United States Office of Office of Solid Waste EPA154015-941505

Environmental Protection Research and and Emergency October 1994

Agency Development Response

 EPA Ground Water Issue

Natural Attenuation of Hexavalent

Chromium in Groundwater and Soils

Carl D. Palmer* and Robert W. PuIs**

Introduction

Chromium is an important industrial metal used in diverse

products and processes (Nriagu, 1988a, b). At many locations,

Cr has been released to the environment via leakage, poor

storage, or improper disposal practices (Palmer and Wittbrodt,

1991; Calder, 1988). Within the environment, Cr is found

primarily in two oxidation states: Cr(Vl) and Cr(lll). Cr(Vl) is

relatively mobile in the environment and is acutely toxic,

mutagenic (Bianci et al., 1983; Beyersmann et al., 1984;

Bonatti et al., 1976; Paschin et al., 1983), teratogenic (Abbasi

and Soni, 1984), and carcinogenic (Mancuso and Hueper,

1951; Mancuso, 1951; Waterhouse, 1975; Yassi and Nieboer,

1988; One, 1988). In contrast, Cr(lll) has relatively low toxicity

(van Weerelt et al., 1964) and is immobile under moderately

alkaline to slightly acidic conditions.

Concerns about the impact of chromium on human health and

the environment require an evaluation of the potential risk of

chromium entering the groundwater flow system and being

transported beyond compliance boundaries. At sites where

such potential exists, active remedial measures such as

excavation or pump-and-treat have been undertaken.

Experience at sites where pump-and-treat remediation of

chromium-contaminated groundwater is currently underway

suggests that, although it is feasible to remove high levels of

chromium from the subsurface, as concentrations decrease it

becomes more difficult to remove the remaining chromium

(Wittbrodt and Palmer, 1992). While several new remedial

technologies are being investigated, there is still concern

about the cost of such remediation technology; and, at many

sites, there is a debate about the need for expensive

remediation.

Researchers have identified natural reductants that can

transform the more toxic hexavalent form of chromium to the

less toxic trivalent form. Under alkaline to slightly acidic

conditions, this Cr(lll) precipitates as a fairly insoluble

hydroxide, thereby immobilizing it within the soil. Such

“natural attenuation” of hexavalent chromium is of great

interest because it suggests that strict waterquality standards

do not have to be attained everywhere within and beneath the

site. If natural attenuation does occur, pump-and-treat

remediation could desist after the most contaminated

groundwater has been removed, even if the maximum

contaminant level (MCL) has not be achieved. Under certain

circumstances, expensive remedial measures may not even

be necessary.

In this paper, what is known about the transformation of

chromium in the subsurface is explored. This is an attempt to

identify conditions where it is most likely to occur, and

describe soil tests that can assist in determining the likelihood

of natural attenuation of Cr(Vl) in soils.

The Geochemistry of Chromium

Chromium exists in oxidation states ranging from +6 to -2,

however, only the +6 and +3 oxidation states are commonly

encountered in the environment. Cr(Vl) exists in solution as

monomeric ions H2CrOj, HCIOj (bichromate), and CrO~-

(chromate), or as the dimeric ion Cr20~- (bichromate)

● Environmental Science and fngirreerirrg, Oregon Graduate Institute

of Science& Technology

“ Robert S. Kerr Environmental Research Laboratory, U.S. EPA

Superfund Technology Support Center for

Ground Water

Robert S. Kerr Environmental

Research Laboratory

Ada, Oklahoma
(e.g. Palmer and Wittbrodt, 1991; Richard and Bourg, 1991).

The monomeric species impart a yellow color to the water

when the ICr(Vl)] is greater than 1 mg/L. Water that contains

high levels of Cr20~” has an orange color.

The monomeric chromate species are related through a series

of acid dissociation reactions

H2Cr~ 2 HCrO~+H+ ; K,

HCrOj i? C~” +H+ ; K2

the pK values are -0.86 and 6.51, respectively (Allison et al.,

1990). The bichromate is the result of the polymerization of

the monomeric bichromate ions to form the dimer, CrzO~”,

HCrOj +HCrOj 2 CrpO~- +& O ; K~

where pKd is -1.54 (Allison et al., 1990). The relative

concentration of each of these species depends on both the

pH of the contaminated water (Fig. 1) and the total

concentration of Cr(Vl) (Fig. 2). Significant concentrations of

~Cr@ only occur under the extreme condition of pH c 1.

Above pH 6.5, Cr@generally dominates. Below pH 6.5,

HCrO~ dominates when the Cr(Vl) concentrations are low (40

mM); but CrpO~- becomes significant when concentrations are

greater than 1 mM, or it may even dominate when the total

Cr(Vl) concentrations are greater than 30 mM.

In the Cr(lll)-H20 system, Cr(lll) exists predominantly as CF’+

below pH 3.5. With increasing pH, hydrolysis of CF+ yields

CrOH2+, Cr(OH)j , Cr(OH );, and Cr(OH~ (Rai, et al., 1987).

At high concentrations, these ions impart a green color to the

solution. Under slightly acidic to alkaline conditions, Cr(lll)

can precipitate as an amorphous chromium hydroxide.

Amorphous Cr(OH), can crystallize as Cr(OH)~. 3HZ0 or

Cr20, (eskolaite) under different conditions (Swayambunathan

et al., 1989). In the presence of Fe(lll), trivalent chromium can

precipitate as a solid solution. If the pH within the contaminant

plume is between 5 and 12, the aqueous concentration of

Cr(lll) should be less than 1 pmole/L (d.05 mg/L) (Fig. 3).

There are several mineral phases that contain Cr(Vl) that may

be present at chromium contaminated sites. Palmer and

Wittbrodt (1990) identified PbCr04 (crocoite), PbCrOQ . MO

(iranite), and K,CrO, (tarapacaite) in chromium sludge from a

hardchrome plating facility. CaCrO, was found at a seepage

face in a drainage ditch where there was high evaporation.

Most of the contaminated groundwater was at equilibrium with

BaCrO, (hashemite). BaCr04 forms a complete solid solution

with BaSOd (Rai et al., 1988) and can be a major impediment

to the remediation of chromium contaminated sites by pumpand-treat (Palmer and Fish, 1992; Wittbrodt and Palmer,

1992).

u) a) .-

0

a)

$0.6

‘5 t

i

IL 0.2 - .-.:-.’, -,.

0.0 “ ‘ “ “ ‘ .

o 2 4 6 6 10 12

pH

Figure 1. Distribution of Cr(Vl) species as a function of pH.

)0

Figure 2. Fraction of bichromate (HCrOJ and bichromate (CrzO~_) at

pH 4 as a function of the total Cr(Vl) concentration.

-2

K L

Cr(W3tamJ

-4 ‘,

1 —.— —- MCL . — - —- —- — -—-—-—-—- -—-—

-6 ‘\

\

-8 I

‘

= Fe0.75 Cr 0.2 5 (0H)3(am) i

pH

Figure 3. Cr(lll) concentration in equilibrium with Cr(OH),(,~) and

Fe,,, Cr02,(OH),,based on data from Rai et al., (1 987) and

Sass and Rai (1987), respectively.

2
Reduction of Hexavalent Chromium

Cr(Vl) is a strong oxidant and is reduced in the presence of

electron donors. Electron donors commonly found in soils

include aqueous Fe(l I), ferrous iron minerals, reduced sulfur,

and soil organic matter.

The reduction of Cr(Vl) by ferrous iron can be written as

HCrO~ + 31%2+ + 7H+ + Cr3+ + 3F23+ + 4H20

This reaction is very fast on the time scales of interest for most

environmental problems with the reaction going to completion

in less than 5 minutes even in the presence of dissolved

oxygen (Eary and Rai, 1988). Only when the pH is greater

than 10 or when PO~ concentrations exceed 0.1 molar does

the rate of oxidation of F&+ by dissolved oxygen exceed the

rate of oxidation by Cr(Vl) (Eary and Rai, 1988). When the pH

of the groundwater is greater than 4, Cr(lll) precipitates with

the Fe(lll) in a solid solution with the general composition

CrXFel..JOH), (Sass and Rai, 1987; Amonette and Rai, 1990).

If the reduction of Cr(Vl) by Fe(n) is the only source of Fe(lll)

and Cr(lll), a solid solution with the composition

CrO,z~FeO,JOH), forms via the reaction

HCrO~ + 3F62+ + 3H20 + 50H+ dCr0.z5Feo.75(OH)3

(Eary and Rai, 1988; Sass and Rai, 1987). The volubility of

CrXFe,-X(OH)~ decreases as the mole fraction of Fe(lll) in the

solid increases. Therefore, if the pH is between 5 and 12, the

concentration of Cr(lli) is expected to be less than 10+ molar.

Numerous minerals in geologic materials contain ferrous iron

that is potentially available for the reduction of hexavalent

chromium. These iron-containing minerals may be silicates,

oxides, or sulfides. Common ferrous iron-containing silicates

include olivine; pyroxenes such as augite and hedenbergite;

the amphiboles hornblende, cummingtonite, and grunerite;

micas such as biotite, phlogopite, and glauconite; chlorites,

and the smectite nontronite. Iron oxides such as magnetite

(Fe2+Fe~+OJ contain iron as a major constituent, however,

hematite (Fe~+03) can contain small amounts of (FeO). in

sulfide minerals such as pyrite (FeS2), both the ferrous iron

and the sulfide are active in reducing hexavalent chromium.

Lancy (1966) suggested that pyrite could be used for treating

spent cooling waters that contain Cr(Vl) as a corrosion

inhibitor. He stated that the reduction of Cr(Vl) occurs at the

pyrite surface rather than in solution. Lancy (1966) found that

reduction by pyrite occurred even in slightly alkaline solutions;

however, the pyrite had to be continuously abraded to remove

surface coatings. Blowes and Ptacek (1992) conducted batch

tests in continuously agitated reaction vessels containing a

solution of 18 mg/L Cr(Vl) and pyrite both in the presence and

in the absence of calcite. In the experiments that used both

pyrite and calcite, 509!. of the Cr was removed in less than 6.5

hours. Concentrations were <0.05 mg/L after 20 hours.

Experiments conducted without the calcite attained 50%

removal in 1 hour and concentrations were c 0.05 mg/L in less

than 4 hours.

Cr(Vl) reduction in the presence of iron oxides has been

observed in several experiments. White and Hochella (1989)

found that magnetite and ilmenite reduced Cr(Vl) to Cr(lli).

The reduction of Cr(Vl) in the presence of hematite (Fe20J

was demonstrated by Eary and Rai (1989). They attribute the

reduction to the presence of a small amount of an FeO

component in the hematite. They suggest that reduction

occurs in solution after the FeO component has been

solubilized.

Reduction of Cr(Vl) by ferrous iron-containing silicates has

been reported. Eary and Rai (1989) suggest that the

reduction of Cr(Vl) in the presence of biotite occurs in solution

rather than at the mineral surface. They observed an increase

in the rate of reduction when their suspensions were spiked

with Fe&. They explain their results with the mechanism

proposed by White and Yee (1985) in which Fe3+ is reduced at

the mineral surface by the reaction

[Fe(n), K+]~iOtit~ + Fe3+ ~ [Fe(lll)]biOtit~ + K+ + Fe2+

where the ions in the brackets denote ions within the crystal

structure of biotite. To maintain charge balance, K+ is

released to solution as the iron in the crystal structure is

oxidized. The Cr(Vl) in solution is then reduced by the Fez+.

The Fes+ resulting from this reduction reaction is then

adsorbed to the surface of the biotite where it is again reduced

to F&+, thus setting up a cycle that ultimately results in the

reduction of more Cr(Vl) than is stoichiometricaliy possible for

the amount of iron that is in solution.

There are some key experimental difficulties in studying

groundwater/mineral interactions such as those just described

that have some bearing on the transfer of knowledge to the

field. Although the processes can in some cases be

interpreted from the data on mineral reactions, the rates

themselves may be quite useless. A key difficulty in studying

mineral reactions in the laboratory is that the rate of the

reaction depends on how the solid phase was prepared. For

example, if the samples are ground and simply washed before

use, microparticles can adhere to the larger grain surfaces.

These microparticles have greater specific surface area and

can react at a much faster rate than the larger size particles.

Such experimental artifacts were observed in weathering

studies of pyroxenes (Schott et al., 1981).

Another important reductant in soils is organic matter. In fact,

bichromate reduction has been used as a wet combustion

method for the determination of soil organic carbon (Walkley

and Black, 1934). Bichromate can react with soil organic

carbon according to

2Cr20~- + 3C0 + 16H+ --+ 4Cr3+ + 3C02 + 8~0
The CP+ may hydrolyze and precipitate as Cr-hydroxide or it

may bind to the remaining soil organic carbon. Much of the

soil organic carbon is present as soil humic and fulvic acids.

Redox reaction with these materials has been demonstrated

for several redox reactive species. Reduction of Cr(Vl) by soil

humic and fulvic acids has been demonstrated by Bartlett and

Kimble (1 976), Bloomfield and Pruden (1980), Goodgame et

al. (1984), Boyko and Goodgame (1986), and Stollenwerk and

Grove (1985). The rate of reduction of Cr(Vl) decreases with

increasing pH (Fig. 4), increases with the increasing initial

Cr(Vl) concentration, and increases as the concentration of

soil humic substance increases. At neutral pH, many weeks

may be required for the Cr(Vl) to be completely reduced

Cr(Vl).

In addition to these abiotic reduction pathways, Cr(Vl) can be

reduced by microbes in the subsurface (Martin et al., 1994).

Both aerobic and anaerobic reduction by microbes have be

observed, however, the latter is more common. The

mechanisms for Cr(Vl) by these microbes is not well known. It

may be part of a detoxification mechanism that occurs

intracellularly. Alternatively, the chromate may be utilized as a

terminal electron acceptor-as part of the cell’ s-metabolism. A

third possibility is that reduction is an extracellular reaction

with excreted waste products such as H2S. In addition to two

strains of Gram-positive bacteria, Martin et al., (1994) found a

fungus in contaminated soil that was capable of reducing

Cr(Vl) under anaerobic conditions.

J CtfVl) Raduction bya

o.I. Soil Humk Acid

100 mg SHAA

A

0.015 . A

A A

A

pi+ 5

0.010 ■ A A

I -‘m pH 38.

0.005

■

■

O.000 I

o 200 400 600 800 10001200 1400 1S00 1800 2000

Time (hrs)

Figure 4. Reduction of Cr(Vl) in a suspension of 100 mg/L soil humic

acid (SHA) at pH 3 and 5 (Wittbrodt and Palmer, in

preparation).

Oxidation of Cr(lll)

Any evaluation of the natural attenuation of Cr(Vl) must

consider the potential oxidation of the Cr(lll) to the toxic Cr(Vl)

form. In contrast to the numerous pathways for the reduction

of Cr(Vl), there are very few mechanisms for the oxidation of

Cr(lll). Only two constituents in the environment are known to

oxidize Cr(ill) to Cr(Vl): dissolved oxygen and manganese

dioxides (MnOJ (Eary and Rai, 1987). Studies of the reaction

between dissolved oxygen and Cr(lll) revealed very little

(Schroeder and Lee, 1975) or no (Eary and Rai, 1987)

oxidation of Cr(lll) even for experiments conducted at pH as

great as 12.5 for 24 days. Therefore, the transformation of

Cr(lll) by dissolved oxygen is not likely to bean important

mechanism for the oxidation of Cr(lll).

Oxidation of Cr(lll) has been observed in several soils and

sediments. The oxidation of the Cr(lll) can be relatively slow

requiring several months (Fig. 5). For example, Palmer and

Wittbrodt (1 990) monitored Cr(Vl) concentrations in batch

tests using three different geologic media from a site in

Corvallis, OR. They observed increases in Cr(Vl)

concentrations over a 300-day period with Cr(Vl)

concentrations becoming as great as 7 mg/L in one

experiment. Bartlett and Kimble (1976) did not observe

oxidation of Cr(lll) in their dried soils but Bartlett and James

(1979, 1983a) did observe oxidation in soils that were

maintained in a moist state.

,~-1 ~

d

< s ., , & ‘0

v

at

N.-

U

w 10+

o

==

6

10-f

o 50 100 150 200 250 300 350

TIME (Days)

Figure 5. Cr(lll) oxidized to Cr(Vl) in a suspension of Willamette silt

loam.

Bartlett and James (1979) observed a correlation between the

amount of Cr(lll) oxidized by soils and the amount of

hydroquinone-reduced manganese in soils and suggested the

oxidation of Cr(lll) is the result of interaction with manganese

dioxides. This hypothesis has been experimentally verified

using 13-MnOz or pyrolusite (Eary and Rai, 1987) and S-Mn02

(Fendorf and Zasoski, 1992; Riser and Bailey, 1992). There is

an increase in the rate and amount of Cr(lll) oxidation as pH

decreases, and the surface area to solution volume increases.

Experimental results indicate that the oxidation follows the

reaction

CrO1-?+ + 1.5 &MnOz + HCrO; + 1.5Mn2+

Significant oxidation of Cr(lll) was observed in less than 1

hour (Fendorf and Zasoski, 1992) and continued for more than

600 hours (Eary and Rai, 1987). Eary and Rai (1987)

developed an empirical rate law for the oxidation of Cr(lll) by

4
13-Mn02; however, the zero point charge for this phase is quite

different than birnessite which is more commonly found in

soils. Therefore, this rate law may not be applicable to

manganese dioxides in soils.

Perspective on the Natural Attenuation of Cr(Vl)

If hexavalent chromium can be reduced and immobilized in

the subsurface as a result of interaction with naturally existing

reductants, then expensive remedial measures may not be

required at certain sites. In principle, the natural attenuation of

Cr(Vl) in the subsurface is feasible. There are several natural

reductants that can transform Cr(Vl) to Cr(lil). If the pH of the

contaminant plume is between about 5 and 12, Cr(lll)

precipitates as Cr(OH), or as part of a solid solution with

Fe(lll), thereby keeping Cr(lll) concentrations below 1 pmole/L

(0.05 mg/L). Whether or not natural attenuation at a particular

site is a viable option depends on the characteristics of both

the aquifer and the contaminant plume under investigation.

The potential reductants of Cr(Vl) include aqueous species,

adsorbed ions, mineral constituents, and organic matter.

When a contaminant plume containing hexavalent chromium

enters the subsurface, it displaces the groundwater containing

the dissolved reductants. There is little mixing of the waters

containing the reducing agents and the Cr(Vl)-contaminant

plume. What mixing does occur will be driven by molecular

diffusion at the front of the plume or from the edges of the

plume and diffusion from lower permeability lenses containing

relatively immobile water. Thus, aqueous reductants such as

Fez+ are not going to be important in reducing hexavalent

chromium. Mixing of reductants and Cr(Vl) in the plume are

going to occur primarily though the interactions of the plume

with the immobile soil matrix. Such interactions include

desorption of reductants such as F&+ from mineral surfaces,

direct and indirect surface redox reactions between Cr(Vl) and

the mineral surfaces, and reduction by soil organic matter.

Thus, it is the soil matrix that is most important with regards to

redox transformations of chromium in the subsurface. This

argument is further supported by studies that clearly

demonstrate that groundwater contributes less that 1 YO of the

oxidation capacities (equivalents of Cr oxidized per gram of

soil) and reduction capacities (equivalents of Cr reduced per

gram of soil) of aquifer systems while the soil matrix

contributes the remaining fraction (Barcelona and Helm,

1991). Thus, any discussion of redox transformations of

chromium in the subsurface must focus on the soil matrix.

Three key factors must be addressed in considering the

potential use of natural attenuation of Cr(Vl) in the subsurface.

Firstly, the reduction capacity of the aquifer, R,, must be great

enough to reduce all of the Cr(Vl) that passes through it. If XC

is the distance from the source to the point of compliance (Fig.

6), the total mass of Cr(Vl) from the source, MO, must be less

than the total mass of Cr(Vl), M,, that can be reduced by the

aquifer material between the source and XC:

M O< M, = X#pb RC (1)

where A is the cross-sectional area of the plume normal to the

direction of groundwater flow and R is the dry bulk density of

the aquifer. As XC increases, the mass of Cr(Vl) that can be

reduced increases. A key difficulty in applying this criterion is

in providing a reasonable estimate of Mo

. In the absence of

other reactions such as adsorption or precipitation, the

minimum rate of movement of the Cr(Vl) front through the

aquifer, Vtin, computed by assuming the reductant reacts

instantaneously with the Cr(Vl), is

Vmin =

Vw

,+ p~ RC

0“ c.

(2)

where p~ and f3v are the dry bulk volumetric water content of

the porous medium, VW is the velocity of the groundwater, and

CO is the concentration of the chromium in the contaminant

plume.

CROSSSECTIONAL POINT OF

AREA, A COMPUANCE

Figure & Cr-plume moving from the source area to the point of

compliance. The initial Cr(Vl) concentration in the source

area is MO, vw is the groundwater velocity, v~in is the velocity

of the Cr(Vl) front assuming instantaneous reduction of the

Cr(Vl), and XC is the distance from the source area to the

point of compliance.

The second key factor in the application of natural attenuation

of Cr(Vl) is the rate of reduction relative to the rate of

advective transport in the subsurface. The time for the

reduction reaction to decrease the concentration from its initial

concentration, CO, to some target concentration, C,, such as a

drinking water standard, should be less than the residence

time of the contaminated water in the portion of the aquifer

between the source of the Cr(Vl) and the point of compliance.

For example, if the rate of reduction of Cr(Vl) follows a firstorder rate equation

(3)

the time for the concentration of Cr(Vl) to decrease from C. to

C, must be less than the residence time of the contaminated

parcel of water within the aquifer

5
In(co /%) < E

k

 - Vw

If natural attenuation is to be a viable option, this criterion

must be met. Difficulties in utilizing this criterion arise in

applying the appropriate rate equation and obtaining the

pertinent rate coefficients.

(4)

A third factor concerning the natural attenuation of Cr(Vl) is

the possible oxidation of Cr(lll) to the more toxic hexavalent

form. While contamination is actively entering the subsurface,

conditions may favor the reduction of Cr(Vl) to Cr(lll). After

the source of the active contamination is removed, however,

chemical parameters within the aquifer, particularly pH, may

be altered. Under the new conditions, oxidation of Cr(lll) may

be favored. Thus, soil containing Cr(lll) formed during the

active contamination phase may become a source of Cr(Vl).

Both oxidation and reduction of chromium are occurring

simultaneously within the subsurface as part of a geochemical

cycle. As the Cr(lll) is oxidized to Cr(Vl) by manganese

dioxides in the soil, Cr(Vl) can be reduced to Cr(lll) by some

reductant such as soil organic carbon or pyrite. The rate of

change in ICr(Vl)] (d[Cr(Vl)]/dt) is the sum of the rate of

reduction of Cr(Vl), R,d, and the rate of oxidation of Cr(lll), ROX:

‘[cl’’)] =[%Hrd+[wlox=Rr* +Rc’x “)

If a soil initially contains both Cr(Vl) and Cr(lll), then ICr(Vl)]

increases when Rti + ROX >0 and decreases when

R,d + ROX e O. Ultimately, the ICr(Vl)] will reach a steady state,

i.e., d[Cr(Vl)]/dt = O. At this time, the rate of loss of Cr(Vl) via

reduction is balanced by the rate of production by the

oxidation of Cr(lll):

-R - rd - ROX (6)

Wittbrodt and Palmer (1 994) suggest that the reduction of

Cr(Vl) by soil fulvic acid can be represented by

R,d = -k,~ X; [HCK3~][SHS][H+]p (7)

where [SHS] is the concentration of soil humic substance and

X8 denotes the equivalent fraction of the humic substance that

has been oxidized. Fendorf and Zasoski (1992) suggest that

CrOH2+ is the reactive species in the oxidation of Cr(lll) by

MnO,. For illustrative purposes, assume that the oxidation

reaction follows a rate equation of the form

ROX = kOX [CrOH2+] [A/V]m ~+]n (8)

where (A/V) denotes the surface area of the Mn02 per unit

volume of solution. If we further assume that the solution is

equilibrated with Cr(OH),(,~), then

ROX = kOX (K /K~)(A/V)m [H+]”+’

where KW is the dissociation constant for H20 and K is the

equilibrium constant for the reaction

Cr(OH)x~~) z Cr(OH)’+ + 2(OH-)

(9)

Equating ROX with -R,w and rearranging the terms yields

[Hcro~l’*+x.-~+ln+2-p (lo)

red w

Although some of the specific points of rate equations

presented here are debatable, equation 10 does illustrate

aspects of natural attenuation in soils that contain both a

reductant and Mn02. The key point is that as long as the

supply of reductant and MnOz have not been significantly

depleted, [HCrO~ does not converge to zero with increasing

residence time within the aquifer as one would expect for a

first order reaction that only considers reduction of Cr(Vl).

Rather, [HCrO~ converges to some steady-state

concentration that is> O that may or may not be above the

MCL. This steady-state concentration increases with

increasing kO~k,d, and (A/V) m/[SHS] and it varies with pH.

Thus, in principle, if the rate equations are correct and all of

the parameters are known, one could calculate the steadystate Cr(Vl) concentration and determine if natural attenuation

could achieve compliance goals. Studies of the kinetics of

these coupled processes needs be done to verify the general

forms of the rate equations and to determine the appropriate

rate coefficients.

Determining the Potential for Natural Attenuation

If “natural attenuation” is to be considered an alternative to

expensive remediation efforts, additional characterization is

required to demonstrate that the expectations are likely to be

met. There is no single test that can tell us if natural

attenuation of Cr(Vl) will occur at a particular site. Several

tests are briefly described which have been utilized to address

key factors affecting Cr(Vl) transport in the subsurface and

describe how the results can be utilized in determining the

potential for the natural attenuation of Cr(Vl) in the subsurface.

Ideally, it must be demonstrated that 1 ) there are natural

reductants present within the aquifer, 2) the amount of Cr(Vl)

and other reactive constituents do not exceed the capacity of

the aquifer to reduce them, 3) the rate of Cr(Vl) reduction is

greater than the rate of transport of the aqueous Cr(Vl) from

the site, 4) the Cr(lll) remains immobile, and 5) there is no net

oxidation of Cr(lll) to Cr(Vl). Some of these criteria are

relatively simple while others require additional tests and

interpretation. Additional tests that will be required include

tests of the oxidizing and reducing capacities of the aquifer.
Mass of Cr(Vl) at the Source

It must be demonstrated that the amount of Cr(Vl) in the

aquifer does not exceed the capacity of the soil for reducing

this chromium. Therefore an important first step in evaluating

the potential for natural attenuation is to determine the mass

of Cr(Vl) in the soil. Chromium exists in the subsurface either

in solution or in association with the solid phase. Cr(Vl) in

solution can be determined by the diphenylcarbazide (DPC)

method (APHA, 1989). Aqueous samples are most often

obtained from monitoring wells. Alternatively, water separated

from the soil matrix either by centrifugation or by squeezing.

The pH of these waters should be measured to determine if it

is within the proper range (5.5 to 12) to insure the Cr(lll)

concentrations are less than 1 ~M (0.05 mg/L).

Cr(Vl) associated with the soil matrix maybe adsorbed to

mineral surfaces (particularly iron oxides) or precipitated as

chromate minerals. There is no precise method for

I determining each of these fractions of Cr(Vl); nonetheless,

determinations have been made using sequential extractions.

An initial water extraction serves to remove remaining pore

water and dissolve highly soluble chromium minerals present

in the soil or that may have precipitated as the result of

evaporation during sample handling and storage. This water

extraction also removes some adsorbed ions.

Following the water extraction, a phosphate extraction is used

as a measure of the “exchangeable” chromate in the soil

(Bartlett and James, 1988). The test is conducted by adding

phosphate to the soil and equilibrating for 24 hours. The

water is then separated from the slurry and Cr(Vl) is measured

by the DPC method (Bartlett and Kimble, 1976; Bartlett and

James, 1988c). The increase in the chromate concentration is

the amount of “exchangeable” chromate. Amacher and Baker

(1982) found optimal extraction using 0.01 M monobasic

potassium phosphate (KH,PO,). James and Bartlett (1983b)

used a solution of 0.005 M KHZPOA and 0.05 M KzHPOi to

yield a pH of 7.2. James and Bartlett (1983b) stated that

doing the extraction at pH 7.2 is preferred because there is

less likelihood of chromate reduction than at lower pH.

However, decreasing the pH of the soil slurry can result in

dissolution of BaCrO, from the soil. Moreover, if the pH of the

soil water was initially low, then increasing the pH to 7.2 can

cause precipitation of BaC@d, thereby complicating the

interpretation of the results. When the soil water is not

equilibrated with BaCrO,, the phosphate extraction method of

James and Bartlett (1983b) primarily measures the amount of

adsorbed Cr(Vl) in the soil. The phosphate removes chromate

by both directly competing for the adsorption sites in the soil

and indirectly (in some cases) by increasing the pH.

BaCrO, is a likely chromate mineral phase that can be a

source of Cr(Vl) in contaminated aquifers. There is no direct

test for BaCrOd in soils, however, when the groundwater is

equilibrated with this phase and the source of the Ba2+ is

entirely from the clays in the natural soil, the maximum

amount of BaCrOd in the aquifer is equal to the ammonium

acetate exchangeable Ba2+ (Thomas, 1982) in background

soils. For example, Palmer and Wittbrodt (1990) found that

the amount of exchangeable Ba2+ was useful in estimating the

number of pore volumes required to flush Cr(Vl) from soil

columns.

At many sites, the total Cr(Vl) associated with the soil matrix is

the sum of the BaCrO, and the PO,-extractable Cr(Vl). This

sum, S, is often reported in units of mass per gram of soil.

The total concentration of Cr(Vl) in the soil, Cr(Vl),O, is the sum

of the aqueous Cr(Vl) and the matrix associated Cr(Vl), S,

which can be reported in common units of mass per unit

volume of water by

1000 p~ s

Cr(Vl)tOt = [Cr(Vl)] +— (11)

0,

where the dry bulk density of the soil, p~, is in g-cm+. The

total mass of Cr(Vl) in the site soils can then be estimated by

integrating the concentrations over the volume of

contaminated soil.

Mass of Cr(lll) in the Subsurface

If all of the chromium that entered the soil was Cr(Vl), then

demonstrating the presence of Cr(lll) in the soil would prove

that reduction is occurring. The mass of Cr(lll) in the soil can

provide a measure of the amount of reduction that has

occurred. Although proof of chromate reduction is necessary,

it is not sufficient for demonstrating that natural attenuation will

adequately protect the environment.

The total amount of Cr(lll) present in the soil is the sum of the

mass in solution as well and mass associated with the solid

phase. Total chromium in solution can be determined by

atomic absorption spectrophotometry (AAS) or inductively

coupled plasma spectroscopy (ICP). When total chromium is

statistically greater than Cr(Vl), Cr(lll) can be simply

determined by difference.

The amount of Cr(lll) associated with the soil matrix has

ostensibly been determined using several techniques. An

ammonium oxalate (0.1 M) extraction serves to remove

amorphous hydroxides of Cr, Fe, and Al (Ku et al., 1978;

Borggaard, 1988). Bartlett (1991) suggests that a K, H-citrate

extraction provides a measure of the Cr(lll) that is potentially

removable by low molecular weight organic molecules. A

dithionate-citrate-bicarbonate (DCB) extraction is conducted

by adding 0.3 M sodium citrate and 0.1 M sodium bicarbonate

to the soil sample and heating to 80°C for 20 minutes. One

gram of sodium dithionate is then added and the soil slurry is

stirred for another 15 minutes. The DCB extraction removes

the crystalline forms of the Cr-, Fe-, and Al-oxyhydroxides (Ku

et al., 1978; Borggaard, 1988). The dithionate reduces

crystalline iron (goethite) in the soil and Cr, Fe, and Al are

complexed by the citrate. In addition to the Cr(lll)

oxyhydroxides, the DCB method also extracts sparingly

soluble Cr(Vi) mineral phases such as BaCrO,, thereby

complicating interpretation of the results. Bartlett (1991) uses

7
40 mL per gram of soil of 0.7 M NaOCl solution (undiluted

laundry bleach) at pH 9.5 to extract chromium. The slurry is

placed in a boiling water bath for 20 minutes before the liquid

is separated and Cr is determined by AAS or ICP. This

method is useful in determining total chromium in the soil

because it readily oxidizes and removes Cr(lll) that is not

removed by other methods.

Identification of Potential Reductants

The presence of Cr(lll) in the soil maybe indicative of active

reduction in the soil, or it may be the result of the

neutralization of acidic waters containing Cr(lll) with

subsequent precipitation of chromium hydroxides. Therefore,

identification of specific reductants within the aquifer is

warranted. The identification of some potential reductants at a

site can be fairly simple in some cases. For example, pyrite

(FeS,), a common constituent in many geological materials, is

readily identifiable by its visual characteristics. Other mineral

phases capable of reducing Cr(Vl) can be identified using

classical petrographic techniques or powder x-ray diffraction.

Scanning electron microscopy (SEM) can be utilized to identify

crystallite morphology. SEMS equipped with energy

dispersive x-ray spectroscopy can also provide information

about the elemental composition of these crystallite. Electron

diffraction patterns obtained from transmission electron

microscopes provide crystallographic information. Such

electron microscopy methods can, however, be relatively

expensive. A fairly simple and inexpensive test for organic

carbon can provide a measure of the amount of carbon

available for reduction of Cr(Vl). Knowledge of the specific

reductant within the aquifer is useful in determining the time

scale for the reduction of Cr(Vl) based on studies that are

reported in the literature. Soils containing iron sulfides or

organic matter are more likely to reduce Cr(Vl) on the time

scales of interest than soils containing ferrous iron silicates.

Reduction Capacity of the Aquifer

Adequate protection of the environment by natural attenuation

of Cr(Vl) requires that the soil possess a large enough

reducing capacity to reduce all the hexavalent chromium in the

source area. Several measures for predicting reduction of

Cr(Vl) in soil are presented by Bartlett and James (1 988) and

Bartlett (1991). A measure of the maximum amount of Cr(Vl)

that can be reduced per unit mass of aquifer, the “total Cr(Vl)

reducing capacity”, can be obtained using the classical

Walkley-Black method for determining soil organic carbon

(Bartlett and James, 1988). In this method, 2 to 3 grams of

soil are reacted with a mixture of 1 N K2CrzOT in IN H#Od for

30 minutes (Walkley and Black, 1934; Nelson and Sommers,

1982; Bartlett and James, 1988). The Cr(Vl) concentration is

measured using the diphenylcarbazide (DPC) method (APHA,

1989) and the decrease in the mass of Cr(Vl) in the reaction

vessel per gram of soil used in the test is the reduction

capacity. Although this method of determining soil organic

carbon has its limitations (e.g. Nelson and Sommers, 1982), it

is a direct measure of how much Cr(Vl) can be reduced by a

soil at extreme acid concentrations. Variations on this method

use heat or a combination of heat and pressure (Nelson and

Sommers, 1982). Barcelona and Helm (1991) used a

modified closed-tube chemical oxygen demand procedure

(U.S. EPA, 1979) to determine reduction capacities.

The extreme conditions of pH and temperature used in the

total Cr(Vl) reducing capacity test may yield a greater reducing

capacity than would be available under most environmental

conditions. The “available reducing capacity” test of

Bartlett and James (1988) determined the reduction capacity

by reacting about 4 to 5 grams of moist soil in a solution of

10 mM H, PO, and K2Cr20, for 18 hours. The H~PO, is added

to buffer the pH and to compete with the Cr(Vl) for the

adsorption sites. When KH2POd is used Bartlett and James

referred to it as the “reducing intensity”. These tests are

designed to determine the reducing capacity at pH values

more likely to be encountered in the field. However, in long

term reduction tests at near neutral pH, Palmer and Wittbrodt

(unpublished data) observed reduction occurring after 250

days (Fig. 7). Such long-term reduction tests are not practical

at most waste sites.

0.08

n

=

g

~ 0.06

~

: 0.04

~

~ 0.02

w

K

~ 0.00

50 100 150 200 250 300

TIME (Days)

Figure 7. Cr(Vl) reduced to Cr(lll) in Willamette silt loam.

Oxidation Capacity

A potential limitation to the use of natural attenuation of Cr(Vl)

in soil is the oxidation of the Cr(lll) to Cr(Vl) by Mn02. If the

oxidizing capacity of the soil is greater than the reduction

capacity, then as the chromium is cycled in the soil it could

exhaust the soil reductant and be oxidized and ultimately

mobilized in the soil. It is important, therefore, to determine

the capacity of the aquifer to oxidize Cr(lll).

Bartlett and James (1988) suggest a relatively simple test for

the amount of Cr(lll) that can be oxidized by a soil. The

method involves adding 2.5 grams of soil to a solution

containing 25 ml of 1 mM CrCl~. After shaking for 15 minutes,

a solution of K&POd . K2HPOA is added to the reaction

vessel, the slurry centrifuged or filtered, and the Cr(Vl)

8
measured using the DPC method. Moist soils should be used

in these tests. Drying the soils alters the surfaces of the

manganese dioxides making them less reactive (Bartlett and

James, 1979).

Barcelona and Helm (1991) used a solution of chromous

(Cr(ll)) ion to measure the oxidation capacity of soils. They

added about 1 g of soil to cuvettes containing the Cr(ll)

solution. The work was performed in a glove box to prevent

oxygen from reacting with the Cr(ll). The cuvettes were

sealed, shaken, and allowed to react for 2 hours. The

samples were centrifuged and the Cr(ll) measured

spectrophotometrically. The loss of Cr(ll) is then used as a

measure of the oxidation capacity of the soils.

Each of these methods has some problems. Palmer and

Wittbrodt (unpublished data) conducted oxidation tests similar

to the Bartlett and James (1 988) method except that the

concentration of Cr(Vl) was monitored over nearly a year.

Cr(Vl) concentrations in these tests continued to increase up

to 100 days and may have been continuing after 300 days.

For one soil, the short-term oxidation test of Bartlett and

James underestimated the amount of oxidation obtained in the

long-term tests by more than an order of magnitude. It is not

clear in the Barcelona and Helm method whether the Cr(ll) is

being oxidized to Cr(lll) or Cr(Vl) or some combination of the

two. If both products are forming, the results are more difficult

to interpret.

If the only mechanism for the oxidation of Cr(lll) in soils is

oxidation by manganese oxides, then using extraction

methods specifically designed for this purpose may be a good

way of determining the oxidation capacity of the soils. One

very simple extraction technique (Chao, 1972; Gambrell and

Patrick, 1982) utilizes 0.1 M hydroxylamine hydrochloride

(NHAOH . HCI) in 0.1 M HNO,. About 0.5 g of soil are added

to 25 mL NH20H . HCI of the solution and shaken for 30

minutes and the Mn concentration is measured. The number

of moles of Cr(lll) that can be oxidized is then computed by

dividing the number of moles of Mn2+ per gram of soil obtained

in the extraction by 1.5. The hydroxylamine hydrochloride test

is fast, easy, and specifically targets the phase that promotes

the oxidation of Cr(lll).

Rates of Oxidation and Reduction

Key factors in the suitability of natural attenuation as an option

for chromium contaminated soils are the rates of oxidation and

reduction of chromium. This information is the most difficult to

obtain. Scientists are only now learning about the form of the

aDDlicable rate eauations and the armomiate rate coefficients

reductant and pH, it is important to obtain rate coefficients that

were acquired under conditions similar to those at the site.

Many rate studies have considered only a limited set of

conditions such as a single pH value or one reductant

concentration. Consequently, the reported rate coefficients

are apparent values that are strictly valid only under the

conditions of the experiment. Thus, the experimental factors

must be taken into account before the rate coefficients can be

applied to field problems.

When MnO, is present, Cr(lll) maybe oxidized back to Cr(Vl)

and the net rate of reduction will be less than that obtained

from experiments that only utilize reductants. Further, many

rate experiments are conducted in stirred reactors that can

abrade reactive surfaces. In soils, the rate of reaction may

become surface limited as adsorbed ions and precipitates

cover the reactive surfaces.

One method of obtaining the net rate of reduction is through

tests on uncontaminated soils obtained from the site. These

soils should be similar to those through which the contaminant

plume will be migrating. Cr(Vl) can be added to the soil slurry

and the Cr(Vl) concentrations monitored over time. The

reaction vessels must exclude light to prevent photoreduction

reactions and the slurry must have the same pH as the

contaminant plume. A key limitation to such experiments is

that they require several months to a year to complete.

Estimating Reduction from Monitoring Well Data

In principle, Cr(Vl) reduction can be estimated from the

decrease in the mass of Cr(Vl) in the aquifer (e.g., Henderson,

1994). The key difficulty in such an approach is to estimate

the mass of Cr(Vl) using the aqueous concentrations. The

total mass of Cr(Vl) in the aquifer is the sum of the mass that

is in solution, the mass that is adsorbed to the aquifer matrix,

and the mass that is precipitated within the aquifer. The mass

of Cr(Vl) in solution is obtained by integrating the Cr(Vl)

concentrations over the volume of the contaminated aquifer

M,q = 8VCV (12)

where V is the volume of aquifer containing a plume with a

Cr(Vl) concentration of C.

The mass of Cr(Vl) adsorbed to the soil matrix, M,m, can be

computed from the adsorption isotherm. For example, if

Cr(VI) follows a Langmuir isotherm, then over a volume of

aquifer, V, with constant aqueous Cr(Vl) concentration, C, M,b

can be computed as

I that may apply. Such kinetic studies are an area where

research has lagged behind the practical need for the

information. While rates can be obtained from the technical

literature, one must be careful to use rates of reduction for

materials that are most likely controlling the Cr(Vl) reduction in

the site soil. In addition, because the rates depend on the

concentration (surface area per liter of solution) of the

[ 1

Md, . Pb :;aKbscc V(l-r3v)

‘KS

(13)

where K~& is the Langmuir adsorption constant, S- is the

maximum amount of contaminant that can be adsorbed to the

soil.
There is no unique amount of Cr(Vl) precipitate for a given

hexavalent chromium concentration. Therefore it is

impossible to estimate mass of this fraction of Cr(Vl) is the

subsurface using only the measured concentrations in

monitoring wells. Thus, natural attenuation of Cr(Vl) from

mass balances using monitoring well data can only be used

when it can be reasonably demonstrated the Cr(Vl)

precipitates cannot form within the aquifer.

Even when it is demonstrated that the formation of precipitates

within the aquifer is unlikely, there are inherent problems with

any monitoring system that can create uncertainties in the

estimated mass of Cr(Vl) during a sampling round. In the

three-dimensional flow field, the highest concentrations from

one sampling period may migrate between the discrete

monitoring points of the next sampling round. The undetected

mass is not included in the total mass estimates in the second

sampling round and may be mistakenly interpreted as mass

loss due to Cr(Vl) reduction.

Summary

Under certain conditions, toxic Cr(Vl) can be reduced to the

less toxic Cr(lll) in soils and precipitated as an insoluble

hydroxide phase. The possibility of relying on such “natural

attenuation” of Cr(Vl) is attractive because of the great

expense of remediating chromium contaminated sites. Before

such an option is adopted, however, it should be

demonstrated that natural attenuation is likely to occur under

the specific conditions at the site being investigated.

If natural attenuation is to be considered a viable option for

chromium contaminated sites, then ideally, it must be

demonstrated that 1) there are natural reductants present

within the aquifer, 2) the amount of Cr(Vl) and other reactive

constituents do not exceed the capacity of the aquifer to

reduce them, 3) the time scale required to achieve the

reduction of Cr(Vl) to the target concentration is less than the

time scale for the transport of the aqueous Cr(Vl) from source

area to the point of compliance, 4) the Cr(lli) will remain

immobile, and 5) there is no net oxidation of Cr(lll) to Cr(Vl).

The most difficult information to obtain is the time scales for

the reduction and oxidation of chromium in the soil.

Demonstrating Cr(Vl) reduction in aquifer by mass balances

that rely primarily on the aqueous concentrations from

monitoring well networks are valid only if it is demonstrated

that Cr(Vl) precipitates are not forming in the aquifer. The

monitoring network must be sufficiently dense that estimates

of Cr(Vl) are accurate.

Several soil tests are described that are useful in determining

the mass of Cr(Vl) and Cr(lll) in the source areas and the

reduction and oxidation capacities of the aquifer materials.

Some simple conceptual models are presented whereby this

information, combined with knowledge of the residence time of

the chromium between the source and the point of compliance

can be used to determine the feasibility of natural attenuation

of Cr(Vl). The major limitation to this approach is the lack of

information about the rate of oxidation and reduction of

chromium under conditions likely to encountered by plumes

emanating from chromium sources. Without better information

about these rate processes under a wider range of conditions

with respect to pH, the use of the natural attenuation option for

contaminated soils will continue to be a highly debated issue.

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