Advances in agronomy volume 99

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 Mole´culaire 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 ‘Bioge´ochimie 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

ix


x

Contributors

J. M. Meyer (183)
CNRS, Universite´ Louis Pasteur, UMR7156 ‘De´partement Environnement,
Ge´ne´tique mole´culaire 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

xi


C H A P T E R

O N E

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

#

2008 Elsevier Inc.
All rights reserved.

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Philip M. Jardine

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

3

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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Philip M. Jardine

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).


5

Influence of Coupled Processes on Contaminant Fate

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.


6

Philip M. Jardine

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

Influence of Coupled Processes on Contaminant Fate

7

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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Philip M. Jardine

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

Influence of Coupled Processes on Contaminant Fate

9

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 />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


10

Philip M. Jardine

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


Influence of Coupled Processes on Contaminant Fate

11

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 HAsVO42À) 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


12

Philip M. Jardine

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 />

Influence of Coupled Processes on Contaminant Fate

13

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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Philip M. Jardine

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


Influence of Coupled Processes on Contaminant Fate

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

16

Philip M. Jardine

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 S2À 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 />tenji/d_corner/d04.html).

Influence of Coupled Processes on Contaminant Fate

17

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 S2À to complex Hg as HgS. Fortunately, HgS is
highly surface reactive which helps contribute to its lower bioavailability


18

Philip M. Jardine

(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

Influence of Coupled Processes on Contaminant Fate

19

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


20

Philip M. Jardine

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


Influence of Coupled Processes on Contaminant Fate

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.