reducing capacity of water extracts of biochars and their solubilization of soil mn and fe

European Journal of Soil Science, 2013 doi: 10.1111/ejss.12071

Reducing capacity of water extracts of biochars and theirsolubilization of soil Mn and Fe

E . R . G r a b e ra , L . T s e c h a n s k ya , B . L e wb & E . C o h e na,c

aInstitute of Soil, Water and Environmental Sciences, The Volcani Center, Agricultural Research Organization, POB 6, Bet Dagan, 50250,Israel, bInstitute of Agricultural Engineering, The Volcani Center, Agricultural Research Organization, POB 6, Bet Dagan, 50250, Israel,and cDepartment of Soil and Water Sciences, The Robert H. Smith Faculty of Agriculture, Food and Environment, Hebrew University ofJerusalem, Rehovot, 76100, Israel

Summary

Biochar, being produced in an oxygen-restricted environment, is chemically more reduced than the originalfeedstock. Consequently, it was hypothesized that reduced biochar components could participate in redox-mediated reactions in the soil. This hypothesis was tested by measuring the reducing capacities of aqueousextracts of biochars and the reduction and solubilization of soil Mn and Fe oxides by the extracts. Thereduction capacity of extracts from biochars produced from three feedstocks (eucalyptus wood, EUC; olivepomace, OP; and greenhouse waste, GHW) at different highest pyrolysis treatment temperatures (HTT; 350,450, 600 and 800◦C) was less for the EUC feedstock than the others, and was greater for biochars produced atlower HTTs. The organic fraction of the extracts apparently was responsible for the major part of the reducingcapacity. Extracts of smaller-HTT biochars, having greater dissolved organic carbon (DOC) contents, hadgreater reducing capacities than extracts of larger-HTT biochars from the same feedstock. Extracts of two GHWbiochars (GHW-450 and GHW-600) solubilized Mn and Fe from soils at pH values below 8. The extract withthe greater reducing capacity (GHW-450) solubilized both metals to a significantly greater extent. Smaller-HTTbiochars produced from agricultural wastes, having a greater variety and concentration of soluble reducingagents, are expected to have more impact on soil redox reactions than larger-HTT biochars. By participatingin chemical and biological redox-mediated reactions in the soil, biochar could influence microbial electronshuttling, nutrient cycling, pollutant degradation, contaminant mobilization and abiotic formation of humicstructures.

Introduction

Different types of biochar used along with organic and inorganic

fertilizers can sometimes improve crop productivity (Lehmann

et al., 2003), enhance soil aggregate structure (Liang et al.,

2010), alter soil microbial populations (Kolton et al., 2011) and

promote plant resistance to disease (Elad et al., 2010; Meller Harel

et al., 2012). Some types of biochars have been suggested as

contributing to plant growth directly as a result of their nutrient

content and nutrient release characteristics (Silber et al., 2010).

Biochars also have been thought to impact plant growth indirectly

by increasing nutrient retention in the soil (Lehmann et al., 2003),

improving soil pH (Yuan & Xu, 2011), raising soil cation exchange

capacity (Silber et al., 2010), improving soil water retention

characteristics (Laird et al., 2010) and promoting beneficial soil

Correspondence: E. R. Graber. E-mail: [email protected]

Received 30 May 2013; revised version accepted 30 May 2013

microbes (Graber et al., 2010). It was also speculated thatbiochar-borne organic compounds could promote plant growthand health directly via chemical hormesis, or indirectly viatheir impact on the soil microbial community (Graber et al.,2010). However, the controlling mechanisms responsible for thesevarious effects remain largely unknown.

Biochar is produced in an oxygen-restricted environment, andits polycondensed aromatic carbon structure and other components(mineral phases and organic molecules that are not integralparts of the condensed carbon framework) are in relativelyreduced states compared with their states in the initial feedstock.We hypothesized therefore, that biochar could take part in awide range of chemical and biological redox-mediated reactionsin the soil and, in this way, influence important processesalong the soil–microbe–plant continuum. These processes includemicrobial electron shuttling, nutrient cycling, root uptake ofnutrients, free radical scavenging, abiotic formation of humicstructures, pollutant degradation and contaminant mobilization or

© 2013 The AuthorsJournal compilation © 2013 British Society of Soil Science 1

2 E. R. Graber et al.

immobilization (Bartlett & James, 1993). Losses of chromiumas Cr(VI) from suspensions of artificially contaminated soils anddecreased chromate-induced toxicity in sunflowers have beenattributed to biochar involvement in redox processes (Choppalaet al., 2012). Promotion by biochar of reductive transformationsof dinitro-herbicides and nitro explosives in the presence of areducing agent has also been observed (Oh et al., 2013).

The water-soluble fraction of many biochars consists ofa variety of soluble salts, colloidal minerals, small organicmolecules belonging to various chemical classes and largemacromolecular organic compounds similar in character to humicsubstances. A number of the small and large organic constituentsmaking up the water-soluble fraction of biochar (Graber et al.,2010; Lin et al., 2011) are the same as or similar to those identifiedin natural dissolved organic matter (DOM) as being redox-active(Fimmen et al., 2007). Such species (bearing quinoid, aromaticand thiol moieties) are important intermediaries in microbialmetabolic processes and facilitate biological cycling of metals withmultiple oxidation states (Lovley et al., 1998). As a consequence,they influence both microbial ecology and function (Visser,1985). Dissolved OM components having multiple oxidationstate heteroatom-containing structures (N or S) or inner-spheremetal-organic charge transfer complexes can also participatein abiotic electron transfer processes (Fimmen et al., 2007).Phenolic compounds with hydroxyl groups in the ortho- and para-position can chemically reduce manganese (Mn) and iron (Fe)oxides under normal soil conditions (Pohlman & McColl, 1989),while carboxylate moieties accelerate chemical oxidation, thusencouraging redox cycling. Moreover, some biochar-borne organiccompounds are ligands possessing multiple carboxyl, phenol,alcohol or enol groups (Graber et al., 2010; Lin et al., 2011),known to form stable metal–organic complexes with metalshaving different oxidation states. Formation of water-solublemetal–organic complexes can increase the concentration of metalsin the aqueous phase and their bioavailability, while formation ofwater-insoluble metal–organic complexes can increase the soilorganic matter content.

The goal of this study was therefore to evaluate the possibilitythat aqueous extracts of various biochars have significant redoxactivity and may solubilize Mn and Fe from soils.

Materials and methods

Biochars

Biochars were produced from three feedstocks: (i) greenhousewastes (GHW) consisting of pepper plant residues, (ii) olivepomace (OP) residues from olive oil production and (iii)eucalyptus (EUC) wood chips. The biochars were preparedin-house at different highest treatment temperatures (HTT; allfeedstocks at 350, 450 and 600◦C, and EUC additionally at 800◦C)in a slow pyrolysis reactor (BEK, All Power Labs, San Francisco,California, USA) operated in indirect retort mode. The biocharsare designated according to the abbreviation used for the feedstockand the HTT: e.g., GHW-350 was produced from greenhouse

waste at an HTT of 350◦C. Biochars were ground and sievedto a powder of < 0.5 mm particles and stored in sealed containers.Ash content (six replicates) was determined by mass loss of oven-dry (105◦C) biochars after heating to 500◦C in air for 12 hours,followed by cooling to ambient conditions. Sample residueafter loss on ignition was digested to dryness in concentratedHNO3 at 150◦C, digested a second time in 1:4 concentratedHNO3:concentrated H2O2, dissolved in HNO3, and analysed forelements by inductively coupled plasma (ICP-AES; ARCOS SOP,Spectro Analytical Instruments, Inc., Mahwah, New Jersey, USA)(Enders & Lehmann, 2012). Biochar elemental analysis (C, H,and N) was determined in triplicate by an EA-1112 ElementalAnalyzer (Thermo Finnigan, Cambridge, Massachusetts, USA),with O being calculated by difference. Biochar specific surfacearea (SSA) was determined following degassing at 120◦C for5 hours by N2-BET adsorption using a Quantachrome MonosorbII instrument (The Israel Ceramic and Silicate Institute, TechnionCity, Haifa, Israel). Biochar characteristics (HTT, SSA, CHNO,elemental ratios, ash content and ash composition) are given inthe Supporting Information (Table S1).

Aqueous extracts of the ten biochars were prepared in triplicateby shaking 2.5 g biochar with 50 ml deionized water (hereafter,water) in 50 ml polypropylene centrifuge tubes without headspacein the dark for 24 hours, followed by sedimentation and filtrationof the liquid solution via membrane filters of 0.22-μm poresize (Durapore PVDF membrane, Millipore Corp., Carrigtwohill,Ireland). Extracts were characterized for dissolved organic carbon(DOC; TOC-VCPN, Shimadzu Corp., Kyoto, Japan), redoxpotential (ORP probe), pH, electrical conductivity (EC), totalphenols and reducing capacity. Other than preparing the extractsin sealed tubes in the absence of headspace, no attempt wasmade to control the atmosphere in contact with the extractsor to exclude oxygen. Reported reduction capacities do notrepresent the potential capacity of fully reduced species but rather,native reducing capacities. Pertinent chemical characteristics ofthe extracts are given in Table 1.

Additionally, larger volumes of aqueous extracts of twobiochars (GHW-450 and GHW-600) were prepared for use inthe experiments involving metal solubilization from soil. Thiswas done by pre-wetting 100 g of biochar with approximately60 ml water, which was then placed in the centre of a large pre-wetted square of Whatman #1 filter paper. The filter paper wasgathered around the wetted biochar and tied with cotton stringto make a ‘biochar bag’. The biochar bag was suspended inone litre of water stirred by magnetic stirrer for 24 hours. Thebag was periodically kneaded gently to ensure contact of thewater with the biochar in the bag. After 24 hours, the aqueoussolution was successively vacuum-filtered through increasinglyfiner filter papers and membranes, with the final filter poresize being 0.22 μm (Durapore PVDF membrane, Millipore Corp.,Carrigtwohill, Ireland). These solutions were characterized asbefore, and additionally by gas chromatograph/mass spectrometer(GC/MS) for organic species, and by ICP-AES or AAS forelements (details below).

© 2013 The AuthorsJournal compilation © 2013 British Society of Soil Science, European Journal of Soil Science

Metal reduction and solubilization by biochar extracts 3

Table 1 Pertinent chemical characteristics of biochar aqueous extracts

Biochar DOCa / mg l−1

Redox potentialb (E h)/ mV pH EC / mS cm−1

Total phenols/ μmol GAE l−1

Total phenols/ μmol GAE mg DOC−1

EUCc-350 74 119 6.9 0.81 17.5 0.236EUC-450 44 90 8.8 1.01 8.2 0.186EUC-600 28 132 9.7 0.77 < MDLd < MDLEUC-800 24 −17 10.6 1.15 < MDL < MDLOPe-350 371 23 9.4 3.22 126 0.338OP-450 404 27 9.7 3.94 149 0.368OP-600 85 32 10.1 5.86 10.8 0.127GHWf-350 248 13 9.9 7.31 133 0.537GHW-450 427 5 9.7 8.51 204 0.478GHW-600 42 −29 10.7 7.71 6.2 0.148SRNOM 520 n.a. n.a. n.a. 434 0.835SRFA 520 n.a. n.a. n.a. 610 1.17PPHA 520 n.a. n.a. n.a. 414 0.796PPFA 520 n.a. n.a. n.a. 217 0.417Water 2 443 4.7 0.001 < MDL < MDL

aDOC, dissolved organic carbon. Reproducibility of DOC analysis is better than 7%.bRedox potential (E h) – reproducibility is better than 10%.cEUC, biochar made from eucalyptus wood chips at highest treatment temperature (HTT) in ◦C as specified by number following EUC-.dMDL, method detection limit. The MDL for total phenols is 3 μmol GAE l−1, and the reproducibility (three replicate analyses) for a given sample isbetter than 5%.eOP, biochar made from olive pomace at highest treatment temperature (HTT) in ◦C as specified by number following OP-.fGHW, biochar made from greenhouse waste at highest treatment temperature (HTT) in ◦C as specified by number following GHW-.Notes: For biochar extracts, properties were measured in 1:20 mass:volume biochar:deionized water mixed for 24 hours and filtered with a 0.22-μm filter.EC, electrical conductivity; GAE, gallic acid equivalents; n.a., not applicable; PPFA, Pahokee Peat Fulvic Acid; PPHA, Pahokee Peat Humic Acid; SRFA,Suwannee River Fulvic Acid; SRNOM, Suwannee River Natural Organic Matter.

Natural dissolved organic matter (DOM) standards

For purposes of comparing the reducing capacities of biocharaqueous extracts to those of natural dissolved organic matter(DOM) solutions, four DOM standards from the InternationalHumic Substances Society (IHSS; Saint Paul, Minnesota, USA)were used: Suwannee River Natural Organic Matter (SRNOM; cat.# IN101), Suwannee River Fulvic Acid (SRFA; cat. # 1R101F),Pahokee Peat Humic Acid (PPHA; cat. # 1S103H), and PahokeePeat Fulvic Acid (PPFA; cat. # 2S103F). Stock aqueous solutions(1 mg DOM ml−1) were made up in water and diluted as needed.PPHA was dissolved in 1 mm NaOH (pH 9) and then neutralizedwith 100 mm HCl.

Soils

Four soils of differing textures, representing major soil types inIsrael, were collected from the cultivated layer (0–25 cm) andused in this study: (i) sand (Typic Xerochrept) from Bet Dagan,the Coastal Plain; (ii) light clay (Chromic Haploxeralf), fromHefetz Haim, in the Pleshet Plains; (iii) clayey loam (CalcicHaploxeralf), from Nahal Oz, in the Northern Negev; and (iv)heavy clay (Chromic Haploxeralf), from Eilon, in the WesternGalilee (the soil classification is that of Soil Survey Staff, 1999).The soils were air-dried, crushed, sieved (< 1 mm), and stored inclosed containers under ambient laboratory conditions. Soil texture

was determined by the hydrometer method (Bouyoucos, 1962)according to the international classification system, taxonomywas specified according to the United States Department ofAgriculture (USDA) soil taxonomic system, soil organic matter(SOM) was determined by the Walkley-Black method (Walkley,1947), pH was determined in the supernatant of a 1:10 w:vsoil:water slurry mixed for 24 hours, and cation exchange capacity(CEC) and % CaCO3 following standard methods (Nelson, 1982;Rhoades, 1982). Selected soil data are tabulated in the SupportingInformation, Table S2.

Total phenols and reducing capacity

Total phenols in the biochar aqueous extracts and DOM solutionswere determined with the Folin-Ciocalteu (FC) assay (Singletonet al., 1999). The FC reaction involves oxidation of phenols andother compounds (Singleton et al., 1999) by a phosphomolybdictungstic acid reagent in a basic medium (pH 10), resulting inthe formation of superoxide ion. The superoxide ion reacts withmolybdate to form molybdenum oxide (MoO4

+) which has avery intense absorbance at 725 nm. The FC reaction mechanismis thought to involve single electron transfer according to thereduction half-reaction Mo(VI) + e → Mo(V). The FC assay,conducted at pH 10, has fast reaction kinetics for phenolic speciesbecause the kinetics of phenol electron donating, via the phenolateanion, is strongly favoured at the high pH (Singleton et al., 1999).



Following convention for reporting total phenols by the FC assay,gallic acid was used to develop a standard curve, and total phenolresults are reported in gallic acid equivalents (GAE) as μmol GAEml−1 extract. FeSO4·H2O standards were also measured by the FCassay to calculate reducing capacity in terms of moles of chargetransfer (molc), where each mole of FeSO4 represents one moleof charge transfer.

The ferric reducing antioxidant power (FRAP) assay (Benzie &Strain, 1996) was also used to determine reducing capacity. TheFRAP assay measures the reduction of the ferric tripyridyltriazinecomplex, Fe(III)-TPTZ, to the ferrous Fe(II)-TPTZ complex at pH3.6, according to the reduction half-reaction Fe(III) + e → Fe(II).Fe(II)-TPTZ has an intense blue colour and is monitored byabsorption at a wavelength of 593 nm and quantified usingFeSO4·H2O standard solutions. The FRAP assay is conventionallydetermined after only 4 minutes reaction time, as it was developedwith compounds which exhibited very fast reaction kinetics.However, the reaction kinetics for many phenols and polyphenolscan be slow because the kinetics of phenol electron donatingis sluggish at the low pH of the assay (Singleton et al., 1999).Therefore, we followed the FRAP reaction for samples andstandards for 240 minutes. Over this time, there was no significantchange in the absorption of the standards. As with the FCassay, the FRAP assay mechanism is nonspecific single electrontransfer, but the two assays have different standard reductionpotentials (E h =+0.43 V and +0.70 V for the reduction half-reactions involved in the FC and FRAP assays, respectively,(http://www.av8n.com/physics/redpot.htm)). Standard reductionpotentials of several humic acid samples were reported to rangebetween +0.15 and −0.30 V (Aeschbacher et al., 2011).

Gas chromatograph/mass spectrometer (GC/MS) analysisof extracts

Aliquots (2 ml) of the GHW-450 and GHW-600 extracts usedfor the solubilization experiments were lyophilized and thedried residue subjected to a two-stage derivatization procedureimmediately prior to analysis by GC/MS. The residue was firstderivatized for 2 hours at 37◦C using 40 μl of methoxyamine inpyridine (20 mg ml−1) to stabilize carbonyl moieties. Followingmethoxyamination, functional groups such as -OH, -COOH, -SHand -NH groups were converted into trimethylsilyl (TMS)-ethers,TMS-esters, TMS-sulfides or TMS-amines, respectively, using70 μl of N-methyl-N-(trimethylsilyl)trifluoroacetamide (MSTFA)by heating for 30 minutes at 37◦C. Following derivatization,the samples were diluted with 125 μl pyridine and filteredby Teflon membrane syringe filter (0.45 μm) to remove solids(inorganic salts in the aqueous extracts). The derivatized extractswere analysed by quadrupole GC/MS (Agilent Technologies,Santa Clara, California, USA) using electron impact ionization.Separation was achieved on a 30-m long, VF-5M, 0.25 μm film,capillary column (Varian Inc., Palo Alto, California, USA) usingthe following conditions: initial oven temperature 60◦C, initialhold 15 minutes, ramp 5◦C minute−1 to a final temperature of

330◦C, final hold 5 minutes, MS source 230◦C, injector 230◦C;MS Quad 150◦C; m/z range 50–600.

Total ion chromatograms (TIC) were analysed with freelyavailable deconvolution software (Automated Mass SpectralDeconvolution and Identification System (AMDIS), developed bythe National Institute of Standards & Technology (NIST)) at aminimum match probability of 75% and retention indices based ona series of alkanes. The deconvoluted mass spectra were comparedwith a specialty library for plant, animal and microorganismmetabolites (Golm Metabolite Database) freely provided by theMax Planck Institute for Metabolic Plant Physiology (Golm,Germany) and with the NIST08 mass spectral library.

Mn and Fe solubilization from soil

Solubilization of Mn and Fe from the sandy soil, and Mn fromthe other three soils, was compared for two biochar aqueousextracts (GHW-450 and GHW-600) and water as a function ofpH, which was modified with buffers as detailed below. Soil(2 g) was weighed into 50-ml disposable polypropylene centrifugetubes equipped with screw caps. Buffer (10 ml) and aqueousbiochar extracts or water (10 ml) were added to the tubes, andthe suspensions (or soilless controls) were shaken in the dark for24 hours on a horizontal table shaker at 120 cycles minute−1. Eachpoint (samples and controls) was made in triplicate; average valuesand bars denoting standard error are given in the Figures. Noattempt to exclude oxygen or otherwise control the atmosphere inthe tubes other than hermetical capping was made. After 24 hours,the tubes were centrifuged (12 000 × g) for 15 minutes, and thesupernatant filtered through 0.22-μm syringe filters (DuraporePVDF membrane, Millipore Corp., Carrigtwohill, Ireland)). Thefiltered solutions were analysed by inductively coupled plasma-atomic emission spectroscopy (ICP-AES; ARCOS SOP, SpectroAnalytical Instruments, Inc., Mahwah, New Jersey, USA: sandysoil) or by atomic absorption spectroscopy (AAS; Analyst 800atomic absorption spectrophotometer, Perkin Elmer, Waltham,Massachusetts, USA: all other soils). Manganese and Fe releasewas calculated as the difference between the total concentrationmeasured in the soil-containing samples (extract or water + buffer+ soil) and the soilless controls (extract or water + buffer),and is denoted in the figures with a � notation. Since Zn(III)is hardly amenable to reduction, yet is more readily complexedby most organic ligands than Mn (Furia, 1980), Zn concentrationswere also determined to help differentiate between reduction andcomplexation mechanisms potentially involved in solubilization.Preliminary experimentation showed that release of the metalsfrom soil was not affected by total salt concentrations in the rangesfound in the aqueous biochar extracts. Moreover, DOC release andsorption by the soils were negligible.

Buffers (100 mm) were made up using the following solutions:(A) glacial acetic acid (8.6 ml made up to 1.5 l in water; (B)sodium acetate trihydrate (27.2 g) made up to 2 l in water; (C)disodium hydrogen phosphate (28.4 g) made up to 2 l in water;and (D) 0.1 m HCl. Buffer pH 4 was made by mixing 153 ml (A)



and 847 ml (B); pH 5 by mixing 643 ml (A) and 357 ml (B); pH 6by mixing 947.8 ml (A) and 52.2 ml (B); pH 7 by mixing 756 ml(C) and 244 ml (D); pH 8 by mixing 955.5 ml (C) and 44.9 ml(D). Despite the use of buffer solutions, it was difficult to achievetarget pH values because of the strong buffering capacities of thesoils and extracts. Therefore, results were compared within buffersystems by using the actual measured pH values at the time ofsampling.

Statistics

GraphPad Software, Inc. (San Diego, California) was used forcarrying out unpaired two-tailed Student’s t-tests. Results reportedas significant were at P < 0.01.

Results

Chemical characteristics and reducing capacity of biocharextracts and DOMs

Extracts of OP and GHW biochars were alkaline and relativelysaline (Table 1) regardless of pyrolysis HTT. Extracts of theGHW biochars in particular had elevated salt contents, becausethe feedstock comprised residues of pepper plants grown on salinewater (3–4.5 dS m−1). The pH of EUC biochar extracts increasedfrom neutral to alkaline as a function of increasing pyrolysis HTT.For a given feedstock, extracts of smaller-HTT biochars (350 and450◦C) had notably larger DOC concentrations than extracts oflarger-HTT biochars (Table 1). At the smaller-HTTs, feedstockhad a particularly strong impact on DOC content, with biocharsfrom the agricultural residues (GHW and OP) having extractswith significantly higher DOC concentrations than those producedfrom wood; the differences between feedstocks decreased forbiochars produced at 600◦C. All the biochar extracts had redoxpotentials (as measured by ORP probe) significantly less than thatof water (Table 1). The concentrations of metals in the GHW-450 and GHW-600 extracts that were used for the solubilizationexperiments were very small (Table 2).

The biochar extracts exhibited large differences in concentrationof total phenols (from not detectable to 204 μmol GAE l−1),with extracts of smaller-HTT biochars having much greaterconcentrations than extracts of larger-HTT biochars (Table 1).While the major part of the difference was mainly related toDOC content of the extracts (linear regression of total phenols(TPs) against DOC: TP = 0.4416 DOC − 10.77; R2 = 0.943),normalization to unit weight of DOC shows that extracts ofsmaller-HTT biochars had greater content of total phenols perunit of DOC (Table 1). On a unit DOC weight basis, total phenolcontents of smaller-HTT biochar extracts were of the same orderas those of natural DOM samples (Table 1).

Reducing capacity was calculated from both the FC assay (usingthe Fe2SO4 standard curve) and the FRAP assay (Figure 1). For allthe biochar extracts except that of GHW-600, the FRAP4 reducingcapacity was significantly less than that determined by the FCassay (Figure 1). Over the 4–240 minute FRAP assay time, the

Table 2 Elemental content of aqueous extracts of biochars GHW-450 andGHW-600 used in solubilization experiment determined by inductivelycoupled plasma-atomic emission spectroscopy (ICP-AES). ICP-AESanalysis was performed in triplicate; differences between replicates wereless than 5%

Element GHW-450 / μmol ml−1 GHW-600 / μmol ml−1

Al 0.0093 0.0006B 0.0545 0.0456Ba 0.0002 0.0000Ca 0.1971 0.0863Cr 0.0001 0.0000Cu 0.0006 0.0000Fe 0.0039 0.0004Li 0.0304 0.0177Mg 1.6910 0.0235Mn 0.0001 0.0000Mo 0.0001 0.0002Na 5.2634 6.4814Ni 0.0001 0.0000P 0.0197 0.0251S 4.1166 4.8339Si 0.1560 0.1720Sr 0.0043 0.0009Ti 0.0014 0.0000Zn 0.0008 0.0000

reducing capacity increased for all the biochar extracts (Figure 1).For the most part, the reducing capacity at FRAP240 was less than,or nearly the same as, that of the FC assay (Figure 1). However,several extracts, notably those of GHW-600, EUC-350 and EUC-450, had significantly larger FRAP240 reducing capacities thanFC reducing capacities (Figure 1).

The increase in FRAP reducing capacity from 4 (FRAP4)to 240 minutes (FRAP240) was generally larger for the biocharextracts than the natural DOM samples (Figure 1). BetweenFRAP4 and FRAP240, the reducing capacity of the DOM samplesincreased by 1.2 (PPHA) to 3.4 (SRNOM) times, while that ofthe biochar extracts increased by 3.0 (OP-600) to 6.1 (GHW-600) times. In contrast to the biochar extracts, FRAP240 reducingcapacities of the DOM samples were always much smaller thantheir respective FC reducing capacities.

A number of compound classes were identified putatively byGC/MS in the aqueous extracts of GHW-450 and GHW-600biochars: polyols, hydroxy acids, benzoic acids, substitutedheterocyclic amines, urea, medium and long-chain alkanoiccarboxylic acids, dicarboxylic acids, sugars, sugar alcohols, sugaracids, anhydrosugars and glycerol-substituted long-chain acids(Table 3). Some of the compounds have structures associatedwith redox activity, in particular the aromatic compounds andheterocyclic amines. Other identified compounds, mainly thedicarboxylic acids, have considerable propensity to form stablecomplexes with metals. Twice as many compounds overallwere identified in the extract of GHW-450 as in the extractof GHW-600 (35 compared with 17). It should be noted that



EUC-350

EUC-450

EUC-600

EUC-800

OP-350

OP-450

OP-600

GHW-3

50

GHW-4

50

GHW-6

00

SRNOMSRFA

PPHAPPFA

0

1

2

3

4

5

6

7

8

Red

ucin

g C

apac

ity /

μmol

c mg

DO

C-1

FC FRAP4 FRAP30 FRAP60 FRAP120 FRAP180 FRAP240

Figure 1 Reducing capacity in μmolc per unit dissolved organic carbon (mg DOC) of the biochar extracts and natural dissolved organic matter (DOM)samples measured by the Folin-Ciocalteu (FC) assay and the FRAP (Ferric Reducing Antioxidant Power) assay. The number following the FRAP designationrepresents the time in minutes from the start of the reaction; thus, FRAP4 is the reading taken 4 minutes after the start of the reaction. Abbreviations forbiochars and DOM samples are as given in the text. Error bars represent the maximum propagated standard error (7% of the mean value).

the major portion of organic carbon in the extracts was notidentifiable by GC/MS, thus the compounds in Table 3 representonly a small fraction of the total carbon present.

Mn and Fe solubilization from soils

Extracts of GHW-450 and GHW-600 solubilized more Mn andFe from the sandy soil than did the control solutions at neutral toacidic pH values, with the amount of metal released into solutionincreasing with decreasing pH (Figure 2). The relative solubi-lization power of the extracts compared with the control solutionalso increased as pH decreased (Figure 2). The same trend wasobserved for Mn solubilization in the other three soils (Figure 3).

At a given pH, the extract of the small-HTT biochar (GHW-450) solubilized significantly more Mn and Fe than did theextract of the large-HTT biochar (GHW-600; Figures 2, 3); forexample, at pH 5.1, releasing 3.7 times more Mn and 12.7 timesmore Fe from the sandy soil (respectively, Mn, 146 ± 8.4 and39.8 ± 0.15 μmol l−1; and Fe, 5.0 ± 0.57 and 0.4 ± 0.01 μmol l−1;Figure 1). The extent of Mn solubilization differed from soil tosoil (Figures 2, 3), reflecting the particular chemistry of each soilin contact with the extracts, but in all cases, the GHW-450 extractreleased significantly more Mn than the GHW-600 extract. Ingeneral in the different soils, enhanced solubilization by thebiochar extracts disappeared at pH greater than 7.5–8.

No difference in Zn release in biochar extracts or controlsolutions was observed in any of the soil-pH systems (not shown).

Discussion

As phenols have fast redox reaction kinetics in the FC assay andslow reaction kinetics in the FRAP assay, the significantly largerreducing capacity measured by the FC assay than that measuredat 4 minutes in the FRAP assay (FRAP4), suggests that phenoliccompounds were responsible for the greater part of the reducingcapacities of the extracts. Phenols are among the compoundsresponsible for the redox activity of natural DOMs (Rimmer &Abbott, 2011; Aeschbacher et al., 2012). In general, the measuredreducing capacities of the DOM and biochar samples were inthe range of the electron donating capacities determined for 15different humic substances over a large range of pH and E h valuesmeasured using mediated electrochemical oxidation (Aeschbacheret al., 2012). For some extracts, the long-time FRAP reducingcapacity (FRAP240) exceeded that of the FC assay, perhapsindicating that those extracts contained additional species whoseoxidation half-reaction potentials were sufficient for reducingFe(III) in the FRAP assay, but not Mo(VI) in the FC assay.

The natural DOM samples had notably slower FRAP kineticsthan the biochar extracts. DOM is known to have sluggishredox kinetics (Aeschbacher et al., 2010) and possibly slowdissociation kinetics because of the time needed for diffusion intothe macromolecular framework. Biochar extracts contain bothlow and high molecular weight organic compounds (Graber et al.,2010; Lin et al., 2011), such that their kinetics and chemicalprocesses may differ from those that characterize natural DOM.



Table 3 Putative identifications by gas chromatograph/mass spectrometer (GC/MS) of compounds in aqueous extracts of greenhouse waste (GHW) biocharsproduced at 450◦C (GHW-450) and 600◦C (GHW-600)

Compound Class GHW-450 GHW-600

Lactic acid Hydroxy acid Ya YHexanoic acid n-Alkanoic acid YHydroxy-acetic acid Hydroxy acid Y YEthandioic acid Dicarboxylic acid – Y4-Hydroxy-butyric acid Hydroxy acid Y –Benzoic acid Benzoic acid Y YUrea Urea Y –Octanoic acid n-Alkanoic acid Y –Glycerol Polyol Y YSuccinic acid Dicarboxylic acid Y Y2-Methyl benzoic acid Substituted benzoic acid Y –2-Methyl butanedioic acid Dicarboxylic acid Y YGlyceric acid Hydroxy acid Y YNonanoic acid n-Alkanoic acid Y –Glutaric acid Dicarboxylic acid Y –2,4-Bis[hydroxy] butanoic acid Hydroxyl acid Y –Benzenepropanoic acid Aromatic organic acid Y –Decanoic acid n-Alkanoic acid Y –Butane-1,2,3,4-tetraol (erythritol) Polyol Y –Pentane-1,2,5-triol Polyol Y –5-Oxo-pyrrolidine-2-carboxylic acid (pyroglutamic acid) Substituted heterocyclic amine Y YPiperidine-2-carboxylic acid Substituted heterocyclic amine – Y2-Hydroxy-pentandioic acid Dicarboxylic acid Y –Erythronic acid Sugar acid – YThreonic acid Sugar acid – Y3-Hydroxy-benzoic acid Substituted benzoic acid Y –1H-benzoimidazole,1-(2-ethoxyethyl)-2-(4-methoxyphenyl) Substituted imidazole Y –2,4,5-Trihydroxypentanoic acid Hydroxy acid Y –Ethane-1,2-diol (ethylene glycol) Diol Y –1,6-Anhydroglucose Anhydrosugar Y –Ribitol Reducing sugar Y –Mannitol Sugar alcohol Y YHexadecanoic acid n-Alkanoic acid Y YMyo-inositol Polyol Y YOctadecanoic acid n-Alkanoic acid Y Y1-Monohexadecanoylglycerol Glycerol ester Y –2,3-Bis[(hydroxyl)propyl]-hexadecanoic acid Carboxylic acid Y –Melezitose Trisaccharide – YTrehalose Disaccharide Y –Monooctadecanoylglycerol Ketoglycerol Y –

aY-identified in the extract.

The GC/MS analysis revealed redox-active small molecules in the

biochar extracts, but large molecules are not amenable to GC/MS

analysis, nor are small polar compounds that lack groups having

active H atoms. Many large polyphenols, tannins and other

macromolecular species such as humic-like substances produced

during the pyrolysis process (Lin et al., 2011) cannot be detected

by GC/MS, yet are expected to play a role in the redox activity

of the extracts. They may also form complexes with metals.

Most of the metals naturally present in the extracts are

assumed to be complexed with organic species or present as

mineral nanoparticles. In general, the redox reaction rate of metal

species is fast, particularly at the low pH conditions of the FRAP

assay. Therefore, it is expected that the original metal content of

the extracts does not play the major role in their redox activity, in

accordance with previous findings for metals in DOM solutions

(Aeschbacher et al., 2011).

While formation of complexes between soil metals and organic

ligands in the extracts potentially may be a mechanism for the

observed enhanced solubilization of Mn and Fe by the biochar

extracts, this possibility is not supported by the results of the Zn



0

30

60

90

120

150

180

(b)

(c)

(d)

(a) GHW-450water

Mn

Sand

0

1

2

3

4

5

6

7GHW-450water

Fe

4 5 6 7 8 9

0

10

20

30

40

50GHW-600water

Solution pH

Mn

4 5 6 7 8 9

0.0

0.1

0.2

0.3

0.4

0.5GHW-600water

Fe

Δ E

lem

ent C

once

ntra

tion

/ μm

ol l-

1

Figure 2 Release of Mn (a, b) and Fe (c, d) froma sandy soil into the extract of greenhouse waste(GHW) biochar produced at a highest treatmenttemperature (HTT) of 450◦C (GHW-450) and 600◦C(GHW-600), and water as a function of pH. Errorbars represent standard errors, and when not presentare smaller than the symbols.

analyses. Zinc(III) is poorly amenable to reduction, yet is morereadily complexed by most organic ligands than Mn (Furia, 1980).As there was no difference in Zn released to the extracts or con-trols, it is suggestive that the cause of Mn (and Fe) enhanced sol-ubilization was indeed reduction and not complexation by organicligands. Manganese release was greater than that of Fe, presum-ably because the reduction potentials of most soil Mn oxides aregreater than those of most soil Fe species (Bartlett & James, 1993).

Biochar feedstocks differ in chemical composition, and cor-respondingly, chemical properties of biochar differ as a func-tion of feedstock. The disparities in reducing capacity amongthe biochars from various feedstocks in this study are thoughtto reflect these chemical differences. Extracts of smaller-HTTbiochars had greater reduction capacities than those of larger-HTT biochars, implying that biochars produced at smaller HTTshave greater potential for participating in redox reactions in soil.Notably, when the soil mineral solubilization results were normal-ized to the DOC content of the extracts, the larger-HTT biocharextract had significantly greater solubilization power than thesmaller-HTT biochar extract (shown for Mn for all four soilsin Figure 4). This result corresponds to the significantly greaterDOC-normalized FRAP240 reducing capacity of the extract fromthe larger-HTT biochar (GHW-600) compared with the extractof the smaller-HTT biochar (GHW-450), and demonstrates howHTT can affect the nature of the redox active compounds inbiochar.

The implications of these results for the role of biochar in thesoil may be far-reaching. Because oxidation of biochar solidsleads to continued release of redox-active, acidic and phenolicorganics of both low and high molecular weight (Abiven et al.,2011), biochar may continue to participate in redox reactions as

it ages in the soil. This could have consequences for soil releaseof nutrient or contaminant metals, and their bioavailability ormobility in the environment.

As well as the importance of oxidation state for the valueof Fe and Mn as micronutrients, their oxidation state can affectthe oxidation states of other metallic species such as chromium(Tokunaga et al., 2007). Reduced Fe and Mn species have beensuggested as catalysts for abiotic reduction of nitrate to nitrite,with nitrite reacting with dissolved organic compounds vianitration and nitrosation of aromatic ring structures to produceorganic N structures (Davidson et al., 2003). Such abiotic redoxreactions could be one of the ways by which biochar influencessoil N cycles and could contribute to the build-up of SOM. In aseries of papers, Rimmer and colleagues (Rimmer, 2006; Rimmer& Smith, 2009; Rimmer & Abbott, 2011) suggested that soilpolyphenols scavenge free radicals, thus terminating the oxidativechain reactions responsible for SOM breakdown. Polyphenolsreleased from biochars could perform similar functions, whichmay be one means by which biochar helps to stabilize soil organicmatter.

Redox reactions between biochar organic compounds and soilmetals could contribute in other ways to SOM stabilization.Phenolic compounds with hydroxyl groups in the ortho- andpara-position are known to reduce Mn and Fe under normal soilconditions and precipitate as polymeric humic-like substances(Pohlman & McColl, 1986, 1989). This mechanism accords withobservations that biochar added to soil resulted in enhancedbreakdown of fresh organic material followed by its incorporationinto soil aggregates and organo-mineral fractions, leading tostabilization of originally labile organic matter (Liang et al.,2010).



5 6 7 80

30

60

90

120

150

180(b)

(c) (d)

(e) (f)

(a)

BC extractwater

GHW-450 GHW-600

clayey loam

light clay

5 6 7 80

20

40

60

80

100

light clay

BC extractwater

5 6 7 80

30

60

90

120

5 6 7 8

0

10

20

30

40

clayey loam

5 6 7 80

40

80

120

160

200

heavy clay

5 6 7 8

0

20

40

60

80

heavy clay

Solution pH

Δ M

n C

once

ntra

tion

/ μm

ol l-

1

Figure 3 Release of Mn into the extract of greenhouse waste (GHW) biochar produced at a highest treatment temperature (HTT) of 450◦C (GHW-450)and 600◦C (GHW-600), and water compared as a function of pH for three soils (light clay (a, b) GHW-450 and GHW-600, respectively: clayey loam(c, d) GHW-450 and GHW-600, respectively: heavy clay (e, f) GHW-450 and GHW-600, respectively). Error bars represent standard errors, and whennot present are smaller than the symbols.

The effect of biochar on soil microbial processes and communi-ties (Graber et al., 2010; Kolton et al., 2011) could also be relatedto redox activity of biochar, with components in the water-solublefraction enabling electron transfers between bacterial cells andFe(III)-bearing minerals. Redox active substances could take partin bacterial reduction and immobilization of contaminants such asCr(VI) (Choppala et al., 2012) or, alternatively, cause the abioticrelease of contaminants such as arsenic that are associated withFe oxides.

The specific effect of biochar addition on release and solu-bilization of soil metals will depend not only on the reductioncapacity and chelating potential of the water-soluble componentsof biochars, but also on prevailing pH conditions. While solu-bilization of soil metals by biochar extracts is enhanced whencompared with water at a given pH, biochars and their extractshaving basic pH values may raise soil pH in acidic sandy soils(Yuan & Xu, 2011) and hence result in overall decreased metalrelease to soil solution. In this way, biochar could result indecreased metal toxicity and contamination risks in acidic sandysoils despite its content of reducing agents and organic ligands.

Conclusions

Water extracts of many biochars have substantially smaller redoxpotentials than water and can solubilize soil Mn and Fe. Thedissolved organic matter fraction appears to be mainly responsiblefor the redox activity of the extracts examined in this study,although metals released from biochars prepared from otherfeedstocks or under different conditions may also be important.While some compounds making up the water-soluble fractionare also able to form complexes with released metals, in thisstudy, reduction was apparently the main process responsible forMn and Fe solubilization from soils. The implications of thisare that many biochars have the possibility of participating ina variety of chemical and biological redox-mediated reactionsin the soil. In this way, biochar could influence importantprocesses occurring in the soil, such as microbial electronshuttling, nutrient cycling, free radical scavenging, pollutantdegradation, contaminant mobilization, and abiotic formation ofhumic structures in soils. Participation in varied redox reactionsmay be among the many mechanisms involved in the impact ofbiochar on the soil environment.



5 6 7 8

0

4

8

12

16(b)(a)

GHW-450

GHW-600

sand

5 6 7 80

5

10

15

20

25

30

35light clay

5 6 7 8

0

4

8

12(c)

Solution pH

heavy clay

5 6 7 8

0

5

10

15

20

25

30(d)clayey loam

Δ M

n co

nc. /

μm

ol M

n m

mol

DO

C-1

Figure 4 Release of Mn per unit dissolved organic carbon (DOC) from four soils, (a) sand, (b) light clay, (c) clayey loam and (d) heavy clay, comparedbetween extracts of greenhouse waste (GHW) biochar produced at a highest treatment temperature (HTT) of 450◦C (GHW-450) and 600◦C (GHW-600).Error bars represent standard errors, and when not present are smaller than the symbols.

Supporting Information

The following supporting information is available in the onlineversion of this article:Table S1. Physical and chemical characteristics of the biocharsused in this study.Table S2. Soil characteristics.

Acknowledgements

The authors would like to acknowledge Dr Stephen Josephfor many interesting conversations on the redox behaviour ofbiochars and other subjects, and Professor Uri Mingelgrin forhelpful discussions on the nature of redox reactions. A criticalreview of an earlier version of the manuscript by Dr MichaelSanders was very insightful and important, and his contributionis gratefully acknowledged. Ms Racheli Rozenfeld assisted invarious aspects of the laboratory experiments. Dr Guy Levy andMs Dina Goldstein are thanked for providing three of the soils andanalyses of their physical-chemical characteristics. This researchwas supported by public grants from the Chief Scientist of theIsrael Ministry of Agriculture and Rural Development, projects301-0693-10 and 261-0848-11, and the Israel-Italy Program 2011(project 301-750-11).

References

Abiven, S., Hengartner, P., Schneider, M., Singh, N. & Schmidt, M.

2011. Pyrogenic carbon soluble fraction is larger and more aromatic

in aged charcoal than in fresh charcoal. Soil Biology & Biochemistry ,

43, 1615–1617.

Aeschbacher, M., Sander, M. & Schwarzenbach, R.P. 2010. Novel

electrochemical approach to assess the redox properties of humic

substances. Environmental Science & Technology , 44, 87–93.

Aeschbacher, M., Vergari, D., Schwarzenbach, R.P. & Sander, M. 2011.

Electrochemical analysis of proton and electron transfer equilibria

of the reducible moieties in humic acids. Environmental Science &

Technology , 45, 8385–8394.

Aeschbacher, M., Graf, C., Schwarzenbach, R.P. & Sander, M. 2012.

Antioxidant properties of humic substances. Environmental Science &

Technology , 46, 4916–4925.

Bartlett, R.J. & James, B.R. 1993. Redox chemistry of soils. Advances in

Agronomy , 50, 151–208.

Benzie, I.F.F. & Strain, J.J. 1996. The ferric reducing ability of plasma

(FRAP) as a measure of “antioxidant power”: the FRAP assay.

Analytical Biochemistry , 239, 70–76.

Bouyoucos, G.H. 1962. Hydrometer method improved for making particle

size analysis of soils. Agronomy Journal , 54, 464–465.

Choppala, G.K., Bolan, N.S., Megharaj, M., Chen, Z. & Naidu, R.

2012. The influence of biochar and black carbon on reduction and



bioavailability of chromate in soils. Journal of Environmental Quality ,41, 1175–1184.

Davidson, E.A., Chorover, J. & Dail, D.B. 2003. A mechanism ofabiotic immobilization of nitrate in forest ecosystems: the ferros wheelhypothesis. Global Change Biology , 9, 228–236.

Elad, Y., Rav David, D., Meller Harel, Y., Borenshtein, M., Ben Kalifa,H., Silber, A., et al. 2010. Induction of systemic resistance in plants bybiochar, a soil-applied carbon sequestering agent. Phytopathology , 100,913–921.

Enders, A. & Lehmann, J. 2012. Comparison of wet digestion and dryashing methods for total elemental analysis of biochar. Communicationsin Soil Science & Plant Analysis , 43, 1042–1052.

Fimmen, R.L., Cory, R.M., Chin, Y.P., Trouts, T.D. & McKnight,D.M. 2007. Probing the oxidation-reduction properties of terrestriallyand microbially derived dissolved organic matter. Geochimica et

Cosmochimica Acta , 71, 3003–3015.Furia, T.E. 1980. Sequestrants in Foods . CRC Press, Cleveland, OH.Graber, E.R., Meller-Harel, Y., Kolton, M., Cytryn, E., Silber, A., Rav

David, D., et al. 2010. Biochar impact on development and productivityof pepper and tomato grown in fertigated soilless media. Plant & Soil ,337, 481–496.

Kolton, M., Meller Harel, Y., Pasternak, Z., Graber, E.R., Elad,Y. & Cytryn, E. 2011. Impact of biochar application to soil onthe root-associated bacterial community structure of fully developedgreenhouse pepper plants. Applied & Environmental Microbiology , 77,4924–4930.

Laird, D.A., Fleming, P.D., Davis, D.D., Horton, R., Wang, B. & Karlen,D.L. 2010. Impact of biochar amendments on the quality of a typicalMidwestern agricultural soil. Geoderma , 158, 443–449.

Lehmann, J., Pereira da Silva, J. Jr., Steiner, C., Nehls, T., Zec, W. &Glaser, B. 2003. Nutrient availability and leaching in an archaeologicalAnthrosol and a Ferralsol of the Central Amazon basin: fertilizer,manure and charcoal amendments. Plant & Soil , 249, 343–357.

Liang, B.Q., Lehmann, J., Sohi, S.P., Thies, J.E., O’Neill, B., Trujillo, L.,et al. 2010. Black carbon affects the cycling of non-black carbon in soil.Organic Geochemistry , 41, 206–213.

Lin, Y., Munroe, P., Joseph, S., Henderson, R. & Ziolkowski, A. 2011.Water extractable organic carbon in untreated and chemical treatedbiochars. Chemosphere, 87, 151–157.

Lovley, D.R., Fraga, J.L., Blunt-Harris, E.L., Hayes, L.A., Phillips, E.J.P.& Coates, J.D. 1998. Humic substances as a mediator for microbiallycatalyzed metal reduction. Acta Hydrochimica et Hydrobiologica , 26,152–157.

Meller Harel, Y., Elad, Y., Rav-David, D., Borenshtein, M., Schulcani, R.,Lew, B., et al. 2012. Biochar mediates systemic response of strawberryto foliar fungal pathogens. Plant & Soil , 357, 245–257.

Nelson, R.E. 1982. Carbonate and gypsum. In: Methods of Soil Anal-

ysis. Part 2. Chemical and Microbiological Properties , 2nd edn

(eds A.L. Page, R.H. Miller & D.R. Keerney), pp. 181–196. ASA andSSSA, Madison, WI.

Oh, S.-Y., Son, J.-G. & Chiu, P.C. 2013. Biochar-mediated reductivetransformation of nitro herbicides and explosives. EnvironmentalToxicology & Chemistry , 32, 501–508.

Pohlman, A.A. & McColl, J.G. 1986. Kinetics of metal dissolution fromforest soils by soluble organic acids. Journal of Environmental Quality ,15, 86–92.

Pohlman, A.A. & McColl, J.G. 1989. Organic oxidation and manganeseand aluminum mobilization in forest soils. Soil Science Society ofAmerica Journal , 53, 686–690.

Rhoades, J.D. 1982. Cation exchange capacity. In: Methods of SoilAnalysis. Part 2. Chemical and Microbiological Properties, 2nd edn(eds A.L. Page, R.H. Miller & D.R. Keerney), pp. 149–157. ASA andSSSA, Madison, WI.

Rimmer, D.L. 2006. Free radicals, antioxidants, and soil organic matterrecalcitrance. European Journal of Soil Science, 57, 91–94.

Rimmer, D.L. & Abbott, G.D. 2011. Phenolic compounds in NaOHextracts of UK soils and their contribution to antioxidant capacity.European Journal of Soil Science, 62, 285–294.

Rimmer, D.L. & Smith, A.M. 2009. Antioxidants in soil organic matterand in associated plant materials. European Journal of Soil Science, 60,170–175.

Silber, A., Levkovitch, I. & Graber, E.R. 2010. pH-Dependent mineralrelease and surface properties of cornstraw biochar: agronomic impli-cations. Environmental Science & Technology , 44, 9318–9323.

Singleton, V.L., Orthofer, R. & Lamuela-Raventos, R.M. 1999. Analysis oftotal phenols and other oxidation substrates and anti-oxidants by meansof Folin-Ciocalteau reagent. Methods in Enzymology , 299, 152–178.

Soil Survey Staff 1999. Soil Taxonomy: A Basic System of Soil Classifi-cation for Making and Interpreting Soil Surveys . Agriculture Handbook

No. 436 , 2nd edn. U.S. Department of Agriculture, Natural ResourcesConservation Service, Washington.

Tokunaga, T.K., Wan, J., Lanzirotti, A., Sutton, S.R., Newville, M. & Rao,W. 2007. Long-term stability of organic carbon-stimulated chromatereduction in contaminated soils and its relation to manganese redoxstatus. Environmental Science & Technology , 41, 4326–4331.

Visser, S.A. 1985. Physiological action of humic substances on microbial-cells. Soil Biology & Biochemistry , 17, 457–462.

Walkley, A. 1947. A critical examination of a rapid method fordetermination of organic carbon in soils – effect of variations indigestion conditions and of inorganic soil constituents. Soil Science,63, 251–257.

Yuan, J.H. & Xu, R.K. 2011. The amelioration effects of low temperaturebiochar generated from nine crop residues on an acidic Ultisol. Soil Use& Management , 27, 110–115.


reducing capacity of water extracts of biochars and their solubilization of soil mn and fe

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