The Transport and Fate of Cr(VI) in the Environmentbaholmen/docs/ENVE290W/National... · The assessment of the transport and fate of any contaminant in the envi- ronment generally

161

1-5667-0608-4/01/$0.00+$1.50© 2004 by CRC Press LLC

5

The Transport and Fate of Cr(VI)

in the Environment

Frederick T. Stanin and Malcolm Pirnie

CONTENTS

5.1 The Presence of Chromium in the Environment .................................1635.1.1 Antropogenic Sources...................................................................1645.1.2 Natural Sources .............................................................................165

5.2 Geochemistry of Chromium ....................................................................1655.2.1 Cr(III)...............................................................................................1665.2.2 Cr(VI)...............................................................................................167

5.3 Oxidation-Reduction of Chromium........................................................1685.3.1 Review of Oxidation-Reduction Reactions ...............................1685.3.2 General Redox Behavior of Chromium

in the Environment .......................................................................1695.3.3 Oxidation of Chromium...............................................................170

5.3.3.1 Oxidation of Cr(III) to Cr(VI) by Dissolved Oxygen and Manganese Dioxides ...............................171

5.3.3.2 Oxidation of Cr(III) to Cr(VI) by H

2

O

2

.......................1725.3.4 Reduction of Cr(VI) to Cr(III) .....................................................172

5.3.4.1 General..............................................................................1725.3.4.2 Fe(II) (Dissolved Fe(II) and Fe(II)-Containing

Minerals)...........................................................................1735.3.4.3 Sulfides .............................................................................1755.3.4.4 Organic Matter ................................................................1755.3.4.5 Cu(I) ..................................................................................1765.3.4.6 Hydrogen Peroxide(H

2

O

2

) .............................................1765.4 Precipitation/Dissolution Reactions of Chromium .............................1765.5 Sorption and Desorption Reactions of Chromium ..............................177

5.5.1 General Discussion of Sorption ..................................................1775.5.2 Sorption of Chromium .................................................................178

5.5.2.1 Sorption of Cr(III) ...........................................................1785.5.2.2 Sorption of Cr(VI) ...........................................................179

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Chromium(VI) Handbook

5.6 General Transport and Fate of Chromiumin Environmental Media...........................................................................1805.6.1 Chromium in the Atmosphere....................................................1815.6.2 Chromium in Aquatic Environments ........................................182

5.6.2.1 Surface Waters .................................................................1825.6.2.2 Groundwater ...................................................................1845.6.2.3 Plumes of Chromium in Groundwater—Case

Studies ..............................................................................1865.6.2.3.1 Nassau County, New York..........................1865.6.2.3.2 Telluride, Colorado.......................................1875.6.2.3.3 Southwestern Michigan—Wood

Treatment Plant .............................................1875.6.2.3.4 Industrial Waste Landfill,

Northern France............................................1885.6.3 Chromium in Soil..........................................................................188

5.6.3.1 Overview of Metals in Soil ...........................................1895.6.3.2 Behavior of Chromium in Soil......................................191

5.6.3.2.1 Sorption of Cr(III) and Cr(VI) ....................1935.6.3.2.2 Reduction of Cr(VI) and Oxidation

of Cr(III) .........................................................1935.6.4 The Uptake and Transformation of Chromium by Biota.......195

5.7 Utilizing Natural Environmental Processes as a Remedy for Soil and Groundwater Contaminated with Chromium ...............1955.7.1 There Are Natural Reductants in the Aquifer..........................1975.7.2 The Amount of Cr(VI) and Other Reactive Constituents

Do Not Exceed the Reductive Capacity of the Aquifer .........1985.7.2.1 Mass of Cr(VI) .................................................................1985.7.2.2 Mass of Cr(III) .................................................................1995.7.2.3 Reduction Capacity of the Aquifer ..............................1995.7.2.4 Oxidation Capacity of the Aquifer ..............................200

5.7.3 The Rate of Cr(VI) Reduction to the Target Concentration Compared to the Rate of Transport of Cr(VI)from Source to Point of Compliance..........................................2005.7.3.1 Rates of Oxidation and Reduction...............................2015.7.3.2 Estimating Reduction from Monitoring

Well Data ..........................................................................2015.7.3.3 Monitoring Reduction Via Stable

Isotopes of Chromium ...................................................202Bibliography ......................................................................................... 204

The knowledge of the transport and fate of contaminants in the subsurfaceenvironment is essential to achieving environmental restoration objectives.Unfortunately, gaining and using this knowledge can be difficult because ofproblems and limitations involving complex hydrogeology, hydrochemistry,and microbiology, or by economical restraints. This chapter discusses the


The Transport and Fate of Cr(VI) in the Environment

163

most important physical, chemical, and biological parameters to understand-ing the transport and fate of Cr(VI), and chromium in general, and howthese parameters can be measured and utilized in environmental restorationprojects. These discussions are drawn from several sources, as referenced.

The assessment of the transport and fate of any contaminant in the envi-ronment generally consists of site characterization, risk assessment, andremediation, which are the three main phases of environmental restorationprograms. Conventional approaches to site characterization may not ade-quately define the need to obtain enough detailed information about naturalprocesses affecting the transport behavior and the ultimate fate of contami-nants. The use of state-of-the-art site characterizations, although more costlyto implement than more conventional means, may ultimately result in sig-nificant savings because of improved technical effectiveness and efficiencyof site cleanup. Also, proper site characterization methods can aid risk man-agement decisions (risk assessments) in determining if remediation is evennecessary and/or to what extent, and choosing the proper cleanup technol-ogies if remediation is needed.

A sound conceptual site model (CSM) is absolutely necessary for transportand fate assessments. An adequate CSM incorporates information on geo-logic, hydrologic, chemical, and biological processes to produce an effectivecontaminant transport evaluation. The practical use of risk assessments andremedial technologies is highly dependent on site-specific knowledge ofthese processes. Therefore, the processes that govern the subsurface behaviorand treatability of contaminants must be understood. This is particularlytrue for Cr(VI).

This chapter discusses the transport and fate of Cr(VI) as well as otherforms of chromium in the environment, most notably Cr(III). These formsor species of chromium are inexorably linked by many environmental trans-port and fate processes. First presented is an overview of the presence ofchromium in the environment and its general geochemistry. The most impor-tant transport and fate processes are discussed, namely oxidation-reduction,precipitation-dissolution, and sorption-desorption. Then, the general aspectsof the transport and fate of chromium in the atmosphere, aquatic environ-ments (surface waters and groundwater), and soil, including the uptake andtransformation of Cr(VI) by biota, are summarized. The chapter concludeswith a discussion of how natural attenuation processes can be implementedas a remedy for sites contaminated by Cr(VI).

5.1 The Presence of Chromium in the Environment

Chromium contamination of soil and groundwater is a significant problemworldwide. The extent of this problem is due primarily to its use in numerousindustrial processes (i.e., metal plating and alloying, leather tanning, wood


164


treatment, etc.), but also its natural presence in rocks enriched in chromium(i.e., ultramafic rocks such as serpentinite). Compared to the results of con-tamination of soil and groundwater by industrial practices, the naturally-occurring concentrations of chromium in soil and groundwater are relativelylow. However, relatively high concentrations of naturally-occurring dis-solved chromium have been observed, usually associated with the verysoluble chromate species (Robertson, 1975). Thus, both anthropogenic (man-made) and natural sources of chromium can lead to locally elevated levelsin soils and waters.

The presence of chromium in the environment is discussed by severalauthors (Davis and Olsen, 1995; Kimbrough et al., 1999; Richard and Bourg,1991; Kotas and Stasicka, 2000). Their presentations are briefly summarizedbelow. Chapter 4 gives additional information on naturally-occurring Cr(VI)in groundwater.

5.1.1 Antropogenic Sources

Chromium is used in several industries, including metallurgy (steel, ferro-and nonferrous alloys), refractory (chrome and chrome-magnesite), andchemical manufacturing (pigments, electroplating, tanning and other),involving numerous commercial processes including electroplating, leathertanning, pulp production, milling, mining (ore refining), and wood preser-vation. The industrial use of chromium generally begins with the mining ofchromite (a naturally-occurring ore), usually as ferrous chromite (FeO

⋅

Cr

2

O

3

)(Hartford, 1983). Then, the ore is either oxidized or reduced during industrialprocessing. Sodium chromate (Na

2

CrO

4

) has usually been produced by theoxidation of chromite. Sodium carbonate, calcium oxide, and calcium chro-mate (CaCrO

4

) are produced as byproducts. In turn, several substances arederived from Na

2

CrO

4

, including dichromates (Na

2

Cr

2

O

7

or K

2

Cr

2

O

7

), Cr(VI)oxide (CrO

3

), chromic acid (H

2

Cr

2

O

7

), and other oxides of Cr (e.g., K

2

CrO

4

),including chromium pigments (barium, calcium, lead, strontium, and zincchromate). Also, chromite ore can be reduced by a variety of methods usingaluminum, silicon, or carbon as reducing agents (Hathway, 1989). Materialsfrom this reduction are used for producing chromium alloys and chromealum (NH

4

Cr(SO

4

)

2

⋅

12H

2

O).Most of the chromium consumed by industry in the U.S. is for the pro-

duction of metal alloys, mainly wrought-stainless and heat-resisting steels(Hartford, 1983). Chromium as part of an iron alloy is insoluble with a zerooxidation state and therefore is not a form of chromium having an environ-mental concern. However, Cr can be oxidized and leached from stainlesssteel into a water-soluble form (Kimbrough et al., 1999).

Chromium from anthropogenic sources can be released to soils and sedi-ments indirectly by atmospheric deposition, but releases are more commonlyfrom dumping of Cr-bearing liquid or solid wastes such as chromate byprod-ucts (“muds”), ferrochromium slag, or chromium plating wastes. Such wastescan contain any combination of Cr(III) or Cr(VI) with various solubilities.



165

The nature and behavior of various forms of chromium found in wastewaterscan be quite variable. The presence, form and concentration of Cr in dis-charged effluents depends mainly on the chromium compounds utilized inthe industrial process, on the pH, and on the presence of other organic andinorganic processing wastes.

Chemicals containing Cr(VI) are principally used for metal plating (whichuse H

2

Cr

2

O

7

), as dyes, paint pigments, and leather tanning (Hartford, 1983).Chromium platers involve the use of H

2

Cr

2

O

7

to plate chromium onto piecesof other metals. Thus, Cr(VI) will dominate in wastewater from the metal-lurgical industry, metal finishing industry (Cr hard plating), refractory indus-try and production or application of pigments (chromate color pigments andcorrosion inhibition pigments). Cr(III) will be found mainly in wastewatersof the tannery, textile (printing, dying) and decorative plating industries.However, there are exceptions to these generalities due to several factors.For example, in tannery wastewater where Cr(III) is the most expected form,the redox reactions occurring in sludge can increase the concentration ofCr(VI). Such transformations are also common in the subsurface, such asoxidation, reduction, sorption, precipitation, and dissolution, which are alldiscussed later in this chapter.

5.1.2 Natural Sources

As previously mentioned, chromium occurs naturally in the environment,most notably in its most concentrated forms as an ore mineral. Chromiumalso occurs naturally as a component of soils (Schacklette and Boerngen,1984), usually as chromite (FeCr

2

O

4

), a relatively insoluble soil mineral(Schmidt, 1984). The main source of such chromium in natural soils is theweathering of their parent materials. The average amount of this element invarious kinds of soils ranges from 0.02 to 58 micromoles per gram (Richardand Bourg, 1991; Coleman, 1988). The chromium concentration will be influ-enced by the composition of the parent rock. Granite, carbonates and sandysediments present the lowest chromium content whereas shales, river sus-pended matter, and soils typically exhibit highest levels. Highest chromiumcontents tend to be associated with finest grain size soils (Robertson, 1975)and sediments (Salomons and DeGroot, 1978). Thus, the natural concentra-tion of chromium in the environment varies greatly (Cary, 1982).

5.2 Geochemistry of Chromium

Chromium can exist in several chemical forms with oxidation numbers rang-ing all the way from –2 to

+

6. However, in the environment, chromiumcommonly exists in only two stable oxidation states, Cr(VI) and Cr(III), whichhave greatly contrasting toxicity and transport characteristics. Chromiumspeciation in the environment, particularly in groundwater, is affected


166


primarily by Eh (oxidizing or reducing conditions) and pH (acidic or alkalineconditions). In general, Cr(VI) predominates under oxidizing conditions, andCr(III) predominates under more reducing conditions.

These two different forms of chromium are quite different in their prop-erties: charge, physiochemical characteristics, mobility in the environment,chemical and biochemical behavior, bioavailability, and toxicity. Most nota-bly, Cr(III) is considered to be a trace element essential for the proper func-tioning of living organisms, whereas Cr(VI) exerts toxic effects on biologicalsystems. Also, Cr(VI) compounds are generally more soluble, mobile, andbioavailable in the environment compared with Cr(III) compounds. And, aspreviously mentioned, the more toxic and mobile Cr(VI) predominates inoxidizing environments, while the less toxic and immobile Cr(III) is restrictedto reducing environments. Therefore, it is quite important to distinguishthese forms of chromium rather than discussing this element as “total chro-mium.” The geochemistry of these forms are briefly discussed below. A morecomprehensive discussion of the geochemistry of chromium and chromiumcompounds is presented in Chapter 2.

5.2.1 Cr(III)

In aqueous systems, Cr(III) can be present as Cr

3

+

, Cr(OH)

2

+

, Cr(OH)

2

+

, andCr(OH)

4–

. Additionally, the precipitated phase Cr(OH)

3

predominatesbetween pH 6 and pH 12 (Rai et al., 1987). Under slightly acidic to alkalineconditions, and if Fe(III) is present, Cr(III) can precipitate as an amorphousmixed hydroxide Cr

x

Fe

1–x

(OH)

3

(Eary and Rai, 1988). Amorphous Cr(OH)

3

can crystallize as Cr(OH)

3

⋅

3H

2

O or Cr

2

O

3

(eskolaite) under different condi-tions (Palmer and Puls, 1994). With high redox potential, Cr(VI) predomi-nates with a much higher solubility (Loyaux-Lawniczak et al., 2001). In areducing environment, and in the absence of Fe, Cr(III) precipitates readilyto form Cr(OH)

3

(Rai et al., 1987).In relatively low Eh environments, the main aqueous Cr(III) species are

Cr

3

+

, Cr(OH)

2

+

, Cr(OH)

30

and Cr(OH)

4–

(Rai et al., 1986, 1987). The Cr

3

+

speciesis prevalent only at pH lower than about 4. With increasing pH, hydrolysisof Cr

3

+

yields Cr(OH)

2

+

(generally present in groundwater at a pH of 6 to 8but also in some acidic waters), and Cr(OH)

30

and Cr(OH)

4–

(generally inalkaline groundwater) (Rai et al., 1987). Polymeric species such as Cr

2

(OH)

24

+

,Cr

3

(OH)

45

+

and Cr

4

(OH)

66

+

are never significant in natural systems (Rai et al.,1986, 1987). Cr(III) readily forms complexes with a variety of ligands:hydroxyl, sulfate, ammonium (NH

4

), cyanide and sulphocyanide, fluorideand chloride (to a lesser extent), and natural and synthetic organic ligands(Richard and Bourg, 1991). Only one Cr(III) compound, Cr

2

O

3

, is an oxide,so the role of oxygen is central to the oxidation/reduction (redox) processfor chromium (Kimbrough et al., 1999).

Solubility can significantly limit the concentration of Cr(III) in groundwa-ter at a pH above 4. The low solubility of the Cr(III) solid phases, Cr

2

O

3

and



167

Cr(OH)

3

(Hem, 1977), is likely the major reason why Cr(III) generally makesup a small percentage of the total chromium concentration in natural orcontaminated groundwaters. Cr(III) tends to be essentially immobile in mostgroundwaters because of its low solubility (Calder, 1988).

5.2.2 Cr(VI)

Chromium(VI) exists in the environment as part of several compounds, Cr(VI)is present in solution as monomeric forms: H

2

CrO

40

, HCrO

4–

(bichromate),CrO

42–

(chromate), and CrO

3

(chromium(III) oxide), or as the dimeric ionCr

2

O

72–

(dichromate) (Palmer and Puls, 1994). Under oxidizing conditions,aqueous chromium is present in a Cr(VI) anionic form, HCrO

4–

or CrO

42–

,depending on the pH (CrO

42–

at a higher pH) (Richard and Bourg, 1991).Within the normal pH range in natural waters (i.e., 6 to 8), the CrO

42–

, HCrO

4–

and Cr

2

O

72–

ions are the forms expected. At relatively high Cr(VI) concentra-tions, the Cr

2

O

72–

ion predominates in acidic environments (Richard andBourg, 1991).

It should be noted here that the term Cr(VI) is somewhat of a misnomer.This is because Cr(VI) is not present in the environment as a free cation(whereas Cr(III) does exist in the environment as previously mentioned). Infact, as all Cr(VI) species are oxides, they act like a –2 anion (ion

2–

) ratherthan a Cr(VI) cation (Kimbrough et al., 1999).

The relative concentration of the various Cr(VI) species depends on thepH and the total Cr(VI) concentration (Palmer and Puls, 1994). For example,significant concentrations of H

2

CrO

40

only occur under the extreme conditionof pH around 1. Above a pH of about 6, CrO

42–

generally dominates (Davisand Olsen, 1995). Below pH of about 6, HCrO

4–

dominates when the Cr(VI)concentrations are relatively low, but Cr

2

O

72–

becomes more significant asCr(VI) concentrations increase, or it may even dominate when the totalCr(VI) concentrations are relatively high (Palmer and Puls, 1994). Thesespecies constitute many of Cr(VI) compounds which are quite soluble andthus mobile in the environment (Kotas and Stasicka, 2000). These speciesdiffer in their solubility and in their tendency to be sorbed by soil or aquifermaterials (Calder, 1988).

There are no significant solubility constraints on the concentrations ofCr(VI) in groundwater. The chromate (CrO

42–

) and dichromate ions(Cr

2

O

72–

) are water soluble at all pH. However, chromate can exist as aninsoluble salt of a variety of divalent cations, such as Ba

2

+

, Sr

2

+

, Pb

2

+

, Zn

2

+

,and Cu

2

+

, and these salts have a wide range of solubilities. The rates ofprecipitation/dissolution reactions between chromate, dichromateanions, and these cations vary greatly and are pH dependent. An under-standing of the dissolution reactions is particularly important for assess-ing the environmental effects of chromium because Cr(VI) often entersthe environment by dissolution of chromate salts (e.g., SrCrO

4

) (Rai et al.,1987).


168


5.3 Oxidation-Reduction of Chromium

Chromium in the environment is altered by oxidation-reduction reactions,changing its physical and chemical properties. To understand these changes,it is worthwhile to review the basics of oxidation and reduction.

5.3.1 Review of Oxidation-Reduction Reactions

Oxidation-reduction reactions involve the transfer of electrons. In oxidation-reduction reactions, some components (charged/uncharged atoms) loseelectrons and some components gain electrons from this transfer. The pro-cess of removing electrons from a component (loss of electrons) is calledoxidation and results in a more positive oxidation number. After oxidationhas occurred, the component is said to have been oxidized. The process ofadding electrons to an atom (gain of electrons) is called reduction, and resultsin a more negative oxidation number. After reduction has occurred, the com-ponent is said to have been reduced. The component which gains electronsin an oxidation-reduction reaction is called the oxidizing agent, whereasthe component which loses electrons is the reducing agent. Due to theconservation of mass, and also the conservation of electrons and oxidationnumbers, electrons lost by the oxidized component are gained by the oxi-dizing agent. Therefore,

oxidation is always accompanied by reduction, and theoxidation and reduction always takes place to an equal degree.

For this reason,reference is usually made to combined oxidation-reduction reactions (redoxreactions) rather than separate oxidation or reduction reactions. It is some-times useful to consider the oxidation reaction and reduction reaction sep-arately as “1/2 reactions.”

The potential for an electron transfer is best measured by the redox poten-tial (Eh), which is sometimes expressed as the redox intensity factor (pe),the negative log of the electron activity (a

e

),

pe

=

−

log(a

e

) (5.1)

Also,

Eh

=

(pe)(2.3

RT

/

F

)

=

0.059 pe

where

R

=

gas constant

T

=

temperature

F

= Faraday constant


The Transport and Fate of Cr(VI) in the Environment 169

essentially the ratio of electron donors (oxidizing agents) to electron acceptors(reducing agents) (Kimbrough et al., 1999).

5.3.2 General Redox Behavior of Chromium in the Environment

The oxidation and reduction of chromium in the environment is discussed byseveral authors, including Kimbrough et al. (1999), Richard and Bourg (1991),and Calder (1988). Their presentations are summarized below, followed byseparate, more detailed discussions of chromium oxidation and reduction.

The distribution between Cr(III) and Cr(VI) in the environment, includingaquatic environments such as groundwater, will be regulated by redoxreactions and redox conditions (Richard and Bourg, 1991). To understandthe distribution of Cr(III) and Cr(VI), Eh-pH diagrams for chromium inaqueous environments are usually employed, such as the one discussedearlier in Chapter 3.

The redox transformation of Cr(III) to Cr(VI), the oxidation of Cr(III), or thereduction of Cr(VI) to Cr(III), requires another redox couple (of oxidizing/reducing agent) which accepts or gives the necessary electrons. In naturalaquatic environments, the significant redox couples (reducing agents/oxida-tion agents) are (Richard and Bourg, 1991):

• H2O/O2 (aq)• Mn(II)/Mn(IV)• NO2/NO3

• Fe(II)/Fe(III)• S2–/SO4

2–

• CH4/CO2

In the case where O2 is the oxidizing agent, oxidation of chromium requiresdonation of electrons to oxygen, while reduction of oxygen requires the accept-ing of electrons from chromium. For a given chromium compound, the redoxreactions involving other chemical agents (i.e., oxidizing and reducing agents)are governed by the agents’ capacity for donating or accepting electrons. Sev-eral oxidation and reduction reactions of chromium with common environ-mental agents are given in several sources, including Kimbrough et al. (1999).

Oxidation Example:

Cr(III) → Cr(VI)

2Cr2O3 + 3O2 → 4CrO3

The concentration of these oxidizing and reducing agents affects the oxi-dation-reduction of chromium. Many oxidizing agents are known to oxidizeCr(III) to Cr(VI), but only a few of them are found in the environment (i.e.,groundwater) in sufficient concentration to do so. On the other hand, the


170 Chromium(VI) Handbook

many reducing agents for Cr(VI) reduction to Cr(III) are typically found atsufficient levels. For example, ozone (O3) can theoretically oxidize Cr(III) tochromate (a reaction with Eh = 0.87 volts), but the concentration of ozone inthe environment is usually insufficient (it is relatively unstable) to accom-plish this oxidation (Grohse et al., 1988). However, the reduction of Cr(VI)by Fe(II), even though less favored thermodynamically (a reaction with Ehonly 0.56 volts) is feasible because iron concentrations are generally sufficientin the environment (Rai et al., 1989).

At lower pHs, chromates exist as chromic acid (H2CrO4) and hydrogenchromate (HCrO4

–). When the concentration of CrO42– is high, chromates are

transformed to dichromate (H2Cr2O7 or HCr2O7–), which are strong oxidizing

agents and are thus rapidly reduced in the presence of reducing agents atlow pH or high Eh. At high pH, chromates exist in the form of CrO4

2–, whichis a poor oxidizing agent and hence, with lower Eh values, are more stable.

Cr(VI) can be transported great distances in groundwater due in part toits high solubility, whence it may be transformed by reduction to, and pre-cipitated as, Cr(III) if the transported Cr(VI) enters an area with relativelylow Eh. Cr(VI) can be reduced readily to Cr(III) in the presence of organicmatter, especially where pH is low (Bartlett and Kimble, 1976; Bloomfieldand Pruden, 1980). Cr(VI) can also be reduced by Fe(II) and dissolved sul-fides (Schroeder and Lee, 1975).

Cr(III) generally is not transported great distances by groundwater becauseof its low solubility. However, Cr(III) can be transformed to the more solubleCr(VI) if the redox conditions along the transport pathway change fromreducing to oxidizing. Under natural conditions, Cr(III) has been found tobe oxidized to Cr(VI) by manganese (Schroeder and Lee, 1975; Bartlett andJames, 1979). In the laboratory, Cr(III) can exist as highly soluble organiccomplexes, particularly under low pH conditions (Bartlett and Kimble, 1976;James and Bartlett, 1983). Therefore, if Cr(III) is in a complexed form, it couldbe present at much higher concentrations in groundwater than if it is uncom-plexed. However, the existence of Cr(III) complexes has not been docu-mented under field conditions.

5.3.3 Oxidation of Chromium

There are several sources of oxygen for the oxidation of chromium. In theenvironment, water is the most important source; manganese dioxide, ozone,hydrogen peroxide (H2O2), manganese dioxide, and lead dioxide are othernotable sources of oxygen (Kimbrough et al., 1999). Oxidation of chromiuminvolving these sources requires the presence of water, thus water chemistryis important to the understanding of chromium oxidation. Generally, highvalues of Eh in water correspond to strongly oxidizing conditions. A sum-mary of Cr(III) oxidation via dissolved oxygen, manganese dioxides, andhydrogen peroxides are discussed below, based on papers by Kotas andStasicka (2000); Loyaux-Lawniczak et al. (2001); Palmer and Puls (1994);Davis and Olsen (1995); and Richard and Bourg (1991).



5.3.3.1 Oxidation of Cr(III) to Cr(VI) by Dissolved Oxygenand Manganese Dioxides

The redox potential of the Cr(VI)/Cr(III) couple is high enough so that onlya few oxidants are present in natural systems capable of oxidizing Cr(III) toCr(VI). Only dissolved oxygen and manganese dioxides (MnO2) are knownto oxidize Cr(III) to Cr(VI) (Eary and Rai, 1987) and manganese dioxides arethe more common oxidant; oxidation of Cr(III) by dissolved oxygen withoutany mediate species has been reported to be negligible (Schroeder and Lee,1975; Eary and Rai, 1987), whereas mediation by manganese oxides has beenfound to be the effective oxidation pathway in environmental systems(Schroeder and Lee, 1975; Bartlett and James, 1979; Nakayama et al., 1981;Saleh et al., 1989; Johnson and Xyla, 1991).

Dissolved oxygen can oxidize Cr(III) into Cr(VI), (Rai et al., 1986; Schroederand Lee, 1975; Eary and Rai, 1987; Nakayama et al., 1981) but laboratorystudies indicate that this can be relatively slow requiring several months(Palmer and Wittbrodt, 1990), especially in slightly acidic to basic environ-ments (Eary and Rai, 1987). Such slow kinetics enable Cr(III) to be involvedin other reactions (sorption or precipitation) that are much faster. Therefore,the oxidation of Cr(III) in the environment by dissolved oxygen is unlikely.

Manganese oxides are likely to be responsible for most Cr(III) oxidationin aquatic environments. Fendorf and Zasoski (1992) suggest that CrOH2+ isthe reactive species in this Cr(III) oxidation. Bartlett and James (1979)observed a correlation between the amount of Cr(III) oxidized by soils andthe amount of reduced manganese in soils, thereby suggesting the oxidationof Cr(III) is the result of interaction with manganese dioxides, which hasbeen verified by laboratory studies. Experimental results indicate that theoxidation follows the reaction (Palmer and Puls, 1994):

CrOH2+ + 1.5δ − MnO2 → HCrO4

− + 1.5Mn2+(5.2)

2 CrOH2+ + 3δ − MnO2 → 2HCrO4

− + 3Mn2+

Manganese oxides are present in the subsurface as grain coatings, depositsin cracks or fractures, or as finely disseminated grains; sometimes this pres-ence is a result of bacterial activities. The mechanisms for the reaction withMnO2 occurring at the manganese oxide surfaces (by adsorption of Cr(III)on active surface sites) are very complex and not yet fully understood (Earyand Rai, 1987; Fendorf and Zasoski, 1992). The Cr(III) oxidation rate is likelyrelated to the amount and surface area of manganese oxides (Schroeder andLee, 1975; Eary and Rai, 1987), and lab studies indicate this rate to be initiallyrapid, but then slowing down significantly. Also, there is an increase in therate and amount of Cr(III) oxidation as pH decreases, and the surface areato solution volume increases.

Richard and Bourg (1991) explain that the oxidation of Cr(III) by manga-nese dioxides is likely to occur as a result of three sequential steps (Rai et al.,1986; Schroeder and Lee, 1975; Bartlett and James, 1979; Eary and Rai, 1987;



Amacher and Baker, 1982). First, the Cr(III) would be sorbed onto MnO2

surface sites. Then, the Cr(III) would be oxidized to Cr(VI) by Mn(IV) on thesurface sites, however, all the Mn(IV) reaction sites are probably not acces-sible to Cr(III) (Rai et al., 1986; Amacher and Baker, 1982). Finally, the reactionproducts, Cr(VI) and Mn(II), would be desorbed. Richard and Bourg (1991)give theoretical stoichiometries that have been suggested for this oxidation:

2Cr3+ + 3δ ⋅MnO2 (s) + 2H2O = 2HCrO4− + 3Mn2+

+ 2 H+ (Amacher and Baker, 1982) (5.3)

and

Cr(OH)2+ + 3δ ⋅MnO2 (s) + 3H2O = HCrO4− + 3MnOOH (s)

+ 3H+ (Eary and Rai, 1987) (5.4)

The solid MnOOH(s) would decay into aqueous Mn2+ afterwards.

5.3.3.2 Oxidation of Cr(III) to Cr(VI) by H2O2

In an attempt to mobilize Cr(III) by oxidizing it to Cr(VI), H2O2 was appliedto groundwater in laboratory studies by Davis and Olsen (1995). Also, Pettineet al. (2002) have studied the oxidation of Cr(III) with H2O2 in basic solutions(Pettine and Millero, 1990 and 1991). They found that H2O2 controls the rateof oxidation of Cr(III) in surface waters.

5.3.4 Reduction of Cr(VI) to Cr(III)

5.3.4.1 General

Cr(VI) is a strong oxidant and therefore can be reduced in the presence ofelectron donors. The most common forms of chromium dissolved in naturalwaters, within the environmentally normal range of pH, are CrO4

2–, HCrO4–

and Cr2O72– ions, (Kotas and Stasicka, 2000) which form many of the Cr(VI)

compounds that can be quite readily reduced to Cr(III) forms in the presenceof electron donors like organic matter and inorganic compounds in theirreduced state, many of which are quite common in soil, water, and theatmosphere (Stollenwerk and Grove, 1985). The major factors in this reduc-tion to Cr(III) are dissolved Fe(II), minerals with Fe(II), sulfides (reducedsulfur), and organic matter (Kotas and Stasicka, 2000; Palmer and Puls, 1994;Wielinga et al., 2001). Studies of reaction kinetics by Fendorf et al. (2001)indicate that Fe(II) and dissolved sulfides will probably dominate the reduc-tion of chromate (Wielinga et al., 2001). Loyaux-Lawniczak et al. (2001)report that photoreduction is another pathway to reduce Cr(VI) in theenvironment (Kieber and Heiz, 1992; Hug et al., 1997), but this mechanismis probably only important in the atmosphere or in upper surface waters.



Much of the information of Cr(VI) reduction is based on laboratory studies,and many are referenced herein. However, it should always be noted thatthe application of these experimental observations to field situationsremains dubious.

Cr(VI) can be reduced by biological and chemical (abiotic) processes. It isdifficult to determine which processes are responsible for the reduction ofmetal contaminants. However, it is probable that the reaction rates will deter-mine the reduction process and its specific pathway. By comparing reductionrates involving Fe(II) and sulfides with those reported for direct microbialreduction, the chemical reduction of chromate by Fe(II) is more than 100 timesfaster than the observed biological reduction rate, thus chemical reduction ofCr(VI) will probably be the main process for chromate reduction when eitherferrous iron or sulfide are present, and these are present in the environmentunder anaerobic conditions (Wielinga et al., 2001). Yet, both aerobic andanaerobic reduction by microbes have been observed, the latter being morecommon (Palmer and Puls, 1994). The specific mechanisms for Cr(VI) reduc-tion by these microbes is not well known, but the chromate may actually beused as an electron acceptor for cell metabolism (Palmer and Puls, 1994).

The major factors in the reduction of Cr(VI) to Cr(III), namely dissolvedFe(II), minerals with Fe(II), sulfides (reduced sulfur), and organic matter, arediscussed below, based on papers by Pettine et al. (2002); Palmer and Puls(1994); Wielinga et al. (2001); and Richard and Bourg (1991). The roles ofcopper and H2O2 are also discussed.

5.3.4.2 Fe(II) (Dissolved Fe(II) and Fe(II)-Containing Minerals)

Fe(II) is a major factor in the reduction of Cr(VI) to Cr(III)—experimentalresults of Davis and Olsen, 1995 from column tests were consistent withother published observations (Schroeder and Lee, 1975) that found Cr(VI)to be reduced to Cr(III) by Fe(II). Dissolved Fe(II) ions in environmentalwaters can be generated by the discharges of some industrial wastes, butalso can result from the weathering of Fe(II)-containing minerals. There arenumerous minerals in geologic materials that contain Fe(II) for Cr(VI)reduction. These minerals include silicates, oxides, or sulfides (Palmer andPuls, 1994):

• Silicates: Olivine, pyroxenes (augite and hedenbergite), amphiboles(hornblende, cummingtonite, and grunerite), micas (biotite, phlo-gopite, and glauconite), chlorites, and clays (the smectite nontron-ite).

• Oxides: Magnetite, ilmenite, and hematite.• Sulfides: Pyrite, in which both the iron(II) and the sulfide are active

in reducing Cr(VI).

Cr(VI) reduction via Fe(II) in silicate minerals (e.g., biotite in solutionrather than at the mineral surface) was reported by Eary and Rai. (1989).



This is described by Palmer and Puls (1994) as a rather complex process. Thepresence of Fe3+ increases the reduction rate, and the Fe3+ is reduced at themineral surface. The iron in the crystal structure is oxidized, K+ is releasedto solution, and Cr(VI) in solution is then reduced by the Fe2+. The Fe3+

resulting from this reduction reaction is then sorbed to the surface of thebiotite where it is again reduced to Fe2+.

Palmer and Puls (1994) state that Cr(VI) reduction in the presence of ironoxides has been observed in several experiments (Eary and Rai, 1989; Whiteand Hochella, 1989). In the case of hematite, the reduction is suggested tooccur in solution after the FeO component has dissolved. The reduction ofCr(VI) via Fe(II) in pyrite is described in work by Lancy (1966), who sug-gested that pyrite could be used for treating spent cooling waters that containCr(VI) as a corrosion inhibitor, because reduction of Cr(VI) occurs at thepyrite surface rather than in solution. This reduction was described as occur-ring even in slightly alkaline solutions; however, the pyrite had to be con-tinuously abraded to remove surface coatings. Batch testing reported byBlowes and Ptacek (1992) involving pyrite both in the presence and in theabsence of calcite showed faster removal of Cr(VI) with no calcite. Also,Loyaux-Lawniczak, et al. (2001) demonstrated that Fe(II)-Fe(III) hydroxysaltgreen rusts can reduce Cr(VI); ferrihydrite is the Cr(III)-bearing solid phasethat is formed from this reduction.

Under neutral to alkaline pH conditions, Fe(II) controls the reduction ofCr(VI) in natural anaerobic systems (Pettine et al., 1998), while at acidic pHlevels, other reductants may be more efficient than Fe(II). The involvementof Fe(II) in Cr(VI) reduction, where the Fe(II) (as FeO) comes from hematiteor biotite, can be expressed as follows (Richard and Bourg, 1991):

[3FeO] + 6H+ + Cr(VI) (aq) = Cr(III) (aq) + 3Fe(III) (aq) + 3H2O (5.5)

This can be a relatively rapid reaction from the standpoint of environmen-tal situations; laboratory studies report this reaction being complete in lessthan 5 min (Eary and Rai, 1988). In acidic waters, the end products of thisreaction are Fe(III) and Cr(III) (Stollenwerk and Grove, 1985), whereas underneutral to alkaline conditions, Cr(OH)3 is probably the end product becauseof the very low solubility of Fe(OH)3 (Rai et al., 1988). In groundwaters ofpH more than 4, Cr(III) precipitates with the Fe(III) in a solid solution withthe general composition CrxFe1–x(OH)3 (Eary and Rai, 1988; Rai et al., 1988;Sass and Rai, 1987).

Wielinga et al. (2001) reported that the Cr(VI) reduction via Fe(II) (orsulfide) depends on microbial activity. A diverse and widely distributedgroup of bacteria are able to couple the oxidation of organic compoundsor H2 to the reduction of iron (hydr)oxides (Lovely, 1993; Coates et al.,1996). Thus, in many environments where iron reduction is the predominantterminal electric accepting process (TEAP) in microbial respiration, theindirect reduction of Cr(VI) (chromate) via reaction with a respiratory



byproduct is likely a dominant reductive pathway as shown below (Wielingaet al., 2001):

3C3H5O3− + 12Fe(OH)3 → 3C2H3O2

− + 12Fe2+ + 3HCO3

− + 8H2O + 21OH− (5.6)

3Fe2+ + HCrO4− + 8H2O → 3Fe(OH)3 + Cr(OH)3 + 5H+ (5.7)

This reduction of Cr(VI) (chromate) is via a coupled, two-step, biotic-abiotic reaction pathway in which Fe(II) produced during iron respirationcatalyzes the reduction of Cr(VI). Thus, attenuation of chromate in saturated,subsurface environments may be in large part attributable to iron reduction.In addition, the capacity for soils to reduce and immobilize Cr(VI) could bedramatically underestimated if this biotic-abiotic process is not considered.Wielinga et al. (2001) emphasize that the implications of these reactions isimportant—the primary terminal electron acceptor is continually regenerated.Fe(II) produced in the first reaction listed (Equation 5.6) is cycled back toFe(III) in the second listed reaction (5.7) thereby acting as an electron shuttle(a catalytic role) between the bacteria and chromium. Thus, a significantamount of Cr(VI) could potentially be reduced even with little available Fe.With the rapid cycling of Fe(II) back to Fe(III), evidence such as high porewater Fe(II) concentrations in pore water could be hidden.

5.3.4.3 Sulfides

Sulfides (reduced sulfur) can be a major factor in the reduction of Cr(VI) toCr(III). Although most sulfides are not soluble, some dissolved sulfides canbe present in the environment due to processes including the discharge ofindustrial wastes, the decomposition of organic matter, and sulfate reduction(a common process in the biodegradation of chlorinated solvent chemicals).Laboratory studies have reported that the reduction of Cr(VI) involvingsulfides is initially rapid, slows down in a few minutes, but reaches comple-tion after one day (Schroeder and Lee, 1975). The rates of reduction of Cr(VI)with H2S have been studied by Pettine et al. (1994) and Pettine et al. (1998).

5.3.4.4 Organic Matter

Organic matter is an important reductant in soils. Much of the organic matter(i.e., organic carbon) in soil is present as humic and fulvic acids. Organicmatter, important in the reduction of Cr(VI), is also present as simple amino-acids (Schroeder and Lee, 1975). The reduction of Cr(VI) by soil humic andfulvic acids has been demonstrated by several researchers as referenced inPalmer and Puls (1994). This reduction, with an intermediate Cr(V) species,is favored by acidic conditions (Bloomfield and Pruden, 1980; Stollenwerkand Grove, 1985; Cary et al., 1977; Grove and Ellis, 1980).



The rate of reduction of Cr(VI) decreases with increasing pH, increaseswith the increasing initial Cr(VI) concentration, and increases as the concen-tration of soil humic substance increases (Palmer and Puls, 1994). At a verylow pH, laboratory studies indicate that the half-life for Cr(VI) reductionwith humic acids is approximately three days, whereas several days arerequired within the pH range of 4 to 7 (Eckert et al., 1990).

5.3.4.5 Cu(I)

Reduced copper [Cu(I)] may also play a role in the reduction of Cr(VI) toCr(III), discussed by Pettine et al. (2002) especially in atmospheric and sur-face waters with low pH and low ionic strength (Abu-Saba et al., 2000). Thereduction of Cr(VI) with Cu(I) has been produced in the laboratory byradiolysis experiments in dilute solutions in the presence of Cr(II).

5.3.4.6 Hydrogen Peroxide (H2O2)

The role of H2O2 in the reduction of Cr(VI) is discussed by Pettine et al.(2002) who describe H2O2 as an oxidant of Cr(III) at pH > 7.5, a reductant atlower pH, and its strength as a reductant being greatly increased at low pH.In acid wastes receiving freshwater, and in atmospheric aqueous media withpH ranging from about 1 to 5, the reduction of Cr(VI) with H2O2 is thermo-dynamically possible (Seigneur and Constantinou, 1995), and has been usedin treatment processes for removing Cr(VI) from wastewaters (Eary and Rai,1988). In the latter, the reduction of Cr(VI) with H2O2 includes a preliminaryconversion of Cr(VI) to Cr(III) and its subsequent precipitation.

5.4 Precipitation/Dissolution Reactions of Chromium

In addition to oxidation-reduction (redox) mechanisms for Cr(VI) reductionas discussed above, chromium can undergo precipitation dissolution reac-tions (Bodek et al., 1988), which are governed by the solubility of the chromiumcompound and the kinetics of the dissolution. Kimbrough et al. (1999) listand discuss these reactions.

Most Cr(III) species that are water-soluble do not occur naturally and areunstable in the environment. The principle Cr(III) reaction in water is theformation of chromium hydroxides of varying solubilities. The precipitationof Cr(III) as the mixed iron-chromium hydroxide (Cr,Fe)(OH)3, discussedpreviously, enhances the precipitation of Cr(III) in waters with neutral pHlevels. The kinetics of this reaction is rather rapid, making it an importantsolubility controlling compound (Sass and Rai, 1987).

The Cr(VI) ions chromate (CrO42–) and dichromate (Cr2O7

2–) are watersoluble at all pHs. However, chromate can exist as the insoluble salt of avariety of divalent cations, such as Ba2+, Sr2+, Pb2+, Zn2+, and Cu2+. The rates



of precipitation/dissolution reactions between chromate, dichromate anions,and these cations vary greatly and are pH dependent. An understanding ofthe dissolution reactions is particularly important for environmental assess-ments because Cr(VI) often enters the environment by dissolution of suchchromate salts. Dissolution of somewhat soluble chromate salts (e.g., SrCrO4)is particularly important because they provide a continual source ofchromate anions.

5.5 Sorption and Desorption Reactions of Chromium

5.5.1 General Discussion of Sorption

McLean and Bledsoe (1992) and Calder (1988) give general discussions ofsorption, which are summarized here. It is first important to define anddistinguish some important terms, which are used interchangeably, properlyand improperly, in the literature. The general term sorption actually com-prises two processes: (1) adsorption, the process by which a solute clings toa solid surface; and (2) absorption, the process by which the solute diffusesinto a porous solid and clings to interior surfaces.

Sorption is important in the transport and fate of a constituent. An equi-librium distribution coefficient (Kd) is used in the estimation of the retarda-tion of a constituent’s migration in groundwater. As with redox andprecipitation reactions, sorption reactions are highly influenced by the com-plex environmental conditions inherent in the subsurface. Therefore, generalassumptions about sorption cannot be made. Such variables as pH, surfacearea, density of active sites, among others, influence sorption equilibrium(Kimbrough et al., 1999). Sorption studies also can are used to evaluate theeffect that changing a soil solution parameter, (e.g., adjustment of pH, ionicstrength, addition of competing cations, or addition of inorganic or organicligands) has on the retention of a constituent by the aquifer matrix.

Laboratory studies generate sorption isotherms, which describe equilib-rium conditions of sorption. These isotherms are the relationship betweenthe amount of constituent sorbed and the equilibrium concentration of theconstituent. If the isotherm is linear, a single coefficient (Kd) can be definedto describe sorption. For metals, such as chromium, the relationship is sel-dom linear. Soil processes are never at equilibrium; soil systems are dynamicand are thus constantly changing. Therefore, other equations with two ormore coefficients must be used. Nonlinear sorption behavior of metals insoil are usually expressed by the Langmuir and the Freundlich equations,even though sorption of metals by soils violates many of the assumptionsassociated with these equations.

For nonlinear sorption, groundwater transportation equations must besolved iteratively using a concentration-dependent retardation factorbecause the retardation of the contaminant will vary with time due to



changing solution concentrations as the plume of contaminated groundwaterpasses a particular portion of the aquifer segment. This makes comparisonsand predictions more difficult than for the linear adsorption model. Equilib-rium studies predict whether a reaction will occur but give no indication ofthe time necessary for the reaction to take place. Therefore, kinetic studieshave been performed to establish the proper time interval for use in equilib-rium sorption/desorption studies. Many mathematical transport modelsnow allow a kinetic term for sorption.

Adsorption occurs because dissolved ionic species are attracted to mineralsurfaces that have a net electrical charge due to imperfections or substitutionsin the crystal lattice or chemical dissociation reactions at the particle surface.This electrical charge varies with pH.

The importance of sorption to the transport and fate of constituents ingroundwater is that it retards the advance of the contaminant with respectto the groundwater velocity, and can also reduce the contaminant concen-tration. However, sorption is reversible, meaning that sorbed contaminantscan be released back into the aqueous medium, causing an increase in con-centrations after periods of decreasing concentrations. Davis and Olsen(1995) showed that in laboratory experiments with columns containing pre-dominantly Cr(III) that were not augmented by additives, less than 2.5% ofthe total Cr was leached, while in soil bearing primarily Cr(VI), over 80% oftotal Cr was leached; Cr(VI) readily dissolved or desorbed from contami-nated soils, while Cr(III) occurred in a predominantly nonleachable form.

5.5.2 Sorption of C4hromium

Several papers, notably Calder (1988); Richard and Bourg (1991); andDavis and Olsen (1995) give detailed discussion of the sorption of thedifferent forms of chromium in the subsurface. Their work is discussedbelow. Also, Calder (1988) gives examples of Cr(III) and Cr(VI) partition-ing ratios (e.g., Kd) that can be used to describe the sorption of chromiumand describes the effect of different values pH and chromium concentra-tions on sorption.

5.5.2.1 Sorption of Cr(III)

Cr(III) is rapidly, strongly and specifically sorbed in soil by Fe and Mn oxides,clay minerals, and sand (Bartlett and Kimble, 1976; Schroeder and Lee, 1975;Korte et al., 1976; Griffin et al., 1977; Rai et al., 1984; Dreiss, 1986). Accordingto experimental data, this sorption is rapid, with about 90% of chromiumbeing sorbed by clay minerals and iron oxides in 24 h. Furthermore, thesorption of Cr(III) increases with increasing pH (Griffin et al., 1977; Rai et al.,1984) (as the clay surfaces become more negatively charged) and increasingorganic matter content of soils (Paya Perez et al., 1988); whereas the adsorp-tion of Cr(III) decreases when other inorganic cations or dissolved organicligands are present.



Partitioning ratios (Kd) for Cr(III) have been estimated to be very high.Cr(III) sorption is nonlinear, however, so the Kd sorption model cannot beused for assessing retardation except for a particular concentration. If thesepartitioning ratios were equivalent to Kd’s, retardation factors of 500 to 6100could be estimated, indicating the relative immobility of Cr(III) due to sorptionat a pH of around 4. Above this pH, Cr(III) would also be relatively immobilebecause of its low solubility. It is probable that Cr(III) mobility would beenhanced by the formation of complexes due to their decreased sorption orincreased solubility compared with the uncomplexed form of Cr(III).

5.5.2.2 Sorption of Cr(VI)

Chromate ions (the anionic forms HCrO4– and CrO4

2–) can be sorbed by Mn,Al, and Fe oxides and hydroxides (positively charged surfaces), clay mineralsand natural solids and colloids (Rai et al., 1986; James and Bartlett, 1983;Stollenwerk and Grove, 1985; Rai et al., 1988; Griffin et al., 1977; MacNaugh-ton, 1975; Davis and Leckie, 1980; Music et al., 1986; Zachara et al., 1987).These substances commonly coat aquifer materials (Stumm and Morgan,1981). Amorphous iron is the adsorbate found at the highest concentrationsin most aquifer materials. The batch experiments of James and Bartlett (1983)confirm that iron hydroxides strongly sorb to Cr(VI). Batch sorption datafrom experiments on Cr(VI) conducted by Davis and Olsen, 1995 conformedbest to a Freundlich isotherm, but Langmuirean behavior at a neutral pH,and a decreasing Kd of Cr(VI) with increasing concentration, has beenreported (Griffin et al., 1977), probably due to competitive inhibition forsurface sites at higher Cr concentrations. Davis and Olsen, 1995 note thattheir observed linear sorption is probably due to the lower range of Crconcentrations used in their experiments.

This sorption is pH dependent (Richard and Bourg, 1991). At dilute con-centrations, adsorption of Cr(VI) increases as pH decreases, no matter whatthe sorbent (Rai et al., 1986; Bartlett and James, 1979; Rai et al., 1988; Griffinet al., 1977; Rai et al., 1984; Davis and Leckie, 1980; Zachara et al., 1987). Thissuggests that Cr(VI) sorption is favored on sorbents which are positivelycharged at low to neutral pH. Interestingly, compared to clay, sandy materialhas a greater preponderance of positively charged surfaces over the pH 5–7.5range, resulting in a greater affinity for CrO4

– and thus a higher Kd of Cr(VI)on sand than for clay. Lower pH values result in higher Kd values, based onpublished Cr(VI) sorption data for sandy soils compared with soils contain-ing kaolinite and montmorillonite clays.

Sorption of Cr(VI) by clays, soils, and natural aquifer materials is low tomoderate in pH ranges common to groundwater. Sorption of Cr(VI) charac-teristically decreases with increasing pH due to the decrease in positivesurface charge of the sorbing medium. Furthermore, Cr(VI) sorption hasbeen found to be nonlinear (Stollenwerk and Grove, 1985; Griffin et al., 1977),fitting the Langmuir adsorption model. Similar to the discussion of Cr(III)previously, if Cr(VI) adsorption were linear, the calculated partitioning ratios



would correspond to retardation factors of 2.5 to 329 (much lower than thosecalculated for Cr(III)), indicating that Cr(VI) mobility at pH 7 could rangefrom high to low—lower below a pH of 7 and higher above a pH of 7 (andabove a pH of 8.5, Cr(VI) would be entirely unretarded).

Competing anions have a drastic effect on Cr(VI) sorption, with the effectbeing variable, depending on dissolved concentrations of the competinganion and CrO4

2–, on their relative affinities for the solid surface, and onsurface site concentration (Rai et al., 1986). The competitive sorption ofCr(VI) with cations and anions has been investigated by much research (Raiet al., 1986; Stollenwerk and Grove, 1985; Rai et al., 1988; Rai et al., 1984;Music et al., 1986; Zachara et al., 1987; Benjamin and Bloom, 1981). Theelectrostatic sorption of anions is enhanced by cation sorption (due toenhanced positive surface charge). Also, chromates either increase or haveno effect on the sorption of heavy metals (Cd2+, Co2+, Zn2+); competition forsurface sites is relatively minor (Benjamin and Bloom, 1981). Additionally,sorption of chromates in the presence of a mixture of ions is lower than intwo-solution systems, particularly when H2SiO4

2– is present. The effectappears to be additive (Rai et al., 1986; Zachara et al., 1987).

Richard and Bourg (1991) note that the kinetics of Cr(VI) sorption are notwell documented. The sorption of chromates on soils apparently follows atwo-step reaction rate (Amacher et al., 1988). Also, sorption of Cr(VI) doesnot seem totally reversible. Amacher et al. (1986) attributed this lack ofreversibility to reduction of Cr(VI) to Cr(III), possibly by organic matter fromthe soil they studied.

5.6 General Transport and Fate of Chromiumin Environmental Media

Kimbrough et al. (1999) discuss a generalized intermedia transport schemefor environmental chromium. Chromium is directly emitted from industrialactivity either into the air, into water systems (e.g., streams, sewers, lakes,etc.), or to the ground. Airborne chromium eventually settles out into soilor water. In a given parcel of soil, there can be a mixture of Cr(VI) and Cr(III),both naturally occurring and anthropogenic. Cr(VI), but not Cr(III), can beleached out of the soil and enter groundwater, which in turn can becomepart of an aquifer and also migrate to surface waters. As Cr(VI) is leachedfrom the soil, the remaining Cr(III) can slowly oxidize to Cr(VI) to reestablishthe equilibrium of the soil (Bartlett, 1991). In surface waters, Cr(VI) canmigrate in the dissolved form, while both Cr(III) and Cr(VI) can migratewhile being bound with dissolved organic carbon (DOC) or suspendedparticles. And, chromium can migrate from the aqueous phase to sedimentsfrom a dissolved state or with DOC or particles. In the sediment, dissolvedCr(VI) can be immobilized if it enters a stable anaerobic portion, but Cr(VI)in aerobic sediments can be redissolved.



The following are discussions of each of these environmental compart-ments included in the general transport and fate of chromium in environ-mental. Presented are discussions of chromium in the atmosphere, aquaticenvironments (surface waters and groundwater), and soil, and the uptakeof chromium by biota.

5.6.1 Chromium in the Atmosphere

Kotas and Stasicka (2000) and Kimbrough et al. (1999) give detailed discus-sions of the behavior of chromium in the atmosphere, and their work issummarized here. The majority of chromium in the atmosphere (approxi-mately 60 to 70%) is due to anthropogenic sources. Chromium from naturalsources accounts for the remaining amounts (Seigneur and Constantinou,1995). Human activities that can product chromium in the atmosphereinclude metallurgical industries, refractory brick production, electroplating,combustion of fuels, the production of chromium chemicals (i.e., chromatesand dichromates, pigments, chromium trioxide, chromium salts), the cementindustry, production of phosphoric acid in thermal processes, and combus-tion of refuse and sludges (Nriagu, 1988). Natural sources of chromiuminclude volcanic eruptions, erosion of soils and rocks, airborne sea salt par-ticles, and smoke from forest wildfires (Pacyna and Nriagu, 1988). Averageatmospheric concentrations of chromium range from 1 ng/m3 in rural areasto 10 ng/m3 in polluted urban areas (Nriagu, 1988).

The atmosphere has become a major pathway for long-range transfer ofchromium to different ecosystems (Nriagu, 1988; Spokes and Jickells, 1995).Atmospheric chromium-containing particles are transported over varyingdistances by the wind, before they fall or are washed out from the air ontoland and water surfaces, and the distance of transport depends on meteo-rological factors, topography, and vegetation (Nriagu, 1988; Spokes andJickells, 1995). Wet precipitation and dry fallout of chromium from theatmosphere is greatly affected by particle size; the chromium oxidation stateis less important.

The atmospheric transport and fate of chromium largely occurs in theliquid phase and solids phases (i.e., droplets and particles) or, more generally,aerosols instead of as a gas (Seigneur and Constantinou, 1995). The size ofthe particles is important not only to the transport of chromium in theatmosphere, but to health effects as well. Only particles with diameters lessthan 10 µm are respirable; their retention in the lungs can pose a carcinogenicrisk (Friess, 1989).

Chromium in aerosols is generally removed from the atmosphere by bothdry deposition and wet deposition. In dry deposition, the particles settle andare captured by the soil or surface waters via gravitational sedimentation,impaction, or interception. Wet deposition is the process where aerosol parti-cles are actively entrained or scavenged by atmospheric moisture, such as rain,snow, fog, or dew. Chromium can also be introduced, or reintroduced, intothe atmosphere via wind resuspension of chromium-containing soil particles.



The two stable oxidation states of chromium in the atmosphere are Cr(III)and Cr(VI). Atmospheric particles do not contribute to the chemical reactionsthat control the occurrence and ratio between Cr(III) and Cr(VI). Instead,precipitation, complex formation, and oxidation reactions influence theabundance and ratio of Cr(III) and Cr(VI).

Computer simulations by Seigneur and Constantinou (1995) have led tothe conclusion that typical atmospheric conditions favor the Cr(VI) reductionto Cr(III). This is the likely case because of the presence and concentrationsof reducing agents in the air (i.e., V2+, Fe2+, H2S, HSO3

–, NO2–, and organic

materials) as well as the acidity of the atmosphere. Cr(VI) can be reducedrapidly in the atmosphere based on theoretical (Seigneur and Constantinou,1995) and experimental (Grohse et al., 1988) studies. Estimates of atmo-spheric half-life for Cr(VI) reduction to Cr(III) range from 16 h (Grohse et al.,1988) to 4.8 days. The few materials capable of oxidizing Cr(III) to Cr(VI) inthe atmosphere, such as ozone, occur in concentrations too low to producemeasurable conversions in the atmosphere.

5.6.2 Chromium in Aquatic Environments

5.6.2.1 Surface Waters

Kotas and Stasicka (2000) and Kimbrough et al. (1999) give detailed discus-sions of the behavior of chromium in surface aquatic environments, and theirwork is summarized here. Chromium in natural waters originates fromnatural sources or from manmade pollution. Natural sources include theweathering of rock constituents, wet precipitation and dry fallout from theatmosphere, and run-off from the terrestrial systems. Manmade pollutionsources to waters (mostly surface waters such as rivers) include the dischargeof industrial wastewaters (i.e., from the metallurgical, electroplating, tan-ning, and dying industries), from sanitary landfill leaching, and from watercooling towers (Nriagu, 1988). The number and type of chromium speciespresent in effluents depend on the character of the industrial processes usingchromium.

In natural waters, chromium exists in its two stable oxidation states, Cr(III)and Cr(VI). The presence and ratio between these two forms depend onvarious processes, which include chemical and photochemical redox trans-formation, precipitation/dissolution reactions, and adsorption/desorptionreactions. Simplistically, Cr(III) should be the only form present in anaerobicor subanaerobic conditions, whereas in aerobic aqueous environments,Cr(VI) should be the only form present. However, the presence of Cr(III)and Cr(VI) is also dependent on the pH of the water. Under neutral tobasic conditions, Cr(III) will tend to precipitate out, while under acidconditions, Cr(III) will tend to solubilize. While Cr(VI) ions (i.e., chromateand dichromate) are extremely water soluble at all pHs, they can precipitatewith a number of divalent cations. In waters of intermediate pH values, theCr(III)/Cr(VI) ratio is largely dependent on the concentration of oxygen.



In contrast to the atmosphere, many aqueous environments do containoxidizing agents, such as MnO2 and Mn3+ in sufficiently high concentrationsto produce measurable yields of Cr(VI). In oxygenated surface waters, notonly are pH and oxygen concentration important, but the nature and con-centrations of reducing agents, oxidation mediators, and complexing agentsplay important roles. These factors seem to be responsible for the occurrenceof significant Cr(III) quantities in many oxygenated surface waters (Kieberand Heiz, 1992; Cranston and Murray, 1978 and 1980; Pettine et al., 1991).At times, Cr(III) can be the predominant chromium component in oxygen-ated waters (Chuecas and Riley, 1966). Several mechanisms for this mightinclude Cr(VI) reduction via Fe(II), Cr(VI) reduction via H2O2, and dissolvedorganic matter. Other reducing agents might include hydrogen sulfide (H2S),sulfur (S), NH4, nitrate (NO3), and vanadium (V2+) (Eary and Rai, 1988; Bodeket al., 1988). The photochemical generation of Cr(III) has also been suggested(Kieber and Heiz, 1992; Kaczynski and Kieber, 1993).

Both Cr(III) and Cr(VI) have been shown to bind with naturally occurringdissolved organic carbon (DOC). Organically-bound Cr(III) can stay in solu-tion at higher pH than unbound Cr(III) (Palmer and Wittbrodt, 1991) andorganically bound chromium can also sorb to and desorb from the organicportion of suspended and settled sediments. Therefore, chromium migratesas either dissolved ions or as attached to particles, or both.

The effect of sunlight can be important in surface water chemistry. Theoxidation and reduction of chromium are affected by sunlight (Kaczynski andKieber, 1994). Sunlight appears to degrade organically-bound chromium, andthis process releases inorganic chromium. Also, sunlight acts indirectly byassisting the reduction of iron, which also results in the formation of H2O2

(Kieber and Heiz, 1992; Beaubien et al., 1994) affecting the oxidation state ofchromium. Additionally, sunlight aids the oxidation of manganese (Bartlett,1991), also affecting the oxidation state of chromium.

Chromium transport and fate in surface waters can be discussed by usingthree subsystems: rivers, lakes, and oceans. The transport pathways arecontrolled by specific conditions prevailing in each of these subsystems,including temperature, depth, degree of mixing, oxidation conditions, andamount or organic matter. Chromium as a component of suspended particlesis the most important transport mechanism in rivers. Dissolved chromium inriver water decreases during its passage through turbid coastal environments.

Lakes generally have relatively high levels of biologic activity and highratios of sediment-to-water surface area, which greatly influence transportof metals. The high level of organic matter creates the medium for reductionand the formation of complexes, favoring the reduction of Cr(VI) to Cr(III),which is afterwards rapidly precipitated or sorbed onto the sediment min-erals. And, chromium in sediments can be remobilized into the surroundingpore water via oxidation or solubilization of Cr(III) sediments. The mostcomplex transport pathways of chromium are in seasonally anaerobic lakes(Beaubien et al., 1994; Achterberg et al., 1997) where deep basinal water inthe summer months becomes anaerobic due to the coupling of high biological



activity and thermal stratification. Therefore, depth and season heavily influ-ence the concentration and speciation of chromium. Dissolved chromiumusually decreases in the summer months, and the areas dominated by Cr(VI)versus Cr(III) become more segregated to surface and deep layers, respec-tively. This distribution of chromium species is consistent with what wouldbe expected from seasonal increases in temperature, a decrease in pH, andthe oxygen content in the basinal water. The aerobic regime favors Cr(VI),and Cr(III) is favored in anaerobic areas.

Chromium generally enters oceans via rivers and from atmospheric fall-out. Atmospheric inputs result in more homogeneous distribution of chro-mium in the ocean water compared with river inputs; the latter are thesubject of eustarine removal processes and ocean circulation patterns Spokesand Jickells, 1995). It has been proposed that chromium sources to the oceansare mostly as particles (suspended solids from river and aerosols). In oceanwaters, dissolved and precipitated chromium exist together in equilibrium.Dissolved chromium is generally removed from the aqueous phase andincorporated into biologic material (i.e., siliceous and carbonaceous skele-tons) and by adsorption onto sediment particles. This removal occurs bothin the water column and at the sediment–water interface, resulting in deepand bottom–water enrichment of dissolved chromium.

Except in estuaries, chromium concentrations in seawater are dominatedby chromates, probably due to the generally oxidizing conditions in the oceanand low suspended concentration of particles. Reduction of Cr(VI) occurs inanaerobic basins and the oxygen-free zones, where increased Cr removalmay be due to Cr(III) adsorption onto bottom sediments (Smith et al., 1995).Chromium cycling in the water column occurs in response to nutrient bio-geochemistry. When Cr scavenged by particles is deposited on the oceanfloor, diagenetic processes can lead to remobilization of Cr either as chromateor as organic Cr(III) complexes. The remobilization of Cr(III) from sedimentcan also occur by its oxidation, carried out mostly by manganese dioxide.

5.6.2.2 Groundwater

Richard and Bourg (1991) and Calder (1988) give detailed discussions of thebehavior of chromium in groundwater environments, and their work issummarized here. The mobility of chromium in groundwater depends onits solubility and its tendency to be sorbed by soil or aquifer materials. Thesefactors, in turn, depend on the groundwater chemistry and the characteristicsof soil or aquifer material in contact with the chromium-containing ground-water. Otherwise, much of the basic considerations discussed for surfacewaters also apply for the understanding of the transport and fate of chro-mium in groundwater.

Large plumes of chromium-contaminated groundwater in shallow aqui-fers have been well documented (Stollenwerk and Grove, 1985; Deutsch,1972; French et al., 1985; Perlmutter and Lieber, 1970; Wiley, 1983). Sourcesof this contamination can be the same as those listed for surface waters. Such



plumes in sand and gravel aquifers have been reported to reach lengths ofup to 1300 m (Perlmutter and Lieber, 1970). Impacts of many of the plumeshave been serious enough to necessitate the abandonment of local ground-water supplies.

Groundwater contamination by chromium can be extensive in permeableaquifers (i.e., sand and gravel and fractured rock aquifers) because ground-water velocities in these materials are relatively high (i.e., about 0.1 and 5 mper day, respectively). On the other hand, groundwater velocities in aquifersof much lower permeability (i.e., clayey materials) tend to be low, perhapson the order of a few centimeters or less per year. Thus, chromium-contam-inated groundwater in these settings cannot extend far from the source ofthe chromium.

Cr(III) tends to be relatively immobile in most groundwater because of theprecipitation of Cr(III) compounds of low-solubility (e.g., Cr(OH)3(s),FeCr2O4(s), (Fe1–x, Crx)(OH)3(ss)) in neutral to alkaline pH range (i.e., abovepH 4). This results in low Cr(III) dissolved concentrations. Also, in neutralto slightly acidic conditions (i.e., especially below pH 4), Cr(III) is removedfrom solution by sorption. Calder (1988) reported that sorption of Cr(III)increases with increasing pH. Furthermore, it has been speculated that Cr(III)may be mobile in groundwater if it is in a complexed form, although thishas not been documented in the field. Precipitation and sorption can beinhibited by complex formation with dissolved ligands such as naturalorganic matter (Gerritsee et al., 1982).

Cr(VI) tends to be moderately to highly mobile in most shallow ground-water aquifers. This tends to be due to two major factors: (1) the lack ofsolubility constraints; and, (2) the low to moderate sorption of Cr(VI)anionic form in neutral to alkaline waters. In soils or sediments with highcontent of Fe and Mn oxides conditions, Cr(VI) should be removed bysorption processes (Rai et al., 1986; Eary and Rai, 1987; Stollenwerk andGrove, 1985; Rai et al., 1988; Cary et al., 1977). But sorption is significantlydepressed by competing background anions (Rai et al., 1988) so that Cr(VI)can expected to be highly mobile. In alkaline environments, sorption is notstrong enough to keep Cr(VI) from migrating through soil or sediments.Cr(VI) sorption generally increases with decreasing pH, so sorption ofCr(VI) can be very significant in neutral to acidic groundwater. As previ-ously discussed, Cr(VI) sorption can be strongly nonlinear, such that sorp-tion decreases with increasing Cr(VI) concentration. Also, sorption alsoappears to be rate-dependent, so the kinetics of the sorption process wouldvery important, especially in high-velocity groundwater regimes. Also,Cr(VI) reduced to Cr(III) with subsequent precipitation and sorption isbelieved to control the mobility of Cr(VI) (Rai et al., 1988): Cr(VI) reductionto Cr(III), which is afterwards rapidly precipitated or sorbed. In Fe(II)-richand dissolved organic matter-rich environments, the reduction of Cr(VI) ismore likely to occur and the resulting aqueous Cr(III) concentration will becontrolled by the solubility of Cr(III). In such cases, Cr(VI) should not migratesignificantly.



The impact of water chemistry on the presence and movement of chro-mium in groundwater was demonstrated in field experiments involvinginjections of 100 µmol/l of Cr(VI) into various zones of a gravel aquifer(Kent et al., 1989). Some chromium disappeared from the aqueous phase inthe anaerobic part of the aquifer due to reduction to the less soluble Cr(III)form. Cr(VI) (chromate) generally migrated at about the same rate asgroundwater flow, except in areas with low pH and low concentrations ofanions, where it was retarded due to competition with these forms forsorption sites.

The relationship of chromium and manganese is particularly interestingin the transport and fate of chromium in groundwater. Cr and Mn form apair of chemical elements with contrasting tendencies (Murray et al., 1983).Under oxidizing conditions, Cr(VI) is soluble as CrO4

2– while Mn(IV) isscavenged as MnO2(s). Under reducing conditions, Cr(III) is removed fromsolution as Cr(OH)3(s) while Mn(II) is soluble as Mn2+. These contrastingtendencies for the solubility of Cr and Mn have been observed in shallowgroundwater of the Western San Joaquin Valley in California (Deverel andMillard, 1988). In the alluvial-fan geologic zone, dissolved Cr concentrationsare high, whereas dissolved Mn is low. However, in the basin-trough zone,Cr concentration is low and Mn concentration is high.

5.6.2.3 Plumes of Chromium in Groundwater—Case Studies

There are a number of documented cases of Cr plumes in groundwater. Thefollowing are accounts of major occurrences of Cr in groundwater as sum-marized by Calder (1988) and Loyaux-Lawniczak et al. (2001).

5.6.2.3.1 Nassau County, New YorkThe best known and the first major published case study in North Americaof Cr in water supply wells was in Nassau County, Long Island (Lieber et al.,1964). A number of investigators have studied the site, and this case dem-onstrates the types of uncertainties that complicate predictions of Cr migra-tion in groundwater. The source of Cr was an aircraft plant that used Crsolutions for anodizing and plating metals. The site is located on a verypermeable sand and gravel aquifer with groundwater velocities estimatedto be approximately 0.15 to 0.5 m per day (Perlmutter and Lieber, 1970). Theestimated length and width of the plume was 1300 m by 300 m, with amaximum Cr concentration in groundwater of 40 mg/l. The groundwaterpH was 4.6 to 6.2 which is the range where Cr(VI) could be significantlysorbed by the aquifer materials. Calder (1988) promotes three hypothesesthat can account for the apparent retardation of the Cr plume:

1. Greater sorption as a result of the lowering of the chromium con-centration due to remediation efforts (i.e., concentration-dependenceof chromium sorption).



2. The slow reduction of Cr(VI) to Cr(III), with subsequent precipita-tion of Cr(III), particularly in deeper groundwater.

3. Slow kinetics of the sorption process in the more permeable por-tions of the aquifer, such that chemical equilibrium, and thereforemaximum sorption, would not occur until the aquifer had beenexposed to chromium-containing groundwater over a long periodof time.

5.6.2.3.2 Telluride, ColoradoThis is the site of a heavy metal mining and milling operation near Tellu-ride, Colorado (Grove and Stollenwerk, 1985). The source of chromiumwas a tailings pond, which, since 1977, apparently discharged chromium-containing wastes into the groundwater system. Water in the tailing pondhad chromium concentrations of 8.8 mg/l (Stollenwerk and Grove, 1985).The USGS initiated a study of the site in October 1978, providing anexcellent opportunity for comparison of field observations with laboratoryexperiments and computer simulation (Stollenwerk and Grove, 1985). Theshallow aquifer was gravel and sand alluvium, and the estimated ground-water velocities were approximately 5 m per day. In 1979, a chromiumplume at least 520 m long was observed, with a maximum chromiumconcentration of 2.7 mg/l (Grove and Stollenwerk, 1985). The groundwaterpH was approximately 6.8 (Stollenwerk and Grove, 1985). Laboratory andfield investigations determined that the chromium was retarded by a factorof 10 relative to the groundwater velocity. Even with such an appreciableretardation rate, the high groundwater velocity resulted in a relativelymobile chromium plume.

5.6.2.3.3 Southwestern Michigan—Wood Treatment PlantAn incidence of groundwater contamination by chromium from a woodtreatment plant was reported in southwest Michigan in 1980 (French et al.,1985). The wood was treated with a 2% aqueous solution containing 47.5%chromium trioxide, 34% arsenic pentoxide, and 18.5% copper oxide. Effluentfrom a sump pit was discharged to the ground adjacent to the treatmentbuilding until 1980. An effluent sample from the pit contained 1600 mg/lchromium, of which 1500 mg/l was Cr(VI). The site is located on a permeableoutwash plain consisting of gravelly sands with up to 17% silt and clay-sizedparticles. The water table was at a depth of about 8 m. The groundwaterhad a moderately alkaline pH of up to 8.4, and groundwater velocities wereestimated at approximately 0.15 m per day. The highest chromium concen-tration in the plume was 6.58 mg/l, and chromium was detected in thefacility’s supply well as high as 2.5 mg/l. The chromium plume, defined bytotal chromium concentrations above 50 parts per billion, extended approx-imately 600 m from the discharge area, with a width of approximately 200 mand a vertical thickness of 20 m. The length of the plume was found to beconsistent with estimated groundwater velocities, suggesting that Cr was



essentially unretarded. It was assumed that the chromium was almostentirely Cr(VI). A purge well-spray irrigation system was established torestore the aquifer to drinking water standards.

5.6.2.3.4 Industrial Waste Landfill, Northern FranceThe site, an industrial waste landfill located in northern France, wasoperational from 1905 to 1982, producing materials including dichromates,chromic acid, sulfuric acid, and phosphates. Chromite and pyrite werethe main primary minerals used at the facility, and Cr mineral processingwastes were collected in a slag heap, which was covered in 1990 by ageomembrane to limit runoff. The groundwater table is normally locatedat 2 m depth, with an annual fluctuation of 1 m, within the infill. Thewater table aquifer, in a silt layer, is well separated from deeper aquifersby a green clay unit. The hydraulic conductivity was estimated to be about7 × 10–7 m/s. The chromium concentrations in the source area were about210 mg/l, with the migrating plume extending approximately 160 mdowngradient. The plume does not extend to the downgradient boundaryof the site 375 m away. The Fe(II) distribution in groundwater is quitevariable. It is virtually absent in the major area of the Cr plume, then theFe(II) concentration abruptly increases near the downgradient end of theCr plume (1680 mg/l). Concentrations of total copper, cadmium, zinc, andsulfate ions show a similar distribution to Fe(II). The pH values in thegroundwater are mostly neutral (6.5 to 7.3) in the major portion of theplume, becoming more acidic (approximately 4) at the downgradient por-tion of the Cr plume, possibly due to the oxidation of pyrite that was usedin massive amounts in this area of the site. Loyaux-Lawniczak et al. (2001)explain the distribution of metals in groundwater by postulating thatCr(VI) (produced by leaching of the ore residue slag heap) migrates ingroundwater flow and then enters into a reducing zone (with Fe(II)present), where the Cr(VI) is reduced by Fe(II). It has been widely acceptedthat the kinetics of this reaction in solution is fast, and that with excessFe(II), all of the Cr(VI) is reduced. Cr(III) is especially immobilized in theclay fraction of the soil; analyses of this clay fraction revealed that mont-morillonite flakes are the chromium-bearing mineralogical phase. In sum-mary, Cr(VI) migration in groundwater is retarded horizontally by a redoxmechanism involving chromate ions and ferrous ions or Fe(II)-bearingminerals, and vertically by a thick green clay unit.

5.6.3 Chromium in Soil

The discussion of chromium behavior in soil presented here is preceededby an overview of the presence and behavior of metals in soil. The discus-sion specific to the behavior of chromium in soil includes a general over-view followed by discussions of the sorption, oxidation, and reduction ofchromium.



5.6.3.1 Overview of Metals in Soil

McLean and Bledsoe (1992) present an overview of metals in soil, and theirwork is discussed below. Metals are found in soil within one or more soil“pools”:

1. Dissolved in the soil solution2. Occupying exchange sites on inorganic soil constituents3. Specifically sorbed on inorganic soil constituents4. Associated with insoluble soil organic matter5. Precipitated as pure or mixed solids6. Present in the structure of secondary minerals7. Present in the structure of primary minerals

Metals exist in soil solution as either free (uncomplexed) metal ions (e.g.,Cr3+), or in soluble complexes with inorganic or organic ligands, or associatedwith mobile inorganic and organic colloidal material. A complex is a molec-ular unit where a central metal ion is bonded by a number of associatedatoms or molecules ( e.g., Cr(OH)4

–), and these associated atoms or moleculesare termed ligands (i.e., OH– is a ligand). Metals will form soluble complexeswith inorganic ligands such as SO4

2–, Cl–, OH–, PO43–, NO3

–, and CO32–.

Soluble complexes with organic ligands are not as well defined. The freemetal ion is generally the most bioavailable and toxic form of the metal.With complex formation, the resulting metal species may be positively ornegatively charged or be electrically neutral. The presence of a complexspecies of a metal in the soil solution can significantly affect the migrationof metals through the soil relative to the free metal ion. In addition todissolved metal complexes, metals also may associate with mobile colloidalparticles (size of 0.01 to 10 µm). Colloidal particles include iron and manga-nese oxides, clay minerals, and organic matter. These surfaces have a highcapacity for metal sorption.

The extent of migration of metals from the ground surface into and throughthe subsurface depends on the retention capacity of the soil is exceeded, andit is directly related to the solution and surface chemistry of the soil and tothe specific properties of the metal and associated waste matrix. The mech-anisms for retention of metals in soil include sorption and precipitation.Retention of cationic metals is related to soil properties such as pH, redoxpotential, surface area, cation exchange capacity, organic matter content, claycontent, iron and manganese oxide content, and carbonate content. Anionicmetal retention has been correlated with pH, iron and manganese oxidecontent, and redox potential. Consideration must also be given to the typeof metal, its concentration, the presence of competing ions and complexingligands, and the pH and redox potential of the soil-waste matrix. Also, themigration of metals can depend on the type of wastes that may be associatedwith the metal. Therefore, because of the differing varieties of soils, and the



many different forms of metals themselves and the wastes containing them,evaluating the extent of metal retention by a soil is site specific, soil specific,and waste specific.

Precipitation and sorption are the main metal retention mechanisms insoil. Precipitation is where metals precipitate to form a solid (three-dimen-sional) phase in soils, which might be a pure solid or a mixed solid (e.g.,(FexCr1–x)(OH)3); the latter forms when various elements co-precipitate. Sorp-tion of metals is the accumulation of ions at the solid phase—aqueous phaseinterface. Sorption of metals in the soil matrix often involves organic matter,clay minerals, iron and manganese oxides and hydroxides, carbonates, andamorphous aluminosilicates. Binding of metals to organic matter rangesfrom weak forces of attraction to formation of strong chemical bonds. Soilorganic matter can be the main source of soil cation exchange capacity.Organic matter content generally decreases with depth in soil, so that themineral (inorganic) constituents of soil will become a more important surfacefor sorption with increasing depth. There have been numerous studies ofthe adsorptive properties of clay minerals, in particular montmorillonite andkaolinite, and iron and manganese oxides. Griffin and Shimp (1978) foundthe relative mobility of 9 metals through montmorillonite and kaolinite tobe: Cr(VI) > Se > As(III) > As(V) > Cd > Zn > Pb > Cu > Cr(III). Also, Jenne(1968) concluded that Fe and Mn oxides are the principal soil surface thatcontrol the mobility of metals in soils and water. In arid soils, carbonateminerals may immobilize metals by providing sorbing and nucleating sur-face (Santillan-Medrano and Jurinak, 1975; Cavallaro and McBride, 1978;McBride, 1980; Jurinak and Bauer, 1956; McBride and Bouldin, 1984; Dudleyet al., 1988, 1991).

Generally, the sorption capacity for anions (some metals form anioniccontaminants) is lower than the cation sorption capacity of soils. The sorptioncapacity (both exchange and specific sorption) of a soil is determined by thenumber and kind of sites available. Sorption process are affected by varioussoil factors (i.e., pH, redox potential, clay, soil organic matter, oxides, andcalcium carbonate content), by the form of the metal added to the soil, andby the solvent introduced along with the metal. Therefore, interactions ofthese influences on sorption may increase or decrease the migration of metalsin soil.

Although the principles affecting sorption and precipitation are similar forcationic and anionic metals, the following is a list with a brief descriptionof factors affecting the behavior of cationic metals in soils.

Competing cations: Trace cationic and anionic metals are preferentiallysorbed over the major cations (Na+, Ca2+, Mg2+) and major anions (SO4

2–,NO3

–). However, when the specific sorption sites become saturated, exchangereactions dominate and competition for these sites with soil major ionsbecomes important.

Complex formation: Metal cations form complexes with inorganic andorganic ligands, whereby the ligand forms a coordinate bond with the metalatom. The resulting association has a lower positive charge than the free



metal ion (and might even be uncharged or negatively charged). The effectof complex formation on sorption is dependent on the type and amount ofmetal present, and type and amount of ligands present, soil surface proper-ties, soil solution composition, pH, and redox. The presence of complexingligands may either increase metal retention or greatly increase metal mobil-ity. Data from the literature that do not consider the presence of complexingligands at the site, both organic and inorganic, may lead to significant errorin estimating metal mobility.

pH: The pH of the soil system is a very important parameter, directlyinfluencing sorption/desorption, precipitation/dissolution, complex forma-tion, and oxidation–reduction reactions. The pH affects several mechanismsof metal retention of soils both directly and indirectly. Sorption increaseswith pH for all cationic metals, but retention does not significantly increaseuntil the pH gets above 7. As true for all oxyanions (i.e., Cr(VI)), sorptiondecreases with pH. The pH dependence of sorption reactions of cationicmetals is due in part to the preferential sorption of the hydrolyzed metalspecies in comparison to the free metal ion. Many sorption sites in soils arepH dependent (i.e., Fe and Mn oxides, organic matter, carbonates, and theedges of clay minerals). As the pH decreases, the number of negative sitesfor cation sorption diminishes while the number of sites for anion sorptionincreases. Also, as the pH becomes more acidic, metal cations also face com-petition for available permanent charged sites by Al3+ and H+. Jenne (1968)stated that hydrous oxides of Fe and Mn play a principle role in the retentionof metals in soils. Solubility of Fe and Mn oxides is also pH-related. BelowpH 6, the oxides of Fe and Mn dissolve, releasing sorbed metal ions tosolution (Essen and El Bassam, 1981). In soils with significant levels ofdissolved organic matter, increasing soil pH may actually mobilize metaldue to complex formation. A word of caution is warranted here, however.The complexity of the soil-waste system (several types of surfaces and solu-tion compositions) may render generalizations just given to be not true. Forexample, cationic metal mobility can actually increase with increasing pHdue to the formation of metal complexes with dissolved organic matter.

Oxidation-Reduction: Many metals have more than one oxidation state, andare directly affected by changes in the oxidation-reduction (redox) potentialof the soil. Redox reactions can greatly affect contaminant transport, inslightly acidic to alkaline environments. In general, oxidizing conditionsfavor retention of metals in soils, while reducing conditions contribute toaccelerated migration.

5.6.3.2 Behavior of Chromium in Soil

Kimbrough et al. (1999); Kotas and Stasicka (2000); and McLean and Bledsoe(1992) present overviews of the behavior of chromium in soil, and their workis summarized here. More details concerning the chemistry of chromium insoils and sediments are provided in review articles by Cary (1982); andRichard and Bourg (1991).



Chromium commonly exists in two oxidation states in soils, Cr(III) andCr(VI). Forms of Cr(VI) in soils are as hydrogen, chromate ion (HCrO4

–

predominant at pH < 6.5 or CrO42– predominant at pH 6.5), and as dichro-

mate (Cr2O72–) predominant at higher concentrations and at pH 2 to 6. The

dichromate ion is more toxic to humans than chromate ion. In neutral-to-alkaline soils, Cr(VI) is mostly soluble (e.g., Na2CrO4) but also in moderately-to-sparingly soluble chromates (e.g., CaCrO4, BaCrO4, PbCrO4) (Bartlett andKimble, 1976; James, 1996). In more acidic soils (pH < 6), HCrO4

– becomesa dominant form. The chromate ions (CrO4

2– and HCrO4–) are the most

mobile forms of Cr in soils. They can be taken up by plants and easily leachedout into the deeper soil layers causing groundwater and surface water pol-lution (Calder, 1988; James and Bartlett, 1984; Handa, 1988). Some minoramounts of Cr(VI) are bound in soils. This binding depends on the miner-alogical composition and pH of the soil. The CrO4

2– ion can be sorbed bygoethite, FeO(OH), aluminum oxides and other soil colloids with a posi-tively charged surface (Richard and Bourg, 1991; James and Bartlett, 1983and 1988). The HCrO4

– ion, which occurs in more acidic soils, may also beheld in soils, or remain soluble (James and Bartlett, 1983). Reviews of theprocesses that control the fate of chromium in soil and the effect theseprocesses have on remediation are given in Bartlett (1991) and Palmer andWittbrodt (1991).

In a study of the relative mobilities of 11 different trace metals for a widerange of soils, Korte et al. (1976) found that clayey soil, containing free ironand manganese oxides, significantly retarded Cr(VI) migration. Cr(VI) wasthe only metal that was highly mobile in alkaline soils. The study alsoshowed that free iron oxides, total manganese, and soil pH influencedCr(VI) immobilization, whereas soil properties such as cation exchangecapacity, surface area, and percent clay had no significant influence onCr(VI) mobility.

Chromium in soils is naturally present mostly as insoluble Cr(OH)3 or asCr(III) sorbed to soil components, which prevents leaching into the ground-water or its uptake by plants (Bartlett and Kimble, 1976). The dominantchromium form depends strongly on pH; in acidic soils (pH<4) it isCr(H2O)6

3+, whereas at pH < 5.5 it is its hydrolysis products, mainly solubleCrOH2+ (Ritchie and Sposito, 1995); both these forms are easily sorbed byclays. The process is intensified by an increase in pH, which can be inter-preted in part due to an increase of negative charge on the clays. Here humicacids contain donor groups forming stable Cr(III) complexes. The Cr(III)sorption to humic acids renders it insoluble, immobile, and unreactive; thisprocess is the most effective within the pH range of 2.7 to 4.5 (James, 1996).In contrast, mobile compounds such as citric and fulvic acid, form solubleCr(III) complexes which control its oxidation to Cr(VI) in soils (Bartlett andKimble, 1976; Bartlett and James, 1979; James and Bartlett, 1983; James, 1996;Wittbrodt and Palmer, 1995).

Both reduction of Cr(VI) to Cr(III) and the sorption of Cr(VI) can occur insoil, even simultaneously. Therefore, this causes a problem in assigning one



mechanism of observed attenuation of Cr(VI) in the subsurface (Bartlett,1991). The following are discussions of each of these mechanisms.

5.6.3.2.1 Sorption of Cr(III) and Cr(VI)Cr(VI) is an anion, and its association with soil surfaces is thus limited topositively charged exchange sites, which decrease in number with increasingsoil pH. Therefore, sorption of Cr(VI) decreases with increasing soil pH. Ironand aluminum oxide surfaces will adsorb chromate ions (CrO4

2–) at acidicand neutral pH (Davis and Leckie, 1980; Zachara et al., 1987; Ainsworthet al., 1989). The sorption of Cr(VI) in groundwater by alluvium aquifermaterials has been found to be due to the iron oxides and hydroxides coatingthe alluvial particles (Stollenwerk and Grove, 1985). But the sorbed Cr(VI)was desorbed by adding uncontaminated groundwater, indicating nonspe-cific sorption of Cr(VI). The presence of chloride and NO3 have little effecton Cr(VI) sorption, whereas sulfate and phosphate tend to inhibit sorption(Stollenwerk and Grove, 1985).

Chromates can be sorbed by iron, aluminum oxides, amorphous alumi-num, hydroxides, organic complexes, and other soil components, which mayprotect Cr(VI) from reduction (Bartlett, 1991; Bartlett and James, 1988). Also,Cr(III) materials can be sorbed onto silicacious components. In the aqueousphase of soils, Cr(III) that is not sorbed by the solid phase would generallyhydrolyze to the hydroxide and precipitate. Chromate would be far lesslikely to precipitate and so would be expected to be more mobile. In thissituation precipitation reactions are closely tied to oxidation and reductionreactions. In anaerobic sediments, oxidation is unlikely to take place andchromium hydroxide [Cr(OH)3] could be immobilized as long as the sedi-ments are physically stable (Eary and Rai, 1989).

Soil pH determines both the speciation of Cr(VI) and the charge charac-teristic of the surface with which it reacts (James and Bartlett, 1983). AbovepH 6.4, HCrO4

– dissociates to CrO42– as the dominant form of Cr(VI) and the

charge characteristic of the surface with which it reacts (James and Bartlett,1983), and the CrO4

2– may in turn be sorbed. Sorption of chromates can bea reversible process suggested by leaching of Cr(VI) from soils (Baron et al.,1996). However, such reversibility depends on the chemistry of the leachateand of the soil or sediment.

5.6.3.2.2 Reduction of Cr(VI) and Oxidation of Cr(III)Oxidation and reduction reactions can convert Cr(III) to Cr(VI) and viceversa (Bartlett and Kimble, 1976; Bartlett and James, 1979; James and Bartlett,1983; James and Bartlett, 1988; Wittbrodt and Palmer, 1995; James, 1994;Powell et al., 1995; Deng and Stone, 1996). These processes depend on pH,oxygen concentration, presence of appropriate reducers and mediators act-ing as ligands or catalysts. Mobile forms for Cr(VI) (HCrO4

– and CrO42–) can

be reduced by different reducers such as Fe(II) or S2–.



It has been thought that Cr(VI) can be reduced to Cr(III) under normalsoil pH and redox conditions. However, Bloomfield and Pruden (1980) rein-vestigated earlier claims that Cr(VI) is readily reduced to Cr(III) under suchnormal soil conditions, and they found that the analytical methods used inprevious investigations (Bartlett and Kimble, 1976) were unreliable becausethe soil extracts probably contained organic matter capable of reducingCr(VI). They also found that the reduction of Cr(VI) in soil of normal pHwas not particularly rapid under aerobic conditions.

The reduction of chromates by Fe, V2+, sulfides, and organic materials iswell demonstrated (Cary, 1982), and the kinetics of Cr(VI) reduction hasbeen reported to follow a simple first-order reaction kinetics (Amacher andBaker, 1982; Bartlett and James, 1988). Soil organic matter has been identifiedas the important electron donor (i.e., the principal reducing agent) in thisreaction (Bartlett and Kimble, 1976; Bloomfield and Pruden, 1980), and thereduction of Cr(VI) in the presence of organic matter proceeds at a slow rateat normal levels of pH and temperatures found in the environment (Bartlettand Kimble, 1976; James and Bartlett, 1983). This relatively slow rate ofCr(VI) reduction increases with decreasing soil pH (Bloomfield and Pruden,1980; Cary et al., 1977). In the absence of soil organic matter, Fe(II)-containingminerals reduce Cr(VI), however, tests have shown that this reduction occursin subsurface soil with a low pH (<5) (Eary and Rai, 1991). Reduction byFe(II) is more favorable under anaerobic conditions since oxygen can oxidizethe iron (II) (Fendorf and Li, 1996). However, high concentrations of Cr(VI)may quickly exhaust the available reducing capacity of the soil, and excessCr(VI) may persist for years in soils (Baron et al., 1996). In general, it hasbeen noted that chromates are relatively stable and mobile in soils that aresandy or have low organic content (Cary, 1982; Bloomfield and Pruden, 1980;Frissel et al., 1975).

Under conditions prevalent in some soils, Cr(III) can be oxidized (Bartlettand James, 1979). Only a few oxidants present in soils and sediments (i.e.,dissolved oxygen and MnO2) are capable of oxidizing Cr(III) to Cr(IV). Theoxidation of Cr(III) by MnO2 (which serves as an electron acceptor) has beenshown to occur in soils (Eary and Rai, 1987; Johnson and Xyla, 1991; Fendorfand Zasoski, 1992), and aerobic sediments, but not in anaerobic sediments.Oxidation of Cr(III) by dissolved oxygen has been found to be insignificant(Rai et al., 1989) when compared with MnO2, which is the most likely oxidantof Cr(III) in soils. Thus, if soluble Cr(III) is added to an “average” soil, aportion of the soluble Cr(III) will become immediately oxidized by MnO2 toCr(VI) (Cary, 1982). The rest of the Cr(III) may remain reduced for longperiods of time, even in the presence of electron-accepting manganeseoxides, perhaps because soluble Cr(III) can form complexes with low-molec-ular mass organic molecules and then be oxidized where redox conditionsare optimal. Added organic matter also may facilitate oxidation of Cr(III) toCr(VI). This has implications to remediation strategies. The addition oforganic residues potentially as a remediation strategy for Cr(VI)-contami-nated soils containing high levels of oxidized manganese may result in the



formulation of unstable Mn(III) organic complexes that not only temporarilyprevent Cr(III) oxidation but also promote the desired reduction of Cr(VI)(Bartlett and James, 1988).

5.6.4 The Uptake and Transformation of Chromium by Biota

Kotas and Stasicka (2000) and Kimbrough et al. (1999) present overviews ofthe uptake and transformation of chromium by biota, and their work issummarized here. Chromium can be taken up by biota from the air, water,and soil. Most studies in this realm do not distinguish between the oxidationstates of chromium. However, it is known that the Cr(VI) form is moreavailable for living organisms than Cr(III), and the uptake of Cr(VI) by biotais a main role in the removal of Cr(V) from water and soil systems. Thefollowing discussion of the role of biota in the transport and fate of chromiumis presented for bacteria, plants, aquatic animals, and terrestrial animals.

Microorganisms accumulate chromium (Coleman, 1988) and reduce Cr(VI)to Cr(III) (Campos et al., 1995; DeLeo and Ehrlich, 1994). Although highlevels of Cr(VI) are toxic to microorganisms (Bartlett, 1991), chromium isimportant to yeast metabolism (Coleman, 1988; Anderson et al., 1977). How-ever, there is not much evidence in the literature of bioaccumulation ofchromium as Cr(VI) in bacteria, since most studies report chromium bioac-cumulation in terms of total chromium. There are conflicting views concern-ing the uptake and translocation of Cr(VI) in plants. Also, whether chromiumis an essential element in plants has been debated in the literature. The WorldHealth Organization says it is unknown whether chromium is an essentialnutrient for all plants, yet all plants contain the element. On the other hand,Richard and Bourg (1991) suggested that Cr(III) is an essential nutrient inplant metabolism (amino and nucleic acid synthesis). The literature on chro-mium bioaccumulation in aquatic animals (e.g., finned fish) suggests thatCr(VI) is not expected to accumulate and increase in the aquatic food chain.And, there is no indication of biomagnification of chromium within theterrestrial animal food chain (soil-plant-animal) (Clay, 1992; ATSDR, 1992).

5.7 Utilizing Natural Environmental Processes as a Remedy for Soil and Groundwater Contaminated with Chromium

Natural attenuation is a term that describes the naturally-occurring environ-mental processes that act without human intervention to reduce the mass,toxicity, mobility, volume, or concentration of contaminants. These processescan be grouped into two classes, destructive and nondestructive processes.Destructive processes include biotransformation and abiotic chemical reac-tions. Nondestructive processes include sorption (sorption and absorption),dispersion, dilution from recharge, and volatilization. Natural attenuation



is sometimes referred to by several other names, such as intrinsic remedia-tion, intrinsic bioremediation, natural restoration, or passive bioremediation.For the purposes of this chapter, the term natural attenuation will be used,because some of the synonyms used such as intrinsic bioremediation actuallyrefer to only one of many processes responsible for natural attenuation.

The implementation of natural attenuation processes as a remedy for soiland groundwater contamination is termed Monitored Natural Attenuation(MNA). The United States Environmental Protection Agency (USEPA) hasstated its position on the use of MNA for the remediation of contaminatedsoil and groundwater in their Final OSWER Directive “Use of MonitoredNatural Attenuation at Superfund, RCRA Corrective Action, and Under-ground Storage Tank Sites” (OSWER Directive Number 9200.4–17P), datedApril 21, 1999. The USEPA defines MNA as the reliance on natural attenu-ation processes (within the context of a carefully controlled and monitoredclean-up approach) to achieve site-specific remediation objectives within atime frame that is reasonable compared to that offered by other more activemethods. MNA is generally not seen as a viable stand-alone remedy, but ismore commonly viewed as a possible component of an overall remedialstrategy for a contaminated site. Nonetheless, MNA is increasing beingviewed as a viable alternative for the management of contaminated sites inthe U.S. and other countries.

Palmer and Puls (1994) present an outline of how the natural attenuationof Cr(VI) in the environment, especially in groundwater, can be evaluatedand how MNA can be implemented as a remedy for contamination. Theirwork is summarized here. Also Ellis et al. (2002) and Blowes (2002) givesummaries of the use of stable isotopes of Cr as an evaluation methodologyfor the implementation of MNA, and their work is also summarized herein.

Minerals containing reduced forms of iron and sulfur are abundant inmany aquifers, and these minerals reduce Cr(VI) to Cr(III) and promote theprecipitation of insoluble solids such as Cr(OH)3. Also, organic carbon-richmaterials can also reduce Cr(VI). In aquifers where reduced sediments areabundant and the concentrations of Cr(VI) are low, the attenuation capacityof the aquifer may be sufficient to prevent chromium migration. This mech-anism of the natural attenuation of Cr(VI) could be used as a remedialstrategy. The appropriateness of MNA depends on the groundwater flowsystem, the rate of chromium migration, and the ability of the aquifer mate-rials to reduce Cr(VI). It can be difficult, however, to distinguish measuredchromate decreases caused by attenuation reactions from those caused bymixing or dispersion (a dilution effect) in the aquifer.

However, for such a strategy to be adopted by a regulatory agency, thelikelihood that natural attenuation is likely to occur under the specific con-ditions at the site being investigated will most likely have to be demon-strated. There is no single test that can tell us if natural attenuation of Cr(VI)will occur at a particular site. Several tests are briefly described which havebeen utilized to address the factors affecting Cr(VI) transport in the subsur-face and describe how the results can be utilized in determining the potential



for the natural attenuation of Cr(VI) in the subsurface. Such a demonstrationwill likely comprise at least the following three items:

1. There are natural reductants in the aquifer2. The amount of Cr(VI) and other reactive constituents do not exceed

the reductive capacity of the aquifer3. The rate of Cr(VI) reduction to the target concentration is greater

than the rate of transport of Cr(VI) from source to point of compliance

Some of these criteria are relatively simple while others require additionaltests and interpretation. Each are discussed below.

5.7.1 There Are Natural Reductants in the Aquifer

In principle, the natural attenuation of Cr(VI) in the subsurface is feasibleas a result of interaction with naturally existing reductants. There are severalnatural reductants that can transform Cr(VI) to Cr(III). Potential reductantsof Cr(VI) include

• Aqueous species• Sorbed ions• Mineral constituents• Organic matter

During the migration of a Cr(VI) plume in groundwater, there is littlemixing of the waters containing the reducing agents and the Cr(VI). Whatmixing does occur will be driven by molecular diffusion at the front or edgesof the plume, and diffusion from lower permeability lenses containing rel-atively immobile water. Thus, reductants that are primarily dissolved ingroundwater such as Fe2+ are not going to be important in reducing Cr(VI).The mixing of reductants and Cr(VI) is instead going to occur primarilythrough the interactions of the Cr(VI) plume and the immobile soil matrix.Such interactions include

• Desorption of reductants such as Fe2+ from mineral surfaces• Direct and indirect surface redox reactions between Cr(VI) and the

mineral surfaces• Reduction by soil organic matter

The presence of Cr(III) may indicate either active reduction in the soil orneutralization of acidic waters containing Cr(III) with subsequent precipita-tion of chromium hydroxides. Therefore, specific reductants in the aquifershould be identified, and this can be fairly straightforward. Reductants (suchas pyrite, FeS2) can be readily identifiable by visual characteristics or by



standard petrographic techniques (i.e., powder x-ray diffraction, scanningelectron microscopy, polarized light microscopy). Also, testing for organiccarbon can provide a measure of the amount of carbon available for reduc-tion of Cr(VI).

Knowledge of the specific reductant within the aquifer can be useful indetermining the time scale for the reduction of Cr(VI). This is discussed later.Based on literature studies, soils containing iron sulfides or organic matterare more likely to reduce Cr(VI) on the time scales of interest than soilscontaining iron(II) silicates.

5.7.2 The Amount of Cr(VI) and Other Reactive ConstituentsDo Not Exceed the Reductive Capacity of the Aquifer

Studies demonstrate that groundwater contributes less than 1% of theoxidation capacities (equivalents of chromium oxidized per gram of soil)and reduction capacities (equivalents of chromium reduced per gram ofsoil) of aquifer systems while the soil matrix contributes the remainingfraction (Barcelona and Helm, 1991). Thus, any discussion of Cr(VI) reduc-tion or Cr(III) oxidation in the subsurface must focus on the soil matrix.Several soil tests can be useful in determining the mass of Cr(VI) andCr(III) and also the reduction and oxidation capacities of the aquifermaterials.

5.7.2.1 Mass of Cr(VI)

It must be demonstrated that the amount of Cr(VI) in the aquifer does notexceed the capacity of the soil for reducing Cr(VI) to Cr(III). Therefore animportant step in evaluating the potential for natural attenuation is to deter-mine the mass of Cr(VI) in the aquifer. Aqueous samples are most oftenobtained from monitoring wells, or water can separated from the soil matrixeither by centrifugation or by squeezing. The pH of the water should bemeasured to determine if it is within the proper range (5.5 to 12) to insurethe Cr(III) concentrations are less than 1 µ mol/l (0.05 mg/l).

Cr(VI) associated with the soil matrix may be sorbed to mineral surfaces(particularly iron oxides) or precipitated as chromate minerals. There is noaccurate method for determining each of these fractions of Cr(VI); but,sequential extractions have been used, where an initial water extraction servesto remove remaining pore water and dissolve highly soluble chromium min-erals present in the soil or that may have precipitated due to evaporationduring sample handling. Then, a phosphate extraction is used as a measureof the “exchangeable” chromate in the soil (Bartlett and James, 1988); thewater is separated from the slurry and Cr(VI) is measured by the DPC Method(Bartlett and Kimble, 1976; Bartlett and James, 1988). The increase in thechromate concentrations is the amount of “exchangeable” chromate. Thephosphate removes chromate by both directly competing for the sorptionsites in the soil and indirectly (in some cases) by increasing the pH. Palmer



and Puls (1994) report that at many sites, the total Cr(VI) associated with thesoil matrix is the sum of BaCrO4 and the phosphate-extractable Cr(VI). Thissum is then added to the aqueous Cr(VI) to get the total concentration ofCr(VI) in the soil (Cr(VI)tot).

5.7.2.2 Mass of Cr(III)

The presence of Cr(III) in the soil may demonstrate that Cr(VI) reduction isoccurring. Thus, the mass of Cr(III) in the soil could provide a measure ofthe amount of reduction that has occurred. The total amount of Cr(III)present in the soil is the sum of the mass in solution as well as mass associatedwith the solid phase. Total chromium in solution can be determined byatomic absorption spectroscopy (AAS) or inductively coupled plasma spec-troscopy (ICP). When total chromium is statistically greater that Cr(VI),Cr(III) can be simply determined by difference.

5.7.2.3 Reduction Capacity of the Aquifer

For the natural attenuation of Cr(VI) to be demonstrated, the soil mustpossess sufficient reducing capacity (RC) to reduce all the Cr(VI) in the sourcearea. Thus, the total mass of Cr(VI) from the source (MO) must be less thanthe total mass of Cr(VI) (MTOT) that can be reduced by the aquifer materialbetween the source and a point of compliance where XC is distance betweenthem. Where A is the cross-sectional area of the plume normal to the directionof groundwater flow, and ρb is the dry bulk density of the aquifer, thisrequirement can be expressed as

MO < MTOT = XCAρbRC (5.8)

As the distance XC increases, the mass of Cr(VI) that can be reduced alsoincreases. A key difficulty in applying this criterion is in providing a reason-able estimate of the total mass of Cr(VI) in the source area (MO).

The total Cr(VI) reducing capacity can be obtained using the classicalWalkley-Black method for determining soil organic carbon (Bartlett andJames, 1988; Walkley and Black, 1934). Although this method has its limita-tions (Nelson and Sommers, 1982), it is a direct measure of how much Cr(VI)can be reduced by a soil at extreme acid conditions. Variations on this methodhave been described (Barcelona and Helm, 1991; Nelson and Sommers, 1982).The extreme conditions of pH and temperature used in the total Cr(VI)reducing capacity test may yield a greater reducing capacity than would beavailable under most environmental conditions. The “available reducingcapacity” test of Bartlett and James (1988) is designed to determine thereducing capacity at pH values more likely to be encountered in the field.However in long-term reduction tests at near neutral pH, reduction has beenobserved to be occurring after 250 days. Such long-term reduction tests arenot practical at most waste sites.



5.7.2.4 Oxidation Capacity of the Aquifer

As part of a geochemical cycle, both oxidation and reduction of Cr areoccurring simultaneously within the subsurface. As the Cr(III) is oxidizedto Cr(VI) by manganese dioxides in the soil, Cr(VI) can be reduced to Cr(III)by some reductant such as soil organic carbon or pyrite. Ultimately, asteady state situation will be reached where the rate of loss of Cr(VI) viareduction is balanced by the rate of production by the oxidation of Cr(III).Under certain conditions, oxidation of Cr(III) may be favored over thereduction of Cr(VI). For example, soil containing Cr(III) formed by thereduction of Cr(VI) may become a source of Cr(VI). Therefore, a potentiallimitation to the use of natural attenuation of Cr(VI) as a remedial optionis the oxidation of the Cr(III) to Cr(VI) by oxidants such as MnO2. If theoxidizing capacity of the soil is greater than the reduction capacity, thenas the chromium is cycled in the soil, the reductants could be exhausted,the Cr(III) could oxidized, and ultimately, Cr(VI) could be mobilized in thesoil. It is therefore important to determine the capacity of the aquifer tooxidize Cr(III).

Bartlett and James (1979, 1988) suggest a simple test for the amount ofCr(III) that can be oxidized by a soil. Barcelona and Helm (1991) give amethod to measure the oxidation capacity of soils. Each of these methodshas some problems. If the sole Cr(III) oxidation mechanism is by manganeseoxides, then using extraction methods specifically designed for this purposemay be useful (Chao, 1972; Gambrell and Patrick, 1982).

Fendorf and Zasoski (1992) suggest that CrOH2+ is the reactive species inthe oxidation of Cr(III) by MnO2. The key point to consider when consideringnatural attenuation in soils that contain both a reductant and MnO2 is thatas long as the supply of reductant and MnO2 have not been significantlydepleted, (the HCrO4

– concentration) does not converge to zero with increas-ing residence time within the aquifer as one would expect for a first-orderreaction that only considers reduction of Cr(VI). Rather, HCrO4

– convergesto some steady-state concentration that is >0 that may or may not be abovethe MCL. This steady-state concentration increases with a preference of oxi-dation over reduction and it varies with pH.

5.7.3 The Rate of Cr(VI) Reduction to the Target Concentration Compared to the Rate of Transport of Cr(VI) from Sourceto Point of Compliance

In principle, if the rate equations are correct and all of the parameters areknown, one could calculate the steady-state Cr(VI) concentration and deter-mine if natural attenuation could achieve compliance goals. The major lim-itation to this approach is the lack of information about the rate of oxidationand reduction of chromium under conditions likely to be encountered byplumes emanating from chromium sources. Without better information



about these rate processes under a wider range of conditions with respectto pH, the use of the natural attenuation option for contaminated soils willcontinue to be a highly debated issue.

5.7.3.1 Rates of Oxidation and Reduction

If natural attenuation is to be a viable option, the time for the reductionreaction to decrease the concentration from its initial concentration (CO) tosome target concentration (CS), such as a drinking water standard (i.e., MCL),should be less than the residence time of the contaminated water in theportion of the aquifer between the source of the Cr(VI) and the point ofcompliance. Difficulties in utilizing this criterion as mentioned above arisein applying the appropriate rate equation and obtaining the pertinent ratecoefficients. Rates can be obtained from the technical literature, but one mustuse reduction rates based on materials that are most likely controlling theCr(VI) reduction at the site under evaluation. In addition, because the ratesare concentration-dependent, and are related to the specific reductant andthe pH level, it is important to obtain rate coefficients that were acquiredunder conditions similar to the site. As far as reductants are concerned,literature studies indicate soils containing iron sulfides or organic matter aremore likely to reduce Cr(VI) on the time scales of interest than soils contain-ing iron(II) silicates.

One method of obtaining the net rate of reduction is through tests onuncontaminated soils (background soil) at the site that are similar to thosethrough within the contaminant plume. Cr(VI) can be added to the back-ground soil in the laboratory and the Cr(VI) concentrations monitored overtime. The reaction vessels must exclude light to prevent photoreductionreactions and the slurry must have the same pH as the contaminant plume.A key limitation to such experiments is that they require several months toa year to complete.

5.7.3.2 Estimating Reduction from Monitoring Well Data

In principle, Cr(VI) reduction can be estimated from the decrease in the massof Cr(VI) in the aquifer (Henderson, 1994). The key difficulty in such anapproach is to estimate the mass of Cr(VI) using aqueous concentrations.The total mass of Cr(VI) in the aquifer is the sum of the mass that is insolution, the mass that is sorbed to the aquifer matrix, and the mass that isprecipitated within the aquifer. The mass of Cr(VI) in solution (Maq) isobtained by integrating the Cr(VI) concentrations over the volume of thecontaminated aquifer,

Maq = θvCV (5.9)

where V is the volume of aquifer containing a plume with a Cr(VI) concen-tration of C. (and θv represents the porosity). The mass of Cr(VI) sorbed to



the soil matrix (Mads) can be computed from an sorption isotherm. Thereis no unique amount of Cr(VI) precipitate for a given Cr(VI) concentration.Therefore it is impossible to estimate mass of this fraction of Cr(VI) in thesubsurface using only the measured concentrations in monitoring wells.Thus, natural attenuation of Cr(VI) from mass balances using monitoringwell data can only be used when it can be reasonably demonstrated theCr(VI) precipitates cannot form within the aquifer. Even when it is dem-onstrated that the formation of precipitates within the aquifer is unlikely,there are inherent problems with any monitoring system, creating uncer-tainties in the estimated mass of Cr(VI) during a sampling round.

5.7.3.3 Monitoring Reduction Via Stable Isotopes of Chromium

Variations in the isotopic ratios of light elements are sensitive indicators ofchemical processes that occur in natural systems. For example, the 34S/32Sratio in dissolved sulfate increases when bacteria reduce sulfate to sulfide.Reduction reactions tend to enrich products in the lighter isotopes becausethey preferentially react (Hoefs, 1987), and the residual reactants becomeprogressively enriched in the heavier isotopes as reduction proceeds (Boettcheret al., 1990; Thode and Monster, 1965; and Johnson et al., 1999). Ellis et al.(2002) now show that the 53Cr/52Cr ratio also changes during reduction ofCr(VI) to Cr(III). They show that abiotic reduction of Cr(VI) resulting fromreaction with the mineral magnetite, estuarine sediments, and freshwatersediments leads to a consistent 53Cr/52Cr shift. This observation indicatesthat 53Cr/52Cr ratios increase systematically with progressive Cr(VI) reduc-tion in groundwater.

The measurements of chromium isotope ratios provides a new tool forevaluating the extent and rate of the natural attenuation of chromate. Rela-tive to conventional methods, Blowes (2002) states that the use of isotopeshas two advantages: (1) Each 53Cr/52Cr determination provides a measure ofthe amount of reduction that has already occurred, and there is thus no needto see whether Cr(VI) mass decreases over time; and (2) the measurement ofreduction integrates spatially over a flow path, whereas analyses of aquifersolids give information on a much smaller spatial scale.

53Cr/52Cr measurements can also help to evaluate in situ approaches ofremediation, which have been developed for site where natural attenuationis insufficient to prevent chromate migration (i.e., permeable reactive barri-ers, injection of chemical reactants to reduce chromate, and injection of areductant to react with the aquifer materials to form reduced minerals in theaquifer). The new finding by Ellis et al. (2002) that 53Cr/52Cr ratios revealthe extent of abiotic reduction suggests that 53Cr/52Cr measurements canassist in the evaluation of the effectiveness of all these approaches to chro-mate reduction in groundwater.

Chromium has four stable isotopes of masses 50 (4.35%), 52 (83.8%),53 (9.50%), and 54 (2.37%) (Rotaru and Birck, 1992; Handbook of Chemistry



and Physics, 1996). Ellis et al. (2002) measured chromium isotope fraction-ation during reduction of Cr(VI) by slurries of magnetite and two sedimentsamples. Because magnetite is a likely reducing agent in some aquifer sed-iments (Anderson et al., 1994), the magnetite experiment provided a simpleanalog for a natural aquifer. (In these experiments) at pH 6 to 7, sorption ofCr(VI) was negligible. As reduction proceeded and the Cr(VI) concentrationsdecreased, δ 53 Cr (defined in Chapter 2) values of the remaining unreducedCr(VI) increased, indicating preferential reduction of the lighter isotopes.These experiments showed that reduction of Cr(VI) results in Cr stable-isotope fractionation.

To estimate initial isotopic compositions of the contaminants, Ellis et al.(2002) measured δ 53Cr values in samples of plating baths in active use atdifferent sites, the chromic acid supply used to make up plating baths,laboratory chromium reagents, and basaltic rock standards. All of thesesamples yielded δ 53Cr values close to 0%. Therefore, Ellis et al. (2002)suggested that Cr released as plating waste generally has an initial δ 53Crvalue slightly greater than zero. If so, detection of Cr(VI) reduction ingroundwater systems would be relatively simple, as the initial δ 53Cr valuewould be known and groundwater values greater than that would directlyindicate the extent of reduction. If, on the other hand, plating wastes havevariable δ 53Cr values, then it may be possible to distinguish differentcontamination sources via their δ 53Cr values. In studies of groundwatercontamination sites, Ellis et al. (2002) showed that all of the groundwaterCr(VI) analyses indicated enrichment in the heavy isotope relative to theplating baths, and Cr(VI) reduction had preferentially removed lighterisotopes from the groundwater. The variation in δ 53Cr values at each ofthe sites suggested that reduction of Cr(VI) was occurring and had pro-gressed to different degrees in different parts of the contaminant plumes.The highest δ 53Cr values were found in samples with lowest Cr(VI) con-centration at (the sites studied). This result was to be expected becausethe fringe areas of the contaminant plumes likely would have greaterdegrees of reduction than the plume cores, where chromium concentra-tions are high and the reducing power of the aquifer materials had beendepleted.

Stable Cr isotope ratios can thus serve as indicators of the extent of Cr(VI)reduction in groundwater. Cr(VI) reduction by bacteria or reducing agentsother than those studied by Ellis et al. (2002) could induce greater or lesserisotopic fractionation than (they) observed. And, processes other than reduc-tion (i.e., sorption, precipitation, and uptake by plants and algae) could alsobe responsible for removing chromium from solution (James and Bartlett,1983, 1984, and 1988). If such processes and/or Cr(III) oxidation also induceisotopic fractionation, this could complicate the interpretation of δ 53Cr mea-surements. However, as with S and Se isotopes (Johnson et al., 1999) it couldbe expected that the dominant cause of chromium isotope fractionation isreduction.



Bibliography

Abu-Saba, K.E., Sedlak, D.L., and Flegal, A.R., 2000, Mar. Chem., vol. 69, pp. 33–41.Achterberg, E.P., van den Berg, C.M.G., Boussemart, M., and Davison, W., 1997,

Speciation and cycling of trace metals in Esthwaite Water: a productiveEnglish lake with seasonal deep-water anoxia, Geochim. Cosmochim. Acta,vol. 61, pp. 5233–5253.

Agency for Toxic Substances and Disease Registry (ATSDR), 1992, Toxicological Profilefor Chromium, Draft, U.S. Dept. of Health and Human Services, Washington,D.C.

Ainsworth, C.C., Girvin, D.C., Zachara, J.M., and Smith, S.C., 1989, Chromate ad-sorption on goethite: effects of aluminum substitution, Soil Sci. Soc. Am. J.,vol. 53, pp. 411–418.

Amacher, M.C. and Baker, D.E., 1982, Redox reactions involving chromium, plutoniumand manganese in soils, DOE/DP/04515-1, Institute for Research on Land andWater Resources, Pennsylvania State University and U.S. Department of Ener-gy, Las Vegas, NV, 166 p.

Amacher, M.C., Kotuby-Amacher, J., Selim, H.M., and Iskandar, I.K., 1986, Retentionand release of metals by soils, evaluation of several models, Geoderma, vol. 38,pp. 131–154.

Amacher, M.C., Selim., H.M., and Iskandar, I.K., 1988, Kinetics of chromium(VI) andcadmium retention in soils: a non-linear multireaction model, Soil Sci. Soc. Am.,vol. 52, pp. 398–408.

Anderson, L.D., Kent, D.B., and Davis, J.A., 1994, Environ. Sci. Technol., vol. 28, pp. 178.Anderson, R.A., Polansky, M.M., Brantner, J.H., and Roginski, E.E., 1977, Chemical

and biological properties of biologically active chromium, in Int. Symp. TraceElement Metabolism in Man and Animals, Friesing, Federal Republic of Germany,pp. 269–271.

Barcelona, M.J. and Helm, T.J., 1991, Oxidation-reduction capacities of aquifer solids,Environ. Sci. Technol., vol. 25, pp. 1565–1572.

Baron, D. Palmer, C.D., and Stanley, J.T., 1996, Identification of two iron-chromateprecipitates in a Cr(VI)-contaminated soil, Environ. Sci. Technol., vol. 30, pp.964–968.

Bartlett, R.J. and James, B., 1979, Behavior of chromium in soils: III: Oxidation, J.Environ. Qual., vol. 8, pp. 31–35.

Bartlett, R.J. and James, B.R., 1988, Mobility and bioavailability of chromium in soils,in Chromium in the Natural and Human Environments, Nriagu, J.O. and Nieboer,E., Eds., John Wiley and Sons, New York, NY, pp. 267–304.

Bartlett, R.J. and Kimble, J.M., 1976, Behavior of chromium in soils, I. Trivalent forms,II. Hexavalent forms, J. Environ. Qual., vol. 5, pp. 379–386.

Bartlett, R.J., 1991, Chromium cycling in soils and water: links, gaps and methods,Environ. Health Prespective, vol. 92, pp. 17–24.

Beaubien, S., Nriagu, J., Blowes, D., and Lawson, G., 1994, Chromium speciation anddistribution in the Great Lakes, Environ. Sci. Technol., vol. 28, pp. 730–738.

Benjamin, M.M. and Bloom, N.S., 1981, Effect of strong binding of anionic adsor-bates on adsorption of trace metals on amorphous iron oxyhydroxide, inAdsorption from Aqueous Solution, Tewari, P.H., Ed., Plenum Press, New York,NY, pp. 41–60.



Bloomfield, C. and Pruden, G., 1980, The behavior of Cr(VI) in soil under aerobicand anaerobic conditions, Environ. Pollut. (Series A), vol. 23, pp. 103–114.

Blowes, D.W. and Ptacek, C.J., 1992, Geochemical remediation of groundwater bypermeable reactive walls: Removal of chromate by reaction with iron-bearingsolids, Proc. Subsurface Restoration Conference, June 21–24, 1992, Dallas, TX,pp. 214–216.

Blowes, D., 2002, Tracking Hexavalent Cr in Groundwater, Science, vol. 295.Bodek, I., Lyman, W.J., Reehl, W.F., and Rosenblatt, D.H., 1988, Environmental Inor-

ganic Chemistry Properties, Processes and Estimation Methods, Bodek, I. Ed., Per-gamon Press, New York, NY.

Boettcher, J., Strebel, O., Voerkelius, S., and Schmidt, H.L., 1990, J. Hydrol., vol. 114,pp. 413.

Calder, L.M., 1988, Chromium contamination of groundwater, in Chromium in theNatural and Human Environments, Nriagu, J.O. and Nieboer, E., Eds., John Wileyand Sons, New York, pp. 215–231.

Campos, J., Martinez-Pacheco, M., and Cervantes, C., 1995, Hexavalent-chromiumreduction by chromate-resistant Bacillus sp. Strain, Antione von Leeuwenhoek,vol. 68, pp. 203–208.

Cary, E.E., Allaway, W.H., and Olson, O.E., 1977, Control of chromium concentrationin food plants, II. Chemistry of chromium in soils and its availability to plants,J. Agric. Food Chem., vol. 25, pp. 305–309.

Cary, E.E., Ed., 1982, Chromium in Air, Soil, and Natural Waters, Elsevier Biomedical,New York, NY, pp. 49–63.

Cavallaro, N. and McBride, M.B., 1978, Copper and cadmium adsorption charac-teristics of selected acid and calcareous soils, Soil Sci. Soc. Am. J., vol. 42,pp. 550–556.

Chao, T.T., 1972, Selective dissolution of manganese oxides from soils and sedi-ments with acidified hydroxylamine hydrochloride, Soil Sci. Amer. Proc.,vol. 36, pp. 764–768.

Chuecas, D.L. and Riley, J.P., 1966, The spectrophotometric determination of chromi-um in sea water, Analyt. Chim. Acta, vol. 35, pp. 240–246.

Clay, R.E., 1992, Multimedia Environmental Distribution of Gaseous, Dissolved and parti-cle-Bound Pollutant, Master of Science Thesis, Chemical Engineering Depart-ment, University of California Los Angeles (UCLA).

Coates, J.D., Phillips, E.J.P., Lonergan, D.J., Jenter, H., and Lovely, D.R., 1996, Appl.Environ. Microbiol., vol. 62, pp. 1531.

Coleman, R.N., 1988, Chromium toxicity: effects on microorganisms with specialreference to the soil matrix, in Chromium in the Natural and Human Environ-ments, Nriagu, J.O. and Nieboer, E., Eds., John Wiley and Sons, New York, NY,pp. 335–368.

Cranston, R.E. and Murray, J.W., 1980, Chromium species in the Columbia Riverestuary, Limnol. Oceanogr., vol. 25, pp. 1104–1122.

Cranston, R.E. and Murray, J.W., 1978, Determination of chromium species in naturalwaters, Analyt. Chim. Acta, vol. 99, pp. 275–282.

Davis, A. and Olsen, R.L., 1995, The Geochemistry of chromium migration andremediation in the subsurface, Ground Water, vol. 33, pp. 759–768.

Davis, J.A. and Leckie, J.O., 1980, Surface ionization and complexation at theoxide/water interface: III. Adsorption of anions, J. Colloid Interface Sci.,vol. 74, pp. 32–43.



DeLeo, P.C. and Ehrlich, H.L., 1994, Reduction of hexavalent chromium by pseudomonasfluorescens LB300 in batch and continuous cultures, Appl. Microbiol. Biotechnol.,vol. 40, pp. 756–759.

Deng, B. and Stone, A.T., 1996, Surface-catalyzed chromium(VI) reduction: theTiO2

−Cr(VI)-mandelic acid system, Environ. Sci. Technol., vol. 30, pp. 463–472.Deutsch, M., 1972, Incidents of chromium contamination of groundwaters in Mich-

igan, in Water Quality in a Stressed Environment, Pettyjohn, W.A., Eds., BurgessPublishing Co., Minneapolis, MN, pp. 149–159.

Deverel, S.J. and Millard, S.P., 1988, Distribution and mobility of selenium and othertrace elements in shallow groundwater of the Western San Joaquin Valley,California, Envir. Sci. Technol., vol. 22, pp. 697–702.

Dreiss, S.J., 1986, Chromium migration through sludge treated soils, Ground Water,vol. 24, pp. 312–321.

Dudley, L.M., McLean, J.E., Furst, T.H., and Jurinak, J.J., 1991, Sorption of Cd andCu from an acid mine waste extract by two calcareous soils: column studies,Soil Sci., vol. 151, pp. 121–135.

Dudley, L.M., McLean, J.E., Sims, R.C., and Jurinak, J.J., 1988, Sorption of copper andcadmium from the water-soluble fraction of an acid mine waste by two calcar-eous soils, Soil Sci., vol. 145, pp. 207–214.

Eary, L.E. and Rai, D., 1991, Chromate reduction by subsurface soils under acidicconditions, Soil Sci. Soc. Am. J., vol. 55, pp. 676–683.

Eary, L.E. and Rai, D., 1988, Chromate removal from aqueous wastes by reductionwith ferrous iron, Environ. Sci. Technol., vol. 22, pp. 972–977.

Eary, L.E. and Rai, D., 1987, Kinetics of chromium (III) oxidation to chromium(VI)by reaction with manganese dioxide, Environ. Sci. Technol., vol. 21, pp.1187–1193.

Eary, L.E. and Rai, D., 1989, Kinetics of chromate reduction by ferrous ions derivedfrom hematite and biotite at 25°C, Am. J. Sci., vol. 289, pp. 180–213.

Eckert, J.M., Stewart, J.J., Waite, T.D., Szymezek, R., and Williams, K.L., 1990, Anal.Chim. Acta, vol. 236, pp. 357–362.

Ellis, A.S., Johnson, T.M., and Bullen, T.D., 2002, Chromium isotopes and the fate ofhexavalent chromium in the environment, Science, vol. 295, pp. 2060–2062.

Essen, J. and El Bassam, N., 1981, On the movility of cadmium under aerobic soilconditions, Environ. Pollut. Ser. A., pp. 15–31.

Fendorf, S., Wielinga, B.W., and Hansel, C.M., 2001, Int. Geol., vol. 42, pp. 691.Fendorf, S.E. and Li, G., 1996, Kinetics of chromate reduction by ferrous iron, Environ.

Sci. Technol., vol. 30, pp. 1614–1617.Fendorf, S.E. and Zasoski, R.J., 1992, Chromium(III) oxidation by δ-MnO2, 1. Char-

acterization, Environ. Sci. Technol., vol. 26, pp. 79–85.French, W.B., Gallagher, M., Passero, R.N., and Straw, W.T., 1985, Hydrogeologic

investigation for remedial action related to a chromium-arsenic-copper dis-charge to soil and groundwater, in Innovations in Water and Wastewater Fields,Glysson E.A., Swan D.E., and Way E.J., Eds., Ann Arbor Press, Ann Arbor, MI,pp. 209–229.

Friess, S.L., 1989, Carcinogenic risk assessment criteria associated with inhalation ofairborne particulates containing chromium(VI/III), Sci. Tot. Environ., vol. 86,pp. 109–112.

Frissel, M.J., Poelstra, P., and Reisinger, P., 1975, The Behavior of Chromium in Aquaticand Terrestrial Food Chains: EUR 5375e, Boite postale 1003, Luxembourg, pp. 27–42.

Au: Pls. provide chapter title forreference




Gambrell, R.P. and Patrick, W.H., 1982, Manganese, in Methods of Soil Analysis, Part2, Chemical and Microbiological Properties, Page, A. L., Ed., 2nd ed., AgronomyNo. 9, Part 2, American Society of Agronomy, Soil Science Society of America,Madison, WI, pp. 313–322.

Gerritsee, R.G., Vriesema, R., Dalenberg, J.W., and DeRoos, H.P., 1982, Effect ofsewage sludge on trace element mobility in soils, J. Envr. Qual., vol. 11,pp. 359–364.

Griffin, R.A. and Shimp, N.F., 1978, Attenuation of pollutants in municipal landfill leachateby clay minerals, EPA/600/2–78/157, U.S. Environmental Protection Agency,Washington, D.C.

Griffin, R.A., Au, H.K., and Frost, R.R., 1977, Effect of pH on adsorption ofchromium from landfill leachate by clay minerals, J. Envr. Sci. Hlth. Ser. A.,vol. 12, pp. 431–449.

Grohse, P.M., Gutknecht, W.F., Hodson, L., and Wilson, B.M., 1988, The fate of hexava-lent chromium in the atmosphere: Research Triangle Institute Report on CARBContract #A6–096–32, Report Number ARB/R–89/379, California Air ResourcesBoard, Sacramento, CA.

Grove, D.B. and Stollenwerk, K.G., 1985, Modeling the rate-controlled sorption ofhexavalent chromium, Water Resources Res., vol. 21, pp. 1703–1709.

Grove, J.H. and Ellis, B.G., 1980, Extractable chromium as related to soil pH andapplied chromium, Soil Sci. Soc. Am. J., vol. 44, pp. 238–242.

Handa, B.K., 1988, Occurrence and distribution of chromium in natural waters ofIndia, in Chromium in Natural and Human Environments, Nriagu, J.O. andNieboer, E., Eds., Wiley Interscience, New York, NY, pp. 189–215.

Handbook of Chemistry and Physics, 1996, 77th ed., CRC Press.Hartford, W., 1983 Chromium chemicals, in Kirk-Othmer Encyclopaedia of Chemical

Technology, Grayson, M., Ed., 3rd ed., vol. 6, Wiley-Interscience, New York, NY,pp. 83–120.

Hathway, J.A., 1989, Role of Epidemiologic studies in evaluating the carcinogenicityof chromium compounds, Sci. Total Environ., vol. 86, pp. 169–179.

Hem, J.D., 1977, Reactions of metal ions at surfaces of hydrous iron oxide, Geochim.Cosmochim. Acta, vol. 41, pp. 527–538.

Henderson, T., 1994, Geochemical reduction of hexavalent chromium in the TrinitySand, Ground Water, vol. 32, pp. 477–486.

Hoefs, J., 1987, Stable Isotope Geochemistry, 3rd ed., Springer-Verlag, New York, NY.Hug, S.J., Laubscher, H.U., and James, B.R., 1997, Environ. Sci. Technol., vol. 31,

pp. 160–170.James, B.R. and Bartlett, R.J., 1983, Behavior of chromium in soils: V. Fate of organ-

ically complexed Cr(III) added to soil; VI. Interaction between oxidation-re-duction and organic complexation; VII. Adsorption and reduction ofhexavalent forms, J. Environ. Qual., vol. 12, pp. 169–181.

James, B.R. and Bartlett, R.J., 1988, Mobility and bioavailability of chromium in soil, inChromium in Natural and Human Environments, Nriagu, J.O. and Nieboer, E.,Eds., Wiley Interscience, New York, NY, pp. 265–305.

James, B.R. and Bartlett, R.J., 1984, Plant-soil interactions of chromium, J. Environ.Qual., vol. 13, pp. 67–70.

James, B.R., 1994, Hexavalent chromium solubility and reduction in alkaline soilsenriched with chromite ore processing residue, J. Environ. Qual., vol. 23,pp. 227–233.

Au: Pls. provide chaptertitle forreference



James, B.R., 1996, The challenge of remediation chromium-contaminated soil, Environ.Sci. Technol., vol. 30, pp. 248–251.

Jenne, 1968, Control of Mn, Fe, Co, Ni, Cu, and Zn concentrations in soils andwater–the dominant role of hydrous manganese and iron oxides, Adv. In Chem.,vol. 7, pp. 337–387.

Johnson, C.A. and Xyla, A.G., 1991, The oxidation of chromium(III) to chromium(VI)on the surface of manganite (Gamma-MnOOH), Geochim. Cosmochim., vol. 55,pp. 2861–2866.

Johnson, T.M., Herbel, M.J., Bullen, T.D., and Zawislanski, P.T., 1999, Geochim. Cos-mochim. Acta, vol. 63, pp. 2775.

Jurinak, J.J. and Bauer, N., 1956, Thermodymanics of zinc adsorption on calcite, colo-mite and magnesite-type minerals, Soil Sci. Soc. Am. Proc., vol. 20, pp. 466–471.

Kaczynski, S.E. and Kieber, R.J., 1993, Aqueous trivalent chromium photoproductionin natural waters, Environ. Sci. Technol., vol. 27, pp. 1572–1576.

Kaczynski, S.E. and Kieber, R.J., 1994, Hydrophobic C18 bound organic complexesof chromium and their potential impact on the geochemistry of chromium innatural waters, Environ. Sci. Technol., vol. 28, pp. 799–804.

Kent, D.B., Davis, J.A., Maest, A.S., and Rea, B.A., 1989, Field and laboratory studiesof transport of reactive solutes in groundwater, in Proc. 6th Int. Conf. Water-RockInteraction, Miles, D.L., Ed., Miles, Balkema, Rotterdam, Netherlands, pp.381–383.

Kieber, R.J. and Heiz, G.R., 1992, Indirect photoreduction of aqueous chromium (VI),Environ. Sci., Technol., vol. 26, pp. 307–312.

Kimbrough, D.E., Cohen, Y., Winer, A.M., Creelman, L., and Mabuni, C., 1999, Acritical assessment of chromium in the environment, in Critical Reviews, Envi-ronmental Science and Technology, vol. 29, pp. 1–46.

Korte, N.E., Skopp, J., Fuller, W.H., Niebla, E.E., and Aleshii, B.A., 1976, Trace elementmovement in soils: influence of soil physical and chemical properties, Soil Sci.,vol. 122, pp. 350–359.

Kotas, J. and Stasicka, Z., 2000, Chromium occurrence in the environment and meth-ods of its speciation, Environ. Poll. vol. 107, pp. 263–283.

Lancy, L.E., 1966, Treatment of spent cooling waters, U.S. Patent No. 3,294,960.Lieber, M., Perlmutter, N.M., and Frauenthal, H.L., 1964, Cadmium and hexavalent

chromium in Nassau County ground water, J. AWWA, vol. 56, pp. 739–747.Lovely, D.R., 1993, Annu. Rev. Microbiol., vol. 47, pp. 263.Loyaux-Lawniczak, S., Lecomte, P., Ehrhardt, J., 2001, Behavior of hexavalent chro-

mium in a polluted groundwater: redox processes and immobilization in soils,Environ. Sci. Technol., vol. 35, pp. 1350–1357.

MacNaughton, M.G., 1975, Adsorption of chromium(VI) at the oxide-water interface,Interim Report: Envir. Chem. Div., Environics Directorate, Air Force Civil Eng.Center, Tyndal AFB, FL.

McBride, M.B. and Bouldin, D.R., 1984, Long-term reactions of copper(II) in a con-taminated calcareous soil, Soil Sci. Soc. Am. J., vol. 48, pp. 56–59.

McBride, M.B., 1980, Chemisorption of Cd2+ on calcite surfaces, Soil Sci. Soc. Am. J.,vol. 44, pp. 26–28.

McLean, J.E. and Bledsoe, B.E., 1992, Behavior of metals in soils, EPA/540/S-92/018.Murray, J.W., Spell, B., and Paul, B., 1983, The contrasting geochemistry of manganese

and chromium in the Eastern Tropical Pacific Ocean, in Trace Metals in Seawater,Wong, C.S., Boyle, E., Bruland, K., Burton, J.D., and Goldberg E.D. Eds., PlenumPress, New York, NY, pp. 643–670.

Au: Pls. provide chapter title




Music, S., Ristic, M., and Tonkovic, M., 1986, Sorption of chromium(VI) on hydrousiron oxides, Z. Wass. Abwass. Forsch., vol. 19, pp. 186–196.

Nakayama, E., Kuwamoto, T., Tsurubo, S., Tokoro, H., and Fujinaga, T., 1981, Chem-ical speciation of chromium in sea water: 1. Effect of naturally occurringorganic materials on the complex formation of chromium(III), Analyt. Chim.Acta, vol. 130, pp. 289–294.

Nelson, D.W. and Sommers, L.E., 1982, Total carbon, organic carbon, and organicmatter, in Methods of Soil Analysis, Part Two, Dinauer R.C., Ed., AmericanSociety of Agronomy, Inc. and Soil Science Society of America, Inc., Madison,WI, pp. 539–580.

Nriagu, J.O., 1988, Historical perspectives; and, Production and uses of chromi-um, in Chromium in the Natural and Human Environments, Nriagu, J.O. andNieboer, E., Eds., John Wiley and Sons, New York, NY, vol. 20, pp. 1–20and pp. 81–104.

Pacyna, J.M. and Nriagu, J.O., 1988, Atmospheric emissions of chromium from nat-ural and anthropogenic sources, in Chromium in Natural and Human Environ-ments, Nriagu, J.O. and Nieboer, E., Eds., Wiley Interscience, New York, NY,pp. 105–125.

Palmer, C.D. and Wittbrodt, P.R., 1990, Geochemical characterization of the UnitedChrome Products Site, Final Report, in Stage 2 Deep Aquifer Drilling TechnicalReport, United Chrome Products Site, Corvallis, Oregon, CH2M Hill, Corvallis,OR.

Palmer, C.D. and Wittbrodt, P.R., 1991, Processes affecting the remediation of chro-mium-contaminated sites, Environ. Health Perspectives, vol. 92, pp. 25–40.

Palmer, C.D. and Puls, R.W., 1994, Natural attenuation of hexavalent chromium ingroundwater and soils, Ground Water, EPA/540/S–94/505.

Paya Perez, A.B., Gotz, L., Kephalopoulos, S.D., and Bignoli, G., 1988, Sorption ofchromium species on soil, in Heavy Metals in the Hydrocycle, Astruc, M. andLester, J.N., Eds., Selper, London, pp. 59–66.

Perlmutter, N.M. and Lieber, M., 1970, Dispersal of plating wastes and sewage con-tamination in ground water and surface water, South Farmingdale MassapequaArea, Nassau County, New York, U.S.G.S. Water Supply Paper 1879-G.

Pettine, M. and Millero, F.J., 1990, Limnol. Oceanogr., vol. 35, pp. 730–736.Pettine, M. and Millero, F.J., 1991, Mar. Chem., vol. 34, pp. 29–46.Pettine, M., Campanella, L., Millero, F.J., 2002, Reduction of hexavalent chromium

by H2O2 in acidic solutions, Environ. Sci. Technol., vol. 36, n. 5.Pettine, M., D’Ottone, L., Campanella, L., Millero, F.J., and Passino, R., 1998, The

reduction of chromium(VI) by iron(II) in aqueous solutions, Geochim. Cosmo-chim. Acta, vol. 62, pp. 1509–1519.

Pettine, M., Millero, F.J., and La Noce, T., 1991, Chromium(III) interactions in seawaterthrough its oxidation kinetics, Mar. Chem., vol. 34, pp. 29–46.

Pettine, M., Millero, F.J., and Passino, R., 1994, Mar. Chem., vol. 46, pp. 335–344.Powell, R.M., Puls, R.W., Hightower, S.K., and Sabatini, D.A., 1995, Coupled iron

corrosion and chromate reduction: mechanism for subsurface remediation,Environ. Sci. Technol., vol. 29, pp. 1913–1922.

Rai, D., Eary, L.E., and Zachara, J.M., 1989, Environmental chemistry of chromium,Sci. Tot. Environ., vol. 86, pp. 15–23.

Rai, D., Sass, B.M., Moore, D.A., 1987, Chromium (III) hydrolysis constants andsolubility of chromium (III) hydroxide, Inorg. Chem., vol. 26, pp. 345–349.



Rai, D., Zachara, J.M., Eary, L.E., Ainsworth, C.C., Amonette, J.E., Cowan, C.E.,Szelmeczka, R.W., Resch, C.T., Schmidt, R.L., Girvin, D.C., and Smith, S.C.,1988, Chromium reactions in geological materials, Interim Report, Electric PowerResearch Institute (EPRI) EA-5741, EPRI, Palo Alto, CA.

Rai, D., Zachara, J.M., Eary, L.E., Girvin, D.C., Moore, D.A., Resch, C.T., Sass, B.M.,and Schmidt, R.L., 1986, Geochemical behavior of chromium species, Interim ReportElectric Power Research Institute (EPRI) EA EA–4544, EPRI, Palo Alto, CA.

Rai, D., Zachara, J.M., Schwab, A.P., Schmidt, R.L., Girvin, D.C., and Rogers, J.E.,1984, Chemical attenuation rates, coefficients, and constants in leachate migra-tion, Vol. 1, A critical review, Final Report Electric Power Research Institute(EPRI) EA EA-3356, EPRI, Palo Alto, CA.

Richard, F.C. and Bourg, A.C.M., 1991, Aqueous geochemistry of chromium: a review,Wat. Res., vol. 25, pp. 807–816.

Ritchie, G.S.P. and Sposito, G., 1995, Speciation in soil, in Chemical Speciation in theEnvironment Ure, A.M. and Davidson, C.M., Eds., Blackie Academic and Pro-fessional, Glasgow, Scotland, pp. 201–233.

Robertson, F.N., 1975, Hexavalent chromium in the groundwater in the ParadiseValley, Arizona, Ground Water, vol. 13, pp. 516–527.

Rotaru, M., Birck, J.L., and Allegre, C.J., 1992, Nature, vol. 358, pp. 465.Saleh, F.Y., Parkerton, T.F., Lewis, R.V., Huang, J.H., and Dickson, K.L., 1989, Kinetics

of chromium transformation in the environment, Sci. Tot. Environ., vol. 86,pp. 25–41.

Salomons, W. and DeGroot, A.J., 1978, Pollution history of trace metals in sediments,as affected by the Rhine River, in Environmental Biogeochemistry, Krumbein,W.E., Ed., Ann Arbor Science, Ann Arbor, MI, pp. 149–162.

Santillan-Medrano, J. and Jurinak, J.J., 1975, The chemistry of lead and cadmium insoils: solid phase formation, Soil Sci. Am. Proc., vol. 29, pp. 851–856.

Sass, B.M. and Rai, D., 1987, Volubility of amorphous chromium(III)-Iron(III) hydrox-ide solid solutions, Inorg. Chem., vol. 26, pp. 2228–2232.

Schacklette, H.T. and Boerngen, J.G., 1984, Element concentration in soils and othersurficial materials, U.S. Geol. Surv. Prof. Paper 1710, Washington, DC.

Schmidt, R.L., 1984, Thermodynamic properties and environmental chemistry of chromium,U.S. Dept. Energy, DE-AC06-76RLD-1830, Washington, DC.

Schroeder, D.C. and Lee, G.F., 1975, Potential transformation of chromium in naturalwaters, Water Air Soil Pollut., vol. 4, pp. 355–365.

Seigneur, C. and Constantinou, E., 1995, Environ. Sci. Technol., vol. 29, pp. 222–231.Smith, L.A., Means, J.L., Chen, A., Alleman, B., Chapman, Ch.C., Tixier, J.S., Brauning,

S.E., Gavaskar, A.R., and Royer, M.D., 1995, Remedial options for metals-contaminated sites, Contaminant Fate and Migration, CRS Press, Inc., Boca Raton,FL, pp. 17–34, Chap. 3.

Spokes, L.J. and Jickells, T.D., 1995, Speciation of metals in the atmosphere, in ChemicalSpeciation in the Environment, Ure, A.M. and Davidson, C.M., Eds., BlackieAcademic and Professional, Glasgow, Scotland, pp. 137–168.

Stollenwerk, K.G. and Grove, D.B., 1985, Adsorption and desorption of hexavalentchromium in an alluvial aquifer near Telluride, Colorado, J. Environ. Qual.,vol. 14, pp. 150–155.

Stumm, W. and Morgan, J.J., 1981, Aquatic Chemistry, John Wiley and Sons, New York,NY.

Thode, H.G. and Monster, J., 1965, in Fluids in Subsurface Environments—a Sympo-sium, American Association of Petroleum Geologists, Tulsa, OK, pp. 367–377.


Au: Pls. provide chaptertitle forreference



Walkley, A. and Black, L.A., 1934, An examination of the Degtjareff method fordetermining soil organic matter and a proposed modification of the chromicacid titration method, Soil Sci., vol. 37, pp. 29–38.

White, A.F. and Hochella, M.F., 1989, Electron transfer mechanisms associated withthe surface oxidation and dissolution of magnetite and ilmenite, Proc. 6thInternat. Sympos. On Water-Rock Interaction, pp. 765–768.

Wielinga, B., Mizuba, M., Hansel, C., and Fendorf, S., 2001, Iron promoted reductionof chromate by dissimilatory iron-reducing bacteria, Environ. Sci. Technol., vol.35, n. 3.

Wiley, K.G., 1983, Hydrogeological investigation examining the identification andcleanup of chromium contaminated groundwater in Richland Township,Kalamazoo County, Michigan. M.S. Thesis, Wright State University.

Wittbrodt, P.R. and Palmer, C.D., 1995, Reduction of Cr(VI) in the presence of excesssoil fulvic acid, Environ. Sci. Technol., vol. 29, pp. 255–263.

Zachara, J.M., Girvin, D.C., Schmidt, R.L., and Resch, C.T., 1987, Chromate adsorptionon amorphous iron oxyhydroxide in presence of major ground water ions,Environ. Sci. Technol., vol. 21, pp. 589–594.



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