soil mineral organic matter microbe interactions: impacts

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Pedobiologia 49 (2005) 609—635

KEYWORDBiogeochereactionsprocesses;BiodiversitInterfacialinteractionMicroorganMinerals;Organic mSoil;Terrestrialecosystem

0031-4056/$ - sdoi:10.1016/j.

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www.elsevier.de/pedobi

INTERNATIONAL SYMPOSIUM ON IMPACTS OF SOIL BIODIVERSITY ON BIOGEOCHEMICAL PROCESSES INECOSYSTEMS, TAIPEI, TAIWAN, 2004

Soil mineral–organic matter–microbe interactions:Impacts on biogeochemical processes andbiodiversity in soils

Pan-Ming Huanga,�, Ming-Kuang Wangb, Chih-Yu Chiuc

aDepartment of Soil Science, University of Saskatchewan, 51 Campus Drive, Saskatoon, SK, Canada S7N 5A8bDepartment of Agricultural Chemistry, National Taiwan University, No. 1, Section 4, Roosevelt Road,Taipei 10764, TaiwancResearch Center for Biodiversity, Academia Sinica, Nankang, Taipei 11529, Taiwan

Received 6 October 2004; accepted 29 June 2005

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SummarySoils are the central organizer of the terrestrial ecosystem. Their colloidal andparticulate constituents, be they minerals, organic matter, and microorganisms, arenot separate entities; rather, they are constantly interacting with each other.Interactions of these components control biogeochemical reactions, namely, theformation of short-range-ordered metal oxides, catalysis of humic substanceformation, enzymatic stability and activity, mineral transformation, aggregateturnover, biogeochemical cycling of C, N, P, and S, and the fate and transformationof organic and inorganic pollutants. Furthermore, the impacts of mineral–organicmatter–microorganism interactions and associated biogeochemical reactions andprocesses on biodiversity, species composition, and sustainability of the terrestrialecosystem deserve close attention for years to come. This paper integrates thefrontiers of knowledge on this subject matter, which is essential to uncovering thedynamics and mechanisms of terrestrial ecosystem processes and to developinginnovative management strategies to sustain ecosystem health on the global scale.& 2005 Elsevier GmbH. All rights reserved.

Introduction

Soil is a life-sustaining material, which is astructurally porous and biologically active medium

5 Elsevier GmbH. All rights rese

sk.ca (P.-M. Huang).

that has evolved over time on the continental landsurfaces on our planet Earth (Fig. 1). This materialis formed and continues to develop through weath-ering processes driven by biogeochemical, climatic,geological, topological, and chronological influ-ences. Soil is geoderma, namely, the skin of theplanet Earth. It is the pedosphere, which overlaps

rved.

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Figure 1. Soil as the life-sustaining material which is the structurally porous and biologically active medium (Spositoand Reginato, 1992).

Figure 2. Diagram illustrating the interaction of bacteriaand fungi with mineral particles in a soil aggregate(Theng and Orchard, 1995). Bacterial cells with a coat ofextracellular polysaccharides (EPS) are enveloped by clayparticles. The pore space where clays and bacteriainteract, bounded by silt- and sand-size particles, canbe relatively enriched in organic matter including EPSresidues. Fungal hyphae are attached to the outsidesurface of an aggregate. Inset shows an enlarged view ofa bacteria cell with its complement of EPS. At circumneutral soil pH conditions, the cell has a net negativesurface charge. Most clay particles adhere to the cellsurface by bridging through polyvalent cations, repre-sented by Mn+ (a), although some may be attacheddirectly by electrostatic interactions, either in a face-to-face (b), or edge-to-face (c) association.

P.-M. Huang et al.610

with the lithosphere, hydrosphere, atmosphere,and biosphere and is, thus, an integral part of theecosystem. Therefore, soil physical, chemical andbiological process should have a profound impacton ecosystem sustainability.

Soil is the central organizer of the terrestrialecosystem (Odum, 1989; Coleman et al., 1998).Minerals, organic components and microorganismsare among major solid components of soils. Thesecomponents are not separate entities but rather aunified system constantly in association with eachother in the environment (Huang et al., 1995). Theassociation of microorganisms with soil mineral andorganic colloids is depicted in Fig. 2. Interactionsamong these components have enormous impactson physics, chemistry, and biology of soil andsurrounding ecosystems (Huang and Schnitzer,1986; Huang et al., 1995). Fundamental under-standing of mineral–organic matter–microorganisminteractions is essential for understanding, restor-ing, enhancing, and sustaining ecosystem integrityon a global scale. The study of the interactions ofthese soil components has to be considered fromthe molecular level to field/landscape systems andis essential to stimulate further research to uncoverthe dynamics and mechanisms of soil processes andthe impacts on biogeochemistry and biodiversity inthe terrestrial ecosystem. In view of the continualmineral–organic component–microorganism inter-actions and interactive physical, chemical, andbiological processes in the terrestrial ecosystem, anew Commission ‘‘Soil Physical/Chemical/Biologi-cal Interfacial Interactions’’ was created within thestructure of the International Union of Soil Sciencesin 2004.

This paper integrates the existing knowledge,especially the latest advances on the fundamentalsof soil mineral–organic matter–microorganism in-teractions, and addresses a variety of their poten-tial impacts on biogeochemical processes and

biodiversity. It begins with the topics on theformation of short-range-ordered (SRO) metaloxides, mineral catalysis of humic substanceformation, effects of soil colloids on enzymaticstability and activity, and microbial mediation ofmineral transformation, because mineral and or-ganic colloids, enzymes, and microbes exert great

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Soil mineral–organic matter–microbe interactions 611

control over soil physical, chemical, and biologicalprocesses. We then address the topics on dynamicsof aggregate turnover, biogeochemical cycling of C,N, P, and S, the fate and transformation of organiccontaminants and metals, and potential impact onbiodiversity in terrestrial ecosystems. Further-more, future prospects of major research thrustson the subject matter are also discussed topromote research and education in this importantand exciting area of science.

Formation of short-range-ordered metaloxides

Metal oxides are ubiquitous in soils and play asignificant role in influencing soil behavior, and,thus, have great impacts on the ecosystem. SRO Aland Fe oxides, especially their nanoparticles areundoubtedly among the most reactive components

Figure 3. A proposed reaction scheme for Al hydroxide mineet al., 2002). Aluminum transformation in aqueous systems maI is common when systems are neutralized rapidly within seformation of Al13 (AlO4Al12(OH)24(H2O)12

7+). Reaction Path II iscompared with Reaction Path I. Reaction Path III is applicableformation of Al13 aggregates through an anion bridging mechordered and highly reactive intermediate phases and stable

of acidic and neutral soils (Bigham et al., 2002;Huang et al., 2002). Biomolecules exert a verysignificant influence on the formation and transfor-mation of metal oxides and the resultant alterationof their surface properties pertaining to thespeciation, fate and bioavailability of nutrients,toxic inorganic and organic substances, and themicrobial events (Huang and Violante, 1986;Schwertmann et al., 1986; Sposito, 1996; Huangand Wang, 1997; Huang et al., 2002).

As the third most abundant element on theEarth’s surface (after O and Si), Al is a majorelement in all-mineral soils. There are only a fewcrystalline Al oxide, hydroxide, or oxyhydroxideminerals, and only one, i.e., gibbsite occurs to anygreat extent in soils. Aluminum, however, alsoforms a series of highly reactive soluble species andpoorly crystalline to noncrystalline minerals andcolloids (Fig. 3).

The model depicted in Fig. 3 incorporates multiplereaction paths. Reaction path I is believed to be

ral formation from hydrolyzed aluminum solution (Huangy proceed through multiple reaction paths. Reaction pathconds or hours. Reaction Paths II and III both involve theapplicable when systems are neutralized relatively slowlywhen Al13 forms under conditions that promote the rapidanism. These reactions result in the formation of poorlycrystalline Al hydroxides.

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Figure 4. The X-ray diffraction patterns of hydrolyticprecipitation products of Al, showing how four differentorganic acids influence the transformation to morecrystalline phases. The initial concentration was1.1� 10�3 M at an OH/Al molar ratio of 3 and the solutionwas aged for 40 d at room temperature in the presence of10�4 M organic acids (Kwong and Huang, 1979b).


common when systems are neutralized rapidly withinseconds or hours. Under these conditions, little orno Al13 polynuclear species ðAlO4Al12ðOHÞ24ðH2OÞ

7þ12 Þ

forms; the formation of Al(OH)3 nuclei is rapid, andgibbsite appears within days or weeks of aging.Reaction Paths II and III both involve the formation ofAl13. Reaction path II is applicable when systems areneutralized relatively slowly compared with Reactionpath I. The individual Al13 ions can remain in solutionfor months to years, but eventually they transfer byone of the three pathways. If no Al13 nucleiare present, the soluble Al13 ions can slowlydissociate into Al3+ ions and deposit onto Al(OH)3nuclei (Path IIa). These soluble Al13 may alsoaggregate and form Al13 nuclei (Path IIb). If someAl13 ions are already present, the remaining Al13 ionscan be deposited on them within weeks and months(Path IIc). Path III is applicable when Al13 forms underconditions that promote the rapid formation of Al13aggregates via an anion bridging mechanism, i.e.,outer-sphere associations. These aggregates rapidlyrearrange themselves into Al13 nuclei. The Al13 nucleiin paths IIb, IIc, and III then transform into a poorlyordered phases, i.e., microcrystalline boehmite(pseudoboehmite), which then transforms into gibb-site. The relative importance of each pathwaydepends greatly on reaction conditions. These solubleAl species and poorly ordered to noncrystallineAl species formed in the reaction pathways are highlyreactive and, thus, important in the environment.The transformation of Al via the various pathwaysdescribed above are strongly influenced by the natureand concentrations of soluble inorganic and organicions and solid state ions, such as clay minerals andhumic substances (Huang and Violante, 1986; Huang,1988; Bertsch and Parker, 1996; Krishnamurti et al.,1999, 2004; Huang et al., 2002).

The influence of organic substances has beenstudied most extensively on the particular Al solidphases that form (Huang et al., 2002). Theinfluence of a particular organic acid is generallyrelated to the stability constant of the complexthat the acid forms with Al (Kwong and Huang,1979a). Therefore, p-hydroxybenzoic acid, whichforms a low stability complex with Al, does notinhibit the crystallization of Al hydroxides, whereasaspartic, malic, and citric acids increasingly retardcrystallization (Fig. 4). This is because organic acidsreplace water molecules that would otherwisecoordinate with Al3+ ion and the extent to whichthis occurs depends on the chemical affinity of theorganic acid for the Al, i.e., the stability constantof an Al-organic complex. Humic substances such asfulvic acid (FA) and humic acid (HA) also influencethe transformation of Al by promoting the forma-tion of SRO Al hydroxides (Kodama and Schnitzer,

1980; Singer and Huang, 1990). FA and HA resemblealiphatic acids such as citric and malic acids, in thatthey contain COOH and aliphatic OH groups. Theyalso resemble quercetin and tannic acid, becausethey contain phenolic hydroxyl and ketonic CQOgroups. Through these functional groups, FA and HAform stable complexes with Al and inhibit anddisrupt the crystallization of Al hydroxides.

Organics substantially enhance the specific sur-face area and alter surface charge characteristicsof Al transformation products (Kwong and Huang,1978, 1979c, 1981). The intermediate transforma-tion products of Al, which include soluble mono-nuclear and polynuclear Al species and colloidalSRO hydroxides, are the most reactive Al species ininfluencing biogeochemical processes in the terres-trial ecosystem. These Al species influence envir-onmental quality and ecosystem health byimpacting acidification of the environment, forma-tion of humic substances, development of soilstructure, dynamics of organic C cycling, transfor-mations of nutrients and pollutants, biologicalproductivity and food chain contamination, micro-bial and enzymatic activity, and human and animalhealth (Huang et al., 2002).

Interactions of soil minerals with organic sub-stances and microorganisms also have an enormousimpact on the formation and transformation of SROFe oxides (Schwertmann et al., 1986; Cornell and

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Schwertmann, 2003). Microorganisms may influ-ence Fe transformation through reduction andoxidation reactions. Soil organic matter (SOM) andmicroorganisms play a vital role in influencingredox cycling of Fe, hydrolytic reactions of Fe,crystallization of Fe precipitation products, and thesubsequent formation of pedogenic Fe oxides(Huang and Wang, 1997). Furthermore, the finescale morphology, mean surface roughness, fractaldimension, specific surface, microporosity, andsurface charge properties of Fe oxides are greatlyinfluenced by organic substances (Huang and Wang,1997; Liu and Huang, 1999). The surface of Feoxides is the region of their interactions with thesoil solution, organic and inorganic particles, plantroots, microorganisms, and other soil biota. Surfaceproperties of Fe oxides should have profoundimpacts on microaggregate formation, water flux,nutrient and pollutant flux, soil biota habitat, andthe ability of soils to promote plant growth, torespond to management, and to resist degradation.Therefore, surface properties of Fe oxides formedunder the influence of organic substances deserveclose attention in advancing our understanding oftheir surface chemistry pertaining to biogeochem-ical processes in the terrestrial ecosystem.

Mineral catalysis of formation of humicsubstances

Humic substances are formed through biotic andabiotic processes. A variety of biomolecules, suchas carbohydrates, phenolic compounds, and aminoacids, can participate as raw materials. Soil mineralsurfaces play a vital role in the catalysis of abioticformation of humic substances.

The Maillard reaction (Maillard, 1913) is per-ceived to be a major pathway in humificationbecause of significant similarities between humicsubstances and melanoidins formed through thispathway involving sugar-amino acid condensations(Ikan et al., 1996). The presence of characteristicproducts of the Maillard reaction (alkyl pyrazines)

Figure 5. Formation of the Amadori compound from D-glucoNote that a molecule of water is split off. The Amadori compwhich is perceived as one of the pathways in humification.

was detected in archaeological plant remains up to1500 years in age (Evershed et al., 1997). Despitethe importance of the Maillard reaction, themechanisms and rates of the Maillard reaction innature remain vague (Ikan et al., 1996). The greatappeal of the Maillard reaction in humificationprocesses lies in the two proposed precursors,sugars and amino acids, which are among the mostabundant constitutions of terrestrial and aquaticenvironments (Anderson et al., 1989). A majorcriticism of the Maillard reaction has been that it isvery slow under ambient conditions (Hedges, 1988).In order to elucidate some details of the process,Jokic et al. (2001a) applied molecular topologicalanalysis to investigate the initial reaction betweenD-glucose and glycine to form the Amadori com-pound fructosylglycine, which is an intermediateproduct in the Maillard reaction (Fig. 5). Theircalculations show that fructosylglycine and waterand D-glucose and glycine as separate entitiesare very close to each other in terms of theirground state energy. Therefore, the potentialenergy barrier is high and the reaction between D-glucose and glycine alone to form fructosylglycineis very slow at room temperature. However, it hasrecently been demonstrated that the action of d-MnO2 under ambient environmental conditionssignificantly accelerates the Maillard reaction,lending credence as an important abiotic pathwayfor the formation of humic substances (Jokic et al.,2001b).

Besides the mineral catalysis of the Maillardreaction, soil minerals play an important role inaccelerating abiotic polymerization of phenoliccompounds, the polycondensation of phenoliccompounds and amino acids, and the subsequentformation of humic substances (Wang et al., 1986;Huang, 2000). Kumada and Kato (1970), Wang andLi (1977), and Filip et al. (1977) are among thepioneers in the study of browning of polyphenolscatalyzed by clay-size layer silicate. Since the early1980s, Huang and co-workers have studied thesequence of catalytic power of layer silicates andtheir reaction sites in the polymerization ofphenolic compounds and the subsequent formation

pyranose (D-glucose) and glycine (Mossine et al., 1994).ound is an intermediate product in the Maillard reaction

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Table 1. Effect of birnessite on glucose and glycine solutions at an initial pH 7.00 (Jokic et al., 2004a)

Treatmenta Absorbance pH Dissolved Mn (mgmL�1) pH+pE

400 nm 600 nm

45 1CGlucose+glycine+birnessite 10.470.3 0.7470.01 7.8970.08 82867108 10.370.3Glucose+glycine 0.3470.02 0.09370.005 7.0370.05 0 13.870.2Glucose+birnessite 0.04370.002 0.02770.001 7.4770.08 2797775 12.070.2Glucose 0.03870.002 0.02270.001 6.6970.04 0 13.270.2Glycine+birnessite 0.02170.001 Colourlessb 7.4670.04 20.970.8 13.970.1Glycine 0.02070.001 Colourlessb 6.9570.05 0 13.270.2

25 1CGlucose+glycine+birnessite 0.11470.007 0.02370.001 7.3970.11 2341756 13.870.2Glucose+glycine 0.02570.001 0.01170.001 7.0670.05 0 14.070.3Glucose+birnessite Colorlessb 6.8570.09 1001717 13.070.3Glucose Colorlessb 6.6570.07 0 13.470.2

aThe reaction periods were 15 and 30 days, respectively, at 45 and 25 1C.bThe supernatant was colorless with an absorbance o0.01 at the wavelengths studied.


of humic substances (Shindo and Huang, 1985a, b;Wang and Huang, 1986, 1988, 1994). Among Al, Feand Mn oxides, hydroxides and oxyhydroxides, Mnoxides are the most effective catalysts in thetransformation of phenolic compounds because oftheir high oxidation potentials, high specific sur-face area, and high surface reactivity (Shindo andHuang, 1982, 1984; Wang and Huang, 2000a).Manganese oxides (birnessite, cryptomelane, andpyrolusite), which are common in soils, act as Lewisacids that accept electrons from phenolic com-pounds, resulting in the formation of semiquinones,oxidative polymerization, and formation of humicsubstances. The abiotic ring cleavage of polyphe-nols catalyzed by SRO oxides of Al, Fe and Mn (Wangand Huang, 2000a, b) may account, in part, for thecarboxylic group contents and the origin of thealiphaticity of humic substances in soils (Schnitzer,1977; Preston et al., 1982). Therefore, the cataly-tic power of these oxides in affecting the Cturnover and humic substance formation via abioticprocesses in soil and related environments deservescontinued close scrutiny.

Both the Maillard reaction and the polymerizedpolyphenol model are considered as separatesignificant pathways for the formation of humicpolycondensates. In nature, however, it is likelythat these two pathways do not occur separately,but rather closely interacting with each other, sincesugars, amino acids, and polyphenols coexist in soilsolutions and natural waters. Jokic et al. (2004a)recently reported that soil mineral colloids such asd-MnO2 significantly accelerate humification pro-cesses in systems containing glucose, glycine andcatechol at ordinary temperatures and pH. Thedata clearly show the significant increase in

browning exhibited by systems containing d-MnO2

in contrast to non-d-MnO2 containing systems(Table 1). The promotional effect of d-MnO2 onthe system consisting of carbohydrate, amino acid,and polyphenol is a complex process involvingsurface mineral sorption and condensation. Thispoints to a linking of the polyphenol and Maillardreactions in humification pathways, which is asignificant advancement in the understanding ofnatural humification processes in the environment.

Effects of soil colloids on enzymaticactivity

Extracellular enzymes are rapidly sorbed atmineral and humic colloids in the terrestrialecosystem. Mineral colloids have a high affinityfor enzymes. However, the ability of the sorbedenzymes to retain their catalytic effectiveness isgreatly influenced by the nature and properties ofmineral colloids. Humic substances also have theability to retain enzyme and influence the activityof enzymes.

Mineral colloid-enzyme interactions have beenwell documented (e.g., Theng, 1979; Burns, 1986;Naidja et al., 2000, 2002). Adsorption of enzymesby mineral colloids may proceed through electro-static, ionic, covalent, hydrophobic, hydrogenbonding, and van der Waals forces. When enzymesare adsorbed on mineral colloids, changes in thetertiary structures (i.e., the folding of the helix orcoil into a compact, globular molecule stabilized byinterfold hydrogen bonding, van der Waals, andhydrophobic interactions) of the enzymes and their

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Table 2. Immobilization of a laccase (from Trametes versicolor) and a peroxidase (from horseradish) on differentsupport surfaces (Gianfreda and Bollag, 1994)

Enzyme and support Enzymatic activity

Protein adsorbeda

(mg per %)Units adsorbedb Specific activityc Residual specific

activityd (%)

LaccaseGlass beads 0.452/56 28.8 63.7 236Montmorillonite 0.622/71 19.8 31.8 118Kaolinite 0.566/64 13.1 23.1 85.5Soil 0.644/73 15.7 24.4 90.4

PeroxidaseGlass beads 0.092/17 8.4 91.6 93.8Montmorillonite 0.224/43 23 102.8 105.2Kaolinite 0.120/23 9.5 78.9 80.7Soil 0.162/31 15 92.6 94.8

aDifference between proteins initially added to 200mg of support (0.88mg laccase and 0.52mg of peroxidase) and those recovered inthe supernatant and washings.bExpressed as mmol O2 consumed min�1 for laccase and mmol guaiacol transformed min�1 for peroxidase.cUnits adsorbed/protein adsorbed.dCalculated as percentage of the specific activity (sa) of the free enzyme (laccase, sa ¼ 27mmolmin�1mg�1; peroxidase,sa ¼ 97.7 mmolmin�1mg�1).


active sites decrease the activity of the enzyme oreliminate it altogether (Burns, 1986). However,there are notable exceptions to the adsorption-decline in activity rule. Various materials havedifferent enzyme immobilization capabilities(Table 2). There is considerable variation in theretained activities of the two enzymes (laccase andperoxidase) on four materials (glass beads, mon-tmorillonite, kaolinite, soil). The residual specificactivities (calculated as percentages of the specificactivities of the free enzyme) of laccase andperoxidase immobilized on all materials are high.Furthermore, laccase immobilized on montmorillo-nite shows specific activities higher than that of thefree enzyme. This may be attributed to stericmodification of the immobilized enzyme. There-fore, the performance of enzymes in the terrestrialecosystem is substantially influenced by mineralcolloids. The role of SRO oxides, hydroxides, andoxyhydroxides of Al, Fe, and Mn in influencingenzyme activities involved in the transformationsof natural organics and xenobiotics has been welldocumented (Huang, 1990; Naidja et al., 1997,2000; Huang et al., 1999; Violante and Gianfreda,2000).

Fewer studies investigated the adsorption ofenzymes to humic substances, probably becauseof the physicochemical and biochemical complexityand dynamic nature of SOM. Nevertheless, it hasbeen reported that SOM can inhibit enzymes(Vuorinen and Saharinen, 1996) and enzyme activ-ity may be reduced by adsorption on humic

polymers (Gianfreda et al., 1998). Enzyme inhibi-tion of humic substances has been well demon-strated for an oxidoreductase (Pflug, 1980; Sarkarand Bollag, 1987), a protease (Ladd and Butler,1969), an invertase and a phosphatase (Malcolmand Vaughan, 1979). On the other hand, Kang et al.(2002) and Park et al. (2000) demonstrated that,although high concentrations of humic-like poly-mers tend to inhibit laccase-mediated transforma-tion of xenobiotics (including catechol), lowconcentrations of HA might enhance the enzymatictransformation of phenolic compounds. Further-more, it has been reported that enzymes can bestabilized by SOM (Conrad, 1942; Burns, 1986;Nannipieri and Gianfreda, 1999).

Mechanisms proposed to account for the stabilityof enzymes as influenced by SOM include ionexchange, H-bonding, covalent bonding, lipophylicreactions, and entrapment within three-dimen-sional micelles during organic matter genesis.However, the intrinsic mechanisms of interactionsof enzymes with humic substances still remain to beestablished. Functional groups of enzymes impli-cated in covalent bonding to humic polymersinclude terminal and basic amino, carboxyl, sulfhy-dryl, phenolic, and immidazole groups. All thesemay be involved in the stabilization processes,provided they do not form part of the active sites ofthe enzymes and are not crucial to the retention ofits tertiary structure.

Enzyme–humic complexes can also be bound tomineral colloids and this may further enhance

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enzyme stability. Violante and Gianfreda (2000)demonstrated the enhanced stability of enzymesbound to organo-mineral complexes. Furthermore,a significant portion of immobilized enzyme in soilsis associated with humic substances, not byadsorption or occlusion, but by the formation ofenzyme-phenolic copolymers during soil formation(Burns, 1986; Huang, 1990). Soil minerals withdifferent structural configuration and surface prop-erties may play a catalytic role in the formation ofenzyme-phenolic co-polymers and the subsequentimpact on the activity and stability of enzymes inthe terrestrial ecosystem.

Microbial mediation of soil mineraltransformation

Weathering of minerals in the terrestrial ecosys-tem can be enhanced by microbial activity by asmuch as 106 folds (Kurek, 2002). Microbes candissolve minerals by direct and indirect actionunder aerobic and anaerobic conditions (Ehrlich,2002; Kurek, 2002). The modes of microbial attackof minerals include: (1) direct enzymatic oxidationor reduction of a reduced or oxidized mineralcomponent, (2) indirect attack with a metabolicallyproduced redox agent or inorganic and organicacids, (3) indirect attack by metabolically producedalkali, usually in the form of ammonia, (4) indirectattack with a metabolically produced ligand thatform a highly soluble product with a mineralcomponent, and (5) indirect attack by biopolymers.Microorganisms may be dispersed in the soilsolution or grow in biofilms on the surface of thesusceptible minerals. Weed et al. (1969) haveshown that fungi can adsorb K from solution andthus shift K equilibrium in suspensions of micas andtransfer them to vermiculites. Such process canalso occur for many major and trace elements(Robert and Berthelin, 1986). Some microbes canpromote the transformation of one mineral intoanother by diagenesis, which can be an indirecteffect of aerobic and anaerobic microbial metabo-lism (Ehrlich, 2002; Kurek, 2002).

In view of their active surfaces, bacteria haverecently been included in geochemical modeling ofmetal speciation and transport. The large surfacearea-to-volume ratio and charge characteristics ofbacteria facilitate their metal-binding capacity,resulting in precipitation of mineral phases on theircell walls or other surface polymers (McLean et al.,2002). The mechanisms by which bacteria initiatethe formation of minerals vary widely with species.There may be a combination of biochemical and

surface-mediated reactions in the formation pro-cess of minerals. Bacteria surface layers maypassively adsorb and may indirectly serve as anucleation template. Bacteria may also directlyinitiate mineral precipitation by producing reactivecompounds (e.g., enzymes, metallothioneins, side-rophores), which bind metals or catalyze metaltransformation. Moreover, bacteria can instigatethe spontaneous precipitation of metals by alteringthe chemistry of the microenvironment (Beveridgeet al., 1997; Douglas and Beveridge, 1998). Forexample, reactive inorganic ligands such as sulfidemay be produced as a cellular metabolic by-product. Catalyzed by sulfate-reducing bacteria,sulfide reacts with metals to form metal sulfides, acommon reaction in anoxic environments. Anotherexample of microbially mediated fine-grainedmineral development is the formation of Mn oxides.Microbial oxidation of Mn(II) is a major process thatcan produce Mn oxide coatings on soil particles 105

times faster than abiotic oxidation (Tebo et al.,1997). Manganese oxides are highly reactive miner-als and help restrict the mobility of metals insoils through adsorption on their surfaces. BiogenicMn oxides have significantly larger specific surfacearea and higher Pb adsorption capacity thanabiotically precipitated Mn oxides (Nelson et al.,1999). Thus, bioformation of minerals can playa role in developing innovative managementstrategies for remediation of metal-contaminatedsoils.

Dynamics of aggregate turnover

Living and nonliving organic components varyingin size, chemical composition, and extent ofdecomposition have the capacity to promote andstabilize aggregations of soil particles at size scalesranging over nine orders of magnitudes (Fig. 6).Thus, they are essential in maintaining the properbalance of components of SOM to ensure thestability of the entire soil matrix. Mechanisms ofstabilization of soil structure can operate overlarger distances to bind microaggregates intomacroaggregates. Due to the distances involved,the stabilization of macroaggregates is related tothe presence of nonliving particulate organicmatter (POM) capable of spanning distances4100 mm or the existence of a network of a fungalhyphae and plant roots that physically enmeshesmicroaggregates (Fig. 7). The death of roots andhyphae growing within and through macroaggre-gates produces biochemical binding agents capableof stabilizing the structure of macroaggregates.

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Figure 6. Size scales associated with soil mineral particles, organic components, pores and aggregations of mineral andorganic components. The definition of pore size have used those developed by IUPAC (micropores o2 nm, mesopores2–50 nm and macropores450 nm). Alternatively, the pore sizes corresponding to the lower (cm ¼ �1500 kPa) andupper (cm ¼ �100 kPa) limits of water availability to plant may be used to define the boundaries between the differentclasses of pore size. cm is soil water metric potential (Baldock, 2002).


However, SOM responsible for the stabilizationof soil structure is subject to decomposition bymicrobial activity. Thus, aggregation is a dynamicprocess in soils (Baldock, 2002). Biological activityhas the potential not only to stabilize soil struc-tures through the production of organic substancescapable of binding soil particles, but also todestabilize soil structure by decomposition oforganic binding agents. The balance between these

two processes dictates soil structural stability. Acontinual addition of organic matter is essential toensure that the contents of aggregating agents areadequate to maintain soil structural stability.

The model proposed by Golchin et al. (1997)can be used to relate the physical and chem-ical properties of soil organic materials andtheir distribution and cycling to the stabilizationof soil aggregates. The major organic materials

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Figure 7. A conceptual model of aggregate hierarchy in soils where organic materials play an important role in thestabilization of aggregates (Jastrow and Miller, 1997).


contributing to the structural stability of soils areperceived to be free and occluded POM fractionsand the biomolecules synthesized and exuded byroots, mycorrhizal fungi, and other microorganismsin soils. The model of Golchin et al. (1997), whichdepicts the dynamics of soil aggregation, is mostapplicable to the aggregation of soils with appreci-able contents of clays that protect aggregatingagents from rapid decomposition (Baldock andSkjemstad, 2000). The protection mechanisms thatstabilize organic binding agents from microbialattack are less effective in sandy soils, i.e., a moredynamic process. The large size of mineral particlesand pores in sandy soils reduces the effectivenessof polysaccharides and other unaltered and alteredbiomolecules to stabilize soil structure. Thesebiomolecules are important to the adherence ofsoil particles to plant roots and fungal hyphae andsoil microorganism to mineral particles. However,they do not have the capacity to span the distancesbetween the large primary particles and stabilizesoil structure.

Biogeochemical cycling of C, N, P, and S

Soil minerals play a stabilizing role in SOM andrelated turnover of C, N, P, and S. The interactionof Al and Fe with organic components is of primaryimportance in determining the content of organic

matter in tropical and temperate soils (Wada,1995). The Al and Fe that complex and stabilizeorganic matter against microbial decomposition arereleased from soil minerals during soil formation.Their supply rates affect the content of SOM to agreat extent as illustrated by the relationshipbetween pyrophosphate-extractable C and pyro-phosphate-extractable Al plus Fe (Fig. 8). Closeinteractions between soil mineralogy and soilorganic C have also been demonstrated by Torn etal. (1997). The amount of C stabilized by non-crystalline minerals is much greater than that bycrystalline minerals. A positive relationship be-tween noncrystalline minerals and organic C existsin soils through the climate gradient. Since pyr-ophosphate-extractable Al and Fe are in the formof SRO minerals, the findings of Wada and Higashi(1976) and Wada (1995) are in accord with those ofTorn et al. (1997). Therefore, interactions of SROminerals with organic matter should largely controlthe storage and turnover of SOM and atmosphere-ecosystem carbon fluxes during long-term soildevelopment.

More than 95% of the N and S and between 20%and 90% of the P in surface soils are present in SOM(Guggenburger and Haider, 2002). The turnover oforganic C is closely associated with the dynamics ofN, P, and S in soils. The stabilization and degrada-tion of SOM and its nutrients are greatly influencedby the surface reactivity of soil mineral colloids.Interactions of minerals with SOM largely result in

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Figure 9. N Is XANES spectra of (a) fulvic acid isolatedfrom a glucose–glycine-d-MnO2 system and (b) thelyophilized solid phase. The peaks are assigned topyridinic (398.6 eV), pyridon (400.7 eV), amide(401.3 eV) and pyrolic (402.0 eV) moieties (Jokic et al.,2004b).

Figure 8. Relationship between pyrophosphate-extractable C and pyrophosphate-extractable (Al+Fe) in JapaneseAndisols of different ages (Wada and Higashi, 1976).


the stabilization of organic substances. Some soilminerals such as birnessite (d-MnO2) enhance thecycling of the organically bound nutrients bycatalytic degradation of biomolecules such asamino acids (Wang and Huang, 1987) and associatedhumification (Wang and Huang, 1992; Jokic et al.,2001b, 2004b). The catalysis of the Maillardreaction, condensation of sugar and amino acid byd-MnO2 and the resultant humification and theformation of unknown N are illustrated in Fig. 9.

Under most vegetation, the mass of humus in thesoil profile exceeds the combined content oforganic matter in the forest floor and above groundbiomass (Schlesinger, 1977, 1997). Table 3 shows aglobal inventory of plant detritus and SOM, totaling1456� 1015 g C. Alternative estimates based on soilgroups or climatic regions are similar (Post et al.,1982; Eswaran et al., 1993; Batjes, 1996). Theglobal estimates of SOM, divided by the estimate ofglobal litterfall, suggests a mean residence time ofabout 30 years for the total pool of soil organic C,but the mean residence time varies over severalorders of magnitude between the surface litter andthe various humus fractions (Schlesinger, 1997).The turnover times of microbial by-products andresistant plant residues adsorbed on soil particlesare in the order of years. Fulvic acids (FAs) haveturnover times in the neighborhood of hundred ofyears, whereas humic acids (HAs) and humins aremuch longer, i.e., thousands of years (Paul and Van

Veen, 1978). The distribution and annual transfersof C in the various fractions for a grasslandChernozem are shown in Fig. 10. Although the HAs

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Table 3. Distribution of soil organic matter by ecosystem types (Schlesinger, 1977)

Ecosystem type Mean soil organicmatter (kg Cm�2)

World area(ha� 108)

Total world soilorganic carbon (mtC� 109)

Amount in surfacelitter (mt C� 109)

Tropical forest 10.4 24.5 255 3.6Temperate forest 11.8 12 142 14.5Boreal forest 14.9 12 179 24.0Woodland andshrubland

6.9 8.5 59 2.4

Tropical savanna 3.7 15 56 1.5Temperategrassland

19.2 9 173 1.8

Tundra and alpine 21.6 8 173 4.0Desert scrub 5.6 18 101 0.2Extreme desert,rock, and ice

0.1 24 3 0.02

Cultivated 12.7 14 178 0.7Swamp and marsh 68.6 2 137 2.5Totals 147 1456 55.2


and humins constitute the majority of the organic Cin this system, they contribute only a smallproportion to the annual C cycling within the soilbecause of their very slow turnover rate. Theundecomposed litter (Fig. 10) also includes the soilbiomass and microbial metabolites. These, alongwith the plant residues, constitute the activefraction of organic matter that has a prominentrole in the cycling of elements such as C, N, and Sannually.

Jenkinson et al. (1991) estimated the additionaldegradation effects on SOM if the global annualmean temperature rises during the next 60 years by3 1C. According to their estimate, about 100Gtshould be additionally evolved as CO2 from SOM(1600Gt C). This will increase the present CO2

concentration in the atmosphere by 14% whereasthe combustion of fossil fuels (5.4Gt C year�1)would add only about 330Gt C to the atmosphere.Furthermore, transformations of C, N, and S in soilsas influenced by land management and the impacton global climatic change should not be overlooked(Lal, 1998; Guggenburger and Haider, 2002).

The release of the stored organic nutrientsdepends on the mean residence time of SOMassociated with the mineral phases. The meanresidence time of SOM varies widely with the typeof the organo-mineral associations and the spatiallocation within the aggregate structure of soil(Table 4). Some SOM with rapid turnover isapparently microbial residues, which are oftenloosely bound to mineral colloids. By contrast,SOM of slow turnover is tightly bound to sesqui-

oxides by the formation of inner-sphere complexes.Primary particles (layer silicates and Al and Feoxides, hydroxides, and/or oxyhydroxides) areclustered to larger entities, i.e., the secondary soilparticles or aggregates in soil environments. Thearrangement of primary and secondary particleswhich is defined as soil structure results in theformation of a series of pores ranging in size fromnanometers to centimeters. Soil structure deter-mines the pore-size distribution and, hence, theaccessibility of substrates adsorbed on surfaces of aseries of pores to exoenzymes, microbes, andfaunal grazers. Pools I and II (Table 4) are generallyplant fragments which may be divided into easilyavailable cell constituents and lignocelluloses atvarious degrees of degradation. When these plantfragments are degraded or reduced in size, themicrobially available components are exhaustedand the residues are more resistant to microbialattack. Pool III includes soil biomass and rapidlyavailable organic matter within large aggregates.The pools I to III, which on average account for20–30% of the total C in SOM, must be renewedcontinuously by fresh plant residues in order tomaintain a relatively constant nutrient level andrelease by mineralization. The balance betweendecay and renewal processes controls the avail-ability of N, P, and S and is sensitive to landmanagement. Pool IV is physically protected SOMwhose stability is affected by cultivation, sincephysical disturbance such as ploughing destroysmacroaggregates. Pool V is chemically stabilizedSOM having the longest mean residence time and

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Figure 10. Turnover of litter and soil organic fractions in the 0–20 cm layer of a Chernozem in a grassland soil(Schlesinger, 1977). Note that mean residence time can be calculated for each fraction from measurements of thequantity in the soil and the annual production or loss (respiration) from the fraction. Carbon pools (kg Cm�2) and annualtransfers (kg Cm�2 year�1) are indicated. Total profile content of C to 20 cm is 10.4 kg Cm�2.


accounting for 50–70% of the total C in SOM and isgenerally considered to be little affected by landmanagement.

Transformation of organic contaminants

Adsorption and desorption of organic contami-nants by soils are key processes in the transforma-tion of organic contaminants in the terrestrialecosystem (Huang, 1990, 1994; Chiou, 2002). Theavailability of organic contaminants for uptake byorganisms, for chemical, biochemical, and micro-bial transformations, and for transport in solutionor in gaseous phases is affected by adsorption–

desorption processes. The main factors affectingthe adsorption–desorption of organic contaminantsin soils are the physicochemical characteristics oforganic contaminants, the nature and properties ofthe soil colloids, and physicochemical–biologicalinterfacial interactions in soil environments. Soil

mineral colloids, especially SRO metal oxides andsilicates, and humic substances, which have largeand physico-chemically active surface areas, pro-vide a major retention sink for many organiccontaminants. Furthermore, since a vast majorityof humic substances are bound with soil mineralcolloids, the chemistry of humus–clay complexesplay an important role in adsorption–desorptionreactions of organic contaminants. For example,Liu and Huang (2004) reported that the complexa-tion of Al, Fe, and Mn oxides with humic macro-molecules substantially decreases the adsorption of2,4-dichlorophenoxyacetic acid (2,4-D) by metaloxides (Table 5), which is caused by the resultantalteration of the surface properties.

The degradation of organic contaminants maydecrease or diminish when they are adsorbed onsoil colloids due to the lower bioavailability ofchemicals involved in binding processes (Alexander,1995). The availability of sorbed organic contami-nants to microorganisms varies with the chemical

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Table

4.

Com

parison

ofestimated

mea

nresiden

cetimes

ofsoilorga

nicmatterin

soilphy

sica

lfrac

tion

s(G

ugge

nburge

ran

dHaider,20

02)

Pool

Mea

nresiden

cetime(yea

rs)

Jenk

insonan

dRay

ner(197

7)Pa

rton

etal.(198

7)Buy

anov

skyet

al.(199

4)Carter(199

6)

IDec

omposab

leplant

material,

0.24

Metab

olic

plant

residue

s,0.1–

1Ve

getative

frag

men

ts2–

0.2mm,0.5–

1Litter,1–

3

IIRe

sistan

tplant

material,3.33

Structural

plant

residue

s,1–

5Ve

getative

frag

men

ts0.05

–0.02

5mm,

1–3

Free

POM

(light

frac

tion

),1–

15

IIISo

ilbiomass,

2.44

ActiveSO

Mpoo

l,1–

5OM

inag

greg

ates

2–1mm,1–

4Microbialbiomass,

0.1–

0.4

IVPh

ysically

stab

ilized

OM,72

Slow

SOM

poo

l,25

–50

OM

inag

greg

ates

1–0.1mm,2–

10Interm

icroag

greg

ateOMa,5–

50V

Che

mically

stab

ilized

OM,28

57Pa

ssiveSO

Mpoo

l,10

00–15

00OM

infine

silt,�40

0Intram

icroag

greg

ateOMb

phy

sica

llyseque

stered

,50

–10

00OM

infine

clay,�10

00Che

mically

seque

stered

,10

00–30

00

aOrgan

icmatterstored

withinmac

roag

greg

ates

but

external

tomicroag

greg

ates;includ

esco

arse

occlud

edPOM

andmicrobialorga

nicmatter.

bOrgan

icmatterstored

withinmicroag

greg

ates;includ

esfine

occlud

edPOM

andmicrobialderived

orga

nicmatter.


properties of the contaminants, the nature ofsorbent, the mechanism of sorption, and propertiesof degrading organisms (Guerin and Boyd, 1992).

The mechanism of sorption may change withresidence time of xenobiotics present in soil,altering their bioavailability (Alexander, 1995).For instance, a strain of Pseudomonas mineralizedas much as 45% of phenanthrene during a 5-dayincubation with the compound that was in soilfor less than 1 day (Hutzinger and Alexander, 1995).When phenanthrene was aged with soil for 315days, only 5% of the compound was subjectto microbial mineralization. Phenanthrene mole-cules adsorbed on soil colloids in the initial fastadsorption are apparently more accessible todegradation in contrast to those adsorbed afterthe slow diffusion of the chemical across organicmatter and its subsequent sequestration in inac-cessible microsites within the soil matrix.

Contaminants sorbed by soil particles may beavailable to certain microorganisms and unavail-able to others. For example, soil-sorbed naphtha-lene is degraded by Pseudomonas putida strain17,484, but it does not undergo degradation by aGram-negative soil isolate, designated NP-Alk(Guerin and Boyd, 1992; Crocker et al., 1995).Apparently, the Pseudomonas strain directly miner-alizes surface-localized, labile sorbed naphthalene.By contrast, degradation of naphthalene by NP-Alkappears to be mainly controlled by the passivedesorption of the surface-sorbed compound (Guerinand Boyd, 1992).

Some researchers have provided evidence con-tradictory to the general principle that bindinginteractions should decrease the rates of transfor-mation of organic contaminants (Dec et al., 2002).Enhanced degradation of organic contaminants bymicrobes in the presence of soil colloids isattributed to their increased concentration in thevicinity of the microbial cells attached to organicand/or mineral colloids. However, both organicmatter and mineral colloids have the ability toinduce degradation through surface-catalyzed re-actions of adsorbed organic pollutants (Huang,1990, 2000; Senesi and Miano, 1995). Therefore,both biotic and abiotic degradation processes oforganics need to be taken into consideration instudying the effects of soil colloids on microbialdegradation of organic contaminants in the terres-trial ecosystem.

The attachment of microorganisms to solidsurfaces is a common phenomenon in soils. Micro-organisms may use many different mechanismsto achieve and maintain the attachment (Decet al., 2002). In most cases, microorganismsmanage to strengthen the attachment by means

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Table 5. Rate coefficients of the fast adsorption of2,4-D by various samples based on the 1st-order equationat 298 K (Liu and Huang, 2004)

Sample Rate coefficients� 103, h�1a

Al oxide 113.8Fe oxide 133.2Mn oxide 98.7Al oxide–humic complex 45.6Fe oxide–humic complex 43.2Mn oxide–humic complex 44.6HA 39.4

aStandard errors were o6%.


of extracellular polymeric adhesives such as poly-saccharides and proteins, which are frequentlytheir own metabolic products. Both electrostaticand hydrophobic interactions are involved inmicrobial attachment (Fletcher, 1996). Sorption ofmicroorganisms may be extremely extensive espe-cially in soils with high contents of clay and organicmatter. The effect of adsorption of microorganismon their ability to degrade organic contaminants isdifficult to predict. The microbial degradation maybe enhanced or reduced, or may remain unchangedas the cells undergo the adsorption to solidsurfaces. This apparently depends on the natureand properties of solid surfaces and the kinds ofmicroorganisms. However, the utilization of sorbedorganic compounds by microorganisms that adhereto the same surfaces is a common phenomenon (DiGrazia et al., 1990; Alexander, 1999). Our currentunderstanding of the effects of mineral and organiccolloids on microbial activity under field conditionsis limited (Haider, 1995), which may vary with thetype of soil components and microorganism. Theinfluence of structural configuration and surfaceproperties of soil minerals and organic matter onmicrobial activity and ability to degrade organiccontaminants with different structure and func-tionality should be a challenging area for futureresearch.

Soil is a living system in which enzymes arepresent either free in solution or bound ontomineral colloids, humic substances, and humus–mineral colloid complexes. Extracellular enzymesin solution are, however, rapidly bound to thosehighly reactive soil colloids. This immobilization is amajor factor that gauges the performance ofenzymes in the degradation of organic contami-nants in soils. The transformation and degradationof anthropogenic organic compounds includingchlorinated phenols, detergents, pesticides, andother organic contaminants can be mediated by

various enzymes (Munnecke et al., 1982; Bollag,1992; Naidja et al., 2000). The lifetime andstability of enzymes can be increased by immobi-lization on soil colloids (Burns, 1986; Dec et al.,2002). Enzymes immobilized by mineral colloidscan be used more efficiently with a possible longertime, higher stability, and possible reusability(Ruggiero et al., 1989; Naidja et al., 2000).However, information on the involvement of im-mobilized enzymes in the transformation of organiccontaminants in soils is sparse (Dec et al., 2002). Tocope with the problems caused by increasingamounts of agrochemical and industrial effluentsdisposed in soil, a fundamental understanding ofenzyme–soil colloid interactions and their role inremediation of anthropogenic organic compoundsin the environment is essential for the developmentof effective, durable, and practical treatmentmethods. However, the mechanisms of immobiliza-tion of enzymes on mineral colloids, humic sub-stances, and humus–clay complexes and theresultant changes of protein conformation ofenzymes remain relatively unclear. A wide rangeof soil minerals have yet to be investigated thatmay have a high capability to immobilize enzymes,adsorb substrates and reaction products and maybe catalytically active in the transformation oforganic pollutants. Furthermore, the role of abioticcatalysis of the minerals (Naidja and Huang, 1996)in enzyme–mineral colloid complexes is still at anemerging stage and warrants an in-depth research.

Transformation of metals and metalloids

The transformation of metals and metalloids insoils is governed by interactions of physicochem-ical, biochemical, and biological processes. Theimpacts of these interactive processes on metaltransformation are especially important in therhizosphere soil where the type and concentrationof substrates are different from those of the bulksoil because of root exudation (McLaughlin et al.,1998; Huang and Germida, 2002). This results incolonization by different populations of bacteria,fungi, protozoa, and nematodes. Soil–plant–biotainteractions, in turn, result in intense biological,biochemical, and physicochemical processes in therhizosphere. Therefore, reactions and processes inthe rhizosphere, which is the essential pathway ofmetal contamination of the terrestrial food chain,can only be understood satisfactorily with inter-disciplinary collaborative research efforts. Metaltransformation in soil and associated environmentsis governed by redox reactions, complexation,

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Table 6. Species and amounts of LMWOAs in tobacco rhizosphere soils treated with various Cd concentrations (Chouet al., 2003)

Treatment LMWOA(mmol kg �1 soil)

Acetic Lactic Glycolic Maleic Succinic Total

10a 260.8d 154.3c 797.3c 503.5b 173.7c 1889.65 146.4bc 145.3c 671.3ab 484.5ab 131.6ab 1579.11 113.5b 113.1b 594.6a 429.0a 114.6a 1364.9Control 59.9a 66.9a 488.5a 384.1a 98.1a 1097.4LSD 0.05 37.3 27.0 148.3 103.7 30.4

aCd concentration (mg Cd kg�1 soil).


adsorption–desorption, precipitations-dissolution,biomineralization, microbial extractions, and my-corrhizal infection.

Redox reactions are important in influencing thechemical speciation of a number of contaminantmetals and metalloids, such as Cr, Pu, Co, Pb, Cu,As, and Se (Oscarson et al., 1981; Bartlett andJames, 1993; Sparks, 2003). Redox reactions alsoare important in controlling the transformation andreactivity of Mn and Fe oxides in soils, which haveenormous capacities to adsorb metal and metalloidcontaminants and are major sinks of these con-taminants (Huang and Germida, 2002). Severaldissolved organic substances have been shown toeffect the reduction of ferric and maganic oxides;polyphenolic and hydroxy/carboxylic functionalgroups, such as found in FAs and HAs are particu-larly efficient in this regards (Campbell et al.,1988). Furthermore, reduction of sulfate to sulfidein anaerobic environments also affects the trans-formation of metal and metalloid contaminantsthrough the formation of highly insoluble metalsulfides (Alloway, 1995). The transformation, mo-bility, bioavailability, and toxicity of metals andmetalloids, which exist in more than one oxidationstate, are significantly influenced by redox reac-tions. The critical redox potentials for the trans-formation of some metal contaminants in soils havebeen summarized (Massechelyn and Patrick, 1994).However, there have been limited studies of howrhizosphere processes affect redox reactions ofmetals and metalloids.

A series of organic ligands occur in the rhizo-sphere due to root exudates and microbial meta-bolites (Lynch, 1990a, b; McLaughlin et al., 1998).In view of the stability constants of the complexesof metals with these ligands (NIST, 1997), a largefraction of metal ions in the solution may actuallybe complexed with a series of organic ligandscommonly present in the rhizosphere. Although thehypothesis that the toxicity or bioavailability of ametal is related to the activity of the free aquo ion

is gaining popularity in the studies of soil–plantrelations (Parker et al., 1995), new evidence is nowemerging that free metal ion hypothesis may not bevalid in all situations (Linehan et al., 1989; Tessierand Turner, 1995). Krishnamurti et al. (1996)reported that, after 2 weeks of crop growth in afield experiment, more Cd is complexed with thelow-molecular-weight organics acids (LMWOAs) atthe soil–root interface, compared with the bulksoil. More recent data show that the kind andamount of root exudates vary with the level of Cd incontaminated soils (Chou et al., 2003; Wang et al.,2003) as illustrated in Table 6. Hence, metalspeciation in relation to its bioavailability asinfluenced by physicochemical and biological inter-facial interactions in the rhizosphere warrantscareful research for years to come.

Metal transformation is significantly influencedby adsorption–desorption reactions in soils, whichare affected by biogeochemical processes, espe-cially in the rhizosphere. Factors influencingadsorption–desorption of metals in the rhizosphereinclude solution pH, LMWOAs, inorganic ligands,solution ionic strength, cation concentrations, andthe nature and properties of soil particles. Little isknown on the adsorption–desorption of metals inthe rhizosphere. Most studies investigated thesereactions in bulk soils and under modified condi-tions to simulate rhizosphere conditions. It hasbeen shown that the increase in Cd release in thepresence of LMWOAs can be explained by thesurface complexation of the particulate-bound Cdin soil with LMWOAs, which is reflected in theincrease in the release of Cd from the soils with theincrease in the stability constant of Cd–LMWOAscomplex (Krishnamurti et al., 1997). Also, the ratecoefficients of the Cd release from the soils aresubstantially influenced by LMWOAs and follow thesame order as that of the Cd availability index (CAI)values of the soils (Krishnamurti et al., 1995).Furthermore, amounts of Cd release from soils byrenewal of LMWOAs (Fig. 11) also follow the same

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Figure 11. Time function of Cd release from the Luseland soil as influenced by selected low molecular weight organicacids (LMWOAs) (citric, fumaric, oxalic,succinic, and acetic acids) at a concentration of 10�2 M. The cumulative amountof Cd released by renewal of LMWOAs after every 2-h reaction period is shown as solid line (Krishnamurti et al., 1997).


trend as the CAI values of the soils. More researchshould be conducted to uncover the dynamics ofthe adsorption–desorption reactions as influencedby biochemical and biological processes in soil andrelated environments.

For most of the trace metals, direct precipitationfrom solution homogeneous nucleation appears tobe less likely than adsorption–desorption by virtueof the low concentrations of these metals in soilsolution in well aerated drylands soils (McBride,1989; Christensen and Huang, 1999). On the otherhand, in reducing environments when the sulfideconcentration is sufficiently high, precipitation oftrace metals as sulfide may play a significant role inmetal transformation (Robert and Berthelin, 1986).In aerobic soils, although precipitation of tracemetals through homogeneous nucleation is unlikely,heterogeneous nucleation catalyzed by mineral,organic, and microbial surfaces may have asignificant role in metal transformation (Huangand Germida, 2002).

All microorganisms contain biopolymers, such asproteins, nucleic acids, and polysaccharides, whichprovide sites for binding metal ions (Hughes andPoole, 1989). These binding sites include negativelycharged groups such as carboxylate, thiolate,phosphate, and groups such as amines, whichcoordinate to the metal center through lonepairs of electrons. Because of the ability of thesebiopolymers to bind metals, large concentrationsof metals are frequently associated, not onlywith living microbial biomass, but also with deadcells (Berthelin et al., 1995). Many metals bindwith varying tenacity to the largely anionic outersurface layers of microbial cells. Binding of metalsby microbial cells alter cell wall compositionand induces morphological, ultrastructural andsurface charge changes (Venkateswerlu et al.,1989; Collins and Stotzky, 1996). Some metals arebound by cell walls to a greater extent than bymineral colloids (Table 7). This indicates thatmicrobial cell walls and membranes may act as

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Table 7. Metals bound by native B. subtilis walls, E. coli envelopes, kaolinite, and smectite (Walker et al., 1989)

Metal Amount of metal bound (mmol g�1) (oven-dry weight)a

Walls Envelopes Kaolinite Smectite

Ag 423715 17673 0.4670.02 4370.3Cu 530713 97279 570.03 19774Ni 654725 19073 470.2 173710Cd 683719 22176 670.2 170.02Pb 543711 25475 370.2 11876Zn 973713 529732 3771 6572Cr 435737 10272 870.5 3975

aThe data represent the average of three to five determinations for each sample from duplicate experiments and the standard error.


foci for accumulation of metals in the terrestrialecosystem.

Microbial formation of minerals, i.e., biominer-alization is an important pathway of mineralformation in nature (Beveridge, 1989). Researchon biomineralization indicates that specific mole-cular interactions at inorganic–organic interfacescan result in the controlled nucleation and growthof inorganic crystals (Mann et al., 1993). A centraltenet of biomineralization is that the nucleation,growth morphology, and aggregation (assembly ofthe inorganic crystals) are regulated by organizedassemblies of organic macromolecules, the organicmatrix. Control over the crystallochemical proper-ties of the biominerals is achieved by specificprocesses involving molecular recognition at inor-ganic–organic interfaces. Electrostatic binding orassociation, geometric matching (epitaxis), andsterochemical correspondence are important inthese recognitions. Compared with other lifeforms, bacteria may have a greater capacity tosorb and precipitate metals from solution resultingin mineral formation as they have the highestsurface area to volume ratio (Beveridge, 1989).Although in most environments, soluble metals arepresent in low concentrations, bacteria cells have aremarkable ability to concentrate metal ions fromsolutions. Therefore, the impact of biomineraliza-tion on the transformation, dynamics, toxicity, andfate of metal contaminants in the environmentshould not be overlooked.

Microorganisms have a range of metal transfor-mation abilities, which are often specific forcertain metals and capable of accumulating metalsagainst large concentration gradient (Kurek, 2002).For example, some microorganisms make side-rophores which are biomolecules capable of bindingFe (Nielands, 1981). Other microorganisms producemetallothioneins that are small cysteine richprotein that strongly bind Cd, Cu, and Zn. Ligands

of this type and related biomolecules are significantin influencing the transformation, transport, bioa-vailability, and toxicity of metals in soils, especiallyin the rhizosphere.

Most plants in natural habitats form associationwith mycorrhizae. Arbuscular mycorrhizal fungi(AMF) are obligate symbionts, and infection ofplant roots exerts a metabolic load on the hostplant (Reid, 1990). Mycorrhizal fungi producemycelium inside root cortical cells during infectionand colonization of host plant root. They may formstorage structures termed vesicles, and may alsoform other structures which are referred to asarbuscules and serve as the site of ion exchangebetween the host plant and the mycorrhizal fungus.The fungi also form extra cellular hyphae thatpenetrate out of the root and expand the volume ofsoil the root can penetrate. The influence ofmycorrhizal infection on metal uptake dependsupon the ability of the fungal symbiont to absorbmetals and transfer them to the symbiotic rootsthrough extensive vegetative mycelium. Mycorrhi-zal fungi also release LMWOAs into soils and thusenhance the solubilization of particulate-boundmetals (Huang and Germida, 2002). However, thereare substantial gaps in knowledge on the mechan-isms with which fungi influence the transformationand uptake of metals in the terrestrial ecosystem.

Biodiversity in terrestrial ecosystems

Soil is a focal point of the terrestrial ecosystem(Odum, 1989; Coleman et al., 1998). Physical,chemical, and biological processes are interactingprocesses in the terrestrial ecosystem. The func-tioning and stability of the terrestrial ecosystemare determined by plant biodiversity and speciescomposition (Schulze and Mooney, 1993; Tilman

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et al., 1996; Hooper and Vitousek, 1997). However,the ecological mechanisms by which they areregulated and maintained are not well understood.Therefore, it is essential to identify these mechan-isms to ensure successful management for con-servation and restoration of diverse naturalecosystems.

van der Heijden et al. (1998) reported thatbelow-ground diversity of AMF is a major factorcontributing to the maintenance of plant biodiver-sity and to ecosystem functioning. This indicatesthe need to protect AMF and to consider these fungias one of the items in future management practicesto maintain diverse ecosystems. Further, theyindicate that conservation of the fungal gene poolis likely to be a prerequisite for maintenance offlouristic diversity in grasslands and other ecosys-tems such as boreal forests, where fungal web isknown to affect allocation of resources betweenplant species. Mycorrhizal community is sensitiveto perturbations, particularly those associatedwith cultivation and nutrient enrichment (Helgasonet al., 1998). The findings of van der Heijden et al.(1998) demonstrate that biodiversity and ecosystemfunctions can be driven by microbial interactions.

Flouristically rich systems may display greaterstability under stress (Tilman and Downing, 1994),are more likely to mitigate global problems posedby atmospheric CO2 enrichment (Naeem et al.,1994), and are more productive (Tilman et al.,1996). Van der Heijden et al. (1998) reported thatmicrobial interactions can influence not only plantdiversity but also productivity. Both the plantbiodiversity, as measured by Simpson’s diversityindex (Fig. 12a), and productivity above and belowground (Fig. 12b and c) increased with increasingAMF-species richness. The lowest plant productivitywas found in those plots without AMF or with only afew AMF species. In contrast, plant productivity washighest when eight or 14 AMF species were present.This research indicates that plant productivity in agiven ecosystem can be dependent on the diversityof symbionts. The results also indicated a mechan-istic explanation for the effects of mycorrhizal-species richness on plant productivity. IncreasedAMF-species richness led to a significant increase inthe length of mycorrhizal hyphae in the soil (Fig.12d), to a decreased soil phosphorus concentration(Fig. 12e) and to an increased phosphorus content inplant material (Fig. 12f). Therefore, increasing AMFbiodiversity results in more efficient exploitation ofsoil phosphorus and a better use of the resourcesavailable in the system. The loss of AMF biodiver-sity, which may occur in agricultural systems(Naeem et al., 1994; Steitwolf-Engel et al., 1997),could decrease ecosystem productivity.

The findings of van der Heijden et al. (1998) haveclearly demonstrated the impact of the loss ofbelow-ground biodiversity on the decrease of plantbiodiversity and biological productivity of soils.This represents an understudied field of research,which requires increasing attention. A factor ofcentral importance of soil to ecosystem functions isthat soils on a global scale have a range ofcharacteristics, which enable an enormous arrayof microorganisms, plants, animals, and humans toco-exist and thrive. Humans have exploited theability of soils to provide massive amounts of food.More than 40% of the net primary production of theworld are exploited by humans (Vitousek et al.,1986). The exploitation is increasing with theaddition of 87.6 million people to the globalpopulation every year, with the rate additionsteadily increasing (Brown and Flavin, 1996). Theimpact of population and the accompanying in-tensification of agriculture and industrialization onecosystem functions and human health is, thus, ofincreasing concern.

The diversity of terrestrial vegetation systems iseverywhere under stress. The present reduction inbiodiversity on earth and its potential threat toecosystem stability and productivity are of concern(Tilman et al., 1996; van der Heijden et al., 1998).A recognition of the significance of plant biodiver-sity in ecosystem stability and productivity hasencouraged biologists to investigate the mechan-isms that determine and affect species compositionin plant communities. Future development ofinnovative land management strategies is essentialto protect the terrestrial ecosystem and to supplyfood to sustain human health. Below-groundmicrobial diversity substantially influences plantbiodiversity and ecosystem variability (van derHeijden et al., 1998). Furthermore, microbialevents are, in turn, significantly affected by soilsurface-reactive particles (Stotzky, 1986, 2002;Theng and Orchard, 1995). These reactive particlescan influence the metabolic processes, growth,adhesion, and ecology of below-ground biologicalentities by interactions of their reactive surfaceswith microbes. They can also indirectly affectthese microbial events by modifying the spatialarrangement of soil particles especially the poresize distribution within soil aggregates as well asthe physicochemical environments in which viruses,microbes, and macrobiota reside. Therefore, theeffect of mineral–organic component–micro-organism interactions on below-ground microbe-species richness and the impact on above-groundbiodiversity and biological productivity should,thus, be an issue of intense interest for yearsto come.

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Figure 12. The effect of AMF-species richness on different parameters (van der Heijden et al., 1998). Effects on: (a)the plant biodiversity, as measured by Simpson’s diversity index (fitted curve is y ¼ 0:271þ 0:077x20:003x2; r2 ¼ 0:63;Pp0:0001); (b) shoot biomass (y ¼ �0:334x2 þ 8:129x þ 72:754; r2 ¼ 0:69; Pp0:0001); (c) root biomass(y ¼ �0:265x2 þ 6:772x þ 96:141; r2 ¼ 0:55; Pp0:0001); (d) length of external mycorrizal hyphae in soil(y ¼ 0:001x320:046x2 þ 0:756x þ 2:979; r2 ¼ 0:60; Pp0:0001); (e) soil phosphorus concentration (y ¼ 0:065x221:593x þ 14:252; r2 ¼ 0:67; Pp0:0001); and (f) total plant phosphorus content (linear relationship;y ¼ 61:537x þ 1156:281; r2 ¼ 0:48; Pp0:001), in macrocosms simulating North American old-field ecosystems. Squaresrepresent means (7s.e.m.).


Conclusions

Minerals, organic matter, and microorganismshave significant higher order interactions in terres-trial ecosystems. These interactions exert strongcontrol over soil physical, chemical, and biologicalprocesses.

Organic substances and microorganisms substan-tially affect the formation and transformation ofSRO sesquioxides. These oxides exert great influ-ences on biogeochemical processes in the terres-

trial ecosystem. Soil mineral surfaces play a vitalrole in the catalysis of abiotic formation of humicsubstances. Soil minerals, especially noncrystallineminerals have the ability to chemically stabilizeSOM. Organic substances also act as binding agentsto promote and stabilize aggregation. Therefore,noncrystalline minerals retard the microbial degra-dation of SOM, which in turn inhibits the transfor-mation of noncrystalline mineral phases to morecrystalline minerals. Hence, there are distinctinteractive mechanisms between soil minerals,

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organic matter, and microbes, which should havea direct influence on the stability and cycling ofC, N, P, and S and global change.

To date, our understanding of the impacts of soilmineral–organic matter–microbe interactions onthe transformation and degradation of anthropo-genic organic compounds in soils is still relativelylimited. The knowledge on detailed surface struc-tures of the enzyme–mineral colloid and enzyme–-

humus complexes at the molecular level and themechanisms of their catalysis in the transformationof organic contaminants need to be advanced. Moreresearch should examine new enzymes and newmineral composite supports to be used for catalyticdegradation of a wide range of industrial andagricultural organic contaminants in soils.

Microorganisms and organics can profoundlyaffect the transformation of metals, metalloidsand minerals in the terrestrial ecosystem. Biomi-neralization is an important pathway of mineralformation in natural environments. More attentionshould be paid to environmental impact of biomi-neralization on the fate of metals and metalloids.Furthermore, little is known on the influence ofphysico-chemical and microbiological interactionson the transformation, dynamics and toxicity ofmetals and metalloids in soil and related environ-ments. The role of these interactions in remedia-tion of metal- and metalloid-contaminated soils isan emerging issue in environmental protection.

To sum up, more research should be conducted touncover the effects of physical–chemical–biologicalinteractions on biogeochemical reactions, which,in turn, govern ecosystem biodiversity, sustainabil-ity and productivity. Use of advanced analyticalinstrumentation, i.e., synchrotron-based X-ray andinfrared spectroscopies and atomic force micro-scopy and the like in this area of scientific pursuitsshould shed new light on the mystery of soilmineral–organic matter–microorganism interac-tions. Fundamental understanding of these inter-actions at the molecular level and the impacts onthe terrestrial ecosystem would facilitate ourdevelopment of innovative management strategiesto regulate the behavior of the terrestrial ecosys-tem on a global scale. Future research on thisextremely important and exciting area of scienceshould be stimulated to restore as well as sustainthe ecosystem integrity.

Acknowledgements

We acknowledge support by Discovery Grant 2383of the Natural Sciences and Engineering Research

Council of Canada and Grants NSC 92-2811-B-002-075 of the National Science Council, Taiwan.

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