000C acts as an insulator, a semiconductor, and a conductor, respectively. The interfaces between the mineral phase and the amorphous C can have a high defect structure, including nanoscale pores (Fig. 1), where reactions can occur preferentially. The high reactivity of the surfaces of biochar particles in soils is partly attributed to the presence of a range of reactive functional groups, some of which are pH-dependent (Cohen-Ofri et al. 2007; Cheng et al. 2008; Amonette and Joseph 2009; Cheng and Lehmann 2009; Keiluweit et al. 2010).
Methods of storage and incorporation
The initial weathering of biochar particles may occur during storage, should the biochar come in contact with moist air (Boehm 2001). This phenomenon, known as 'ageing', occurs as a result of the oxidation of exposed C rings with a high density of [pi]-electrons (Contescu et al. 1998) and free radicals (Montes-Morala et al. 2004). If biochar is mixed with decomposable organic material, such as compost, weathering reactions (e.g. C oxidation, dissolution of mineral components) during storage are enhanced (Yoshizawa et al. 2007; Dias et al. 2010).
[FIGURE 1 OMITTED]
The method for biochar incorporation into the soil may potentially modify the structure and the particle size of biochar, and this may affect the mineralisation rate, as seen for wildfire charcoal (Nocentini et al. 2010), and the water-holding capacity. Ploughing results in greater soil mechanical disturbance than other methods, such as deep banding or direct drilling. Mechanical disturbance of soils amended with biochar has been shown to promote biochar decomposition for several weeks following disruption (Kuzyakov et al. 2009). Those authors suggested that the destruction of aggregates and exposure of native organic matter to microbial attack facilitated the co-metabolic decomposition of biochar. Most of the common methods used to date (spreading and incorporation with rotary plough, deep banding) involve the incorporation of large volumes of biochar (>5 t/ha) into the soil to a depth of 60-100mm (Blackwell et al. 2009; Major et al. 2009a).
Initial reactions of blochar when placed in the soil
Little research has been undertaken to determine biochar weathering and reactions that occur within the first few weeks after application of biochar to soils (Singh and Cowie 2008; Kuzyakov et al. 2009). The ageing of biochar, which might have started before its addition to soil, continues once it is incorporated in the soil, at a rate that is partly governed by conditions of moisture (Nguyen and Lehmann 2009) and temperature (Cheng and Lehmann 2009; Nguyen and Lehmann 2009). An immediate evolution of biochar-derived C[O.sub.2] from soils has been observed within the first 2 weeks (Singh and Cowie 2008; Hilscher et al. 2009; Kuzyakov et al. 2009) after amendment and tends to decrease exponentially with time (Singh and Cowie 2008; Kuzyakov et al. 2009). As with mineral weathering, the presence of water will have a major role in processes such as dissolution, hydrolysis, carbonation and decarbonation, hydration, and redox reactions, affecting biochar weathering in soil, as well as interactions with soil biota. The rates at which these reactions occur depend on the nature of the reactions, type of biochar, and pedoclimatic conditions. These reactions are discussed below.
The dissolution and leaching of soluble salts and organic compounds present in the biochar will be among the first reactions, especially if soils are moist and there is a rain event (Shinogi et al. 2003; Major et al. 2009b). The initial dissolution of soluble salts (e.g. K and Na carbonates and oxides) may produce a pH increase in the water-film around the biochar particles. However, in leaching environments, the pH will tend to decrease as these salts are lost from the system (Yao et al. 2010); the magnitude of this decrease is determined by the intensity of the leaching and the acid-buffering capacity of the system. The pH around biochar particles would also initially increase due to the Lewis basicity of [pi]-electrons (Contescu et al. 1998) and then decrease as acidic functional groups are formed on biochar surfaces (Cheng et al. 2006; Cheng and Lehmann 2009). In high-ash biochars, the pH increase due to basic salts may be larger than the pH decrease induced by surface oxidation even over longer time scales (Nguyen and Lehmann 2009).
Over time, some organic compounds present in the biochar can be released to solution; for example, a range of biopolymers and low molecular weight compounds have been detected in the leachates of an Acacia saligna biochar (produced at 400[degrees]C with a heat treatment time of 30 min) and a biochar-mineral complex produced at low temperature (Henderson, unpubl, data; Fig. 2). The amount and type of organic compounds released are highly dependent on biochar characteristics and environmental conditions. Hockaday (2006) and Hockaday et al. (2007) detected condensed aromatic ring structures in the pore water of soils on which charcoal was deposited 100 years before, during a fire event in a mixed-hardwood forest. This contrasts with the findings of Kuzyakov et al. (2009), who did not detect presence of any dissolved organic C derived from [sup.14]C-labelled, black C residues from a rye-grass biochar that had been mixed with soil.
[FIGURE 2 OMITTED]
The precipitation reactions of inorganic compounds may be important, especially in soils subjected to wetting and drying cycles with minimal leaching. As the soil dries out, the ionic activity in solution increases. Once the ionic activity product reaches the saturation point, new precipitates are formed. These processes are enhanced within small pores, such as those present in biochars (Downie et al. 2009), as they have slow water percolation. This concentrates reaction products in solution and thus promotes precipitation (Lasaga 1998). In acid soils, a layer of either iron (hydr)oxide or alumina may deposit/ precipitate on part of the biochar (see section below: Observed changes over multiple years). In calcareous soils very low accumulations of Fe and Al oxy-hydroxides have been observed on the surface and in the pores of the biochar (N. Foidl et al., unpubl, data), which were dominated by the presence of carbonates.
Decomposing organic detritus can be regarded as an electron-pump supplying electrons to more oxidised species present in the soil system (Chesworth 2004). Biomass that has been converted to biochar is still thermodynamically unstable under the oxidative conditions of most surface soils (Fig. 3; Macias 2004; Macias and Camps Arbestain 2010), although it is evident that it remains in the soil as a meta-stable material with a much longer residence time than the original biomass from which it was formed. The chemical stability of charcoal is mainly due to the condensed aromatic structure, which is what distinguishes it from other more readily degradable aromatic substances such as lignin. Low-temperature biochars have, however, a considerable fraction of non-aromatic C (e.g. aliphatic C) (McBeath and Smemik 2009), which may make these biochars more susceptible to microbial attack, and thus to oxidation, than high-temperature biochars (Singh and Cowie 2008; Nguyen et al. 2010).
In spite of the high stability of aromatic C, it has redox activity and mainly functions as a reducing agent, [O.sub.2] being the most common electron-acceptor species. The oxidation of biochar is largely initiated through abiotic reactions boosted by the electron-donating properties of areas with a high density of [pi]-electrons (Contescu et al. 1998), and followed by the subsequent formation of O-containing functional groups at the surface of the biochar, some of them acidic in nature (see next section). The presence of free radicals in biochars--the amount being dependent on pyrolysis process conditions (Feng et al. 2004)--increases the reactivity towards oxidation, often to the point of being pyrophoric (Amonette and Joseph 2009). With increasing aromaticity, however, the stability of free radicals increases. In the absence of [O.sub.2], alternative electron acceptors (e.g. MnOOH, Mn[O.sub.2] in Fig. 3) may be able to oxidise aromatic C, as suggested by Nguyen and Lehmann (2009). Microorganisms can use aromatic compounds as the sole source of C (Hofrichter et al. 1999; Boonchan et al. 2000) or degrade them through co-metabolic decomposition (Kuzyakov et al. 2009).
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Aromatic compounds are not only potential [pi]-electron donors, but also [pi]-electron acceptors. They become stronger [pi]-donors as the number of aromatic rings increases, whereas the presence of N substituents in the rings can create either electronrich systems (e.g. indoles) or electron-depleted ones (e.g. 2,4,6-trinitrotoluene) (Keiluweit and Kleber 2009). The role of N substituents present in the different types of biochars in [pi]-[pi] electron donor-acceptor-type interactions (referred to as EDA) deserves further investigation. When n orbitals overlap with d orbitals from many transition-series metals, stable complexes are formed, and these are important for catalysis.
Overall, from a thermodynamic standpoint (Kennedy 2001), the entropy (the degree of order) and the enthalpy (heat of formation) change continuously and these systems go from one steady-state to another. The free energy of the biochar surfaces may thus change as microbial and root activity, oxygen and water content, and soil temperature change. The free energy of the internal structure may also continually change as dissolved organic and inorganic compounds move into, and out of, the macropores.
Acidic and basic surface charge of biochars
Biochar surfaces can initially have acidic as well as basic properties (Cheng et al. 2008; Amonette and Joseph 2009), which have an important influence on both the wettability of biochar particles and the surface retention of ions through electrostatic interactions (Boehm 2001). Although acidic and basic sites usually coexist, the concentration of basic sites decreases through oxidative processes as the biochar particle weathers (Cheng et al. 2008; Cheng and Lehmann 2009). The functional groups likely to be formed with this process are carboxylic, lactonic, phenolic, carbonyl, o-quinone-like structures, and ether-type oxygen (Boehm 2001). The first 3 groups have Bronsted acidic properties and provide a pH-dependent cation exchange capacity to biochar particles.
Recent studies using X-ray photoelectron spectroscopy (XPS) of the structural changes occurring in biochar particles 1 year after their application to soils showed a predominant increase of carbonyl and, to a lesser extent, of carboxylic functional groups with prolonged weathering (see section below: Observed changes over multiple years). The same trend was observed by Cheng et al. (2008), investigating the changes in the molecular form and surface charge of aged charcoal particles sampled from historical charcoal blast furnace sites. The smaller increase observed in carboxylic groups--although still substantial and having a key role in nutrient retention--compared with carbonyl groups (Cheng et al. 2008; see Observed changes over multiple years) could be due to the lower stability of the former. Once in solution, these groups can become partly decarboxylated through hydrolysis reactions (Yan et al. 1996).
While the nature of the acidic surface sites is well understood, the nature of basic sites on C surfaces is a controversial issue (Leon y Leon et al. 1992; Contescu et al. 1998; Montes-Moran et al. 2004). Traditionally, oxygen-containing groups (e.g. pyrone type, chromene, diketone, or quinone groups) and delocalised [pi]-electrons of the basal planes are assumed to have a basic nature (Contescu et al. 1998; Montes-Moran et al. 2004). While pyrone undergoes Bronsted acid-base reactions, proton transfer processes to and from the other surface groups occur simultaneously with redox reactions (Contescu et al. 1998). Stabilised free radicals present in the basal planes may also contribute to basicity (Montes-Moran et al. 2004).
Sorption of organic compounds on biochar surfaces
Once biochar is added to the soil, it can adsorb organic molecules (Pignatello et al. 2006), including residual herbicides and pesticides (Smernik 2009; Spokas and Reicosky 2009). Numerous mechanisms can be involved in the interactions of biochars with organic compounds, including H-bonding, cation-bridging, covalent bonding, and hydrophobic types of interactions. In addition, there is increasing evidence of the important role of mechanisms involving non-covalent EDA interaction between [pi]-systems of organic compounds and sorption sites at soil mineral surfaces, natural organic matter, and biochar (Keiluweit and Kleber 2009).
Reaction rates for the adsorption processes of organic molecules onto biochar particles depend on the characteristics of the adsorbate as well as on the surface properties of the adsorbent. Polar compounds are more likely to be adsorbed on surfaces with high oxygen functionality, although Zhu et al. (2005) found that these functionalities were not the principal driving force for sorption of polar compounds. The adsorptive properties of biochar particles may become attenuated with time as the external space available for adsorption reactions diminishes with the increasing presence of organic compounds at the surface (Pignatello et al. 2006). As organic and mineral matter builds up on the surface of the biochar, the type of compounds involved in surface reactions and the adsorption rates will change (Kwon and Pignatello 2005; Keiluweit and Kleber 2009). In some instances, adsorption processes might be difficult to distinguish from surface oxidation reactions (Lehmann et al. 2005; Dias et al. 2010).
Biochar-soil mineral-soil organic-matter interactions
Interactions between organic matter and clay mineral surfaces in soil are complex and depend on the type of clay (2 : 1, 1 : 1), the distribution of different functional groups on the clay (siloxane, OH) and the organic matter (COOH, C=O, C-O, CN), the polarity of these compounds, and the composition and concentration of cations and anions in solution (Kleber et al. 2007). Similar complex reactions are likely to take place on biochar surfaces, especially for those biochars that have high mineral content. Based on the literature related to organic matter and mineral interactions, and on the mechanisms described in previous sections, several mechanisms for interactions between biochar and organic matter and/or minerals in soil can be hypothesised as follows:
(i) Surface hydrophobic and hydrophilic interactions of biochar, organic compounds, and clay minerals, following the conceptual model of organo-mineral interactions proposed by Kleber et al. (2007). This occurs through direct electrostatic interactions, H bonding, cation bridging, and ligand exchange reactions in the hydrophilic zone (Yariv and Cross 2002), whereas the bi-layer formed in the hydrophobic zone is entropically driven (Kleber et al. 2007);
(ii) EDA interactions can occur between aromatic compounds (including biochar) and mineral surfaces, as well as between 2 aromatic compounds, as described in detail by Keiluweit and Kleber (2009);
(iii) Soluble organic compounds released from the biochar particles and/or from other organic matter in soil can become intercalated within 2 : 1 and 1 : 1 clay minerals. Replacement of interlayer water in smectites by neutral organic molecules and binding of organic compounds in tubular kaolinites through strong hydrogen bonds and/or strong dipole interaction to silicate layers have been reported (Lagaly 1984; Matusik et al. 2009);
(iv) Deprotonated multidentate organic acids are known to form complexes with transition metals (Violante and Gianfreda 2000) and also silicic acid (Marley et al. 1989).
Macropore filling and interactions with microorganisms, dissolved organic matter and minerals
Most biochars have a high concentration of macropores with a diameter >1 [micro]m that extend from the surface of the particle into its interior (Downie et al. 2009). After continuous irrigation or heavy rain events, a high water-potential gradient can be generated between the inner and outer part of the pore channels (Hillel 2004). Capillary forces draw soil solution into the microvoids, if biochar is not too hydrophobic, and small mineral and organic particles carried in suspension accumulate in the pore channels (Fig. 4). This process is favoured if the electrolyte concentration of soil water is low, as this enhances clay dispersion (Buurman et al. 1998), and if coarser material does not clog the pore system. The depth to which these particles penetrate the biochar depends on the diameter, tortuosity, connectivity, and length of the macropores, and also on the size of the particles. Reactions in pores can be very complicated given the heterogeneous nature of the surfaces of the solid biochar, mineral and organic phases/ compounds, and the complexity of the dissolved organic and inorganic matter in the liquid phase that surround the solids (Hammes and Schmidt 2009). These include dissolution and precipitation reactions (see Overview) and sorption reactions described in previous sections.
The porous structure of biochar is likely to provide microorganisms with a highly suitable habitat to colonise. Ogawa (1994) has noted that fungi grow from within the pores out into the soil. Thies and Rilling (2009) reported the existence of a range of microbial communities within pores. These microbes might decompose organic matter adsorbed on the biochar surface and within pores (Zackrisson et al. 1996); however, the extent of this decomposition will be mainly related to accessibility and the amount of chemical energy required for the enzymatic cleavage of chemical bonds (Kogel-Knabner et al. 2008).
Liang et al. (2010) observed that, in spite of the greater microbial biomass detected in charcoal-rich soils, these have a lower microbial metabolic quotient, and suggested the presence of a less-active fraction of microorganisms in the charcoal-amended soil than in soils without charcoal. The composition of these communities will also depend on the pH and Eh at the micro-site. As oxygen is depleted within the inner pores, facultative aerobes may begin to use anaerobic respiratory pathways, and the redox potential will decrease as the concentration of the most favourable electron acceptors is depleted. However, biochar addition to soil may increase soil aeration because of its highly porous nature, and hence decrease the total amount of anaerobic microsites per unit volume of soil.
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Interactions with plant roots and microorganisms in the rhizosphere
A complex set of reactions may occur between plant roots, microorganisms, and biochars. In most agricultural applications, biochars will have interacted with soils, following the reactions described in the previous section, before the root hairs come into contact, increasing the complexity of these interactions. Microscopic, chromatographic, and spectroscopic studies of biochar and root growth in biochar-amended systems suggest the following interactions may occur.
Once the root system encounters a biochar particle, root hairs can enter the water-filled macropores or bond onto the biochar surface (Fig. 5a, b). Biochars and minerals associated with biochar particles can adsorb organic compounds released from the growing roots, following the mechanisms described above (in Sorption of organic compounds on biochar surfaces). These include low-molecular weight organic compounds (free exudates), high-molecular-weight gelatinous material (mucilage), and sloughed-off cells and tissues and their lysates (Violante and Gianfreda 2000). These compounds can bind either to the mineral layers on the biochar or directly to the biochar surfaces. The rhizosphere constitutes an important redox interface, as the reducing and oxidising environments are intimately associated with one another (Bartlett 1999). Biochar might change the complex redox potential existing in this zone, altering reactions involving free radicals. These redox reactions include transformations of C, N, and S, which can involve reactions with Fe and Mn species.
[FIGURE 5 OMITTED]
Biochar has also been shown to adsorb toxic compounds of natural (MacKenzie and DeLuca 2006; Pignatello et al. 2006; Cheng and Lehmann 2009) and anthropogenic (Smemik 2009; Spokas and Reicosky 2009) origin (see Sorption of organic compounds on biochar surfaces) and decrease their activity. This is the case for phenolics produced during decomposition of ericaceous vegetation, the net result of which is an increase in nutrient availability through enhanced nitrification (Wardle et al. 1998; Berglund et al. 2004; MacKenzie and DeLuca 2006). High concentrations of phenolic acids, such as p-coumaric and p-hydroxybenzoic, have also been reported to impair root elongation and affect the metabolism of IAA, a major auxin (Marschner 1999). The activity of other organic compounds, such as tyrosinase, a phenol oxidase enzyme that assists in the breakdown of organic matter, has been suggested to be promoted in the presence of highly microporous charcoal and Fe and Mn oxides (Amonette et al. 2006). Biochar may have a more fundamental effect on plant root and soil microbe growth, through a range of organic molecules that could be present in the volatile fraction of biochars. Some of these compounds (e.g. butenolide) can trigger germination (Dixon 1998), while others (e.g. sesquiterpenes; Bridgwater and Boocock 2006) can promote the growth of mycorrhizal fungi (Amonette and Joseph 2009). Auxofuran could play a major role in assisting plant uptake of nutrients (Riedlinger et al. 2006).
Root redox potential and membrane-associated redox activities influence a wide range of other cell functions, including the activity of many enzymes, the expression of many genes (Quaggiotti et al. 2004), and the activity of cytosolic redox buffers and several plasma membrane oxidoreductases (Arvieu et al. 2003). Changes in redox potential may also influence the plant growth regulator, IAA, in the region of the embryo that will become the root (Tarkka et al. 2008). Biochar is likely to adsorb the ethylene produced by the plants, as has been suggested in sand cultures using activated charcoal (Michael 2001). Conversely, some evidence for the emission of ethylene by biochars suggests a more active role of biochar in inhibiting the nitrification process in soils that leads to nitrous oxide emissions (Spokas and Reicosky 2009). More information is needed from biochar-amended soil studies.
Possible negative effects of biochar addition
Little is known about the potential toxic effects of biochar in soil, and specifically of those related to the presence of heavy metals and plant-available organic compounds that have condensed on the surfaces of biochar during their manufacture. A discussion of possible effects is given in Lehmann and Joseph (2009). The following is a brief summary.
Condensates on the surface of biochars may contain compounds such as polycyclic aromatic hydrocarbons, cresols, xylenols, formaldehyde, acrolein, and other toxic carbonyl compounds that can have bactericidal or fungicidal activity (Painter 2001). However, Ogawa (1994) has shown that these substances can, and do, serve as C and energy sources for selected microbes. McClellan et al. (2007) found that residual volatiles on biochar made using a flash carboniser proved toxic to plants. When the biochar soaking time was increased these toxic effects were removed.
Observed changes over multiple years
Biochars produced at 450[degrees]C (chicken litter) and 550[degrees]C (papermill waste and greenwaste) were characterised before application and at 1 and 2 years atter application to a Ferrosol (Australian Soil Classification System) in a subtropical environment at Wollongbar (NSW, Australia, 28[degrees]50'S, 153[degrees]25'E). A summer sweet corn-winter pulse rotation was established at the site. The feedstocks, pyrolysis conditions, sampling procedures, and analytical techniques are described in the Accessory publication (available on journal website) and are also reported in Chan et al. (2008) and Van Zwieten et al. (2010).
The surfaces of the weathered biochar particles were found to be heterogeneous, covered with a wide variety of mineral and organic matter, and very different from the surfaces of the fresh biochars, as shown in the SEM images in Fig. 6. At these surfaces, high concentrations of clay minerals consisting of various combinations of Al, Si, C, Fe, and Ti, and trace amounts of Ca, Mg, Mn, K, Na, P, and S, were found using detailed electron microprobe analysis of cross-sectioned samples of a greenwaste biochar (Fig. 7). This figure demonstrates that most of the inorganic phases are located at the external surfaces of a greenwaste biochar particle. The P-rich components are mostly associated with high Fe concentrations, but also with high Al and Si, suggesting either precipitation (e.g. Fe, Al-phosphates) and/or sorption interactions on Al and Fe oxy hydroxides and aluminosilicate minerals. Nitrogen, S, and Ca compounds were identified on the inner fiat surface of the weathered greenwaste biochar. This could indicate that a range of calcium polysaccharides, proteins, and amino acids have complexed on the surface. Overall, examination of sectioned samples of the different biochars using SEM and transmission electron microscopy (TEM) (data not shown) indicated that only some of the pores had a range of organic and mineral matter after the first year, whereas most of the pores were filled after a second year in the soil.
[FIGURE 6 OMITTED]
[FIGURE 7 OMITTED]
Energy dispersive X-ray spectrometry (EDS) analysis, in conjunction with scanning TEM was performed across the organomineral/biochar interfaces to examine both the pores and the surfaces of the 3 biochars aged for 1 year in soil. Figure 8 shows a cross-section of an organo-mineral particle complexed onto the surface of a chicken litter biochar particle (aged 1 year). EDS measurements showed a phase rich in Al and O within a pore of the C matrix adjacent to the interface. In this region, the phase was also rich in C and had small amounts of Ca and Si. Adjacent to this phase, there was a phase containing a mixture of Al and Fe oxy-hydroxides and aluminosilicate particles. These inorganic particles appeared to be associated with C compounds and had a significant Ca content. The region was also enriched in P, similar to results obtained with greenwaste biochar using the electron microprobe (Fig. 7). The EDS maps and line scans obtained for the aged papermill waste biochar (data not shown) also revealed a range of Al/Si/O and Fe phases bonded to the amorphous C phase, which were not seen in the fresh papermill waste biochar. The Fe concentration in the papermill waste was also lower than that observed in the other biochars.
Results of XPS analysis of the surfaces (Table 1) indicate that the fresh greenwaste biochar and poultry manure biochar consisted only of C and O, with O/C ratio <0.2. After 1 and 2 years in the soil, the C content of this biochar decreased. The total surface O and O bound to C increased, leading to O/C ratios of ~0.75. The aged greenwaste biochar showed the largest decrease in surface C, whereas the aged papermill waste had the highest C content among the 3 aged biochars. With respect to C bonding states (Table 1), C-C/C-H/C=C (~284.9 eV) was the major component in all fresh and aged biochars, but the relative content of these bond types sharply decreased after application to soil. The amount of C in more oxidised states was increased, which indicated that the oxidation of the surface and/ or adsorption of soil organic matter had occurred. There was little difference between the first and second year for the relative oxidation status of the C species (Table 1). The surfaces of all of the aged biochars had increased N content mainly associated with proteins and amino acids (~400.2eV) and N[H.sub.4.sup.+] (~401.5 eV) and N-C (~399.5 eV) compounds (Table 2). The greatest increase in surface N content was in the papermill waste biochar; the high content of Ca on the papermill waste surfaces may result in Ca bridging with the nitrogenous compounds in the soils.
[FIGURE 8 OMITTED]
The [sup.13]C CP-NMR spectra of the fresh and 1-year weathered biochars (Fig. 9) shows that the weathered PM biochar contains 2 additional peaks in the O-alkyl area, compared with the fresh biochar, suggesting either incomplete separation of biochar from soil organic matter or the existence of organo-biochar interactions at the surface of the biochar particle. The weathered greenwaste biochar is very similar to the original biochar, except for a slightly less prominent O-aryl shoulder (~150ppm) and slightly less alkyl C (~30ppm). This could be related to a faster decomposition of the non-aromatic C fraction. Differences in NMR spectra between the fresh and the weathered chicken litter biochar are most likely due to the presence of artefacts from soil particles in the latter. Finally, the degree of aromatic condensation of the fresh and aged biochars were determined using the method of McBeath and Smemik (2009), which is based of the effect of biochar ring currents on sorbed [sup.13]C-benzene. This indicated that the papermill waste and chicken litter biochars taken from the soil after 1 year had a lower degree of aromatic condensation than for the corresponding fresh papermill biochars, whereas the greenwaste biochar showed the opposite trend (Fig. 10). This suggests that ageing of papermill and chicken litter biochars results in biochar surfaces with a lower degree of aromatic condensation, through oxidation of the surface or precipitation or sorption of inorganic or organic materials on the surface.
Biochar and soil fauna
Some worms, termites, larvae, and other insects appear to ingest or live inside biochar (Fig. 11) breaking it up and/or coating it with organic compounds. Breaking up the biochar can result in exposed surfaces being oxidised and then reacting with mineral or organic matter as discussed in Initial reactions of biochar when placed in the soil. Eckmeier et al. (2007) noted that earthworms ingest particles <2 mm (but do not digest them) and redistribute them elsewhere in the profile (concentrated at ~0.8 m depth) by excretion, as shown by thin sections of soil with small biochar particles present in earthworm casts. Bioturbation can thus have an important role in the physical mixing of biochar particles and its re-distribution within the soil profile. Over time, biochar will tend to move down through the soil profile (Major et al. 2009a) to areas where microbial activity is lower.
Conclusions: towards an understanding of the dynamics of biochar reactions
The limited data available indicate different biochars undergo a range of complex reactions in soil that vary depending on the particular properties of the biochars and soils involved. Factors such as nutrient composition and availability, organic matter content, soil mineralogy and texture, pH/Eh conditions, presence of toxins, soil biota, the type of plants grown, the proximity of the biochars to the rhizosphere, and the temporal variation in soil moisture can all impact the nature of biochar reaction.
Although not conclusive, data from both pot and field trials suggest the following series of time-sequenced reactions takes place:
1. When high mineral ash biochars are added to moist soil (or are followed by a rain event) there is a change in the pH, EC, and Eh around the particle, probably within the first week, as the minerals dissolve and/or ions are exchanged on the surfaces of the surrounding clay particles.
2. For high-porosity biochars, a rain event can also result in the inclusion and enhanced surface interaction of a range of soil mineral and organic compounds with the biochar particles (Fig. 4). The limited data available indicate that the type of reactions that take place and the stability of the compounds retained on the surface depend on the type of biochar and the conditions under which it was produced (Spokas and Reicosky 2009; Nguyen et al. 2010). Any significant reaction involving the biochar, the soil mineral, and dissolved organic matter could lead to a substantial change in the properties of the biochar (e.g. pH and EC).
3. When biochars produced at low temperatures are added to moist soils, a considerable quantity of soluble organics can be released to the soil solution (Fig. 1). As noted by Dixon (1998) and Light et al. (2009), some of the labile organic compounds on the surfaces of biochar can stimulate both seed germination and growth of fungi. Little is known, however, about the toxicity of these compounds.
4. The enhanced C[O.sub.2] emissions observed during the first months after biochar addition to soil are partly attributed to biochar surface oxidation. There are several different biotic and abiotic reaction mechanisms involved. High mineral ash biochars have been shown to oxidise faster than low mineral ash biochars. It is possible that the mineral matter in the biochar provides nutrients for microorganisms to grow faster as well as catalysing the breakdown of organic matter (Amonette et al. 2006). The higher the temperature of pyrolysis, the higher the resistance of the C in biochar against microbial decomposition due to the increased condensation in aromatic C.
5. Following surface oxidation of biochars, the potential for hydrophilic interactions of biochars with a range of soil organic and inorganic compounds increases. This is more significant in high mineral ash biochars (Lima and Marshall 2005). Greater reactivity of biochars with mineral matter could further promote physical protection of biochar and, thus, long-term stability (Brodowski et al. 2006).
6. Once roots and root hairs interact with the biochars, a much wider range of reactions can occur, through the uptake of nutrients, and the release of root exudates, which enhances both complexation reactions and microbial activity in the rhizosphere.
[FIGURE 9 OMITTED]
[FIGURE 10 OMITTED]
[FIGURE 11 OMITTED]
[FIGURE 12 OMITTED]
Figure 12 summarises the reactions that can take place in and on the surfaces of biochar. The data obtained from microscopic and spectroscopic examination of the surfaces have been used to posit the possible mineral and organic structures that can occur during ageing of biochar. Formation of micropores, mesopores, and macropores has been identified between and within the different organo-mineral phases. As more data are obtained from detailed examination of aged biochars of known properties, this model will be refined. Greater understanding of these interactions will permit more accurate estimation of long-term stability and therefore climate change mitigation benefit from application of biochars.
S.J. is very grateful for the support of VenEarth LLC and his colleagues at Anthroterra Pty. M.C.A. is very grateful for financial support from the Ministry of Agriculture and Forestry of New Zealand. We acknowledge Professor Felipe Macias, from the Universidade de Santiago de Compostela, for his assistance in producing Fig. 3. The authors also thank the anonymous reviewers for their valuable suggestions.
Manuscript received 5 January 2010, accepted 24 May 2010
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S. D. Joseph (A,K), M. Camps-Arbestain (B), Y. Lin (A), P. Munroe (A), C. H. Chia (A), J. Hook (C), L. van Zwieten (D), S. Kimber (D), A. Cowie (E), B. P. Singh (F), J. Lehmann (G), N. Foidl (H), R. J. Smernik (I), and J. E. Amonette (J)
(A) School of Material Science and Engineering, University of NSW, Sydney 2052, Australia.
(B) New Zealand Biochar Research Centre, Private Bag 11222, Massey University, 4442 Palmerston North, New Zealand.
(C) NMR Facility, Analytical Centre, University of NSW, Sydney, NSW 2052, Australia.
(D) lndustry and Investment NSW, Wollongbar, NSW 2477, Australia.
(E) National Centre for Rural Greenhouse Gas Research, University of New England, Armidale, NSW 2351, Australia.
(F) Forest Science Centre, Industry and Investment NSW, PO Box 100, Beecroft, NSW 2119, Australia.
(G) Department of Crop and Soil Sciences, College of Agriculture and Life Sciences, Cornell University, Ithaca, NY 14853, USA.
(H) Venearth LLC, San Francisco, USA.
(I) School of Agriculture, Food and Wine, DP 636, The University of Adelaide, Adelaide, SA 5000, Australia.
(J) Pacific Northwest National Laboratory, Richland, WA 99354, USA.
(K) Corresponding author. Email: email@example.com
Table 1. XPS results of fresh and aged greenwaste biochar after 1 and 2 years of weathering in an Australian Ferrusol n.d., Not detected Transition Assigned Peak energy (eV) structure Fresh 1 year 2 year OIs O-C 533.42 n.d. n.d. Chemisorbed oxygen? 531.46 n.d. n.d. Silicates n.d. 532.84 n.d. Subtotal OIs CIs C-O 288.30 288.08 287.77 C O 286.50 286.58 286.24 C~C/C-H 284.96 284.89 284.87 COOH n.d. 289.17 289.71 Subtotal CIs NIs N-C-COOH n.d. 400.51 400.25 N[H.sub.4.sup.+] n.d. n.d. 401.49 Subtotal NIs Organics only Subtotal organics Si2p Silicates n.d. 103.02 103.05 AI2p Silicates n.d. 74.61 74.90 Fe2p3/2 FeS? n.d. 712.84 713.05 Total Transition Assigned Abundance (atom %) structure Fresh 1 year 2 year OIs O-C 10 0 0 Chemisorbed oxygen? 6 0 0 Silicates 0 47 0 Subtotal OIs 15 47 19 CIs C-O 2 6 12 C O 14 12 20 C~C/C-H 69 14 36 COOH 0 3 6 Subtotal CIs 85 34 74 NIs N-C-COOH 0 3 1 N[H.sub.4.sup.+] 0 0 0 Subtotal NIs 0 3 1 Organics only Subtotal organics 100 84 94 Si2p Silicates 0 7 2 AI2p Silicates 0 8 3 Fe2p3/2 FeS? 0 1 0 Total 100 100 100 Transition Assigned Organics only structure Fresh 1 year 2 year OIs O-C 10 0 0 Chemisorbed oxygen? 6 0 0 Silicates 0 56 0 Subtotal OIs 15 56 21 CIs C-O 2 7 13 C O 14 14 22 C~C/C-H 69 16 38 COOH 0 4 6 Subtotal CIs 85 40 78 NIs N-C-COOH 0 3 1 N[H.sub.4.sup.+] 0 0 0 Subtotal NIs 0 3 1 Organics only Subtotal organics 100 100 100 Si2p Silicates AI2p Silicates Fe2p3/2 FeS? Total Transition Assigned CIs only structure Fresh 1 year 2 year OIs O-C Chemisorbed oxygen? Silicates Subtotal OIs CIs C-O 2 16 17 C O 17 34 28 C~C/C-H 81 41 48 COOH 0 9 8 Subtotal CIs 100 100 100 NIs N-C-COOH N[H.sub.4.sup.+] Subtotal NIs Organics only Subtotal organics Si2p Silicates AI2p Silicates Fe2p3/2 FeS? Total Table 2. XPS results for the O1s, C1s, and N1s core level of fresh and aged chicken litter biochars after l and 2 years of weathering in an Australian Ferrosol n.d., Not detected Transition Assigned Peak energy (eV) structure Fresh 1 year 2 year O1s O-C 533.71 n.d. 533.01 Silicates? 532.33 532.30 n.d. Subtotal O1s C1s C=O 288.01 288.01 288.32 C-0 286.61 286.50 286.72 C-C/C-H 285.01 284.94 285.05 COOH n.d. 289.24 289.60 Carbonate 290.58 n.d. n.d. Subtotal CIs N-C 399.10 398.86 399.49 N1s N-C-COOH n.d. n.d. 400.74 N[H.sub.4].sup.(+)] 402.12 400.20 402.10 Subtotal NIs Organics only Subtotal organics Transition Assigned Abundance (atom %) structure Fresh 1 year 2 year O1s O-C 9 0 32 Silicates? 19 42 0 Subtotal O1s 25 42 32 C1s C=O 4 6 9 C-0 10 15 18 C-C/C-H 36 17 22 COOH 2 2 6 Carbonate 3 0 0 Subtotal CIs 55 40 55 N-C 1 1 1 N1s N-C-COOH 0 0 2 N[H.sub.4].sup.(+)] 0 3 1 Subtotal NIs 1 4 4 Organics only Subtotal organics 81 86 87
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|Author:||Joseph, S.D.; Camps-Arbestain, M.; Lin, Y.; Munroe, P.; Chia, C.H.; Hook, J.; van Zwieten, L.; Kimbe|
|Publication:||Australian Journal of Soil Research|
|Date:||Sep 1, 2010|
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