Advances in Agronomy, Volume 99

CONTRIBUTORS Numbers in Parenthesis indicate the pages on which authors contributors begin V. C. Baligar (345) USDA-A...

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CONTRIBUTORS

Numbers in Parenthesis indicate the pages on which authors contributors begin

V. C. Baligar (345) USDA-ARS-Sustainable Perennial Crops Lab, Beltsville, Maryland 20705-2350 Guilhem Bourrie´ (227) INRA, UR 1119, Soil and Water Geochemistry, Europoˆle de l’Arbois, B.P. 80, F-13545 Aix-en-Provence (France) J. F. Briat (183) CNRS, Universite´ Montpellier II, SupAgro, INRA, UMR5004 ‘Biochimie et Physiologie Molećulaire des Plantes’, Place Pierre Viala, F-34060 Montpellier cedex I, France N. K. Fageria (345) National Rice and Bean Research Center of EMBRAPA, Caixa Postal 179, Santo Antoˆnio de Goia´s, GO, CEP. 75375-000, Brazil Rebecca E. Hamon (289) Plant Chemistry Section, Agricultural and Environmental Chemistry Institute, Faculty of Agricultural Sciences, Universita` Cattolica del Sacro Cuore, Via Emilia Parmense 84, I-29100, Piacenza, Italy Alfred E. Hartemink (125) ISRIC - World Soil Information, 6700 AJ Wageningen, The Netherlands P. Hinsinger (183) INRA, SupAgro, UMR1222 ‘Biogeóchimie du Sol et de la Rhizosphe`re’, Place Pierre Viala, F-34060 Montpellier cedex 1, France Philip M. Jardine (1) Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, TN 37831 P. Lemanceau (183) INRA, Universite´ de Bourgogne, UMR1229 ‘Microbiologie du Sol et de l’Environnement’, CMSE, BV 86510, F-21034 Dijon cedex, France Enzo Lombi (289) Plant and Soil Science Laboratory, Department of Agricultural Science, Faculty of Life Sciences, University of Copenhagen, Thorvaldsensvej 40, 1871 Frederiksberg C, Denmark

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Contributors

J. M. Meyer (183) CNRS, Universite´ Louis Pasteur, UMR7156 ‘De´partement Environnement, Geńe´tique molećulaire et Microbiologie’, F-67000 Strasbourg, France David R. Parker (101, 289) Soil and Water Sciences Section, Department of Environmental Sciences, University of California, Riverside, California 92521 A. Robin (183) INRA, Universite´ de Bourgogne, UMR1229 ‘Microbiologie du Sol et de l’Environnement’, CMSE, BV 86510, F-21034 Dijon cedex, France Angelia L. Seyfferth (101) Department of Environmental Sciences, University of California, Riverside, California 92521 Fabienne Trolard (227) INRA, UR 1119, Soil and Water Geochemistry, Europoˆle de l’Arbois, B.P. 80, F-13545 Aix-en-Provence (France) G. Vansuyt (183) INRA, Universite´ de Bourgogne, UMR1229 ‘Microbiologie du Sol et de l’Environnement’, CMSE, BV 86510, F-21034 Dijon cedex, France

PREFACE

Volume 99 contains seven comprehensive and timely reviews dealing with plant, soil, and environmental sciences. Chapter 1 is an excellent review on the influence that complex hydrological, geological, and biological processes have on inorganic contaminant fate and transport, with emphasis on field-scale studies. Chapter 2 focuses on the uptake and fate of perchlorate in plants. Chapter 3 is a timely review on the soil and environmental issues related to the use of sugarcane for bioethanol production. Chapter 4 is a comprehensive review on iron dynamics in the rhizosphere including the impact of plants and microorganisms on iron status and iron-mediated interactions in the rhizosphere. Chapter 5 deals with a reevaluation of the Fe cycling in soils in light of recent advances in understanding the geochemistry of green rusts and fougerite. Chapter 6 is a thorough review of recent advances on using isotopic dilution techniques in trace element research including a discussion of methods, benefits, and limitations. Chapter 7 deals with liming of tropical Oxisols and includes factors affecting lime requirements and methods and frequency of lime applications. I thank the authors for their fine contributions. DONALD L. SPARKS University of Delaware

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Influence of Coupled Processes on Contaminant Fate and Transport in Subsurface Environments Philip M. Jardine Contents 1. Introduction and Rationale 2. Chapter Objectives and Outline 3. General Overview on the Impact of Coupled Processes on Subsurface Fate and Transport 3.1. The importance of subsurface media structure 3.2. Influence of subsurface hydrologic processes on biogeochemical reactions 3.3. Influence of the subsurface capillary fringe on couple hydro-bio-geochemical reactions 4. Influence of Coupled Processes on Inorganic Contaminant Fate and Transport 4.1. General overview 4.2. Inorganic metals 4.3. Inorganic radionuclides 4.4. Inorganic ligands 4.5. General inorganics 4.6. Modeling coupled processes involving dissolved aqueous phase inorganic constituents 5. Influence of Coupled Processes on Organic Contaminant Fate and Transport 5.1. General overview 5.2. Chlorinated solvents 5.3. Hydrocarbons 5.4. Pesticides and herbicides 5.5. Modeling coupled processes involving organic constituents 6. Concluding Remarks Acknowledgments References

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Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, TN 37831 Advances in Agronomy, Volume 99 ISSN 0065-2113, DOI: 10.1016/S0065-2113(08)00401-X

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2008 Elsevier Inc. All rights reserved.

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Abstract The following chapter emphasizes subsurface environmental research investigations over the past 10 to 15 years that couple hydrological, geochemical, and biological processes as related to contaminant fate and transport. An attempt is made to focus on field-scale studies with possible reference to laboratory-scale endeavors. Much of the research discussed reflects investigations of the influence of coupled processes on the fate and transport of inorganic, radionuclide, and organic contaminants in subsurface environments as a result of natural processes or energy and weapons production endeavors that required waste disposal. The chapter provides on overview of the interaction between hydro-biogeochemical processes in structured, heterogeneous subsurface environments and how these interactions control contaminant fate and transport, followed by experimental and numerical subsurface science research and case studies involving specific classes of inorganic and organic contaminants. Lastly, thought provoking insights are highlighted on why the study of subsurface coupled processes is paramount to understanding potential future contaminant fate and transport issues of global concern.

1. Introduction and Rationale Until recently, worldwide waste disposal practices were an afterthought to the desire for economic expansion and national security and defense. In an age full of fear, greed, and the desire for global superiority, waste disposal practices regarding weapons, energy, and food production, and the quest for a higher standard of living, were of little consequence and were deemed an effort that future generations would confront. Unfortunately, cleanup technologies have been slow in development and the resolution of the legacy waste problem persists. An excellent example exists within several government agencies within the United States (U.S.) such as the Department of Energy (DOE) and the Department of Defense (DoD) which face a daunting challenge of remediating huge below ground inventories of legacy radioactive, toxic metal, and mixed organic wastes. The scope of the problem is massive, particularly in the high recharge, humid regions east of the Rocky Mountains, where the off-site migration of contaminants continues to plague soil water, groundwater, and surface water sources. Even in semiarid regimes west of the Rocky Mountains, the threat of contaminant migration through seemingly ‘‘dry’’ porous media persists due to slow water movement along fine sediment layers as a result of tension-driven anisotropic flow. Industrial activities have also contributed to massive legacy waste problems that are associated with accidental and intentional spills and disposal activities. The cleanup of these activities by DOE, DoD, and the U.S. Environmental Protection Agency (EPA) has

Influence of Coupled Processes on Contaminant Fate

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been ongoing for several decades with the pace slowing due to budget cuts and priority shifts in the U.S. government spending portfolio. In this context, it is not surprising that determining the best course of action— large-scale cleanups, focused hotspot remediation, or no action (natural attenuation)—remains exceedingly difficult from a technical standpoint. If a natural system has sufficient capacity for clean-up of contaminants by in situ processes (e.g., adsorption, dilution, precipitation, biodegradation, chemical transformation), perhaps natural attenuation processes should be considered as the first option. The current reality (i.e., 2008) is that contaminated sites are closing rapidly and many remediation strategies have chosen to leave contaminants in-place with little consideration of whether the decision is appropriate. In situ barriers, surface caps, and bioremediation are often the remedial strategies of choice. By choosing to leave contaminants in-place, we must accept the fact that the contaminants will continue to interact with subsurface and surface media. Contaminant interactions with the geosphere are complex and investigating long-term changes and interactive processes is imperative to verifying risks. Since contaminants may be left in-ground, it is critical to understand immobilization and remobilization processes that may operate during long-term stewardship as it is our societal responsibility to ensure a healthy environment for future generations. A deeper understanding of the relevant spatial and temporal scales that govern the fate of transport mechanisms is needed in order to make informed decisions about the applicability of various remediation options including natural attenuation. Understanding the spatial and temporal scales at which coupled hydrobio-geochemical processes operate is essential to designing an efficient and effective monitoring program for long-term stewardship.

2. Chapter Objectives and Outline In the following chapter we emphasize subsurface environmental research investigations that combine hydrological, geochemical, and biological processes as related to contaminant fate and transport. We do not consider coupled subsurface deformation, mechanical, or thermal processes as related to chemical distribution and reactivity. This information can be found in Bai and Elsworth (2000). We attempt to discuss only fieldscale studies with possible reference to laboratory-scale endeavors. A review of environmental investigations involving coupled processes at the laboratory scale can be found in Geesey and Mitchell (2008). Much of the research discussed in this chapter reflects investigation of the influence of coupled processes on the fate and transport of contaminants in subsurface environments as a result of natural processes or energy and weapons production endeavors that require waste disposal. Many of the approaches and research

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findings from these studies have potential application to future investigations on the environmental consequences of contaminant dissemination as a result of shifts in energy and climate policy and man-made changes to the global hydrologic cycle. Section 3 provides an overview of the interaction between hydrological, geochemical, and microbial processes in structured, heterogeneous subsurface environments and how these interactions control contaminant fate and transport. Next, Section 4 highlights recent field relevant research on the influence of these coupled processes on inorganic contaminant fate and transport, and Section 5 provides numerous examples of field-scale research on the impact of coupled processes on organic contaminant fate and transport. Lastly, Section 6 provides concluding remarks of how the study of subsurface coupled processes is paramount to understanding potential future contaminant fate and transport issues of global concern.

3. General Overview on the Impact of Coupled Processes on Subsurface Fate and Transport 3.1. The importance of subsurface media structure Undisturbed subsurface soils and geologic material consist of a complex continuum of pore regions ranging from large macropores and fractures at the millimeter scale to small micropores at the submicrometer scale. Structured media, common to most subsurface environments throughout the world, accentuates this physical condition which often controls the hydrological, geochemical, and microbial processes affecting transport phenomena. More often than not, subsurface media structure controls the rate and extent of geochemical and microbial reactions, all of which ultimately influence contaminant fate and transport processes. Geochemical and biological reactions and activity may, in turn, influence media structure and the hydrodynamics of the system (e.g., biogeochemical pore plugging, earthworm channels). Therefore, the extent and magnitude of subsurface biogeochemical reactions is often controlled by the spatial and temporal variability of the media structure which controls the system hydrodynamics. The physical properties of the media (e.g., structured, layered) coupled with its antecedent water content and the duration and intensity of precipitation events, dictate the avenues of water, solute, and microbe movement as well as their interaction within the subsurface. In humid environments where structured media is commonplace, transient storm events invariably result in the preferential migration of water (Gerke et al., 2007; Hornberger et al., 1991; Jardine et al., 1989, 1990a,b; 1998, 1999a, 2001, 2002; 2006, 2007; Mayes et al., 2003; Shaffer et al., 1979; Shuford et al., 1977; Vogel et al., 2006; Wilson et al., 1989, 1993, 1998).

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Highly conductive voids within the media (e.g., fractures, macropores) carry water around low permeability, high porosity matrix blocks or aggregates resulting in water bypass of the latter (Fig. 1A). Subsurface preferential flow is also a key mechanism controlling water and solute mobility in arid environments (Hendrickx and Yao, 1996; Ho and Webb, 1998; Liu et al., 1998; Mayes et al., 2003, 2005; Pace et al., 2003, 2007; Porro et al., 1993; Ritsema et al., 1993, 1998; Tompson et al., 2006). Lithologic discontinuities and sediment layering promote perched water tables and unstable wetting fronts that drive both lateral and vertical subsurface preferential flow (Fig. 1B). Water that is preferentially flowing through media often remains in intimate contact with the porous matrix, and physical and hydrologic gradients drive the exchange of mass from one pore regime to another. Mass exchange is time dependent and is often controlled by diffusion to and from the matrix. The preferential movement of water and mass through the subsurface therefore significantly impacts geochemical and microbial processes by controlling the extent and rate of various reactions with the solid phase. It imposes kinetic constraints on biogeochemical reactions and limits the surface area of interaction by partially excluding water and mass from the matrix porosity. These concepts are likewise conveyed in the subject area hydropedology which provides a link between the disciplines of pedology (e.g., soil B A

1 cm

10 cm

Structured saprolite Laminated sediments

Figure 1 An example of structured media from (A) humid and (B) semiarid climatic regimes showing a fractured shale-derived saprolite and a layered sediment consisting of laminated coarse- and fine-grained material, respectively. The fractured saprolite in (A) consists of macroporous fast-flowing fractures that surround low permeability, high porosity matrix blocks. The laminated sediments in (B) are irregularly spaced depositional layers of fine- and coarse-grained minerals that have drastically different hydrologic characteristics that often results in tension-driven anisotropic lateral flow along fine layers.

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macro- and micromorphology) and subsurface hydrology and other disciplines involved with land, air, and water interfaces (Kutilek and Nielsen, 2007). The coupling of such processes suggests that anisotropy is a general characteristic of soils and that the formulation of physically meaningful transport parameters requires quantitative knowledge of soil micromorphology. As suggested by Kutilek (1978, 1990), the assumption that soil is an isotropic body is only an approximation of reality. Coupling of hydropedology with geochemistry and microbiology provides new insights into the role of solute and contaminant fate and transport as a function of hydrology and soil structure.

3.2. Influence of subsurface hydrologic processes on biogeochemical reactions Subsurface geochemical and microbial reactions are directly linked to the system hydrodynamics. Soil moisture conditions that promote the onset of preferential flow and thus higher volumetric flux per unit area will minimize geochemical and microbial interfacial reactions due to decreased residence times during transport and potential bypass of the soil matrix (Estrella et al., 1993; Jardine et al., 1988, 1993a; Jarvis, 2007; Jarvis et al., 2007; Kung, 1990a,b; Maraqa et al., 1999). Conversely, soil moisture conditions that do not promote preferential flow will, in general, enhance geochemical retardation and microbial interfacial reactions. In the presence or absence of preferential flow, water content variations affect the extent and rate of geochemical and microbial reactions very differently. The extent of contaminant retardation by the solid phase via geochemical mechanisms (e.g., sorption, redox alteration, and complexation) will be more pronounced when flow is restricted to smaller pore size regimes (e.g., mesopores/micropores). Jardine et al. (1988, 1993a,b) have found that the reactivity of reactive contaminants and chelated radionuclides increased dramatically with a slight decrease in pressure head or water content. The larger surface area and potential reactivity of smaller sized pores versus macropores allow geochemical reactions to proceed to a more significant extent in the subsurface media. Microbial activity and transport in the subsurface are also controlled by physical and chemical interactions with the solid phase as well as the availability of nutrients, sources of carbon, and possible electron acceptors. Hydraulic conductivities can have a severe influence on nutrient transport and delivery within the subsurface and can often be the most limiting aspect of bioremediation. Biotransformation, biosorption, and electron transfer reactions are typical processes that govern the fate and transport of microbes in the subsurface. Unlike solutes that can reside within nearly all of the pore structure of subsurface media, microbes (i.e., bacteria and viruses) are too large to reach a significant fraction of the micropore regime and are restricted to the mesopore and macropore domains. Usually, less than 5–10% of the


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void volume in structured media is accessible to bacteria (Bales et al., 1989; Champ and Schroeter, 1988; Harton, 1996; Harvey et al., 1989, 1993; Jardine et al., 1998; McKay et al., 1993a,b; Smith et al., 1985; Wilson et al., 1993) because most of the soil porosity is contained within the micropore domain. However, microbial activity may actually accelerate solute mass transfer from micropores to larger meso- and macropores. Although bacteria cannot physically access most of the micropore regime, they can form biofilms around the soil aggregates and matrix blocks. These biofilms are permeable to the transfer of water and solutes between the various pore domains. It is possible that active biofilms that surround micropore domains accelerate the mass transfer of contaminants and solutes to the more biologically active pore regions. This may occur since microbial processes maintain a steep concentration gradient between the small and the large pores. The mass transfer process from one pore class to another may still remain quite slow however, and can often be the rate limiting factor governing the success of contaminant bioremediation strategies. Delivery of nutrients to microbes colonizing surfaces of low-permeability media might be diffusion controlled, whereas in high permeability media (coarse grained, fractured, or macroporous) it may primarily be advectively controlled. The rates of microbial growth and activity and propensity to alter or degrade contaminants may be quite different for the two distinct hydrologic regimes (Kieft et al., 1997; Sinclair and Ghiorse, 1989). Thus, faster flowing fracture dominated regimes will most likely be physically more appealing for sustained bioreduction as long as a suitable electron donor can be supplied. In contrast, bioreduction processes in slower flowing matrix regimes will most likely be limited by rate-dependent mass transfer of contaminants from smaller pores into larger pores. Accumulation of biomass on the surfaces of flow paths within geologic media may cause a decrease in the effective pore diameter which restricts flow and solute transport of growth promoting nutrients to organisms (Geesey et al., 1987). Another important consideration regarding bioremediation in structured and unstructured media is that the mechanisms and rates of bacteria retention are proportional to the degree of gas saturation since bacteria are preferentially sorbed to the gas–water interface versus the solid–water interface ( Jewett et al., 1999; Powelson and Mills, 1996, 1998; Schafer et al., 1998; Wan et al., 1994). Bacteria tend to accumulate at the air–water interface and thus the extent of bacterial retardation in the subsurface increases markedly with decreasing water content of the porous media. This mechanism of retention is enhanced by the corresponding loss or decrease of preferential flow and the corresponding increase in available surface area of both the solid surface and the air–water interface. The degree of sorption to the air–water interface is controlled mainly by the hydrophobicity of the cell surface, and the sorption process is essentially irreversible because of capillary forces (Wan et al., 1994).

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Depending on the water content of the subsurface media, unsaturated preferential flow may still significantly contribute to microbial bypass of the soil matrix ( Jewett et al., 1999; Powelson and Gerba, 1994; Powelson and Mills, 1998; Schafer et al., 1998). Wilson (unpublished data, University of Tennessee) found that only 6–15% of the cross-sectional area of an undisturbed block of structured coastal plain sandy sediment exhibited flow during unsaturated bacterial transport, with 88% of this flow occurring through just 4% of the area. Particle size distribution, rather than porosity, was the most significant property controlling microbial transport as areas dominated by fine sand tended to accumulate bacteria. Thus, subtle variations in particle size and arrangement (i.e., media structure) control unsaturated preferential flow paths and the degree of gas saturation which allows for the accumulation of bacteria within the subsurface. Powelson and Gerba (1994) found that virus removal by soil was three times more effective during unsaturated flow relative to saturated conditions; however, the column displacement retardations of virus transport were only 0.8–8% of that predicted by adsorption coefficients determined from batch studies. Chemical adsorption, precipitation, ion exchange, redox, complexation/ chelation, colloid formation, and microbially mediated transformation in subsurface media need to be defined in terms of hydrodynamic parameters which are often time-dependent nonlinear processes. Microbial metabolism can also alter pH, redox potential, and chemistry of the surrounding pore water causing geochemical changes (e.g., mineral precipitation). Cunningham and Fadel (2007) examined the correlation between subsurface groundwater hydraulic conductivity and the degradation rate constant for reactive contaminant transport in heterogeneous aquifers. The authors found that a negative correlation between hydraulic conductivity and the rate of contaminant degradation resulted in fingering of the contaminant plume and the persistence of more contaminant mass relative to a positive correlation. The spatial variability of the degradation rate was thought to be a function of the variability in activity of bacteria responsible for biodegradation which in turn could be the result of geochemical and mineralogical heterogeneities in an aquifer setting. Chapelle (2000) provided an overview of the significance of microbial processes on hydrological and geochemical conditions within groundwater. The author provides examples of electron donor- and electron acceptor-limited subsurface systems and the influence that microbes have on transient geochemical conditions and changes in mineral porosity, and thus groundwater flow and mass exchange.

3.3. Influence of the subsurface capillary fringe on couple hydro-bio-geochemical reactions The capillary fringe is an ill-defined boundary condition separating the water table from the unsaturated zone, without defining it as a significant part of either (Fig. 2). It is the subsurface layer in which groundwater is


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Infiltration

Unsaturated zone

Capillary fringe

Saturated zone (groundwater)

Figure 2 A schematic of the capillary fringe which is a dynamic boundary separating the water table from the unsaturated (vadose) zone. It is a subsurface layer in which groundwater is pulled from the saturated zone into the vadose zone by capillary forces. The blue color represents an aqueous phase and the shade white represents a gaseous phase (from http://oceanworld.tamu.edu/resources/oceanography-book/groundwater. html).

pulled from the saturated zone into the vadose zone by capillary forces. Pores at the base of the capillary fringe are filled with water due to tension saturation. If pore size is small and relatively uniform, it is possible that soils can be completely saturated with water for several feet above the water table. Alternately, the saturated portion will extend only a few inches above the water table when pore size is large. Capillary action supports a vadose zone above the saturated base within which water content decreases with distance above the water table. Subsurface capillary fringe regimes are an extreme example of couple processes undergoing constant dynamic changes due to recharge inputs or lack thereof, and groundwater fluctuations due to changes in surface water stage height. Because of the dynamic condition associated with most system hydrologic cycles, the capillary fringe is a temporally and spatially variable

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regime that is more often than not in a state of nonequilibrium. As such, geochemical and microbial processes are constantly changing within the capillary fringe as the influx of nutrients and oxygenated storm water impact the subsurface. Silliman et al. (2002) and Berkowitz et al. (2004) discuss the importance of the capillary fringe on local flow, chemical migration, and microbiology. They stress the impact of physical heterogeneity and the exchange of water and solutes between the capillary fringe and the region below the water table and how this alters subsurface geochemical and microbial processes. The authors suggest that physical heterogeneity with the subsurface media adjacent to the water table can lead to (1) increased flow and exchange of solutes between the capillary fringe and the underlying saturated zone, (2) preferential transport of solutes moving into the capillary fringe during infiltration events, (3) enhanced horizontal chemical flux above the water table, and (4) increased contact between gas (trapped and free flowing) and liquid phases in the region bounding the water table (Berkowitz et al., 2004; Silliman et al., 2002). Recent column studies by Qafoku et al. (2004) suggested that capillary fringe fluctuations at the DOE’s Hanford Reservation could promote the kinetically limited desorption of U into area groundwater and surface waters. Ronen et al. (2000) investigated the influence of groundwater recharge events from surface precipitation on the capillary fringe of a sandy, phreatic aquifer. During rainy seasons, abrupt changes in media water content and the increases in the height of the water table were observed. Within a 4-m interval, water table heights varied as much as 33% and 50% before and after the rainy season, respectively. Saturated conditions were detected in some regions of the capillary fringe while unsaturated conditions were found in other regions even though they were below the water table. The residence time of recharge water in the unsaturated water table regions (below the water table) was estimated to be several years. This was attributed to the entrapment of air within the pore structure of the media. Thus, multiphase flow and transport processes appear significant in the capillary fringe which will have a dramatic influence on hydrologic, geochemical, and microbial processes in the subsurface.

4. Influence of Coupled Processes on Inorganic Contaminant Fate and Transport 4.1. General overview Metals and radionuclides offer unique challenges for remediation of contaminated subsurface environments since they typically cannot be degraded into innocuous products as can organic contaminants. Because of this, inorganic contaminants are often leached deep into the subsurface where they are unreachable by conventional remedial technologies. Unlike metals


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and radionuclides, some inorganic ligand contaminants such as nitrate and perchlorate can be degraded or transformed into innocuous products via biotic and abiotic pathways. For inorganic metals and radionuclides, one option being explored is the use of microbes to transform soluble inorganic contaminants into sparingly soluble species via elemental redox changes, thereby immobilizing them in situ. Microbes can also alter inorganic contaminants in ways other than transformation, as they can alter pH, redox, and the chemical environment of subsurface systems which in turn can influence metal and radionuclide speciation and reactivity indirectly. Often these metal and radionuclide speciation changes are reversible and this is one reason why long-term stewardship and monitoring of metal and radionuclide contaminated sites is so important. Successful implementation of such a strategy requires an enhanced fundamental understanding of coupled hydrological, geochemical, and microbial processes that control contaminant migration in subsurface environments as a function of space and time. In this section, the influence of coupled processes on subsurface metal, radionuclide, and co-contaminant fate and transport are discussed. The section is divided into specific elemental inorganic contaminant types with a focus on recent field-scale relevant examples. The section ends with a discussion of recent modeling strategies that incorporate coupled processes in the simulation of the fate and transport of dissolved aqueous phase inorganic constituents in subsurface environments.

4.2. Inorganic metals 4.2.1. Arsenic The redox-sensitive toxic metal arsenic (As) is often times significantly impacted by coupled processes in subsurface environments. Sources of As are both natural and anthropogenic and it exists in the metallic state and several ionic forms (Lambert and Lane, 2004; Mansfeldt and Dohrmann, 2004; Polizzotto et al., 2005). Common inorganic species are the negatively charged arsenates (H2AsVO4– and HAsVO42) and zero-charged arsenite (H3AsIIIO30). Arsenic has been used as a medicinal agent, a pigment, a pesticide, and an agent of criminal intent. It typically accumulates in oxic sediments that contain mineral oxides of Fe and Mn since As forms strong inner sphere bonds with these mineral surfaces (Wang and Mulligan, 2006). Suboxic and anoxic environments favor the reduction of As(V) to As(III) via both geochemical and microbial pathways. Elemental arsenic is not toxic; however, most compounds of this element are extremely poisonous since very few organ systems escape its toxic effects. Arsenic in groundwater has emerged into the largest environmental health disaster of the past several decades with an estimated 100 million people worldwide at risk of exposure to unacceptable arsenic levels in drinking water (Bhattacharya et al., 2007; Ohno et al., 2007; O’Shea et al., 2007). This has become a major public

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health issue in the developing world, primarily Bangladesh and surrounding countries, where many thousands of individuals are suffering from precancerous arsenic-related disease (Fig. 3). Fortunately, several technologies are available for As removal from groundwater, ranging from simple flocculation to sophisticated ion exchange and reverse osmosis (Naidu and Bhattacharya 2006). A low cost, but effective, method for As removal in drinking water is through the use of natural Fe-rich mineral phases (i.e., Oxisols, Bauxsols, and Laterites). Polizzotto et al. (2005) investigated coupled processes responsible for the release and transport of As into aquifers of Bangladesh where nearly 57 million people drink water with As levels exceeding the limits set by the World Health Organization (WHO). The high concentrations of As are indigenous to the area and contaminated sediments wash from the mountains each year and are deposited in flood plains during the rainy season. The near-surface soil As is released to the aqueous phase through cyclic, seasonal redox cycles that impact the biogeochemistry of the subsurface (Saha and

CHINA

NEPAL

BHUTAN Brahmaputra

Ganges BANGLADESH INDIA

Dhaka Meghna Kolkata (Calcutta)

Bay of Bengal

BURMA

Areas where majority of wells contain more than 50 micrograms/liter of arsenic.

Figure 3 Schematic diagram showing the prevalence of groundwater arsenic contaminations above 50 ppb in drinking water for Bangladesh and India. The current USEPA MCL for As is 10 ppb (from http://earthtrends.wri.org/updates/node/176).


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Ali, 2006; Swartz et al., 2004). During the rainy season, subsurface conditions are ideal for microbially mediated iron and metal reduction and As is released from the solid phase. Polizzotto et al. (2005) hypothesized that Fe(III)-respiring bacteria are mobilizing both As(V) and As(III) that is bound to soil ferric oxides by the reductive dissolution of iron-arsenate minerals (Horneman et al., 2004; Islam et al., 2004; Kent and Fox, 2004; Nicholas et al., 2003). However, Kocar et al. (2006) suggests that As retention and release from Fe(III)-oxides is controlled by complex pathways of Fe biotransformation and that reductive dissolution of As-bearing ferrihydrite can promote As sequestration rather than desorption under certain environmental conditions. In the studies of Polizzotto et al. (2005), the reduction of Fe is most likely driven by microbial metabolism of sedimentary organic matter which is present in the soil at concentrations as high as 6% C (Harvey et al., 2005; Nickson et al., 2000). Arsenic released by oxidation of pyrite, due to water draw-down via irrigation and the entry of air, was considered a negligible contributor to As release into the groundwater. Voltammetric measurements in field studies have indicated that more than 95% of the dissolved As is as As(III) (van Geen et al. 2006). The release of As into the aqueous phase coupled with the fact that groundwater recharge is sufficient to continually supply As to the aquifer appears to have created a rather unfortunate situation since As retardation is limited in the aquifer due to insufficient mineralogical and geochemical conditions. Similar investigations of As mobility in Bangladesh soils by Van Geen et al. (2006) found that elevated local recharge in areas where the permeability of surface soils was high, prevented As from accumulating in groundwater. Conversely, dissolved As concentrations were found to be high in regions where local recharge was restricted by surface covers of low permeability. Nickson et al. (2005) found that in central Pakistan, a semiarid environment, canal irrigation has resulted in widespread water-logging of soils and evaporative concentrations of salts has caused As concentrations to significantly increase in groundwater. Efforts to remove As from groundwater prior to use involve filtration and in situ aeration (S.E. Fendorf, Stanford University, personal communication). In situ aeration causes an increase in groundwater dissolved oxygen (DO), which in turn causes Fe(II) to precipitate to amorphous Fe(III)oxides. The newly formed Fe(III) solid phase serves as an excellent sorbent for removing toxic levels of As from solution. Zheng et al. (2005) used hydrological and geochemical data to propose that deeper aquifers, low in As, could be used as a viable source of drinking water as long as withdrawals do not exceed recharge rates comparable to 1 cm/year. Likewise, Yu et al., (2003) suggested that replacing 30% of the existing wells in Bangladesh with deeper wells would reduce As health effects by 70% provided that As concentrations in the deep wells remained low. Efforts to construct deeper tube-wells to 60 m rather than the traditional 30 m is underway since low

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concentrations of Fe(II) and As exist in these deeper groundwater ( Jakariya et al., 2007; von Bromssen et al., 2007). However, infiltration of shallow high-As groundwater into these deeper groundwater sources is of concern due to increase pumping of the latter and shifts in the vertical subsurface hydraulic gradient (Jakariya et al., 2007; Stollenwerk et al., 2007). Hohn et al. (2006) investigated the fate and transport of As(V) in an Fereducing, sandy aquifer at the USGS site in Cape Cod, MA. An oxygenated injectate solution containing As(V) and nonreactive Br was added to the aquifer and numerous geochemical entities were measured downgradient for an extended period of time. Elevated DO in the injectate caused significant Fe(II) oxidation and subsequent adsorption of As(V) onto the freshly precipitated Fe(III)-oxides. Anoxic conditions returned to the aquifer once the injectate was terminated and an increase in As(III) was observed in downgradient monitoring wells. Sediment microbial assays and elevated hydrogen concentrations in groundwater suggested the presence of Asreducing microorganisms that were converting As(V) to As(III). Microbial reduction of As(V) coupled to oxidation of organic C or hydrogen has been shown to be important processes in some systems (Ahmann et al., 1997; Oremland et al., 2000). The investigations of Hohn et al. (2006) showed however, that even in the presence of biological reduction, both As(III) and As(V) transport were delayed relative to Br suggesting geochemical retardation of both species via precipitation and/or sorption. Arsenic is a major contaminant of acid mine drainage that typically results from historical mining activities. Acid mine drainage or acid rock drainage refers to the outflow of acidic water from abandoned metal mines or coal mines. Acid rock drainage occurs naturally within some environments as part of the weathering of sulfide-bearing rocks but is exacerbated by large-scale earth disturbances characteristic of mining and other large construction activities. This highly acidic water is caused by the biological oxidation of sulfidic materials and frequently contains high concentrations of redox-sensitive metals such as As and Fe that interact with the subsurface. The importance of microbial activity in sulfide dissolution and acid generation at mining sites has received significant attention over the years due to (1) the potential for contaminant mobilization and (2) the economic prospects of bioleaching. Biological processes in acid mine drainage are complex and are typically controlled by a variety of coupled physical and chemical processes. Edwards et al. (1999) investigated the impact of seasonal variations and various environmental conditions on microbial populations in acid mine drainage systems. They found that the relative proportions and the absolute numbers of microbial populations were spatially and seasonally correlated with geochemical (e.g., pH and conductivity) and physical conditions (e.g., temperature and rainfall). Studies by Edwards et al. (1999) showed that high concentrations of dissolved solutes occurred in the summer months and correlated with high archaeal populations and lower


15

bacterial populations. Eukaryotes were essentially absent during the winter months but increased during the rest of the year in low pH environments (pH 0.5) which correlated with decreasing water temperatures and increasing numbers of prokaryotes. Routh et al. (2007) also investigated the biogeochemical impacts on As dynamics in mining soils from Northern Sweden where soil and groundwater are heavily contaminated with As. The authors found that although oxic conditions prevailed, As-rich surface and groundwater samples contained predominately As(III). Microbially activity was believed to be responsible for the abundant proportions of reduced As (III) since the microorganism A. bolidensis was isolated from the area and it is known that this organism is capable of reducing As(V) to As(III). 4.2.2. Mercury The redox-sensitive toxic element mercury (Hg) is also significantly impacted by coupled processes in subsurface environments. Major uses of Hg in industry are historically for the production of caustic soda and chlorine as well as certain pesticides and antifouling paints. Massive quantities of Hg were also used in the 1950s and early 1960s at the Oak Ridge Tennessee Y12 Plant for the first production-scale separation of lithium isotopes (6Li) during the development of the hydrogen bomb. As of 2005, the world’s largest user of Hg is small-scale gold mining in underdeveloped countries, accounting for nearly 30% of the global Hg demand (Hogue, 2007). The world’s second largest user is China for the production of vinyl chloride (20% of the global demand). Previous and current releases of Hg to the environment have been enormous, with coal-fired electric power plants being the largest current source of human-induced Hg air emissions in the USA (40% of total emissions) (Schnoor, 2004). Atmospheric releases of Hg from coal burning are expected to become worse, since coal is cheap and abundant and has become the fuel of choice in much of the world. Coal burning is powering the economic boom in China and India, and the worldwide demand for coal is projected to rise significantly over the next decade. At the U.S. DOE Y-12 Plant in Oak Ridge, Tennessee, USA, nearly 950,000 kg elemental Hg was disseminated throughout the environment due to historical releases during the 1950s and 1960s, with the environmental implications of these releases still persisting today, some 60 years later (Burger and Campbell, 2004; Burger et al., 2005; Southworth et al., 2000, 2002). The UNEP estimates that small-scale gold mining activities account for the release of 650–1000 metric tons of Hg/year, which is about a third of all Hg releases to the environment from humans. Since Hg has no known metabolic function and it is not easily eliminated by humans or animals, it is considered extremely toxic (Eisler, 1987). Ecological and toxicological effects, however, are highly dependent on speciation (Clarkson, 2002) where Hg attacks the central nervous system, especially sensory, visual, and auditory aspects of coordination. Various forms

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of Hg (e.g., methylmercury—MeHg) can be potent neurotoxins that bioaccumulate as they track through the food chain (Akagi et al., 1995; Kurland et al., 1960; Montuori et al. 2006; Southworth et al., 2000, 2002; Ullrich et al., 2007a,b). Bioaccumulation and toxicity of Hg are strongly connected to its complex biogeochemical cycle within the environment (Fig. 4). Subsurface Hg is often highly reactive with soil and sediment and with a variety of aqueous phase ligands. In terrestrial environments, OH–, Cl–, and S2 ions have the largest influence on ligand formation with Hg where Hg(OH)2, HgCl2, HgOHþ, HgS, and Hg0 are the predominant inorganic Hg forms under oxidized conditions and HgSHþ, HgOHSH, and HgClSH are the predominant forms of Hg under reduced conditions (Barnett et al., 1995, 1997; Gabriel and Williamson, 2004). Hg forms strong inner-sphere complexes with soil and sediments, particularly those with high clay and organic matter (Liu et al., 2006; Miretzky et al., 2005; Wallschlager et al., 1998a,b), with adsorption increasing with increasing pH and decreasing with increased ligand complexation (e.g., Cl–). This is consistent with increasing evidence that Hg is primarily transported from subsurface environments to surface

H2O, O3

Hg2+

Hg+ (vapor)

Oxidation

Volatilization 55~60%

Gold mining by dredging (raft of gold miners)

Rivers in forests pH 4.7~6.0

Mercury discharged in the environment

40~45%

River Hg2+ Bottom sediment Organification

Hg (CH3)+

pH 6.0~7.1 Hg0 (metallic mercury)

Uptake by fish Retention by sediment

pH: an indicator showing acidity or alkalinity; pH7 means neutrality and smaller figures indicates higher acidity.

Figure 4 A schematic example of mercury biogeochemical cycling in terrestrial, aquatic, and atmospheric regimes (from http://www.nimd.go.jp/archives/english/ tenji/d_corner/d04.html).


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waters via particulate forms versus dissolved forms (Barringer et al., 2006; Hultberg et al., 1994; Kolka et al., 2001; Slowey et al., 2005). Both inorganic and organic colloids (Fe-oxides, clays, and DOC) make up this particulate material, all of which have a strong affinity for a variety of Hg species. Anaerobic and aerobic microbial activity via bacteria and fungi can synthesize the potent neurotoxin methyl-mercury (CH3Hgþ) (Choi and Bartha, 1993; Compeau and Bartha, 1985; Gray et al., 2004; Jackson, 1998; Regnell et al., 2001; Slowey and Brown, 2007; Watras et al., 1995; Zhang and Planas, 1994). (CH3)2Hg is sparingly soluble and highly volatile (Cotton and Wilkinson, 1988; Gavis and Ferguson, 1972), whereas CH3Hgþ is quite soluble and poses severe bioaccumulation problems, even at very low concentrations (Bakir et al., 1973; Kurland et al., 1960; Kuwabara et al., 2007; Mason et al., 1995; Southworth et al., 2000, 2002; Wiatrowski and Barkay, 2005). Southworth et al. (2000, 2002) found that the concentration of bioaccumulated CH3Hgþ in fish was more than 10,000-fold greater than its concentration in the surface water where the fish resided. Three major sources of CH3Hgþ to freshwater ecosystems have been identified by Rudd (1995) which consists of precipitation, runoff from wetlands, and in-lake/ stream methylation. The methylation, demethylation, and oxidation of Hg are typically all secondary in magnitude relative to Hg2þ reduction to Hg0 in terrestrial environments (Carpi and Lindberg, 1998). The formation of Hg0 and subsequent volatilization is an important terrestrial reaction that can regulate much of the Hg load to surface waters where bioaccumulation is a major threat. This concept has also guided several remedial strategies that take advantage of microbial reduction of Hg(II) to Hg0 in waste streams and soil (Takeuchi et al., 2001; Wagner-Dobler, 2003). Once formed, the migration of Hg0 is dependent on soil structure and soil ambient air temperature (Carpi and Lindberg, 1998; Lindberg et al., 1979; Schluter, 2000). Various strands of bacteria are known to metabolically mediate the reduction of Hg in subsurface environments (Hansen et al., 1984; Schluter, 2000; Takeuchi et al., 2001). Although Hg volatilization helps to decrease surface water Hg loads, dry deposition of Hg contributes significantly to the atmosphere/surface exchange and biogeochemical cycling of Hg (Gosar et al., 2006; Lindberg et al., 1992). Lechler et al. (1997) investigated Hg migration processes at the Carson River Superfund site in west-central Nevada, USA, where Hg contaminated soils, water, and biota exist due to historical amalgamation milling processes of Ag-Au ores. Their results suggested that Hg was preferentially leached from Hg-Au amalgam particles and subsequently adsorbed onto fine-grained sediments which were deposited downstream. In reducing environments, Hg was converted to relatively insoluble HgS where microbially mediated sulfate reduction most likely provided ample concentrations of the reduced ligand S2 to complex Hg as HgS. Fortunately, HgS is highly surface reactive which helps contribute to its lower bioavailability

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(Barnett and Turner, 2001). Bonzongo et al., (2006), however, found that naturally occurring hydrologic processes within the Carson River caused a buildup of certain anions and oxyanions which interfered with the transformation of Hg within the S cycle. The authors found that low-flow conditions were characterized by high water pH values, high concentrations of oxyanions, and decreased microbial-mediated Hg methylation in the sediments; whereas the reverse was observed during high-flow conditions. The results suggested that changing flow regimes likely affected the rates of MeHg production through a coupling of factors such as a high pH which favors MeHg demethylation, and the occurrence of high concentrations of oxyanions that can interfere with microbial sulfate reduction and MeHg production due to Hg complexation by various anionic ligands. These findings were consistent with the observations of Pettersson et al. (1995) who noted that MeHg transport was highly correlated with humic materials and that the MeHg humic/TOC ratio decreased significantly during high flow conditions suggesting rapid drainage of groundwater storage and a slow microbial production of MeHg during times of watershed depletion. Barringer and Szabo (2006) provided an overview of investigations into Hg in groundwater, soils, and seepage along the New Jersey, USA, southern coastal plain. Investigations by health departments and the USGS in the region, in response to potential human exposure risk, have shown that Hg concentrations in water from more than 600 domestic groundwater wells exceeded the maximum concentration of Hg allowable in drinking water. Through extensive observation and compilation of data, Barringer and Szabo (2006) concluded that soil disturbance caused the downward vertical migration of colloidal organic and inorganic Hg from surface soils to subsoils and that septic system effluent provided dissolved constituents that enhanced Hg mobility through the vadose zone to the saturated zone. Without disturbance, Hg infiltration would be typically limited to the upper 0.5–1.0 m of the soil profile, although deeper migration may occur if fractures or macropores are present (Henke et al., 1993). The coupled hydrological and geochemical processes controlling Hg migration along the New Jersey coastal plain were further complicated by methylation of Hg in the shallow aquifer where redox conditions, organic C, and SO4 were optimal to stimulate the activity of SO4-reducing bacteria. It is the methylated forms of Hg that pose the largest health risk due to enhanced bioavailability relative to other Hg species. Huge quantities of Hg0 were used in the 1950s and early 1960s at the Oak Ridge Tennessee Y-12 Plant to enrich 6Li during the development of the hydrogen bomb. Major releases of Hg to the environment during this period included an estimated 35,000 kg to air, 120,000 kg to floodplain and reservoir sediments, 194,000 kg to onsite soil and rock, and 590,000 kg unaccounted for and presumed lost to the environment (Southworth, ORNL, personal communication, 2007). Present day Hg losses to nearby


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surface waters are about 75 kg/year where the sources are predominately leakage from traps, junction boxes, and building footers in the historical Hguse areas at the Y-12 plant (Southworth et al., 2000, 2002). There is also evidence of significant Hg0 discharges from deep underlying karsts bedrock and near surface clay hardpans that reside under armored fine sediments at the site. This hydrologically active system maintains a strong hydraulic relationship between groundwater and surface water sources which creates significant intermingling of the two water sources and significantly impacts the off-site fate and transport of Hg. At the Oak Ridge site, Hg flux during the rainy season and during storm events appears to be dominated by resuspension of Hg-rich particulates from streambeds and inputs of dissolved Hg in the Y-12 plant storm-drain network. Barnett et al. (1995, 1997) found that the Hg sources from floodplain soils at the site were sparingly soluble mercuric sulfide and metallic Hg, and Liu et al. (2006) noted that much of this Hg was associated with organic matter. Most important, however, is that long-term studies on the Oak Ridge Reservation suggest that significant reductions in waterborne inorganic Hg inputs have not reduced microbially mediated methylmercury (MeHg) concentrations in fish. It appears that a small amount of inorganic Hg goes a long way to produce sufficient methylmercury to allow Hg bioaccumulation to persist. Current research strategies are investigating techniques that decrease in-stream formation of methylmercury without having to further eliminate inorganic Hg inputs. Strategies include (1) blocking key inorganic precursors for microbial production of MeHg via chlorination to eliminate Hg(II) transport, additions of sulfide and other complexants to bind Hg(II), the addition of chemicals to inhibit the photoreduction of Hg(II), (2) reducing net methylation via changes in microbial ecology as a result of simulation and changes in biochemistry, and (3) blocking the uptake or assimilation of MeHg from food by fish or invertebrates via food chain manipulation. Montgomery et al. (2000) present evidence that (2) above can significantly influence MeHg formation in surface waters where flooded reservoir sites were found to have higher levels of autochthonous material (algae/bacteria, i.e., potential sources/methylators of Hg) on fine particular matter relative to freshwater lakes. As well, Driscoll et al. (1995) found that high concentrations of dissolved organic C may complex MeHg, diminishing its bioavailability. Branfireun (2004) investigated the influence of coupled processes on the spatial variability of MeHg in peatlands, with a focus on microtopographical features. Since peatlands show distinctive topographical self-organization (Foster et al., 1983) where pore-water chemistries are known to have considerable vertical and horizontal spatial variability (Hunt et al., 1997), Branfireun (2004) investigated MeHg in porewater beneath several peatland microtopographical landscapes. Concentrations of MeHg were 3.5 times higher in shallow hollows versus deeper hollows which were related to biogeochemical changes associated with water table fluctuations. Branfireun

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suggests that these differences in MeHg concentration at the water table are likely due to subsurface processes that influence both microbial metabolism and inorganic Hg bioavailability in the different landforms. The spatial variability of MeHg in these systems was thought to be a complex synergy of local hydrology and accompanying groundwater–surface water interactions, plant and moss ecology, pore water geochemistry, and microbial consortia. Gray et al. (2006) investigated Hg speciation and microbial transformation of historical mine waste in southwest Texas, USA, and evaluated the propensity for Hg transport into the surrounding ecosystem. The mine waste was found to contain variable amounts of cinnabar, metacinnabar, Hg0, and Hg sorbed onto solid particulates. Stable Hg isotope analysis (see Ridley and Stetson, 2006) revealed that the net methylation rate was high indicating significant microbial Hg methylation at the site which was positively correlated with the geochemical constituents Hg2þ, organic C, and total S. Methylation of Hg was primarily a microbially mediated process that was enhanced in anaerobic, saturated environments and was favored by the highly bioavailable Hg, the presence of sulfate-reducing bacteria (SRB), and ample amounts of nutrients and organic C. Hydrologic factors limited Hg methylation at this site as the arid environment and lack of precipitation inhibited microbial activity downstream from the source. The authors noted that during periods of precipitation, the potential for Hg methylation production increased across the watershed. 4.2.3. Selenium Selenium (Se) is an essential nutrient for the health of humans and animals with recent research even suggesting that Se may reduce liver disease and prevent/cure cancer. Low Se status in humans has been associated with several chronic diseases (Li et al., 2007) such as hypertension (Mihailovic et al., 1998), coronary heart disease (Yoshizawa et al., 2003), cancer (Rayman, 2005), diabetes (Faure, 2003), and many other pathological symptoms. However, excess Se can be toxic to both humans and animals as well. Selenium from the soil is absorbed by plants which may be eaten by livestock over extensive periods resulting in chronic Se toxicity. Chronic Se toxicity in livestock is called ‘‘alkali disease’’ and is characterized by a lack of vitality, roughness of coat, loss of hair, hoof soreness, and so on. Early signs of selenium toxicity in humans include nausea, weakness, and diarrhea. With continued intake of selenium, changes in fingernails and hair loss result, and damage to the nervous system may occur. Soils with high concentrations of Se are widespread in the Rocky Mountain and Great Plains regions of the western USA and in western States with semiarid climates where irrigation is utilized for agricultural production. With regard to irrigation in these regions, excess water is typically applied to fields to flush out salts leached onto the surface soils. This excess water either


21

infiltrates into the soil or runs off into nearby basins, ponds, or streams. The irrigation water can mobilize trace elements such as Se through the soil profile, polluting groundwater and surface water sources. The high evaporation rates of semiarid environments can concentrate Se in waters to levels that are toxic to fish and sensitive bird species. The effects of selenium toxicity to fish and birds include impaired reproduction and deformed embryos. Monthly maximum discharge limits have been established for Se in irrigation drainage by the State of CA and the U.S. EPA (Green et al., 2003), and as a result, farmers and drainage districts on the western side of the San Joaquin valley are required to reduce Se concentrations in irrigation drainage discharged to the San Joaquin River. The enormous economic and health impacts posed by Se in drainage waters have prompted investigations of the biologically enhanced volatilization of Se from dewatered seleniferous sediments and what impact coupled hydrological and geochemical processes have on the rates and mechanisms of biovolatilization. Biovolatilization occurs with several metals that undergo methylation when they are taken up by plant or microbial cells. This can potentially make the metal more toxic relative to the elemental form (see Hg discussion above). Studies by de Souza et al. (2001), Frankenberger and Arshad (2001) and Frankenberger and Karlson (1994, 1995) have demonstrated that 30–70% of Se entering wetlands in central California, USA, was volatilized as dimethyl-Se (i.e., (CH3)2Se) as a result of microalgae and bacteria activity. Flury et al. (1997) investigated the potential for long-term depletion of Se from dewatered sediments by taking advantage of the concept that microbial methylation of Se to volatile (CH3)2Se may contribute to a significant loss of Se from seleniferous soils. Field experiments were initiated to investigate the likelihood that microbially mediated volatilization of Se could be used as a bioremediation approach to dissipate Se. Microbial activity within the field plots was stimulated using different organic C and protein amendments and periodic tillage and irrigation. Over a period of 100 months, Flury et al. (1997) observed that 68–88% of the Se in the upper 0–15 cm of the soil profile had dissipated. By monitoring coupled processes, Se depletion was found not to correlate with rainfall events or temperature changes. Since rainfall occurred primarily during the cooler winter months, Se leaching was primarily during this period; whereas, volatilization dominated during the summer months. The highest amount of Se depletion occurred with the amendment of protein casein; however, statistical significance was lacking with regard to nonamendment plots. The results suggested that irrigation and tillage were more important than the addition of organic C or protein amendments and thus soil structure and hydrology were key processes controlling microbial activity and therefore Se methylation and volatilization. Modeling endeavors confirmed that Se depletion from soil was kinetically controlled where the rate limiting mechanisms changed as a function of time.

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In a similar manner as Flury et al. (1997), Frankenberger and Karlson (1995) performed field investigations on biologically enhanced volatilization of Se from dewatered seleniferous sediments in central California, USA, where Se contamination in agricultural drainage waters is of significant concern. Field plots were amended with various organic materials including citrus peels, cattle manure, barley straw, and grape pomace, and several subplots were fertilized with nitrogen and zinc. Over a 22-month period, the greatest emission of gaseous Se was observed during the summer months and the lowest emissions were noted during the cooler winter months. Irrigation and tillage resulted in a 30% loss in soil Se, while plots with manure application lost nearly 60% of their indigenous Se, where cattle manure was the most effective organic amendment. Frankenberger and Karlson (1995) found that the most important parameters responsible for Se volatilization were aeration, an available C source, moisture, and warm temperatures. Thus, microbial-enhanced volatilization of Se proved to be an effective means of detoxifying contaminated sediments when soil physical, hydrological, and geochemical conditions were suitable to support microbial activity. Frankenberger and Arshad (2001) also performed laboratory and field studies to investigate microbial transformation of toxic Se species into nontoxic forms. Investigations considered microbial reduction of toxic oxyanions of Se, such as SeO42 and SeO32, into insoluble Se0 and methylation of these species into volatile (CH3)2Se. Microorganisms such as Enterobacter cloacae could be stimulated with organic amendments and were found to actively reduce Se oxyanions present in contaminated irrigation water into insoluble Se0. The authors found that the process of Se biomethylation in soil sediments and water was active and highly dependent on specific C amendments such as pectin and proteins, pH, temperature, moisture, and aeration. Further, the process of biomethylation was found to be protein/peptide limited rather than N or C limited. Additional research by Zhang and Frankenberger (2003) suggested the optimum conditions for rapid Se(VI) removal from contaminated irrigation waters were a pH of 6–9, high amounts of sulfate, low amounts of nitrate, and significant amounts of organic C amendments. Siddique et al. (2007) also found that Se-reducing bacteria were present in Se contaminated sediments that were associated with coal tailings and could reduce Se(IV) and Se(VI) to insoluble Se0. The presence of Se0 was confirmed with SEM EDX and numerous Se-reducing bacteria were isolated from the sediments. 4.2.4. Chromium In a similar manner as As, Hg, and Se, the redox-sensitive toxic metal chromium (Cr) can also be significantly impacted by coupled processes in subsurface environments. Subsurface Cr exists as both an anion and a cation depending on its oxidation state. Common inorganic species are negatively charged chromate (HCrVIO2– and CrVIO22) and the positively charged


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chromium ion, Cr(III). Whereas Cr(III) is an essential element in humans at small doses, Cr(VI) is a powerful oxidant that can quickly reduce to Cr(V) which is a known carcinogen that can lodge in living tissue to form cancerous growths. Cr(VI) has long been used as a corrosion inhibitor and in leather production creating both airborne and aqueous waste. Reduced forms of Fe (e.g., Fe(0) and Fe(II)) and synthetic and natural organics can directly reduce toxic Cr(VI) to the less mobile, less toxic Cr(III) form (Anderson et al., 1994; Deng and Stone, 1996a,b; Fendorf and Li, 1996; Ginder-Vogel et al., 2005; Jardine et al., 1999; Mayes et al., 2000; Powell et al., 1995). Direct microbial reduction of Cr(VI) is also possible if the contaminant concentration does not exceed a toxic effect on the organism (Bank et al., 2007; Cummings et al., 2007; Middleton et al., 2003; Sani et al., 2002). Indirect microbial reduction of Cr(VI) by subsurface dissimilatory bacteria is more common in mixed systems whereby biogenic Fe(II), formed from microbial-induced Fe(III)-oxide reduction, serves to reduce Cr(VI) to Cr(III) (Hansel et al., 2003; Wielinga et al., 2001; Wilkins et al., 2007). Hazen and Tabak (2005) performed field biostimulation investigations at a Cr(VI) contaminated site at the Hanford 100 H area in Richland, WA, USA, where the vadose and saturated zones were contaminated with Cr(VI) due to historical reactor operations. The Hanford 100 area resides adjacent to the pristine Columbia River, and potential migration of Cr(VI) into the river is problematic since Cr(VI) is a noted carcinogen. Conversion of mobile anionic Cr(VI) to sparingly soluble, cationic Cr(III) is highly desirable since Cr(III) precipitates as Cr(OH)3 in pH environments above 5, and the Hanford sediments have pH values near 8. Reoxidation of precipitated Cr(III) to Cr(VI) is unlikely, even in the presence of strong oxidants such as oxygen and even Mn-oxides, as Cr(OH)3 is typically scavenged and stabilized by subsurface Fe(III)-oxides (Fendorf and Zasoski, 1992; Hansel et al., 2003; Stewart et al., 2003a,b). Hazen and coworkers created bioreducing conditions in the Hanford 100 area groundwater by injection of 13C labeled poly-lactate (hydrogen releasing compound—HRC) using a dipole injection/extraction technique. Microbial (direct count and molecular analyses), geophysical (pre- and postinjection seismic and radar), and geochemical (anions, Cr, metals) analyses of groundwater, coupled with stable isotope monitoring (13C), allowed for accurate tracking of microbial processes to confirm that Cr(VI) was successfully removed from groundwater using the HRC as an electron donor and C source. The reduction of Cr(VI) occurred either directly by stimulated bacteria or, more likely, it occurred indirectly via biogenic Fe(II) which is formed by microbial reduction of subsurface Fe(III)-oxides. It is also possible that the formation of hydrogen sulfide contributed to Cr(VI) reduction since the HRC can potentially depress the redox potential of the aquifer toward sulfate-reducing conditions. These results are consistent with laboratory investigations of Tokunaga et al.

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(2001, 2003a,b) and many others that have demonstrated that substrates such as lactate can stimulate Fe-reducing bacteria which produce biogenic Fe(II) which in-turn can effectively reduce Cr(VI) to sparingly soluble Cr(III). The studies of Hazen and Tabak (2005) also used 16S rDNA microarray analysis to perform groundwater microbial community characterization and they found significantly increased microbial diversity as a result of the injected HRC which included nitrate, iron, and sulfate reducers. It is important to note that Cr(VI) reduction occurred even though Fe and sulfate TEA were not depleted from the system. This suggests that multiple electron donors can be simultaneously utilized in the subsurface, most likely due to the presence of microbial biofilms that develop during stimulation.

4.3. Inorganic radionuclides 4.3.1. Uranium Coupled processes have been shown to significantly impact the subsurface fate and transport of the redox-sensitive toxic metal/radionuclide uranium (U). The radioactivity of U found in nature is typically weak, and the chemical toxicity effects are vastly greater than the radiological effects. Uranium poisoning and toxicity are considered rare and typically limited to cases of accidental exposure by uranium miners and workers. These instances have indicated that uranium affects the proximal tubules of the kidney at very high acute doses; however, at lower doses there is generally no diminution in kidney function. Uranium has been, and continues to be used as a nuclear fuel or is converted into plutonium via ‘‘breeder’’ reactors to generate nuclear explosive material. Depleted U (238U) is a huge legacy waste problem for the DOE where massive volumes of solid and liquid waste were generated during the Cold War era. Over the past couple of decades, depleted uranium has also been used by the DoD in most of their ammunitions since the Gulf War. The U containing munitions are preferred since they are pyrophoric and self-sharpening on impact thus resulting in incredible heat and energy focused on a minimal area, for example, armor piercing ammunitions. The munitions are being used domestically (firing ranges) and overseas (war efforts) with hundreds of thousands of tons of the munitions scattered throughout the world. The most common oxidation states of uranium in the environment are U(IV) and U(VI), and their two corresponding oxides are uranium dioxide (UO2) and uranium trioxide (UO3), respectively. The UO22þ ion represents the U (VI) redox state and it is highly soluble, surface reactive, and known to form compounds with ligands such as carbonate, hydroxyl, sulfate, and organics. The U(IV) redox species is sparingly soluble and generally precipitates to form uraninite. The U.S. DOE is faced with considering options for remediating numerous sites contaminated with uranium in highly heterogeneous,


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difficult to characterize subsurface environments (NRC, 1999). The scale of this legacy waste problem is massive and includes 120 DOE sites in the USA alone and many other facilities in Europe and Russia (Lloyd and Renshaw, 2005). In this context, it is not surprising that determining the best course of action—large-scale cleanups, focused hotspot remediation, or no action (natural attenuation)—remains exceedingly difficult from a technical standpoint. Stabilization of U in situ is preferred due to the vast spatial domain of the problem (Lloyd and Renshaw, 2005). Over the past decade, research efforts have sought to use various geochemical- and microbial-based methods to convert mobile U(VI) to sparingly soluble, immobile U(IV). The concept is challenging since subsurface conditions are typically not conducive to U(VI) reduction due to the presence of co-contaminants that act as competing electron acceptors (e.g., O2, NO3–; see Finneran et al., 2002a,b; Istok et al., 2004; Wu et al., 2007) or the lack of sufficient electron donor and C sources to stimulate ample microbial activity. Vrionis et al. (2005) investigated the impact of coupled hydrological, geochemical, and microbial processes on the bioremediation of U at a historical DOE Uranium Mill Tailing Remedial Action (UMTRA) site where uranium ore was processed in the 1950s and 1960s. The UMTRA sites were U processing plants, located in the western and southwestern portions of the USA, that were closed in the 1960s and the tailing piles from mill operations were abandoned in-place. The legacy waste remains at many of the sites and efforts to immobilize U(VI) within the subsurface have involved a variety of methods including biostimulation (Anderson et al., 2003; Senko et al., 2002). Vrionis et al. (2005) utilized a multiple well injection scenario to deliver the electron donor acetate into the subsurface in an effort to stimulate microbially mediated U(VI) reduction. Both horizontal and vertical geochemical gradients were observed at the site, with more reduction of Fe(III)-oxides and sulfate occurring near the injection source and at greater depths. Downgradient from the array of injection wells, acetate utilization created Fe-reducing conditions, and as a result, an increase in abundance of 16S rRNA gene sequences belonging to the dissimilatory Fe(III)- and U(VI)-reducing family Geobacteraceae were noted (Chang et al., 2005; Vrionis et al., 2005). The highest levels of contaminant reduction were correlated with the maximal recovery of Geobacteraceae gene sequences; however, reduction zones were spatially heterogeneous due to the method of acetate injection or heterogeneities in the groundwater hydrologic flow-field. Laboratory studies of Finneran et al. (2002a,b) using UMTRA sediments from the site, also indicated that the addition of acetate to the sediments stimulated anaerobic conditions and the loss of U(VI) from solution. The reduction of U(VI) occurred simultaneously with the formation of Fe(II) and prior to sulfate reduction. 16S rDNA analyses of the simulated microorganisms revealed that U reduction occurred as the microbial communities shifted toward organisms known to

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reduce both Fe(III) and U(VI), such as Geobacteraceae which were greatly enriched (40% of total detectable bacterial community) (Holmes et al., 2002). Davis et al. (2006) also investigated natural in situ processes affecting the transport of U(VI) at another UMTRA site under nonbiostimulated conditions. The researchers noted that the upgradient portion of the contaminated aquifer had very little dissolved Fe(II), few metal-reducing bacteria, and the U(VI) that was present in solution was controlled by the U(VI) adsorption to the solid phase (vs dissolution of a U precipitate). However, in the downgradient portion of the aquifer where redox conditions were more anoxic, Fe(II) concentrations increased, diverse populations of Fe(III)reducing bacteria were observed, and significant reduced U(IV) was detected on the solid phase; all indicative of microbially mediated U(VI) reduction and immobilization to U(IV). Schryver et al. (2006) investigated the relationship between hydrologically impacted groundwater geochemistry and microbial community structure at the Shiprock uranium mill tailing disposal site in New Mexico, USA. The authors applied both nonlinear and generalized linear data analysis methods to relate microbial biomarkers, such as phospholipids fatty acid (PLFA), to groundwater geochemical characteristics where the primary contaminants and solutes of concern were U(VI), SO42, and NO3–. Neural network models were found to greatly outperform the generalized linear models for describing the data. Modeling results suggested that riverine influences (i.e., nearby river impact on groundwater hydrology and geochemistry) and U(VI) distribution were important in predicting the distribution of the microbially based PLFA classes. Nonlinear principal components were then extracted from the PLFA data using a variant of the feed-forward neural network technique which grouped samples according to similar geochemistry. The PLFA indicators of Gram-negative bacteria and eukaryotes were associated with groundwater with lower concentrations of contaminants. Groundwaters with significantly higher concentrations of contaminants were associated with terminally branched saturated and branched monounsaturates that are indicative of microbial metal reducers, actinomycetes, and Gram-positive bacteria. These findings indicate that microbial community composition at this U contaminated site is strongly coupled to the groundwater geochemistry (i.e., also observed by Palumbo et al., 2004) which is spatially and temporally altered by surface and subsurface hydrologic processes. Groudev et al. (2001a,b) performed laboratory and field investigations to evaluate the propensity for bioremediation of agricultural soils contaminated with radionuclides (U, Ra, and Th) and toxic metals (Cu, Cd, and Pb) in southeastern Bulgaria that have resulted from previous mineral processing and mining activities. Stimulation of heterotrophic and chemolithotrophic aerobic microbes in the near surface resulted in dissolution of contaminants with subsequent hydrologic transport to lower horizons where contaminants


27

became immobilized as sparingly soluble compounds primarily as a result of anaerobic SRB (Gadd, 2004; Geets et al., 2005). The activity of these organisms was enhanced by perturbations in various hydrological and geochemical environmental factors including water, oxygen, and nutrient content. Extensive U bioremediation research has been underway at the former S-3 ponds located within the Y-12 facility in Oak Ridge, Tennessee in the eastern USA, where massive unlined surface impoundments were used to dispose of acidic, highly buffered U, Tc, and NO3 bearing waste during a period from the early 1950s to the early 1980s. The liquid waste was pipelined to the ponds at a rate of 10 million liters/year for 32 years, and during this period infiltration was the primary release mechanism to the surrounding soils and groundwater. In 1984, attempts were made to neutralize and bio-denitrify the S-3 ponds and they were capped in 1988. Subsurface contaminant fate and transport processes are complex at the site since multiregion flow and transport mechanisms are the norm due to fractured weathered saprolite derived from interbedded shale and limestone sequences. The media consist of highly permeability fractures that surround low permeability, high porosity matrix blocks on the centimeter scale, and thus the media is not only conducive to significant preferential flow, it is also a source/sink for contaminants (Fig. 1A). Contaminants, such as U, Tc, NO3, PCE, and toxic metals (e.g., Al, Ni, and Hg) migrate away from the capped waste disposal units following both geologic strike and bedding plane dip, with density effects also being quite significant over great distances from the source. Near source groundwater concentrations of NO3 can be as high as 40,000 ppm and U concentrations can be as high as 60 ppm, with solid phase concentrations near or above 1000 mg U/kg and in some places above 12,000 mg U/kg solid. Elevated nitrate concentrations and significant U have been detected vertically to several hundred feet owing to rapid movement through the saprolite and underlying bedrock. The nitrate plume extends nearly 100 ha down the valley in relation to the source. Since the year 2001, research activities at this site have focused on plot-scale biostimulation studies in an effort to induce in situ bioreduction and immobilization of subsurface U and Tc contamination. The investigations have shown that microorganisms indigenous to the subsurface environments can be stimulated to transform contaminants, such as U and Tc, into chemical species that are less mobile in groundwater. These studies have also tested novel geophysical, hydraulic, and tracer techniques for characterizing and monitoring subsurface coupled processes and groundwater flow. For example, they have tested new inexpensive surface geophysical techniques in which seismic waves and electrical currents are used to create three-dimensional (3D) images of the subsurface geology and of contaminated groundwater plumes. One investigation at the Oak Ridge Y-12 site combined subsurface transport, microbiology, and geochemistry to identify the conditions that

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are conducive to the bioremediation and immobilization of U and Tc (Istok et al., 2004). The investigations focused on the potential to stimulate indigenous microbial communities that could reduce a mixture of mobile U(VI) and Tc(VII) in the presence of elevated initial nitrate (120 mM) cocontamination in the shallow unconfined aquifer. Microbial reduction of these contaminants is desirable since the reduced forms of these radionuclides (e.g., U(IV) and Tc(IV)) are sparingly soluble and significantly less mobile than their oxidized species. The investigations of Istok et al. (2004) utilized small-scale field ‘‘push-pull’’ tests where electron donor such as ethanol, glucose, and acetate were injected radially into the aquifer, and then slowly removed as a function of time. The authors found that when electron donor was added, rapid nitrate utilization via denitrification was observed with nearly simultaneous reduction of Tc(VII). Once Fe-reducing conditions were achieved in the subsurface, U(VI) reduction began to occur (Fig. 5). Changes in viable biomass, community composition, metabolic status, and respiratory state of organisms sampled from down-well microbial samplers during these tests were consistent with enhanced microbial growth, creation of anaerobic conditions, and an increase in the abundance of metal-reducing organisms (e.g., Geobacter and Anaeromyxobacter) (North Fe(II) NO−2 pH

500 250 0

9 8 7

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Figure 5 Loss of uranium (U) and technetium (Tc) from groundwater following the injection of ethanol during a field ‘‘push-pull’’ biostimulation experiment at the Oak Ridge Y-12 S-3 ponds site. Note the formation of Fe(II) in solution which results from the reduction of Fe(III)-oxides which acts as a competing terminal electron acceptor (see Istok et al., 2004).


29

et al., 2004; Peacock et al., 2004; Wilkins et al., 2006). These results were further supported by the observations of Petrie et al. (2003) who used phylogenetic analysis of 16S rRNA gene sequences extracted from MPN dilutions to show that the predominant members of Fe(III)-reducing consortia from background sediments were closely related to members of the Geobacteraceae family, while the Fe(III)-reducing bacterium Anaeromyxobacter sp., Paenibacillus sp., and Brevibacillus sp. predominated in the Fe(III)-reducing consortia of the contaminated sediments. Analysis of core samples taken before and after biostimulation using variable-temperature Fe-57 Mossbauer spectroscopy revealed an overall loss of Fe from the system and major changes to the distribution of the Fe-oxide mineral forms relative to prebiostimulated conditions (Stucki et al., 2007). Within biostimulated cores, goethite spectral components were greatly diminished in intensity whereas hematite spectral components were greatly enhanced suggesting preferential loss of goethite from biostimulated samples. This was most likely due to microbially induced reduction of Fe(III) within the goethite minerals to soluble Fe(II) moieties. This is supported by additional data of Stucki et al. (2007) that showed that the Fe(II):Fe(III) ratio in the nonoxide phase (aluminosilicate clay minerals) increased during the biostimulation process. The biogenic Fe(II) that was formed could also have contributed to the reduction of U(VI) and Tc(VII) observed in these systems. In another research effort at the Y-12 site in Oak Ridge, Wu et al. (2006a,b) investigated the rates and mechanisms by which naturally occurring microorganisms transformed solution and solid phase U(VI) to U(IV) in the presence of dynamic flow conditions and complex geochemical reactions. They used a double-dipole, forced gradient injection-extraction strategy where tracers and electron donor (i.e., ethanol) were intermittently injected into the inner loop recirculation zone (inner dipole), and clean water was injected into the outer loop recirculation well (outer dipole) which was designed to protect experimental reactions within the inner loop. Geophysical measurements, including multielectrode resistivity and tomographic seismic refraction, were used to guide monitoring well placement within the inner loop, and to confirm the location of subsurface regimes that were most hydrologically active and most highly contaminated (Chen et al., 2006; Watson et al., 2005). Since the research effort was focused near the source where nitrate concentrations were in the thousands of parts per million, the overall strategy combined an aboveground removal of PCE, NO3, high concentrations of Al, Ca, and Mg (Wu et al., 2006a), and a belowground biostimulated reduction zone for immobilization of solution and solid phase U(VI) (Wu et al., 2006b). The addition of ethanol initially stimulated denitrification of solid-phase matrix ‘‘entrapped’’ NO3 which was subsequently followed by U(VI) reduction as sulfate-reducing conditions were invoked (Fig. 6). Continued additions of electron donor allowed for sustained U reduction over a 13-month investigation, and

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X-ray Absorption Near Edge Structure (XANES) on sediment samples acquired after biostimulation confirmed the formation of significant solidphase U(IV) that was not present prior to biostimulation (Kelly et al., 2008). Wu et al. (2006b) maintained the system (1) at a pH below 6.2, (2) at low bicarbonate levels, and (3) with residual sulfate to suppress methanogenesis and minimize U remobilization. The research was further expanded to consider the application of functional gene arrays (FGAs) to the analysis of the in situ U bioreduction processes (He et al., 2007). The array, known as the GeoChip, is the most comprehensive FGA available for environmental studies and allows for the investigation of microbial community gene functionality and processes in groundwater and soil contaminated with metals, ligands, and organics with excellent resolution. Analysis of groundwater via the FGA before and during subsurface biostimulation showed statistically significant positive hybridization signals with dissimilatory Fe (III)-reducing bacteria (FeRB) such as Geobacter spp. and SRB, such as Desulfovibria spp., which reached their highest levels during the biostimulation period when U(VI) reduction was observed. These results are consistent with classical microbial monitoring methods that have shown these two groups of microorganisms are capable of U(VI) reduction via direct enzymatic or indirect chemical mechanisms (Liger et al., 1999; Lovley et al., After biostimulation 2 pH

pH

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Before biostimulation 2

Figure 6 Groundwater pH, nitrate, and uranium concentration profiles within a field facility at the Y-12 S-3 ponds site before and after biostimulation using ethanol as an electron donor. The pictorial insets show sediment samples acquired before and after biostimulation where sediments on the left are unaltered and samples on the right have been reduced. The black color may be indicative of the U(IV) species uraninite.


31

1993a,b; Petrie et al., 2003; Tebo and Obraztsova, 1998; Truex et al., 1997; Wu et al., 2006b). Microarray analysis indicated that in situ U reduction activity correlated with the abundance of multiheme C-type cytochome genes (R ¼ 0.6), dissimilatory sulfite reductase genes (dsrAB) (R ¼ 0.8) similar to those from Desulfovibrio-like and Geobacter-like species. These results are also consistent with findings from 16S rRNA gene-based library studies which are very labor intensive relative to the FGAs. Fields et al. (2006) investigated changes in microbial community structure along this particular contaminant plume (e.g., NO3, Tc, and U) by monitoring shifts in microbial phylogenetics and functionality and changes in geochemistry. Clonal libraries of multiple genes were analyzed from groundwater that varied in contaminant concentration along the plume and compared this information to over 100 geochemical parameters using principal components analyses. The analyses-suggested sites could be grouped as low, intermediate, and extreme contaminant levels where the low and extreme sites were functionally less diverse than sites with intermediate contaminant concentrations. The ‘‘extreme’’ sites were characterized by not only high contaminant concentrations, but highly buffered acidity; whereas the ‘‘low’’ sites were characterized by low ionic strength conditions and limited nutrients. Both these conditions were thought to contribute to the observation of similar functionality even though they were phylogenetically distinct. Since U transformations can be influenced by both biological and chemical transformation reactions, Christensen et al. (2004) used a stable isotope technique to distinguish sources and pathways. They used ratios of U isotopes to implicate leaking waste storage tanks at the Hanford site in Richland, WA, USA, and their contribution to vadose zone pore water and groundwater. The authors showed that both stable and slowly decaying radioactive isotopes could be used as signatures for source identification and the transformation of metals and radionuclides contaminants in heterogeneous subsurface environments. Brooks et al. (2003) recently showed the pronounced influence of Ca2þ on the bioreduction of U(VI) at circum-neutral pH values. The authors provided evidence for the formation of a Ca–UO2–CO3 complex that is resistant to microbial reduction via metal-reducing bacteria since it is less effective than uncomplexed U(VI) at being a terminal electron acceptor. However, Stewart et al. (2007) has shown that the presence of Fe(III)-oxides strongly influence the complexation reaction between Ca2þ and U(VI) due to the significance of Ca2þ adsorption (resulting in less Ca–UO2–CO3) and the presence of a competing terminal electron acceptor. The authors found that ferrihydrite acts as a competitive electron acceptor and thus, like Ca, decreases U(VI) reduction. However, with increasing Ca2þ concentrations, U(VI) reduction was enhanced in the presence of ferrihydrite (relative to its absence) and U(VI) reduction becomes almost independent of Ca concentration. Several other studies have documented the inhibition of U(VI)

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Philip M. Jardine

reduction and/or its reoxidation from reduced U(IV) in the presence of nitrate and NOx (Elias et al., 2003; Finneran et al., 2002b; Istok et al., 2004; Senko et al., 2002; Wu et al., unpublished, ORNL) and dissolved O2 (Wu et al., 2007). Gu et al. (2005) also found that although the presence of humics can accelerate the reduction of U(VI), it can also accelerate the reoxidation process from U(IV) to U(VI) under certain circumstances. 4.3.2. Technetium Technetium is a radioactive chemical element with no stable isotopic forms. Most technetium produced on Earth is a fission by-product of 235U in nuclear reactors and is extracted from nuclear fuel rods. Its short-lived gamma-emitting nuclear isomer 99 mTc (half life, t1/2 ¼ 6 h) is used in nuclear medicine for a variety of diagnostic tests. The 99 mTc isomer decays to 99Tc which has a t1/2 ¼ 212,000 years and is used as a source of beta particles. Although it has a low chemical toxicity, its radioactive toxicity can be potentially harmful. The two most common redox states in the environment are pertechnetate, Tc(VII)O4, under oxic conditions and Tc(IV) under anoxic conditions. Tc(VII) is a contaminant of concern at a number of U.S. DOE facilities, including sites at Oak Ridge, TN; Paducah, KY; Hanford, WA; and Portsmouth, OH, due to its large migration tendency in groundwater. In oxygenated and suboxygenated environments, Tc(VII) is highly soluble, poorly sorbed by sediment minerals, and is therefore highly mobile in the subsurface (Bondietti and Francis, 1979; Gu and Schulz, 1991; Schulte and Scoppa, 1987; Wildung et al., 1986). In reducing environments, Tc(VII) is readily reduced, either chemically or biologically, to Tc(IV) or Tc(V) species, which have a much lower solubility and thus are retained by sediments and soil humic materials (Bondietti and Francis, 1979; Gu and Schulz, 1991; Lloyd et al., 1998, 2000; Schulte and Scoppa, 1987; Wildung et al., 1986, 2000). A variety of anaerobic microorganisms have been shown to be capable of reducing Tc(VII)O4– in solution to solid-phase Tc(IV) precipitates in the presence of various electron donors (Lloyd et al., 1998, 2000; Wildung et al., 2000). Therefore, bioreduction of Tc(VII) has been proposed as an option to remove or impede the migration of technetium in the subsurface. On the other hand, it is also known that reduced Tc(IV) species can readily form complexes with a number of organic and inorganic ligands such as carbonate, citrate, EDTA, and DOC under reducing conditions and thus render it soluble and mobile in groundwater (Geraedts et al., 2002; Gu and Ruan, 2007; Maes et al., 2004). Istok et al. (2004) investigated the propensity for bioremediation of Tc (VII) at the Oak Ridge Tennessee USA Y-12 site (described above in Section 4.3.1). Investigations focused on the stimulation of indigenous microbial communities to reduce a mixture of mobile U(VI) and Tc(VII) in the presence of elevated initial nitrate (120 mM) co-contamination in the shallow unconfined aquifer. Microbial reduction of Tc(VII) to Tc(IV) is


33

desirable since the reduced form of this radionuclide is sparingly soluble and significantly less mobile than the oxidized species. The investigations utilized small-scale field ‘‘push-pull’’ tests where electron donor was injected radially into the aquifer, followed by slow removal with time. With the addition of electron donor, rapid nitrate utilization via denitrification was observed with a nearly simultaneous loss of Tc(VII) from solution. These results suggest direct microbial enzymatic reduction of Tc(VII), although Mossbauer spectroscopy analyses of postbiostimulated core sample suggested the formation of biogenic Fe(II) (Stucki et al., 2007) which may have contributed to the reduction of Tc(VII). Wildung et al. (2004) observed presumptive evidence of such a process while investigating Tc reduction in Atlantic Coastal Plain sediments from a shallow sandy aquifer that exhibited a Fe(II)/Tc(VII) concentration of ethylene>VCTCE with TCE concentrations being 10-fold higher in the waste trenches relative to downgradient sampling wells, whereas VC, 1,2-DCE, and ethylene concentrations in the waste trenches were similar or slightly higher than the downgradient groundwater monitoring wells. TCE concentrations disappeared within 10 m from the end of the waste trenches and 1,2-DCE disappeared just prior to the seep, 50 m downgradient the trench source. VC and ethylene were still present at the seep, with ethylene showing peak concentrations at this locale. These results indicated that anaerobic reduction of the chlorinated organics was occurring. In addition, the presence of high concentrations of methane throughout the site was also an indication of anaerobic metabolism (Lenczewski et al., 2003). The accumulation of VC in many locations was consistent with the notion that conversion of VC to ethane is usually the rate limiting step during reductive dechlorination of chlorinated solvents (Ballapragada et al., 1997). Another interesting finding at the site was that the concentration of chlorinated organics was slightly higher in the matrix relative to the fracture regime, whereas the concentration of ethylene was lower in the matrix relative to the fracture regime. These results suggest that the more hydrologically active fracture regime was slightly more effective in the anaerobic reduction of the chlorinated organics. Temporal variability of TCE and its degradation products was slight, with a general increasing concentration trend of chlorinated organics and dissolved gases as the site hydraulic gradient increased (a response to increased storm events during the winter and spring months). This may imply that the intrinsic bioremediation scenario at the site was less effective at higher discharge rates. Measurements

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of redox potential at the site indicated that iron-reduction, sulfate reduction, and potentially methanogenesis were occurring and are conducive to dechlorination of TCE. Bacterial enrichments of groundwater samples revealed the presence of methanotrophs, methanogens, iron-reducing bacteria, and SRB, all of which have previously been implicated in anaerobic biodegradation of TCE. 16S ribosomal DNA (rDNA) sequences from DNA extracted from groundwater were similar to sequences of organisms previously implicated in the anaerobic biodegradation of chlorinated solvents. The combined data strongly suggest that anaerobic dechlorination of TCE to VC and ethylene was occurring (Lenczewski et al., 2003). The presence of methane oxidizers (methanotrophs) also suggested possible zones of oxygenated groundwater, which was confirmed with on-site DO measurements. Groundwater DO was found to increase during the winter and spring months when the site hydraulic gradient increased due to oxygenated recharge from atmospheric precipitation during this period. Thus, a second biological removal mechanism of chlorinated organic compounds may have been occurring at the site, which involved the oxidation of TCE to ethane via methane oxidizing bacteria. Since both ethane and VC-ethylene are all present as possible degradation products of the TCE, it is probable that both mechanisms were operative. Skubal et al. (1999) investigated temporal changes in redox zonation at a mixed hydrocarbon/solvent contaminated aquifer in an effort to quantify the propensity for TCE natural attenuation in situ. Predominant TEAPs as measured by dissolved hydrogen, suggested temporal variations in reoxygenation along the plume transect. It was postulated that the intrusion of oxygen was possibly due to recharge, fluctuations in the water table, and/or microbial activity. Microbial analyses revealed a correlation between bacterial phylogeny, TEAP, and groundwater hydrogen concentrations. An increase in the water table and evidence of methanogenesis corresponded to an order of magnitude increase in archaeal 16S rRNA relative to when it was unsaturated (creation of capillary fringe). Spatial and temporal variations in TEAPs and microbial community structure suggest that the potential for TCE dechlorination varies seasonally within the plume, with reductive reactions (formation of DCE and VC) more likely in the shallow saturated zone or the capillary fringe during wet cycles, and aerobic co-metabolism of TCE and its products more likely in the deep aerobic subsurface or vadose zone where it could be supported by organic co-contaminants or methane from methanogenesis. Nevertheless, natural dechlorination processes at the site were limited despite the abundance of electron donor and C sources. Significant vertical and horizontal variations in TEAP zonation and associated organic contaminant degradation processes have been noted in a variety of settings (Christensen et al., 1994; Norris et al., 1994). At a sewage-effluent plume in Cape Cod, MA, the vertical dispersivity was found to be insignificant relative to the longitudinal dispersivity


55

(Garabedian et al., 1991; LeBlanc et al., 1991). This circumstance limited vertical mixing and allowed for a sharp gradient in oxygen and other solutes within the subsurface over long time periods (Smith et al., 1991) thus influencing contaminant fate and transport processes in the different zones. McGuire et al. (1999) also observed significant temporal variations in TEAP zonation within a mixed fuel/chlorinated solvent plume. Simultaneous changes in microbial community structure were noted during the same time period as the TEAP shifts, with methanogen abundance being highest where methanogenesis and sulfate reduction were the predominant TEAP (Haack and Reynolds, 1999; Reynolds and Haack, 1999). Song et al. (2002) used a time-series stable C isotope technique to monitor the degradation of groundwater TCE and its by-products at the Idaho National Laboratory in western USA following subsurface biostimulation using lactate. Large kinetic isotope effects were observed during the dechlorination process which indicated a microbially mediated reaction scenario. The observed changes in the 13C/12C ratio indicated microbialmediated biodegradation of the VOCs, since a constant isotopic ratio would be more indicative of geochemical and hydrological impacts on VOC concentration changes. Morrill et al. (2005) also observed substantial C isotope enrichment in c-DCE at a DoD contaminated site in San Antonio, TX, and the authors were able to calculate the rate of VOC transformation using the stable C isotope technique. Similarly, Chu et al. (2004) suggested that VOC compound-specific isotopic fractionation could assist in determining whether aerobic or anaerobic degradation of VC and c-DCE occur during in situ reductive dechlorination; however, metabolic versus cometabolic reactions could not be distinguished since their isotopic fractionation changes were too similar. Lee et al. (2007) studied stable C isotope fractionation of chloroethenes by dehalorespiring isolates and found that a wide range of isotopic enrichment factors were associated with functionally similar and phylogenetically diverse organisms. Because of this, the authors cautioned that although compound-specific isotope fractionation is a powerful tool for evaluating the progress of in situ bioremediation in the field, care must be exercised when applying enrichment factors for the interpretation of dechlorination results. Numerous investigations have shown that fluctuations in recharge to shallow contaminant plumes can create temporal variability in geochemical conditions that are reflected in microbial population changes. Recharge events that deliver electron acceptors such as O2, NO3, SO4, and Fe(III) to anaerobic, contaminated subsurface environments are likely important considerations for assessing the propensity for organic contaminant natural attenuation (Vroblesky and Chapelle, 1994). McGuire et al. (2005) noted recharge-induced geochemical changes in a sandy aquifer contaminated with waste fuels and chlorinated solvents. Multiple regression analysis indicated that dominant chemical associations and their interpreted

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processes (anaerobic and aerobic microbial processes, mineral precipitation and dissolution, and temperature effects) did not change significantly during spring time recharge events; however, the relative importance of each of the processes within the contaminated plume did in fact change. The authors found that after recharge events, the overall importance of aerobic processes increased and during anaerobic periods the zones with multiple electron accepting processes (TEAPs) likely occurred in the same aquifer unit. It was determined that recharge effects on TEAPs occurred primarily at the interface between infiltrating recharge water and the aquifer (capillary fringe) where rising water table elevations may have enhanced the availability of Fe(III)-oxide coatings as electron acceptors for metal-reducing microorganisms. This hypothesis is supported by Lovley and Anderson (2000) who demonstrated that Fe(III)-reducing bacteria can be effective agents in removing aromatic hydrocarbons from groundwater under anaerobic conditions. The investigations of McGuire et al. (2005) make clear that an understanding of site geology, hydrology, and hydrogeochemistry are required to avoid misinterpreting TEAP zonation and its impact on organic contaminant biodegradation. A good example is Yager et al. (1997) who delineated the hydrogeochemical setting of a fractured dolomite aquifer that was contaminated with chlorinated ethenes. Methane and sulfide analysis in groundwater wells suggested methanogenesis or sulfate reduction was a possible TEAP; however, both degradation pathways were discounted using hydrogen gas analyses and recognizing that the source of methane and sulfide was from deeper noncontaminated portions of the aquifer. Although the transient nature of water levels, flow directions, availability of terminal electron acceptors, and contaminant concentrations within chlorinated solvent plumes has been recognized (Christensen et al., 2000; NRC, 2000), temporal variations in aquifer microbial community structure, and the factors that might govern temporal variations, are not well studied in contaminated groundwater. Haack et al. (2004) investigated spatial and temporal changes in microbial community structure associated with recharge-induced chemical gradients in a contaminated aquifer containing waste fuels and chlorinated solvents (McGuire et al., 2002, 2005). Community amplified ribosomal DNA restriction analysis (ARDRA) using 16S rDNA primers and denaturing gradient gel electrophoresis (DGGE) using 16S rDNA primers indicated that (1) communities in the middle of the aquifer where anoxic/contaminated conditions occurred were similar regardless of recharge, (2) communities at the greatest aquifer depths were similar to those in uncontaminated environments after extended recharge, and (3) communities changed in the upper and lower depths of the aquifer during extended periods of low recharge. General aquifer geochemistry was found to be quite important as was TEAPs with regard to the spatial and temporal variability of microbial communities within the aquifer (Haack et al., 2004). Numerous other investigators have also observed significant


57

within-plume microbial heterogeneities at contaminated sites containing chlorinated solvents (Bekins et al., 2001; Christensen et al., 2000; Cozzarelli et al., 2000; Davis et al., 2002; Dojka et al., 1998; Haack and Bekins, 2000; Kinner et al., 2002; Madsen, 2000; Pickup et al., 2001). The propensity for microbial-mediated dechlorination of contaminant organic solvents in the presence of bioavailable groundwater sulfate has resulted in a variety of conflicting findings. Investigations of subsurface hydrology-induced geochemical impacts on the degradation of chlorinated solvents indicated that the presence of sulfate influenced dechlorination via no inhibition (El Mamouni et al., 2002; Hoelen and Reinhard, 2004), partial inhibition (Cabirol et al., 1998; Townsend and Suflita, 1997), or complete inhibition (Nelson et al., 2002; Warner et al., 2002). The lack of dechlorination during sulfate-reducing conditions may result from several factors including (1) competition for electron donor by SRB, (2) dechlorination enzyme inhibition by sulfate, (3) use of the sulfate as a terminal electron acceptor versus the CAHs, and (4) scenarios conducive to larger populations of SRB relative to dechlorinators and thus reaction kinetics toward sulfate reduction versus dechlorination. Similar situations may also apply to the competition for H2 in the presence of nitrate and Fe(III)-reducing conditions (Aulenta et al., 2006). Dybas et al. (1998, 2002) conducted plot- and field-scale bioaugmentation experiments designed to demonstrate the remediation of nitrate and carbon tetrachloride in an aquifer at Schoolcraft, MI. Pseudomonas stutzeri (strain KC) was utilized in the endeavor since it is a denitrifying bacterium that degrades carbon tetrachloride (CT) to CO2 and other inert compounds (Criddle et al., 1990) without producing CF. Subsurface activity of the organisms was maintained by adjusting the pH of the groundwater to more alkaline conditions and using pulsed additions of acetate to the groundwater followed by additions of acetate-free water. Significant losses of both nitrate and CT were observed. Uniform efficiencies of nitrate and CT removal over a 15-m vertical depth profile were observed despite significant variability in groundwater hydraulic conductivity.

5.3. Hydrocarbons 5.3.1. Crude Crude oil, also known as petroleum, is a naturally occurring NAPL within geologic Earth deposits and consists primarily of a complex mixture of alkane hydrocarbons of various lengths ranging from approximately C5H12 to C18H38. The largest quantities of petroleum are used primarily for producing fuel oil and gasoline with nearly 85% of the hydrocarbons present in petroleum being converted into energy-rich fuels, including gasoline, diesel, jet, heating, and other fuel oils and liquefied petroleum gas. Petroleum is also used in the production of many pharmaceuticals

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products, solvents, fertilizers, pesticides, and plastics. With the notion of the world’s oil supplies beginning a downward trend, coupled with current-day stress factors concerning the world’s clean water supply, the presence of oil has significant social and environmental impacts ranging from routine activities such as seismic exploration and drilling to accidents and polluting wastes. Oil spills on land and water bodies are mostly caused by accidents involving tankers, barges, pipelines, refineries, and storage facilities and are caused by human error or carelessness, and sometimes by natural disasters such as hurricanes or earthquakes. Cozzarelli et al. (1999) investigated scale-effects on biogeochemical reactions in a physically and chemically heterogeneous aquifer that was contaminated with gasoline. The aquifer was composed of two hydrologic units with a shallow local aquifer of perched water and an underlying regional sandy aquifer. Vertical heterogeneity was pronounced as reactive groundwater species concentrations varied by an order of magnitude, which significantly impacted the propensity for gasoline biodegradation. Microbially mediated degradation of hydrocarbons was noted to vary over short vertical distances and time, and Cozzarelli et al. (1999) found that anaerobic processes dominated within the low-permeability clay units with nitrate reduction and aerobic hydrocarbon degradation occurring to a greater extent in the more permeable sandy layers where the availability of electron acceptors was plentiful. Because of the limited availability of electron acceptor in the low-permeability layers, hydrocarbon degradation was limited relative to the more permeable sand layers. Degradation processes were still evident in the lower hydrologically conductive clay layer and were linked to the presence of sulfate and iron reduction within this unit. The authors noted that the chemical effects resulting from the microbial degradation of the hydrocarbons led to discrete zones where secondary minerals, such as iron-sulfide may precipitate from solution. Vertical heterogeneity at the site was such that small-scale geochemical changes had to be quantified in order to evaluate changes in biogeochemical processes with depth, and the impact of hydrologic processes was different for the perched water regime versus the regional aquifer due to different hydrodynamics of the two zones. Cozzarelli et al. (1999) found that recharge water entering the perched water was depleted in oxygen and nitrate as it mixed with contaminated groundwater of the shallow, higher permeability zone. Thus, recharge events can be a significant driver of groundwater redox changes, especially in organic contaminated aquifer that are often anaerobic (McGuire et al., 2000; Vroblesky and Chapelle, 1994). Hydrocarbons that concentrate near the capillary fringe serve as abundant electron donors for microbial respiration; however, microbial activity is often limited by the availability of electron acceptors. In reduced environments, recharge events can initiate changes in TEAPs by providing an influx of electron donors such as oxygen, nitrate, and sulfate (Vroblesky and Chapelle,


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1994). Microbial terminal electron processes are often examined for natural and enhanced biodegradation of contaminants (Baedecker et al., 1993; Bjerg et al., 1995; Chapelle et al., 1996; Levine et al., 1997; Lyngkilde and Christensen, 1992; Vroblesky et al., 1997); however, it is often difficult to distinguish between the direct influence of recharge on TEAPs versus dilution of electron donors (organic contaminants and dissolved hydrogen gas) when recharge mixes with indigenous groundwater. As capillary fringe zones are frequently aerobic due to the intrusion of oxygen from recent recharge, many of the aromatic compound degrading genotypes present in groundwater (Hosein et al., 1997; Stapleton and Sayler, 1998; van der Meer et al., 1998) are linked to oxidative processes requiring molecular oxygen and, therefore, are only active in oxygenated areas such as the plume fringe (Davison and Lerner, 2000). At a crude-oil spill site near Bemidji, MN, Essaid et al. (1995) noted variations in hydrologic flow paths due to restricted recharge through nonaqueous oil bodies. Flow path variations resulted in large changes in the overall depth of the anaerobic portion of the plume, where ferric iron reduction and methanogenesis were the dominant TEAPs. Bekins et al. (1999) found that in regimes where methanogenic conditions were prevalent at the site, a shift in the number and types of solid phase culturable organisms was also present. Methanogenic intervals were noted to have an increased number of methanogens and heterotrophic fermenters and fewer iron reducers. Haack et al. (2004) found that a particular type of microbial community within various locations of the Bemidji aquifer was equally influenced by aquifer geochemistry and the dominating terminal electron acceptor that were present. At the same study site in Bemidji, MN, Bekins et al. (2001) investigated controls of coupled processes on the spatial distribution of subsurface microbes and its impact on the propensity of natural attenuation at a crude oil spill site. Microbial populations were analyzed along with aquifer permeability, pore-water chemistry, nonaqueous phase oil content, and extractable sediment Fe-oxides. Vertical profiles through anaerobic portions of the aquifer exhibited regimes that had progressed from iron-reducing conditions to methanogenesis. Methanogenic conditions existed both within the nonaqueous phase contaminated regimes and below the oil where hydrocarbon concentrations were high and aquifer permeability was high. These results suggested that advective transport played an important role in which zones first supported methanogenic activity. It was also found that Fe(II) concentrations and proximity to the water table were also important factors in controlling regimes of methogenesis and hydrocarbon degradation. Sustained methogenesis was found only to occur below the lowest water table elevation during seasonally oscillating conditions of the capillary fringe. Bennett et al. (2000) investigated microbial controls of mineralgroundwater equilibria in a petroleum-contaminated aquifer. The authors investigated the relationships between mineral alteration, groundwater

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chemistry, and microbial colonization. Scale effects were observed for microbial influences on mineral weathering processes at (1) the macroscale through perturbations of groundwater geochemistry and thus mineralwater equilibria, and (2) the microscale where attached organisms influence local-scale mineral-water equilibria releasing trace nutrients from dissolving minerals. In a similar manner, Thorn and Aiken (1998) investigated contaminant hydrocarbon dynamics in an unconfined glacio-fluvial aquifer and found that during oxic conditions, carbonate dissolution was controlled by heterotrophic respiration resulting in the production of excess carbon dioxide. During anoxic conditions, calcite overgrowths occurred on calcite mineral surfaces and were thought to be due to consumption of groundwater acidity by iron-reducing bacteria. At the microscale, microbes tended to preferentially colonize feldspars in anoxic regimes apparently due to the availability of P in apatite inclusions, and feldspar dissolution was found to be accompanied by precipitation of secondary minerals. As groundwater oxygen increased downgradient the oil pools, aerobic microorganisms became dominate resulting in carbonate dissolution and Fe(III)-oxide precipitation and microbial colonization did not appear to be an important mechanism under aerobic conditions. 5.3.2. Btex Benzene, toluene, ethylbenzene, and xylene (BTEX) are a group of volatile organic compounds (VOCs) found in petroleum hydrocarbons and other common environmental contaminants that can have major human health effects and target the central nervous system. BTEX compounds are common groundwater and soil contaminants that occur near petroleum and natural gas production sites, gasoline stations, and other storage areas containing petroleum-related products. They are considered one of the major causes of environmental pollution because of widespread occurrences of leakage from underground petroleum storage tanks and spills at petroleum production wells, refineries, pipelines, and distribution terminals (Fries et al., 1994). It is estimated that more than 35% of the 1.4 million gasoline storage tanks in the USA are leaking into subsurface soil and groundwater (Harwood and Gibson, 1997) resulting in extensive belowground BTEX contaminant plumes (Fig. 13). BTEX compounds can undergo aerobic metabolism which includes biodegradation by a variety of pathways. Whereas benzene is degraded to a substituted catechol, toluene degradation may follow many separate biodegradative pathways. Many microbially mediated pathways also exist for ethylbenzene which can be degraded to 3-ethylcatechol, and xylene compounds can be metabolized to mono-methylated catechols. Anaerobic pathways of BTEX biodegradation are also prevalent in subsurface environments depleted of DO (Heider and Fuchs, 1997) with toluene and ethylbenzene biodegradation generating benzoyl-CoA as an intermediate,

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Drinking well

Gas tank

Volatilization Gas leak Mobile phase

Microbial degradation

Sorption on to soil Dissolved phase Groundwater flow (Q)

Figure 13 Schematic illustration of BTEX contamination resulting from leakage of gasoline from faulty and poorly maintained underground storage tanks. Once released to the environment, BTEX can volatilize (evaporate), dissolve, attach to soil particles, or degrade biologically (from http://www.envirotools.org/factsheets/btex.shtml).

which is the most common central intermediate of anaerobic aromatic metabolism (Heider and Fuchs, 1997). Kao et al. (2006) investigated the influence of coupled hydrologic, geochemical, and microbial processes on the propensity of natural attenuation processes to remediate a petroleum-hydrocarbon spill site. Numerous lines of evidence were used to suggest that natural biodegradation was the major factor observed in contaminant reduction, which included (1) significant depletion of DO, nitrate, and sulfate; (2) production of groundwater Fe(II), S2, and CO2; (3) decrease in BTEX along the transport path coupled with limited spreading via dispersion; (4) increased alkalinity and microbial populations; and (5) preferential removal of key BTEX components along the transport pathway. As such, successful bioremediation of BTEX contaminants often depends on knowledge of the subsurface mineralogy and aqueous phase geochemistry. Multivariate statistics and artificial neural networks have been used to link geochemistry with microbial community analyses and thus the propensity for biodegradation (Feris et al., 2004; Lee et al., 2001). Maurer and Rittmann (2004) have shown that abiotic geochemical processes such as precipitation and dissolution of calcite and surface interactions with iron sulfide minerals are important in the destruction and attenuation of BTEX. At the Bemidji, MN, crude-oil spill site, Cozzarelli et al. (2001) found that groundwater redox dynamics and changes in Fe reduction had a pronounced influence on the behavior of a subsurface hydrocarbon plume. Pore-scale analysis indicated that the hydrocarbon plume had been growing over a two-decade period due to the depletion of solid phase

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Fe(III)-oxides. However, Lovley et al. (1989) showed bacteria catalyzed oxidation of aromatic hydrocarbons may be coupled to ferric-iron reduction in systems containing ferric oxide minerals. Thus, the depletion of ferric iron that was observed by Cozzarelli et al. (2001) may have also affected the development of other TEAP zones within the contaminated plume as xylene contaminants were shown to migrate along a thin layer within the aquifer that had undergone methanogenic conditions. At another crude oil contaminated site, Baedecker et al. (1993) found that hydrocarbon degradation was initially associated with Fe(III) reduction processes; however, as ferric iron was depleted (Tuccillo et al., 1999), methanogenic zones were formed in areas of high contaminant flux (Bekins et al., 1999). Cozzarelli et al. (2001) also noted that plume-scale observations differed from these smaller-scale observations since the largerscale observation suggested that the extent of the Fe(II) and BTEX plume had changed very little during the second half of the second decade of study. Richnow et al. (2003a,b) quantified in situ microbial degradation of aromatic hydrocarbons in a contaminated anoxic aquifer using geochemical isotopic fractionation. The isotopic fractionation technique confirmed xylene and dimethylbenzene biodegradation as well as the distinction between biodegraded aromatics and untransformed aromatics. Several other studies have also used the stable isotope technique for C and H to track the biodegradation of benzene and other aromatic hydrocarbon contaminants in groundwater (Gray et al., 2002; Griebler et al., 2004; Kuder et al., 2005; Mancini et al., 2003; Morasch et al., 2004; Steinbach et al., 2004). The technique is highly sensitive and informative since the elemental isotopic ratios will change during biodegradation due to preferential enrichment or depletion of one of the isotopes, whereas isotopic ratios remain constant in response to geochemical and hydrological impacts such as adsorption, dilution, and evaporation. 5.3.3. Coal-tar/creosote Coal-tar is a highly viscous liquid that is produced when coal is carbonized or glasified to make coke or coal gas, respectively. Coal-tar products are used in medicines to treat skin diseases such as psoriasis, and are used as wildlife repellents, insecticides, and fungicides. Coal-tar derivatives are also used for roofing, road paving, and coking. Coal-tars are complex mixtures of phenols, polycyclic aromatic hydrocarbons (PAHs), and various heterocyclic compounds. Coal-tar creosote is a thin oily liquid that is typically used as a wood preservative and is classified as a DNAPL since it has a density slightly greater than water. It may consist of as many as 200 different organic compounds and on average is composed of 85% PAHs by mass, 10% phenolic compounds, and 5% heteroaromatic type compounds (Mueller et al., 1989). Williams et al. (2001) investigated the hydro-bio-geochemical characteristics of a subsurface coal-tar distillate plume in a sandstone aquifer described by Bridge (1997). The authors found that redox conditions and


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the production of methane and CO2 suggested natural attenuation of the coal-tar plume; however, the kinetics of degradation were found to be slow. The presence of the electron acceptor sulfate in the plume suggested that methanogenesis was probably limited. Microbial characterization showed a diverse array of microbial communities that had the potential for both aerobic and anaerobic degradation of the coal-tar contaminant. The authors found that microbial activity was greatest at the leading edge of the plume and that degradation in the core of the plume was limited, possibly due to phenol toxicity. It was proposed that once the plume was hydrologically diluted due to groundwater dispersion, natural microbial attenuation of the hydrocarbons could proceed. Pickup et al. (2001) investigated the influence of coupled processes on the biodegradation potential of phenol and other tar acids in a contaminated aquifer in the West Midlands, UK (Thornton et al., 2001). The potential for phenol degradation was found to be influenced by the concentration of the contaminant and the total bacterial cell count that was present in the groundwater. The observed microbial activity complemented results obtained through chemical analysis, and when combined with hydrologic data, provided a realistic profile of plume effects that could be related to the potential for natural attenuation at the site. The authors stressed that microbial data suggesting favorable conditions for natural attenuation without accompanying chemical data may result in an incorrect assertion that microbial attenuation processes are operative (Stapleton and Sayler, 1998; Williams et al., 2001). King and Barker (1999) and King et al. (1999) investigated the influence of hydrological, geochemical, and microbial processes on the fate and transport of coal-tar creosote in the Canadian Forces Base (CFB) Borden site, located near Toronto, Ontario, Canada. The authors tracked a welldefined source of seven representative creosote compounds over a four-year period as they developed into a plume downgradient within the aquifer. It was noted that the different compounds within the common source showed markedly different patterns of plume development and that significant transformations in compound mass occurred during transport which impacted the behavior of the overall contaminant plume. Phenol was found to dissolve quickly from the source, thus migrating downgradient as a discrete slug. The phenol plume was nearly absent after 2 years owing to transformation by microbial degradation. The xylene plume was found to migrate to a maximum distance at around 2 years, after which time the plume receded back toward the source as the rate of xylene mass flux from the source decreased below the rate of xylene microbial degradation. Carbazole exhibited similar behavior as xylene, although the overall kinetic reactions controlling its migration were much slower. King et al. (1999) used several lines of evidence to support that the loss of contaminants were due to microbial degradation reactions. Geochemical redox indicators showed that DO and sulfate decreased in the groundwater within the

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plume while significant increases in dissolved Fe(II), Mn(II), methane, and aromatic acids were observed. Further, measurements of PLFA in the aquifer sediments indicated that microbial biomass and turnover rate were greater within the plume than adjacent to the plume, which is consistent with systems undergoing biodegradative processes. King et al. (1999) also found that the naphthalene plume continued to advance and increase in mass over the 4-year observation period suggesting minimal biodegradation. These results are consistent with Ramaswmi et al. (1994) who investigated the role of physical and chemical mass transfer processes on the biodegradation of PAH compounds that were derived for residual coal-tar that was present within microporous media. The authors found that the release kinetics of naphthalene from coal-tar to the bulk aqueous solution was more rapid than the biotransformation rates for this compound. In a similar manner, Broholm et al. (1999) investigated the transport of biodegradation of 25 organic compounds typical of creosote in a fractured clay till soil from Funen, Denmark. At low contaminant concentrations, significant biodegradation was observed for many of the compounds, with the presence of nitrate and oxygen enhancing the degradation process. Higher concentrations of creosote compounds resulted in little biodegradation and the contaminants were transported through the structured media in a similar manner as nonreactive Br. Thornton et al. (2001) investigated the impact of coupled processes on the distribution and natural attenuation of phenol, cresols, and xylenols in a deep Triassic sandstone aquifer that was contaminated by a historical coaltar distillation plant. Overlapping contaminant plumes existed at the site including phenols, mineral acids, and a highly alkaline condition that resulted from waste NaOH, with their distribution primarily related to historical source releases. The authors found that contaminant degradation was occurring via aerobic respiration nitrate reduction, Mn(IV) and Fe(III) reduction, sulfate reduction, methanogenesis, and fermentation, with accumulation of various products including inorganic carbon, organic metabolites, acetate, methane, dissolved hydrogen, and reduced forms of Fe, Mn, and S. Respiratory processes were found to be rate limiting with regard to the spatial distributions and dynamics of hydrogen and TEAPs, and the aerobic processes were thought to be controlled by the mixing of uncontaminated aquifer groundwater, rich in DO and nitrate, with the contaminant plume. Contaminant transformations by geochemical oxidation reactions were found to be minimal since mineral oxide and sulfate consumption was small relative to their total system mass. Overall biodegradation rates were found to be slow, but were expected to increase with time as contaminant concentrations decreased due to plume hydrologic dispersion and the increased efficiency of intruding DO from outside the plume. The authors suggested that hydrologic transport processes may exert a greater

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Aerial drift Adsorption

Atmospheric deposition Crop removal

Photodegradation

Ru

Volatilization

Chemical degradation

Adsorption desorption

no

ff

Biological degradation

Leaching

Ground water

Figure 14 Schematic illustration of various hydrological, geochemical, and microbial processes that influence the fate and transport of pesticides in terrestrial environments (from http://www1.agric.gov.ab.ca/$department/deptdocs.nsf/all/wat3350).

control on microbial-based natural attenuation of the plume as compared to geochemical factors such as aquifer oxidant availability.

5.4. Pesticides and herbicides Over the past several decades, extensive research has sought to provide an improved understanding and predictive capability of pesticide and herbicide fate and transport in surface water, groundwater, and the vadose zone (Bloomfield et al., 2006; Muller et al., 2007; Sarmah et al., 2004; Fig. 14). Pesticide use as of the year 2000 has increased 50-fold since 1950, and 2.5 million tons of industrial pesticides are now used each year to enhance agriculture production and decrease human disease carrying insects. The major source of nonfarming human exposure to pesticides is through diet and it is believed that long-term chronic exposures can increase the risk of cancer, infertility, and cause disruptions to the endocrine system and possible mutagenic effects. Vinther et al. (2001) investigated the impact of hydrological and geochemical processes on the microbial degradation of pesticides in loamy and sandy soils. Bacterial biomass, enzymatic activity, C utilization patterns, and

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pesticide mineralization were monitored. Bacteria biomass and activity, and C utilization in the macroporous portion of the loamy soil were higher than that of its surrounding matrix with the macroporous portion of the soil having a higher pesticide degradation potential relative to the surrounding matrix portion of the soil. The authors suggested that the higher abundance of nutrients and pesticides in the macropore channels relative to the matrix may have created a more favorable environment for microbial activity and the potential for pesticide degradation. Vinther et al. (1999) also found that macropore-type soils had higher concentrations of nitrate and DOC relative to matrix dominated soils, with the latter having fewer numbers of bacteria. Tariq et al. (2006) evaluated the influence of soil structure, temperature, soil water content, and microbial activity on the persistence of the pesticides carbosulfan, carbofuran, lambda-cyhalothrin, endosulfan, and monocrotophos in a sandy loam soil from Pakistan. Microbially mediated degradation of the pesticides was found to be kinetically controlled with the degradation rate enhanced by an increase in soil temperature and moisture, and the degradation rate slowed by conditions of limited organic carbon. Movement of the pesticides into deeper soil horizons was attributed to preferential hydrologic flow during storm events, where the resident time in the media was short, causing microbial degradation processes to be minimal due to their kinetic nature (Ghodrati and Jury, 1992; Jury et al., 1987; O’Dell et al., 1992). Bolduan and Zehe (2006) investigated the microbial degradation kinetics of the herbicide isoproturon in soil macro- and micropores within a soil catchment in SW Germany. The authors found that herbicide degradation kinetics within earthworm constructed soil macropores was as rapid as nearsurface organic-rich topsoils. This observation may have been the result of organic rich coatings that can develop within the macroporous channels. The authors also noted that herbicide degradation rates for the soil matrix that surrounded the macropores, was an order of magnitude lower than that observed in the macropore domains. This was attributed to the lower microbial activity that was present in the soil matrix (i.e., microporosity). This study stresses the importance of media structure on controlling the degradation rates and preferential transport of herbicides, and other organic contaminants, in macropores versus slow transport through micropores. Pivetz and Steenhuis (1995) investigated the influence of soil structure on the transport and biodegradation of the pesticide 2,4-dichlorophenoxyacetic acid (2,4-D). The authors stressed that preferential flow of pesticides in macropores can lead to a decreased residence time through the soil which can enhance the possibility of groundwater pollution. However, they point out that macropores may present a more favorable environment for biodegradation due to greater oxygen, nutrients, substrate supply, and higher microbial populations, particularly in earthworm burrows, relative to the


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soil matrix where micropores exists. Pivetz and Steenhuis (1995) noted that pesticide degradation rates within macropores was significantly greater than within micropores, presumably due to increased microbial activity and numbers in the larger pore type. Nevertheless, the authors found that both macropore and micropore flow paths resulted in pesticide degradation with the rate of biodegradation increasing with time in each flow path type. Muller et al. (2007) provided an extensive review of the influence of effluent agricultural irrigation on the fate and transport of pesticides in soil. Treated effluents for agricultural land use, otherwise known as reclaimed water, include municipal wastewater, farm effluents (dairy, piggery), and effluents from food and plant processing. These effluents typically have high concentrations of natural and synthetic DOC (e.g., natural humics vs surfactants and solvents) which can influence the geochemical nature and thus transport of pesticides. Muller et al. (2007) stressed the importance of soil properties on the transformation and transport of the organic contaminants and in particular the influence of DOC on pesticide mobility and biodegradation. Co-transport of pesticides via complexation with DOC may result in accelerated transport through soil by decreasing pesticide sorption, or the DOC may enhance pesticide biodegradation, thereby decreasing its mobility, by providing an energy source for microorganisms that are capable of pesticide degradation.

5.5. Modeling coupled processes involving organic constituents Numerous multicomponent reactive transport models involving nonaqueous phase constituents have been developed over the years that couple hydrodynamic transport with multiprocess, time-dependent geochemical and microbial reactions. The multiphase flow and multicomponent reactive transport simulator, PARSim has been linked to a mixed chemical kinetic and equilibrium model (KEMOD) to allow simulation of multiple flowing phases with a full complement of reactions (Arbogast et al., 1996). The model (RPARSim/KEMOD) allows for a more general, nonequilibrium phase transfer for KEMOD-style reactions where the reactants and products are in different phases. The model has been parallelized in order to enhance computational efficiency and the need to simulate larger-scale, more realistic environmental problems. Other kinetic-based models designed to deal with subsurface DNAPL issues tend to emphasize substrate-limited biodegradation. The EPA code, BIOPLUME III is a 2D contaminant transport model that couples DNAPL advective-dispersive transport with sorption, first-order decay, and biodegradation through instantaneous, zero-order, first-order, or Monod kinetics. The model is based on the USGS BIOMOC code where the hydrocarbon source and each active

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electron acceptor (e.g., O2, NO3–, Fe(III), SO42, and CH4) are simulated as separate plumes. Another numerical simulator named MT3D99, which is an enhancement of MODFLOW, also couples advective-dispersive transport in soil systems with nonequilibrium sorption, time-dependent nonaqueous phase liquid dissolution, and rate-limited microbial processes. It considers BIOPLUME-type reactions, monad reactions, and daughter products, thus enabling the simulation of multispecies transport in a similar manner as the transport code RT3D. Kinetically limited hydrocarbon biodegradation using multiple electron acceptors and time-dependent transport of bacteria, electron acceptors, and hydrocarbons are explicitly considered. MT3D99 also maintains a dual porosity option, where the soil media is divided into an advective dominated mobile domain and a diffusion dominated immobile domain. An empirical first-order parameter accounts for mass transfer between the domains. Recently, Barry et al. (2002) reviewed modeling investigations on the fate of oxidizable organic contaminants in groundwater. A comprehensive modeling framework was specified, including geochemical reactions and interphase mass transfer processes such as sorption/ desorption, nonaqueous phase liquid dissolution, and mineral precipitation/ dissolution, all of which can be equilibrium or kinetically controlled. As well, the framework was specified to simulate microbially mediated transformation/degradation processes and microbial population growth and decay. Microbial degradation reactions allowed for limitations based on the availability of nutrients and electron acceptors (i.e., changing redox states), as well as concomitant secondary reactions. Phanikumar et al. (2005) developed a 3D numerical model to describe microbial transport and biodegradation of CT at the Schoolcraft, MI, site (see Section 5.2 above for experimental details of this investigation). The model simulates transport and reaction of solution and sorbed CT, acetate, electron acceptor nitrate, mobile and immobile microbes, and nonreactive tracers (e.g., Br). Microbial processes included growth, decay, attachment, detachment, and endogenous respiration. The model was found to predict observed acetate and nitrate concentration profiles quite well; however, a lower CT degradation rate, relative to that determined in laboratory studies (Phanikumar et al., 2002), was needed to describe the CT concentrations observed in situ after the inoculation event. Essaid et al. (2003) used the USGS multicomponent solute transport and biodegradation code BIOMOC and inverse modeling code UCODE to simulate BTEX dissolution and biodegradation at a crude oil spill site in Bemidji, MN. Historical experimental data from 1986 to 1997 was used and the inverse modeling strategy successfully described the results when coupled transport and degradation processes were used and a single dissolution rate coefficient was used for all BTEX components. Model parameters consistent with subsurface coupled processes were used including hydraulic


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conductivity, dispersion, dissolution kinetics, and anaerobic microbial degradation rates. The calibrated simulations reproduced the general large-scale evolution of the plume, but did not reproduce small-scale spatial and temporal variabilities in concentration. It was suggested that anaerobic degradation removed 77% of the aqueous phase BTEX versus 17% for aerobic processes. Prommer et al. (2006) investigated field-scale reactive transport modeling strategies to simulate capillary fringe controls on natural attenuation of phenoxy acid herbicides (e.g., mecoprop MCPP) in a landfill plume. Attenuation processes were noted in transition zones between the anaerobic plume core and the overlying aerobic water body. The location of the transition zone was controlled by vertical transverse dispersion processes occurring downgradient the contaminant source term. Simulations involved a 2D vertical cross section to quantify the combined physical, geochemical, and microbial processes influencing the herbicide fate and transport processes. The capillary fringe regime was found to control the aerobic degradation of the phenoxy acids. Yabusaki et al. (2001) also used a multicomponent reactive transport code to assess the in situ destruction of chlorinated hydrocarbons by a Fe0 PRB. Both equilibrium and time-dependent hydrocarbon degradation, iron dissolution, secondary mineral precipitation, and a variety of complexation reactions were considered. Dominant precipitants in the PRB zone were Fe-oxides, siderite, aragonite, brucite, and iron sulfide which are PRB-mediated mineral phases observed in the experimental findings of Gu et al. (2002a) and Wilkins et al. (2006). Model predictions suggested that mineral precipitants could account for a 3% increase in mineral volume per year which could have significant implications for the long-term performance of subsurface barriers of this type. The authors suggested that the inclusion of transport (hydrodynamics) within the simulation was paramount to understanding the interplay between rates of transport and rates of reactions and therefore a more accurate assessment of barrier longevity and performance and the understanding of mechanisms responsible for contaminant destruction and immobilization. Recently, Lichtner and Wolfsberg (2004), Hammond et al. (2005), Lu and Lichtner (2005), and Mills et al. (2005) described a newly developed high-performance simulator, PFLOTRAN, which is a massively parallel 3D multiphase, multicomponent simulator of subsurface flow and reactive transport. Since PFLOTRAN was built on top of PETSc, the Portable, Extensible Toolkit for Scientific Computation (Balay et al., 1997), the code exhibits excellent performance on the world’s largest-scale supercomputers, such as that at Oak Ridge National Laboratory which is a Cray XT3/4 system consisting of 11,706 dual-core Operon processor nodes. The PFLOTRAN code solves a system of mass and energy conservation equations for a number of phases including water, supercritical CO2, black oil, and a

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gaseous phase. It describes coupled thermal, hydrologic, and chemical processes in variably saturated, nonisothermal, porous media in multiple spatial dimensions. The code has been written with parallel scalability in mind and can run on single processors to the largest massively parallel computer architectures. Many enhancements are planned for PFLOTRAN over the next several years, such as methods to upscale soil processes that are typically nonlinear and scale dependent, in an effort to simulate large-scale (watershed and basin scale) contaminant fate and transport scenarios (Mills, ORNL, personal communication).

6. Concluding Remarks This chapter has shown that subsurface contaminant fate and transport processes are invariably impacted by coupled hydrological, geochemical, and microbial processes. Many times the assessment of process contribution to the overall flux of contaminants is difficult since the interactions of the processes are often nonlinear, time dependent, and complex. Published multiscale research endeavors in these areas over the past several decades, such as those described above, have provided excellent scientific knowledge and prediction of contaminant fate and transport issues that can be used for decision-making and assessment of natural attenuation or manipulative remediation strategies. This chapter has focused on the impact of coupled processes on legacy waste issues that have plagued society for many years. It is the author’s belief that the continued investigation of subsurface coupled processes is imperative in order to deal with future environmental issues of global concern. Because of a ‘‘business as usual’’ mentality among the industrial societies, four main topical areas are perceived as standouts with regard to imparting adverse environmental consequences on the world over the next century ( Jack Parker, University of Tennessee, personal communication, 2007). These topical areas include (1) energy availability, (2) climate change, (3) water quality and supply, and (4) land use change. Each one of these topical areas not only exhibits huge potential environmental impacts upon the earth’s terrestrial and aquatic ecosystems, but they potentially will have massive economic and societal implications on the world human population for many years to come. Continued improvements to our conceptual and predictive understanding of these environmental issues will require fundamental knowledge of the coupled processes that dictate behavior responses of associated contaminants in the geosphere. Each of the four perceived environmental issues listed above are briefly discussed below with the intention of providing the reader with some complex environmental issues that are in need of investigation and resolution.


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The notion of the world’s oil supplies beginning a downward trend have prompted the consideration of alternative fuel sources and types (Altun et al., 2006; Hill, 2007; Kok, 2002; Petroll and Tveiten, 2007; Sanderson et al., 2006; Smeets et al., 2007). As easily accessible sources of oil become depleted, alternative crude sources such as those found in deep shale formations will become more attractive. However, the environmental consequences of such action are unpredictable, with current technologies creating vast regions of organic contaminated soil and imparting massive greenhouse effects which soon may be a tremendous financial liability (Kahru et al., 2002; Pollumaa et al., 2001). In addition, in situ shale oil extraction creates large thermal, pressure, and hydraulic gradients that may allow organic and inorganic contaminants to escape from the retort zone into groundwater and surface water sources (Kahru and Pollumaa, 2006). Other alternative energy sources such as biofuels and an increased use of nuclear power plants may also become much more attractive as crude production declines (Boczar et al., 1998; Hahn-Hagerdal et al., 2006; Petroll and Tveiten, 2007). Each of these energy sources imparts its own potential environmental impacts (Converse, 2007; Hill, 2007; Jonsson and Hillring, 2006), where spent-fuel nuclear waste disposal issues continue to plague society in our current environment. Subsurface contaminant fate and transport issues associated with future energy production strategies will most certainly be an issue of global environmental concern. Climate change as a result of anthropogenic emissions is a strongly debated topic and one that most certainly will not be resolved until it is far too late for immediate corrective action (Alcamo et al., 2005; Oppenheimer and Petsonk, 2005). Lal (2007) discusses the daunting environmental consequences of the emerging carbon civilization on the planet Earth. Because of political agendas and ignorance toward technical and scientific realities related to what appear to be certain indicators of climate change, short- and long-term consequences of climate change on the world’s environment are largely unknown. Not only will aboveground processes be influenced by changing climates, but belowground processes as we currently understand them will be altered as well. Belowground temperature increases may accelerate organic C turnover rates and possibly disrupt agricultural and silviculture productivity with unpredictable consequences to the environment. Subsurface solute fate and transport issues associated with future climatic change impacts will again most certainly be an issue of global environmental concern. Future shortages in water supply and quality are foreseen even without the influence of climate change ( Jury and Vaux, 2005; Lal, 2007; Tao et al., 2003). Stress on the world’s water supply will severely impact agriculture and energy productivity as well as human health and quality of life (Vitale et al., 2003) which will in turn create unpredictable environmental consequences of global concern. New strategies will be necessary to optimize

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water use and management of reservoirs and watersheds, and advancements in science and technology will be needed to optimize water recycling and reuse (Arnell and Delaney, 2006; Shelef and Azov, 1996). Groundwater and surface water quality issues will be of increasingly significant concern with regard to human and animal health, agricultural production, and potential disastrous ecological change due to toxins and salinity effects. Important water management problems can only be adequately addressed from a holistic view of the water cycle in a world with hydro-geo-bio- complexity. Understanding and predicting subsurface solute fate and transport processes will be an important component when addressing such water quality issues. Land use change is also envisioned to impart adverse environmental consequences on the world over the next century ( Johnson et al., 2007; Lal, 2007; Liu and Chen, 2006; Tomich et al., 2004). As above- and belowground environments succumb to the aggressive influence of economic productivity, severe environmental impacts to the world’s hydrologic cycle and biodiversity functions are foreseen. The loss of massive sectors of aboveground biomass and diversity, increased soil erosion, and conversion of wetlands into human habitats are examples of changing land characteristics that could have huge impacts on the global hydrologic cycle which will in turn impact global biogeochemical processes in terrestrial and aquatic ecosystems (Liu and Chen, 2006; Zhao et al., 2006). Emerging infections of humans and wildlife are often related to land use change as evolutionary relationships between hosts and pathogens are altered ( Johnson et al., 2007). Ecological changes in aquatic systems typically involve eutrophication which broadly enhances infection and pathology of human and wildlife parasites. Vast municipal landfills, covering large tracks of land, plague many underdeveloped countries and pose a severe threat to human health since dump sites are unrestricted, unmanaged, and publicly accessible. Such dump sites are often ravaged for basic necessities required for survival or for material that is saleable in order to generate income, thereby exposing unsuspecting individuals to unacceptable levels of heavy metals and toxins that are present in the dump site soil, water, and air. It is estimated that 25% of all deaths in poor countries is linked to environmentally related illnesses. The influence of land use change on subsurface solute and contaminant fate and transport processes is therefore an area of concern since disruption of the soil structure, chemical and microbial environment, and increased propensity for soil organic C loss and erosion will create an environment in severe nonequilibrium with a pathway toward stability that is currently unpredictable. Scientific investigation of the environmental consequences of future energy production, climate change, water quality and supply, and land use issues will require an improved experimental and predictive capability of coupled subsurface processes that are spatially and temporally variable across scales ranging from molecular to basin levels. The demand for energy,


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resources, and potable water will require a keen understanding of the relationships between hydrological, geochemical, and biological processes in subsurface environments. As stated by Geesey and Mitchell (2008), continued and expanded research in these areas is necessary for the (1) protection of world’s aquifers and surface waters from contamination, (2) appropriate disposal of hazardous waste, (3) protection of ecosystems from chemical, radioactive, or biological contamination, (4) sustained agricultural productivity, (5) identification and wise use of energy and mineral resources, and (6) mitigation of global climate change.

ACKNOWLEDGMENTS This research was sponsored by the U.S. Department of Energy (DOE), Office of Science, Biological Environmental Research, Environmental Remediation Sciences Division (ERSD). The Environmental Sciences Division (ESD) of the Oak Ridge National Laboratory (ORNL) is managed by UT-Battelle, LLC, for the U.S. Department of Energy under Contract DE-AC05-00OR22725. The author wishes to thank Dr. Donald L. Sparks, editor of this book, for the opportunity to prepare the following chapter, and to thank Beth Bailey of the ESD for reference compilation and editing. The author is also grateful for the financial and moral support provided by the DOE ERSD technical representatives Todd Anderson, Paul Bayer, David Lesmes, and Michael Kuperberg.

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Uptake and Fate of Perchlorate in Higher Plants Angelia L. Seyfferth and David R. Parker Contents 1. Introduction 1.1. Perchlorate in the environment 1.2. Toxicological issues 1.3. Objectives of review 2. Perchlorate Levels in Plants 2.1. Plants growing on highly contaminated sites 2.2. Market surveys 3. Perchlorate Uptake Studies 3.1. Phytoremediation 3.2. Mechanistic studies 4. Conclusions and Future Research References

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Abstract Perchlorate recently emerged as a drinking water contaminant, and its high water solubility and relatively unreactive nature under ambient conditions make it a persistent and mobile contaminant. Perchlorate interferes with iodine uptake by the human thyroid, potentially leading to adverse effects on normal metabolism and cognitive function in sensitive groups. There is an interest in the fate of perchlorate in higher plants because phytoremediation is a promising remediation option, and because there is mounting concern about human exposure to perchlorate from contaminated produce. Perchlorate is taken up by many higher plants and is mainly stored in leaves, although perchlorate is also found in smaller quantities in fruits, stems, seeds, and roots. Transpiration plays a key role in delivery of perchlorate to plant roots, and it appears that perchlorate traverses the root cell membrane via the same ion transporter as for nitrate. Certain plants are able to metabolize high concentrations (mg/liter) of perchlorate within their leaves (phytodegradation) by way of chlorate and chloride intermediates to chloride, although this process Department of Environmental Sciences, University of California, Riverside, California 92521 Advances in Agronomy, Volume 99 ISSN 0065-2113, DOI: 10.1016/S0065-2113(08)00402-1

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is slower than ex situ microbial degradation in the root zone (rhizodegradation). However, it is currently unknown whether higher plants will metabolize smaller quantities (i.e., mg/liter concentrations) of perchlorate in vivo. More research is needed in order to determine the extent of translocation, phytodegradation, and exudation of perchlorate and its metabolites, as well as the ability of modified stems and roots to store perchlorate.

1. Introduction In the last 10 years, perchlorate (ClO 4 ) has emerged as a controversial environmental contaminant. Its detection in ground and surface waters has increased with the advent of new analytical techniques and, more recently, it has been found in a range of concentrations in vegetation (El Aribi et al., 2006; FDA, 2005; Sanchez et al., 2005a,b, 2006), organisms (Smith et al., 2004, 2006), beverages (El Aribi et al., 2006), vitamins (Snyder et al., 2006), breast milk (Kirk et al., 2005, 2007), and baby formula (Pearce et al., 2007). Perchlorate ingestion interferes with iodine uptake by the thyroid gland, and may result in a lower production of key thyroid hormones in humans. The US Environmental Protection Agency has recently adopted a reference dose (RfD) of 0.7 mg per kg of body weight per day (see http://www.epa.gov/ IRIS/subst/1007.htm), but the extent of exposure through consumption of contaminated produce is currently unknown. Perchlorate is taken up by a wide variety of plants, including fresh produce, and the mechanisms involved in its uptake and fate in plants have only recently started to unfold. From both an ecological and a human health standpoint, an understanding of the uptake and fate of perchlorate in higher plants is becoming increasingly important.

1.1. Perchlorate in the environment 1.1.1. Chemical properties Perchlorate is an inorganic ion consisting of one chlorine (VII) atom surrounded by four oxygen atoms (Fig. 1) with a delocalized negative charge (Urbansky, 1998). From a thermodynamic standpoint, perchlorate is a potent oxidant, but its tetrahedral symmetry leads to very sluggish kinetics under typical, ambient conditions (Urbansky, 1998, 2002). The high solubility of perchlorate salts and the unreactive nature of perchlorate under ambient conditions make it both a favored industrial oxidant as well as a persistent environmental contaminant. 1.1.2. Production and use Ammonium perchlorate and/or potassium perchlorate salts are produced and used mainly as oxidizing additives in solid rocket propellant, munitions, and explosives, but perchlorate salts are also used in analytical chemistry,

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O

O Cl O

O

Figure 1

Tetrahedral structure of the perchlorate anion.

airbag manufacture, leather tanning, pyrotechnics, and road flares (Urbansky, 1998). Most of the environmental contamination is due to the manufacture and use of ammonium or potassium perchlorate for the defense industry. Due to the short shelf life of the propellants, unused portions must be disposed of and this is usually done with high-pressure water washes. Since perchlorate has historically not been regulated, large volumes of perchlorate-laden wastes have been legally discharged over the last 50 years (Urbansky, 1998). Because of the chemical properties of perchlorate, environmental disposal is largely responsible for the widespread contamination of ground and surface waters, especially in the western United States where many manufacturers and users of perchlorate salts are located. For instance, the low-level contamination in the Colorado River south of Lake Mead is associated with discharges over the last 50 years from PEPCON and Kerr McGee, which are the two largest manufactures of ammonium perchlorate salts in the United States.

1.1.3. Natural occurrence In addition to its anthropogenic sources, perchlorate also forms naturally under certain environmental conditions. The Atacama Desert of Chile is notoriously associated with natural geologic deposits of perchlorate salts. These soils are nitrate-rich and have been a source of nitrate fertilizer across the United States and elsewhere, but many commercial fertilizers from this Chilean saltpeter contain low levels ( romaine > green leaf > red leaf > iceberg (Sanchez et al., 2005b). In another study of citrus from California and Arizona, perchlorate was highest in leaves (669–4930 mg/kg DW) of lemon trees (Citrus limon L. Burm f.), less in fruit (64–195 mg/kg DW), and even less in branches ( 0.05. Modified from Seyfferth et al. (2008) Linear regression analyses were conducted separately for each genotype using actual values (n ¼ 9). Superscripts on r2 values indicate the level of significance:

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smartweed (Polygonum spp.). An increase of nitrate from 3.6 to 36 mM did result in a decrease in perchlorate uptake from ca. 750 to 350 mg/kg FW, but the high variation within replicates may explain the statistically nonsignificant result (Tan et al., 2006). In other work, Nzengung et al. (1999, 2004) observed that black willow grown with Hoagland solution (nitratebased N source) accumulated more perchlorate in leaves than plants grown with Miracle-GroÒ (ammonium- and urea-based N source). Nzengung et al. (2004) also observed that after spiking 36Cl-labeled perchlorate into either Hoagland or Miracle-Gro solutions in which black willow was growing, 96% of the 36Cl activity remained in Miracle-Gro solution whereas just 76% remained in the Hoagland solution after perchlorate was completely depleted from solution. The authors conclude that plants take up more perchlorate when the N source is nitrate (Hoagland solution) compared with the ammonium (Miracle-Gro); however, there may be other chemical differences between Hoagland solution and Miracle-Gro that may be responsible for their observations (e.g., phosphate or urea concentrations). Moreover, nitrate likely inhibits rhizodegradation of perchlorate to a greater extent than ammonium, thus allowing the more sluggish plant uptake to occur. Thus, their observation is most likely a reflection of a rate differential between rhizodegradation and plant uptake in the presence of different forms of N. Nzengung and McCutcheon (2003) observed the that old-leaf perchlorate concentrations steadily decreased whereas new-leaf concentrations increased, and the authors conclude that perchlorate may move from old leaves to new leaves in some plant species (Fig. 4.). However, it is not clear whether this was due to phloem retranslocation of perchlorate per se or due to phytodegradation of perchlorate in older leaves along with concurrent uptake and translocation of perchlorate in new leaves. In recent work in our laboratory using 37Cl-enriched perchlorate, not more than 1% of the perchlorate stored in outer leaves was transported to new leaves of lettuce that was grown at solution concentrations of 10 or 50 mg/liter (Seyfferth and Parker, 2008, unpublished results). More data is needed in order to determine the mechanism of translocation and the extent (if any) of phloem retranslocation of perchlorate or perchlorate metabolites in various plant species. Data is clearly lacking on the extent of leaf or root exudation of perchlorate in higher plants. As previously mentioned, there is some evidence of exudation of perchlorate metabolites from roots after phytodegradation, but to date, there is no clear evidence of nondegraded perchlorate exudation from roots. Tan et al. (2006) grew smartweed for 30 days in 10,000 mg/liter perchlorate, and allowed one set of plants to desorb in deionized water for 15 days. Upon analyzing various plant organs before and after desorption, the authors found a significant difference between perchlorate root concentrations, but not in stems or leaves, although leaf concentration tended to increase after desorption. The differences in total

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Solution concentration of perchlorate (mg/liter)

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Solution perchlorate (mg/liter) Old leaf perchlorate (mg/kg FW) New leaf perchlorate (mg/kg FW)

Figure 4 Temporal perchlorate concentrations in solution, old leaves, and new leaves of black willow (Salix caroliniana Michx.). Adapted from Nzengung and McCutcheon, 2003 and reprinted with permission from John Wiley and Sons, Inc. (2008).

plant perchlorate concentrations before and after desorption were just 4 2% (Tan et al., 2006). More research is needed in order to conclusively determine whether root exudation of perchlorate exists in higher plants. In addition, there have been no studies on perchlorate exudation from leaves in plants that use salt glands for osmoregulation. Martinelango et al. (2006) noted that rinsing removed from 38% to 73% of perchlorate from four seaweed samples, and the authors conclude that surface adsorption may play a role in observed perchlorate concentrations in seaweed. However, given the known nonreactivity of perchlorate, this observation much more likely due to osmotic effects or to flushing out of apoplastic perchlorate. In another study, no significant difference in perchlorate concentration of field-grown wheat samples was found before and after rinsing with deionized water ( Jackson et al., 2005). The fate of perchlorate in dead leaves is not fully understood. In one study, dead leaves collected under willow and sweet gum from the Longhorn Army Ammunition Plant contained no perchlorate despite perchlorate detection in leaves of live trees (Nzengung and McCutcheon, 2003). In another study, leaf litter collected from beneath live trees had significantly

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less perchlorate than those collected from under dead trees (Smith et al., 2004). It is unknown weather perchlorate can be retained by higher plants through resorption of perchlorate upon leaf senescence, or whether the lack of perchlorate in dead leaves simply reflects microbial decomposition or leaching of fallen leaves, or both. In addition, it remains unknown if there is any metabolism of perchlorate in plants after harvesting (i.e., postharvest phytodegradation).

4. Conclusions and Future Research It is clear that a wide variety of plant species as well as plant cultivars accumulate perchlorate mostly in their leaves, followed by fruits, stems, branches, and roots. Overall, the amount of perchlorate found within plants depends on (1) the concentration of perchlorate in the growth media, (2) the duration of plant growth, (3) the plant species, (3) the presence of competing ions in solution, (4) the amount of water transpired, (5) the extent of phytodegradation within the plant, (6) the extent of exudation from the plant, and (7) the portion of the plant analyzed. Evidence suggests that transpiration plays a key role in perchlorate delivery to plant roots, where perchlorate is actively taken up by the nitrate ion transporter through cotransport with protons. The most promising conditions for perchlorate phytoremediation are (1) using plants within which perchlorate is fully phytodegraded to chloride or (2) stimulating rhizodegradation under anaerobic conditions by low nitrate availability and high concentration of electron donors. Of most concern to human health are produce in which mostly leaves are consumed. More research is needed in order to understand the extent of perchlorate metabolism within produce and the extent to which edible modified stems and roots (e.g., tubers) accumulate perchlorate, both pre- and postharvest. Additional research utilizing isotopic labeling to determine the extent of translocation, phytodegradation, and exudation of environmentally relevant levels of perchlorate and its metabolites should also be conducted.

REFERENCES Bao, H. M., and Gu, B. H. (2004). Natural perchlorate has a unique oxygen isotope signature. Environ. Sci. Technol. 38, 5073–5077. Barber, S. A. (1995). ‘‘Soil Nutrient Availability: A Mechanistic Approach,’’, 2nd edn. pp. 89–90. John Wiley & Sons Inc, New York. Blount, B. C., Pirkle, J. L., Osterloh, J. D., Valentin-Blasini, L., and Caldwell, K. L. (2006). Urinary perchlorate and thyroid hormone levels in adolescent and adult men and women living in the United States. Environ. Health Perspect. 114, 1865–1871.


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Cataldo, D. A., Garland, T. R., and Wildung, R. E. (1978). Plant root absorption and metabolic fate of technetium in plants. In ‘‘Technetium in the Environment’’ (G. Desmet and C. Myttenaere, Eds.), Elsevier, London. CDHS (1997). ‘‘Determination of Perchlorate by Ion Chromatography.’’ Department of Health Services, California. Dasgupta, P. K., Martinelango, P. K., Jackson, W. A., Anderson, T. A., Tian, K., Tock, R. W., and Rajagopalan, S. (2005). The origin of naturally occurring perchlorate: The role of atmospheric processes. Environ. Sci. Technol. 39, 1569–1575. Dasgupta, P. K., Dyke, J. V., Kirk, A. B., and Jackson, W. A. (2006). Perchlorate in the United States. Analysis of relative source contributions to the food chain. Environ. Sci. Technol. 40, 6608–6614. Duncan, P. B., Morrison, R. D., and Vavricka, E. (2005). Forensic identification of anthropogenic and naturally occurring sources of perchlorate. Environ. Forensics 6, 205– 215. El Aribi, H., Le Blanc, Y. J. C., Antonsen, S., and Sakuma, T. (2006). Analysis of perchlorate in foods and beverages by ion chromatography coupled with tandem mass spectrometry (IC-ESI-MS/MS). Anal. Chim. Acta 567, 39–47. Ellington, J. J., Wolfe, N. L., Garrison, A. W., Evans, J. J., Avants, J. K., and Teng, Q. C. (2001). Determination of perchlorate in tobacco plants and tobacco products. Environ. Sci. Technol. 35, 3213–3218. EPA (1999). ‘‘Method 314.0 Determination of Perchlorate in Drinking Water Using Ion Chromatography.’’ US Environmental Protection Agency (EPA), National Exposure Research Laboratory, Cincinnati, OH. FDA (2005). Exploratory Survey Data on Perchlorate in Food. Food and Drug Administration http://www.cfsan.fda.gov/~dms/clo4data.html. Foubister, V. (2006). Chilean fertilizer leaves perchlorate legacy. Anal. Chem. 78, 7914– 7915. Greer, M. A., Goodman, G., Pleus, R. C., and Greer, S. E. (2002). Health effects assessment for environmental perchlorate contamination: The dose response for inhibition of thyroidal radioiodine uptake in humans. Environ. Health Perspect. 110, 927–937. Hurd, R. G. (1958). The effect of pH and bicarbonate ions on the uptake of salts by disks of red beet. J. Exp. Bot. 9, 159–174. Hutchinson, S. L. (2003). ‘‘A Study on the Accumulation of Perchlorate in Young Head Lettuce.’’ US Environmental Protection Agency, Athens, GA EPA/600/R-03/003. Jackson, P. E., Gokhale, S., Streib, T., Rohrer, J. S., and Pohl, C. A. (2000). Improved method for the determination of trace perchlorate in ground and drinking waters by ion chromatography. J. Chromatogr. A 888, 151–158. Jackson, W. A., Joseph, P., Laxman, P., Tan, K., Smith, P. N., Yu, L., and Anderson, T. A. (2005). Perchlorate accumulation in forage and edible vegetation. J. Agric. Food Chem. 53, 369–373. Kirk, A. B., Smith, E. E., Tian, K., Anderson, T. A., and Dasgupta, P. K. (2003). Perchlorate in milk. Environ. Sci. Technol. 37, 4979–4981. Kirk, A. B., Martinelango, P. K., Tian, K., Dutta, A., Smith, E. E., and Dasgupta, P. K. (2005). Perchlorate and iodide in dairy and breast milk. Environ. Sci. Technol. 39, 2011–2017. Kirk, A. B., Dyke, J. V., Martin, C. F., and Dasgupta, P. K. (2007). Temporal patterns in perchlorate, thiocyanate, and iodide excretion in human milk. Environ. Health Perspect. 115, 182–186. Krynitsky, A. J., Niemann, R. A., and Nortrup, D. A. (2004). Determination of perchlorate anion in foods by ion chromatography-tandem mass spectrometry. Anal. Chem. 76, 5518–5522. Marschner, H. (2003). ‘‘Mineral Nutrition of Higher Plants’’ 2nd edn. Academic Press, London.

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Martinelango, P. K., Tian, K., and Dasgupta, P. K. (2006). Perchlorate in seawater: Bioconcentration of iodide and perchlorate by various seaweed species. Anal. Chim. Acta 567, 100–107. Motzer, W. E., Mohr, T. K. G., McCraven, S., and Stanin, P. (2006). Stable and other isotope techniques for perchlorate source identification. Environ. Forensics 7, 89–100. Nzengung, V. A., and McCutcheon, S. C. (2003). Phytoremediation of perchlorate. In ‘‘Phytoremediation: Transformation and Control of Contaminants’’ (S. C. McCutcheon and J. L. Schnoor, Eds.), pp. 863–885. John Wiley & Sons Inc., Hoboken. Nzengung, V. A., Penning, H., and O’Niell, W. (2004). Mechanistic changes during phytoremediation of perchlorate under different root-zone conditions. Int. J. Phytoremediation 6, 63–83. Nzengung, V. A., Wang, C. H., and Harvey, G. (1999). Plant-mediated transformation of perchlorate into chloride. Environ. Sci. Technol. 33, 1470–1478. Olsen, C. (1953). The significance of concentration for the rate of ion absorption by higher plants in water culture 4: The influence of hydrogen ion concentration. Physiol. Plantarum 6, 848–858. Park, J. W., Rinchard, J., Liu, F. J., Anderson, T. A., Kendall, R. J., and Theodorakis, C. W. (2006). The thyroid endocrine disruptor perchlorate affects reproduction, growth, and survival of mosquitofish. Ecotox. Environ. Safe 63, 343–352. Parker, D. R., Seyfferth, A. L., and Kiel Reese, B. (2008). Perchlorate in groundwater: A synoptic survey of ‘‘pristine’’ sites in the coterminous United States. Environ. Sci. Technol. 42, 1465–1471. Pearce, E. N., Bazrafshan, H. R., He, X. M., Pino, S., and Braverman, L. E. (2004). Dietary iodine in pregnant women from the Boston, Massachusetts area. Thyroid 14, 327–328. Pearce, E. N., Leung, A. M., Blount, B. C., Bazrafshan, H. R., He, X., Pino, S., ValentinBlasini, L., and Braverman, L. E. (2007). Breast milk iodine and perchlorate concentrations in lactating Boston-area women. J. Clin. Endocr. Metab. 92, 1673–1677. Sanchez, C., Krieger, R., and Blount, B. (2007). Potential perchlorate exposure from horticultural crops irrigated with Colorado River water. Hortscience 42, 884–885. Sanchez, C. A., Krieger, R. I., Khandaker, N., Moore, R. C., Holts, K. C., and Neidel, L. L. (2005a). Accumulation and perchlorate exposure potential of lettuce produced in the Lower Colorado River region. J. Agric. Food Chem. 53, 5479–5486. Sanchez, C. A., Crump, K. S., Krieger, R. I., Khandaker, N. R., and Gibbs, J. P. (2005b). Perchlorate and nitrate in leak vegetables of North America. Environ. Sci. Technol. 39, 9391–9397. Sanchez, C. A., Krieger, R. I., Khandaker, N. R., Valentin-Blasini, L., and Blount, B. C. (2006). Potential perchlorate exposure from Citrus sp irrigated with contaminated water. Anal. Chim. Acta 567, 33–38. Seyfferth, A. L., and Parker, D. R. (2006). Determination of low levels of perchlorate in lettuce and spinach using ion chromatography-electrospray ionization mass spectrometry (IC-ESI-MS). J. Agric. Food Chem. 54, 2012–2017. Seyfferth, A. L., and Parker, D. R. (2007). Effects of genotype and transpiration rate on the uptake and accumulation of perchlorate (ClO4) in lettuce. Environ. Sci. Technol. 41, 3361–3367. Seyfferth, A. L., Henderson, M. K., and Parker, D. R. (2008). Effects of common soil anions and pH on the uptake and accumulation of perchlorate in lettuce. Plant Soil 302, 139– 148. Shrout, J. D., Struckhoff, G. C., Parkin, G. F., and Schnoor, J. L. (2006). Stimulation and molecular characterization of bacterial perchlorate degradation by plant-produced electron donors. Environ. Sci. Technol. 40, 310–317.


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Smith, P. N., Theodorakis, C. W., Anderson, T. A., and Kendall, R. J. (2001). Preliminary assessment of perchlorate in ecological receptors at the Longhorn Army Ammunition Plant (LHAAP), Karnack, Texas. Ecotoxicology 10, 305–313. Smith, P. N., Yu, L., McMurry, S. T., and Anderson, T. A. (2004). Perchlorate in water, soil, vegetation, and rodents collected from the Las Vegas Wash, Nevada, USA. Environ. Pollut. 132, 121–127. Smith, P. N., Severt, S. A., Jackson, W. A., and Anderson, T. A. (2006). Thyroid function and reproductive success in rodents exposed to perchlorate via food and water. Environ. Toxicol. Chem. 25, 1050–1059. Snyder, S. A., Vanderford, B. J., and Rexing, D. J. (2005). Trace analysis of bromate, chlorate, iodate, and perchlorate in natural and bottled waters. Environ. Sci. Technol. 39, 4586–4593. Snyder, S. A., Pleus, R. C., Vanderford, B. J., and Holady, J. C. (2006). Perchlorate and chlorate in dietary supplements and flavor enhancing ingredients. Anal. Chim. Acta 567, 26–32. Sorensen, M. A., Jensen, P. D., Walton, W. E., and Trumble, J. T. (2006). Acute and chronic activity of perchlorate and hexavalent chromium contamination on the survival and development of Culex quinquefasciatus. Environ. Pollut. 144, 759–764. Sturchio, N. C., Bohlke, J. K., Beloso, A. D., Streger, S. H., Heraty, L. J., and Hatzinger, P. B. (2007). Oxygen and chlorine isotopic fractionation during perchlorate biodegradation: Laboratory results and implications for forensics and natural attenuation studies. Environ. Sci. Technol. 41, 2796–2802. Sundberg, S. E., Ellington, J. J., Evans, J. J., Keys, D. A., and Fisher, J. W. (2003). Accumulation of perchlorate in tobacco plants: Development of a plant kinetic model. J. Environ. Monit. 5, 505–512. Susarla, S., Bacchus, S. T., McCutcheon, S. C., and Wolfe, N. L. (1999). ‘‘Potential Species for Phytoremediation of Perchlorate.’’ US Environmental Protection Agency, Athens, GA EPA/600/R-99/069. Susarla, S., Bacchus, S. T., Harvey, G., and McCutcheon, S. C. (2000). Phytotransformations of perchlorate contaminated waters. Environ. Technol. 21, 1055–1065. Tan, K., Anderson, T. A., Jones, M. W., Smith, P. N., and Jackson, W. A. (2004). Accumulation of perchlorate in aquatic and terrestrial plants at a field scale. J. Environ. Qual. 33, 1638–1646. Tan, K., Anderson, T. A., and Jackson, W. A. (2006). Uptake and exudation behavior of perchlorate in smartweed. Int. J. Phytoremediat. 8, 13–24. Toulon, V., Sentenac, H., Thibaud, J. B., Soler, A., Clarkson, D., and Grignon, C. (1989). Effect of HCO 3 concentration in the absorption solution on the energetic coupling of Hþ-cotransports in roots of Zea Mays L. Planta 179, 235–241. Urbansky, E. T. (1998). Perchlorate chemistry: Implications for analysis and remediation. Bioremed. J. 2, 81–95. Urbansky, E. T. (2002). Perchlorate as an environmental contaminant. Environ. Sci. Pollut. Res. 9, 187–192. Urbansky, E. T., Magnuson, M. L., Kelty, C. A., and Brown, S. K. (2000). Perchlorate uptake by salt cedar (Tamarix ramosissima) in the Las Vegas Wash riparian ecosystem. Sci. Tot. Environ. 256, 227–232. Urbansky, E. T., Brown, S. K., Magnuson, M. L., and Kelty, C. A. (2001). Perchlorate levels in samples of sodium nitrate fertilizer derived from Chilean caliche. Environ. Pollut. 112, 299–302. Van Aken, B., and Schnoor, J. L. (2002). Evidence of perchlorate (ClO4) reduction in plant tissues (poplar tree) using radio-labeled (36ClO4). Environ. Sci. Technol. 36, 2783– 2788. Wolff, J. (1998). Perchlorate and the thyroid gland. Pharmacol. Rev. 50, 89–105. Yu, L., Canas, J. E., Cobb, G. P., Jackson, W. A., and Anderson, T. A. (2004). Uptake of perchlorate in terrestrial plants. Ecotox. Environ. Safe 58, 44–49.

C H A P T E R

T H R E E

Sugarcane for Bioethanol: Soil and Environmental Issues Alfred E. Hartemink* Contents 127 128 128 129 133 137 140 143 147 147 149 151 152 153 154 156 158 160 161 161 162 164 166 169 172 172

1. Introduction 2. Changes in Soil Chemical Properties 2.1. Data sources and types 2.2. Monitoring over time 2.3. Samples from different land-use systems 2.4. Soil organic matter dynamics 2.5. Leaching, denitrification, and inorganic fertilizers 2.6. Nutrient balances 3. Changes in Soil Physical Properties 3.1. Compaction and aggregate stability 3.2. Soil erosion 4. Changes in Soil Biological Properties 4.1. Macrofauna 4.2. Microbes 5. Environmental Issues 5.1. Herbicides and pesticides 5.2. Inorganic fertilizers 5.3. Air and water quality 6. Discussion and Conclusions 6.1. Sugarcane for bioethanol 6.2. Effects on the soil 6.3. Effects on air and water 6.4. Sugarcane yields 6.5. The potential for precision farming Acknowledgments References

*

ISRIC - World Soil Information, 6700 AJ Wageningen, The Netherlands

Advances in Agronomy, Volume 99 ISSN 0065-2113, DOI: 10.1016/S0065-2113(08)00403-3

#

2008 Elsevier Inc. All rights reserved.

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Abstract Cultivation of sugarcane for bioethanol is increasing and the area under sugarcane is expanding. Much of the sugar for bioethanol comes from large plantations where it is grown with relatively high inputs. Sugarcane puts a high demands on the soil because of the use of heavy machinery and because large amounts of nutrients are removed with the harvest; biocides and inorganic fertilizers introduce risks of groundwater contamination, eutrophication of surface waters, soil pollution, and acidification. This chapter reviews the effect of commercial sugarcane production on soil chemical, physical, and biological properties using data from the main producing areas. Although variation is considerable, soil organic C decreased in most soils under sugarcane and, also, soil acidification is common as a result of the use of N fertilizers. Increased bulk densities, lower water infiltration rates, and lower aggregate stability occur in mechanized systems. There is some evidence for high leaching losses of fertilizer nutrients as well as herbicides and pesticides; eutrophication of surface waters occurs in high-input systems. Soil erosion is a problem on newly planted land in many parts of the world. Trash or green harvesting overcomes many of the problems. It is concluded that sugarcane cultivation can substantially contribute to the supply of renewable energy, but that improved crop husbandry and precision farming principles are needed to sustain and improve the resource base on which production depends.

1. Introduction Bioenergy is energy from biofuels. Biofuel is produced directly or indirectly from biomass such as wood, charcoal, bioethanol, biodiesel, biogas (methane), or biohydrogen (FAO, 2006). It is big business. Demand for biofuels is surging because of the rise in crude oil prices and the global search for renewable energy (Valdes, 2007) and global biofuel production tripled between 2000 and 2007. Currently, the most important biofuel crops are corn, rapeseed, soybean, sugarcane, and oil palm whereas suitable trees for bioenergy production include eucalyptus, poplar, and willow. Biofuel production itself needs fossil energy. Currently, agriculture accounts for about 15% of the global energy demands (fertilizers, transport etc.) but it is estimated that agriculture can produce half to several times the current global energy demand (Smeets et al., 2007). The environmental impact of the shift toward growing crops for energy is still to be assessed. It is a complex matter with economic interests and other factors interacting on several scales. For example, the cultivation of biofuel crops is competing with food crops and may drive up commodity prices (UNEP, 2007)—over the last few years, world food prices have increased because of market demand for corn, wheat, and soybean. There

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is further concern that the expansion of biofuel crops takes place at the expense of rainforest and has negative effects on biodiversity and the environment. Sugarcane as a biofuel crop has much expanded in the last decade, yielding anhydrous ethanol (gasoline additive) and hydrated ethanol by fermentation and distillation of sugarcane juice and molasses (Gunkel et al., 2007; Pessoa et al., 2005). By-products are bagasse and vinasse (stillage or dunder), which is the liquid waste sometimes used for fertigation purposes. Bagasse, a by-product of both sugar and ethanol production, can be burned to generate electricity or be used for the production of biodegradable plastic. It provides most of the fuel for steam and electricity for sugar mills in Australia and Brazil. One hectare of sugarcane land with a yield of 82 t ha1 produces about 7000 liter of ethanol. Brazil currently produces about 31% of the global production and it is the largest producer, consumer, and exporter of ethanol for fuel (Andrietta et al., 2007). The industry employs more than one million people (Pessoa et al., 2005). The value of the sugar and ethanol industry reached $8 billion in 2006, some 17% of Brazil’s agricultural output (Valdes, 2007). Between 1990 and 2005, global average sugarcane yields increased from 61 to 65 Mg ha1 (http://faostat.fao.org). In 1990, global production was 1050 million Mg and in 2005, production of sugarcane was 1225 million Mg. Much is grown on large plantations but in some countries sugarcane is grown by smallholders, for example in Thailand where there are more than 100,000 farmers growing sugarcane (Sthiannopkao et al., 2006). In Brazil, less than 20% of the sugarcane is produced on small farms; most is grown in the southeast with over 60% of the production in the Saõ Pula district (FAO, 2004). In some countries, sugarcane is the main source of revenue and in Mauritius, sugarcane occupies 90% of the arable land (Ng Kee Kwong et al., 1999). Globally, the area harvested increased by 2.6 million ha in the period 1990–2005; the largest expansion was in India and Brazil. It is expected that the area under sugarcane in Brazil will expand by 3 million ha over the next 5 years whereas the area under sugarcane in China is forecast to rise by 5% or more than 100,000 ha year1. Brazil has a long tradition of growing sugarcane. In sixteenth century, it was the world’s major supplier of sugar (Courtenay, 1980). In 1975, the area under sugarcane in Brazil was 1.9 million ha (de Resende et al., 2006), now there is about 6.2 million ha under sugarcane in Brazil compared to 21 million ha soybean and 14 million ha corn. Other big sugarcane producers are India (4.2 million ha), China (1.4 million ha), Thailand (1.1 million ha), and Pakistan (0.9 million ha) whereas the sugarcane areas in Australia, Cuba, Indonesia, Mexico, and South Africa cover some 0.5–0.6 million ha in each country. In the United States, there are about 170,000 ha in Louisiana and 167,000 ha in Florida. Against the trend, the area under sugarcane in Hawaii has decreased from

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about 100,000 ha in the 1930s to 6000 ha in 2007, and also the area under sugarcane in Cuba has been more than halved in the last 15 years. Traditionally, sugarcane was harvested manually; the senescent leaves (trash) and stalks were removed by people using big knives. Green harvesting was common in Brazil up to 1940s (de Resende et al., 2006), but the large volume of trash makes manual harvesting difficult (Boddey et al., 2003). As labor shortages developed, it became common practice to burn of the dead leaves prior to the harvest (preharvest burning). In the last two decades, preharvest burning has been replaced by mechanical green- or trash harvesting by cutter-chopper-loader harvesters that leave the trash on the field. Most of the sugarcane in Australia and parts of the West Indies is now arvested like this (Graham et al., 2002a). Up to the 1960s, Australian sugarcane was harvested manually but a decade later, following severe labor shortages, nearly all sugarcane was harvested mechanically (Brennan et al., 1997). Currently, about 30% of the Brazilian sugarcane is greenharvested, the rest is harvested manually with preharvest burning. All sugarcane in the United States is mechanically harvested but over 90% of the fields are burned after the green harvesting, to get rid of the trash blanket. Sugarcane is grown as a ratoon crop: the whole above ground biomass is harvested each year and harvests may continue for a number of years (ratoons). Yields decline with ratooning and, after some years, the land is ploughed and new sugarcane is planted. Much of the world sugarcane is grown with a high degree of mechanization. Also, large amounts of biomass are annually removed with the harvest and herbicides and pesticides are used extensively. Irrigation and large amounts of inorganic fertilizers are often required for high yields. As a consequence, soil properties are likely to change under sugarcane cultivation and the high biocide inputs may affect the environment. Environmental concerns and policies are key factors affecting the future of sugarcane production (Valdes, 2007). There is a also risk that the sugar industry is expanding on marginal lands where the costs or preventing or repairing environmental damage may be high (Arthington et al., 1997). This chapter reviews the main soil and environmental issues under continuous sugarcane cultivation. Most of this work predates the surge of sugarcane production for bioethanol but the results are very relevant for the new situation.

2. Changes in Soil Chemical Properties 2.1. Data sources and types There is fair a body of literature on changes in soil properties under sugarcane cultivation, especially in conference proceedings and books. Increasingly, there have been publications on soil and environmental issues


129

in international scientific journals in English. Changes in soil properties under continuous sugarcane have been investigated in two ways. First, soil properties are monitored over time at the same site and this generates Type I data using chronosequential sampling. There are few such data sets because they require long-term research commitment and detailed recordings of soil management and crop husbandry practices. In the second approach, soils under adjacent different land-use systems are sampled at the same time and it is called biosequential sampling (Tan, 1996) generating Type II data (Sanchez et al., 1985). The assumption is that the soils of the cultivated and uncultivated land are the same and that differences in soil properties can be attributed to differences in land use and management (Hartemink, 2003). A considerable number of studies have focused on soil chemical and physical changes, and there are only few studies that included soil biological changes (Table 1). Several studies have been conducted in Brazil, Australia, and South Africa; although sugarcane is important and extensively grown in many other countries, fewer studies have been reported in the literature. Well-researched soil types are Fluvents, Inceptisols, Alfisols, and Oxisols; less data are available from Vertisols, although they are extensively used for sugarcane (Ahmad, 1983).

2.2. Monitoring over time Few studies have monitored soil chemical properties under continuous sugarcane cultivation. In Fiji, Haplic Acrustox were sampled under native vegetation prior to planting sugarcane, and again 6 years later (Masilaca et al., 1985). Exchangeable K decreased, soil P levels were increased in two of the three topsoils, and in one-third of the Oxisols, the topsoil pH had declined from 5.5 to 4.6 (Table 2). Schroeder et al. (1994) measured soil pH over 5 years on sugarcane farms on soils derived from sedimentary rocks in South Africa. These soils had received about 140 kg N ha1 year1 and pH declined by 0.4 units. Soil pH in the VMC milling district in the Philippines declined from 5.0 to 4.7 over a 19-year period under sugarcane (Alaban et al., 1990). The decline in pH was accompanied by a decrease in organic C from 14 to 10 g kg1; also available P and levels of exchangeable cations decreased (Table 3). In Papua New Guinea, Hartemink (1998a,c) compiled soil data at a plantation on Fluvents and Vertisols. Soil chemical data were available from the early 1980s and early 1990s (Table 4). A significant decrease was found in the pH, available P, and CEC of the Fluvents and even in Vertisols, the pH had decreased significantly. A decrease of 0.2–0.4 pH unit was found to a depth of 0.60 m after 10 years of continuous sugarcane (Table 5).

130 Table 1

Studies focusing on changes in soil chemical, physical, and biological properties under sugarcane cultivation Dataa

Soil property investigated Soil order

Alfisols

Country

Australia

Brazil India Swaziland Andosols Fluvents

USA Hawaii Australia

Brazil Fiji USA Hawaii Iran Mexico Papua New Guinea

Chemical

Physical

Biological

p

p

p

p

p

Type I

p

p

p p

p p

p p

p

p

p

p

p

p

p

p

p p

p

p

p

p

p

p p

Type II

p p

p

p

References

Blair, 2000; Bramley et al., 1996; Pankhurst et al., 2005a,b; Skjemstad et al., 1999 Caron et al., 1996; Tominaga et al., 2002 Sundara and Subramanian, 1990 Henry and Ellis, 1995; Nixon and Simmonds, 2004 Zou and Bashkin, 1998 Bramley et al., 1996; Braunack et al., 1993; Pankhurst et al., 2005a,b; Skjemstad et al., 1999 de Resende et al., 2006 Masilaca et al., 1985 Juang and Uehara, 1971; Trouse and Humbert, 1961 Barzegar et al., 2000 de la F et al., 2006 Hartemink, 1998a,c

Inceptisols

Australia

India

Oxisols

Iran South Africa Brazil

Fiji USA Hawaii South Africa

Spodosols Ultisols

Vertisols

Swaziland Australia USA Australia Brazil Indonesia Mexico

p

p

p

p

p

p

p

p

p

p

p

p

p

p

p

p

p

p

p

p p

p

p

p p

p

p

p

p

p

p

p

p

p

p

p p

p

p p

131

p

p

p

p

p p

p

Bramley et al., 1996; Noble et al., 2003; Pankhurst et al., 2005a,b; Skjemstad et al., 1999 Singh et al., 2007; Srivastava, 2003; Suman et al., 2006 Barzegar et al., 2000 Dominy et al., 2002 Caron et al., 1996; Ceddia et al., 1999; Cerri and Andreux, 1990; de Souza et al., 2005; Nunes et al., 2006; Razafimbelo et al., 2006; Silva et al., 2007 Masilaca et al., 1985 Juang and Uehara, 1971; Trouse and Humbert, 1961 Dominy and Haynes, 2002; Dominy et al., 2002; Haynes et al., 2003 Henry and Ellis, 1995 McGarry et al., 1996a,b Muchovej et al., 2000 Pankhurst et al., 2005a,b Ceddia et al., 1999 Sitompul et al., 2000 Carrillo et al., 2003; de la F et al., 2006 Hartemink, 1998b,c (continued)

Table 1

(continued) Dataa

Soil property investigated Soil order

Country

Papua New Guinea South Africa

Not specified

Zimbabwe Australia

India Mexico Philippines South Africa Trinidad a

Chemical

Physical

Biological

p

p

p

p

p

p

p

p

p p

p

p

p

Type I

p

p

p

p p

Type II

p p

p

p

p

References

Graham and Haynes, 2005, 2006; Graham et al., 2002b Rietz and Haynes, 2003 Garside et al., 1997; King et al., 1953; Maclean, 1975; Magarey et al., 1997; Moody and Aitken, 1995, 1997; Wood, 1985 Srivastava, 1984; Yadav and Singh, 1986 Campos et al., 2007 Alaban et al., 1990 Schroeder et al., 1994; Swinford and Boevey, 1984 Georges et al., 1985

Type I are data whereby soil dynamics are followed with time on the same site; Type II are data whereby different land use was sampled simultaneously [see Hartemink (2006)].

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Table 2 Changes in soil chemical properties at sugarcane plantations in Fiji CEC and exchangeable cations (mmolc kg1)

Sampling Site depth (m) pH

C P N (g kg1) (g kg1) (mg kg1) CEC

A

–26.9 þ3.8 –0.6 –14.2 þ1.3 þ0.5 –17.0 þ7.8 –0.2

B

C

0–12 30–40 70–80 0–12 30–40 70–80 0–12 30–40 70–80

–0.7 –0.8 –0.6 –0.3 þ0.1 –0.1 þ0.2 þ0.1 þ0.1

–2.1 –0.2 0 –2.2 þ0.1 –0.2 –0.3 þ0.3 0

þ62.0 þ3.0 –1.0 –2.0 þ1.0 þ2.0 þ64.0 þ6.0 –4.0

–96.0 þ0.3 –18.0 –38.0 þ5.0 þ7.0 –3.0 þ35.0 þ28.0

Ca

Mg

K

–19.9 þ2.4 þ0.2 –26.9 –3.3 –2.9 –29.6 –1.4 0

–1.3 –0.1 –0.2 –10.6 –1.4 –2.2 þ1.8 –0.4 –0.3

–1.8 –0.1 –0.2 –1.1 –0.2 0 þ0.5 –0.9 –0.2

Soils were Oxisols and had been under sugarcane for 6 years. Type I data, modified from Masilaca et al. (1985).

Table 3 Changes in soil chemical properties on sugarcane plantations in the Philippines Exchangeable cations (mmolc kg1)

Sampling period

pH

Organic C (g kg1)

Available P (mg kg1)

Ca

Mg

K

1969–1970 1988–1989

5.0 4.7

13.3 9.9

27.3 17.3

85.7 47.4

11.6 11.1

3.7 3.4

Type I data, modified from Alaban et al. (1990).

2.3. Samples from different land-use systems One of the longest data sets on soil changes under sugarcane cultivation is from the coastal tableland in Alagoas, Brazil (Silva et al., 2007). Soil samples were taken Oxisols in undisturbed forest and compared with soils that had been under sugarcane for 2, 18, and 25 years. Under forest, soil organic C was about 26 g kg1 in the upper 0.20 m soil layer but had decreased to 19 g C kg1 after 2 years of sugarcane cultivation. After 18 and 25 years, soil organic C levels were similar to those under forest in both topsoil and subsoil. In South Africa, an experiment established in 1939 on a Vertisol at the Experimental Station at Mount Edgecombe, has trash-burned and unburned treatments and with or without inorganic fertilizers. Fertilized plots received 140 kg N ha1, 20 kg P ha1, and 140 kg K ha1. Soil organic matter was lowest when crop residues (trash) were removed and

134

Alfred E. Hartemink

Table 4 Soil chemical properties (0–0.15 m) of Fluvents and Vertisols under sugarcane in the 1980s and 1990s Fluvents (n ¼ 7 pairs)

Vertisols (n ¼ 5 pairs)

Soil chemical properties

1982– 1983

1991– 1994

Difference

1982– 1984

pH H2O (1:2.5 w/v) Available P (mg kg1) CEC (mmolc kg1) Exchangeable Ca (mmolc kg1) Exchangeable Mg (mmolc kg1) Exchangeable K (mmolc kg1)

6.3

5.9

p < 0.001

37.2

29.0

p ¼ 0.04

412

354

p < 0.001

229

213

100 11.0

1991– 1994

Difference

6.4

6.0

p < 0.001

35.4

24.6

ns

450

403

ns

ns

269

250

ns

94

ns

109

95

ns

9.5

ns

10.1

ns

13.0

ns ¼ not significant. Type I data, modified from Hartemink (1998c).

highest when residues were retained and inorganic fertilizers were applied. Soil pH decreased from 5.8 under natural grassland to 5.2 under sugarcane with fertilizer applications and also as a result of the trash retention. In soils where there was no trash or inorganic fertilizers, there was no significant decline in pH. Acidification was accompanied by a decrease in the levels of Ca and Mg (Graham and Haynes, 2005; Graham et al., 2002a). Several studies have been conducted in Australia where Type II data are termed samples from ‘‘paired sites’’ or ‘‘paired sampling,’’ sampling ‘‘old and new soils,’’ comparing ‘‘cropped and undeveloped’’ land, or comparing ‘‘virgin and cultivated’’ soils (Hartemink, 2006). King et al. (1953) compared soil chemical properties of uncultivated soils with those that had been under sugarcane for 22 years in the Bundaberg area. The cultivated soils contained on average 22 g C kg1 whereas the C content of virgin soils was 48 g kg1. In proportion, total N contents of the soils under sugarcane were also less than half of the N contents in virgin soils. Maclean (1975) found significant differences in topsoil pH between sugarcane and uncultivated land and also topsoil P, Ca, and Mg levels were significantly lower in soils under sugarcane. In the subsoil, available P and exchangeable Mg were significantly lower, but below 0.3 m depth, there was no significant difference between soils under sugarcane and uncultivated soils. Wood (1985) sampled cultivated and adjacent uncultivated land at 19 sites in a range of different soil types. The cultivated sites had been cropped with sugarcane for at least 30 years whereas the uncultivated sites were road reserves, cleared

Table 5 Change in pH H2O with depth based on samples from the same site at different times and from the different land use sampled at the same time Type I data

Type II data

Sampling depth (m)

Sample pairs

1986

1996

Difference

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60

9 9 7 7

6.2 6.2 6.5 6.6

5.8 5.9 6.1 6.4

p < 0.001 p < 0.001 p ¼ 0.02 p ¼ 0.01

a Soils were continuously cultivated with sugarcane for at least 10 years ns ¼ not significant. Modified from Hartemink (1998a).

Sampling depth (m)

Sample pairs

Natural grassland

Continuous sugarcanea

Difference

0–0.15 0.15–0.30 0.30–0.50 0.50–0.70 0.70–0.90

5 5 5 5 5

6.3 6.3 6.6 6.7 6.9

5.8 6.1 6.4 6.6 6.8

p ¼ 0.02 p ¼ 0.02 p ¼ 0.05 ns ns

136

Alfred E. Hartemink

Table 6 Changes in soil chemical properties on sugarcane plantations in North Queensland, Australia

Land use

Sugarcane

Sampling depth (m)

0–0.10 0.10–0.20 0.20–0.30 0.30–0.40 Uncultivated 0–0.10 0.10–0.20 0.20–0.30 0.30–0.40

CEC and exchangeable cations (mmolc kg1)

P C pH (g kg1) (mg kg1) CEC Ca

Mg

K

5.0 7.0 4.9 6.5 4.9 5.6 5.0 4.0 5.2 15.0 5.2 8.1 5.1 5.9 5.1 4.9

7.3 5.1 5.6 8.1 14.1 12.3 12.4 15.3

2.0 1.4 1.1 1.0 2.9 1.6 1.3 1.3

35 26 15 9 14 8 7 3

37.0 37.0 39.0 41.3 56.3 47.5 46.8 51.7

15.2 15.5 17.1 18.7 32.8 26.1 23.1 25.0

Average data of various soil types. Sugarcane was cultivated for at least 30 years. Type II data, modified from Wood (1985).

land, or forest. A slightly lower pH was found under sugarcane and differences in soil reaction in the 0.20–0.30 m soil horizon were significant (Table 6). Organic C levels in soils under sugarcane were less than half of the levels in uncultivated soils. Exchangeable cations and the CEC were significantly lower in soils under sugarcane but these soils had significantly higher levels of available P due to high application rates of P fertilizers. Bramley et al. (1996) sampled Dystropepts, Ustropepts, Tropaquepts, Natrustalfs, Haplustalfs, and Fluvents that had been under sugarcane for 20 years or more. Soil fertility decline differed between soil orders and depths. Organic C declined in the Fluvents, but no significant changes were found in the other soils. A significant decline in soil pH was found only in Ustropepts. Skjemstad et al. (1999) investigated the same soils and found little changes in total soil organic C and in the light fraction ( tot

L < tot

p

E>L

E¼L

E tartrate > malate (Hue et al., 1986). The decrease in Al-activity by addition of organic matter has been reported by Kochain (1995). The functional groups involved in metal complexation by organic matter are COOH and OH (Wong and Swift, 2003). Surface application or surface incorporation of organic matter also decreased phytotoxic subsoil Al3þ activities because dissolved organic matter (DOM) that leached into the subsoil formed nontoxic Al–DOM complexes (Hue, 1992; Hue and Licudine, 1999; Liu and Hue, 1996; Willert and Stehouwer, 2003). The combined application of CaCO3 and organic matter in lime-stabilized biosolids decreased subsoil acidity and increased subsoil Ca saturation, compared with CaCO3 alone (Brown et al., 1997; Tan et al., 1985; Tester, 1990; Willert and Stehouwer, 2003). This effect was attributed to increases in Ca mobility caused by Ca–DOM complexes (Willert and Stehouwer, 2003). Additional benefits of organic matter addition to acid soils are improving nutrient cycling and availability to plants through direct additions as well as through modification in soils’ physical and biological properties. A complementary use of organic manures and chemical fertilizers has proven to be the best soil fertility management strategy in the tropics (Fageria and Baligar, 2005a; Makinde and Agboola, 2002). Enhanced soil organic matter increases soil aggregation and water-holding capacity, provides source of nutrients, and reduces P fixation, toxicities of Al and Mn, and leaching of nutrients (Baligar and Fageria, 1999). Build-up of organic matter through additions of crop and animal residues increases the population and species diversity of microorganisms and their associated enzyme activities and respiration rates (Kirchner et al., 1993; Weil et al., 1993). The use of organic compost may result in a soil that has greater capacity to resist the spread of plant pathogenic organisms. The improvement in overall soil quality may produce more vigorous growing and high yielding crops (Brosius et al., 1998).

5.6. Conservation tillage The mechanical manipulation of the soil profile for crop production is known as tillage. Conservation tillage is minimum manipulation of the soil profile for crop production and 30% or more soil surface is covered with crop residues. In conservation tillage, weeds are controlled by herbicides and this practice helps in conservation of water and nutrients, and reduces soil erosion and labor cost. Conservation tillage reduces oxidation

371

Ameliorating Soil Acidity

of organic matter or conserves soil organic matter content and consequently decreases the adverse effects of soil acidity. Additional research is needed to assess the long-term conservation tillage effects on soil acidity.

5.7. Crop species and genotypes within species Crop species and genotypes within species differ significantly in relation to their tolerance to soil acidity (Baligar and Fageria, 1997; Devine, 1976; Fageria and Baligar, 2003b; Fageria et al., 1989, 2004; Foy, 1984; Garvin and Carver, 2003; Reid, 1976; Sanchez and Salinas, 1981; Yang et al., 2000). Hence, lime requirements also vary from species to species and among cultivars within species. Many of the plant species tolerant to acidity have their center of origin in acid soil regions, suggesting that adaptation to soil constraints is part of the evolution processes (Foy, 1984; Sanchez and Salinas, 1981). A typical example of this evolution is the acid soil tolerance of Brazilian upland rice cultivars. In Brazilian Oxisols, upland rice grows very well without liming, when other essential nutrients are supplied in adequate amount and water is not a limiting factor (Fageria, 2000, 2001a). Experimental results obtained on Brazilian Oxisols with upland rice are good examples of crop acidity tolerance evaluation. Fageria et al. (2004) reported that grain yield and yield components of 20 upland rice genotypes were significantly decreased at low soil acidity (limed to pH 6.4) as compared with high soil acidity (without lime, pH 4.5), demonstrating the tolerance of upland rice genotypes to soil acidity. In Table 13, data are presented showing grain yield and panicle number of six upland rice genotypes at two acidity levels. These authors also reported that grain yield gave significant negative correlations with soil pH, Ca saturation, and base saturation. Further, grain Table 13 Grain yield and panicle number of six upland rice genotypes at two soil acidity levels in Brazilian Oxisols Grain yield (g pot1)

Panicle number (pot1)

Genotype

High acidity (pH 4.5)

Low acidity (pH 6.4)

High acidity (pH 4.5)

Low acidity (pH 6.4)

CRO97505 CNAs8983 Primavera Canastra Bonança Carisma Average

74.3 55.2 53.0 51.6 48.8 50.8 66.7

52.0 42.9 47.2 38.9 36.5 17.5 47.0

38.0 29.0 25.0 32.0 26.3 43.3 38.7

28.3 25.7 21.7 26.3 20.7 17.7 28.1

Source: Compiled from Fageria et al. (2004).

372

N. K. Fageria and V. C. Baligar

yield had significant positive correlations with soil Al and H þ Al, confirming that upland rice genotypes were tolerant to soil acidity. Fageria (1989) reported stimulation of growth of Brazilian rice cultivars at 10 mg Al3þ L1 in nutrient solution compared with a control (no Al) treatment. Okada and Fischer (2001) suggested that the mechanism for the genotype difference of upland rice for tolerance to soil acidity is due to the relationship between regulation of cell elongation and legend-bound Ca at the root apoplast. A substantial number of plant species of economic importance are generally regarded as tolerant to acid soil conditions of the tropics (Sanchez and Salinas, 1981). In addition, there are cultivars within crop species that are tolerant to soil acidity (Fageria et al., 2004; Garvin and Carver, 2003; Yang et al., 2004). Yang et al. (2005) reported significant differences among genotypes of rye, triticale, wheat, and buckwheat to Al toxicity. These crop species or cultivars within these species can be planted on tropical acid soils in combination with reduced rates of lime input. Combination of legume–grass pasture and agroforestry system of management are the other important soil acidity management components useful in tropical ecosystems. For example, Pueraria phaseoloides is used as understory for rubber, Gmelina arborea or Dalbergia nigra, plantations in Brazil, presumably supplying nitrogen to the tree crops (Sanchez and Salinas, 1981) A detailed discussion of combination of legume–grass pasture and agroforestry in tropical America is given by Sanchez and Salinas (1981). These authors reported that when an acid-tolerant legume or legume–grass pasture is grown under young tree crops, the soil is better protected, soil erosion is significantly reduced, and nutrient cycling is enhanced. Some important annual food crops, cover or green manure crops pasture species, and plantation crops tolerant to tropical acid soils are listed in Table 14. Acid soil tolerant crops are useful to establish low input management systems.

5.8. Interaction of lime with other nutrients Recognition of the importance of nutrient balance in crop production is an indirect reflection of the contribution of interactions to yield. The highest yields are obtained where nutrients and other growth factors are in a favorable state of balance. As one moves away from this state of balance, nutrient antagonisms are reflected in reduced yields (Fageria et al., 1997). Nutrient interactions can occur at the root surface or within the plant and can be classified into two major categories. In the first category are interactions which occur between ions because the ions are able to form a chemical bond. Interactions in this case are due to formation of precipitates or complexes either in soil or in the plant. For example, this type of interaction occurs where the liming of acid soils decreases the concentration of almost all micronutrients in soil solution except molybdenum. Such reduction in ion concentration in soil solution decreases the uptake. Increasing soil pH

373


Table 14 Some important crop species, pasture species, and plantation crops tolerant to soil acidity in the tropics Annual crop species

Pasture species

Plantation crops

Rice Peanut Cowpea Potato Cassava Pigeon pea Millet Kudzu Mucuna Crotolaria

Brachiaria Andropogon Panicum Digitaria Napiergrass Jaraguagrass Centrosema Stylosanthes

Banana Oil palm Rubber Coconut Cashewnut Coffee Guarana Tea Leucaena Brazilian nut Eucalyptus Papaya

Source: Sanchez and Salinas (1981), Caudle (1991), Fageria et al. (1997), and Brady and Weil (2002).

due to lime will have a more marked effect on Zn than Cu uptake, mainly because Cu is more complexed and protected from precipitation by soluble organic matter (Robson and Pitman, 1983). The second form of interaction is between ions whose chemical properties are sufficiently similar that they compete for site of adsorption (soil components, cell walls), absorption, transport, and function on plant root surfaces or within plant tissues. Such interactions are more common between nutrients of similar size, charge, geometry of coordination, and electronic configuration (Robson and Pitman, 1983). Generally, three types of interactions occur among essential nutrients in plants, and these interactions are known as synergistic, antagonistic, and neutral. If upon addition of two nutrients, an increase in plant growth or yield that is greater than that achieved by adding only one occurs, the interaction is synergistic or positive. Similarly, if adding the two nutrients together produced less plant growth or yield as compared to individual ones, the interaction is negative or antagonistic. When there is no significant change in plant growth or yield with the addition of two nutrients, there is no interaction (Sumner and Farina, 1986). Data in Table 15 show that lime and P fertilization have positive as well as negative interactions depending on crop species. In the case of upland rice and wheat, shoot dry weight yields were higher at zero level of lime compared with 2 and 4 g lime kg1 of soil. Hence, interaction between lime and P in this case was negative. Whereas, shoot dry matter yield of common bean and corn increased with increasing lime as well as P rates, indicating that interaction between lime and P in this case was synergistic. An increasing response to applied P with

374


Table 15 Dry matter yield of shoots of upland rice, wheat, common bean, and corn as affected by lime and phosphorus applied to a Brazilian Oxisols

Lime rate (g kg1)

P rate (mg kg1)

Rice (g pot1)

Wheat (g pot1)

Common bean (g pot1)

Corn (g pot1)

0 0 0 2 2 2 4 4 4

0 50 175 0 50 175 0 50 175

0.72 15.08 18.63 0.73 15.23 13.23 0.33 10.20 13.20

0.17 5.83 8.40 0.20 5.20 6.50 0.20 4.90 5.97

1.25 8.30 10.60 1.30 9.00 10.60 1.70 10.50 12.00

1.10 4.93 8.73 1.43 7.13 11.47 1.10 6.60 9.93

Source: Compiled from Fageria et al. (1995).

increasing rates of added lime has been attributed to either an improved rate of supply of P by the soil or an improved ability of the plant to absorb P when Al toxicity has been eliminated (Friesen et al., 1980b). Liming also improves the microbiological activities of acid soils, which, in turn, can increase dinitrogen fixation by legumes and liberate N and other nutrients from incorporated organic materials (Fageria et al., 1995). Nutrient interaction is also evaluated by studying the influence of increasing nutrient concentrations on the uptake of other nutrients and corresponding plant growth. When plant uptake of a given nutrient decreased and corresponding plant growth also decreased, interaction is negative. However, if plant growth and nutrient concentration increased with increasing nutrient supply, the interaction is positive. Liming of acid soils is mainly to supply Ca and Mg to plants and neutralize phytotoxic levels of soil Al3þ. Hence, interaction of Ca2þ, Mg2þ, and Al3þ with other elements is an important aspect for nutrient interaction discussion. Although absolute Ca requirements for plant growth and metabolism are low, it has great significance in providing balance for levels of other plant nutrients, maintaining membrane integrity, and reducing potential for toxicity of other elements (Wilkinson et al., 2000). Potassium, Al3þ, Mg2þ, Mn2þ, and Hþ and heavy metals can reduce Ca uptake by binding at the exterior surface of plasma membrane which increases the Ca requirement (Marschner, 1995). Fertilization with NO3 generally enhances Ca and Mg concentrations in plants driven by the need for cation–anion balance (Wilkinson et al., 2000). Lime and P interactions are mainly associated with Al toxicity, which limits root growth and proliferation, and nutrient uptake. Aluminum absorbed by roots can also precipitate


375

root-absorbed P and hinder its subsequent translocation to plant tops (Foy, 1983). Mora et al. (2005) reported that the most important Al detoxifying mechanisms in ryegrass were apparently physiological Al-PO4 precipitation inside the root and chemical AlSOþ 4 complex formation in the nutrient solution. The results of solution culture experiments have shown that negative effects of Al on root growth can be reduced by increasing Ca2þ concentration in the growth medium (Nichol et al., 1993; Zysset et al., 1996). Baligar et al. (1992) also reported that Ca ameliorates Al toxicity in wheat plants. Huang and Grunes (1992) reported that adequate supply of Ca and Mg might also ameliorate Al toxicity in wheat plants. Aluminum toxicity continues to be associated with P nutrition of plants (Foy, 1984). Santana and Braga (1977) found that P concentrations in rice tops decreased with increasing Al saturation of the soil. Helyar (1978) concluded that Al toxicity effects were largely associated with Al interference with P metabolism and with Al binding to pectins in root cell wall, stopping root elongation. Fageria and Baligar (1999) studied the interactions between calcium and P, Mg, Zn, Cu, Mn, Fe, and B in common bean and found that uptake of P was quadratically increased, whereas uptake of Mg, Zn, Cu, Mn, Fe, and B significantly decreased as soil Ca increased from 4.9 to 12.5 cmolc kg1 of soil. Calcium Al3þ interactions are important in acid soils (Foy, 1984). Below pH 5.5, Ca2þ Al3þ antagonism is probably the most important factor affecting Ca uptake by plants. Lance and Pearson (1969) reported that reduced Ca uptake was the first externally observed symptom of Al damage on cotton seedlings. Franco and Munns (1982) reported that increasing Ca2þ concentrations from 8 to 80 mg L1 decreased Al toxicity in bean plants. Simpson et al. (1977) attributed poor root growth of alfalfa to Al3þ Ca2þ interaction. Aluminum Feþ3 interactions are frequently reported in the literature (Foy, 1984). Alam (1981) reported Al-induced Fe deficiency in oat and postulated that Al interfered with the reduction of Fe3þ to Fe2þ within the plant, a process essential for normal Fe metabolism and utilization. Clark et al. (1981) found that Fe-deficiency chlorosis was a common symptom of Al toxicity in sorghum.

6. Criteria to Determine Liming Material Quantity Use of adequate lime rate to correct soil acidity and production of maximum yield of a crop species is an important consideration for economical and ecological reasons. Quantity of liming material required is determined on the basis of soil pH, base saturation, and aluminum saturation adjustment at appropriate levels. Appropriate levels of these acidity indices

376


vary with soil type, soil fertility, plant species, and crop genotypes within species (Fageria and Baligar, 2003a). In addition, crop response curves related to lime rate and yield is another criterion that can be used to define lime requirement for any given crop species. Crop response curves to lime levels should be determined for each crop species under different agroecological regions to make liming recommendations effective and economical.

6.1. Soil pH Soil pH or hydrogen ion activity is the most common acidity index used in soil testing program for assessing lime requirements of crops grown on acid soils. Weaver et al. (2004) reported that soil pH buffering capacity, since it varies spatially within crop production fields, may be used to define sampling zones to assess lime requirement, or for modeling changes in soil pH when acid-forming fertilizers or manures are added to a field. The pH is defined as the negative logarithm of the hydrogen ion concentration or activity. It is determined by means of a glass, quinhydrone, or other suitable electrode or indicator at a specified soil to solution ratio in a specified solution, usually distilled water, 0.01 M CaCl2, or 1 M KCl (Soil Science Society of America, 1997). In soil testing laboratories of Brazil, a soil to solution ratio of 1:2.5 is commonly used to determine soil pH (EMBRAPA, 1997). The pH measured in the soil solution represents the active acidity of the soil. Hydrogen and aluminum ions adsorbed by the soil, as well as other soil constituents that generate hydrogen ions, constitute the reserve acidity. The active acidity is neutralized by the addition of lime where more hydrogen ions from the reserve pool go into solution. This results in the resistance of soil to changes in the pH of the soil solution. This is termed ‘‘buffering capacity.’’ The pH of acid soils is lower than 7.0, that of neutral soils is equal to 7.0, and that of alkaline soils is above 7.0. At neutrality, the concentrations of OH ions and Hþ ions are equal, since these ions are derived in equal quantities from the ionization of water. The pH of most agricultural soils is in the range of 4.0–9.0 (Fageria et al., 1990, 1997). Soil acidity is classified into several groups based on soil pH. Slightly acid soils have a pH range of 6.0–7.0, moderately acid soils pH range from 5.5 to 6.0, strongly acid soils pH from 4.5 to 5.0, and extremely acid soils have a pH range below 4.5 (Fageria and Gheyi, 1999). These acidity classifications are arbitrary, and care should be taken when defining adequate pH for crop yields, particularly the extractant used could make a difference. Soil pH in water is higher than that in CaCl2 solution. Soil pH measurements not only indicate the acidity level of a soil but also be used as an initial basis for the prediction of the chemical behavior of soils, particularly in relation to nutrient availability and the presence of toxic elements. Most plant-essential nutrients in soil reach maximal or near-maximal

377


availability in the pH range 6.0–7.0, and decrease both above and below this range (McLean, 1973). Optimal pH values for annual crops cultivated on Brazilian Oxisols are given in Table 16. A curve showing the relationship between lime rate and pH change of a Brazilian Oxisol is presented in Fig. 2. Oxisols with application of lime 0–18 Mg ha1 increase pH significantly and quadratically. Figure 3 shows the relationship between pH of an Oxisol and shoots dry weight and grain yield of common bean. Maximum shoot dry weight and grain yields were obtained at a pH of 6.4. Improvement in grain yield was associated with increasing numbers of pods, grains per pod, and weight of 100 seeds when pH was increased from 5.3 to 6.4 (Fig. 4).

6.2. Base saturation Base saturation is another important chemical property of soils used as a criterion for liming recommendations. Base saturation is defined as the proportion of the CEC occupied by exchangeable bases. It is calculated as follows (Fageria et al., 2007):

Table 16

Optimal soil pH for important crop species grown on Brazilian Oxisols

Crop species

Wheat Common bean Upland rice

Type of experiment

Plant part measured

Soil pH in H2O (1:2.5) Reference

Greenhouse Shoot dry 6.0 weight Field Grain yield 6.6 Field

Grain yield 5.6

Common bean Corn Soybean Upland rice

Field

Grain yield 6.2

Field Field Field

Grain yield 6.4 Grain yield 6.8 Grain yield 5.5

Upland rice Upland rice

Greenhouse Grain yield 5.4 Greenhouse Grain yield 5.9

Corn

Field

Grain yield 6.2

Soybean

Field

Grain yield 6.4

Fageria et al. (1997) Fageria and Stone (2004) Fageria and Baligar (2001) Fageria and Baligar (2001) Fageria (2001b) Fageria (2001b) Fageria and Stone (1999) Fageria (2000) Fageria et al. (1990) Gonzales-Erico et al. (1979) Raij and Quaggio (1997)

378


Shoot dry weight and grain yield (kg ha−1)

Shoot

Grain a

a

6.4

6.8

3000 b 2000

a

a

b 1000

0

5.3

6.4

6.8

5.3

Soil pH in H2O

Figure 3 Relationship between soil pH and shoot dry weight and grain yield of common bean grown on a Brazilian Oxisol (Fageria and Santos, 2005).

b 200

100

4

a c

b Weight of 100 grains (g)

Number of pods (m−2)

a

Number of grains (pod−1)

a

300

3

2

1

a

a

6.4

6.8

b 30

20

10

0 5.3

6.4

6.8

5.3

6.4

6.8

5.3

Soil pH in H2O

Figure 4 Relationship between soil pH and yield components of common bean grown on Brazilian Oxisols (Fageria and Santos, 2005).

P Base saturationð%Þ ¼

ðCa; Mg; K; NaÞ 100 CEC

where CEC is the sum of Ca, Mg, K, Na, H, and Al expressed in cmolc kg1.

379


In Brazil, Naþ is generally not determined because of a very low level of this element in Brazilian Oxisols (Raij, 1991). Hence, Na is not considered in the calculation of CEC or base saturation. For crop production, base saturation levels in soil may be grouped into very low (lower than 25%), low (25–50%), medium (50–75%), and high (>75%) (Fageria and Gheyi, 1999). Very low and low base saturation means a predominance of adsorbed hydrogen and aluminum on the exchange complex. Deficiencies of calcium, magnesium, and potassium are likely to occur in soils with low CEC and very low to low percent base saturation. Quantity of lime required by the base saturation method is calculated by using the following formula (Fageria et al., 1990):

CECðB2 B1 Þ df Lime rateðMg ha Þ ¼ TRNP 1

where CEC: cation exchange capacity or total exchangeable cations (Ca2þ, Mg2þ, Kþ, Hþ þ Al3þ) in cmolc kg1, B2: desired optimum base saturation, B1: existing base saturation, TRNP: total relative neutralizing power of liming material, and df: depth factor, 1 for 20 cm depth and 1.5 for 30 cm depth. For Brazilian Oxisols, the desired optimum base saturation for most of the cereals is in the range of 50–60%, and for legumes it is in the range of 60–70% (Fageria et al., 1990). However, there may be exceptions, like upland rice, which is very tolerant to soil acidity and can produce good yield at base saturation lower than 50%. Specific optimal base saturation values for important annual crops grown on Brazilian Oxisols are given in Table 17. Nature of the soil alters the optimum base saturation required by any given crop species. A relationship between lime rate and base saturation in a Brazilian Oxisol is given in Fig. 5. Bean yield was having significant quadratic response in relation to base saturation (Fig. 6). Maximum yield was obtained with base saturation of 73% at 0–10 cm soil depth, with base saturation of 62% at 10–20 cm soil depth and at 67% base saturation when averaged across two soil depths. Hence, at topsoil layer, higher base saturation was required compared with that at lower soil layer (Fageria, 2008).

6.3. Exchangeable aluminum, calcium, and magnesium levels Aluminum has long been recognized as a toxic element for plant growth (Cronan and Grigal, 1995; Foy, 1984). In soil–plant systems, plant-available Al is determined by soil extraction procedure to predict the risk of Al toxicity and the need for liming (Thomas and Hargrove, 1984). In addition, Ca and Mg contents of the Oxisols are important in determining growth of plants. Hence, exchangeable aluminum, calcium, and magnesium contents

380


Table 17 Optimal base saturation for important annual crops grown on Brazilian Oxisols Type of experiment

Plant part measured

Common bean Common bean Upland rice

Field

Grain yield 60

Field

Grain yield 69

Field

Grain yield 40

Common bean Corn Soybean Upland rice Upland rice Common bean Corn Wheat Soybean Cotton Sugarcane Soybean

Field

Grain yield 70

Field Field Field Field Field

Grain yield Grain yield Grain yield Grain yield Grain yield

59 63 50 30 71

Field Field Field Field Field Field

Grain yield Grain yield Grain yield Grain yield Cane yield Grain yield

60 60 60 60 50 61

Crop species

Base saturation (%)

Reference

Fageria and Santos (2005) Fageria and Stone (2004) Fageria and Baligar (2001) Lopes et al. (1991) Fageria (2001a) Fageria (2001a) Lopes et al. (1991) Sousa et al. (1996) Fageria and Stone (2004) Raij et al. (1985) Lopes et al. (1991) Raij et al. (1985) Raij et al. (1985) Raij et al. (1985) Gallo et al. (1986)

of the soil are taken into account to determine the rate of lime required for a crop grown on an Oxisol. The equation used for lime rate determination is (Fageria et al., 1990; Raij, 1991):

Lime rateðMg ha1 Þ ¼ ð2 Al3þ Þ þ ½2 ðCa2þ þ Mgþ Þ where values of Al3þ, Ca2þ, and Mg2þ are expressed in cmolc kg1. If values of Ca2þ and Mg2þ cations are more than 2 cmolc kg1, only Al multiplied by factor 2 is considered. This criterion was originally suggested by Kamprath (1970) for tropical soils and is still largely used for liming recommendation for Brazilian acid soils (Paula et al., 1987; Raij, 1991; Raij and Quaggio, 1997). Alvarez and Ribeiro (1999) recommended that the factor used to multiply Al should be varied according to soil texture. These authors suggested that in sandy soil with clay content of 0–15%, the factor 0–1 should be used; for medium-texture soils with clay content of 15–35%, a factor 1–2 should be used; for clayey soil with clay content of 35–60%,

381


100

Base saturation (%)

80

60

40 Y = 20.0658 + 8.1171X - 0.2493X 2 R 2 = 0.9635**

20

0

3

6

9

12

18

Lime rate (Mg ha-1)

Figure 5 Relationship between lime rate and base saturation of Brazilian Oxisols.

a factor 2–3 should be used; and for heavy clayey soil with clay content of 60–100%, a factor of 3–4 should be used.

6.4. Aluminum saturation Crops grown in soils with acceptable levels of basic cations do not show Al toxicity symptoms even when the levels of KCl-extractable Al are considered high (Kariuki et al., 2007). Hence, the mere presence of Al in the soil is not an indicator of Al toxicity ( Johnson et al., 1997). A more reliable measure of the potential for Al toxicity is Al saturation (Kariuki et al., 2007). It has been widely reported in the literature that differences in Al tolerance are found among plant species and cultivars within species (Fageria and Baligar, 2003a; Foy, 1992; Kochain, 1995; Okada and Fischer, 2001; Yang et al., 2004). It is evident, therefore, that crop tolerance to Al should be taken into account in estimating the amounts of lime needed to correct Al toxicity. Cochrane et al. (1980) suggested that crop aluminum tolerance should be considered along with levels of exchangeable Ca and Mg, in determination of lime requirement. Aluminum saturation, or the proportion of aluminum among the cations, is calculated by using the following formula:

Al saturationð%Þ ¼

Al 100 ECEC

382


Average of two soil depths 3000

2000

Y = 1248.41 + 56.6337X - 0.4203X 2 R 2 = 0.7736**

1000

0

Grain yield (kg ha-1)

10–20 cm 3000

2000 Y = 1347.17 + 58.6230X - 0.4696X2 R2 = 0.7391**

1000

0 0–10 cm 3000

2000 Y = 1173.82 + 54.2814X - 0.3709X2 R2 = 0.7956**

1000

0

Figure 6

20

40 60 Base saturation (%)

80

Influence of base saturation on grain yield of dry bean (Fageria, 2008).

where ECEC is in cmolc kg1, which is the sum of exchangeable Al3þ, Ca2þ, Mg2þ, and Kþ in cmolc kg1. After determining Al saturation, the following formula is used to calculate lime rate (Cochrane et al., 1980):

3þ Al TASðAl3þ þ Ca2þ þ Mg2þ Þ Lime rateðMg ha Þ ¼ 1:8 100 1

383


where 1 M KCL extracts Al, Ca, and Mg and concentrations are expressed in cmolc kg1 and TAS is target Al saturation, which varied from crop species to species. Critical Al saturation values for important plant species are given in Table 18. These values can be used as a reference guide to calculate the lime rate for different crop species. This approach is very useful where lime is difficult to obtain and rather costly and Al-tolerant cultivars are available.

6.5. Crop responses Methods discussed earlier for lime rate determination provide reference guides for determination of lime rates for correcting soil acidity-related constraints for crops. The best criterion, however, for determining lime rate is actual testing of crop responses to applied lime rates. Crop responses to liming are determined by soil, climate, plant species, and cultivar within Table 18 Critical soil aluminum saturation for important field crops at 90–95% of maximum yield

Crop

Type of soil

Critical Al saturation (%)

Cassava Upland rice Cowpea Cowpea Peanut Peanut Soybean Soybean Soybean Soybean Corn Corn Corn Corn Mungbean Mungbean Coffee Sorghum Common bean Common bean Common bean Cotton

Oxisols/Ultisols Oxisols/Ultisols Oxisols/Ultisols Oxisols Oxisols/Ultisols Oxisols Oxisols Oxisols Oxisols/Ultisols Not given Oxisols Oxisols/Ultisols Oxisols/Ultisols Oxisols Oxisols/Ultisols Oxisols/Ultisols Oxisols/Ultisols Oxisols/Ultisols Oxisols/Ultisols Oxisols/Ultisols Oxisols/Ultisols Not given

80 70 55 42 65 54 19 27 15

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