Report on Bioavailability of

Chemical Wastes With Respect

to the Potential for Soil

Bioremediation

Eugene L. Madsen, Ph.D.

Department of Microbiology, Cornell University

Ithaca, NY
Although the information in this document has been funded by the U.S. Environmental Protection Agency under contract number T28006: QT-DC-99-003260 to

Cornell University, it does not necessarily reflect the views of the Agency and no

official endorsement should be inferred.

This document has been reviewed by a panel of five external experts: Dr. Kevin

Johnson of Southern Illinois University; Dr. Todd Anderson of Texas Tech University; Dr. Katherine Banks of Purdue University; Dr. Teresa Fan of University of

California, Davis; and Dr. James Shine of Harvard University.

Acknowledgments

The author is grateful for the support of the U.S. Environmental Protection

Agency and Dr. Robert Menzer. During preparation of this report, insightful discussions with P. Baveye, M. Alexander, J. Pignatello, S. Hawthorne, W. Ball, J.

Kreitinger, and input from four anonymous reviewers occurred. Remarkable patience and expertise in manuscript preparation was contributed by Patti Durfee.
iii

Contents

Executive Summary .............................................................................................. 1

1 Introduction: Scope and Goals ................................................................... 3

2 Definitions of Biodegradation and Bioremediation, and the Many

Facets of Bioavailability ............................................................................. 5

2.1 Biodegradation ................................................................................... 5

2.2 Bioremediation .................................................................................... 5

2.3 The Susceptibility of Inorganic Materials, Including Metals,

Nonmetals, and Radionuclides, to Microbe-Based Bioremediation....... 6

2.4 Summary ........................................................................................... 10

3 Bioavailability .......................................................................................... 12

3.1. Defining Bioavailability: Nine Definitions ........................................ 12

3.2 Biosensor Technology Offers a Means Towards Direct

Measurement of Bioavailability ......................................................... 19

3.3 Summary ........................................................................................... 20

4 A Survey of Field Projects Using Bioremediation To Treat

Contaminants in Soils ............................................................................. 22

5 Mechanisms of Persistence of Organic Compounds ............................... 31

5.1 An Evolutionary Perspective on the Persistence of

Organic Compounds ......................................................................... 31

5.2 Placing “Bioavailability” Within the Established Framework of

Persistence Mechanisms................................................................... 32

5.3 Illustrating Mechanisms of Persistence: Soil Organic

Matter (SOM) ................................................................................... 34

5.3.1 Summary (Soil Organic Matter Persistence) ............................ 35

6 Paradigms for the Composition and Structure of Soil and the PhysicalChemical State of Contaminants Therein ................................................ 37

6.1 Soil Complexity .................................................................................. 38

6.2 A Thermodynamic Overview of Inorganic Soil Reactions .................. 38
iv

6.3 Models of Soil Structural Characteristics and Methodological

Limitations in Environmental Microbiology ........................................ 39

6.4 Interactions Between Geosorbents and Contaminants........................ 42

6.4.1 Inferences of Geosorbent Structure Based on Indirect

Observations of Hydrophobic Organic Compounds ............... 42

6.4.2. Direct Observations of Contaminated Geosorbents ................ 45

6.4.2.1 Hydrophobic Organic Compounds .......................... 45

6.4.2.2 Inorganic Contaminants........................................... 47

6.5 Summary ........................................................................................... 48

7 Uptake of Soil Constituents by Plants and Microorganisms .................... 50

7.1 Principles of Soil Solution Chemistry and Uptake by Plants of

Inorganic Compounds ....................................................................... 50

7.2 Movement of Solutes From Soil Solution to Roots ............................ 51

7.3 Examples of Nutrient Uptake by Plant Roots....................................... 52

7.3.1. Phosphorus (P) ...................................................................... 53

7.3.2. Iron (Fe) ................................................................................. 53

7.4 Contrasts Between Uptake Mechanisms in Microorganisms

and Plants ......................................................................................... 54

7.5 Membrane Transport in Microorganisms .......................................... 56

7.6 Uptake of Insoluble Organic Substrates by Microorganisms: Wood .. 56

7.7 Phosphorus and Iron Uptake by Bacteria ........................................... 57

7.8 Summary ............................................................................................ 58

8 Reviewing the Facts: Examining Relationships Between Contaminant

Sequestration and Bioremediation ........................................................... 60

8.1 Ambiguity Is the Rule: An Historical Overview of the Impact

of Solid Surfaces on Microbial Activity .............................................. 60

8.2 Selection and Justification of Criteria for Identifying the Highest

Quality Investigations Pertinent to Bioavailability and

Biodegradation .................................................................................. 62

8.2.1 Further Scrutiny of Artifacts That May Be Caused by Soil

Sterilization and Contaminant Addition in Organic Solvents ... 67
v

8.3 Scrutinizing Selected Investigations Describing the Bioavailability

of Contaminants and Their Biodegradation ....................................... 69

8.3.1. Summary ................................................................................ 94

8.4 Influence of Bioavailability on Phytoremediation of MetalContaminated Soils ........................................................................... 95

8.5 A Synthesis: Evaluating the Relationships Between Bioavailability

and Bioremediation Based on Sections 2 to 8.4 of This Report. .......... 96

9 Overcoming Constraints on Site Cleanup ............................................... 99

10 Conclusions, Implications, and Possible Areas of Future Research ....... 101

11 Literature Cited ..................................................................................... 105
vi
1

EXECUTIVE SUMMARY

Based on conservative, reasonably thorough and careful evaluation of scientific

studies described in this report, there is no doubt that chemical wastes in soil can be,

and often are, in a state of reduced bioavailability. An analysis of the literature on

bioremediation research concludes that bioremediation of chemical wastes in soils

and sediments is rarely 100 percent efficient, due at least in part to the reduced bioavailability of the chemical. Reduced bioavailability simply means that a chemical

waste’s diminished “effective concentration” is proportionately balanced by a lingering reservoir of the chemical waste in soil and sediments. This lingering reservoir

remains in the soil habitat regardless of which combinations of conceptual or actual

sequestration mechanisms (e.g., complexation into bound residues, diffusion into

soil pores, NAPL partitioning) apply.

Soil is, by definition, a thermodynamically unstable, kinetically constrained

medium whose chemical composition, (solid, liquid, and gaseous phases) is constantly changing. Thus, the “nonbioavailable” chemical wastes in this lingering

reservoir are always subject to release into soil solution where the wastes are

resubjected to a variety of transport and/or transformation processes (e.g., immobilization, biodegradation, uptake by receptors).

Considerable effort has been expended in investigating the hypothesis that

chemical wastes have diminished bioavailability in soils and sediments. Results of

these efforts have been ambiguous because of the immense diversity in types and

properties of chemical wastes, geosorbents, biota, experimental approaches, and

the idiosyncrasies in mechanisms by which biota interact with chemical wastes. In

this report, 17 studies were selected from the vast literature on bioavailability to

illustrate the range in quality of reported research. Methodologies were analyzed,

and conclusions were critiqued. From this analysis, criteria for identifying the highest quality investigations pertinent to bioavailability are proposed. These are

based on the realism and environmental relevance of the study, the absence of

experimental artifacts, and the consistency of the results.

An accurate estimate of risks to human and environmental health posed by

chemical wastes in soils is a crucial step toward: (1) identifying pragmatic, economically feasible environmental cleanup goals; (2) establishing operational definitions of “treatment” by bioremediation technology; (3) realistically classifying
2

polluted sites based on planned land-use scenarios; (4) developing public acceptance of risk-based contaminant cleanup efforts; (5) developing public acceptance

of cleanup goals that are above the “original, pristine state” of the contaminated

site; and (6) legitimizing the concept of “environmentally acceptable endpoints.”

From a practical, regulatory point of view, establishing the foundation of reduced bioavailability is crucial. If the reduced bioavailability of chemical wastes in

soil becomes widely accepted, then proper quantitative measures of bioavailability

reduction could be developed to accurately estimate the risks posed by chemical

wastes in soils and sediments to human health and ecological processes.
3

Section 1

Introduction: Scope and Goals

Commercial, industrial, and military activity, largely in the 19th and 20th centuries,

have led to environmental contamination problems that can threaten human health

and ecosystem function (Alexander, 1999; Madsen, 1998b; National Research Council, 1997; Young and Cerniglia, 1995). The degree to which such contaminants are

available to biota has far-reaching scientific, toxicological, and regulatory implications (e.g., Lee et al., 2000). Bioremediation, the use of organisms (especially naturally

occurring microorganisms or plants) to eliminate these pollution problems has immense potential for meeting society’s pollution-control needs (Alexander, 1999;

Madsen, 1998b; National Research Council, 1993, 1997; Schwarzenbach et al., 1999;

Van der Lelie et al., 2001;Young and Cerniglia, 1995). However, in recent years, evidence has begun to accumulate suggesting that pollution-eliminating processes may

be thwarted by a variety of physical, chemical, or other spontaneous reactions that

diminish the availability of environmental contaminants for uptake and transformation by biota (Alexander, 1995, 1999).

This report critically evaluates current knowledge of the relationships between

the bioavailability of chemical wastes and their susceptibility to bioremediation.

The context of these relationships is that all terrestrial and aquatic habitats are

located in either the surface or subsurface portions of the landscape. However, due

to this report’s focus, soil habitats will be emphasized. Although model systems

using laboratory incubations of environmental samples or cultures of individual

microorganisms provide insights about these relationships, the primary focus of

this report is the behavior of chemical wastes and naturally occurring (or inoculated) microorganisms in situ –– in real-world contaminated field sites. When

appropriate, information about the nascent field of phytoremediation (plant-mediated decontamination processes) will be included. The chemical wastes of interest

are those that confront society today. These include organic (e.g., petroleum hydrocarbons, pesticides, chlorinated solvents) and inorganic (e.g., metals, radionuclides, oxyanions) compounds. This report will not attempt to comprehensively
4

survey the behavior of all contaminants. Instead, compounds will be selected from

both categories (organic and inorganic) to reveal how microorganisms and plants

interact with chemical wastes. There is no doubt that, under favorable conditions,

many contaminants and microorganisms or plants can interact in ways that successfully lead to biodegradation and bioremediation. The issue addressed by this

report is how bioavailability influences the degree to which the bioremediation

outcome is successful (according to the standards of regulatory agencies and

society).

This report is divided into 10 sections, including this introduction. Section 2

will begin by defining biodegradation and bioremediation; because of their novelty,

fundamental microbial reactions that govern the fate of inorganic materials are

reviewed in detail. Section 3 defines bioavailability. Next, selected case studies of

completed bioremediation field projects are presented—seeking evidence for a

reduction in the efficacy of organic- and inorganic-contaminant bioremediation

that may have been caused by bioavailability limitations. To interpret how and why

bioremediation may be inefficient, Section 5 reviews information about the mechanisms by which organic compounds persist in soils. In Section 6, paradigms for

geosorbents and the physical and chemical state of associated contaminants in

field sites will be summarized. Section 7 reviews physiological principles describing

the uptake of sequestered organic and inorganic chemicals by microorganisms and

plants. Assisted by recent scientific reviews, Section 8 critically evaluates the methodologies that have been used to generate information about the influence of

bioavailability on bioremediation. Section 9 examines studies designed to enhance

the efficiency of remediating organic compounds in soil by alleviating bioavailability

limitations. The report concludes (Section 10) with recommendations about new

frontiers and areas of future research.

This document is heuristic—it strives to synthesize new combinations of information of the highest quality from a variety of sources. The sponsor of this report (the

US EPA) prescribed that it be a concise compilation of current knowledge, not an

encyclopedic review. In compiling and processing information for this report, difficult

decisions were made to include some while excluding other scientific studies. The

author apologizes for any omissions of crucial pertinent scientific results. Throughout this report, selected portions of the text will be highlighted in bold because they

figure prominently in the logic that will lead to this report’s conclusions.
5

Section 2

Definitions of Biodegradation and

Bioremediation, and the Many Facets of

Bioavailability

2.1 Biodegradation

“Biodegradation” is the partial simplification or complete destruction of the

molecular structure of environmental pollutants by complex, genetically regulated

physiological reactions catalyzed largely by microorganisms (Alexander, 1999;

Madsen, 1991; Madsen, 1998b; Young and Cerniglia, 1995); and plants (Bhadra et

al., 1999; Bizily et al., 1999; Burken and Schnoor, 1998; Siciliano and Germida, 1998).

Many of these reactions are predictable based on established laws of thermodynamics applied under environmental geochemical conditions (e.g., temperature,

pressure, and oxidation-reduction potential) that prevail in terrestrial and aquatic

habitats (Madsen, 1998a). Microbial biodegradation is routinely measured by applying chemical and physiological assays to laboratory incubations of flasks containing pure cultures of microorganisms, mixed cultures, or environmental samples

(soil, water, or sediment; Madsen, 2002). Plant-mediated biodegradation and/or

contaminant accumulation also is routinely documented in appropriately scaled

laboratory test systems (Bhadra et al., 1999; Bizily et al., 1999; Van der Lelie et al.,

2001; Vangronsveld and Cunningham, 1998).

2.2 Bioremediation

“Bioremediation” is the intentional use of biodegradation or contaminantaccumulation processes to eliminate environmental pollutants from sites where

they have been released. Bioremediation technologies use the physiological potential of microorganisms and plants, as documented most readily in laboratory

assays, to eliminate or reduce the concentration of environmental pollutants in field
6

sites to levels that are acceptable to site owners and/or regulatory agencies (Burken

and Schnoor, 1998; Madsen, 1998b; National Research Council, 1997; Siciliano and

Germida, 1998). Bioremediation may be approached using in situ technology (applied directly to contaminated sites) or ex situ bioreactor methodologies (after

contaminants and/or accompanying soil, sediment, or water are removed from contaminated sites) (Crawford and Crawford, 1996; Flathman et al., 1994; Hinchee and

Olfenbuttel, 1994; Madsen, 1998b; Norris et al., 1994; Rittman et al., 1994; Skipper

and Turco, 1995; Siciliano and Germida, 1998). The open, complex, poorly defined

settings where bioremediation occurs and the often wide array of possible contaminant attenuation pathways may obscure the precise mechanisms that operate during field bioremediation operations (Madsen, 1991; National Research Council,

1993; National Research Council, 2000).

Phytoremediation is a technology that seeks to use several plant-based processes to remove, transfer, stabilize, or destroy contaminants in soil, sediment, and

groundwater (Cunningham et al., 1997; Ensley, 2000; EPA, 2001; Van der Lelie et al.,

2001; Vangronsveld and Cunningham, 1998). The mechanisms of phytoremediation

include rhizosphere biodegradation, phytoextraction, phytodegradation, and

phytostabilization. Phytoremediation strategies can use long-lived perennial species in a stable system or short-lived annual species that are repeatedly planted and

harvested.

2.3 The Susceptibility of Inorganic Materials, Including

Metals, Nonmetals, and Radionucleotides, to MicrobeBased Bioremediation

Biological processes can influence inorganic environmental contaminants (Babu

et al., 1992; Brierley, 1990; Chapatwala et al., 1995; Hinchee et al., 1995; Kalin et al.,

1991; Lenhard et al., 1995; Lovley, 1993, 1995a, b; Lovley, 2000; McHale and McHale,

1994; Saouter et al., 1995; Summers, 1992; Thompson-Eagle and Frankenberger, 1992;

Videla and Characklis, 1992; and Whitlock, 1990). Unlike organic compounds that are

often susceptible to partial structural alteration or complete detoxification to carbon

dioxide by microorganisms, the majority of inorganic contaminant compounds are

subject only to changes in speciation that may influence contaminant mobility. Microorganisms or plants may cause precipitation, volatilization, sorption, and solubilization of inorganic compounds. Mechanistically, these reactions can be the result of

direct enzymatic processes such as oxidation, reduction, methylation, or uptake.

Reaction mechanisms also can be indirect (nonenzymatic)—resulting from production of metabolites or biomass that can strongly influence the behavior of inorganic

contaminants via redox, acid/base, coprecipitation, sorption, and other geochemical

means. In many phytoremediation scenarios, contaminant-attenuation mechanisms

are driven by transpiration-based hydrodynamic interception of contaminants in soil

solution and groundwater. Direct physiological transformation of inorganic com-
7

pounds by plants is an area of current research, as is plant-microbe interactions in the

rhizosphere. Because the science and technology of microbial processes that influence inorganic compounds are well developed, microbial-, and not plant-based,

bioremediation is emphasized below. However, some microbial reaction types may

apply to plants, and certainly apply to the rhizosphere.

One nearly universal means by which microorganisms have been shown to

reduce concentrations of inorganic contaminants in water (e.g., Cu, Ni, Zn, Cd, Pb)

is by immobilizing aqueous-phase inorganics in microbial biomass and/or microbial

exopolymers (Diels, 1997; Macaskie and Basnakova, 1998). The mechanisms range

from nonspecific electrostatic sorptive interaction between cationic metals and

anionic extracellular polysaccharides (Williams et al., 1998) to highly specific active

transport systems that cause metals to accumulate in high concentrations within

microbial cells (Chen and Wilson, 1997). The utility of these sequestration reactions

for soil is in doubt but they are promising in engineered wastewater treatment

systems where metal-laden water flows over fixed biofilms that can be removed

from the treatment system so that the toxic inorganics can be recovered.

Many inorganic contaminants, especially metals, become relatively soluble or

mobile at low pH. In contrast to the various bioremediation approaches that rely on

immobilization reactions, the opposite (washing soils and sediments free of inorganic

contaminants) can theoretically be achieved by directing low pH waters through

contaminated sites. The acidification step can be mediated by a variety of microbial

processes that include the oxidation of elemental sulfur (Lovley and Coates, 1997).

The often highly abundant nontoxic metals, iron and manganese, exist in reduced and oxidized states. The oxidized states [Fe(III), Mn(IV)] react chemically to

form oxyhydroxide precipitates that serve as physiological electron acceptors for

anaerobic microbial food chains (Lovley, 1995a; 2000). The end products of Fe- and

Mn-reduction [Fe(II) and Mn(II)] are relatively soluble and may migrate to aerobic

habitats, where reoxidation and precipitation also can be catalyzed by microorganisms (Ghiorse, 1994). The behavior of many of the toxic metals discussed below is

intimately tied to the microbially mediated cycling of Fe and Mn because the toxic

metals may be immobilized (through coprecipitation and sorptive reactions with

many Fe and Mn oxides) or solubilized [by being reduced via chemical reactions

with Fe(II) and Mn(II)] (Martinez and McBride, 2001). Thus, most problem inorganic compounds (e.g., Pb, Cr, U, Ni, Hg, Cd, Sr) undergo immobilization reactions

via sorption and precipitation.

Chromium is a metal whose key oxidation states are (VI) and (III) (NRC, 2000). In

aqueous environments, chromium (VI) predominates as the mobile and highly toxic

anions, chromate (CrO4

2-) and dichromate (Cr2

O7

2-). Reduced chromium (III) is less

toxic and less mobile because of precipitation reactions as oxides and hydroxides at

pH 5 and above. A variety of both aerobic and anaerobic microorganisms have been

shown to enzymatically reduce Cr(VI) to Cr(III), but the physiological reason for this
8

ability has not been adequately investigated. Among the hypotheses explaining the

reduction reactions are: survival (i.e., detoxification), cometabolism (i.e., fortuitous

enzymatic reactions), and the use of Cr(VI) as a physiological electron acceptor (to

date, only equivocal evidence for the latter hypothesis has been obtained). Direct

microbial detoxification (reduction) of Cr(VI) is unlikely to be a useful remediation

technology in anaerobic subsurface habitats because the reduction occurs spontaneously in the presence of sulfide, Fe(II), and some organic compounds. Although

microbial production of sulfide, Fe(II), and reduced organic compounds is generally

reliable, additional research is required before judging if Cr(VI) reduction has the

potential to serve as a useful bioremediation tool (NRC, 2000).

Mercury is a toxic metal whose predominant forms include mercuric ion [Hg(II)],

elemental mercury [Hg(0)], and the biomagnification-prone organic mercury compounds, monomethyl- and dimethyl-mercury (NRC, 2000). All transformations of

mercury by microorganisms are considered detoxification reactions that are intended to mobilize mercury away from microbial cells. Most reactions are enzymatic, carried out by both aerobes and anaerobes, and involve uptake of Hg(II)

followed by its reduction to volatile forms (elemental Hg(0), methyl-, and dimethylmercury). Mercuric ion [Hg(II)] also forms highly insoluble precipitates with sulfide; thus, one indirect microbial detoxification strategy involves the stimulation of

sulfate-reducing microorganisms. Although engineered bioreactors that first reduce the mercuric ions and then purge the volatile Hg from water have been designed, no successful applications of this technology have yet been implemented

in soil environments (NRC, 2000).

In addition to mercury, microorganisms are capable of methylating other metals

(Cd, Pb, Sn, Te, Se; Ehrlich, 1996; Se methylation is discussed in detail below). Additional methylation reactions may occur as a result of nonbiological transmethylation

by microbially produced methylated donor compounds such as trimethyl tin (Ehrlich,

1996). These donors may react with ionic forms of Pd, Th, Pt, and Au, but the resultant

reduced metals may not be chemically stable. The significance of these unusual metal

methylation reactions for bioremediation is unknown.

Arsenic is a toxic element capable of existing in five different valence states [As

(-III), (0), (II), (III), and (V)]. Forms of arsenic range from sulfide minerals (e.g.,

As2

S3) to elemental As to arsenic acid to arsenite (AsO2

-

) to arsenate (AsO4

3-) to

various organic forms that include methylated arsenates and trimethyl arsine. Clearly,

the chemical and microbiological reactions of As are complex (Ehrlich, 1996;

Frankenberger and Losi, 1995). Both anionic forms (arsenite and arsenate) are highly

soluble and highly toxic—interfering with various enzyme functions and oxidative

phosphorylation, respectively. All forms of As are toxic. Microorganisms transform

arsenic for several fundamental physiological reasons: (1) under anaerobic conditions, arsenate [As(V)] can be used as a final electron acceptor; (2) under aerobic

conditions, reduced As (e.g., arsenite) oxidation had been shown to generate energy (ATP); and (3) under both anaerobic and aerobic conditions, As can be detoxi-
9

fied by methylation, oxidation, or reduction mechanisms that mobilize As away from

microbial cells. Engineered bioremediation strategies that rely on mobilizing methylated As from water have been implemented (Frankenberger and Losi, 1995).

Although selenium is an important and beneficial micronutrient for plants, animals, humans, and some microorganisms (largely because of its role in some key

amino acids), this element can be toxic at greater than trace concentrations. In

natural environments, selenium has four predominant inorganic species: Se(VI)

(selenate, SeO4

2-), Se(IV) (selenite, SeO32-), Se(0) (elemental selenium), and Se(-II)

(selenide) (Ehrlich, 1996; Lovley, 1995b; Frankenberger and Losi, 1995). Like arsenic, selenium also has many volatile organic forms; these include: dimethyl

selenide, dimethyl diselenide, methane selanone, methane selenol, and dimethyl

selenyl sulfide. Each of these compounds exhibits its own chemical and biochemical behavior, mobility, and toxicity. The various forms of selenium are transformed

by microorganisms according to their physiological needs and the ambient thermodynamic conditions. Reduced inorganic selenium compounds have been shown to

be oxidized under aerobic conditions, though not linked to microbial growth. Oxidized selenium (selenate) can serve as a final electron acceptor for anaerobic microorganisms—resulting in production of selenide and/or elemental Se. Methylation

of the various Se compounds is thought to be a protective detoxification mechanism that mobilizes Se away from microbial cells. Thus like As, the environmental

fate of Se is governed by complex interactions between chemical and physiological

processes. For instance, anaerobic microbial reduction of selenate and selenite to

insoluble elemental selenium represents an important mechanism for immobilizing

and removing Se from aqueous solution. Furthermore, the various volatile methylated forms of Se are sufficiently mobile that aerobic deselenification (largely via

dimethylselenide formation) of highly contaminated California soils has been demonstrated in field experiments (Frankenberger and Losi, 1995).

Oxyanions are water soluble, negatively charged chemical species in which a

central atom is surrounded by oxygen. Nitrate (NO3

-

) is a naturally occurring oxyanion

commonly found at low concentrations in soils, sediments, and both surface and

groundwaters as a result of the biogeochemical cycling of organic matter. Nitrate is

a readily utilizable form of nitrogen that can be assimilated into amino acids by

plants and microorganisms (NRC, 2000). Although serving to supply nitrogen, an

essential nutrient, nitrate also is a serious health concern for at least two reasons:

(1) it can be a chemical or microbiological precursor for nitrite, which spontaneously combines with amino compounds to form highly carcinogenic nitrosamines;

and (2) nitrate can be reduced to nitrite in stomachs of infants, which can cause the

respiratory stress disease, methemaglobinemia.

Nitrate is produced from ammonia by nitrifying microorganisms under aerobic

conditions. The major microbial process that destroys nitrate is dissimilatory reduction to dinitrogen gas (denitrification). Genetic, biochemical, physiological, ecological, environmental, and engineering aspects of denitrification (Zumft, 1997; Atlas and
10

Bartha, 1997) have been examined for decades. Nitrate is used as a physiological

electron acceptor under oxygen-free (anaerobic conditions). The denitrification process is widespread among microorganisms, and occurs reliably in every anaerobic

habitat with abundant carbon and electron sources. Denitrification may not be applicable to surface soil because of the presence of O2

 and the tendency of nitrate to

leach. However, denitrification is a well-established bioremediation process and is

used routinely in sewage treatment plants to curb eutrophication.

The oxyanions chlorate (ClO3

-

) and perchlorate (ClO4

-

) or their precursors (chlorine

dioxide, hypochlorite, and chlorite) are produced by a variety of paper manufacturing,

water disinfection, and both aerospace and defense industries. Although not naturally

occurring, these highly oxidized forms of chlorine are energetically very favorable physiological electron acceptors for microorganisms. Compared to denitrification, knowledge of chlorate and perchlorate biodegradation reactions is quite limited. However,

laboratory studies using both pure bacterial cultures and environmental samples (soil,

freshwater sediments, and sewage) have shown that in the presence of common electron donors (carbohydrates, carboxylic acids, amino acids, even H2

 and H2S), microorganisms can grow at the expense of perchlorate and chlorate; thus reducing them to the

nontoxic chloride ion (Malmqvist et al., 1991; vanGinkel et al., 1995; Rikkon et al., 1996).

Furthermore, a bioreactor has recently been engineered to successfully convert chlorate and perchlorate to chloride (Wallace et al,. 1997).

Uranium is a radioactive element that releases alpha, beta, and gamma radiation that

can be toxic. Uranium can exist in the oxidation states of (0), (III), (IV), (V), and (VI),

though in nature insoluble U(IV)O2 and soluble U(VI)O2

2+ predominate. Recently, U(VI)

has been shown to serve as a terminal electron acceptor for anaerobic microorganisms;

thus, in anaerobic habitats growth-linked reduction (hence immobilization) of U should

be a reliable process (Lovley, 1995a). Although the reverse process, microbial oxidation

of U(IV) to U(VI) under aerobic conditions has been demonstrated, this process has not

been shown to be physiologically beneficial to the responsible microorganisms. Direct

chemical oxidation of U(IV) by molecular oxygen [creating U(VI)] may also influence the

robustness of U bioremediation strategies (NRC, 2000). Another radioactive element,

plutonium (Pu) also has been shown to be susceptible to microbial transformation. Ironreducing microorganisms were found to reduce insoluble Pu4+ to the more soluble Pu3+;

thus, rendering the soluble form more susceptible to mobilization (NRC, 2000). These

microbially mediated oxidation/reduction reactions provide tools for emerging

bioremediation strategies for U, Pu, and other radionuclides.

2.4 Summary

Biodegradation is a complex series of metabolic processes that simplify the molecular structure of organic compounds. Biodegradation is catalyzed largely by microorganisms. Plants also can cause biodegradation reactions, but they are more suited for

uptake and accumulation reactions. Inorganic contaminants (metals, non-metals,
11

oxyanions, and radionuclides) cannot be biodegraded, but their environmental mobility

can be altered through oxidation-reduction, sorption, methylation and precipitation

reactions mediated by microorganisms or plants. Each inorganic pollutant features a

unique set of direct and indirect biotic and abiotic reaction pathways that may be

exploitable by bioremediation technologies. Biodegradation, accumulation, and an altered mobility of contaminants can be rigorously documented in laboratory experiments. These processes are the basis for potential site cleanup technology.
12

Section 3

Bioavailability

3.1 Defining Bioavailability: Nine Definitions

Microorganisms and plants are the catalysts that effect bioremediation. But, as

discussed in Section 1.0, the “bioavailability” of chemical waste in soils, sediments,

and waters may influence, even regulate, bioremediation efficacy. The term

“bioavailability” means many things in many contexts. This section compiles definitions of bioavailability from nine recent authoritative sources. The quotes that

appear below are intentionally extensive to reveal both the many facets of the term

“bioavailability” and the biases of the sources.

The first definition (Southgate et al., 1989) is taken from a 1988 conference,

“Proceedings of Bioavailability 88: Chemical and Biological Aspects of Nutrient

Bioavailability,” organized by the Food Group of Royal Society of Chemistry, The

Working Party on Food Chemistry of the Federation of European Chemical Societies, and the Federation of European Nutritional Societies.

“Until relatively recently, it has been a common assumption that

the nutritional value of foods and diets was more or less synonymous with their nutritional composition as determined by chemical

analysis. The limitations of this simple picture first became obvious

in relation to trace element nutrition, but recently many lines of research have emphasized the need to understand and quantify the

intestinal absorption and subsequent metabolism of all the major

nutrient classes...

Bioavailability has been defined as the proportion of the nutrient that is digested, absorbed, and metabolized through normal pathways... This definition has two important corollaries. First (bio) availability is not a property of the diet of food per se but the response of

the individual to the diet or food. Second, observed (bio)availability,
13

therefore, represents an integration of the processes whereby an

ingested nutrient becomes available...”

This definition was selected because the authors were scholars concerned with

human nutrition—they were not concerned with the scientific and/or political issues of the behavior and fate of environmental pollutants. Yet, the nutritional essence of “bioavailability” establishes themes that will echo throughout this report:

(1) chemical assays of inorganic and organic materials may overestimate what is

actually biologically absorbed and metabolized; and (2) bioavailability is not an

inherent property of substances under examination, rather bioavailability reflects

the response of a biological system to many integrated processes.

The second definition (Alexander, 1999), from a recent textbook on biodegradation and bioremediation, makes the abrupt transition from human nutrition to pollutant chemicals in soils and sediments. Chemically extractable pollutant compounds

are contrasted with biologically available pollutant compounds, as assessed with

biodegradation assays. Five theoretical mechanisms are advanced, which suggest

that reduced bioavailability of organic soil pollutants for microbial biodegradation

is caused by reactions between organic pollutants and the constituents of soils

and sediments.

“The availability of many chemicals is affected by a series of illdefined, often uncharacterized processes. In some of these processes,

the compound is readily evident and can be easily removed from the

soil, sediment, or aquifer by conventional extraction procedures. The

evidence for reduced bioavailability of these compounds is the

marked decline in the rate of biodegradation. In other processes, the

compound is still present, but can only be removed from the environmental sample by highly vigorous extraction techniques. The

evidence for reduced bioavailability of such a compound is the marked

decline in the rate of biodegradation with time or the almost complete

resistance of the molecule to microbial destruction...

At this stage in the development of knowledge of a subject that

is still characterized by confusion, it might be prudent to envision

three separate categories of molecules of reduced bioavailability.

These are in addition to compounds that are poorly available because they are sorbed to solid surfaces or present within NAPLs in

the immediate vicinity of microbial cells having the requisite catabolic enzymes:

(a) Nonsorbed compounds in micropores at some distance from

cells having the requisite enzymes...,

(b) Compounds that either enter into, and are retained within,

nanopores or that partition into the solids themselves..., and
14

(c) Chemicals that complex with humic materials or other environmental constituents to form molecular species that, although

containing the parent molecules or metabolites generated from

them, are in fact new molecular species.”

The third definition (Scow and Johnson, 1997), from a review in the soil sciences

literature, reiterates that several processes act on organic and metallic pollutants in soils

to reduce their toxicological and microbiological availability. An analogy between the

availability of native soil organic material and pollutant (xenobiotic) chemicals is drawn.

“The term bioavailability designates the state of that fraction of

a chemical that is available for uptake and/or transformation by living organisms. Although associated primarily with ecotoxicology,

and usually in reference to organic and metallic pollutants, the term

bioavailability is also relevant to native organic material. Thus, the

“problem” of bioavailability, has existed for microorganisms far longer

than has the presence of xenobiotic chemicals in the environment.

Sorption, insolubility, and related processes are largely responsible for controlling bioavailability of many pollutants to microorganisms in soils and sediments.”

Definition number four (Baveye and Bladon, 1999), derived from a recent international conference on the environmental behavior of organic xenobiotic compounds, evokes the historical, toxicological, and physical-chemical basis for reduced bioavailability. The form of chemicals, the release of chemicals from subsurface environments, and the proportion reaching organisms are emphasized.

“The concept of biological availability was apparently first proposed in 1975 at a National Science Foundation workshop on ecosystem processes and organic contaminants. and was originally

based mainly on physical chemistry... Since then, physical chemists

and biologists have developed independent interpretations of the

concept. Traditionally, chemists have defined bioavailability in terms

of the chemical form in which the compound or element of interest

occurs at a given time. Alternatively, definitions derived by biologists have assumed that the chemical form in the bulk phase is relevant only to the presence of a biological receptor; thus, they have

defined bioavailability based on the portion of the compound that

could pass into an organism under a given set of conditions...

At this juncture, it is not necessary for the sake of our argument

to choose between the approach to bioavailability that focuses solely

on the amounts of xenobiotics that cross the organism’s external

boundary...or the approach that relates these absorbed amounts to

the capacity of the organism’s environment to supply/release the
15

xenobiotics... Both perspectives have merits, which have to be assessed on the basis of the measurements that one can perform on

organisms in subsurface materials.”

Definition number five (Maier, 2000) was prepared for the International Society

for Biotechnology. Emphasizing microbial metabolism of organic environmental

pollutants, this definition examines the relationships between “perceived” contaminant concentrations and details of microbial metabolism such as nutritional

status, enzyme-induction, energy expenditure, and growth. Three causes for reduced bioavailability are mentioned.

“Bioavailability can be defined as the amount of contaminant

present that can be readily taken up by microorganisms and degraded. One reason bioavailability as defined here is so important is

that it governs the rates of biodegradation. Bioavailability controls

biodegradation because microbial cells must expend energy to induce catabolic gene systems used in biodegradation and if the perceived contaminant concentration is too low, induction will not occur. Indigenous soil microbial populations are typically slow growing organisms often exposed to nutrient poor environments... The

presence of contaminants can have a significant impact on the metabolic status of such cells. Three cases can be envisioned that would

result from differing bioavailability of contaminants. In the first case,

biodegradation will not occur because the amount of bioavailable

contaminant is insufficient to justify the energy expenditure to induce biodegradation. In the second case, at low bioavailable concentrations, microbial cells may biodegrade contaminants but in a

resting stage rather than a growing stage. While in this case, biodegradation will occur, it will be limited because new cells are not being

produced. In the third case, there is enough bioavailable contaminant to induce biodegradation in a growing stage. It is this case that

will allow for optimal rates of remediation.

There are several constraints that can limit the bioavailability of

organic compounds in the environment. These are low aqueous solubility, sorption, and micropore exclusion.”

The sixth definition (Bosma et al., 1997) is from a research article focusing on how

mass transfer in soil and aquifers limits the bioavailability of organic compounds.

These authors indicate that reduced bioavailability is a likely explanation for inefficient biodegradation of contaminants in “old polluted sites.” Three causative mechanisms for reduced bioavail-ability are suggested.

“Particularly in old polluted sites, part of the contaminants appear to be inaccessible for biodegradation...these observations indicated a reduced availability of pollutants in soils and sediments
16

contaminated for a prolonged period of time, pollutant—and not

nutrient—availability being the obvious cause. The decrease of the

bioavailability in the course of time is often referred to as “aging” or

“weathering.” It may result from (i) chemical oxidation reactions incorporating them into natural organic matter... (ii) slow diffusion

into very small pores and absorption into organic matter, or (iii) the

formation of semi-rigid films around non-aqueous-phase liquids

(NAPL) with a high resistance toward NAPL-water mass transfer.”

Definition number seven (Alexander, 1997) is derived from a text addressing the

topic of managing contaminated soils. The author points towards limitations in the

use of biodegradation as an assay for the bioavailability of organic compounds in

soils because assimilation, toxicity, and biodegradation are distinctive biological

processes.

“A note of caution is required in regard to the term bioavailability.

It is sometimes considered as synonymous with toxicity to one or

another species, sometimes as equivalent to biodegradation by microorganisms, and sometimes as synonymous with uptake or assimilation. A compound may be assimilated and, although toxic, may

not cause injury because it is not transported to the tissue, cell or

intracellular site where the toxicity can be expressed. A chemical may

be taken up into microbial cells but still not be biodegraded because

that organism does not contain the requisite catabolic enzymes.

Uptake or assimilation is thus a better means of assessing bioavailability, but because of the few studies of uptake per se and the many

more of toxicity and biodegradation, the term bioavailability also will

be used here to include toxicity and biodegradation.”

A recent “state-of-the-science” summary document, addressing the toxicological and environmental impacts of organics, explosives, and metals in soils [the

eighth definition (Loehr et al., 1997)], stressed that bioavailability is a measure that

is dependent on “the receptor, the route of entry, time of exposure, and the matrix

containing the chemical.” This comprehensive definition attributes diminished

bioavailability of contaminants to “physicochemical processes that lead to sequestration.” Three categories of bioavailability are suggested. Furthermore, this definition alludes to bioavailability’s potential impact on endpoints achieved by remediation

efforts.

“Bioavailability is a measure of the potential of a chemical for entry

into ecological or human receptors. It is specific to the receptor, the

route of entry, time of exposure, and the matrix containing the chemical.

It is increasingly recognized that the response of an ecosystem

or any at-risk population is not controlled by the total concentration

of a chemical in the media in which a receptor resides, but instead is
17

controlled by only that portion which is biologically available. Thus,

the definition of “how clean is clean?” or, alternatively, what adverse

effects may be exhibited by exposure to chemicals in soils, is largely

determined by the physicochemical processes that lead to sequestration and the biological processes that may lead to chemical release and accumulation in an ecosystem. In order to set realistic soil

or sediment quality limits for regulatory purposes or to establish

endpoints for remediation processes, the physical, chemical and biological mechanisms that govern chemical release and biological uptake from soils and sediments must be more fully understood.

Aspects of bioavailability include:

• Physicochemical availability—the fraction of a chemical that is

not sequestered and rendered inactive in a solid or other stable

phase;

• Direct bioavailability—all or part of the non-sequestered chemical that is directly available for entry into specific receptors, depending upon the route and duration of exposure; and

• Organic-induced bioavailability—the fraction of a chemical,

perhaps initially considered to be sequestered, that is available

after processing of the soil/sediment by living organisms.

Ultimately, bioavailability is defined by field conditions and all

tests and surrogates for bioavailability must be validated by field

measurements.”

The final definition (Linz and Nakles, 1997) is taken from a document apparently

aimed at using scientific arguments about the behavior of contaminants and their

bioavailability in soil to influence environmental regulatory policy. Contaminant

release from soil and specific exposure routes for receptors are emphasized. This

lengthy excerpt illustrates that it might be possible to extend agricultural, pharmaceutical, and waste stabilization policies toward establishing “environmentally acceptable” concentrations of contaminants in soil.

“The definition of “bioavailability” was also highlighted as important. As defined in this report, the “availability” of contaminants in soils

has two components: (1) the rate and extent to which the contaminant is

released from the soil into the surrounding groundwater and soil vapor

and (2) the rate and extent to which the contaminant is assimilated by

ecological and human receptors following dermal contact, ingestion, or

inhalation (i.e., bioavailability). The work group stressed the importance

of defining “bioavailability” in terms of the specific receptors of interest

(e.g., microorganisms, humans, animals) and the likely exposure routes
18

of importance to these receptors (e.g., ingestion of water or soil or dermal contact with water or soil). This matrix of receptors and exposure

routes identifies the biological systems for which the mechanisms of

bioavailability should be investigated...

The work group noted that the concept of environmentally acceptable endpoints currently exists within the existing regulatory framework

as evidenced by the use of TCLP (Toxicity Characteristic Leaching Procedure) as a means to determine the hazardous classification of wastes.

This procedure classifies the potential hazard of a waste based on the

fraction of the contaminants that leach from it and not on the total

contaminant concentration that is present. Furthermore, the US EPA has

accepted waste stabilization and solidification as acceptable technologies for site remediation. This acceptance required the agency to acknowledge that the relative mobility of the contaminants in the waste

(without treatment) was reduced following the application of the technologies and that this reduction was permanent over time. These concepts are consistent with this document, namely that hydrocarbons

remaining in contaminated soil following bioremediation are not mobile.

In effect, the contaminated material has been “biostabilized.”

Agriculture and other industries integrate information associated

with the presence of chemicals and their toxicological and environmental impact:

• Agriculture. The methodologies for evaluating the delivery and

fate of nutrients in soil as well as the fate and effects of pesticides in soil have been formally documented (e.g., pesticides are

regulated by the US EPA under FIFRA [Federal Insecticide, Fungicide, and Rodenticide Act]). These methodologies and evaluation techniques may be transferable to the determination of environmentally acceptable concentrations of contaminants in soil.

• Pharmaceutical Industry. The development of pharmaceutical

products encompasses many of the issues that must be addressed for contaminants in soil. For example, the development

of slow-release capsules required an understanding of the rate

of release of the chemicals to the receptor and the response of

the receptor to that dose of chemicals. These same issues must

be addressed to make the argument that the release of contaminants from soil that are ingested, inhaled, or contacted dermally

by humans may be acceptable, even if the concentration of the

contaminants in soil is not zero. The methodology that has been

followed by the pharmaceutical industry in this area may be

applicable to this environmental issue.”
19

3.2. Biosensor Technology Offers a Means Towards Direct

Measurement of Bioavailability

Recent biotechnological developments have opened the possibility of directly

measuring the bioavailable fraction of chemicals in environmental samples. These

developments and their implications are briefly described below.

Many types of microorganisms have evolved networks of enzymes that contribute to metabolic pathways for the transformation of both organic and inorganic

toxic contaminant chemicals. Regulation of genes coding for such metabolic pathways is essential for survival and efficient metabolic function of the host microorganism. In a typical genetic regulatory system, gene expression (transcription of

mRNA) is activated when the proper inducer compound is sensed in the environment of the microorganism. This type of transcriptional control is achieved through

the interaction of a transcriptional activator protein with an inducer compound

(often the toxic chemical itself), as well as with RNA polymerase and the DNA that

codes for the appropriate detoxifying proteins.

During the past decade, several groups of investigators have used genetic

engineering techniques to combine the genes involved in sensing environmental

contaminants with reporter genes, such as the lux operon, that triggers a lightproducing bacterial luciferase reaction. Thus, biosensors of this type can be engineered by placing reporter genes under the control of transcriptional activators

(along with their corresponding promoters; Willardson et al., 1998). Under appropriate conditions, a direct correlation between the concentration of a contaminant

chemical and reporter enzyme activity (e.g., light) can be established. These

biosensors have been developed for detection of organic (e.g., Heitzer et al., 1994;

King et al., 1991; Willardson et al., 1998) and inorganic (e.g., Corbisier et al., 1999;

McGrath et al., 1999) chemicals in environmental samples.

Because whole bacterial cells are often the detection system in the biosensors,

they, by definition, measure the amount of contaminant that is bioavailable. Thus,

to some degree these biosensor techniques provide a means of direct detection of

bioavailable environmental contaminants. Biosensors offer the potential of being

free of the pitfalls characteristic of chemical extraction and analytical procedures.

However, despite possible advantages, biosensors also have limitations. These

include: (1) responses specific to the membrane structure and uptake system of the

genetically engineered microorganisms (usually E. coli). The diverse, uncharacterized

microorganisms actually active in soil may differ from E. coli in their membrane

permeability traits; thus, the biosensor may not be a valid surrogate; (2) the biochemistry and specificity of sensor reactions may not be sufficiently understood to

contend with false positive or false negative signals that may occur in complex

geochemical settings; and (3) the physical location of biosensor microorganisms is

unlikely to realistically mimic that of the native soil microorganisms. Thus, much
20

progress will likely be required before biosensor measurements will supercede the

indirect bioavailability measures discussed above.

3.3 Summary

Bioavailability is a concept that has recently been borrowed from the discipline

of nutrition and applied to studies of pollutant compounds. Bioavailability is used,

conceptually, to interpret laboratory- and field-based measurements of the behavior of environmental pollutants in the presence of biota (microorganisms and plants)

in soils, sediment, or waters. Contrary to many implicit and explicit views (e.g.,

definition number 8, above, and the promise of biosensors), bioavailability of chemical

wastes to soil organisms cannot be directly measured. Instead, bioavailability is an

emergent trait, a malleable trait. Information about bioavailability emerges from the

context of real-world or laboratory experimental systems containing biota, contaminant compounds, and a geochemical matrix (soil or sediment or water). It is the

specific, detailed, three-way interactions between system components that allow

inferences about bioavailability to be drawn. Figure 1 provides a graphical conceptual summary of this definition of bioavailability.

 Figure 1 displays directly measurable system components. Thermodynamics (large

circle) provides a stage for potential reactions that could occur between contaminants

(small circle), biota (oval), and soils or sediments (cube). The success of biodegradation

processes, hence bioremediation efforts (one-way arrow, down), depends on specific

interactions (two-way arrows) between system components. Inferences about bioavailability can be drawn only after system behavior and output (biodegradation/

bioremediation) have been observed. Because information about bioavailability is determined by the experimental conditions, the environmental relevance of experimental

conditions used in bioavailability assays must be carefully scrutinized.

Based on the nine definitions listed earlier, there are six hypothesized mechanisms that reduce the availability of contaminant compounds for biotic uptake and

metabolism in soils, sediments, and waters. These mechanisms, which stem from

interactions between the contaminants and the soils or sediments are as follows:

(i) sorption reactions that bind contaminants to solid phase natural surfaces;

(ii) partitioning reactions that place contaminants in non-aqueous phase liquids

(NAPLs)—including NAPLs with “semi-rigid” surface films;

(iii) spatial separation of contaminants in micro- or nano-pores;

(iv) complexation reactions that create new covalent bonds between contaminants

and humic substances;

(v) insolubility of the contaminants; and
21

(vi) partitioning reactions that place contaminants in natural organic matter.

A major goal of this report is to scrutinize these six (and related) hypotheses

(see especially, Sections 5, 6, and 8).

FIGURE 1 Conceptual representation of the relationships between chemical wastes,

biota, and soils, or sediments. The bioremediation outcome is governed by thermodynamic instabilities of unique sets of contaminants in unique combinations with

biota and soil or sediment. The outcome allows inferences about bioavailability to

be drawn (see text, Section 3.3, for explanation).
22

Section 4

A Survey of Field Projects Using

Biodegradation To Treat

Contaminants in Soils

This section is designed to document the efficacy of bioremediation field projects

and to pose the question: “Does availability of chemical wastes constrain bioremediation

as a site cleanup technology?” A comprehensive summary of the performance of

bioremediation in field projects is beyond the scope of the report. A sampling of representative projects is provided in Table 1.1

The first five entries in Table 1 describe completed microbial bioremediation

projects in which soil was contaminated by mixtures of petroleum-related materials.

In some cases, the remediation was implemented on freshly spilled materials (entries 3 and 4). In the other cases, the petroleum was left in contact with soil for many

years (precise duration uncertain) before the active cleanup efforts. The cleanup

endpoints achieved from these remediation activities (800; 700; <50; 10; 100; and

5,409 ppm, respectively, for entries 1-5 of Table 1) were highly variable. One can

make reasonable guesses about causes for the variability. These include: concentrations initially loaded, climatic factors, soil characteristics, degree of mixing, variable microbiological populations, and limited availability of contaminants and nutrients. Distinguishing between these many possible causes is an overarching

theme of this report.

1

A note on sources. Information in Table 1 was not easy to obtain. Unlike peer-reviewed

literature published largely by academicians, the practical results of applied bioremediation

technology are largely reserved for reports in the “grey literature,” where experimental

details and rigorous treatment controls are sometimes omitted. Furthermore, the authors of

“grey literature” bioremediation reports may be biased toward favoring the results of their

efforts both to please their clients (owners of contaminated sites), and to support claims of

the effectiveness of bioremediation technology. Quality of data and experimental design

were two of the criteria for selecting entries in Table 1. Another criterion was field realism.

Although academic studies may feature high-quality scientific procedures, these may be

applied to unrealistic “model systems” whose relevance for the field behavior of contaminant

compounds may be suspect.
23

Entry #6 in Table 1 reports that phenol-contaminated sediment (phenol is a

water soluble, readily biodegradable compound) was cleaned up by microorganisms to below 1 ppm; while tars failed to fall below 183 ppm after 71 days.

Studies examining microbial metabolism of pentachlorophenol (PCP)-contaminated

soils also appear in Table 1. Entry #7 reported a PCP decline from 416 to 150 ppm after 13

weeks of land farming. In the land-treatment system described in entry #8, PCP declined

from 55 to 7 ppm after a nearly 8-year mixture of active and passive treatment.

The next four microbial-related entries in Table 1 (numbers 8-11) describe efforts that attempted to use microbial processes to eliminate polycyclic aromatic

hydrocarbons (PAHs) from soils or sediments. One of these studies (Entry #9)

revealed that PAHs in coal coking waste were very difficult to analyze: despite

mixed aeration in a lagoon, sample variability prevented any trends in biodegradation from being discerned. In the remaining three studies (Entries 8, 10, 11) concentrations of PAHs declined substantially, though residual concentrations were approximately 200 and 159 ppm (for Entries 8 and 11, respectively).

Phytoremediation technology is relatively new to bioremediation. Consequently,

documented full-scale field studies are still relatively rare. The majority of reports

published to date (e.g., Huang et al., 1997; see publications accompanying

Cunningham et al., 1997; Ensley, 2000; Van der Lelie et al., 2001; Vangronsveld and

Cunningham, 1998) focus primarily on the ability of plants to accumulate contaminants; whereas case studies on the performance of the technology in significantly

reducing contaminant concentrations in field soils are quite difficult to identify. The

next three entries in Table 1 (numbers 12-14) are among the few field-scale phytoremediation efforts reported to date. Entry #12 reported that total and leachable

lead in a silt loam soil decreased somewhat after treatment with Indian mustard and

sunflower plants (478 ppm of Pb remained). Entry #13 provides a record of how a

planting of corn and white mustard withdrew some metals from soil (based on

tissue analyses) but project performance, in terms of reduction of metals in bulk

soil, was not provided. Entry #14 of Table 1 notes that a commercial effort using

plants to reduce petroleum hydrocarbon concentrations below a regulatory threshold of 500 ppm was successful.

The final two entries in Table 1 (numbers 15 and 16) are among the many studies

devoted to clean up of chlorinated solvent-contaminated groundwaters. The microbiological basis for such bioremediation approaches is either aerobic

cometabolism of TCE (to short-lived compounds) or anaerobic reductive dechlorination of PCE, TCE, and daughter products (DCE and VC) to nontoxic ethene (NRC,

2000). Because chlorinated solvents are very volatile (and surface soils are easily

excavated) bioremediation technology for chlorinated solvents is usually focused

on groundwater (deep) habitats. Entry #15 displays results of a recent engineered

bioremediation test over a 12 m distance. Steffen et al.’s goal was to prove that

injection of an aerobic bacterium could effectively metabolize TCE and related
24

Entry Habitat/ Age of Contaminant Treatment Degree of Author’s Reference

Location Spill Cleanup Interpretation

1 Soil in land farm PD Petroleum Crushing, sieving, PAH concentra- cited in

(diesel fuel, mixing for 1 year tion declined (Angehrn

heating oil) in a lined land from 10,000 et al., 1998)

treatment facility to 800 ppm

2 Soil adjacent to PD Bunker fuel oil Land farming Total petroleum Successful (Compeau

petroleum tank (plowing and hydrocarbons project; 150,000 et al., 1991)farm irrigation) for declined steadily cubic yards of

14 weeks from 5,000 to soil were treated

700 ppm

3 Soil impacted Fresh Diesel fuel Land treatment TPH Complete (Troy et al.,

by New Jersey (plowing and concentration remediation 1993)

fuel spill irrigation) for dropped from achieved

147 days >700 to <.50 ppm

TABLE 1 Survey of Endpoints Achieved by Field Bioremediation Projects
25

TABLE 1 Survey of Endpoints Achieved by Field Bioremediation Projects (Continued)

Entry Habitat/ Age of Contaminant Treatment Degree of Author’s Reference

Location Spill Cleanup Interpretation

4 Soil surrounding a Fresh No. 6 fuel oil Active TPH Biodegradation (Fogel, 1993)

fuel oil storage treatment: in situ concentration during active

tank mixing, plowing, declined from treatment was

aeration for 6 60,000 to more than 6 timesmonths. 23,700 ppm. more rapid thanSubsequent TPH during passivepassive treatment concenration treatmentfor 27 months declined from

23,700 to10,100 ppm

5 Soil in Southern PD Petroleum Land treatment: TPH declined Site closure (Jerger et al.,

California hydrocarbons tilling, plowing from 13,860 to successfully 1993)

5,409 ppm in 14 achievedmonths

6 Sediment in PD Phenol, tars Lined land Phenol All materials (Portier and

surface farming unit: concentration passed screening Christiansen,

impoundment, mixing, aeration declined from criteria for closure 1993)

Plaque mine, LA for 71 days 137 to <1 ppm.

Tar concentrationdeclined from1,455 to 183ppm
26

Entry Habitat/ Age of Contaminant Treatment Degree of Author’s Reference

Location Spill Cleanup Interpretation

7 Soil in 60 years Pentachloro– Land farming PCP Rate declined (Compeau et

Minnesota phenol (PCP) (plowing and concentration with al., 1991)wood treatment irrigation) for declined from concentration;plant 13 weeks 416 to 150 ppm 17,000 cubic

yards of soilwere treated

8 Soil in creosote- PD PAHs, PCP Active land Total PAH Engineered land cited in:

impacted wood treatment: mix- concentration treatment was (Loehr andtreating site, ing, aeration for declined from effective in Webster, 1997)

Northeastern US 1 year 3,000 to 900 eliminating and

ppm. stabilizingPCP declined contaminantsfrom 55 to 16ppm.PAHs declinedfrom ~900 to~200 ppm.

TABLE 1 Survey of Endpoints Achieved by Field Bioremediation Projects (Continued)
27

TABLE 1 Survey of Endpoints Achieved by Field Bioremediation Projects (Continued)

Entry Habitat/ Age of Contaminant Treatment Degree of Author’s Reference

Location Spill Cleanup Interpretation

9 Sediment in PD PAHs in coal Mixing, Total PAH Much (Leavitt et

mixed tank coking waste aeration for 5 concentrations analytical al.,1991)systems lagoon months began in the variability;

25 to 50 ppm conclusionsrange and ended evasiveat 25 to 500ppm range

10 Soil in Geelong, PD PAHs Land farming Initial PAH All PAHs below (Connolly,

Vicotria, Australia (mixing and concentration 2 ppm were 1999)

aeration) for was between 0.8 persistent and18 months and 174 ppm; considered not

average PAH bioavailable.

mass declined Bioremediation

by >1/3 was limited by:

low bioavailability,low solubility, lowconcentration oflow MW PAHs,desiccation,elevated pH, anduneven nutrientdistribution
28

Entry Habitat/ Age of Contaminant Treatment Degree of Author’s Reference

Location Spill Cleanup Interpretation

11 Soil and PD PAHs Tillage, nutrient PAHs declined cited in:

creosote-rich amendment, from 8,340 to (Loehr and

sludge in a wood irrigation for 159 ppm Webster, 1997)

treating site, 55 monthsWestern MT

12 Poorly drained silt PD Lead (Pb) Indian mustard Total average Pb Plants Federal

loam soil. Ensign- released during and sunflower declined from phytostabilized RemediationBickford open burn/open were planted in a 635 ppm (4/98) the site. Further TechnologyCompany, detonation 1.5 acre area to 478 ppm treatment is Round Table

Simsbury, CT activities (10/98); leaching planned (frtr) Web Site.

assay for Pb Blaylock (1997)showed a declineof 0.95 ppm

13 Twin Cities Army PD Heavy metals, Corn and white Plant yields and During first year, Federal

Ammunition antimony, mustard were metal tissue results were less Remediation

plant, Arden Hills, arsenic, planted to 0.2 acre content were than anticipated Technology

M N beryllium, lead, plots. Fertilizers reported; Round Table

thallium and EDTA were changes in soil (frtr) Web Site

added to the soil concenterations

were notreported

TABLE 1 Survey of Endpoints Achieved by Field Bioremediation Projects (Continued)
29

Entry Habitat/ Age of Contaminant Treatment Degree of Author’s Reference

Location Spill Cleanup Interpretation

14 Former Chevron PD Petroleum 0.6 acre site was Contaminant Site closure was US EPA

Bulk Petroleum hydrocarbons tilled, fertilized, concentration obtained after first Remediation and

processing plant, (TPHs) and seeded. Plant range in top 2 feet growing season Characterization

Astoria, OR species of soil dropped Innovative

unspecified from 170- Technologies

1,000 ppm to (REACHIT)210-450 ppm Web Site

15 Shallow sand/clay PD VOCs: TCE, 305 and 243 Average Using several Steffen et

aquifer in dichloroethane, liters of a concentrations evaluation al.,1999

Pennsauken, NJ vinyl chloride biodegrading fell from 2.2 ppm parameters,

aerobic bacterium to between 0.5 removal ranged

were injected into and 0.05 ppm in from 44 to 78 %groundwater; the test areacontaminant losswas measured inboth test andcontrol areas

16 Deep sand aquifer PD Tetrachloro- Groundwater DNAPL 40 % degradation Lollar et al.,

at Dover Air Force ethene, samples were replenishment of of TCE in plume 2001

Base, Dover, DE Trichloroethene analyzed for contaminants compared to

stable isotope prevented full source area

fractionation to quantification ofdocument effortsreductivedechlorination ofPCE and TCE

TABLE 1 Survey of Endpoints Achieved by Field Bioremediation Projects (Continued)

Abbreviations: PD - age uncertain; probably decades old; TPH - Total Petroleum Hydrocarbons; PCP - pentachlorophenol; PCE - perchloroethene; TCE - trichloroethene; DCE

- dichloroethene; VOC - volatile organochlorides; VC - vinyl chloride; DNAPL - dense nonaqueous phase liquid; PAH - Polycyclic Aromatic Hydrocarbon
30

chloroethenes. The bacterium moved out from the injection well, and average concentrations of volatile organochlorides (VOCs) fell during the 2-day treatment.

Residuals ranged from 0.5 to 0.05 ppm. Entry #16 (Table 1) examined a new technique (stable isotope fractionation) as a tool for proving that PCE and TCE were

anaerobically biodegraded in an aquifer in which a dense nonaqueous phase liquid

(DNAPL) source of PCE and TCE resided. Under the circumstances, complete

cleanup could not be assessed—but a significant (40 percent) proportion of the

TCE was shown to have been metabolized.

Conclusions about the 16 bioremediation projects shown in Table 1 are:

 (1) residual contaminants remained in all soils and subsurface waters; (2) mixing

and aeration accelerated organic contaminant loss; and (3) in some circumstances,

cleanup goals and site closure were achieved.

To augment the survey of field bioremediation projects presented in Table 1,

another recent source of information was consulted. In Appendix D of “Treatment

Technology for Site Clean-Up: Annual Status Report (Ninth edition)” (EPA, 1999), a

summary of treatment technology status-report updates was presented. This site-bysite compilation listed 8 years of additions, changes, and deletions of Records of

Decisions for treatment technologies used at US EPA Superfund Sites. Of the 373

changes presented, 62 concerned bioremediation treatment. Of these 62 bioremediation

projects, 3 were found to be ineffective in bench-scale biodegradation tests and 9

were abandoned because treatment goals could not be met in the field.

It is clear from the above-mentioned EPA report on site cleanup technologies and

from the data in Table 1 as well as from related compilations [e.g., laboratory studies

of PAH biodegradation in soils (Hughes et al., 1997); field observations of pesticide

and other chemical persistence in soil (Alexander, 1997)] that both microbial biodegradation and phytoremediation projects are far from 100 percent efficient. Indeed, even when the soil described in entry 1 of Table 1 was re-spread onto experimental field soils, an average of 82 percent of the oily residues remained 2 1/2 years

later (Angehrn et al., 1999). The question is “Why?” What causes substances to

persist in soil? Certainly bioavailability is one of the candidate explanations for

ineffective bioremediation, but the influence of bioavailability must be evaluated

simultaneously with other alternative hypotheses. The next section summarizes

the past and present understanding of mechanisms by which organic compounds

resist microbial metabolism.
31

Section 5

Mechanisms of Persistence of Organic

Compounds

5.1 An Evolutionary Perspective on the Persistence of

Organic Compounds

The Oxford English Dictionary defines “persistence” as “the action or fact of

persisting; firm or obstinate continuance in a particular course in spite of opposition;

continued existence in time or (rarely) in space; endurance; continuous occurrence.”

Within a biogeochemical context where concern is focused on the behavior of organic compounds in soils, sediments, and waters, “persistent” compounds are ones

that, once found, continue to be found. This tendency to persist occurs despite the

fact that soils, sediments, and waters in the real world are dynamic, open systems

where processes such as dilution, volatilization, photolysis, sorption, advection, and

microbial biodegradation all contribute toward the disappearance of organic compounds. Biodegradation is often unique among these processes because enzymecatalyzed biochemical rearrangements often convert organic compounds completely

to nontoxic carbon dioxide (Madsen, 1991). Such microbial mineralization reactions

contribute to the growth of microorganisms; thus, such processes are robust.

Millions (Wackett, 1996; Wackett and Hershberg, 2000) of naturally occurring

organic compounds (synthesized by plants, animals, and microorganisms) are constantly released into the biosphere. These compounds are continuously destroyed

and transported by the above-mentioned processes that prevent persistence. Consequently over broad temporal and seasonal scales, naturally occurring organic

compounds are generally maintained in steady-state concentrations in soils, sediments, and waters. A widespread readily biodegradable substance, such as glucose (released by microbial attack of the products of photosynthesis), is turned

over rapidly and occurs in soil at very low concentrations; while biochemically

resistant soil organic matter recycles relatively slowly and occurs at relatively high

concentrations. Soil organic matter is a “persistent” naturally occurring material. A
32

small subset of naturally occurring organic compounds are toxic (e.g., aflatoxin,

botulinum toxin), but occur so rarely or are restricted to such rare habitats (e.g.,

peanuts, anaerobic foods) that these toxins seldom threaten human health and/or

ecosystem function. In summary, compounds of low (or rare) toxicity in balanced

dynamic steady-state are the rule for naturally occurring organic compounds.

Industrial chemicals are subject to the same forces that govern the concentrations of naturally occurring organic compounds. However, two additional factors

need to be considered when attempting to understand the environmental fate and

persistence of industrial chemicals. First, some industrial chemicals exhibit newly

synthesized, novel molecular structures that are unlike naturally occurring compounds. These may not be susceptible to enzymatic attack and/or may not have

provided sufficient selection pressure to allow the assembly of efficient metabolic

pathways in microbial systems. In the absence of well-integrated enzymatic systems, microorganisms might not attack novel synthetic chemical structures and

persistence could result. Second, industrial chemicals are often released nonuniformly

in time and space, often at high concentrations. This patchy, discontinuous distribution can often overwhelm naturally occurring processes that might otherwise

successfully maintain a dynamic balance between production and destruction. At

high concentrations, even readily biodegradable compounds can noticeably persist in soils, sediments, and waters. Regardless of the tendency to persist, industrial chemicals that are nontoxic (e.g., polyethylene) may not threaten human health

or ecosystem function, but the combined traits of persistence and toxicity in an

industrial chemical (e.g., PCBs, dioxin) can pose serious environmental dilemmas.

5.2 Placing “Bioavailability” Within the Established

Framework of Persistence Mechanisms

In an insightful review written more than a quarter century ago, M. Alexander

(1973) prepared a list of factors contributing to the persistence (recalcitrance) of

organic compounds. This list was as follows:

I. Property of Organic Molecule

a. Chemical resists attack by all existing enzymes (proper enzymes have not yet

evolved),

b. Molecule unable to pass through cell wall hence unavailable for enzyme

induction and intracellular attack,

c. Site on molecule for enzymatic attack obscured by intramolecular folding,

d. Molecule composed of a variety of subunits connected by a variety of

linkages (thus it is improbable for all of the proper enzymes to be coincident

in time and space),
33

e. Molecule does not yield energy or carbon for microbial growth (cometabolism

may occur, however), and/or

f. Molecule or its products are toxic.

II. Property of environment

a. Environment exceeds the tolerance range of environmental factors affecting

survival and activity of organisms ordinarily responsible for mineralization

of the compounds. Factors include temperature, pH, moisture, light, salinity,

toxins, redox potential, and hydrostatic pressure,

b. Environment lacks growth factors or nutrients essential for microbial activity,

c. Environment inactivates enzymes that would otherwise attack the molecule,

d. Environment is sufficiently voluminous to reduce the concentration of the

molecule below a threshold required for enzyme induction and/or microbial

activity,

e. Environment renders the molecule inaccessible to enzymatic attack by shielding the molecule in microsites on solids or coating the molecule with inert

substances,

f. Environment contains resistant organic substances which form complexes

with the molecule, thus rendering the molecule resistant, and/or

g. Environment contains metallic cations which form complexes with the molecule, thus rendering the molecule resistant.

Readers of this report may perhaps be surprised to note that “bioavailability”

was not explicitly mentioned in the above listing. Instead, the hypotheses for explaining resistance of organic compounds emphasized properties of the organic

molecule and of the environment.

The last four items listed above (II. d, e, f, g under properties of environment)

are manifestations of interactions between contaminants and soil or sediment. Moreover, threshold concentration, shielding in microsites, organic complexation, and

metallic complexation fully encompass the six bioavailability-based mechanisms

identified in Section 3.3.

All six hypothesized mechanisms by which bioavailability controls biodegradation (sorption, NAPL-partitioning, micro- or nanopores, absorption into native

organic matter, organic complexation, and insolubility) stem from interactions

between the contaminant and the soil or sediment. Therefore, progress in comprehending the influence of bioavailability on bioremediation requires understanding the fundamentals of interactions between contaminants and soils and sedi-
34

ments (Section 6 of this report) and the mechanisms of biotic uptake of contaminants (Section 7 of this report).

5.3 Illustrating Mechanisms of Persistence: Soil Organic

Matter (SOM)

To facilitate the task of interpreting data describing the persistence of chemical

wastes in soils (Section 8 of this report), it may be instructive to first review our

understanding of the persistence of a ubiquitous, nontoxic naturally occurring material, soil organic matter (SOM). Because of its agronomic significance in maintaining

soil physical conditions and nutrient status, decades of scholarly research have been

devoted to understand why SOM resists microbial decomposition processes (e.g.,

Alexander, 1973; Allison 1965; Brady and Weil, 1999; Stevenson, 1994; Stout et al.,

1981; vanVeen and Paul, 1981). As will become clear in Section 6, ever-improving

models are needed to understand soil systems. A synopsis of the current model for

SOM is presented below.

Organic matter in soil is composed of recognizable plant material, unrecognizable plant material in various stages of decay, soil biota (microorganisms, soil fauna),

and the persistent material, humus. Humus is divided into two pools (McBride,

1994; Stevenson, 1994): biochemically familiar plant-derived compounds such as

polysaccharides, polypeptides, and lignins; and humic substances (amorphous

polymeric compounds that are inherently resistant to decomposition). In the threedimensional soil matrix, the bulk of soil organic matter occurs as water insoluble

forms that have been traditionally characterized based on their extractability. Four

main “associations” between the inorganic matrix of soil and soil organic matter

have been identified (Stevenson, 1994): (1) insoluble macromolecular complexes;

(2) macromolecular complexes bound together by di- and tri-valent cations such as

Ca2+, Fe3+, and Al3+; (3) organic-clay mineral complexes linked by polyvalent cations (clay-metal-humus), H-bonding, and other ways; and (4) organic substances

intercalated within the interlayers of expanding-type clay minerals.

Soil organic matter does not accumulate indefinitely in well-drained soils. Instead, it reaches equilibrium levels governed by several “not entirely satisfactory

explanations” (Stevenson, 1994): by the ability of dark, high molecular weight

humic substances to resist microbial attack; by nutrient limitations on humic substance synthesis by microorganisms; and by the degree of protection afforded

both pools of humus (familiar classes of compounds and humic substances) through

associations with polyvalent cations and clays.

The dynamic state of soil organic matter must be recognized. When plant materials reach the soil, the nine main classes of plant-derived molecules are decomposed by microorganisms in the following order of decreasing rates: sugars, starches,

simple proteins > crude protein > hemicelluloses > cellulose > fats, waxes > lignins.
35

Stevenson (1994) has reviewed the considerable information on soil decomposition

processes obtained using 14C-labeled plant residues. For soils of the temperate

zone, approximately one-third of the plant-derived carbon remains behind in soil

after the first growing season, mostly as a mixture of “labile” and “stable” components of humus. Although attempts to measure the age of the stable organic matter

fractions in soil have been made (Stout et al., 1981; Stevenson, 1994), absolute age

determinations are often thwarted by the dynamic nature of humus: the standing

pool of humus is continuously decomposed while new humus is re-synthesized.

Nonetheless, mean residence times of the stable pool of modern humus have been

estimated at several hundred to somewhat over 1,000 years. According to Stevenson

(1994), the mean residence time of the “labile” pool of plant carbon (that remaining

after the first season of decomposition) is initially short, but then approaches that

of the stable, native humus carbon.

The biochemical processes by which high molecular weight humic substances

form are not fully understood. Sposito (1989) suggests four stages in the transformation

of plant biomass into humus: (1) decomposition of biomass (including lignin) into

simple organic compounds; (2) microbial metabolism of the simple compounds; (3)

cycling of C, H, N, O between organic matter and microbial biomass; and (4) microbially

mediated polymerization of the cycled organic compounds. The principal humus-forming compounds involved in stages (3) and (4) are believed to be phenolic high molecular

weight structures, especially lignin constituents, which are converted to a reactive

class of compounds containing quinones that readily condense randomly (Flaig, 1975,

as cited in Stout et al, 1981; Sposito, 1989; Stevenson, 1994). A key trait of humic

substances is their dissimilarity to the biomolecules from which they are derived.

Soil management practices have long been known to influence SOM persistence. When either forest or prairie soils are brought under cultivation, a general

decline in SOM occurs (Brady and Weil, 1999; van Veen and Paul, 1981). The

decline in SOM through cultivation and cropping practices can only partially be

attributed to a reduction in the added quantity of plant residues for humus synthesis. It is thought that loss of SOM occurs because cultivation aerates the soil (thus,

stimulating microbial respiratory activity) and exposes previously inaccessible organic matter (such as in micropores) to microbial attack (Stevenson, 1994)].

5.3.1 Summary (Soil Organic Matter Persistence)

 Key facts related to SOM persistence include:

(i) Plant-derived biomass (consisting of many familiar classes of organic structures) is continually replenished in the soil system;

(ii) Microbial processes convert dead biomass to two humus pools—recognizable

(e.g., polysaccharide, polypeptide, and lignin) and nonrecognizable (humic

substances);
36

(iii) Complexation reactions by metallic soil cations influence both pools of humus—thereby rendering them resistant to microbial attack;

(iv) Both pools of humus associate with clay minerals (inert substances)—thereby

rendering the humus resistant to microbial attack;

(v) The “humic substances” pool of SOM also resists enzymatic attack because it

is composed of a variety of unusual molecular subunits randomly connected

by diverse chemical linkages; and

(vi) SOM persistence can be partially overcome by simple mixing of soil—this is

thought to boost aeration and accessibility of SOM to microbial attack.

Items (iii)-(v), above, correspond to items II(g), II(e), and I(d) in Alexander’s

(1973) scheme for explaining persistence (Section 5.2). In addition, entries I(a)-I(d)

of Alexander’s (1973) scheme contribute, mechanistically, to the persistence of

humic substances. Therefore, SOM provides an illustration of the many mechanisms of persistence and their ability to act in concert. Another crucial point about

the origin of SOM is that approximately one-third of the initial mixture of naturally

occurring biodegradable plant materials entering soil resists biodegradation in the

first season of decay. Furthermore, much of this residue enters the stable humus

pool with turnover time of hundreds to more than 1,000 years. To the degree that

plant-derived compounds resemble chemical wastes, lessons from the persistence

of SOM can guide inquiry into the behavior and bioavailability of chemical wastes

in soil.
37

Section 6

Paradigms for the Composition and

Structure of Soil* and the PhysicalChemical State of Contaminants Therein

Ideally, in all remediation (including bioremediation) scenarios, the reactions of

both inorganic and organic contaminants in real-world sites would be known prior

to proposing management and technical cleanup procedures. These reactions are

governed by: (1) properties of the contaminants (these are of wide-ranging complexity because they may occur as mixtures whose individual components may

have multiple reaction pathways); (2) properties of the soils or sediments

(geosorbents*; as discussed throughout this section, especially Section 5.4,

geosorbents are extremely complex); and (3) interactions between contaminants

and geosorbents. In reality, however, site management strategies employ an empirical, iterative, observational approach in which new information is constantly gathered to improve and adjust issues tied to the hazards of contaminants and their

removal/detoxification (National Research Council, 2000).

What follows are several discussions designed to foster an awareness of the

compounded complexities of contaminant-geosorbent interactions. This awareness will later (Section 8) be brought to bear on the major focus of this report

(bioavailability and bioremediation).

* Note: Information presented in this section will treat the terms “soil,” “sediment,” and

“geosorbent” (Luthy, et al., 1997) as synonyms. It will be assumed that the facts and

principles described here apply to all naturally occurring porous geological matrices.
38

6.1 Soil Complexity

Scholarly inquiry into the intrinsic properties of soil (pedology), separate from

their impacts on plant growth (edaphology), developed significantly in Europe,

Russia, and the United States in the 19th century (Brady and Weil, 1999). At least

two complementary approaches to soil science have progressed simultaneously

since then: field approaches to natural history and soil genesis; and laboratory

approaches (chemical, biological, mineralogical, and physical determinations) applied to soil samples. Despite advances in both approaches throughout the 20th

century, McBride (1994) has written that “much of soil science is empirical rather

than theoretical in practice. This fact is a result of the extreme complexity and

heterogeneity of soils, which are impossible to fully describe or quantify by simple

chemical or physical models.”

Soils are natural bodies, whose lateral and vertical boundaries usually occur as

gradients between mixtures of materials of atmospheric, geologic, aquatic, and/or

biotic origin. Soils are open systems subject to fluxes in energy (e.g., sunlight,

wind) and materials (e.g., aqueous precipitation, erosion, deposition, and inputs of

organic compounds from activities of plants, human beings, and other animals).

Furthermore, soil’s intrinsic complexity stems from its nature as an assemblage of

solid, liquid, gaseous, organic, inorganic, and biological constituents whose chemical

composition and random three-dimensional structure have not been completely

characterized. In addition to physical complexity, the microbial (bacteria, fungi,

algae, protozoa, and viruses) physiological processes in soil and their multitude of

interactions are dauntingly complicated. Compounding the challenge of understanding in situ soil processes is the fact that abiotic reactions (e.g., precipitation,

dilution, hydrolysis; Section 6.2) also must be considered when attempting to understand soil geochemistry. Furthermore, in a field setting, plants and animals also

effect geochemical change.

Attention also must be paid to the fact that soil properties described above are

subject to dynamic changes in time and space. No field setting is homogeneous or

static. Regarding spatial inhomogeneity, the physical, chemical, nutritional, and

ecological conditions for soil biota undoubtedly vary from the scale of micrometers

to kilometers. Regarding temporal variability, in situ processes that directly and

indirectly influence fluxes of material into, out of, and within soil are dynamic.

Climate-related influences (such as temperature, sunlight, evaporation, and precipitation) are probably major variables that cause temporal, variations in biogeochemical processes in soil.

6.2 A Thermodynamic Overview of Inorganic Soil Reactions

Lindsay (1979) provided a unifying thermodynamic overview of soil in which

dissolved substances in soil solution are in constant dynamic equilibria with six
39

independent chemical influences: solid mineral phases; exchangeable ions and surface adsorption; nutrient uptake by plants; soil air; organic matter and microorganisms; and water flux. The mineral phases of soil (typically 90 percent of the solid

matter) have been described as “rock on its way to the ocean” (Lindsay, 1979).

Primary minerals (the parent material from which soils are derived ) were often formed

under conditions of high pressure and temperature. At the earth’s surface, subject to

oxidative and hydrolytic weathering, the primary minerals become secondary minerals as ionic species in solution are leached away and the remaining mineral structures

seek lower free energy levels in their atomic arrangements. Soils contain numerous

minerals, some of which are crystalline, while others are amorphous or meta-stable.

These minerals both respond to and control the dynamic pool of dissolved constituents in soil solution. A detailed discussion of soil mineralogy and equilibria is beyond

the scope of this report (for this see, Dixon et al., 1977; Lindsay, 1979; McBride, 1994;

Sposito, 1987); but it is critical to appreciate that soil and sediment habitats are in

constant chemical transition, albeit at rates that are slow in human terms. Many of

the mineral components are thermodynamically unstable, and this instability is

compounded by additional reaction pathways imposed by (micro)biological processes—

especially those driven by plant-derived carbonaceous materials added via photosynthesis (see Sections 5.3 and 6.1).

6.3 Models of Soil Structural Characteristics and

Limitations in Environmental Microbiology

This section will review information describing the three-dimensional arrangements of soil components. Ladd et al. (1996) have reviewed relationships between

soil components and the biological activity occurring therein. These authors emphasized the vastness in the range of scale of soil constituents (nine orders of

magnitude, from atoms to rocks) and the hierarchical features of soil aggregates

that form the three-dimensional fabric of soil. Broadly, six size-based categories of

aggregation were described (Ladd et al., 1996): (1) amorphous minerals develop at

the nanometer-to-angstrom scale; (2) clay microstructure colloids form at 10-7m;

(3) “quasicrystals,” “domains,” and “assemblages” form (10-7 to 10-5 m) between

clay, silt, and smaller particles; (4) macroaggregates (0.1 to 250 μm) occur between

sand, silt, and smaller particles; (5) macroaggregates (250 μm to 25 mm) occur

between gravel, sand, and smaller particles; and (6) clods (> 25 mm) occur between

rocks, gravel and smaller particles.

It is the aggregation, the aggregate behavior, of soil that contributes to its complexity. Tisdall and Oades (1982) insightfully presented a schematic model of the

aggregate organization of soil (Figure 2). Emphasized in Figure 2 are both the hierarchical scales of soil aggregates and the mechanistically crucial binding agents responsible for aggregate formation. It is clear in Figure 2 that the various components

of SOM (Section 5.3, this report) play a major role in creating soil structure. Roots,
40

FIGURE 2 Models, at five different scales, of soil components and their contribution to soil structure (from Tisdall & Oades, 1982).
41

hyphae, plant debris, fungal debris, bacteria, and humic materials are specifically

mentioned in Figure 2 because of their structural contributions to the soil matrix.

Documentation of “soil micromorphology” (or “soil fabric”; Ringrose-Voase

and Humphreys, 1994) has played a major role in establishing and reinforcing the

type of model of soil aggregate organization shown in Figure 2. Microscopic procedures applied, whenever possible, to intact soil samples (Foster, 1993; Ladd et al.,

1996) include: transmitted light through soil thin sections, transmission electron

microscopy (TEM), scanning electron microscopy (SEM), electron microprobe analysis (EMP), and environmental SEM. These approaches have provided direct observations of intimate associations in soil aggregates of solid surfaces, root hairs,

fungi, bacteria, extracellular polysaccharides, clay films (cutans), humic substances,

and cellular debris. Such “ultrastructural” studies have revealed that the soil biomass occupies only 0.001 percent of the soil volume. Microorganisms, though present

in large numbers (approximately109 cells per gram) are neither uniformly nor randomly distributed but, as revealed by TEM of soil sections, have been found

clumped near or within cellular residues or in micropores (Ladd et al., 1996). Electron microscopy/EMP approaches also have been applied to controlled model systems, and provided important insights into the mechanisms by which microorganisms can transform (i.e., precipitate and/or dissolve) inorganic materials (e.g., Warren and Ferris, 1998). However, one well-recognized limitation of high-resolution

soil microscopy is that each image surveys such a small soil volume that accruing

information truly representative of bulk soil remains a challenge (Foster, 1993).

 Within the discipline of environmental microbiology, there has been longstanding interest in understanding small-scale relationships between microorganisms and their habitats. Microscopic environmental microbiological inquiries into

the identity and biogeochemical activity of microorganisms have been conducted.

The goals of these inquiries have been approached by using various combinations

of confocal scanning laser microscopy (for three-dimensional images), autoradiography (to locate both added model radioactive substrates and to document their

uptake by microorganisms), fluorescent antibody staining of individual microorganisms, and probing of naturally occurring microorganisms with fluorescent-labeled oligonucleotides that hybridize with 16S rRNA molecules (Amann, et al.,

1995; Amann and Kuhl, 1998; Ghiorse et al., 1996; Lee et al., 1999; Ouverney and

Fuhrman, 1997, 1999). In a now slightly dated review, Madsen (1996) summarized

the accumulating attempts to discover both the in situ biogeochemical activities of

microorganisms in soil and their identity. Conclusions were that methodological

limitations of sample preparation, incubation, and analysis have prevented achieving the goal of knowing “who, what, when, and where” of microorganisms in soil

habitats (Madsen, 1996). A recent exception to this statement is the discovery that

anaerobic methane oxidation in deep sea sediments can be carried out by an intimate association between sulfate-reducing and methanogenic bacteria (Hinrichs et

al., 1999; Boetius et al., 2000, Orphan et al., 2001). Existing methods for growing
42

and describing microorganisms in soil typically overlook at least 90 percent of

those that are detectable microscopically (Amann et al., 1995; Madsen, 1998). Thus,

not only has the complexity of the soil habitat prevented its full chemical characterization, but there may be a vast, undiscovered diversity of microorganisms,

perhaps with novel metabolic properties, residing in soil. A complete census of

naturally occurring microorganisms has never been achieved in any habitat

(Madsen, 1998). Moreover, microorganisms responsible for biogeochemically significant processes in natural habitats have almost never been identified (Madsen,

1998).

6.4 Interactions Between Geosorbents and Contaminants

There is a broad diversity of physical and chemical forces that governs the

interactions between geosorbents (soils and sediments) and contaminants. These

interactions range from colloidal and electrochemical phenomena to acid-base and

sorptive reactions (e.g., McBride, 1994; Sposito, 1989). Given the perhaps “impossible” complexity of soil and sediment systems (Sections 6.1 and 6.3 of this report), it

is likely that an accurate and comprehensive understanding of soil reactions will only

be approached by combining the results of many types of inquiry. Physical, chemical,

mineralogical, and microbiological characterization of extracted soil constituents can

augment direct microscopic observations of soil micromorphology (Section 6.3). As

was emphasized, conceptually, in Sections 3.1 and 3.3, the central issues of this report

focus on the details of specific interactions between specific geosorbents and specific contaminants. Considering the wide-ranging properties of potential geosorbents

(e.g., types of organic matter, types of minerals, and their respective abundances) and

potential contaminants (e.g., organic, inorganic, soluble, insoluble, ionic, hydrophobic), a thorough examination of the many possible permutations of geosorbentcontaminant interactions is beyond the scope of this report. Nonetheless, both organic and inorganic contaminants will be discussed below. A particular class of organic contaminants, hydrophobic organic compounds, will serve to illustrate how

inferences of geosorbent structure can be drawn from indirect observations of contaminant behavior. Examples of direct observations of organic and inorganic contaminants in soils also will be presented.

6.4.1 Inferences of Geosorbent Structure Based on Indirect

Observations of Hydrophobic Organic Compounds

In a recent critical review on sequestration of hydrophobic organic compounds

(HOCs) by geosorbents (Luthy et al, 1997), a group of experts stated that “currently there are no direct observational data revealing the molecular-scale locations in which nonpolar organic compounds accumulate when associated with

natural soils or sediments. Hence, macroscopic observations are used to make
43

inferences about sorptive mechanisms and the chemical factors affecting the sequestration of HOCs by geosorbents.” Many such macroscopic observations are

derived from laboratory operations that: (1) react soil or sediment with a solution of

known composition at fixed temperature and pressure for a prescribed period of

time; and (2) chemically analyze the reacted soil, the soil solution, or both to determine their compositions (Sposito, 1989). For decades, these types of experiments

have been carried out on combinations of HOCs and field-derived or artificial

geosorbents. By analyzing the rates and extents of sorption and desorption, while

varying key components of the geosorbent [e.g., organic matter, mineral surfaces,

and non-aqueous phase liquid (NAPL) content], many investigators have independently developed theories about geosorbent properties. These have often involved

mathematical models that explicitly and quantitatively define a variety of “sites” for

sorption and desorption reactions based on equilibrium and kinetic interactions

between contaminants and geosorbents. These types of analysis have given rise

to many theories that contribute explanations for contaminant behavior (e.g., “twosite model,” “two site-two region model,” “gamma model of continuous distribution of rate constants,” “organic matter diffusion model,” “sorption-retarded pore

diffusion model”; see Brusseau et al., 1991; Connaughton et al., 1993; Pignatello

and Xing, 1996; Scow, 1997; Young and Weber, 1995). These models often evoke

physical characteristics of sorbents such as “intraparticle nanopores” (Pignatello

and Xing, 1996) and “glassy versus rubbery organic matter” (Graber and Borisover,

1998; Pignatello and Xing, 1999; White and Pignatello, 1999).

The recent consensus critical review mentioned above (Luthy et al., 1997) has

summarized the “state-of-the-science” for HOC-geosorbent interactions. Figure 3

graphically presents current ideas. Please note that there is a great deal of similarity

between Figure 3 (from Luthy et al., 1997) and Figure 2 (the soil model presented by

Tisdall and Oades, 15 years earlier). The components of Luthy et al’s. (1997)

geosorbent model (Figure 3) are: mineral-phase geosorbent (the particles); amorphous sorbent organic matter (“SOM” in Figure 3); dense SOM; clay particles and/

or oxide coatings; NAPLs (fresh, and aged or weathered); mesopores; micropores;

combustion residue (e.g., soot) and water or gas in pores. Each of the model components was carefully scrutinized and justified based on a variety of physical, chemical, and contaminant-behavior techniques.

As summarized by Luthy et al. (1997), the key measurement parameters that provide clues to geosorbent properties are: sorption-kinetics (is equilibrium reached rapidly or slowly? is there evidence for hysteresis?); extent of equilibrium partitioning of

the contaminant between aqueous and solid phases (is this partitioning linear or nonlinear?); heats of sorption (low or high?); activation energy (low or high?); competition

(do other sorbates influence sorbtion reactions?); sorbate characteristics (do molecular

steric effects influence sorption?); and extractability (can solvents alter geosorbent

behavior by removing key components?). Based on a systematic examination of the

patterns of HOC partitioning behavior in the presence of many types of geosorbents,
44

FIGURE 3 Conceptual model of geosorbent domains developed by Luthy et al.,

(1997). The circled letters refer to representations of sorption mechanisms (described in text, Section 6.4.1.). The geosorbent domains include different forms of

sorbent organic matter (SOM), combustion residue particulate carbon such as soot,

and anthropogenic carbon including Nonaqueous Phase Liquid (NAPLs).

Luthy et al. (1997) presented five hypothetical mechanisms for explaining how HOCs

interact with geosorbents. Each of these five mechanisms is represented graphically in

Figure 3 as a capital letter from A to E. Case “A” is absorption into amorphous or “soft”

natural organic matter or NAPL. Case “B” is absorption into condensed or “hard”

organic polymeric matter or combustion residue (e.g., soot). Case “C” is adsorption

onto water-wet organic surfaces. Case “D” is adsorption to exposed, water-wet mineral

surfactants (e.g., quartz). Case “E” is adsorption into microvoids or microporous minerals (e.g., zeolites) with porous surfaces at water saturation.

The reader should note the correspondence of the five hypothetical sorption

mechanisms (letters A to E in Figure 3) mentioned earlier with the six hypothesized mechanisms that reduced bioavailability from Section 3.3. The consensus
45

critical review by Luthy et al. 1997) concluded by recommending that “the discipline is in need of a more sophisticated understanding of...mechanisms at the

microscale.” Some progress in this area has been recently made (see below).

6.4.2 Direct Observations of Contaminated Geosorbents

Direct, small-scale observations of biological and geological materials have led

to profound discoveries in virtually all areas of scientific endeavor (e.g., histology,

mineralogy, biochemistry, and the engineering and material sciences) since the

invention of the microscope (Goldstein et al., 1992). The trend in recent decades

has been in at least two directions: toward greater resolution (e.g., atomic force

microscopy, angstrom resolution in x-ray crystallography of proteins) and toward

combining spectroscopic with visual analyses so that both the spatial arrangement

and the composition of materials can be simultaneously integrated.

Although new technical advances in microscopic characterization of environmental materials yield new information, methodological limitations are slow to be

fully overcome. As an example, Myneni et al. (1999) recently studied the macromolecular structure of humic substances using high-resolution spectromicroscopy at

the Advanced Light Source at Lawrence Berkeley Laboratory. These investigators

achieved heretofore unattainable image resolution that demonstrated the influence

of solution chemistry, mineralogy, and source of origin on the size and shape of

humic substances isolated from aquatic and soil environments. Though such details of the macromolecular structure of humic substances were new, the true threedimensional arrangement of humic substances in intact soil matrices remains virtually unexplored.

6.4.2.1 Hydrophobic Organic Compounds

During preparation of this report, several articles were encountered that directly examined organic contaminants in geosorbent matrices. Examples follow.

Gillette et al. (1999) used microprobe two-step laser desorption/laser ionization

mass spectrometry (μL2 MS) to identify and characterize trace PAHs on geosorbents.

The geosorbents examined included soil spiked with PAHs and also both biotreated

and field-derived contaminated samples (soil and sediment). PAH occurrence was

mapped across the surface of individual particles at 40 μm increments. Results

indicated that sorption phenomena were heterogeneous: individual geosorbent

particles seemed to be composed of different subparticle-size regions having different affinities for PAHs.

Ghosh et al. (2000) continued the μL2 MS studies but added Fourier transform

infrared (FTIR) micro-spectroscopy and scanning electron microscopy/wave length

dispersive x-ray spectroscopy (SEM/WDX) to their inquiry. These multiple assays
46

were used to examine PAH contaminants across sections of sediment particles

derived from Milwaukee Harbor. PAH concentrations in coal- and wood-derived

particles were found to be several orders of magnitude higher than on silica particles. Surface analyses revealed that PAHs were associated with organic-rich areas

on sand particles, and not associated with bare silica regions. Cross-sectional

analyses of particles (silica and coal) showed that PAHs were far more abundant (30

to 100 times) on exterior, compared to interior, regions of these particles. Thus, nearsurface sorption processes (<5 μm) and not deeper absorption processes were

indicated by this study. Ghosh et al. (2000) also fractionated the Milwaukee sediment according to particle size and density. This showed that, although organic

material (coal and wood) only comprised 5 percent of the total sediment weight, this

fraction contained 62 percent of the PAHs. When PAH desorption from the various

sediment fractions was measured, the coal/wood organic material was found to

release less than 10 percent of the bound PAHs to Tenax resin. This contrasted with

more than 80 percent release of the sorbed PAHs from the silt/clay inorganic fraction. Ghosh et al. (2000) suggested that the PAHs associated with wood and coal

may not be bioavailable.

In a followup study, Ghosh et al. ( 2001) used desorption kinetics, thermal

program desorption-mass spectrometry (TPD-MS), and mathematical modeling to

further characterize the behavior of PAHs present in the Milwaukee Harbor sediments. Desorption activation energies and rates were calculated for a range of

PAHs. The model that best described desorption was one in which the PAHs were

located like a rind on the outer regions of sorbent particles. The authors concluded

that PAHs associated with clay/silt particles desorb rapidly and are characterized

by low desorption activation energies. In contrast, PAHs associated with coalderived material desorb at much slower rates and are characterized by high desorption activation energies. The study concluded that “PAHs associated with media

having large activation energies may thus comprise the unavailable fraction in

sediments and the PAHs may pose less risk than PAHs in clay and silt fraction.”

Two other studies will be highlighted here to illustrate recent experimental

attempts to explore contaminant-geosorbent interactions. Unlike Gillette et al. (1999)

and Ghosh et al. (2000, 2001) who largely examined geosorbents that were already

contaminated, the other studies began with uncontaminated geosorbents and added

model PAH compounds.

Guthrie et al. (1999) implemented a study designed to determine the structural

composition and molecular interactions of pyrene with soluble and insoluble organic

matter fractions of sediments. Sediments treated with a biocide and untreated sediments were incubated with [13C]-pyrene in aerated microcosms over a 60-day period.

These investigators used pyrolysis-GC/MS and 13C-NMR to follow subsequent interactions with naturally occurring organic fractions (humic acid and humin). The

unique spectroscopic probing, 13C-NMR, allowed Guthrie et al. (1999) to determine

that the added pyrene remained intact throughout the experiment. Yet, during the 60-
47

day period, the pyrene became increasingly resistant to solvent extraction, while

becoming associated with the humic acid and humin sediment fractions. Furthermore,

pyrolysis GC/MS indicated that the pyrene-organic matter association involved

“adsorbtion or encapsulation (not covalent binding)” and that this association was

enhanced in the presence of viable microorganisms (Guthrie et al., 1999).

Schultz et al. (1999) designed a study to elucidate the relationships between

soil organic matter (SOM) structure and sorption behavior of phenanthrene, another PAH. Pyrolysis-GC/MS was used to directly characterize the heat-labile molecular components of SOM in five geosorbents—three surface soils and two subsurface sediments. Principal components analysis of the pyrolyzed fragmentation

patterns allowed clear distinctions to be drawn between the types of SOM in each

geosorbent. Sorption and desorption parameters were evaluated for the geosorbents,

and correlation analyses were performed between the sorptive properties and the

soil organic matter components. Phenanthrene thermal desorption profiles also

were determined for each geosorbent; these suggested the presence of both mineral and organic-matter components in geosorbents that impede release of PAHs.

6.4.2.2 Inorganic Compounds

Inorganic contaminants [such as heavy metals, metalloids, radionuclides, and

oxyanions (e.g., nitrate, chlorite)] exhibit a wide spectrum of properties (e.g., oxidation state, speciation) that allow them to react with inorganic or organic geosorbent

components, or both. Many sources of information describing behavior, reactivity,

mineralogy, and other aspects of inorganic chemical contaminants have been published (e.g., Adriano, 1992; Alleman and Leeson, 1999; Alloway, 1995; National

Research Council, 2000; Vandegrift et al., 1992; Vangronsveld and Cunningham,

1998). Efforts to characterize inorganic contaminants in soil have progressed significantly in recent years. Many direct microscopic and analytical procedures have

been applied to many combinations of geosorbents and contaminants. Selected

examples are presented below.

During uranium processing, aqueous, solid, and airborne radioactive wastes

were released to soil at the US Department of Energy’s Fernald, Ohio, site. Morris et

al. (1996) examined contaminated soil samples using a combination of x-ray absorption, optical luminescence, and Raman vibrational spectroscopies, along with ancillary techniques such as energy dispersive scanning electron microscopy and powder x-ray diffraction. The objective was to ascertain the oxidation state, the chemical

form, and the physical state (surface precipitate, secondary mineral or absorbate) of

the uranium. The procedures used had experimental advantages of: not requiring

invasive sample preparation; a spatial resolution range from 1 cm2

 to less than 100

μm2; and spectroscopic techniques that applied to both amorphous samples and

submolecular coatings. The x-ray absorption spectroscopy provided definitive evidence that the bulk of the oxidation state distribution of uranium (75-95 percent)
48

favored the hexavalent species. Furthermore, the uranium minerals that were identified, similar to autunite and schoepite, often appeared as platey tabular grains ranging in size from 10 to 100 μm or in association with geothite or quartz. Considerable

weathering (especially oxidation and both phosphate and hydroxide precipitation) of

the initially released uranium had occurred. Additional uranium minerals also were

noted, and their photodecomposition properties and field distribution suggested that

uranium binding by organic ligands also may have occurred at the Fernald site.

Welter et al. (1999) used a combination of XAFS (X-ray Absorption Fine Structure Spectroscopy), XANES (X-ray Absorption Near Edge Structure), and SEM/

EDX (Scanning Electron Microscopy/Energy-Dispersive X-ray fluorescence spectroscopy) to examine the chemical speciation of Pb in two soil samples from a

battery manufacturing plant in Hanover, Germany. Pb was heterogeneously distributed in the soil, concentrations ranged from between 50 and 140 g/kg. To contend

with the inhomogeneity, the soil samples were mixed and ground prior to being

analyzed. By comparing XAFS soil signals to those of combinations of authentic

Pb minerals, Welter et al. (1999) were able to quantify the amount of Pb carbonate,

Pb oxide, and Pb sulfate in the two soil samples. SEM revealed individual Pb minerals in the soil matrix.

Galvez-Clouthier and Dubé (1998) implemented a study designed to document

the associations between heavy metal contaminants (Pb, Zn, Cu, Cd) and natural

sediment constituents in the Lechine Canal, Quebec. X-ray diffraction, TEM (Transmission Electron Microscopy) and other geochemical measures revealed that the

sediments consisted mainly of silt- and clay-sized fractions composed of feldspar,

kaolinite, chlorite, calcite, and dolomite, as well as minor amounts of Fe minerals,

amorphous metal oxides, and organic matter. Each of these constituents bound

heavy metals to varying degrees, as assessed by sequential chemical extractions of

residual contaminants from oxide-, carbonate-, organic-, and exchangeable-phases

of the sediments. The heavy metal partitioning patterns for varying sediment size

fractions also were evaluated. Results indicated that no particular mineral phase or

size fraction accumulated particular heavy metals. However, significant concentrations of the heavy metals had accumulated in the sediments over the last century,

and these posed a high risk for metal release into the water column.

6.5 Summary

Section 6 of this report has presented a glimpse into the facts, principles, and

challenges of understanding soil and sediments, and their interactions with organic

and inorganic chemical contaminants. Of the multitude of geosorbents, contaminants, and their interactions, only a few were highlighted here. Many of the longstanding questions about real-world geosorbent matrices remain unanswered: they

are of complex composition, in complex three-dimensional arrangements, that occur

heterogeneously in open biogeochemical systems. Layered onto geosorbent com-
49

plexity is the biology and biochemistry and ecology of the biota that dwell in soils

and sediments. Also, layered onto this are many potential reaction pathways of

organic and inorganic chemical wastes. There is simply much more to be learned

about geosorbents and contaminants (Sections 6.0 and 6.3), especially because

geosorbent composition, contaminant composition, and the geochemical context

for biogeochemical reactions are all likely to be site specific.

Despite the above seemingly “impossible” complexities (Section 6.1), both

geosorbents and contaminants abide by predictable laws of thermodynamics. We

also have robust chemical, mineralogical, and biochemical principles to constrain

the potential reactions (Section 6.2). Furthermore, a growing battery of new microscopic and spectroscopic procedures (Sections 6.3, 6.4) are being applied at an

accelerating rate to real-world contaminated geosorbents. These procedures have

already provided new information that delivers new insights and distinguishes

between competing hypothetical models for geosorbents and their dynamic reactions with chemical waste materials. For instance, microscale spectroscopic assays

suggest that relatively shallow regions of the “soot” component of sediments

feature prominently in governing the release and availability of hydrophobic organic contaminants. Additional progress is likely to be made via an iterative dialogue between hypothesis refinement, application of new analytical technologies,

and an accruing database that catalogs the properties of geosorbent matrices and

the reactions of contaminants therein.
50

Section 7

Uptake of Soil Constituents by Plants and

Microorganisms

This section will proceed from the general to the specific, from text book principles to recent research articles. After the pivotal role of soil solution in all soil

processes is presented, then detailed uptake mechanisms of inorganic and organic

compounds by plants and microorganisms will be discussed. Bioavailability issues

will be emphasized by focusing on factors that limit accessibility of insoluble materials for transport across the extracellular surfaces of plant roots and microorganisms.

7.1 Principles of Soil Solution Chemistry and Uptake by

Plants of Inorganic Compounds

The complexity of the soil habitat was discussed in Section 6.0 of this report

(see especially Sections 6.1, 6.2, 6.5). Text in Section 6.2 drew on Lindsay’s (1979)

observation that “the soil solution is the focal point. The liquid phase that completely envelops the solid phases...is the medium from which plants absorb their

nutrients.”

Regarding inorganic, mineral phase-forming soil components, Lindsay (1979)

further stressed that “two very important parameters influence the availability of an

element to plants. These are: (1) the intensity factor, which is the concentration of

an element in soil solution, and (2) the capacity factor, which is the ability of solid

phases in soils to replenish that element as it is depleted from solution.” Soils may

have a high bulk content of a given nutrient (such as iron) and therefore have high

capacity. However, the solution-phase concentration of nutrients is regulated by

their dominant mineral form. Often, the most thermodynamically stable mineral phase

is the least soluble phase and, if the equilibrium concentration of the stable phase

is below the critical concentration for root uptake, plant nutrient deficiency will

result. Lindsay (1979) emphasized that nutrient concentrations found in soil solu-
51

tion are influenced by other key reservoirs, especially solid-phase sorption sites

and organic matter. Nonetheless, “mineral phases ultimately control the level of

nutrients in solution” (Lindsay, 1979). Thus, characterization of soil minerals and

understanding their stability and reactions with inorganic contaminants is crucial

for predicting the composition of soil solution. This view was reinforced by Sposito

(1989) who stated that “from the perspective of soil chemistry, the bioavailability of

an element is determined by competition among plant root systems, the soil solution, and solid-phase particles.” Sposito (1989) also emphasized that aqueous chemical complexation reactions influence mineral nutrient bioavailability to plants.

Sposito (1989) presented data depicting a direct relationship between the free-ion

species of metals in solution and uptake of metals by plants. Sposito (1989) concluded that: “A chemical element is bioavailable, if it is present as, or can be

transformed readily to, the free-ion species.” Uncharged metal-organic complexes

also are generally bioavailable.

As will be discussed below, the above principles of soil chemistry set the stage

for understanding nutrient bioavailability in soil. But each component of soil solution has its own set of reaction pathways, solubility products, and spectrum of

chemical species. Furthermore, plants and microorganisms sometimes feature unusual physiological adaptations for obtaining specific nutrients from soil solution

(especially organic iron-binding agents, the siderophores). Thus, sometimes elaborate interactions between soil chemistry and soil biota add to the challenge of

understanding contaminant bioavailability issues.

7.2 Movement of Solutes From Soil Solution to Roots

Plant uptake of ionic nutrients requires that they come into contact with the root

surface (Havlin et al., 1999). There are three primary mechanisms by which nutrient

ions in soil could reach the root surface (Barber, 1995; Marschner, 1995): (1) mass flow

of ions in solution; (2) diffusion of ions in soil solution; and (3) root interception of

solid-phase ions. Havlin et al. (1999) present data depicting the relative significance

of each of these three mechanisms for the uptake of 12 nutrients by corn plants.

Unique soil chemical reactions govern each nutrient; therefore, the relative contribution of each uptake mechanism is highly variable. For instance, Havlin et al. (1999)

indicate that, for corn, root interception accounts for 1 percent of nitrogen (N) uptake,

3 percent of phosphorus (P) uptake, excessive (171 percent) uptake of calcium, and 11

percent of iron uptake. In contrast, diffusion-based uptake for the same four elements

was 0 percent, 94 percent, 0 percent, and 37 percent, respectively.

Mass flow (convection) of nutrients into plant roots occurs when dissolved

substances are transported in the flow of water to roots that results from transpirational uptake by the plant. The amount of nutrients reaching roots by mass flow is

determined by the rate of water consumption by plants and the concentrations of

dissolved soil solution nutrients.
52

Diffusion-based uptake of nutrients by roots from soil solution is governed by

concentration gradients. A high requirement by a plant for a nutrient results in a

large concentration gradient, favoring a high rate of ion diffusion from soil solution

to the root surface. According to Havlin et al. (1999) and Marschner (1995), most P

and potassium (K) uptake in corn is governed by diffusion to the root surface over

distances of 0.02 cm and 0.2 cm, respectively.

Of the three nutrient uptake mechanisms, root interception is most germane

for providing insight into how plants can contend with limited bioavailability. Root

interception relies on direct physical contact between plant root surfaces and soil

solids. Havlin et al. (1999) postulate that root interception occurs via a “contact

exchange” mechanism in which overlapping oscillation volumes between ions (e.g.,

H+) attached to root hair surfaces exchange with ions held on the surfaces of clay

particles and organic matter. Root interception is enhanced both by the growth of

new roots and by mycorrhizal infection (caused by a fungal symbiont whose hyphae link root tissue to soil pores) because these processes allow exploitation of

greater soil volumes.

Thus far, only movement of soil components to the root surface has been

discussed. As a rule, there is great discrepancy between the mineral, nutrient concentrations in soil solution, and the mineral nutrient requirements of plants; therefore, the mechanisms by which plants bring minerals into their tissues must be

selective. The selective uptake mechanisms used by plants can include (Marschner,

1995): unique cation-binding properties of rhizodermal cell walls; and ion channels,

transmembrane pumps, or protein carriers residing in tonoplasts and/or plasma

membranes within plant cells. Marschner (1995) has discussed plant physiological

and anatomical mechanisms that confer such selectivity. Details are beyond the

scope of this report. Nonetheless, it should be recognized that results of soluteuptake studies performed on both lower and higher plants demonstrate the following characteristics (Marschner, 1995): (1) selectivity—preferential uptake and/or

exclusion; (2) accumulation—concentrations of solutes can be much higher in the

cell sap than in soil solution; and (3) genetic variability—distinct differences in

solute-uptake traits among different plant species.

7.3 Examples of Nutrient Uptake by Plant Roots

This subsection of the report directs readers towards two examples of the

detailed mechanisms by which plant physiology and soil-solution chemistry interact to govern solute uptake by plant roots. The first example, phosphorus, has a

fixed oxidation state (+5) but occurs in anionic forms in soil solution that form

insoluble minerals (e.g., hydroxy-apatite). These anionic forms can bind and/or

coprecipitate toxic metals such as Ni, Pb, and U. The second example, Fe, undergoes redox reactions, and has several competing biotic and abiotic reaction pathways in the soil habitat.
53

7.3.1. Phosphorus (P)

According to Barber (1995), the phosphorus content of soils can vary from 0.02

to 0.5 percent, with an average of approximately 0.05 percent. The phosphorus pool

is divided into four general categories: (1) P as ions and compounds in soil solution; (2) P adsorbed onto surfaces of inorganic constituents; (3) P minerals, both

crystalline and amorphous; and (4) P as a component of soil organic matter. In soil

solution, the dominant chemical species are usually either H2

PO4

-

 or HPO4

2-, depending upon ambient pH. This solution-phase P equilibrates rapidly (labile P) and

slowly (nonlabile P) with the adsorbed P pool. Barber (1995) indicates that controversy exists over the amount of P that may be adsorbed on the surfaces of such soil

constituents as iron and aluminum oxides versus the amount that is precipitated as

discrete mineral forms. Many potential P minerals may exist in soil. In basic soils

(above pH 7), calcium phosphates (e.g., fluoro-apatite, hydroxy-apatite) should be

dominant, while in acid soils (pH<7) iron (e.g., strengite) aluminum phosphates

(e.g., variscite) are the dominant forms. Solubilities of the pure crystalline minerals,

and their pH dependence can be used to predict the behavior of P in soil. However,

kinetic barriers to reaching equilibrium in soil, formation in impure minerals, and

metastable states limit the usefulness of such predictions (Barber, 1995).

One-half or more of the total P in many surface soils may be in the organic

form—principally as esters of orthophosphate (e.g., inositol phosphates, phospholipids, nucleic acids) (Havlin et al., 1999; Barber, 19995). The release of organic

P into soil solution (where it can react with other soil constituents or move to the

root and be absorbed) is controlled by the rate of soil organic matter decomposition. Barber (1995) states that, in temperate climates, an organic matter mineralization rate of approximately 2 percent per year is to be expected.

7.3.2. Iron (Fe)

In contrast to phosphorus, iron is cationic and undergoes oxidation/reduction

reactions. In soil, iron can exist in the ferrous (Fe2+) and ferric (Fe3+) states. This

variability in oxidation state, combined with a wide range of complexation and

precipitation reactions (all dependent on pH, redox, and concentrations of soil

anions such as hydroxide and carbonate), contributes to the intricacies of iron’s

behavior in soil (McBride, 1994; Sposito, 1989). In fact, there are more than nine

different possible iron soil minerals (Schwertmann and Taylor, 1977). Iron minerals

commonly found in soils are the oxides or hydroxides (goethite, hematite,

lepidocrocite, maghematite, magnetite, and amorphous forms). According to Barber

(1995), the latter is possibly the most significant form in supplying iron for uptake

by the plant. Although Fe2+ is more soluble than Fe3+, both oxidation states have

such a strong tendency to form insoluble minerals so that total Fe in soil solution is

insufficient to meet plant nutritional requirements. Clearly mechanisms besides
54

simple uptake of iron from soil solution exist—otherwise, plants grown on almost

all soils would be Fe deficient (Havlin et al., 1999).

There are three primary mechanisms by which plant roots actively increase the

availability of iron in soil (Hartwig and Loepert, 1993): (1) exudation of protons to

locally solubilize Fe at low pH, (2) release of reducing agents that convert highly

insoluble Fe3+ to more soluble Fe2+, and (3) by exudation of siderophores (organic

Fe-chelating agents that bind and mobilize Fe3+). Plants have been classified as

either Fe efficient or Fe inefficient. The Fe-efficient plants respond to iron deficiency by releasing protons and reductants into the rhizosphere where the root

hairs absorb the iron, primarily as Fe2+ (Barber, 1995). After transport to the protoxylem and metaxylem, the Fe2+ can be oxidized back to Fe3+, chelated with citrate and

moved into the xylem for transport to the plant shoot (Barber, 1995). According to

Tagaki (1993), the scheme just described is “Strategy I” for overcoming iron deficiency; it can involve enzymatic reduction of Fe3+ to Fe2+ and is prevalent in angiosperms (seed producing plants).

The third mechanism for enhancing iron uptake involves production of

siderophores in the roots of grasses. In Tagaki’s (1993) scheme, these are “Strategy

II”-type plants. The siderophores are organic chelating agents with such high ironbinding affinities that they solubilize the Fe3+ from soil minerals. Once in solution,

movement of the iron-chelate complex is via mass flow or diffusion back to the root

tip region (Barber, 1995), where the Fe is released from the siderophore by being

reduced to the ferrous form, and absorbed by the plant root. Transporter-mediated

internalization of Fe-siderophore complexes (Roemheld and Marschner, 1986) also

has been reported. After being freed following uptake by the root, the siderophore

is released to the soil solution where it can again solubilize ferric iron from the

mineral phase.

7.4 Contrasts Between Substrate Uptake Mechanisms in

Microorganisms and Plants

Soil is inhabited by many types of microorganisms [bacteria, fungi, algae, protists (e.g., protozoa) and viruses]. Most of the discussion below will be restricted to

bacteria because of their relevance to bioremediation. There are many major structural, evolutionary, and physiological distinctions between plants and microorganisms (e.g., Madigan et al., 2000; Lengeler et al., 1999). Only a few of these distinctions are relevant to the uptake of substrates in the soil habitats. Implicit in the

previous discussion of root-uptake mechanisms was the fact that the materials

brought into the plant tissue from soil are used in assimilative metabolism, e.g.,

incorporated into plant structural elements (e.g., cell walls, enzymes, and other

cytoplasmic components). Microorganisms have assimilative nutritional needs that

are similar to plants. However, unlike plants, the majority of soil microorganisms are

nonphotosynthetic—their physiologies rely on uptake and metabolism of organic
55

carbon, not CO2

. This means that soil microorganisms must bring both inorganic

and organic substances across cell walls and cell membranes to grow.

Microorganisms live in intimate contact with soil solids. This direct contact

with soil solids (analogous to the “root interception” uptake mechanisms for plants)

is an important means by which bacteria and fungi acquire insoluble organic and

inorganic substances from the extracellular milieu. In addition, microorganisms (especially bacteria) also have dissimilatory nutritional needs—the primary example of

this is in the generation of ATP using many different final electron acceptors. Like

plants, aerobic microorganisms use oxygen, a sparingly water-soluble gas that

readily diffuses across cell membranes, in respiratory processes to generate ATP.

Because solution-phase oxygen readily passes into cells, one would not expect

enhanced-transport systems for oxygen to be present in the microorganisms. Specific adaptations aimed at overcoming solid-phase bioavailability limitations of the

highly soluble alternative electron acceptors (e.g., nitrate and sulfate, used by

anaerobic microorganisms) also would not be expected. However, in anaerobic

settings, alternative bacterial final electron acceptors can include the highly insoluble oxides of iron and manganese. The mechanisms by which bacteria metabolize these latter insoluble compounds are particularly germane to bioavailability

issues.

Because of the importance of solid phase biogeochemical processes, the mechanisms of electron transfer by microorganisms to poorly soluble minerals for anaerobic respiration is subject to intense, ongoing study (Lovley 2000; Newman and

Kolter, 2000; Seeliger et al., 1998). Currently accepted potential mechanisms (Lovley,

2000) include extracellular electron shuttles either involving naturally occurring

humic substances (Lovley et al., 1996) or bacterially excreted quinones (Newman

and Kolter, 2000) and direct enzymatic reduction of solid oxides at the cell surface

via membrane-bound enzyme systems that span the inner and outer cell surfaces.

Thus, it is important to recognize that in bacteria (prokaryotes), the electron

transport chain used in ATP generation is located in the plasma membrane at the

periphery of the cell (Ehrlich, 1995; Lengler et al., 1999). This contrasts with eukaryotic cells (such as plants) in which electron transport occurs internally in special

organelles called mitochondria. Because of their specialized cellular architecture,

bacteria endowed with appropriate oxido-reductases (enzymes that transfer hydrogen atoms or electrons) in their cell envelope are able to oxidize or reduce insoluble

substrates that cannot be taken into these cells (Ehrlich, 1995; Lengler et al., 1999).

Such substrates include elemental sulfur, iron sulfide, iron oxides, and manganese

oxides. Because the essential enzymes that recognize and act on insoluble substrates are located in exterior regions (e.g., the periplasmic space, the plasma

membrane, or the outer membrane), bacteria are uniquely capable of metabolic

activities that render extracellular materials “bioavailable.” Utilization of endogenous (quinones) and exogenous (humic) substances also can facilitate physiological reactions between bacteria and “nonbioavailable” substrates.
56

7.5. Membrane Transport in Microorganisms

 [The following discussion is derived largely from Lengeler et al., 1999.] Cell

membranes define the internal metabolic system of living cells. Membranes act as a

physical barrier between microorganisms and their environment. Membranes also

carry out essential metabolic, sensory, and reproductive functions. Lengeler et al.

(1999) state that: “The cytoplasmic membrane of bacterial cells consists of a phospholipid bilayer, which functions as a permeability barrier for most solutes. Polar

solutes (e.g., carbohydrates) and charged molecules (ions, carboxylic acids, amino

acids) have a very low rate of passive flux across lipid bilayer membranes. In addition to some small solutes and molecules (such as water, ethanol, ammonia, or

oxygen), only apolar (hydrophobic) compounds (e.g., phenylamine, glycerol, or

fatty acids) pass across lipid bilayers. Bacteria, however, need to transport solutes

at high rates across the cell wall and cytoplasmic membrane for growth and metabolism. The solute transfer across bacterial membranes is mediated by specific membrane proteins called transporters, transport systems, carriers, or, in analogy to

enzymes, permeases. By means of these proteins, the transfer rate across bacterial

membranes can be significantly increased... The presence and activity of carrier

systems in a relatively impermeable membrane is the reason for observed concentration gradients of solutes across the cell membrane. These range from 10-30 fold

(external/internal Na+ and internal/external K+) to more than 10,000 fold (external/

internal free Ca2+), up to a 200,000-fold accumulation of some solutes in the cytoplasm (e.g., maltose and particular amino acids).”

Lengeler et al. (1999) have further summarized the four key mechanisms of

solute transport into prokaryotic cells: (1) diffusion (with or without facilitation by

a permease); (2) secondary transport, in which a permease combines with an electrochemical gradient to adjust concentrations across the membrane; (3) primary

transport, in which solute transport is directly coupled to chemical or photochemical reactions; and (4) group translocation, in which the solute consumes ATP

through phosphorylation reactions inside the membrane. Diffusion-based transport applies to gases (O2

, CO2, NH3), small molecules (water, ethanol), and also

hydrophobic (lipid soluble) molecules such as aliphatic (e.g., butanol) or aromatic

molecules (e.g., benzene). A number of solutes are membrane permeable in their

uncharged form (e.g., protonated organic acids) but do not pass through membranes if charged.

7.6. Uptake of Insoluble Organic Substrates by Microorganisms: Wood

High molecular weight organic polymers such as cellulose and lignin are insoluble. These would appear to be “unavailable” for metabolism, yet they are widespread in nature and are biodegradable. Lengeler et al. (1999) have described some
57

general physiological adaptations that allow microorganisms to exploit the diversity

of naturally occurring polymeric organic substrates. The microbial world features a

wide variety of usually hydrolytic extracellular enzymes that cope with crystallinity,

low solubility, and the association of polymers with solid-phase materials.

Cellulose is the most common organic substrate in nature. It is a linear polymer

of 100 to 100,000 glucose units linked by ß-1,4 bonds. Cellulose chains form intramolecular and intermolecular hydrogen bonds that allow rigid insoluble fibrils to

form. When attacked, this insoluble substrate is eventually converted to soluble

glucose monomers and dimers (cellobiose) that, after cell entry, are intracellularly

metabolized. Attack of cellulose occurs through the action of a battery of extracellular exoenzymes. These enzymes fall into two general categories: endoglucanases

(that sever the linear cellulose polymer, presumably by random attack at sites within

the chains) and exoglucanases that sequentially remove dimeric cellobiose units

from one end of the chain) (Bayer and Lamed, 1992). According to Lengeler et al.

(1999), the binding of exoenzymes to surfaces of the microbial cells, the attachment

of the enzyme to the cellulose, and possibly the uptake of polymeric substrate into

the cell periplasm (between cell wall and cytoplasmic membrane) minimizes both

dilution of the enzymes and losses of hydrolyzed cellulose subunits to competing

microorganisms.

The anaerobic bacterium, Clostridium thermocellum, has evolved a highly

specialized multienzyme cellulase complex, termed a cellulosome. This extremely

high molecular weight structure (>2 million daltons) is positioned at the interface

between the microorganism and the cellulosic substrate. Strong adhesion between

C. thermocellum and the cellulases is required prior to cellulose degradation (Bayer

and Lamed, 1992).

Natural structures like wood are highly complex—consisting of many types of

polymers and monomers in a rigid matrix often bound by lignin. Lignin molecules are

not susceptible to hydrolytic cleavage (Lengeler et al., 1999). Rather, lignin metabolism (rare in bacteria, more common in fungi) requires an oxidizable substrate (e.g.,

glucose) and O2

 to form hydrogen peroxide that activates ligninase enzymes. Thus, a

mingling of different microorganisms, each with their own unique properties, contributes to successful biodegradation of naturally occurring, wood-derived materials.

7.7. Phosphorus and Iron Uptake by Bacteria

Phosphorus is a growth-limiting element under many environmental conditions—soil solution concentrations are in the micromolar range, while intracellular

concentrations are approximately 1,000-fold higher (Lengeler et al., 1999). Inorganic P (orthophosphate) is the preferred P source in bacteria, as its presence

suppresses expression of genes involved in the acquisition of alternative (organic)

P sources. Inorganic P can enter the bacterial cell either through low-affinity (per-
58

mease-based diffusion) or high-affinity (ATP-dependent) transport systems. The

latter are activated by inorganic P limitation.

Microorganisms (especially bacteria and fungi) use mechanisms much like plants

to meet their metabolic demand for insoluble iron in soil. According to Hofte (1993), with

only a few exceptions, all aerobic and facultative anaerobic microorganisms that have

been critically examined produce iron siderophores. These fall into three categories

based on molecular structure: catechols, hydroxamates, and carboxylic acids. The

overall physiological function of microbial siderophores resembles that described for

plants (Section 7.3.2); however, some details are distinctive. In an example described by

Page (1993), ferric iron was scavenged from soil; the siderophore-bound iron was recognized by the cell that produced the siderophore through specific receptor proteins

located in the cell membrane; the ferri-siderophore was transported into the cell and

deferrated by mechanisms that involved iron reduction or ligand hydrolysis.

7.8. Summary

Aqueous-phase soil solution represents a conduit for dynamic equilibrium and

transport processes that are governed by a combination of organic, inorganic,

biochemical, and mineralogical reactions. The soil physical and chemical environment establishes the context for metabolism, growth, reproduction, and nutrient

uptake by plants and microorganisms. Plants and microorganisms dwell in this soil

habitat whose chemical composition is vastly different from their own.

To acquire key inorganic nutrients, plant roots combine three general uptake

mechanisms (mass flow in solution, diffusion, and solid-phase interception) with

sometimes highly selective membrane transport carriers (active and passive) to

regulate the intracellular composition of cellular constituents. Microorganisms also

feature unique structural and physiological traits that allow them to contend with

the soil environment. Microorganisms differ from plants in several major ways.

Most soil microorganisms have heterotrophic nutrition—thus, they must acquire

both organic and inorganic materials from soil. Also, in prokaryotic cell architecture, membrane-associated electron transport systems reside near the cell periphery. This allows some prokaryotes to use insoluble soil components such as Fe and

Mn oxides as physiological final electron acceptors. Endogenous (quinones) and

exogenous (humic) substances also can facilitate extracellular electron transport

reactions for bacteria. Another key characteristic of bacteria and fungi is their

ability to produce extracellular enzyme systems that allow insoluble substrates,

such as cellulose and lignin, to be degraded to constituent components prior to

their uptake and intracellular metabolism. One remarkable ability, shared by microorganisms and plants, is an adaptation that allows them to contend with the unavailability of the essential nutrient, Fe. Both plants and microorganisms acquire

otherwise insoluble Fe3+ by producing highly specialized, organic chelating agents

(siderophores) that solubilize mineral Fe3+ and mobilize it to the cell surface. At the
59

cell surface, the Fe-siderophore complex is destabilized so that, as the Fe is utilized

intracellularly, the siderophore cycles back in soil solution to bind more Fe.

Key insights into the bioavailability of chemical wastes in soil may be achieved

from fundamental knowledge of the mechanisms by which microorganisms and

plants bring soil constituents into their cytoplasm. Highly specialized physiological mechanisms for cellular uptake of some essential soil nutrients have evolved in

both plants and microorganisms. Because much microbial metabolic diversity remains undiscovered (see Section 6.3), the full spectrum of microbial uptake systems

in soil has almost certainly not been described. To the degree that chemical wastes

behave as essential nutrients, one would predict that efficient contaminant uptake

could be achieved. Conversely, to the degree that the properties of chemical wastes

differ from essential nutrients, one would predict low-efficiency uptake by plants

and microorganisms.

Apolar hydrophobic organic compounds in soil should readily pass through cell

membranes; hence, they face no biological transport barriers for cellular uptake. In

contrast, efficient uptake of polar organic and inorganic contaminant compounds can

have a low rate of passive flux into cells. Mechanisms that enhance the uptake (e.g.,

permease or ATP-driven) of polar contaminant compounds could theoretically increase rates of transport into plants and microorganisms, but enhanced uptake would

not be expected unless the contaminants were fortuitously recognized by the uptake

systems.

Plants and microorganisms have developed mechanisms for uptake of materials

from solid phases in soils; thus, entry into soil solution (by dissolution, desorption,

or exchange from soil solids) is not required for acquisition, accumulation, and/or

metabolism of chemical wastes by plants and microorganisms. For those who want

to use bioremediation processes for soil cleanup, solid-phase substrate uptake

capabilities may provide grounds for optimism. However, a less optimistic argument also can be made. The soil habitat poses many constraints on the uptake of

soil constituents by plants and microorganisms. The evolution of elaborate solidphase uptake mechanisms by soil inhabitants can be taken as long-term evolutionary evidence for bioavailability limitations in soil. The subject of bioavailability

limitations and their alleviation will be presented in Section 9 of this report.
60

Section 8

Reviewing the Facts: Examining

Relationships Between Contaminant

Sequestration and Bioremediation

A recent computerized search of the peer-reviewed literature, using “bioavailability” and “bioremediation” as key words, retrieved hundreds of references that

were published over the last two decades. If the goal of this report, “a concise

compilation of current knowledge,” is to be achieved, clearly much of this literature

must be passed over, while focusing on influential research of the highest quality.

But, as will become evident in Section 8.1, even a select subset of the literature on

relationships between biodegradation of organic compounds and bioavailability provides a confusing mass of seemingly conflicting data and interpretations. This section aims to sift through, evaluate, and weigh the significance of key reports. Background information accrued in Sections 2-7 of this report and arguments developed in

Section 8.2 will serve as evaluation criteria.

8.1 Ambiguity Is the Rule: An Historical Overview of the

Impact of Solid Surfaces on Microbial Activity

There is a long, substantive history of investigations that have examined how

microorganisms respond to the presence of solids—both as particles suspended in

an aqueous milieu and as an unsaturated porous geochemical matrix. Much of this

early scholarly information was summarized by Marshall (1976). This excerpt from

his book appeared under the heading “Effect of soil particles on microbial metabolism” within the chapter entitled “Nonspecific interfacial interactions in microbial

ecology: terrestrial ecosystems,”

“Adsorption of organic substrates to soils depends on the nature

of the particulate matter, the organization of the fabric, the clay types,

and the cation status of the soils, as well as on the concentration and
61

molecular structure of the substrate. The availability (emphasis by

Madsen) of substrates to soil microorganisms may be enhanced or

reduced by the presence of soil particulate matter. This may depend

on buffering effects...or on direct sorption phenomena. If the substrates are sorbed, then availability will depend on substrate location

relative to that of the appropriate decomposer microorganisms, on

whether extracellular enzymes are involved and whether these are

sorbed, and on the configuration and juxtaposition of the substrates

and enzymes in the sorbed state. Some contradictory results have

been reported on the metabolism of specific organic substrates in

different soils, but these probably reflect the complexity of soils as

microbial ecosystems.” (Marshall, 1976)

Marshall’s (1976) comment about “some contradictory results” will appear as a

euphemism when compared to summary statements published 14 years later by van

Loosdrecht et al. (1990). In this influential review entitled “Influences of interfaces on

microbial activity,” van Loosdrecht et al. (1990) state: “Although there appears to be a

qualitative consensus that surfaces do influence bacterial metabolism, the experimental

observations are not always consistent; neither has a general explanation been advanced for this influence. The lack of experimental consistency is at least partly due to

the great variation in experimental design with respect to the nature of the solid phase,

the bacteria, the substrate, sterility, and other experimental conditions. The relevant but

disparate literature is summarized in Table 1, wherein it is confirmed that, although

generally surfaces do exert an influence, no systematic trends can be observed.”

Table 1 of van Loosdrecht et al. (1990) was entitled, “Summary of the literature

on the influence of solid surfaces on microbial behavior.” Within this table, the

following categories of observations appeared: “increased growth rate,” “decreased

growth rate,” “increased assimilation and decreased respiration rates,” “increased

respiration,” “decreased respiration,” “increased adhesion of active cells,” “higher

activity of attached cells,” “decreased substrate utilization,” “lower substrate affinity,” “change in pH optimum,” “difference in fermentation pattern,” “increase in

productivity,” “decreased mortality,” and “no effect.”

van Loosdrecht et al. (1990) offered the following concluding remarks: “The

presence of surfaces may positively or negatively (or not at all) affect microbial substrate utilization rates and growth yields. The results often depend on the nature of the

organism, the kind and concentration of substrate, and the nature of the solid surface.

In interpreting the effect of surfaces on bioconversion processes, all possible physical

and chemical interactions (e.g., diffusion, ad- and desorption, ion-exchange reactions,

conformation changes, etc.) of a given compound and its possible metabolites with a

given surface have to be considered before general conclusions can be drawn.”

Particularly germane to this report, Mihelcic et al. (1993) published a review

entitled “Bioavailability of sorbed- and separate-phase chemicals.” This compilation
62

was motivated by the view that “Hydrophobic organic compounds may be sorbed to

soils and sediments or present in a separate phase (e.g., oil and coal tar). Consequently, the effectiveness of bioremediation of soil contaminated with organics may

be affected by physicochemical processes that control phase partitioning between

solid and liquid, and subsequent solute accessibility to microorganisms. Thus, the

bioremediation of soils contaminated with hydrophobic solutes, may depend on the

rate and extent of desorption from a solid surface or dissolution from a separate

phase.” Mihelcic et al.’s (1993) assessment of the literature began with the above

premise and thoroughly compiled information addressing: field and laboratory observations, associations of microorganisms with surfaces, microbial utilization of

sorbed and separate-phase substrate, effects of surfactants on biodegradation, and

models that combine sorption and biodegradation. Throughout their review, Mihelcic

et al. (1993) emphasized the complexity of soil-based experimental systems, potential

experimental artifacts, the range of microbial adaptations to biodegradation reactions,

equivocal experimental results, and the need to carefully evaluate data.

The first paragraph of Mihelcic et al.’s (1993) summary section reads as follows: “The investigation of the physical, chemical, and biological parameters affecting the biodegradation of sorbed- and separate-phase contaminants in soil-water

systems is admittedly complex and presents an investigator with numerous experimental challenges. The analysis of the data reported from both natural and engineered systems clearly shows that the results obtained by any individual investigator are highly dependent on the target substrate being examined, the identity and

concentration of the organism(s), and the nature of the sorbent. To date, there seem

to be few common principles that govern the rate of degradation of a selected pollutant. However, the diversity of the sorbent-water systems examined and the variations

in experimental design employed by investigators renders the existing studies virtually impossible to compare. Models to predict experimental observations are sophisticated in their approach to describing mass transfer kinetics yet are simplistic

in their approach to describing biodegradation and cell growth.”

Thus, ambiguity is the rule when examining the influence of solid surfaces on

microbial activity and the relationship between bioavailability and biodegradation

in soils. It is within this established, yet confusing, scholarly context that this

report aims for progress.

8.2 Selection and Justification of Criteria for Identifying the

Highest Quality Investigations Pertinent to

Bioavailability and Bioremediation

This report has been designed to equip a reader with tools to decipher a vast and

ambiguous literature on bioavailability and biodegradation. Previous sections of this

report serve as a prelude that establishes the foundations for this eighth section:

Section 2 defined bioremediation and how both microorganisms and plants can influ-
63

ence soil contaminants; Section 3 defined bioavailability; Section 4 surveyed results

of bioremediation field projects; Section 5 examined mechanisms of persistence, as

illustrated especially by soil organic matter; and Sections 6 and 7 of this report have

provided a glimpse into what is and is not known about geosorbent-contaminant

interactions and the mechanisms by which microorganisms and plants acquire contaminants in soil. This section of the report directly addresses the relationship(s)

between the bioavailability of chemical wastes in soil and their bioremediation.

Previous sections of this report have emphasized the multifaceted nature of

“bioavailability.” It is not an inherent property of substances under examination.

Rather, it reflects the response of experimentally defined biological systems to many

integrated processes. Bioavailability is an emergent, malleable trait that is inferred

from detailed, three-way interactions between geosorbents, biota, and chemical wastes.

The “malleable” aspects of bioavailability stem from the fact that when scientific

investigators design their experiments, intentional or unintentional choices are made.

Each choice influences the outcome of the experiments. Which geosorbent will be

used (e.g., model system or environmental sample, agricultural or industrial soil,

freshwater sediment, of high organic matter or high clay, contaminated or pristine)?

Which chemical waste will be studied (e.g., organic or inorganic; soluble or insoluble;

volatile; toxic; hydrophobic or hydrophilic, freshly added or historically in place)?

Which of the many possible biota (e.g., pure microbial cultures or mixed cultures,

monocotyledonous or dicotyledonous plants, soil enrichments, microbial communities from contaminated or pristine soil) will be included? How will the geosorbents,

chemical wastes, and biota be pretreated, mixed, and incubated? Are the experimental

choices for system components and incubations realistic? How environmentally significant are the incubation conditions, the biota, and the physical/chemical states of

the contaminants? What artifacts may be created in the data-generating systems or by

the experimental designs and treatments? Can broad, sweeping principles and conclusions be drawn from particular sets of specific, narrowly defined experiments?

The above types of questions and their answers begin to impinge on philosophical issues about the scientific method and what it can achieve. These philosophical concerns are clearly beyond the scope of this report. Nonetheless, based

on the literature reviews cited in Section 8.1, it is clear that, to some degree, the

outcome of experiments examining relationships between bioavailability and biodegradation are arbitrary—virtually any outcome can be achieved when particular combinations of geosorbent, contaminant, and microorganisms or plants are

selected. Given this possibility, interpretation of the multitude of reports must be

grounded in the following characteristics:

(1) realism and environmental relevance,

(2) absence of experimental artifacts, and

(3) consistency of results.
64

Without such a foundation, many of the existing studies on how bioavailability

influences biodegradation provide confusing or misleading results. Moreover,

these criteria should help identify investigations whose results apply to society’s

real-world problems in real-world situations.

Table 2 lists the three characteristics selected for evaluating reports to be scrutinized in this section of this report (see Section 8.3 and Table 4). To identify investigations of the highest quality (realism, absence of artifacts, consistency; first column of

Table 2), seven specific criteria for evaluating experimental procedures will be applied

(middle column, Table 2). The first criterion is that the investigation begin with field

samples and field observations of the persistence of contaminants. This guarantees

the environmental relevance of a study. The second criterion recognizes that hypothesis testing to explain field persistence often requires manipulation of geosorbents

and/or biota in the laboratory—but this should be done in a realistic manner, with

minimal departure of experimental system from field conditions. The third criterion

(realistic biota) recognizes that the microbial world is extremely diverse and responsive to laboratory-imposed conditions (Madsen, 1996, 1998), thus, misleading assays

of biodegradation and/or bioavailability may occur if the biotic component of experimental systems is unlike that found in real-world settings. The fourth criterion acknowledges that exploring the basic physiological, biochemical, and genetic details

of biological processes requires the use of single organisms. This is the major strength

of pure-culture studies. Yet, single-organism studies need to be tempered and interpreted with realism—extrapolation from pure cultures in the laboratory to unknown

complex microbial communities is usually unwise.

The seventh criterion in Table 2 simply expects that meritorious investigations

build on previous results in a way that clearly makes progress toward answering

significant scientific and technological questions. Given the many independent

variables that must be confronted in biodegradation and bioavailability studies

(e.g., Sections 3.1, 6.1, 8.1), discrepancies should be expected; nonetheless, the

validity of results from studies that contradict one another (especially if generated

by the same laboratory) should be questioned and not be considered of the ‘highest quality.”

Entries 5 and 6 in the central column of Table 2 describe criteria for reducing the

credence given to biodegradation/bioavailability investigations based on artifacts

that may be imposed on the geosorbent matrix under study. This is, admittedly, a

controversial issue because compromise is often unavoidable in experimental procedures. If an experimental method is imperfect, do the imperfections erase all validity of the results? The answer is “probably not.” Nonetheless, it is indisputable that

the physical, chemical, and biological status of geosorbent matrices drastically

influence the results of biodegradation and bioavailability assays occurring therein.

Therefore, for the purposes of this review, it is desirable to judge the highestquality investigations as those that avoid drastic alteration of the geosorbent matrix. Useful, supportive, confirmatory information can certainly be generated from
65

TABLE 2 Criteria for Identifying the Highest Quality Investigations Pertinent to Bioavailability

and Bioremediation (See Section 8.2 for Rationale)

Desired Characteristic

Criteria Applicable to

Bioavailability/Biodegradation Reports Rationale

Realism and environmentalrelevance

1. Initial data should be obtained from field

observations of contaminants

1. There is such an abundance of real-world contamination problems, that contrived, laboratory fabrications

of contamination scenarios may be superfluous.

2. Subsequent (e.g., laboratory-based) data

should be obtained without drastic chemicalor physical alteration of geosorbents orcontaminants

2. After documenting the field behavior of contaminants,

hypothesis testing may need to occur via sampleincubation in the laboratory. When brought in fromthe field, neither the geosorbents nor contaminantsshould be unduly altered, unless warranted byexperimental objectives.

3. Biota used in bioassays should represent

those that are native to the field site

3. When the experimental objective is to reveal

biodegradation and/or bioavailability characteristics ofreal-world soil microbial communities, then appropriate, representative biota should be used in theexperiments.

4. When pure cultures are examined, their

relevance and the experimental conditionsselected (such as physiological status)should be defended for their environmentalrelevance

4. If refined hypotheses warrant examination of pure

microbial cultures, then care should be taken to relateexperimental variables of pure-culture manipulationsback to specific hypotheses and their relevance tofield conditions.

Absence of experimentalartifacts

5. Avoidance of soil sterilization procedures 5. Chemical and physical treatments that allow

sterile (abiotic) controls to be examined drasticallyalter the properties of geosorbents.
66

Desired Characteristic

Criteria Applicable to

Bioavailability/Biodegradation Reports Rationale

6. Avoidance of contaminant addition via flash

evaporation from volatile carrier solvents

6. When contaminants are artificially added to

geosorbents in highly volatile organic solvents, therapid evaporation of the solvent may leave thecontaminants in a physical/chemical state and/orlocation that is unlike that of contaminants in fieldsoils. Also, carrier solvents are likely to remain in thegeosorbents, unless they are subject to extensivemeasures that encourage evaporation.

Consistency of results 7. Data from one set of experiments should

support and not conflict with anothersubsequent similar set of experiments

7. Although variability in physical, chemical, and

biological properties of real-world samples has become an accepted means of “explaining” conflictingexperimental results, resorting to such excuses isundesirable. Unresolved inconsistency may detractfrom understanding biodegradation and bioavailabilityissues.

TABLE 2 Criteria for Identifying the Highest Quality Investigations Pertinent to Bioavailability

and Bioremediation (See Section 8.2 for Rationale) (Continued)
67

sterile (entry #5) or spiked (entry #6) geosorbent matrices—but in this report, such

investigations will be considered secondary, not primary, information sources.

8.2.1. Further Scrutiny of Artifacts That May Be Caused by

Soil Sterilization and Contaminant Addition in Organic

Solvents

Soil sterilization (entry #5, central column of Table 2) is required for the preparation of abiotic controls of many experimental designs (Brock, 1978). Yet “sterilization

procedures result in some degree of alteration of soil chemical and physical properties” (Wolf and Skipper, 1994). Measures that have been used to produce abiotic

controls include chemicals that inhibit microbial metabolism (e.g., acid, azide, mercuric chloride) and sterilization (e.g., by autoclaving or γ-irradiation). All of these treatments do far more than simply eliminate biological activity. Cawse (1975) summarized

much of the early literature on the influence of autoclaving and γ-irradiation procedures on soil properties and the activity of reinoculated microorganisms. Influences

include: solubilization of organic matter, release of electrolytes, formation of inhibitory substances, lower pH, and many detailed alterations in the chemical composition

of soil solution. In preparing abiotic controls using low organic matter subsurface

sand, Ball (1989) found that overnight heating or overnight cooling within an autoclave tended to increase the sand’s sorptive properties, but the effect could be avoided

if the sand samples were cooled rapidly after autoclaving. Autoclaving of this same

sand also was found to increase its affinity for binding hexachloroethane (Criddle et

al., 1986; MacKay et al., 1986). Xia (1998) also observed varying effects of autoclaving on phenanthrene sorption by several different subsurface materials. In a study

designed to address PAH bioavailability in soils, Sandoli et al. (1996) discovered that

γ-irradiation-induced changes in geosorbent properties prevented 14C-phenanthrene

mineralization in the presence of otherwise active microorganisms. No evidence for

toxicity was evident in these γ-irradiated soils. Although gamma irradiation was found

to increase the geosorbent’s affinity for phenanthrene by 32 percent, a full explanation of the γ-irradiation-induced attention of the geosorbent was not obtained (Sandoli

et al., 1996). Unfortunately, the impacts of inhibitors and sterilization procedures on

physical, chemical, structural, and sorptive properties of soils are probably as variable as the properties of soils themselves.

Discrepancies in the properties of “freshly added” versus “field contaminated”

pollutants in geosorbent matrices have been noted in many investigations (e.g.,

Alexander, 1999; Burford et al., 1993; Hatzinger and Alexander, 1995; Steinberg et

al., 1987). However, seldom have the two distinctive issues pertinent to these

discrepancies been discussed. The issue that has perhaps received greatest attention to date is that the physical, chemical, and biological behavior of freshly added

contaminants may change over time. Therefore, freshly added contaminants may

not mimic the behavior of “aged material.” It has been postulated that the freshly
68

added contaminant pool may not have had sufficient time to diffuse into micro- or

nano-pores or be absorbed into organic matter (e.g., Alexander, 1995; Luthy et al.,

1997; Reid et al., 2000). Thus, an “aging effect” has been used to argue for kinetically constrained, time-dependent sequestration of environmental contaminants.

The second related but distinctive issue for freshly added contaminants (entry

#6, central column of Table 2) is that the way they are added to geosorbents may

leave them in locations and in physical/chemical states that differ markedly from

their field-derived analogs. When small volumes of contaminant-containing volatile solvents are added (“spiked”) to soil and allowed to evaporate, then crystals of

the contaminant solute may be left behind in the soil matrix. It is unlikely that these

crystals, even if later distributed uniformly throughout the soil via mechanical

mixing, will initially interact with the soil matrix in ways that mimic the interactions

found in field sites (unless the field sites happen to be contaminated via similar

flash-evaporation-type procedures; S. Hawthorne, personal communication). Not

only are the flash-evaporated spiked contaminants likely to be in states and associations and locations unlike those in the field, but complete removal of the carrier

solvent requires heat and vacuum because the solvent’s effective boiling point can

be raised in small pores (J. Pignatello, personal communication). Without such

directed effort to remove residual solvents from spiked soil, they may have unanticipated effects on soil organic matter (e.g., swelling that may be irreversible; J.

Pignatello, personal communication), contaminants, and/or microbial processes.

Burford et al. (1993) carried out an extensive comparison of the extraction efficiencies of freshly added, deuterated PAHs with corresponding nondeuterated PAHs

from historically contaminated soil and sludges. Burford et al’s. (1993) data consistently showed that freshly added PAHs (dissolved in dichloromethane and flash

evaporated) were far more extractable than the native PAHs. The discrepancy between freshly added and long-aged PAHs was greatest for the low molecular weight

PAHs (e.g., naphthalene). Burford et al. (1993) concluded that “it is experimentally

impossible to reproduce the environmental conditions that occur during deposition of pollutants in real-world samples.”

This comment by Burford et al. (1993) is particularly germane to the interpretation

of data describing “aging” of contaminants in geosorbents. Flash-evaporated surrogate contaminants may be dispersed in laboratory-incubated soil experiments. However, once in place, the contaminants will undergo dissolution, diffusion, partitioning,

and other reactions governed by a combination of contaminant properties, the soil,

and the incubation conditions. The many investigations describing “aging effects,”

noted above, have documented time-dependent alteration in biological and/or chemical properties of the contaminants. These investigations have not shown that the

time-dependent (perhaps asymptotic) change approaches the same endpoint that is

found in field-contaminated soils. After all, the real-world weathering processes (water infiltration, freeze/thaw, wind, sunlight, etc.) that occur in uncontrolled field sites

are vastly different from the typical conditions found in laboratory incubations. It is
69

possible that contaminants spiked with geosorbents and subsequently aged in the

laboratory eventually attain the same physical, chemical, and biological properties as

those found in the field. Indeed, “field contaminants” probably occur in a broad

spectrum of states, and are almost certainly neither uniform nor static. However, it is

reasonable to presume that both the rate at which the “field state” is asymptotically

approached, and the asymptote itself, are influenced by the condition of the geosorbent

and the incubation. To the degree that these two variables deviate from real-world

conditions, the state of laboratory-aged contaminants also may deviate from those

found in real-world soils. Currently, there is not enough information addressing the

above-described uncertainties about sterilization, spiking, and aging procedures

and their impact on contaminants in geosorbents. Until the relationships between

field- and laboratory-contaminant disposition in soils are better understood, it will

remain a challenge to interpret laboratory-based experiments that attempt to test

hypotheses about field processes using sterile, spiked soils.

8.3 Scrutinizing Selected Investigations Describing the

Bioavailability of Contaminants and Their

Biodegradation

To some degree, M. Alexander has become an advocate for incorporating

bioavailability considerations into regulatory and/or toxicological interpretation of

environmental contamination. Therefore, it seems appropriate to examine and scrutinize his recent, relevant publications. In 1995 and 1997, Alexander published critical

reviews that advanced the idea that contaminant bioavailability in soil may be the

most relevant criterion for “assessing toxicity, determining risk, and establishing

meaningful regulations for clean up of sites containing hazardous wastes.” Alexander

presents six lines of evidence to support the proposition that organic compounds

become increasingly sequestered (less bioavailable) as their residence times in soils

increase. These lines of evidence appear in Table 3, along with alternative interpretations of the “evidence.” The alternative interpretations (second column) for evidence

presented in the first column of Table 3 simply apply the background knowledge and

principles that have been established in earlier sections of this report. Although all of

the lines of evidence are plausible, they also need to be carefully reexamined for the

possibility of alternative interpretations and also for the quality of supporting experimental data. As indicated in the second column of Table 3, there are many potential

weaknesses in the experiments supporting the six lines of evidence, including: (1)

many factors besides diminished bioavailability can contribute to the long-term persistence of pesticides in soils; (2) observations in the field combined with laboratory

experiments (using soils that have been contaminated for extended periods to examine the cause of contaminant persistence) may be the most robust type of evidence,

given the quality of the data; (3) laboratory aging studies may be rich in artifacts,

especially if sterile, spiked soils are used (see Section 8.2.1); (4) chemical extraction

experiments that clearly demonstrate distinctions between aged and freshly contami-
70

TABLE 3 Lines of Evidence for Time-Dependent Decline in Bioavailability of Organic Soil Contaminants

(from Alexander, 1995, 1997) and Possible Alternative Interpretations

Evidence Alternative Interpretations

(i) A “hockey stick”-like profile of concentration versus time for

pesticides in field soils. The initial rapid loss of pesticides,followed by a plateau, suggests increased sequestration overtime.

(1) Most data were from chlorinated pesticides that are not readily

biodegradable because they do not serve as carbon and energysources for microorganisms. Alternative hypotheses forexplaining persistence of pesticides (such as lack of cosubstrate,simple first order kinetics, nutrient limitations, other limitingenvironmental factors) were discussed in Sections 4.0 and 5.2.Without considering all alternative hypotheses, the“bioavailability” hypothesis cannot be adequately tested.

(ii) Laboratory studies of field samples contaminated with

pesticides and PAHs in which native soil microorganisms wereable to biodegrade a freshly-added, but not an “aged” pool ofthese compounds.

(ii) This evidence can be very convincing (see Table 4)

(iii) Laboratory “aging” studies in which the degree of biodegradation by a microbial inoculum was governed, inversely, by the

duration of prior contact between organic chemicals and sterilesoil.

(iii) Conditions for laboratory “aging” may be artifactual, especially

if sterile soils and contaminants are added via flash-evaporatedorganic solvents (see Section 8.2).

(iv) Chemical extraction assays showing that the degree of

sequestration of organic compounds in soil is proportional tothe duration of compound-soil contact.

(iv) This evidence may or may not be convincing (see Section

8.2.1). A change per se is not adequate. The change in

contaminant extractability must be toward an environmentallyrelevant endpoint.

(v) Analysis of the kinetics of contaminant desorbtion from soils

that suggested the presence of recessed diffusion-limitedbinding sites within pores of soil particles or within organicmatter phases.

(v) This evidence may or may not be very convincing. The test

systems are highly complex and variable, hence, susceptible tomany alternative interpretations (see Sections 6.3 and 6.5).Validity depends on details of specific procedures in specificinvestigations.
71

Evidence Alternative Interpretations

(vi) Bioassays showing that chemicals residing in soil have reduced

toxicity and this reduction may be proportional to thechemical’s soil residence time.

(vi) This evidence may or may not be very convincing. The test

systems are highly complex and variable, hence, susceptible tomany alternative interpretations (see Sections 6.3 and 6.5).Validity depends on details of specific procedures in specificinvestigations, the bioassay, and the specific mechanism oftoxicity.

TABLE 3 Lines of Evidence for Time-Dependent Decline in Bioavailability of Organic Soil Contaminants

(from Alexander, 1995, 1997) and Possible Alternative Interpretations (Continued)
72

nated soils are insightful, though scrutiny of spiking and other procedures is warranted (see Section 8.2); (5) macroscopic observations of the interactions between

contaminants and geosorbents have many varied conceptual interpretations (Sections 6.1 and 6.5); and (6) diminished contaminant toxicity in the presence of soil

solids has many alternative explanations, depending on the bioassay, the soil, and

the mechanism of toxicity.

The quality of experimental data (e.g., field observations, type of model systems, potential experimental artifacts; general conformity to the seven criteria in

Table 2) is of crucial importance in evaluating all “lines of evidence” for or against

a reduction of contaminant bioavailability in soils. Another major consideration is

whether or not “statistically significant” differences between experimental treatments reported in the peer-reviewed literature have real-world ramifications. Baveye

and Bladon (1999) pointed out that small, though statistically significant short-term

treatment differences, may be moot when extrapolated to real-world timeframes of

months to years.

A final, key element to include is provided by an organic substance whose

persistence is well established—soil organic matter. How can the behavior of soil

organic matter enlighten our understanding and expectations for the behavior of

chemical wastes in soil?

Table 4 contains a compilation of some of the most influential and/or recent

experimental studies that have been published to date on the relationship between

bioavailability of organic compounds and bioremediation. The columns in Table 4

(Goals, Compound, Geosorbent, Approach/Methods/Data, Conclusions, Methodological Weaknesses) provide a format to prioritize and evaluate a study using the

criteria established in Table 2.

The first entry in Table 4 (Steinberg et al., 1987) is perhaps the most convincing

study to document the relationship between an aged organic chemical waste and its

availability for bioremediation. Unlike the majority of entries in Table 4, the study by

Steinberg et al., (1987) suffers from no obvious methodological weaknesses—it satisfies all seven criteria in Table 2 for high-quality data. Steinberg et al. (1987) began

with field observations that the fumigant, ethylene dibromide (EDB), had persisted in

field soils for an unexpectedly long time. Subsequent laboratory experiments used

chemical, physiological, and physical measurements to prove that aged, field-applied

EDB was not biodegraded by the native soil microbial community; yet, freshly added

radiolabeled EDB was readily metabolized. Furthermore, measurements of the rate of

release of the aged EDB from soil revealed that it was kinetically retarded and that

release could be accelerated by pulverizing the soil. Entrapment of EDB in soil

micropores was inferred. The study by Steinberg et al. (1987) did not use sterile soils

or flash-evaporated “spiked” soils. Moreover, the study was thorough, internally

consistent, and provided convincing data supporting logical arguments for a causal

relationship between diminished bioavailability and diminished biodegradation.
73

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses

1. Explain

anomalouspersistence ofvolatile,biodegradablepesticide infield soils

1,2-EthyleneDibromide(EDB), a gasphase-appliedfumigant withaqueoussolubility of3,370 ppm

Four tobaccofarm soilstreated withEDB in the field0.9 to 19 yearsprior to analysis

Use chemical andmicrobiological assaysto assess if long-aged(native) EDB wasdistinctive fromfreshly added EDBFreshly added EDBwas volatile; agedEDB was notAged EDB did notreadily equilibratewith aqueoussolution. Rates ofrelease of aged EDBfrom soil intoaqueous solutionwere temperaturedependent.Diffusion-basedmodeling of EDBrelease to aqueoussolution suggested2-3 decades to reach50 % of predictedequilibrium values.

Residual EDB washighly resistant toboth mobilizationand microbialdegradationThere is slowexchange between“native,” aged EDBand the freshlyadded chemicalPulverization of thesoil drasticallyaccelerated chemicalrelease of EDBfrom soilEntrapment of EDBin micropores wasinferred

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil

Reference

Steinberg et al.,

1987
74

The biodegradationtests showed thatfreshly added 14CEDB was readily

converted to acombination of

14CO2 (46 %) and

cell material (54 %),while the long-agedEDB was not

2. Construct,

operate, andmanage a fieldscale, preparedbed, landtreatment unitfor 2 soils;assessperformancebased oncontaminantmobility,toxicity,concentration;determinetreatment to

Crude oil (aged infield 30 to 40years)

Clay soil fromTexas gas storagefacility

Use soils aged in thefield; treat withtillage, nutrients,water (clay soil, 55weeks or silty sand,34 weeks)In clay soil, longcontaminated withcrude oil, totalpetroleum hydrocarbons (6,400 to11,000 ppm) showed no decline andleachability declinedfrom 4.4 to 1.3 ppm

After field treatment,laboratoryincubations (inwhich oxygen andnutrients were addedand viablepetroleum-degradingmicrobial populations were found)failed to stimulatebiodegradationThis, and absenceof inhibitory saltsor metals, suggeststhat bioavailability

Olivera etal.,1998

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
75

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

reachenvironmentally acceptable endpoint

total petroleumhydrocarbons(TPHs). Fieldleachate rangedfrom 0.1 to 75ppm, and only atrace of toxicity(microtox assay)was found

Diesel fuel (agedin field 3 years)

Silty sand fromTexas powerplant

In the silty sandcontaminated withdiesel fuel, TPHdeclined from 1,100to 160 ppm,aqueous release(TCLP) ranged from1.5 to 3.2 ppm, fieldleachate ranged from1.1 to 7.8 ppm, andonly a trace oftoxicity was found.

“Active, preparedbed treatment ofabout 15 weeksreached anenvironmentallyacceptableendpoint”

3. Use laboratory protocols

to quantify

Crude oil (agedin field 30 to 40years)

Clay soil fromTexas gas storagefacility

Build on results ofOlivera et al. (1998)by using laboratory

No degradation ofhydrocarbonsoccurred, nor could

Berg et al.,

1998

limited microbialactivity

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
76

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

rate of releaseof petroleumhydrocarbons;investigateimpacts ofbioremediationon hydrocarbon releasefrom soils;explorepossibilitythat bioremediation islimited byhydrocarbonrelease

measurements ofthe rate of release(ROR) ofpetroleumhydrocarbons frombioremediated fieldoils and theirresidual fractionThe RORprocedure usedXAD2 resins as aconstant sink forcontaminants inlong-term aqueousphase desorptionstudies. The releasedcontaminants weresubsequentlyanalyzed by GC/MS

releases ofhydrocarbon bequantified; “Lack ofbiodegradation wasrelated to inabilityof chemicals to bereleased from thesoils”“This indicated lowbioavailability ofhydrocarbon”

Diesel fuel (agedin fields 3 years)

Silty, sand fromTexas powerplant

Bioremediation byland treatmentdiminished themass of hydrocarData analysis yielded some anomalous (impossible)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
77

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

4. Determine the

limits andextent ofhydrocarbonbiodegradationtoxicity andleachability inpetroleumcontaminatedsoil

Three differentweights of crudeoil, artificiallyweathered at 5 %(wt/vol) freshlyadded

Two soils, onelow, one high inorganic matter,initiallyuncontaminated

Implement an8–11 monthbioremediationtreatment (of moist,nutrient-amendedsoil) and thenperform chemicalanalyses, toxicity,and leachabilityassays. Dependingon soil, weight of oiland hydrocarbonchain length,between 10 % and88 % of the initialoil was biodegraded.BTEX was almostalways belowdetection

bons released fromsoil; an initial rapidrelease wasfollowed by aslow-release phaseBiotreatmentreduced mobilityand eliminatedtoxicity of residualhydrocarbonsResiduals were nottoxic or leachable,or biodegradedfurther

modelingparameters fromthe contaminantrelease curvesBoth contamination andbioremediationevents wereconducted in thelaboratory

Salanitro et al.,1997

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
78

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Biotreatment causeddrastic reduction inleachable BTEX,eliminated toxicityto earthworms, anddiminishedinhibition of seedgerminationPlant growth assaysof toxicity revealedstimulation by oil

5. Test the

hypothesisthat slowdesorption isthe cause oflimitedbiodegradation; and ifthis is thecase,investigateprediction ofbioremediation

Fifteen PAHs Two PAHcontaminated

sediments fromAmsterdamHarbor

Desorption kineticsof 15 PAHs weredetermined beforeand after biotreatment via bioreactors(4 months) or landfarming (2 years)Desorption kineticswere measured bytrapping on TenaxresinsBiotreatment

For readilymetabolized PAHs,the extent ofpossible PAHdegradation couldbe roughlypredicted from theinitially desorbingfractionsHowever, a pool ofPAHs wasdesorbed but not

Cornelissen et al.,

1998

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
79

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

6. Test

hypothesesfor explainingthe longpersistence ofnaphthalenein sediments

Naphthalene(sparinglysoluble,31 ppm)

Sandy, organicmatter-richsurfacesediments

Use chemical andmicrobiologicalassays to attemptto identify factors infield that preventbiodegradation; GCanalysis of fieldsamples, 14CO2

production from14C- naphthalene

added to laboratoryincubated samples;

In laboratoryincubations,naphthalenemetabolism wasoxygen-limited, yetH O2 2 addition in

the field did notstimulate naphthalene loss in situ; noevidence fornutrient limitationwas obtained; soil

γ-irradiated soilswere used in theaging study; thelaboratory agingperiod wasrelatively short (4weeks); chemicalanalysis of fieldsamplesintroduced muchvariability in anattempt to

Madsen et al.,1996

reduced the rapidlydesorbing PAHfractionA pool of nondegrading PAHsalso desorbed; butthis pool remainedunchanged bybiotreatment

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

metabolized,probably due tomicrobial factors,not lack ofbioavailability

based on

desorptionkinetics
80

7. Compare

bioavailabilityand desorption kineticsof weathered(aged) andrecently addedherbicide

Simazine(chlorinatedtriazine herbicidethat is not amicrobial carbonand energysource; sparinglysoluble in water,3.5 ppm)

Agricultural soilreceivingsimazineannually formore than 20years

Sorption coefficientfor aged simazinewas 15 times that offreshly added;simazine concentration in soil solutionin field soils showedthat the agedmaterial was farbelow aqueousequilibriumconcentration;freshly addedsimazine was toxicto sugar beet

Pesticide agingdiminishedbioavailability, asmeasured bymicrobial degradation and plantuptakeDesorption kineticsof aged simazinewere slow compared to therecently addedchemical

Investigators didnot clearlydescribe themeans by whichbiodegradation offreshly added 14Csimazine was

discerned from thetotal simazinepoolSorptioncoefficients for

14C simazine were

measured with

Scribner et al.,1992

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

amendment ofsample, withnutrients, aging of γirradiated soilsfollowed bymineralizationassays, addition ofH O2 2 to fieldsediments followedby GC analysis

slurries enriched insorbed naphthalenemetabolized theaged substrate; noclear evidence todistinguish oxygenlimitation fromsequestration as thecause of persistence

documentbiodegradation insitu; differentenrichmentcultures reacteddifferently to agedsubstrates
81

8. Determine the

effect of agingtime in soilson thebiodegradability andextractabilityof organiccompounds

Phenanthrene(sparingly soluble1.3 ppm) and 4-nitrophenol(readily soluble inwater)

3 soils of varyingproperties(agricultural soil,muck soil,aquifer sand)

Using γ-irradiated,

sterile, spiked soils,the contact timebetween geosorbentand the two 14Clabeled test contaminants was varied

between 0 and 315days; then purecultures of bacteriawere inoculated intothe geosorbent and

14CO2 evolution

from the contaminants wasmeasured;phenanthrene was

“The data showthat phenanthreneand 4-chlorophenoladded to soilbecome increasinglymore resistant withtime to degradationand extraction.”Rates and extents ofmineralization wereinversely proportional to the agingduration

γ-irradiation ofsoil; phenanthrene contaminant was addedin an organicsolvent; purecultures wereused inbiodegradationtests

Hatzinger andAlexander, 1995

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

DMSO as carrierfor the simazine(though simazinesoil solution datawere genuine,carrier-free, fieldmeasurements)

 seedlings, agedsimazine was not
82

9. Use a

nonsterile soilsystem toassessbiodegradability (bioavailability) andextractabilityof differen

Phenanthrene Soil from Sierra

Nevada foothills

Age (0 to 600hours) phenanthrene in a nonsterilesoil with littlebiodegradationactivity, then addactive inoculum, andanalyze 14CO2

evolution kinetics,

Both the hexaneextracts andmineralizationcurves showed arapid (200 h) decline in theavailability ofphenanthrene; mildextraction

Short, agingperiods; phenanthrene was addedin methylenechloride organicsolvent

Schwartz andScow, 1999

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

spiked into soilusing dichloromethane as a carrier,4-chlorophenol wasadded in water;simultaneous to

14CO2 evolution

assay, the labeledcompounds wereextracted from soilusing butanol andsoxhlet procedures;sonication of soilaggregates slightlyenhanced mineralization
83

10. Examine

competitivedisplacementof agedphenanthrene bypyrene inchemicalextraction andbiodegradation assays

Phenanthrene Agricultural soil

and peat soil

Mineralization andchemical extractability of agedradiolabeledphenanthrene (3 to123 days) in twosterile soils, withand without addedpyrene weremeasuredSoils were sterilizedby γ-irradiation;

14C-phenanthrene

was coated onto thewall of a flask andthen allowed toequilibrate throughthe aqueous phasewith soil.

Aging diminishedrate and extent ofmineralization;pyrene additionenhanced phenanthrene mineralization; phenanthreneextractability andKd values also were

altered by pyrene;the pyreneinfluences may havebeen caused bycompetitivedisplacement of thephenanthrene fromglassy organicmatter

γ-irradiation ofsoil; a pureculture was usedin biodegradationtests

White et al., 1999

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

tially aged

phenanthrene

as well as phenanthrene extractability procedures reflect bioavailability;

sorption processesmay prolong thepresence of achemical in soil
84

11. Model

coupledprocesses ofdesorbtionandmineralizationusing twodistinctivemodelorganisms

Naphthalene (asparingly solublehydrophobiccompound; 31ppm)

Agriculturalsoils—dried,ground, andautoclaved

Naphthalenemineralization bytwo differentbacterial cultureswas examined inaqueous solutioncontaining varyingamounts of addedsoil

For one bacterium,sorption limited therate and extent ofslurry-phasenaphthalenemineralization. Thiswas not the case forthe other organism,which metabolizeda significant portionof sorbednaphthalene

Soils were airdried, ground, andautoclaved. 14Cnaphthalene was

added to soilslurries in acetoneas initial carrier;pure cultures wereexamined

Guerin andBoyd, 1992

12. Determine

howmicrobialactivityinfluencesassociationsbetween soilorganicmatter and

Pyrene (solubility0.13 ppm inwater)

A forest soil andan inoculum froma pyrenecontaminatedSuperfund site

The distribution of

14C- pyrene and

byproducts weredetermined bysequential soilextractionprocedures fromforest soil, with orwithout a pyreneOver time, the extractability of pyrene and pyrene products decreased to a greater extent in the metabolically active soil Biological activity Methanol was used as the pyrene solvent It is difficult to interpret the meaning of extractability changes in such a Guthrie and Pfaender, 1998

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

Pyrene was added inmethanol as a carrier(<0.1% by volume)
85

degrading inoculumor a metabolicinhibitor.Bioavailability wasassessed by 14CO2evolution fromradiolabeled pyreneadded to soil inconstantly flushedincubationchambers

may reducebiotoxicity. Theamount of 14CO2

released fromfreshly added andaged (270 days)pyrene was similar

13. Describe the

potentialrelease andsubsequentdegradationof bound 14C

residues fromsoilpreviouslytreated withatrazine

Atrazine, watersolublechlorinatedtriazine herbicidethat was not aready carbon andenergy source forsoil microorganisms

Agricultural soil Form 14C-atrazine

bound residues insoil via a 1-yearlaboratoryincubation;exhaustively extractwith methanol toremove residual 14C;

chemicallycharacterize theresidual 14C in soil;

inoculate the soilwith two differentatrazine-degrading

During an 84-dayincubation, atrazineand four metabolites were releasedfrom the initiallysoil-bound 14C

atrazine; differentbacteria releasedhydroxyatrazine todifferent extents;soil bound residuescan be released

Soil sampleprepared byexhaustiveextraction withmethanol; use ofpure cultures

Khan and Behki,1990

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference

complexexperimentalsystem

contaminants
86

bacteria with andwithout a glucoseamendment; assessmetabolism of thebound 14C

14. Examine the

fate of soilagednonextractable atrazinewith respectto biodegradation,microbiallyfacilitatedrelease, andabioticdesorption

Atrazine (watersoluble)

Agricultural siltloam

After 3 months ofaging atrazine insterile soil, theextractable atrazinewas removed. Next,the soil containingresidual atrazinewas subjected tomineralization anddesorption testsAlmost no (<2.5%)atrazine mineralization occurredViable microorganisms did notenhance release ofpreviously boundatrazine

Bound residues aresubject todesorption andrelease from soilsRates of atrazinemineralization wereexceeded, in nearlyall cases, by thedesorption rates—suggesting no masstransfer limitations

γ-irradiated soilswere used to age(bind) the atrazinein soilTest substrate(atrazine) was notappreciablymetabolized, thusthe bioavailabilityassay was notrobust

Johnson et al.,

1999

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
87

15. Compare

biodegradation ofsorbed- andsolutionphasecompound insoil

2,4-D (a reasonably watersolublecompound; 900ppm)

Two agriculturalsoils and oneclay, all sterilizedby autoclaving

Three separatemathematicalmodels were usedto evaluate themineralization of2,4-D added to soiland metabolized byan inoculated pureculture

The model thatbest fit the datapresumed thatsorbed 2,4-D wasunavailable

Soils wereautoclaved andground; a pureculture was used;2,4-D was freshlyadded to soil withmethanol as initialcarrier. A 5-hincubation timewas used

Ogram et al.,1985

16. Assess the

feasibility ofusingchemicalextractionproceduresto estimatebioavailabilityof pesticidesto earthworms

DDT, DDE,DDD (sparinglysoluble chlorinated insecticideor byproduct)

Seven agriculturalsoils of varyingproperties andpesticideapplicationhistories

Determine pesticideconcentrations viasoxhlet extraction/analysis; measureearthworm uptakeof pesticides; correlate earthwormuptake with C18membranes and25% tetrahydrofuran extracts; agepesticide (0 to 924d) in γ-irradiated

soils; pesticides were

Worms assimilated3% to 66% of thepesticides; thisuptake correlatedwell with chemicalextraction assays.Correlationcoefficients were0.921 or higher forthe C18 membraneand 0.83-09.48 forthe tetrahydrofuranextraction

γ-irradiation ofsoil; addition ofpesticides inorganic solvent;incompletelyvented

Tang et al., 1999

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
88

added indichloromethanesolvent allowed tovolatilize for 48 h

17. Assess the

decline inbioavailabilityas a result ofaging ofpesticides

DDT, DDE,DDD, dieldrin

Three agriculturalsoils of varyingproperties andpesticideapplicationhistories

Use an earthwormuptake bioassay tomeasure pesticidebioavailability infield soils andlaboratory-agedsoils; pesticideuptake assaysinvolved exposingworms to soil, thenharvesting,extraction, and GCanalyses; field agedsoils (49 and 30-year periods) wereused; laboratoryaged soils (0, 90,190 days) were γirradiated andpesticides were

Aging materialsreducesbioavailabilitySoil extraction byTenax resincorrelated well withearthworm uptakefor three of fourpesticides

γ-irradiation ofsoil; addition ofpesticides inorganic solvents,incompletelyventedIn manyinstances, thedata showed thataging had noeffect onbioavailability

Morrison et al.,2000

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
89

added in organicsolvents; chemicalanalysis of thepesticides adsorbedto Tenax resin wasexamined as apossible surrogatefor earthwormuptake

TABLE 4 Survey of Laboratory and Field Studies Aimed at Discerning Mechanistic Relationships Between the Biodegradation

of Organic Compounds and Their Bioavailability in Soil (Continued)

* Departure from the 7 (seven) quality criteria in Table 2

Goals Compound Geosorbent

Approach/

Methods/Data Conclusions

Methodological*

Weaknesses Reference
90

The next two entries in Table 4 were conducted by R. Loehr’s research group at

the University of Texas. Like Steinberg et al. (1987), a field approach was taken, and

subsequent hypotheses were sequentially tested to try to explain the persistence

of organic soil contaminants subjected to bioremediation processes; no methodological weaknesses could be discerned. For Olivera et al. (1998) and Berg et al.

(1998), the contaminants were petroleum hydrocarbons (crude oil and diesel fuel)

that had been in the field for either 3-4 decades or 3 years before being field

bioremediated (land treatment) for either 55 or 34 weeks. Chemical analysis of the

field-bioremediated soils was used to monitor process effectiveness. Simultaneously,

the mobility (leachability) of the residual hydrocarbons and their toxicity were

measured. The behavior of the different hydrocarbons (aged in two different field

soils for different periods) was not uniform: land treatment reduced residual hydrocarbons in the soil contaminated for 3 years, but not for 3 to 4 decades. Nonetheless, low hydrocarbon leachability and toxicity were documented for both soils.

This suggested to the authors that “environmentally acceptable endpoints had

been reached.” Furthermore, alternative hypotheses for explaining the field persistence of the hydrocarbons (e.g., nutrient limitations, lack of viable microbial populations) were tested and dismissed. Thus, diminished bioavailability was implicated

as the cause for persistence of hydrocarbons in biotreated field soils.

Another study addressing bioavailability and biodegradation of petroleum hydrocarbons in soils was that of Salanitro et al. (1997; entry number four of Table 4). The

approach was to freshly contaminate soil with crude oil so that the naturally occurring

microbial populations could biodegrade it. Then chemical analyses, as well as leachability and toxicity (earthworm and seed germination) assays, were performed to characterize the residual hydrocarbons. This study’s only (minor) weakness was that it was done

in a laboratory—it did not include true field observations of persistence or biodegradation activity. Nonetheless, Salanitro et al’s. (1997) study met most of the criteria in Table

2 (real-world microbial populations, absence of artifacts possibly associated with sterilization and flash-evaporated spikes, and consistency). The conclusion seemed sound:

concomitant with the persistence of hydrocarbon residuals in soils, the toxicity and

mobility of petroleum hydrocarbons were substantially reduced by biotreatment.

Cornelissen (1998), the fifth entry in Table 4, gathered field-contaminated,

polycyclic aromatic hydrocarbon (PAH)-rich sediments from Amsterdam Harbor,

subjected them to bioremediation (either bioreactor or land farming), and measured

desorption kinetics of the PAHs before and after biotreatment. This study had no

obvious methodological weaknesses. It found that some PAH compounds persisted regardless of their desorption from sediments. However, for the biodegradable PAH pool, biotreatment was found to effectively degrade the otherwise rapidly

leachable (releasable) PAHs. Cornelissen et al. (1998) concluded that the biodegradable PAH pool was predictable, based on the chemically measurable “rapidly

desorbing fraction.” This implied that the rapidly desorbing fraction was bioavailable

and that the nondesorbing fraction was not.
91

Entry number six in Table 4 (Madsen et al., 1996) began with the goal of testing

several hypotheses for the persistence of a readily biodegradable PAH compound,

naphthalene, in a coal tar-contaminated study site. Cores removed from the site

proved that naphthalene was present in anaerobic freshwater sediments. Furthermore, laboratory incubations proved that the native microbial community could readily

convert freshly added 14C-labeled naphthalene to 14CO2 under aerobic conditions and

that these microorganisms were not nutrient limited. The cause of contaminant persistence was narrowed to two hypotheses: lack of bioavailable naphthalene or lack of

a final electron acceptor (O2

). To address the former hypothesis, irradiated sterile

sediments were aged in the presence of 14C- naphthalene and mineralization was

assayed using inocula from the sediment. However, the maximum aging period was

unrealistically short (4 weeks). In addition, two types of inocula (enriched on either

solid- or liquid-phase naphthalene) responded differently to the aged naphthalene.

Malleability of the microbial community was indicated because sediment microorganisms pre-enriched on soluble naphthalene did not easily mineralize the aged naphthalene, while unenriched sediment microorganisms mineralized aged and unaged naphthalene equally. To test the oxygen limitation hypothesis in the field, H2

O2 was added

to small volumes of enclosed sediments. However, over 3 weeks, the H2O2 treatment

did not stimulate naphthalene loss from soil. Thus, Madsen et al.’s study (1996) was

equivocal. Lack of field stimulation by the H2O2 amendment suggested a bioavailability

limitation, but it was possible that the H2O2 treatment was ineffective. Similarly, the

experiments using aged, irradiated sediments showed that diminished bioavailability

associated with aging was dependent on the inoculum.

To understand field persistence of the herbicide, simazine, Scribner et al. (1992;

seventh entry in Table 4) used chemical and bioassay methods to compare the

partitioning behavior, biodegradability, and toxicity of freshly added versus aged

simazine in soil. When simazine was freshly added in the field, it was found in the

soil solution, but its fractional equilibrium concentration was otherwise very low in

field pore waters. Moreover, freshly added, but not aged, simazine was both biodegradable and toxic to plant seedlings. This study presented laboratory data [much

like that of Steinberg et al. (1987)] contrasting a constant pool of aged simazine that

was not biodegraded in soil by the naturally occurring microflora with a declining

pool of 14C-freshly added simazine; however, the methods leading to such data

were not clearly described. Overall, this study’s results supported the view that an

aged, field-applied pesticide was neither mobile nor bioavailable. The case was not

flawless, however, especially because simazine is not a readily utilized carbon and

energy source for soil microorganisms.

Hatzinger and Alexander (1995; entry #8 of Table 4) published a pivotal series of

experiments that yielded clear, inverse relationships between the extractability and

biodegradability of phenanthrene and 4-nitrophenol that had been aged in sterile

(irradiated) soil up to 315 days prior to chemical and biodegradation analyses.

Unfortunately, using the criteria established in Table 2, the work of Hatzinger and
92

Alexander (1995) has diminished merit because no field observations were made,

sterile soil was used, the phenanthrene was added with a carrier organic solvent,

and pure cultures were used in the mineralization bioassay.

The ninth entry in Table 4 (Schwartz and Scow, 1990) examined the behavior of

a three-ring PAH compound, phenanthrene, which was added to a soil whose

native microflora had unusually low phenanthrene biodegradation activity. After as

long as 600 hours (25 days), chemical extractions and a phenanthrene-degrading

microbial inoculum were used to assess the influence of aging on phenanthrene

availability. Data showed that, although both the amount of hexane-extractable and

biodegradable phenanthrene diminished with aging, the bacteria were able to access a pool of phenanthrene unavailable to the hexane extractant. Thus, some

support for the diminished bioavailability of aged phenanthrene was garnered. But

the results were somewhat ambiguous and the methods did not fully conform to the

criteria of Table 2 (see column 6 of Entry #9 in Table 4).

A collaborative effort by laboratories operated by M. Alexander and J. Pignatello

(White et al., 1997; Entry #10, Table 4) sought to test hypotheses about binding sites

for phenanthrene and how adding another PAH compound (pyrene) could influence

phenanthrene’s chemical and biological behavior. Aging was found to diminish both

the rate and extent of phenanthrene mineralization. Moreover, the added pyrene was

found to boost both extractability of the phenanthrene and its extent of mineralization

by an inoculated bacterium unable to metabolize pyrene. Unfortunately, applying the

criteria in Table 2, the work by White et al. (1999) was not field based and both

irradiated soil and a pure culture were used in the experiments.

In 1992, Guerin and Boyd (entry #11 of Table 4) published an often-cited study

illustrative of how pure cultures of bacteria can reveal contrasting physiological

capabilities. This study failed to conform to any of the seven criteria of quality

shown in Table 2 [no field observations, unrealistic laboratory incubation conditions, altered geosorbents, soil sterilized by autoclaving, the test compound (naphthalene) had a co-solvent (acetone), and two pure bacterial cultures were used].

Nevertheless, diversity and malleability of the microbial world became apparent

when one culture was found to metabolize only aqueous-phase naphthalene; while

the other metabolized a significant portion of the sorbed chemical.

In another laboratory study to determine the fate of PAHs in soil, Guthrie and

Pfaender (1998; entry #12 of Table 4) used the four-ring compound, pyrene, to assess

how microbial activity influenced the conversion of 14C-labeled pyrene that was

added to several operationally established humic acid fractions of soil. Viable soil

microorganisms were found to enhance incorporation of the 14C-labeled pyrene into

humic acid fractions. However, aging the pyrene for 270 days did not significantly

diminish its susceptibility to mineralization by soil microorganisms. The work by

Guthrie and Pfaender (1998) was not conducted in the field and used several approaches that failed to conform with the criteria in Table 2.
93

Khan and Behki (1990) and Johnson et al. (1999) rigorously tackled the issue of

incorporating soil-applied pesticides into humic substances and their potential subsequent release (entries #13 and #14, respectively, in Table 4). Khan and Behki (1990)

formed bound 14C-atrazine residues in an agricultural soil by incubating the herbicide

for 1 year and then used ethanol extraction to ensure all unbound forms of atrazine were

removed. Then Khan and Behki (1990) added microbial cultures to the soil and used

chemical analyses to follow microbial metabolism and extractability of the soil-bound

residue during an 84-day incubation. Atrazine and atrazine metabolites (indicative of

microbial metabolism) were released in ppm amounts. Johnson et al. (1999) essentially

repeated the work of Khan and Behki (1990) with methodological modifications that

included aseptic preparation of bound residues (in γ-irradiated soil) during 3 months of

aging. Data from Johnson et al. (1999) showed that bound residues were extractable

and desorbable from soil but viable microorganisms did not enhance extractability.

Thus, the studies by Khan and Behki (1990) and Johnson et al. (1999) displayed both

consistency and inconsistency: bound residues were released in both studies, but the

role of microorganisms in these releases was not reproduced. Neither of the studies

conformed to the quality criteria in Table 2.

Entry #15 in Table 4 (Ogram et al., 1985) is an often-cited report that used a

combination of mathematical modeling and careful manipulation of soils and pure

bacterial cultures to test the hypothesis that sorption of 2,4-D to soil protects the

herbicide from biodegradation. Ogram et al. used sterile soil (by autoclaving), a pure

bacterial culture, a co-solvent for the 2,4-D, and a brief (5 hours) incubation time for

the experiments. Clearly, these experimental conditions failed to conform with the

“field relevant” criteria of Table 2. It is noteworthy, however, that on numerous occasions Ogram et al.’s (1985) paper has been cited to support arguments that sorbed

substrates may be unavailable for biodegradation. However, in the original publication, the authors were very careful to advise readers not to misuse the findings.

Ogram et al. (1985) stated that “results...for a single strain of bacteria degrading 2,4-D

in a rigorously controlled environment...may not be applicable to conditions one

would expect to find in situ, where mixed populations may be degrading the pesticide

for a much longer time than the 5-hour incubation period used here.”

The final two entries in Table 4 (#16, 17) examined the influence of the residence

time of agricultural chemicals (DDT and related compounds) on bioavailability, assessed via earthworm uptake instead of biodegradation. Tang et al.’s (1995, entry #16

in Table 4) objective was to calibrate chemical extraction techniques with pesticide

bioavailability to earthworms in soil. These researchers were able to build an experimental design around soils to which DDT had been applied 30 and 49 years prior to

the study. Several other combinations of pesticides and agricultural soils also were

prepared in the laboratory using sterile (irradiated) soil aged in the presence of solvent-delivered, flash-evaporated pesticides. Although Tang et al’s. (1999) focus was

on correlating chemical extraction and earthworm uptake measurements, data presented sometimes failed to support the hypothesis that bioavailability diminishes
94

over time. For instance, for a Lima loam soil, aseptic aging of DDT, DDE, and DDD led

to increased pesticide bioavailability to earthworms. In the study by Morrison et al.

(2000; final entry of Table 4), chemical extraction and earthworm uptake procedures

were again applied to aged and unaged pesticides in soils. In several instances,

inconsistent and conflicting data indicated that long residence times of insecticides

in soils did not lead to decreases in earthworm uptake from soil. Thus, Tang et al.

(1999) and Morrison et al. (2000) provide examples of reports that are of secondary

significance for this report. Although some of the data were obtained from fieldcontaminated soil, irradiated soils, flash-evaporated pesticide delivery, and inconsistent and conflicting data (see Table 2) detracted from their results.

8.3.1. Summary

Of the hundreds of research articles published in the last two decades on

bioavailability and biodegradation, details of only 17 were presented in Section 8.3

and Table 4. These were chosen because they were influential and/or recent and

representative of the varying approaches, objectives, and the quality of experimental

designs and methodologies published to date. Each of the studies described in Table

4 had its own merits—otherwise publication would not have occurred. Furthermore,

although none of the studies was designed to specifically address the needs of this

report, roughly the first half of the entries in Table 4 conformed reasonably well with

this report’s quality criteria (field realism, environmental relevance, absence of artifacts, consistency) set forth in Table 2.

Steinberg et al. (1987) showed that field-aged and freshly added EDB pools in

soil were chemically and biologically distinctive, and that pulverization enhanced

desorption of aged EDB; thus, micropore entrapment was indicated as a cause of

reduced EDB bioavailability. Studies by Olivera et al. (1998) and Berg et al. (1998)

showed that petroleum hydrocarbon residues were persistent after land farming

biotreatment had reduced leachability and toxicity and that additional biodegradation was likely prevented by reduced hydrocarbon bioavailability; thus, “environmentally acceptable endpoints” had been achieved by bioremediation treatment.

Although laboratory-based, the study by Salanitro et al. (1997) also demonstrated

that biotreatment of freshly added petroleum hydrocarbons in soil reduced the

mobility and toxicity, hence the bioavailability, of residual contaminants. Cornelissen

et al’s. (1998) investigation showed that a portion of the rapidly desorbing (i.e.,

bioavailable) PAH fraction from harbor sediments corresponded to the readily biodegradable fraction. The combined field and laboratory study by Madsen et al.

(1996) narrowed the cause of naphthalene persistence to be limitations in either

bioavailability or oxygen. Scribner et al’s. (1992) investigation [much like that of

Steinberg et al. (1987)] showed that the chemical and biological properties of freshly

added and aged simazine were distinctive, and an aging-related reduction in simazine bioavailability was the cause. Thus, the above investigations strongly sup-
95

port the hypothesis that chemical assays of materials in soil can overestimate

what is actually biologically absorbed and metabolized. Several bioassays (biodegradation, uptake, and toxicity) and chemical extraction assays indicated that the

reduction in contaminant bioavailability in soil was a time-dependent phenomenon

(freshly added contaminants were distinctive from aged contaminants). Such observations conform with the various operational definitions for bioavailability limitations described in Sections 3.1 and 3.3 of this report.

The entries that appeared in the latter half of Table 4 largely failed to conform with

the quality criteria of Table 2. Nevertheless, each study contributed to broad scholarly information about the potential influences of geosorbents, time, reactions, and

biota on biodegradation and pollutant behavior and/or toxicity in soil.

8.4 Influence of Bioavailability on Phytoremediation of

Metal-Contaminated Soils

One key characteristic of plant-based, cleanup technologies is that plants can

have major physical and hydrologic impacts on contaminated sites (Ensley, 2000;

Schnoor, 2000). Not only do growing plant roots explore and penetrate the threedimensional volume of contaminated soil, but the plant vascular system withdraws

water from the soil depths and delivers it to the atmosphere. Transpirational demand at leaf surfaces is the driving force for the flow of water from soil solution to

the root surface and into and through the plant biomass. As discussed in Section

7.2 of this report, the relative contribution of mass flow (versus diffusion versus

root interception) to nutrient uptake varies with each plant species. Nonetheless,

because all plants transpire, they all establish moisture gradients within the soil

matrix, induce advective mass transfer of water, and convey solutes both through

soil and into their tissues (Schnoor, 2000). If solutes, including metals, are in soil

solution, they assuredly will be conveyed to the plant (see Section 7.1). It is for this

reason that substantial research efforts have been directed at boosting access—

boosting bioavailability—of metals to plants.

Mobilization of metals from soil into soil solution and plants is particularly

amenable to experimentation because organic chelating agents such as ethylenediaminetetraacetic acid (EDTA) have been used for decades for similar purposes.

Phytochelators (Khan et al., 2000) also have been described. Many greenhouse

experiments have demonstrated the effectiveness of added chelating agents in

enhancing potential phytoremediation (Blaylock et al., 1997). For example, when

EDTA and HEDTA were added to pot-grown Helianthus annuus, uptake and translocation of Cd and Ni from contaminated soil was substantially increased (Chen

and Cutright, 2001). Also, Ebbs and Kochian (1998) demonstrated that an EDTA

amendment increased Zn accumulation by Indian mustard planted in metal-contaminated soil.
96

Although entry of metals into soil solution is an important aspect of

phytoremediation strategies, defining soil and plant factors that govern metal uptake

also is a major area of current research (McGrath et al., 2001). In this regard,

“nonbioavailable” forms of metal have begun to be identified. For example, Stanhope

et al. (2000) used isotope dilution techniques in pot studies of Indian mustard to

show that, despite general mobilization of Cd, Zn, Pb, Ni, and Cu from sludge-contaminated soil, EDTA failed to facilitate plant access to a nonlabile pool of soil Cd. In

related studies, researchers have begun to examine the time-dependence of extractability (Martinez and McBride, 20001), animal toxicity (Lock and Janssen, 2001), and

plant uptake (Pedersen et al., 20001; Stacey et al., 2001) of metal contaminants. Results of such studies appear to be delivering a message much like that from microbial

biodegradation studies of organic soil contaminants (see Sections 8.1 through 8.3 of

this report): the relationships between metals, soils, and plants seem to vary substantially depending on the type, composition, and design of the experimental system.

8.5 A Synthesis: Evaluating the Relationships Between

Bioavailability and Bioremediation Based on Sections 2

to 8.4 of This Report

This section presents a series of statements that blend lessons from the mechanistic studies, discussed in Sections 8.1 to 8.4 of this report, with the fundamentals of biodegradation, bioremediation, bioavailability, persistence of organic compounds, paradigms of geosorbent-contaminant interaction, and mechanisms of

contaminant uptake by biota that were presented in Sections 2 to 7 of this report.

• Microorganisms and plants offer a broad array of physiological and/or physical mechanisms for eliminating and/or binding organic and inorganic contaminants in soil and sediment matrices (Section 2).

• Soil and sediment exist in a state of kinetically constrained thermodynamic

disequilibrium in which combinations of biochemical and chemical reactions,

as well as gas-, liquid-, and solid-phase transport processes, cause the biogeochemical cycling of both organic and inorganic compounds (Section 6).

• Naturally occurring, plant-derived organic compounds constantly cycle

through the soil habitat—partially entering a pool of persistent humic substances thought to be protected from microbial attack by a combination of

complex molecular structure, resistance to enzymatic digestion, insolubility,

failure to enter the microbial cell, possible enzyme inactivation, and complexation (masking) (Section 5).

• Although the details of how and why chemicals persist in soil are governed by

each chemical’s molecular structure, a survey of soil bioremediation field

projects (for petroleum hydrocarbons, pentachlorophenol, PAHs, metals; Sec-
97

tion 4) indicates that these organic and inorganic contaminants also routinely

persist in soil, though simple mixing and aeration frequently accelerate organic contaminant loss.

• Six specific mechanisms have been hypothesized as the “cause” of diminished

bioavailability, hence, chemical waste persistence in soils: (1) sorption to

surfaces, (2) partitioning into NAPL phases, (3) micropore- and nanoporeseparation, (4) complexation, (5) insolubility, and (6) partitioning into organic

matter (Section 3).

• Bioavailability can seldom be measured directly. Instead, it emerges from the

specific, detailed, three-way interaction between biota, chemical waste compounds, and a geochemical matrix under study (Section 3).

• Soil biota (microorganisms and plants) feature a diverse array of anatomical

and physiological adaptations (e.g., solid-phase uptake, diffusion-accelerated

uptake, active transport, extracellular enzymatic digestion, extracellular scavenging systems) that allow them to successfully contend with the adverse

nutritional conditions of soil (Section 7). Such physiological and evolutionary

adaptations should not be underestimated when interpreting data or making

predictions related to the availability of chemical wastes for bioremediation in

soil. Many metabolic capabilities of soil biota (especially microorganisms)

probably have yet to be discovered (Section 6.3).

• All six hypothesized “mechanisms of diminished bioavailability” (Section 3)

involve a spatial separation of the chemical substrate from intra- or extracellular metabolism. To some degree, direct physical contact between cells (or

roots) and the metabolized chemical can counteract spatial separation. However, even excretion of extracellular enzymes cannot counteract the size differential between molecules that could diffuse into soil nano- or micropores

(Ball and Roberts, 1991; Newman and Thomasson, 1979) and microbial cells

(~1μm). Thus, regardless of the stage of sophistication of mathematical and

conceptual models of geosorbent matrices (Section 6.4), there is no doubt that

some portions of chemical waste in soil can be—and are—inaccessible to

plants and microorganisms.

• Inaccessibility of chemical wastes in soil, synonymous with diminished

bioavailability, is observed as discrepancies between modest or undetectable

bioassay responses (uptake, toxicity, or biodegradation) to chemicals, compared to ready analytical quantification of the chemicals in the same soil

samples (Section 8.3).

• The prevalence of ambiguous, even conflicting, findings in the existing literature on biodegradation and bioavailability (Sections 8.1 to 8.4) is probably a

reflection of: (1) the simultaneous variability in the methods, test chemicals,

approaches, and experimental systems that have been devised by numerous

~
98

concerned investigators, and (2) the concomitant shifts in prevailing mechanisms (e.g., direct contact, extracellular enzymes, diffusion, active uptake) by

which the biotic components in these investigations access and metabolize the

tested chemicals.
99

Section 9

Overcoming Constraints on Site Cleanup

The propensity for contaminated soil to remain contaminated poses a major

challenge for the proper management of contaminated sites, especially site cleanup

technologies (National Research Council, 1997). Several recent publications on site

cleanup techniques, particularly bioremediation measures, have reviewed current

knowledge on efforts to overcome constraints on contaminant removal (e.g.,

Alexander, 1999; Maier, 2000; Volkering et al., 1997; Scow and Johnson, 1997; West

and Harwell, 1992). According to Maier (2000), surfactants (biosurfactants and

synthetic surfactants), cosolvents, and thermal treatment can aid in the physicochemical removal of organic contaminants from porous media.

All such contaminant-mobilizing strategies have advantages and disadvantages, and none have yet been found to be fully satisfactory. The reasons for this

absence of completely satisfactory performance should be no surprise to the reader

of this report. Soils and sediments are highly complex geochemical matrices and,

given the variable properties of both contaminants and geochemical matrices, the

impact of geosorbent treatment techniques is not likely to be consistent. The task

of mobilizing chemical waste contaminants from geosorbents might be pursued for

“pump and treat” purposes (simply to wash the matrix) or for bioremediation purposes (increasing contaminant availability for biodegradation reactions). Regardless of the goal, understanding the science of mobilizing contaminants out of soils

and sediments is analogous to understanding the science of soil bioremediation

(e.g., Figure 1; Section 3.3). Thermodynamically governed, complex interactions

between the geosorbent matrix and the contaminants cannot be escaped. If Figure

1 was to be adapted to address mobilization of contaminants from soils, the label on

the oval in Figure 1 should be replaced by “contaminant mobilization treatment.”

Then the performance of the system in the revised Figure 1 would be assessed, not

on bioremediation, but simply soil washing efficacy. The “contaminant mobilization

treatments” examined to date have often been plagued with implementation difficulties that range from high cost to inefficiency, secondary chemical effects, and
100

secondary biological effects (toxicity, degradation of water quality, inhibition of

biodegrading microorganisms).

Because potential soil amendments feature a diversity of chemical structures and

their corresponding properties, there are still many promising approaches for

geosorbent decontamination efforts to be devised and tested. Noordman et al. (1998)

have explained that aquifers contaminated with hydrophobic organic contaminants

can theoretically be remediated using dissolved organic matter, cyclodextrins (e.g.,

McCray and Brusseau, 1999), organic cosolvents, or surfactants (Guha et al., 1998).

Rhamnolipid biosurfactants may be superior to the alternatives because they are

naturally occurring compounds with low environmental impact (Noordman et al.,

1998; Torrens et al., 1998). Surfactant treatment of soils and sediments occurs via

emulsification, micellar solubilization, or facilitated transport mechanisms. Volkering

et al. (1997) reviewed many efforts to link surfactant-mediated mobilization of contaminants to bioremediation: results ranged from stimulation to inhibition of both

desorption and biodegradation of polluting compounds. Volkering et al. (1997) concluded that “no general trends can be found... Therefore, more research is necessary

to make the application of surfactants a standard tool in biological soil remediation.”

Regarding phytoremediation of metal-contaminated soil (see Section 8.4 of this

report), there is no doubt that when chelating agents boost the concentration of

metals in soil solution, metal uptake by plants will be enhanced. However, subtle

secondary effects of such amendments on soils, plants, and real-world field sites

have yet to be fully elucidated.
101

Section 10

Conclusions, Implications, and

Possible Areas of Future Research

This section focuses all prior portions of this report upon key facts, principles, and their implications.

Three Scientific Conclusions:

1. Based on a conservative, reasonably thorough and careful evaluation of scientific studies described in this report, there is no doubt that chemical wastes in

soil can be, and often are, in a state of reduced bioavailability.

2. Reduced bioavailability simply means that a chemical waste’s diminished “effective concentration” is proportionately balanced by a lingering reservoir of

the chemical waste in soil and sediments. This lingering reservoir remains in the

soil habitat regardless of which combinations of conceptual or actual sequestration mechanisms (e.g., complexation into bound residues, diffusion into soil

pores, NAPL partitioning) apply.

3. Soil is, by definition, a thermodynamically unstable, kinetically constrained medium whose chemical composition, including solid, liquid, and gaseous components, is constantly changing. Thus, the “nonbioavailable” chemical wastes in

this lingering reservoir (point #2) are always subject to release into soil solution

where the wastes are resubjected to a variety of transport and/or transformation

processes (e.g., immobilization, biodegradation, uptake by receptors).

Implications of Bioavailability:

1. Considerable effort has been expended in investigating the hypothesis that chemical wastes have diminished bioavailability in soils and sediments. Results of these

efforts have been ambiguous because of the immense diversity in types and properties of chemical wastes, geosorbents, biota, experimental approaches, and the

idiosyncrasies in mechanisms by which biota interact with chemical wastes.
102

2. From a practical, regulatory point of view, establishing the foundation of reduced

bioavailability is crucial. If the reduced bioavailability of chemical wastes in soil

becomes widely accepted, then proper quantitative measures of bioavailability

reduction could be developed to accurately estimate the risks posed to human

health and ecological processes by chemical wastes in soils and sediments.

3. An accurate estimation of risks to human and environmental health posed by

chemical wastes in soils (point #2) is, itself, a crucial step toward: (i) identifying

pragmatic, economically feasible environmental cleanup goals; (ii) establishing operational definitions of “treatment” by bioremediation technology; (iii)

realistically classifying polluted sites based on planned land-use scenarios;

(iv) developing public acceptance of risk-based contaminant cleanup efforts;

(v) developing public acceptance of cleanup goals that are above the “original,

pristine state” of the contaminated site, and (vi) legitimizing the concept of

“environmentally acceptable endpoints.”

Possible Areas of Future Research:

1. Microbe- and plant-based bioremediation technologies are not able to remove

100 percent of contaminants from polluted sites. Thus, managing residual contaminants in real-world sites will require progress in the science of contaminant

behavior, in risk assessment, and in site management polices.

2. Several major questions remain about the reservoirs of “nonbioavailable” chemical wastes in soil:

A. Is it possible to determine the rates of soil processes in the field that

convert “nonbioavailable” chemical wastes to available forms?

B. What methods are appropriate?

C. Can these appropriate methods be applied in regionally representative

field sites containing representative classes of organic and inorganic chemical wastes?

D. Can the results of item #2C be used to formulate region-specific, site

management guidelines?

E. When released into soil solution, what are the relative probabilities and

rates of the chemical wastes becoming:

– resequestered?

– biodegraded?

– transported to susceptible receptors?

F. If the released chemical wastes are readily mineralized (if organic) or immobilized (if inorganic), would “natural attenuation” of these originally
103

nonbioavailable contaminants offer effective protection of human health

and the environment?

3. Several major questions remain about the true, field state of solid, liquid, amorphous, and metastable organic and inorganic constituents in soils and sediments:

A. What are the chemical forms and locations of soil constituents?

B. What types of surface chemical and/or other reactions predominate?

C. What effects do these reactions have on sequestration and/or release of

chemical wastes?

D. At what rates (minutes, days, decades, centuries) do these reactions occur?

4. The types of basic scientific questions raised in item #2 also apply to contaminants and their reactions in soils and sediments.

5. The questions raised in item #2 apply to soil biota—especially to the soil

microbial community—because so much of its diversity and potential physiological activities are uncharacterized.

6. Would modeling of the processes in items #2-5 provide predictive, heuristic

tools to regulators, the public, and scientists?

7. Can the risks of release of initially “nonbioavailable” soil contaminants to soil

solution be estimated in a meaningful way that accounts for:

A. Critical pathways at each contaminated site that could lead to exposure to

humans or other key receptors?

B. An integration of known mechanisms that contribute to both the re-release and re-sequestration of soil contaminants?

8. Can knowledge of the mechanisms that lead to persistence of chemical wastes

in soils (i.e., resistance to biodegradation and/or physical/chemical immobilization) be understood well enough to manage the mechanisms at field sites in

ways that either enhance or counteract sequestration reactions, as appropriate

for site management goals (e.g., Verstraete and Devliegher, 1997)?

9. Can researchers be convinced to use methodologies and experimental designs

that avoid artifact-laden experimental results and resist the temptation to publish treatment differences whose practical, long-term implications are moot in

the real world?

10. Do we need to anticipate and measure potentially adverse, secondary effects of

bioremediation technology? These effects could result in nutrient limitations and/

or inadvertent release from soil of undesirable metals or chemical metabolites.
104

11. Research efforts to address bioavailability and bioremediation data have produced a broad spectrum of information about possible reactions that might occur

between geosorbents, contaminant wastes, and biota. This substantial roster of

possibilities needs to be narrowed to identify what actually does happen in realworld field sites. Can the true field behavior of contaminants be systematically

sampled and interpreted so as to enable robust predictions of contaminant behavior based on climate, region, contaminant class, and soil type?
105

Section 11

Literature Cited

1. Adriano, D. C. (ed.). 1992. Biogeochemistry of Trace Metals. Lewis Publishing, Boca Raton, FL.

2. Alexander, M. 1973. Nonbiodegradable and other recalcitrant molecules.

Biotechnol. Bioeng. 15:611-647.

3. Alexander, M. 1995. How toxic are toxic chemicals in soil? Environ. Sci.

Technol. 29:2713-217.

4. Alexander, M. 1997. Sequestration and bioavailability of organic compounds

in soil, pp. 43-136. In: Linz, D. G., and D. V. Nakles, (eds.) Environmentally

Acceptable Endpoints in Soil: Risk-Based Approach to Contaminated Site

Management Based on Availability of Chemicals in Soil. American Academy of Environmental Engineers, Annapolis, MD.

5. Alexander, M. 1999. Biodegradation and Bioremediation. 2nd ed. Academic

Press, New York, NY.

6. Alleman, B. C., and A. Leeson. 1999. Bioremediation of metals and inorganic

compounds, Fifth International In situ and On-site Symposium 5(4). Battelle

Press, Columbus, OH.

7. Allison, F. E. 1965. Soil Organic Matter and Its Role in Crop Production.

Elsevier, Amsterdam.

8. Alloway, B. J. (ed.). 1995. Heavy Metals in Soils, 2nd ed. Blackie Academic

and Professional, London.

9. Amann, R. I., W. Ludwig, and K.-H. Schleifer. 1995. Phylogenetic identification and in situ detection of individual microbial cells without cultivation.

Microbiol. Rev. 59:13-169.
106

10. Amann, R., and M. Kühl. 1998. In situ methods for assessment of microorganisms and their activities. Curr. Opin. Microbiol. 1:352-358.

11. Angehrn, D., M. Schluep, R. Gälli, and J. Zeyer. 1999. Movement and fate of

residual mineral oil contaminants in bioremediated soil. Environ. Toxic. Chem.

18:2224-2231.

12. Angehrn, D., R. Galli, and J. Zeyer. 1998. Physicochemical characterization of

residual mineral oil contaminants in bioremediated soil. Environ. Toxicol.

Chem. 17:2168-2175.

13. Atlas, R.M., and R. Bartha 1997. Microbial Ecology: Fundamentals and

Applications, 4th ed. Benjamin Cummings Publishing Co., Menlo Park, CA.

14. Babu, G. R. V., J. H Wolfram, and K. D. Chapatwala. 1992. Conversion of

sodium cyanide to carbon dioxide and ammonia by immobilized cells of

Pseudomonas putida. J. Ind. Microbiol. 9:235-238.

15. Ball, W. P. 1989. Equilibrium sorption and diffusion rate studies with halogenated organic chemicals and sandy aquifer material. Ph.D. Dissertation.

Stanford University.

16. Ball, W. P., and P. V. Roberts. 1991. Long-term sorption of halogenated organic chemicals by aquifer material. 2. Intraparticle diffusion. Environ. Sci.

Technol. 25:1237-1249.

17. Barber, S. A . 1995. Soil Nutrient Bioavailability: A Mechanistic Approach.

John Wiley & Sons, Inc., New York, NY.

18. Baveye, P., and R. Bladon. 1999. Bioavailability of organic xenobiotics in the

environment: a critical perspective, pp. 227-248. In: Baveye, P., J.-C. Block,

and V. V. Goncharuk (eds.). Bioavailability of Organic Xenobiotics in the

Environment. Kluwer Academic Publishing, Boston, MA.

19. Bayer, E. A., and R. Lamed. 1992. The cellulose paradox: pollutant par

excellence and/or a reclaimable natural resource? Biodegrad. 3:171-188.

20. Berg, M. S., R. C. Loehr, and M. T. Webster. 1998. Release of petroleum

hydrocarbons from bioremediated soils. J. Soil Contam. 7:675-695.

21. Bhadra, R., R. J. Spanggord, D. G. Wayment, J. B. Hughes, and J. V. Shanks.

1999. Characterization of oxidation products of TNT metabolism in aquatic

phytoremediation systems of Myriophyllum aquaticum. Environ. Sci. Technol.

33:3354-3361.

22. Bizily, W. P., C. L. Rugh, A. O. Summers, and R. B. Meagher. 1999. Phytoremediation

of methylmercury pollution: MerB expression in Arabidopsis thaliana confers

resistance to organomercurials. Proc. Natl. Acad. Sci. 96:6808-6813.
107

23. Blaylock, M. J. 1997. Phytoremediation of lead-contaminated soil at a

brownfield site in New Jersey—a cost-effective alternative. In: IBC’s 2nd

Annual International Conference on Phytoremediation, Seattle, WA. International Business Communications: Southborough, MA.

24. Blaylock, M. J., D. E. Salt, S. Dushenkov, O. Zakharova, C. Gussman, Y. Kapulnik,

B. D. Ensley, and I. Raskin. 1997. Enhanced accumulation of Pb in Indian

mustard by soil-applied chelating agents. Environ. Sci. Technol. 31:860-865.

25. Boetius, A., K. Ravenschlag, C. J. Schubert, D. Rickert, F. Widdel, A. Gieseke,

R. Amann, B. B. Jorgensen, U. Witte, and O. Pfannkuche. 2000. A marine

microbial consortium apparently mediating anaerobic oxidation of methane.

Nature 407:623-626.

26. Bosma, T. M. P., P. J. M. Middeldorp, G. Schraa, and A. J. B. Zehnder. 1997.

Mass transfer limitation of biotransformation: Quantifying bioavailability.

Environ. Sci. Technol. 31:248-252.

27. Brady, N. C., and R. R. Weil. 1999. The Nature and Properties of Soils,

Twelfth Edition. Prentice Hall. Upper Saddle River, NJ.

28. Brierley, C. L. 1990. Bioremediation of metal-contaminated surface and

groundwaters. Geomicrobiol. 8:201-224.

29. Brock, T. D. 1978. The poisoned control in biogeochemical investigations, pp.

717-725. In: W. E Krumbein (ed.) Environmental Biogeochemistry and Microbiology, Volume 3. Methods, Metals, and Assessment. Ann Arbor Science, Inc.

30. Brusseau, M. L., R. E. Jessup, and P. S. C. Rao. 1991. Nonequilibrium sorption

of organic chemicals—elucidation of rate-limiting processes. Environ. Sci.

Technol. 25:134-142.

31. Burford, M. D., S. B. Hawthorne, and D. J. Miller. 1993. Extraction rates of

spiked versus native PAHs from heterogeneous environmental samples using supercritical fluid extraction and sonication in methylene chloride. Anal.

Chem. 65:1497-1505.

32. Burken, J. G., and J. L. Schnoor. 1998. Predictive relationships for uptake of

organic contaminants by hybrid poplar trees. Environ. Sci. Technol. 32:3379-

33385.

33. Cawse, P.A. 1975. Microbiology and biochemistry of irradiated soils, pp.

213-267. In: Paul, E. A. and A. D. McLaren (eds.). Soil Biochemistry, Volume

3. Marcel Dekker, New York, NY.

34. Chaney, R. L., M. Malik, Y-M. Li, S. L. Brown, E. R. Brewer, J. S. Angle, and A.

J. M. Baker. 1997. Phytoremediation of soil metals. Curr. Opin. Biotechnol.

8:279-284.
108

35. Chapatwala, K. D., G. R. V. Babu, E. R. Armstead, E. M. White, and J. H.

Wolfram. 1995. A kinetic study on the bioremediation of sodium cyanide and

acetonitrile by free and immobilized cells of Pseudomonas putida. Appl.

Biochem. Biotechnol. 51-52:717-726.

36. Chen, H., and T. Cutright. 2001. EDTA and HEDTA effects on Cd, Cr, and Ni

uptake by Helianthus annuus. Chemosphere 45:21-28.

37. Chen, S., and D. B. Wilson. 1997. Genetic engineering of bacteria and their

potential for Hg-2+ bioremediation. Biodegrad. 8:97-103.

38. Compeau, G. C., W. D. Mahaffey, and L. Patras. 1991. Full-scale bioremediation

of contaminated soil and water, pp. 91-109. In: G. S. Sayler, R. Fox ,and J. W.

Blackburn (eds.). Environmental Biotechnology for Waste Treatment. Plenum Press, New York, NY.

39. Connaughton, D. F., J. R. Stedinger, L. W. Lion, and M. L. Shuler. 1993.

Description time-varying desorption kinetics: Release of naphthalene from

contaminated soils. Environ. Sci. Technol. 27:2397-2403.

40. Corbisier, P., D. van der Lelie, B. Borremans, A. Provoost, V. de Lorenzo, N. L.

Brown, J. R. Lloyd, J. L. Hobman, E. Csoregi, G. Johansson, and B. Mattiasson.

1999. Whole cell- and protein-based biosensors for the detection of bioavailable

heavy metals in environmental samples. Anal. Chem. Acta. 387:235-244.

41. Cornelissen, G., H. Rigterink, M. M. A. Ferdinandy, and P. C. M. Van Noort.

1998. Rapidly desorbing fractions of PAHs in contaminated sediments as a

predictor of the extent of bioremediation. Environ. Sci. Technol. 32:966-970.

42. Crawford, R. L., and D. L. Crawford (eds.). 1996. Bioremediation: Principles

and Applications. Cambridge University Press, New York, NY.

43. Criddle, C. S., P. L. McCarty, M. C. Elliott, and J. F. Barker. 1986. Reduction of

hexachloroethane to tetrachloroethylene in groundwater. J. Cont. Hydrol.

1:133-142.

44. Cunningham, S. D., J. R. Shann, D. E. Crowley, and T. A. Anderson. 1997.

Phytoremediation of contaminated water and soil, pp. 2-17. In: Kruger, E. L.,

T. A. Anderson, and J. R. Coates (eds.). Phytoremediation of Soil and Water

Contaminants. American Chemical Society, Washington, DC, 318 pp.

45. Diels, L. 1997. Heavy metal bioremediation of soil, pp. 283-295. In: Sheehan,

J. (ed.). Methods in Biotechnology, 2. Bioremediation protocols. Humana

Press, Inc., Totowa, NJ.

46. Dixon, J. B., S. B. Weed, J. A. Kittrick, M. H. Milford, and J. L. White (eds.). 1977.

Minerals in Soil Environments. Soil Science Society of America. Madison, WI.
109

47. Ebbs, S. D., and L. V. Kochian. 1998. Phytoextraction of zinc by oat (Avena

sativa), barley (Hordeum vulgare), and Indian mustard (Brassica juncea).

Environ. Sci. Technol. 32:802-806.

48. Ehrlich, H. L. 1996. Geomicrobiology, Third ed. Marcel Dekker, New York,

NY.

49. Ensley, B. D. 2000. Rationale for use of phytoremediation, pp. 3-11. In: Raskin,

I., and B. D. Ensley (eds.). Phytoremedititon of Toxic Metals Using Plants to

Clean Up the Environment. John Wiley & Sons, New York, NY, 304 pp.

50. EPA. 2001. Treatment Technology for Site Cleanup: Annual Status Report.

Tenth Edition. EPA-542-R-01-004. US EPA. Washington, DC.

51. Flaig, W. 1975. Biochemistry of soil organic matter. FAO Soils Bull. 27:31-35.

52. Flathman, P. E., D. E. Jerger, and J. H. Exner. 1994. Bioremediation: Field

Experience. Lewis Publishers, Boca Raton, FL.

53. Fogel, S. 1993. Full-scale bioremediation of No. 6 fuel soil-contaminated soil:

6 months of active and 3 years of passive treatment, pp. 161-175. In: Flathman,

P. E., D. E. Jerger, and J. H. Exner (eds.). Bioremediation Field Experience.

Lewis Publishers, Boca Raton, FL.

54. Foster, R. C. 1993. The ultramicromorphology of soil biota in situ in natural

soils: a review, pp. 381-393. In: A. J. Ringrose-Voase, and G. S . Humphreys

(eds.). Soil Micromorphology: Studies in Management and Genesis. Elsevier

Science, New York, NY.

55. Francis, A. J. 1999. Bioremediation of radionuclide and toxic metal contaminated soils and wastes, pp. 239-271. In: Adriano, D. C. (ed.). Bioremediation

of Contaminated Soils. Soil Sci. Soc. Amer., Madison, WI.

56. Frankenberger, W. T., Jr. and M. E. Losi. 1995. Applications of bioremediation

in the cleanup of heavy metals and metalloids, pp. 173-210. In: Skipper, H. D.

and R. F. Turco (eds.). Bioremediation Science and Applications. SSSA,

Special Publication Number 43, Madison, WI.

57. Galvez-Cloutier, R., and J.-S. Dubé. 1998. An evaluation of freshwater sediments contamination: the lachine canal sediments case, Montréal, Canada.

Part II: Heavy metal particulate speciation study. Water, Air, and Soil Pollut.

102:281-302.

58. Ghiorse, W. C. 1994. Iron and manganese oxidation and reduction, pp. 1079-

1096. In: Weaver, R. W., S. Angle, and P. Bottomley (eds.). Methods of Soil

Analysis, Part 2, Microbiological and Biochemical Properties. ASA-SSSA,

Madison, WI.
110

59. Ghiorse, W. C., D. N. Miller, R. L. Sandoli, and P. L. Siering. 1996. Applications

of laser scanning microscopy for analysis of aquatic microhabitats. Micro.

Res. Tech. 33:73-86.

60. Ghosh, U., J. W. Talley, and R. G. Luthy. 2001. Particle-scale investigation of

PAH desorption kinetics and thermodynamics from sediment. Environ. Sci.

Technol. 35:3468-3475.

61. Ghosh, U., R. G. Luthy, J. S. Gillette, and R. N. Zare. 2000. Microscale location,

characterization, and association of polycyclic aromatic hydrocarbons on

harbor sediment particles. Environ. Sci. Technol. 34:1729-1736.

62. Gillette, J. S., R. G. Luthy, S. J. Clemett, and R. N. Zare. 1999. Direct observation of polycyclic aromatic hydrocarbons on geosorbents at the subparticle

scale. Environ. Sci. Technol. 33:1185-1192.

63. Goldstein, J. I., D. E . Newbury, P. Echlin, D. C. Joy, A. D. Romig, Jr., C. E.

Lyman, D. Fiori, and E. Lifshin. 1992. Scanning electron microscopy and Xray microanalysis: a text for biologists. Materials Scientists, and Geologists. Second ed. Plenum Press, New York, NY.

64. Graber, E. R., and M. D. Borisover. 1998. Evaluation of the glassy/rubbery

model for soil organic matter. Environ. Sci. Technol. 32:3286-3292.

65. Guerin, W. F., and S. A. Boyd. 1992. Differential bioavailability of soil-sorbed

naphthalene to two bacterial species. Appl. Environ. Microbiol. 58:1142-

1152.

66. Guha, S., P. R. Jaffe, and C. A. Peters. 1998. Bioavailability of mixtures of

PAHs partitioned into the micellar phase of a nonionic surfactant. Environ.

Sci. Technol. 32:2317-2324.

67. Guthrie, E. A., and F. K. Pfaender. 1998. Reduced pyrene bioavailability in

microbially active soils. Environ. Sci. Technol. 32:501-508.

68. Guthrie, E. A., J. M. Bortiatynski, J. D. H. Van Heemst, J. E. Richman, K. S.

Hardy, E. M. Kovach, and P. G. Hatcher. 1999. Determination of [13C]pyrene

sequestration in sediment microcosms using flash pyrolysis––GCMS and

13C NMR. Environ. Sci. Technol. 33:119-125.

69. Hartwig, R. C., and R. H. Loeppert. 1993. Evaluation of soil iron, pp. 465-482.

In: L. L. Barton and B. C. Hemming (eds.). Iron Chelation in Plants and Soil

Microorganisms. Academic Press, Inc., New York, NY.

70. Hatzinger, P. B., and M. Alexander. 1995. Effect of aging of chemicals in soil

on their biodegradability and extractability. Environ. Sci. Technol. 29:537-

545.
111

71. Havlin, J. L., J. D . Beaton, S. L. Tisdall, and W. L. Nelson. 1999. Soil Fertility

and Fertilizers: An Introduction to Nutrient Management, 6th ed. Prentice

Hall, Upper Saddle River, NJ.

72. Heitzer, A., K. Malachowsky, J. E . Thonnard, P. R. Bienkowski, D. C. White,

and G. S. Sayler. 1994. Optical biosensor for environmental on-line monitoring of naphthalene and salicylate bioavailability with an immobilized bioluminescent catabolic reporter bacterium. Appl. Environ. Microbiol. 60:1487-

1494.

73. Hinchee, R. E., and R. F. Olfenbuttel. 1994. In situ Bioremediation: Applications

and Investigations for Hydrocarbon and Contaminated Site Remediation.

Battelle Press, Columbus, OH.

74. Hinchee, R. E., J. L. Means, and D. R. Burris (eds.). 1995. Bioremediation of

Inorganics. Battelle Press, Columbus, OH.

75. Hinrichs, H.-U., J. M Hayes, S. P. Sylva, P. G. Brewer, and E. F. DeLong. 1999.

Methane-consuming archaebacteria in marine sediments. Nature 398:802-

805.

76. Höfte, M. 1993. Classes of microbial siderophores, pp. 3-26. In: Barton, L. L.

and B. C. Hemming (eds.). Iron Chelation in Plants and Soil Microorganisms. Academic Press, Inc., New York, NY.

77. Huang, J. W., J. Chen, and S. D. Cunningham. 1997. Phytoextraction of lead

from contaminated soils, pp. 283-298. In: Kruger, E. L., T. A. Anderson, and J.

R. Coats (eds.). Phytoremediation of Soil and Water Contaminants. American Chemical Society, Washington, DC.

78. Huesemann, M. H. 1995. Predictive model for estimating the extent of petroleum hydrocarbon biodegradation in contaminated soils. Environ. Sci.

Technol. 29:7-18.

79. Hughes, J. B., D. M. Beckles, S. D. Chandra, and C. H. Ward. 1997. Utilization

of bioremediation processes for the treatment of PAH-contaminated sediments. J. Indus. Microbiol. Biotechnol. 18:152-160.

80. Jerger, D. E., P. M. Woodhull, P. E. Flathman, and J. H. Exner. 1993. Solidphase bioremediation of petroleum hydrocarbon-contaminated soil: laboratory treatability study through site closure, pp. 177-193. In: Flathman, P. E.,

D. E. Jerger, and J. H. Exner (eds.). Bioremediation Field Experience. Lewis

Publishers, Boca Raton, FL.

81. Johnson, S. E., J. S. Herman, A. L. Mills, and G. M. Hornberger. 1999.

Bioavailability and desorption characteristics of aged, nonextractable atrazine in soil. Environ. Toxicol. Chem. 18:1747-1754.
112

82. Kalin, M., J. Cairns, and R. M. McCready. 1991. Ecological engineering methods for acid mine drainage treatment of coal wastes. Resources Conserv.

Recycl. 5:265-276.

83. Khan, A. G., C. Kuek, T. M. Chaudhry, C. S. Khoo, and W. J. Hayes. 2000. Role

of plants, mycorrhizae and phytochelators in heavy metal contaminated land

remediation. Chemosphere 41:197-207.

84. Khan, S. U., and R. M. Behki. 1990. Effects of Pseudomonas species on the

release of bound 14C residues from soil treated with [14C]atrazine. J. Agric.

Food Chem. 38:2090-2093.

85. King, J. M. H., P. M. Digrazia, B. Applegate, R. Burlage, J. Sanseverino, P. Dunbar,

F. Larimer, and G. S. Sayler. 1991. Rapid, sensitive bioluminescent reporter technology for naphthalene exposure and biodegradation. Science 249:778-781.

86. Ladd, J. N., R. C. Foster, P. Nannipieri, J. M. Oades. 1996. Soil structure and

biological activity, pp. 23-78. In: G. Stotzky, and J-M. Bollag (eds.). Soil Biochemistry, Volume 9. Marcel Dekker, Inc., New York, NY.

87. Leavitt, M. E., D. A. Graves, and C. A. Lang. 1991. Evaluation of bioremediation

in a coal-coking waste lagoon, pp. 71-84. In: Sayler, G. S., R. Fox, and J. W.

Blackburn (eds.). Environmental Biotechnology for Waste Treatment. Plenum Press, New York, NY.

88. Lee, B.-G., S. B. Griscom, J.-S. Lee, H. J. Choi, C.-H. Koh, S. N. Luoma, and

N. S. Fisher. 2000. Influences of dietary uptake and reactive sulfides on metal

bioavailability from aquatic sediments. Science 287:282-284.

89. Lee, N., P. H. Nielsen, K. H. Andreasen, S. Juretschko, J. L. Nielsen, K.-H.

Schleifer, and M. Wagner. 1999. Combination of fluorescent in situ hybridization and microautoradiophy—a new tool for structure-function analyses

in microbial ecology. Appl. Environ. Microbiol. 65:1289-1297.

90. Lengeler, J. W., G. Drews, and H. G. Schlegel. 1999. Biology of the Prokaryotes. Blackwell Scientific, Malden, MA.

91. Lenhard, R. J., R. S. Skeen, and T. M. Brouns. 1995. Contaminants at U.S.

DOE sites and their susceptibility to bioremediation, pp. 157-172. In: Skipper,

H. D. and R. F. Turco (eds.). Bioremediation Science and Applications.

SSSA, Special Publication Number 43, Madison, WI.

92. Lindsay, W. L. 1979. Chemical Equilibria in Soils. John Wiley & Sons, New

York, NY.

93. Linz, D. G., and D. V. Nakles. 1997. Next steps, pp. 467-486. In: Linz, D. G. and

D. V. Nakles (eds.). Environmentally Acceptable Endpoints in Soil. American Academy of Environmental Engineers.
113

94. Lock, K., and. C. R. Janssen. 2001. Ecotoxicity of zinc in spiked artificial soils

versus contaminated field soils. Environ. Sci. Technol. 35:4295-4300.

95. Loehr, R. C., and M. T. Webster. 1997. Effects of treatment on contaminant

availability, mobility, and toxicity, pp. 137-386. In: D. G. Linz and D. V. Nakles

(eds.). Environmentally Acceptable Endpoints in Soil. American Academy

of Environmental Engineers.

96. Loehr, R. C. 1999. Overview, pp. 10-17. In: Anderson, W. C., R. C. Loehr, and B.

P. Smith (eds.). Environmental Availability of Chlorinated Organics, Explosives, and Metals in Soils. American Academy of Environmental Engineers

Publishing, Annapolis, MD.

97. Lollar, B. S., G. F. Slater, B. Sleep, M. Witt, G. M. Klecka, M. Harkness,and J.

Spivack. 2001. Stable carbon isotope evidence for intrinsic bioremediation of

tetrachloroethene and trichloroethene at Area 6, Dover Air Force Base.

Environ. Sci. Technol. 35:261-269.

98. Lovley, D. R. 1993. Dissimilatory metal reduction. Annu. Rev. Microbiol.

47:263-290.

99. Lovley, D. R. 1995a. Bioremediation of organic and metal contaminants with

dissimilatory metal reduction. J. Ind. Microbiol. 14:85-93.

100. Lovley, D. R. 1995b. Microbial reduction of iron, manganese, and other metals. Adv. Agron. 54:176-231.

101. Lovley, D. R. 2000. Fe(III) and Mn(IV) reduction, pp. 3-30. In: Lovley, D. R.

(ed.) Environmental Microbe-Metal Interactions. ASM Press, Washington, DC, 395 pp.

102. Lovley, D. R., and J. D. Coates. 1997. Bioremediation of metal contamination.

Curr. Opin. Biotechnol. 8:285-289.

103. Lovley, D. R., J. D. Coates, E. L. Blunt-Harris, E. J. P. Phillips, and J. C.

Woodward. 1996. Humic substances as electron acceptors for microbial respiration. Nature 382:445-448.

104. Luthy, R. G., G. R. Aiken, M. L. Brusseau, S. D. Cunningham, P. M. Gschwend,

J. J. Pignatello, M. Reinhard, S. J. Traina, W. J. Weber, Jr., and J. C. Westall.

1997. Sequestration of hydrophobic organic contaminants by geosorbents.

Environ. Sci. Technol. 31:3341-3347.

105. Macaskie, L. E., and G. Basnakova. 1998. Microbially-enhanced chemosorption

of heavy metals: a method for the bioremediation of solutions containing longlived isotopes of neptunium and plutonium. Environ. Sci. Technol. 32:184-187.
114

106. Mackay, D. M., W. P. Ball, and M. G. Durant. 1986. Variability of aquifer

sorption properties in a field experiment on groundwater transport of organic solutes: methods and preliminary results. J. Cont. Hydrol. 1:119-132.

107. Madigan, M. T., J. M. Martinko, and J. Parker. 2000. Brock Biology of Microorganisms, 9th ed. Prentice Hall, Upper Saddle River, NJ.

108. Madsen, E. L. 1991. Determining in situ biodegradation: facts and challenges. Environ. Sci. Technol. 25:1662-1673.

109. Madsen, E. L. 1996. A critical analysis of methods for determining the composition and biogeochemical activities of soil microbial communities in situ,

pp. 287-370. In: G. Stotzky, and J-M (eds.). Bollag. Soil Biochemistry, Volume

9. Marcel Dekker, Inc., New York, NY.

110. Madsen, E. L. 2002. Methods for determining biodegradability, pp. 903-913.

In: Hurst, C. J., et al. (eds.). Manual of Environmental Microbiology, 2nd

Ed. American Society for Microbiology, Washington, DC.

111. Madsen, E. L. 1998a. Epistemology of environmental microbiology. Environ.

Sci. Technol. 32:429-539.

112. Madsen, E. L. 1998b. Theoretical and applied aspects of bioremediation: The

influence of microbiological processes on organic compounds in field sites,

pp. 354-407. In: Burlage, R., R. Atlas, D. Stahl, G. Geesey, and G. Sayler (eds.).

Techniques in Microbial Ecology. Oxford University Press, New York, NY.

113. Madsen, E. L., C. L. Mann, and S. Bilotta. 1996. Oxygen limitations and aging

as explanation for the persistence of naphthalene in coal-tar contaminated

surface sediments. Environ. Toxicol. Chem. 15:1876-1882.

114. Maier, R. 2000. Bioavailability and its importance to bioremediation. In: Valdes,

J. J. (ed.). International Society for Environmental Biotechnology: Environmental Monitoring and Biodiagnostics. Kluwer Academic Publishers, pp. 59-

78.

115. Malmqvist, A., T. Welander, and L. Gunnarsson. 1991. Anaerobic growth of

microorganisms with chlorate as an electron acceptor. Appl. Environ.

Microbiol. 57:2229-2232.

116. Marschner, H.1995. Mineral Nutrition of Higher Plants, Second ed. Academic Press, New York, NY.

117. Marshall, K. C. 1976. Interfaces in Microbial Ecology. Harvard University

Press, Cambridge, MA.

118. Martinez, C. E., and M. B. McBride. 2001. Cd, Cu, Pb, and Zn coprecipitates

in Fe oxide formed at different pH; aging effects on metal solubility and

extractability by citrate. Environ. Toxicol. Chem. 20:122-126.
115

119. McBride, M. B. 1994. Environmental Chemistry of Soils. Oxford University

Press. New York, NY.

120. McCray, J. E., and M. L. Brusseau. 1999. Cyclodextrin-enhanced in situ flushing of multiple-component immiscible organic liquid contamination at the field

scale: analysis of dissolution behavior. Environ. Sci. Technol. 32:89-95.

121. McGrath, S. P., B. Knight, K. Killham, S. Preston, and G. I. Paton. 1999. Assessment of the toxicity of metals in soils amended with sewage sludge

using a chemical speciation technique and a lux-based biosensor. Environ.

Tox. Chem. 18:659-663.

122. McGrath, S. P., F. J. Zhao, and E. Lombi. 2001. Plant and rhizosphere processes involved in photoremediation of metal-contaminated soils. Plant and

Soil 232:207-214.

123. McHale, A. P., and S. McHale. 1994. Microbial biosorption of metals: potential in the treatment of metal pollution. Biotechnol. Adv. 12:647-652.

124. Meagher, R. B. 2000. Phytoremediation of toxic elemental and organic pollutants. Curr. Opin Plant Biol. 3:153-162.

125. Mihelcic, J. R., D. R. Lueking, R. J. Mitzell, and J. M. Stapleton. 1993. Bioavailability

of sorbed- and separate-phase chemicals. Biodegradation 4:141-153.

126. Morris, D. E., P. G. Allen, J. M. Berg, C. J. Chisholm-Brause, S. D. Conradson,

R. J. Donohoe, N. J. Hess, J. A. Musgrave, and C. D. Tait. 1996. Speciation of

uranium in Fernald soils by molecular spectroscopic methods: Characterization of untreated soils. Environ. Sci. Technol. 30:2322-2331.

127. Morrison, D. E., B. K. Robertson, and M. Alexander. 2000. Bioavailability to

earthworms of aged DDT, DDE, DDD, and dieldrin in soil. Environ. Sci.

Technol. 34:709-713.

128. Myneni, S. C. B, J. T. Brown, G. A. Martinez, and W. Meyer-Ilse. 1999. Imaging of humic substance macromolecular structures in water and soils. Science 286:1335-1337.

129. National Research Council (NRC). 1993. In situ Bioremediation: When Does

It Work? National Academy Press, Washington, DC.

130. National Research Council (NRC). 1997. Innovations in Ground Water and

Soil Cleanup. National Academy Press, Washington, DC.

131. National Research Council (NRC). 2000. Natural Attenuation for Groundwater Remediation. National Academy Press, Washington, DC.

132. Newman, A. D. C., and A. J. Thomasson. 1979. Rothamsted studies of soil structure III: pore size distribution and shrinkage processes. J. Soil Sci. 30:415-439.
116

133. Newman, D. K., and R. Kolter. 2000. A role for excreted quinines in extracellular electron transfer. Nature 405:94-97

134. Noordman, W. H., W. Ji, M. L. Brusseau, and D. B. Janssen. 1998. Effects of

rhamnolipid biosurfactants on removal of phenanthrene from soil. Environ.

Sci. Technol. 32:1806-1812.

135. Norris, R. D., R. E. Hinchee, R. Brown, P. L. McCarty, L. Semprini, J. T. Wilson,

D. G. Kampbell, M. Reinhard, D. J. Bouwer, R. C. Borden, T. M. Vogel, J. M.

Thomas. and C. H. Ward. 1994. Handbook of Bioremediation. Lewis Publishers, Boca Raton, FL.

136. Ogram, A. V., R. E. Jessup, L. T. Ou, and P. S. C. Rao. 1985. Effects of sorption

on biological degradation rates of (2,4-dichlorophenoxy) acetic acid in soils.

Appl. Environ. Microbiol. 49:582-587.

137. Olivera, F. L., R. C. Loehr, B. C. Coplin, H. Eby, and M. T. Webster. 1998.

Prepared bed land treatment of soils containing diesel and crude oil hydrocarbons. J. Soil Cont. 7:657-674.

138. Orphan, V. J., C. H. House, K-U. Hinrichs, K. D. McKeegan, and E. F. DeLong.

2001. Methane-consuming archaea revealed by directly coupled isotopic

and phylogenetic analysis. Science 293:484-487.

139. Ouverney, C. C., and J. A. Fuhrman. 1997. Increase in fluorescence intensity

of 16S rRNA in situ hybridization in natural samples treated with chloramphenicol. Appl. Environ. Microbiol. 63:2735-2740.

140. Ouverney, C. C., and J. A. Fuhrman. 1999. Combined microautoradiography—

16S rRNA probe technique for determination of radioisotope uptake by specific microbial cell types in situ. Appl. Environ. Microbiol. 65:1746-1752.

141. Page, W. J. 1993. Growth conditions for the demonstration of siderophores

and iron-repressible outer membrane proteins in soil bacteria, with an emphasis on free-living Diazotrophs, pp. 75-110. In: Barton, L. L. and B. C.

Hemming (eds.). Iron Chelation in Plants and Soil Microorganisms. Academic Press, Inc., New York, NY.

142. Pedersen, M. B., C. Kjaer, and N. Elmegaard. 2000. Toxicity and

bioaccumulation of copper to black bindweed (Fallopia convolvulus) in

relation to bioavailability and the age of soil contamination. Arch. Environ.

Contam. Toxicol. 39:431-439.

143. Pignatello, J. J., and B. Xing. 1996. Mechanisms of slow sorption of organic

chemicals to natural particles. Environ. Sci. Technol. 30:1-11.
117

144. Pignatello, J. J., and B. Xing. 1999. Pignatello and Xing’s comment on “Evaluation of the glassy/rubbery model for soil organic matter.” Environ. Sci.

Technol. 33:2837-2838.

145. Portier, R. J., and J. A. Christiansen. 1993. Closure of an RCRA surface impoundment by employing a modified biological treatment approach, pp. 225-

233. In: Flathman, P. E., D. E. Jerger, and J. H. Exner (eds.). Bioremediation

Field Experience. Lewis Publishers, Boca Raton, FL.

146. Reid, B. J., K. C. Jones, and K. T. Semple. 2000. Bioavailability of persistent

organic pollutants in soils and sediments—a perspective on mechanisms,

consequences and assessment. Environ. Poll. 108:103-112.

147. Rikkin, G. B., A. G. M. Kroon, and C. G. VanGinkel. 1996. Transformation of

(per)chlorate into chloride by a newly isolated bacterium: reduction and

dismutation. Appl. Microbiol. Biotechnol. 45:420-426.

148. Ringrose-Voase, A. J., and G. S. Humphreys (eds.). 1994. Soil Micromorphology: Studies in Management and Genesis. Elsevier Science, New York, NY.

149. Rittmann, B. E., E. Seagren, B. A. Wrenn, A. J. Valocchi, C. Ray, and L. Faskin.

1994. In Situ Bioremediation, Second ed. Noyes Publications. Park Ridge,

NJ.

150. Roemheld, V., and H. Marschner. 1986. Evidence for a specific uptake system

for iron phytosiderophores in roots of grasses. Plant Physiol. 80:175-180.

151. Salanitro, J. P, P. B. Dorn, M. H. Huesemann, K. O. Moore, I. A. Rhodes, L. M.

Rice-Jackson, T. E. Vipond, M. M. Western, and H. L. Wisniewski. 1997.

Crude oil hydrocarbon bioremediation and soil ecotoxicity assessment.

Environ. Sci. Technol. 31:1769-1776.

152. Sandoli, R. S., W. C. Ghiorse, and E. L. Madsen. 1996. Regulation of microbial

phenanthrene mineralization in sediments by sorbent-sorbate contact time,

inoculum, and gamma irradiation-induced sterilization artifacts. Environ.

Toxicol. Chem. 15:1901-1907.

153. Saouter, E., M. Gillman, and T. Barkay. 1995. An evaluation of mer-specified

reduction of ionic mercury as a remedial tool of a mercury-contaminated

freshwater pond. J. Ind. Microbiol. 14:343-348.

154. Schlegel, H. G., and H. W. Jannasch. 1992. Prokaryotes and their habitats, pp.

75-125. In: Balows, A., H. G. Trüper, M. Dworkin, W. Harder, and K.-H. Schleifer

(eds.). The Prokaryotes, Second edition. Springer-Verlag, New York, NY.

155. Schlesinger, W. H. 1991. Biogeochemistry: An Analysis of Global Change.

Academic Press, Inc., New York, NY.
118

156. Schnoor, J. L. 2000. Phytostabilization of metals using hybrid poplar trees,

pp. 133-150. In: Raskin, I. and B. D. Ensley (eds.). Phytoremediaiton of Toxic

Metals. Wiley & Sons, Inc., New York, NY.

157. Scholz-Muramatsu, H., et al. 1995. Isolation and Characterization of Dehalospirillum multivorans gen., sp. nov., a tetrachloroethene-utilizing, strictly

anaerobic bacterium. Arch. Microbiol. 163:48-56.

158. Schultz, L. F., M. Young, and R. M. Higashi. 1999. Sorption-desorption behavior of phenanthrene elucidated by pyrolysis—gas chromatography-mass spectrometry studies of soil organic matter. Environ. Toxicol. Chem. 18:1710-1719.

159. Schwartz, E., and K. M. Scow. 1999. Using biodegradation kinetics to measure availability of aged phenanthrene to bacteria inoculated into soil. Environ.

Toxicol. Chem. 18:1742-1746.

160. Schwarzenbach, R. C., R. W. Scholz, A. Heitzer, B. Stäubli, and B. Grossmann.

1999. A regional perspective on contaminated site remediation—fate of materials and pollutants. Environ. Sci. Technol. 33:2305-2310.

161. Schwertmann, U., and R. M. Taylor. 1977. Iron oxides, pp. 145-180. In: Dixon,

J. B., S. B. Weed, J. A. Kittrick, M. H. Milford, and J. L. White (eds.). Minerals

in Soil Environments. Soil Science Society of America. Madison, WI.

162. Scow, K. M., and C. R. Johnson. 1997. Effect of sorption on biodegradation

of soil pollutants, pp. 1-56. In: D. L. Sparks (ed.). Advances in Agronomy,

Volume 58, Academic Press, Inc., New York, NY.

163. Scribner, S. L., T. R. Benzing. S. Sun, and S. A. Boyd. 1992. Desorption and

bioavailability of aged simazine residues in soil from a continuous corn field.

J. Environ. Qual. 21:115-120.

164. Seagren, E.A., B.E. Rittmann and A.J. Valocchi. 1994. Quantitative evaluation

of the enhancement of NAPL-pool dissolution by flushing and biodegradation. Environ. Sci. Technol. 28:833-839.

165. Seeliger, S., R. Cerd-Ruwisch, and B. A. Schink. 1998. A peripasmic and extracellular c-type cytochrome of Geobacter sulfurreducens acts as a ferric iron

reductase and as an electron carrier to other acceptors or to partner bacteria.

J. Bacteriol. 180:3686-3691.

166. Semprini, L. 1997a. In Situ Transformation of Halogenated Aliphatic Compounds under Anaerobic Conditions, pp. 429-450. In: Ward, C. H., J. A.

Cherry and M. R. Scalf (eds.). Subsurface Restoration, Ann Arbor Press,

Inc., Chelsea, MI.

167. Semprini, L. 1997b. Strategies for the aerobic co-metabolism of chlorinated

solvents. Curr. Opinion Biotechnol. 8: 296-308.
119

168. Semprini, L., P. K. Kitanidis, D. H. Kampbell, and J. T. Wilson. 1995. Anaerobic transformation of chlorinated aliphatic hydrocarbons in a sand aquifer

based on spatial chemical distributions. Water Resour. Res. 31:1051-1062.

169. Semprini, L., P. V. Roberts, G. D. Hopkins, and P. L. McCarty. 1990. A field

evaluation of in-situ biodegradation of chlorinated ethenes: Part 2, results of

biostimulation and biotransformation experiments. Ground Water 28:715-727.

170. Sharma, P., and P. L. McCarty. 1996. Isolation and characterization of a facultative aerobic bacterium that reductively dehalogenates tetrachloroethene

to cis-1,2-dichloroethene. Appl. Environ. Microbiol. 62: 761-765.

171. Siciliano, S. D., and J. J. Germida. 1998. Mechanisms of phytoremediation:

biochemical and ecological interactions between plants and bacteria. Environ.

Rev. 6:65-79.

172. Skipper, H. D., and R. F. Turco. (eds.). 1995. Bioremediation: Science and

Applications. Soil Science Society of America, Madison, WI.

173. Smatlak, C. R., J. M. Gossett, and S. H. Zinder. 1996. Comparative kinetics of

hydrogen utilization for reductive dechlorination of tetrachloroethene and

methanogenesis in an anaerobic enrichment culture. Environ. Sci. Technol.

30: 2850-2858.

174. Southgate, D., I. Johnson, and G. R. Fenwick (eds.). 1989. Nutrient Availability: Chemical and Biological Aspects. Royal Society of Chemistry, Cambridge, UK.

175. Sposito, G. 1989. The Chemistry of Soils. Oxford University Press, New York, NY.

176. Stacey, S., G. Merrington, and M. J. McLaughlin. 2001. The effect of aging

biosolids on the availability of cadmium and zinc in soil. Europ. J. Soil Sci.

52:313-321.

177. Stanhope, K. G., S. D. Young, J. J. Hutchinson, and R. Kamath. 2000. Use of

isotopic dilution techniques to assess the mobilization of nonlabile Cd by

chelating agents in phytoremediation. Environ. Sci. Technol. 34:4123-4127.

178. Steffan, R. J., K. L. Sperry, M. T. Walsh, S. Vainberg, and C. W. Condee. 1999.

Field-scale evaluation of in situ bioaugmentation for remediation of chlorinated solvents in groundwater. Environ. Sci. Technol. 33:2271-2281.

179. Steinberg, S. M., J. J. Pignatello, and B. L. Sawhney. 1987. Persistence of 1,2-

dibromoethane in soils: entrapment in intraparticle micropores. Environ.

Sci. Technol. 21:1201-1208.

180. Stevenson, F. J. 1994. Humus Chemistry, Second Edition. Wiley & Sons, New

York, NY.
120

181. Stout, J. D., K. M. Goh, and T. A. Rafter. 1981. Chemistry and turnover of naturally

occurring resistant organic compounds in soil, pp. 1-73. In: Paul, E. A. and J. N.

Ladd (eds.). Soil Biochemistry, Vol. 5. Marcel Dekker, Inc., New York, NY.

182. Summers, A. O. 1992. The hard stuff: metals in bioremediation. Curr. Opinion Biotechnol. 3:271-276.

183. Takagi, S-I. 1993. Production of phytosiderophores, pp. 111-131. In: Barton,

L. L. and B. C. Hemming (eds.). Iron Chelation in Plants and Soil Microorganisms. Academic Press, Inc., New York, NY.

184. Tang, J., B. K. Robertson, and M. Alexander. 1999. Chemical-extraction methods to estimate bioavailability of DDT, DDE, and DDD in soil. Environ. Sci.

Technol. 33:4346-4351.

185. Thompson-Eagle, E. C., and W. T. Frankenberger, Jr. 1992. Bioremediation of

soils contaminated with selenium. Adv. Soil Sci. 17:261-310.

186. Tisdall, J. M., and J. M. Oades. 1982. Organic matter and water-stable aggregates in soil. J. Soil Sci. 33:141-163.

187. Torrens, J. L., D. C. Herman, and R. Miller-Maier. 1998. Biosurfactant (rhamnolipid) sorption and the impact on rhamnolipid-facilitated removal of cadmium

from various soils under saturated flow conditions. Environ. Sci. Technol.

32:776-781.

188. Troy, M. A., S. W. Berry, and D. E. Jerger. 1993. Biological land treatment of

diesel fuel-contaminated soil: Emergency response through closure, pp. 145-

160. In: Flathman, P. E., D. E. Jerger, and J. H. Exner (eds.). Bioremediation

Field Experience. Lewis Publishers, Boca Raton, FL.

189. US EPA. 1999. Treatment technologies for site cleanup: annual status report. 9th ed. EPA-542-R99-00, Number 9, April 1999.

190. Van der Lelie, D., J.-P. Schwitzbuebel, D. J. Glass, J. Vangronsveld, and A.

Baker. 2001. Assessing phytoremediation’s progress in the United States

and Europe. Environ. Sci. Technol. 35:447-452.

191. van Ginkel, C. G., C. M. Plugge, and C. A. Stroo. 1995. Reduction of chlorate

with various energy substrates and inocula under anaerobic conditions.

Chemosphere 31:4057-4066.

192. Van Loosdrecht, M. C. M., J. Lyklema, W. Norde, and A. J. B. Zehnder. 1990.

Influence of interfaces on microbial activity. Microbiol. Rev. 54:75-87.

193. Vandegrift, G. F., D. T. Reed, and I. R. Tasker (eds.). 1992. Environmental

Remediation: Removing Organic and Metal Ion Pollutants. American Chemical Society, Washington, DC.
121

194. Vangronsveld, J., and S. D. Cunningham. 1998. Introduction to the Concepts, pp. 1-15. In: Vangronsveld, J. and S. D. Cunningham (eds.). MetalContaminated Soils: In Situ Inactivation and Phytorestoration. SpringerVerlag, New York, NY.

195. Vangronsveld, J., and S. D. Cunningham. 1998. Metal-Contaminated Soils:

In situ Inactivation and Phytorestoration. Springer-Verlag, New York, NY,

265 pp.

196. vanVeen, J. A., and E. A. Paul 1981. Organic carbon dynamics in grassland

soils. 1. Background information and computer simulation. Can. J. Soil Sci.

61:185-201.

197. Verstraete, W., and W. Devliegher. 1997. Formation of non-bioavailable organic residues in soil: Perspective for site remediation. Biodegradation 7:471-

485.

198. Videla, H. A., and W. G. Characklis. 1992. Biofouling and microbially influenced corrosion. Int. Biodeterior. Biodegrad. 29:195-212.

199. Volkering, F., A. M. Breure, and W. H. Rulkens. 1998. Microbiological aspects

of surfactant use for biological soil remediation. Biodegradation 8:401-417.

200. Wackett L. P., and C. L. Hershberg. 2000. Biocatalysis and Biodegradation.

American Society for Microbiology, Washington, DC, 228 pp.

201. Wackett, L. P. 1996. Co-metabolism: is the emperor wearing new clothes?

Curr. Opin Biotechnol. 7:321-325.

202. Wallace, W., S. Beshear, D. Williams, S. Hospadar, and M. Owens. 1998.

Perchlorate reduction by a mixed culture in an up-flow anaerobic fixed bed

reactor. J. Indust. Microbiol. Biotechnol. 20:126-131.

203. Warren, L. A., and F. G. Ferris. 1998. Continuum between sorption and precipitation of Fe(III) on microbial surfaces. Environ. Sci. Technol. 32:2331-

2337.

204. Welter, E., W. Calmano, S. Mangold, and L. Tröger, 1999. Chemical speciation

of heavy metals in soils by use of XAFS spectroscopy and electron microscopical techniques. Fresenius J. Anal. Chem. 364:238-244.

205. West, C. C., and J. H. Harwell. 1992. Surfactants and subsurface remediation.

Environ. Sci. Technol. 26:2324-2330.

206. White, J. C., and J. J. Pignatello. 1999. Influence of bisolute competition on

the desorption kinetics of polycyclic aromatic hydrocarbons in soil. Environment. Sci. Technol. 33:4292-4298.
122

207. White, J. C., M. Hunter, J. J. Pignatello, and M. Alexander. 1999. Increase in

bioavailability of aged phenanthrene in soils by competitive displacement

with pyrene. Environ. Toxicol. Chem. 18:1728-1732.

208. Whitlock, J. L. 1990. Biological detoxification of precious metal processing

wastewaters. Geomicrobiol. J. 8:241-249.

209. Willardson, B. M., J. F. Wilkins, T. A. Rand, J. M. Schupp, K. K. Hill, P. Keim,

and P. J. Jackson. 1998. Development and testing of a bacterial biosensor for

toluene-based environmental contaminants. Appl. Environ. Microbiol. 64:1006-

1012.

210. Williams, C. J., D. Aderhold, and R. G. J. Edyvean. 1998. Comparison between

biosorbents for the removal of metal ions from aqueous solutions. Water

Res. 32:216-224.

211. Wolf, D. C., and H. D. Skipper. 1994. Soil sterilization, pp. 41-51. In: Weaver, R.

W., S. Angle, P. Bottomley, D. Bezdicek, S. Smith, A. Tabatabai, and A. Wollum

(eds.). Methods of Soil Analysis, Part 2. Microbiological and Biochemical

Properties. Soil Science Society of America, Inc., Madison, WI.

212. Xia, G. 1998. Sorption behavior of nonpolar organic chemicals on natural

sorbents (polyanyi theory). Ph.D. dissertation. Johns Hopkins University.

213. Young, L. Y., and C. E. Cerniglia. 1995. Microbial Transformation and Degradation of Toxic Organic Chemicals. Wiley Liss, Inc., New York, NY.

214. Young, T. M., and W. J. Weber. 1995. A distributed reactivity model for

sorption by soils and sediments. 3. Effects of diagenetic processes on sorption energetics. Environ. Sci. Technol. 29:92-97.

215. Zumft, W. G. 1977. Cell biology and molecular basis of denitrification.

Microbiol. Mol. Biol. Rev. 61:533-616.