United States Office of Office of Solid Waste EPA154015-941505Environmental Protection Research and and Emergency October 1994Agency Development Response

EPA Ground Water Issue

Natural Attenuation of HexavalentChromium in Groundwater and Soils

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

Introduction

Chromium is an important industrial metal used in diverseproducts and processes (Nriagu, 1988a, b). At many locations,Cr has been released to the environment via leakage, poorstorage, or improper disposal practices (Palmer and Wittbrodt,1991; Calder, 1988). Within the environment, Cr is foundprimarily in two oxidation states: Cr(Vl) and Cr(lll). Cr(Vl) isrelatively 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 (Abbasiand 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 moderatelyalkaline to slightly acidic conditions.

Concerns about the impact of chromium on human health andthe environment require an evaluation of the potential risk ofchromium entering the groundwater flow system and beingtransported beyond compliance boundaries. At sites wheresuch potential exists, active remedial measures such asexcavation or pump-and-treat have been undertaken.Experience at sites where pump-and-treat remediation ofchromium-contaminated groundwater is currently underwaysuggests that, although it is feasible to remove high levels ofchromium from the subsurface, as concentrations decrease itbecomes more difficult to remove the remaining chromium(Wittbrodt and Palmer, 1992). While several new remedialtechnologies are being investigated, there is still concernabout the cost of such remediation technology; and, at manysites, there is a debate about the need for expensiveremediation.

Researchers have identified natural reductants that cantransform the more toxic hexavalent form of chromium to theless toxic trivalent form. Under alkaline to slightly acidicconditions, this Cr(lll) precipitates as a fairly insolublehydroxide, thereby immobilizing it within the soil. Such“natural attenuation” of hexavalent chromium is of greatinterest because it suggests that strict waterquality standardsdo not have to be attained everywhere within and beneath thesite. If natural attenuation does occur, pump-and-treatremediation could desist after the most contaminatedgroundwater has been removed, even if the maximumcontaminant level (MCL) has not be achieved. Under certaincircumstances, expensive remedial measures may not evenbe necessary.

In this paper, what is known about the transformation ofchromium in the subsurface is explored. This is an attempt toidentify conditions where it is most likely to occur, anddescribe soil tests that can assist in determining the likelihoodof 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 commonlyencountered 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 Instituteof Science& Technology

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

Superfund Technology Support Center forGround Water

Robert S. Kerr EnvironmentalResearch LaboratoryAda, Oklahoma

(e.g. Palmer and Wittbrodt, 1991; Richard and Bourg, 1991).The monomeric species impart a yellow color to the waterwhen 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 seriesof 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 relativeconcentration of each of these species depends on both thepH of the contaminated water (Fig. 1) and the totalconcentration 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 aregreater than 1 mM, or it may even dominate when the totalCr(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 thesolution. Under slightly acidic to alkaline conditions, Cr(lll)can precipitate as an amorphous chromium hydroxide.

Amorphous Cr(OH), can crystallize as Cr(OH)~. 3HZ0 orCr20, (eskolaite) under different conditions (Swayambunathanet al., 1989). In the presence of Fe(lll), trivalent chromium canprecipitate as a solid solution. If the pH within the contaminantplume is between 5 and 12, the aqueous concentration ofCr(lll) should be less than 1 pmole/L (d.05 mg/L) (Fig. 3).

There are several mineral phases that contain Cr(Vl) that maybe present at chromium contaminated sites. Palmer and

Wittbrodt (1990) identified PbCr04 (crocoite), PbCrOQ . MO(iranite), and K,CrO, (tarapacaite) in chromium sludge from ahardchrome plating facility. CaCrO, was found at a seepageface in a drainage ditch where there was high evaporation.Most of the contaminated groundwater was at equilibrium withBaCrO, (hashemite). BaCr04 forms a complete solid solutionwith BaSOd (Rai et al., 1988) and can be a major impedimentto the remediation of chromium contaminated sites by pump-and-treat (Palmer and Fish, 1992; Wittbrodt and Palmer,1992).

u)a).-0a)

$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~_) atpH 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

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

Fe,,, Cr02,(OH), ,based on data from Rai et al., (1 987) andSass and Rai (1987), respectively.

2

Reduction of Hexavalent Chromium

Cr(Vl) is a strong oxidant and is reduced in the presence ofelectron donors. Electron donors commonly found in soilsinclude 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 mostenvironmental problems with the reaction going to completionin less than 5 minutes even in the presence of dissolvedoxygen (Eary and Rai, 1988). Only when the pH is greaterthan 10 or when PO~ concentrations exceed 0.1 molar doesthe rate of oxidation of F&+ by dissolved oxygen exceed therate of oxidation by Cr(Vl) (Eary and Rai, 1988). When the pHof the groundwater is greater than 4, Cr(lll) precipitates withthe Fe(lll) in a solid solution with the general compositionCrXFel..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 compositionCrO,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 ofCrXFe,-X(OH)~ decreases as the mole fraction of Fe(lll) in thesolid increases. Therefore, if the pH is between 5 and 12, theconcentration of Cr(lli) is expected to be less than 10+ molar.

Numerous minerals in geologic materials contain ferrous ironthat is potentially available for the reduction of hexavalentchromium. These iron-containing minerals may be silicates,oxides, or sulfides. Common ferrous iron-containing silicatesinclude 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). insulfide minerals such as pyrite (FeS2), both the ferrous ironand the sulfide are active in reducing hexavalent chromium.

Lancy (1966) suggested that pyrite could be used for treatingspent cooling waters that contain Cr(Vl) as a corrosioninhibitor. He stated that the reduction of Cr(Vl) occurs at thepyrite surface rather than in solution. Lancy (1966) found thatreduction by pyrite occurred even in slightly alkaline solutions;however, the pyrite had to be continuously abraded to removesurface coatings. Blowes and Ptacek (1992) conducted batchtests in continuously agitated reaction vessels containing asolution of 18 mg/L Cr(Vl) and pyrite both in the presence andin the absence of calcite. In the experiments that used bothpyrite and calcite, 509!. of the Cr was removed in less than 6.5hours. 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 lessthan 4 hours.

Cr(Vl) reduction in the presence of iron oxides has beenobserved 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 (Fe20Jwas demonstrated by Eary and Rai (1989). They attribute thereduction to the presence of a small amount of an FeOcomponent in the hematite. They suggest that reductionoccurs in solution after the FeO component has beensolubilized.

Reduction of Cr(Vl) by ferrous iron-containing silicates hasbeen reported. Eary and Rai (1989) suggest that thereduction of Cr(Vl) in the presence of biotite occurs in solutionrather than at the mineral surface. They observed an increasein the rate of reduction when their suspensions were spikedwith Fe&. They explain their results with the mechanismproposed by White and Yee (1985) in which Fe3+ is reduced atthe 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 crystalstructure of biotite. To maintain charge balance, K+ isreleased to solution as the iron in the crystal structure isoxidized. The Cr(Vl) in solution is then reduced by the Fez+.The Fes+ resulting from this reduction reaction is thenadsorbed to the surface of the biotite where it is again reducedto F&+, thus setting up a cycle that ultimately results in thereduction of more Cr(Vl) than is stoichiometricaliy possible forthe amount of iron that is in solution.

There are some key experimental difficulties in studyinggroundwater/mineral interactions such as those just describedthat have some bearing on the transfer of knowledge to thefield. Although the processes can in some cases beinterpreted from the data on mineral reactions, the ratesthemselves may be quite useless. A key difficulty in studyingmineral reactions in the laboratory is that the rate of thereaction depends on how the solid phase was prepared. Forexample, if the samples are ground and simply washed beforeuse, microparticles can adhere to the larger grain surfaces.These microparticles have greater specific surface area andcan react at a much faster rate than the larger size particles.Such experimental artifacts were observed in weatheringstudies of pyroxenes (Schott et al., 1981).

Another important reductant in soils is organic matter. In fact,bichromate reduction has been used as a wet combustionmethod for the determination of soil organic carbon (Walkleyand Black, 1934). Bichromate can react with soil organiccarbon according to

2Cr20~- + 3C0 + 16H+ --+ 4Cr3+ + 3C02 + 8~0

The CP+ may hydrolyze and precipitate as Cr-hydroxide or itmay bind to the remaining soil organic carbon. Much of thesoil organic carbon is present as soil humic and fulvic acids.Redox reaction with these materials has been demonstratedfor several redox reactive species. Reduction of Cr(Vl) by soilhumic and fulvic acids has been demonstrated by Bartlett andKimble (1 976), Bloomfield and Pruden (1980), Goodgame etal. (1984), Boyko and Goodgame (1986), and Stollenwerk andGrove (1985). The rate of reduction of Cr(Vl) decreases withincreasing pH (Fig. 4), increases with the increasing initialCr(Vl) concentration, and increases as the concentration ofsoil humic substance increases. At neutral pH, many weeksmay be required for the Cr(Vl) to be completely reducedCr(Vl).

In addition to these abiotic reduction pathways, Cr(Vl) can bereduced by microbes in the subsurface (Martin et al., 1994).Both aerobic and anaerobic reduction by microbes have beobserved, however, the latter is more common. Themechanisms for Cr(Vl) by these microbes is not well known. Itmay be part of a detoxification mechanism that occursintracellularly. Alternatively, the chromate may be utilized as aterminal electron acceptor-as part of the cell’ s-metabolism. Athird possibility is that reduction is an extracellular reactionwith excreted waste products such as H2S. In addition to twostrains of Gram-positive bacteria, Martin et al., (1994) found afungus in contaminated soil that was capable of reducingCr(Vl) under anaerobic conditions.

J CtfVl) Raduction bya

o.

I.Soil Humk Acid

100 mg SHAAA

0.015 . AA A

A pi+ 50.010 ■ A A

I -‘mpH 38.0.005

■

■

O.000 Io 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 humicacid (SHA) at pH 3 and 5 (Wittbrodt and Palmer, inpreparation).

Oxidation of Cr(lll)

Any evaluation of the natural attenuation of Cr(Vl) mustconsider the potential oxidation of the Cr(lll) to the toxic Cr(Vl)form. In contrast to the numerous pathways for the reductionof Cr(Vl), there are very few mechanisms for the oxidation ofCr(lll). Only two constituents in the environment are known to

oxidize Cr(ill) to Cr(Vl): dissolved oxygen and manganesedioxides (MnOJ (Eary and Rai, 1987). Studies of the reactionbetween 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 asgreat as 12.5 for 24 days. Therefore, the transformation ofCr(lll) by dissolved oxygen is not likely to bean importantmechanism for the oxidation of Cr(lll).

Oxidation of Cr(lll) has been observed in several soils andsediments. The oxidation of the Cr(lll) can be relatively slowrequiring several months (Fig. 5). For example, Palmer andWittbrodt (1 990) monitored Cr(Vl) concentrations in batchtests using three different geologic media from a site inCorvallis, 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 oneexperiment. Bartlett and Kimble (1976) did not observeoxidation of Cr(lll) in their dried soils but Bartlett and James(1979, 1983a) did observe oxidation in soils that weremaintained in a moist state.

,~-1 ~ d

<s ., ,& ‘0

vatN.-

Uw 10+o==6

10-fo 50 100 150 200 250 300 350TIME (Days)

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

Bartlett and James (1979) observed a correlation between theamount of Cr(lll) oxidized by soils and the amount ofhydroquinone-reduced manganese in soils and suggested theoxidation of Cr(lll) is the result of interaction with manganesedioxides. 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 isan increase in the rate and amount of Cr(lll) oxidation as pHdecreases, and the surface area to solution volume increases.Experimental results indicate that the oxidation follows thereaction

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

Significant oxidation of Cr(lll) was observed in less than 1hour (Fendorf and Zasoski, 1992) and continued for more than600 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 quitedifferent than birnessite which is more commonly found insoils. Therefore, this rate law may not be applicable tomanganese dioxides in soils.

Perspective on the Natural Attenuation of Cr(Vl)

If hexavalent chromium can be reduced and immobilized inthe subsurface as a result of interaction with naturally existingreductants, then expensive remedial measures may not berequired at certain sites. In principle, the natural attenuation ofCr(Vl) in the subsurface is feasible. There are several naturalreductants that can transform Cr(Vl) to Cr(lil). If the pH of thecontaminant plume is between about 5 and 12, Cr(lll)precipitates as Cr(OH), or as part of a solid solution withFe(lll), thereby keeping Cr(lll) concentrations below 1 pmole/L(0.05 mg/L). Whether or not natural attenuation at a particularsite is a viable option depends on the characteristics of boththe 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 chromiumenters the subsurface, it displaces the groundwater containingthe dissolved reductants. There is little mixing of the waterscontaining the reducing agents and the Cr(Vl)-contaminantplume. What mixing does occur will be driven by moleculardiffusion at the front of the plume or from the edges of theplume and diffusion from lower permeability lenses containingrelatively immobile water. Thus, aqueous reductants such asFez+ are not going to be important in reducing hexavalentchromium. Mixing of reductants and Cr(Vl) in the plume aregoing to occur primarily though the interactions of the plumewith the immobile soil matrix. Such interactions includedesorption of reductants such as F&+ from mineral surfaces,direct and indirect surface redox reactions between Cr(Vl) andthe mineral surfaces, and reduction by soil organic matter.Thus, it is the soil matrix that is most important with regards toredox transformations of chromium in the subsurface. Thisargument is further supported by studies that clearlydemonstrate that groundwater contributes less that 1 YO of theoxidation capacities (equivalents of Cr oxidized per gram ofsoil) and reduction capacities (equivalents of Cr reduced pergram of soil) of aquifer systems while the soil matrixcontributes the remaining fraction (Barcelona and Helm,1991). Thus, any discussion of redox transformations ofchromium in the subsurface must focus on the soil matrix.

Three key factors must be addressed in considering thepotential use of natural attenuation of Cr(Vl) in the subsurface.Firstly, the reduction capacity of the aquifer, R,, must be greatenough 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 lessthan the total mass of Cr(Vl), M,, that can be reduced by theaquifer 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 thedirection of groundwater flow and R is the dry bulk density ofthe aquifer. As XC increases, the mass of Cr(Vl) that can bereduced increases. A key difficulty in applying this criterion isin providing a reasonable estimate of Mo. In the absence ofother reactions such as adsorption or precipitation, theminimum rate of movement of the Cr(Vl) front through theaquifer, Vtin, computed by assuming the reductant reactsinstantaneously with the Cr(Vl), is

Vmin =Vw

,+ p~ RC

0“ c.

(2)

where p~ and f3v are the dry bulk volumetric water content ofthe porous medium, VW is the velocity of the groundwater, andCO is the concentration of the chromium in the contaminantplume.

CROSS-SECTIONAL POINT OF

AREA, A COMPUANCE

Figure & Cr-plume moving from the source area to the point ofcompliance. The initial Cr(Vl) concentration in the sourcearea is MO, vw is the groundwater velocity, v~in is the velocityof the Cr(Vl) front assuming instantaneous reduction of theCr(Vl), and XC is the distance from the source area to thepoint of compliance.

The second key factor in the application of natural attenuationof Cr(Vl) is the rate of reduction relative to the rate ofadvective transport in the subsurface. The time for thereduction reaction to decrease the concentration from its initialconcentration, CO, to some target concentration, C,, such as adrinking water standard, should be less than the residencetime of the contaminated water in the portion of the aquiferbetween the source of the Cr(Vl) and the point of compliance.For example, if the rate of reduction of Cr(Vl) follows a first-order rate equation

(3)

the time for the concentration of Cr(Vl) to decrease from C. toC, must be less than the residence time of the contaminatedparcel of water within the aquifer

5

In(co /%) < Ek - Vw

If natural attenuation is to be a viable option, this criterionmust be met. Difficulties in utilizing this criterion arise inapplying the appropriate rate equation and obtaining thepertinent rate coefficients.

(4)

A third factor concerning the natural attenuation of Cr(Vl) isthe possible oxidation of Cr(lll) to the more toxic hexavalentform. While contamination is actively entering the subsurface,conditions may favor the reduction of Cr(Vl) to Cr(lll). Afterthe source of the active contamination is removed, however,chemical parameters within the aquifer, particularly pH, maybe altered. Under the new conditions, oxidation of Cr(lll) maybe favored. Thus, soil containing Cr(lll) formed during theactive contamination phase may become a source of Cr(Vl).

Both oxidation and reduction of chromium are occurringsimultaneously within the subsurface as part of a geochemicalcycle. As the Cr(lll) is oxidized to Cr(Vl) by manganesedioxides in the soil, Cr(Vl) can be reduced to Cr(lll) by somereductant such as soil organic carbon or pyrite. The rate ofchange in ICr(Vl)] (d[Cr(Vl)]/dt) is the sum of the rate ofreduction 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 whenR,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) viareduction is balanced by the rate of production by theoxidation of Cr(lll):

-R - ROXrd - (6)

Wittbrodt and Palmer (1 994) suggest that the reduction ofCr(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 andX8 denotes the equivalent fraction of the humic substance thathas been oxidized. Fendorf and Zasoski (1992) suggest thatCrOH2+ is the reactive species in the oxidation of Cr(lll) byMnO,. For illustrative purposes, assume that the oxidationreaction 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 unitvolume of solution. If we further assume that the solution isequilibrated with Cr(OH),(,~), then

ROX = kOX (K /K~)(A/V)m [H+]”+’

where KW is the dissociation constant for H20 and K is theequilibrium 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 equationspresented here are debatable, equation 10 does illustrateaspects of natural attenuation in soils that contain both areductant and Mn02. The key point is that as long as thesupply of reductant and MnOz have not been significantly

depleted, [HCrO~ does not converge to zero with increasingresidence time within the aquifer as one would expect for afirst order reaction that only considers reduction of Cr(Vl).

Rather, [HCrO~ converges to some steady-stateconcentration that is> O that may or may not be above theMCL. This steady-state concentration increases withincreasing kO~k,d, and (A/V) m/[SHS] and it varies with pH.Thus, in principle, if the rate equations are correct and all ofthe parameters are known, one could calculate the steady-state Cr(Vl) concentration and determine if natural attenuationcould achieve compliance goals. Studies of the kinetics ofthese coupled processes needs be done to verify the generalforms of the rate equations and to determine the appropriaterate coefficients.

Determining the Potential for Natural Attenuation

If “natural attenuation” is to be considered an alternative toexpensive remediation efforts, additional characterization isrequired to demonstrate that the expectations are likely to bemet. There is no single test that can tell us if naturalattenuation of Cr(Vl) will occur at a particular site. Severaltests are briefly described which have been utilized to addresskey factors affecting Cr(Vl) transport in the subsurface anddescribe how the results can be utilized in determining thepotential for the natural attenuation of Cr(Vl) in the subsurface.

Ideally, it must be demonstrated that 1 ) there are naturalreductants present within the aquifer, 2) the amount of Cr(Vl)and other reactive constituents do not exceed the capacity ofthe aquifer to reduce them, 3) the rate of Cr(Vl) reduction isgreater than the rate of transport of the aqueous Cr(Vl) fromthe site, 4) the Cr(lll) remains immobile, and 5) there is no netoxidation of Cr(lll) to Cr(Vl). Some of these criteria arerelatively simple while others require additional tests andinterpretation. Additional tests that will be required includetests 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 theaquifer does not exceed the capacity of the soil for reducingthis chromium. Therefore an important first step in evaluatingthe potential for natural attenuation is to determine the massof Cr(Vl) in the soil. Chromium exists in the subsurface eitherin solution or in association with the solid phase. Cr(Vl) insolution can be determined by the diphenylcarbazide (DPC)method (APHA, 1989). Aqueous samples are most oftenobtained from monitoring wells. Alternatively, water separatedfrom the soil matrix either by centrifugation or by squeezing.The pH of these waters should be measured to determine if itis 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 tomineral surfaces (particularly iron oxides) or precipitated aschromate 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 porewater and dissolve highly soluble chromium minerals presentin the soil or that may have precipitated as the result ofevaporation during sample handling and storage. This waterextraction also removes some adsorbed ions.

Following the water extraction, a phosphate extraction is usedas a measure of the “exchangeable” chromate in the soil(Bartlett and James, 1988). The test is conducted by addingphosphate to the soil and equilibrating for 24 hours. Thewater is then separated from the slurry and Cr(Vl) is measuredby the DPC method (Bartlett and Kimble, 1976; Bartlett andJames, 1988c). The increase in the chromate concentration isthe amount of “exchangeable” chromate. Amacher and Baker(1982) found optimal extraction using 0.01 M monobasicpotassium phosphate (KH,PO,). James and Bartlett (1983b)used a solution of 0.005 M KHZPOA and 0.05 M KzHPOi toyield a pH of 7.2. James and Bartlett (1983b) stated thatdoing the extraction at pH 7.2 is preferred because there isless likelihood of chromate reduction than at lower pH.However, decreasing the pH of the soil slurry can result indissolution of BaCrO, from the soil. Moreover, if the pH of thesoil water was initially low, then increasing the pH to 7.2 cancause precipitation of BaC@d, thereby complicating theinterpretation of the results. When the soil water is notequilibrated with BaCrO,, the phosphate extraction method ofJames and Bartlett (1983b) primarily measures the amount ofadsorbed Cr(Vl) in the soil. The phosphate removes chromateby both directly competing for the adsorption sites in the soiland indirectly (in some cases) by increasing the pH.

BaCrO, is a likely chromate mineral phase that can be asource of Cr(Vl) in contaminated aquifers. There is no directtest for BaCrOd in soils, however, when the groundwater isequilibrated with this phase and the source of the Ba2+ isentirely from the clays in the natural soil, the maximumamount of BaCrOd in the aquifer is equal to the ammoniumacetate exchangeable Ba2+ (Thomas, 1982) in background

soils. For example, Palmer and Wittbrodt (1990) found thatthe amount of exchangeable Ba2+ was useful in estimating thenumber of pore volumes required to flush Cr(Vl) from soilcolumns.

At many sites, the total Cr(Vl) associated with the soil matrix isthe sum of the BaCrO, and the PO,-extractable Cr(Vl). Thissum, 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 sumof the aqueous Cr(Vl) and the matrix associated Cr(Vl), S,which can be reported in common units of mass per unitvolume of water by

1000 p~ sCr(Vl)tOt = [Cr(Vl)] +— (11)

0,

where the dry bulk density of the soil, p~, is in g-cm+. Thetotal mass of Cr(Vl) in the site soils can then be estimated byintegrating the concentrations over the volume ofcontaminated soil.

Mass of Cr(lll) in the Subsurface

If all of the chromium that entered the soil was Cr(Vl), thendemonstrating the presence of Cr(lll) in the soil would provethat reduction is occurring. The mass of Cr(lll) in the soil canprovide a measure of the amount of reduction that hasoccurred. Although proof of chromate reduction is necessary,it is not sufficient for demonstrating that natural attenuation willadequately protect the environment.

The total amount of Cr(lll) present in the soil is the sum of themass in solution as well and mass associated with the solidphase. Total chromium in solution can be determined byatomic absorption spectrophotometry (AAS) or inductivelycoupled plasma spectroscopy (ICP). When total chromium isstatistically greater than Cr(Vl), Cr(lll) can be simplydetermined by difference.

The amount of Cr(lll) associated with the soil matrix hasostensibly been determined using several techniques. Anammonium oxalate (0.1 M) extraction serves to removeamorphous hydroxides of Cr, Fe, and Al (Ku et al., 1978;Borggaard, 1988). Bartlett (1991) suggests that a K, H-citrateextraction provides a measure of the Cr(lll) that is potentiallyremovable by low molecular weight organic molecules. Adithionate-citrate-bicarbonate (DCB) extraction is conductedby adding 0.3 M sodium citrate and 0.1 M sodium bicarbonateto the soil sample and heating to 80°C for 20 minutes. Onegram of sodium dithionate is then added and the soil slurry isstirred for another 15 minutes. The DCB extraction removesthe crystalline forms of the Cr-, Fe-, and Al-oxyhydroxides (Kuet al., 1978; Borggaard, 1988). The dithionate reducescrystalline iron (goethite) in the soil and Cr, Fe, and Al arecomplexed by the citrate. In addition to the Cr(lll)oxyhydroxides, the DCB method also extracts sparinglysoluble Cr(Vi) mineral phases such as BaCrO,, therebycomplicating interpretation of the results. Bartlett (1991) uses

7

40 mL per gram of soil of 0.7 M NaOCl solution (undilutedlaundry bleach) at pH 9.5 to extract chromium. The slurry isplaced in a boiling water bath for 20 minutes before the liquidis separated and Cr is determined by AAS or ICP. Thismethod is useful in determining total chromium in the soilbecause it readily oxidizes and removes Cr(lll) that is notremoved by other methods.

Identification of Potential Reductants

The presence of Cr(lll) in the soil maybe indicative of activereduction in the soil, or it may be the result of theneutralization of acidic waters containing Cr(lll) withsubsequent precipitation of chromium hydroxides. Therefore,identification of specific reductants within the aquifer iswarranted. The identification of some potential reductants at asite can be fairly simple in some cases. For example, pyrite(FeS,), a common constituent in many geological materials, isreadily identifiable by its visual characteristics. Other mineralphases capable of reducing Cr(Vl) can be identified usingclassical petrographic techniques or powder x-ray diffraction.Scanning electron microscopy (SEM) can be utilized to identifycrystallite morphology. SEMS equipped with energydispersive x-ray spectroscopy can also provide informationabout the elemental composition of these crystallite. Electrondiffraction patterns obtained from transmission electronmicroscopes provide crystallographic information. Suchelectron microscopy methods can, however, be relativelyexpensive. A fairly simple and inexpensive test for organiccarbon can provide a measure of the amount of carbonavailable for reduction of Cr(Vl). Knowledge of the specificreductant within the aquifer is useful in determining the timescale for the reduction of Cr(Vl) based on studies that arereported in the literature. Soils containing iron sulfides ororganic matter are more likely to reduce Cr(Vl) on the timescales of interest than soils containing ferrous iron silicates.

Reduction Capacity of the Aquifer

Adequate protection of the environment by natural attenuationof Cr(Vl) requires that the soil possess a large enoughreducing capacity to reduce all the hexavalent chromium in thesource area. Several measures for predicting reduction ofCr(Vl) in soil are presented by Bartlett and James (1 988) andBartlett (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 classicalWalkley-Black method for determining soil organic carbon(Bartlett and James, 1988). In this method, 2 to 3 grams ofsoil are reacted with a mixture of 1 N K2CrzOT in IN H#Od for30 minutes (Walkley and Black, 1934; Nelson and Sommers,1982; Bartlett and James, 1988). The Cr(Vl) concentration ismeasured using the diphenylcarbazide (DPC) method (APHA,1989) and the decrease in the mass of Cr(Vl) in the reactionvessel per gram of soil used in the test is the reductioncapacity. Although this method of determining soil organiccarbon has its limitations (e.g. Nelson and Sommers, 1982), itis a direct measure of how much Cr(Vl) can be reduced by asoil at extreme acid concentrations. Variations on this method

use heat or a combination of heat and pressure (Nelson andSommers, 1982). Barcelona and Helm (1991) used amodified closed-tube chemical oxygen demand procedure(U.S. EPA, 1979) to determine reduction capacities.

The extreme conditions of pH and temperature used in thetotal Cr(Vl) reducing capacity test may yield a greater reducingcapacity than would be available under most environmentalconditions. The “available reducing capacity” test ofBartlett and James (1988) determined the reduction capacityby reacting about 4 to 5 grams of moist soil in a solution of10 mM H, PO, and K2Cr20, for 18 hours. The H~PO, is addedto buffer the pH and to compete with the Cr(Vl) for theadsorption sites. When KH2POd is used Bartlett and Jamesreferred to it as the “reducing intensity”. These tests aredesigned to determine the reducing capacity at pH valuesmore likely to be encountered in the field. However, in longterm reduction tests at near neutral pH, Palmer and Wittbrodt(unpublished data) observed reduction occurring after 250days (Fig. 7). Such long-term reduction tests are not practicalat most waste sites.

0.08n=g~ 0.06~

: 0.04

~~ 0.02wK

~ 0.00

50 100 150 200 250 300TIME (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 theoxidizing capacity of the soil is greater than the reductioncapacity, then as the chromium is cycled in the soil it couldexhaust the soil reductant and be oxidized and ultimatelymobilized in the soil. It is important, therefore, to determinethe capacity of the aquifer to oxidize Cr(lll).

Bartlett and James (1988) suggest a relatively simple test forthe amount of Cr(lll) that can be oxidized by a soil. Themethod involves adding 2.5 grams of soil to a solutioncontaining 25 ml of 1 mM CrCl~. After shaking for 15 minutes,

a solution of K&POd . K2HPOA is added to the reactionvessel, the slurry centrifuged or filtered, and the Cr(Vl)

8

measured using the DPC method. Moist soils should be usedin these tests. Drying the soils alters the surfaces of themanganese dioxides making them less reactive (Bartlett andJames, 1979).

Barcelona and Helm (1991) used a solution of chromous(Cr(ll)) ion to measure the oxidation capacity of soils. Theyadded about 1 g of soil to cuvettes containing the Cr(ll)solution. The work was performed in a glove box to preventoxygen from reacting with the Cr(ll). The cuvettes weresealed, shaken, and allowed to react for 2 hours. Thesamples were centrifuged and the Cr(ll) measuredspectrophotometrically. The loss of Cr(ll) is then used as ameasure of the oxidation capacity of the soils.

Each of these methods has some problems. Palmer andWittbrodt (unpublished data) conducted oxidation tests similarto the Bartlett and James (1 988) method except that theconcentration of Cr(Vl) was monitored over nearly a year.Cr(Vl) concentrations in these tests continued to increase upto 100 days and may have been continuing after 300 days.For one soil, the short-term oxidation test of Bartlett andJames underestimated the amount of oxidation obtained in thelong-term tests by more than an order of magnitude. It is notclear in the Barcelona and Helm method whether the Cr(ll) isbeing oxidized to Cr(lll) or Cr(Vl) or some combination of thetwo. If both products are forming, the results are more difficultto interpret.

If the only mechanism for the oxidation of Cr(lll) in soils isoxidation by manganese oxides, then using extractionmethods specifically designed for this purpose may be a goodway of determining the oxidation capacity of the soils. Onevery simple extraction technique (Chao, 1972; Gambrell andPatrick, 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 30minutes and the Mn concentration is measured. The numberof moles of Cr(lll) that can be oxidized is then computed bydividing the number of moles of Mn2+ per gram of soil obtainedin the extraction by 1.5. The hydroxylamine hydrochloride testis fast, easy, and specifically targets the phase that promotesthe oxidation of Cr(lll).

Rates of Oxidation and Reduction

Key factors in the suitability of natural attenuation as an optionfor chromium contaminated soils are the rates of oxidation andreduction of chromium. This information is the most difficult toobtain. Scientists are only now learning about the form of theaDDlicable rate eauations and the armomiate rate coefficients

reductant and pH, it is important to obtain rate coefficients thatwere acquired under conditions similar to those at the site.Many rate studies have considered only a limited set ofconditions such as a single pH value or one reductantconcentration. Consequently, the reported rate coefficientsare apparent values that are strictly valid only under theconditions of the experiment. Thus, the experimental factorsmust be taken into account before the rate coefficients can beapplied 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 obtainedfrom experiments that only utilize reductants. Further, manyrate experiments are conducted in stirred reactors that canabrade reactive surfaces. In soils, the rate of reaction maybecome surface limited as adsorbed ions and precipitatescover the reactive surfaces.

One method of obtaining the net rate of reduction is throughtests on uncontaminated soils obtained from the site. Thesesoils should be similar to those through which the contaminantplume will be migrating. Cr(Vl) can be added to the soil slurryand the Cr(Vl) concentrations monitored over time. Thereaction vessels must exclude light to prevent photoreductionreactions and the slurry must have the same pH as thecontaminant plume. A key limitation to such experiments isthat they require several months to a year to complete.

Estimating Reduction from Monitoring Well Data

In principle, Cr(Vl) reduction can be estimated from thedecrease in the mass of Cr(Vl) in the aquifer (e.g., Henderson,1994). The key difficulty in such an approach is to estimatethe mass of Cr(Vl) using the aqueous concentrations. Thetotal mass of Cr(Vl) in the aquifer is the sum of the mass thatis in solution, the mass that is adsorbed to the aquifer matrix,and the mass that is precipitated within the aquifer. The massof 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 aCr(Vl) concentration of C.

The mass of Cr(Vl) adsorbed to the soil matrix, M,m, can becomputed from the adsorption isotherm. For example, ifCr(VI) follows a Langmuir isotherm, then over a volume ofaquifer, V, with constant aqueous Cr(Vl) concentration, C, M,bcan be computed as

I that may apply. Such kinetic studies are an area whereresearch has lagged behind the practical need for theinformation. While rates can be obtained from the technicalliterature, one must be careful to use rates of reduction formaterials that are most likely controlling the Cr(Vl) reduction inthe site soil. In addition, because the rates depend on theconcentration (surface area per liter of solution) of the

[ 1Md, . Pb :;aKbscc V(l-r3v)‘KS

(13)

where K~& is the Langmuir adsorption constant, S- is themaximum amount of contaminant that can be adsorbed to thesoil.

There is no unique amount of Cr(Vl) precipitate for a givenhexavalent chromium concentration. Therefore it isimpossible to estimate mass of this fraction of Cr(Vl) is thesubsurface using only the measured concentrations inmonitoring wells. Thus, natural attenuation of Cr(Vl) frommass balances using monitoring well data can only be usedwhen it can be reasonably demonstrated the Cr(Vl)precipitates cannot form within the aquifer.

Even when it is demonstrated that the formation of precipitateswithin the aquifer is unlikely, there are inherent problems withany monitoring system that can create uncertainties in theestimated mass of Cr(Vl) during a sampling round. In thethree-dimensional flow field, the highest concentrations fromone sampling period may migrate between the discretemonitoring points of the next sampling round. The undetectedmass is not included in the total mass estimates in the secondsampling round and may be mistakenly interpreted as massloss due to Cr(Vl) reduction.

Summary

Under certain conditions, toxic Cr(Vl) can be reduced to theless toxic Cr(lll) in soils and precipitated as an insolublehydroxide phase. The possibility of relying on such “naturalattenuation” of Cr(Vl) is attractive because of the greatexpense of remediating chromium contaminated sites. Beforesuch an option is adopted, however, it should bedemonstrated that natural attenuation is likely to occur underthe specific conditions at the site being investigated.

If natural attenuation is to be considered a viable option forchromium contaminated sites, then ideally, it must bedemonstrated that 1) there are natural reductants presentwithin the aquifer, 2) the amount of Cr(Vl) and other reactiveconstituents do not exceed the capacity of the aquifer toreduce them, 3) the time scale required to achieve thereduction of Cr(Vl) to the target concentration is less than thetime scale for the transport of the aqueous Cr(Vl) from sourcearea to the point of compliance, 4) the Cr(lli) will remainimmobile, and 5) there is no net oxidation of Cr(lll) to Cr(Vl).The most difficult information to obtain is the time scales forthe reduction and oxidation of chromium in the soil.

Demonstrating Cr(Vl) reduction in aquifer by mass balancesthat rely primarily on the aqueous concentrations frommonitoring well networks are valid only if it is demonstratedthat Cr(Vl) precipitates are not forming in the aquifer. Themonitoring network must be sufficiently dense that estimatesof Cr(Vl) are accurate.

Several soil tests are described that are useful in determiningthe mass of Cr(Vl) and Cr(lll) in the source areas and thereduction and oxidation capacities of the aquifer materials.Some simple conceptual models are presented whereby thisinformation, combined with knowledge of the residence time ofthe chromium between the source and the point of compliancecan be used to determine the feasibility of natural attenuation

of Cr(Vl). The major limitation to this approach is the lack ofinformation about the rate of oxidation and reduction ofchromium under conditions likely to encountered by plumesemanating from chromium sources. Without better informationabout these rate processes under a wider range of conditionswith respect to pH, the use of the natural attenuation option forcontaminated soils will continue to be a highly debated issue.

References

Abbasi, S.A. and R. Soni, 1984. Teratogenic Effects ofChromium(VI) in the Environment as Evidenced by the Impactof Larvae of Amphibian Rand tigrina: Implications in theEnvironmental Management of Chromium. Int. J.Environmental Studies, 23: 131-137.

Allison, J. D., D.S. Brown, K.J. Novo-Gradac, 1990.MlNTEQA2/PRODEFA2, A Geochemical Assessment Modelfor Environmental Systems: Version 3.0. U.S. EnvironmentalProtection Agency, Athens, GA.

Amacher, M.C. and D.E. Baker, 1982. Redox ReactionsInvolving Chromium, Plutonium and Manganese in Soils.DOE/DP/0451 5-1. Institute for Research on Land and WaterResources, Pennsylvania State University and U.S.Department of Energy, Las Vegas, NV.

Amonette, J.E. and D. Rai, 1990. Identification ofnoncrystalline (Fe, Cr)(OH)3 by Infrared Spectroscopy. Claysand Clay Minerals, 38(2): 129-136.

APHA, 1989. Standard Methods for the Examination of Waterand Wastewater, 17th Edition, L.S. Slesceri, A.E. Greenberg,and R.R. Trussell (Editors). American Public HealthAssociation, Washington, D.C.

Barcelona, M.J. and T.R. Helm, 1991. Oxidation-ReductionCapacities of Aquifer Solids. Environmental Science andTechnology, 25:1565-1572.

Bartlett, R. J., 1991. Chromium Cycling in Soils: Links, Gaps,and Methods. Environmental Health Perspectives, 92: 17-24.

Bartlett, R.J. and J.M. Kimble, 1976. Behavior of Chromium inSoils: 11: Hexavalent Forms. J. Environmental Quality,5(4):383-386.

Bartlett, R.J. and B.R. James, 1979. Behavior of Chromium inSoils: Ill. Oxidation. J. Environmental Quality, 8(1):31-35.

Bartlett, R.J. and B.R. James, 1988. Mobility andBioavailability of Chromium in Soils. IN: Chromium in theNatural and Human Environments, Vol. 20 (J.O. Nriagu and E.Nieboer, editors). John Wiley& Sons, New York: 267-306.

Beyersmann, D.A. Koester, B. Buttner, and P. Flessel, 1984.Model Reactions of Chromium Compounds with Mammalianand Bacterial Cells. Toxicol. Environ. Chem., 8: 279-286.

10

Bianchi, V., A. Zantedeschi, A. Montaldi, and F. Majone, 1984.Trivalent Chromium is Neither Cytotoxic nor Mutagenic inPermealized Hamster Fibroblasts. Toxicological Letters, 23:51-59.

Bloomfield, C. and G. Pruden, 1980. The Behavior of Cr(Vl) inSoil under Aerobic and Anaerobic Conditions. EnvironmentalPollution (Series A), 23: 103-114.

Blowes, D.W. and C.J. Ptacek, 1992. GeochemicalRemediation of Groundwater by Permeable Reactive Walls:Removal of Chromate by Reaction with Iron-Bearing Solids.Proceeding of the Subsurface Restoration Conference, June21-24, 1992, Dallas, TX: 214-216.

Bonatti, S., M. Meini, and A. Abbondandolo, 1976. GeneticEffects of Potassium Chromate in Schizosaccharomycespornbe. Mutat. Res., 38: 147-149.

Borggaard, O. K., 1988. Phase Identification by SelectiveDissolution Techniques. IN: Iron in Soils and Clay Minerals,J.W. Stucki et al. (editors). Reidel Publishing Co., pp. 83-98.

Boyko, S.L. and D.M.L. Goodgame, 1986. The Interaction ofSoil Fulvic Acid and Chromium (Vi) Produces Relatively Long-Lived Water Soluble Chromium(V) Species. Inorg. Chim Acta,123: 189-191.

Calder, L. M., 1988. Chromium Contamination ofGroundwater. IN: Chromium in the Natural and HumanEnvironments, Vol. 20 (J.O. Nriagu and E. Nieboer, editors).John Wiley & Sons, New York: 215-230.

Chao, T.T, 1972. Selective Dissolution of Manganese Oxidesfrom Soils and Sediments with Acidified HydroxylamineHydrochloride. Soil Science Society of America Proceedings,36:764-768.

Eary, L.E. and D. Rai., 1989. Kinetics of Chromate Reductionby Ferrous Ions Derived from Hematite and Biotite at 25”C.American Journal of Science, 289: 180-213.

Eary, L.E. and D. Rai, 1987. Kinetics of Chromium(lll)Oxidation to Chromium(VI) by Reaction with ManganeseDioxides. Environmental Science and Technology,21 (12):1 187-1193.

Eary, L.E. and D. Rai, 1988. Chromate Removal fromAqueous Wastes by Reduction with Ferrous Iron.Environmental Science and Technology, 22(8): 972-977.

Fendorf, S.E. and R.J. Zasoski, 1992. Chromium (Ill)Oxidation by 13-MnOz. 1. Characterization. EnvironmentalScience and Technology, 26: 79-85.

Gambrell, R.P. and W.H. Patrick, 1982. Manganese. IN:Methods of Soil Analysis, Part 2, Chemical and MicrobiologicalProperties, Second Edition. A.L. Page (editor). Agronomy,

No. 9, Part 2, American Society of Agronomy, Soil ScienceSociety of America, Madison,Wl:313-322.

Goodgame, D. M. L., P.B. Hayman, and D.E. Hathway, 1984.Formation of Water Soluble Chromium(V) by the Interaction ofHumic Acid and the Carcinogenic Chromium(VI). Inorg. Chim.Acta, 91:1 13-115.

Henderson, T., 1994. Geochemical Reduction of HexavalentChromium in the Trinity Sand. Ground Water 32(3): 477-486.

James, B.R. and Bartlett, R.J., 1983a. Behavior of Chromiumin Soils. V1. Interactions Between Oxidation-Reduction andOrganic Complexation. J. Environmental Quality 12: 173-176(1983).

James, B.R. and Bartlett, R. J., 1983b. Behavior of Chromiumin Soils: V1l. Adsorption and Reduction of hexavalent forms. J.Environmental Quality 12(2): 177-181.

Ku, H. F. H., B.G. Katz, D.J. Sulam, and R.K. Krulikas, 1978.Scavenging of Chromium and Cadmium by Aquifer Material,South Farmingdale Massapequa Area, Long Island, NewYork. Ground Water 16 (2): 112-118.

Lancy, L. E., 1966. Treatment of Spent Cooling Waters. U.S.Patent 3,294,960.

Mancuso, T. F., 1951. Occupational Cancer and Other HealthHazards in a Chrome Plant. A Medical Appraisal. Il. Clinicaland Toxicological Aspects. Ind. Med. Surg. 20: 393-407.

Mancuso, T.F. and W.C. Heuper, 1951. Occupational Cancerand Other Health Hazards in a Chrome Plant. A MedicalAppraisal. L Lung Cancers in Chromate Workers. Ind. Med.Surg., 20: 358-363.

Martin, C., D.R. Boone, and C.D. Palmer, 1994. Chromate-Resistant Microbes from Contaminated Soil and TheirPotential for Bioaugmented Reduction of Cr(Vl). Proceedingof the Eighth National Outdoor Action Conference andExposition, Minneapolis, MN, May 23-25, 1994. NationalGround Water Association: 191-204.

Nelson, D.W. and L.E. Sommers, 1982. Total Carbon, OrganicCarbon, and Organic Matter. IN: Methods of Soil Analysis,Part 2 (R.C. Dinauer, editor), American Society of Agronomy,Inc. and Soil Science Society of America, Inc., Madison, Wl:539-580.

Nriagu, J. O., 1988a. Historical Perspectives. IN: Chromiumin the Natural and Human Environments, Vol. 20 (J.O. Nriaguand E. Nieboer, editors). John Wiley & Sons, New York: 1-20.

Nriagu, J. O., 1988b. Production and Uses of Chromium. IN:Chromium in the Natural and Human Environments, Vol. 20(J.O. Nriagu and E. Nieboer, editors). John Wiley& Sons,New York: 81-104.

11

One, B.-L, 1988. Genetic Approaches in the Study ofChromium Toxicity and Resistance in Yeast and Bacteria. IN:Chromium in the Natural and Human Environments, Vol. 20(J.O. Nriagu and E. Nieboer, editors). John Wiley& Sons,New York: 351-368.

Palmer, C.D. and W. Fish, 1992. Chemically EnhancedRemoval of Metals from the Subsurface. Proceeding of theSubsurface Restoration Conference, June 21-24, 1992,Dallas, TX: 46-48.

Palmer, C.D. and P.R. Wittbrodt, 1990. GeochemicalCharacterization of the United Chrome Products Site, FinalReport. IN: Stage 2 Deep Aquifer Drilling Technical Report,United Chrome Products Site, Corvallis, OR, September 28,1990, CH2M Hill, Corvallis, OR.

Palmer, C.D. and P.R. Wittbrodt, 1991. Processes Affectingthe Remediation of Chromium-Contaminated Sites.Environmental Health Perspectives, 92: 25-40.

Paschin, Y.V., V.I. Kozachenko, and L.E. Sal’nikova, 1983.Differential Mutagenic Response at the HGPRT Locus in V-79and CHO Cells after Treatment with Chromate. Mutat. Res.122: 361-365.

Rai, D., B.M. Sass, and D.A. Moore, 1987. Chromium(lll)Hydrolysis Constants and Volubility of Chromium(lll)Hydroxide Inorg. 26(3): 345-249.

Rai, D., J.M. Zachara, L.E. Eary, C.C. Ainsworth, J.E.Amonette, C.E. Cowan, R.W. Szeimeczka, C.T. Resch, R.L.Schmidt, S.C. Smith, and D.C. Girvin, 1988. ChromiumReactions in Geologic Materials. Electric Power Institute Res.Inst. Rept. EA-5741, Palo Alto, CA, 287 pp.

Richard, F.C. and A.C.M. Bourg, 1991. AqueousGeochemistry of Chromium: A Review. Water Research,25(7): 807-816.

Riser, J.A. and G.W. Bailey, 1992. Spectroscopic Study ofSurface Redox Reactions with Manganese Oxides. SoilScience Society of America Journal, 56: 82-88.

Sass, B.M. and D. Rai, 1987. Volubility of AmorphousChromium(lll)-lron( ill) Hydroxide Solid Solutions. Inorg.Chem., 26(14): 2228-2232.

Schott, J. R.A. Berner and E. L. Sjoberg, 1981. Mechanism ofPyroxene and Amphibole Weathering–1. ExperimentalStudies with Iron-free Minerals. Geochimica et CosmochimicaAct, 45: 2123-2135.

Schroeder, D.C. and G.F. Lee, 1975. PotentialTransformations of Chromium in Natural Waters. Water, Air,Soil Pollution, 4: 355-365.

Stollenwerk, K.G. and D.B. Grove, 1985. Reduction ofHexavalent Chromium in Water Samples Acidified forPreservation. J. Environmental Quality, 14(3): 396-399.

Swayambunathan, V., Y.X. Liao, and D. Meisel, 1989. Stagesin the Evolution of Colloidal Chromium(lll) Oxide. Langmuir,5(6): 1423-1427.

Thomas, G.W., 1982. Exchangeable Cations. IN: Methods ofSoil Analysis, Part 2, Chemical and Microbiological Properties,Second Edition. A.L. Page (editor). Agronomy, No. 9, Part 2,American Society of Agronomy, Soil Science Society ofAmerica, Madison, Wl: 159-165.

U.S. Environmental Protection Agency, 1979. ChemicalOxygen Demand, EPA Method 410.4, Methods for ChemicalAnalysis of Water and Wastes, U.S. Government PrintingOffice: Washington, D.C.

van Weerelt, M., W.C. Pfeiffer, and M. Fiszman, 1984. Uptakeand Release of 5’Cr(Vl) and 51 Cr(lll) by Barnacles. (E?a/arrussp). Mar. Environ. Res. 11 :201-211.

Walkley, A. and LA. Black, 1934. An Examination of theDegtjareff Method for Determining Soil Organic Matter and aProposed Modification of the Chromic Acid Titration Method.Soil Science, 37:29-38.

Waterhouse, J.A.H., 1975. Cancer among Chromium Platers.Br. J. Cancer, 32: 262.

White, A.F. and M.F. Hochella, 1989. Electron TransferMechanisms Associated with the Surface Oxidation andDissolution of Magnetite and Ilmenite. Proc. 6th InternationalSymposium on Water-Rock Interaction: 765-768.

White, A.F. and A. Yee, 1985. Aqueous Oxidation-ReductionKinetics Associated with Coupled Electron-Cation Transferfrom Iron Containing Silicates at 25°C. Geochimica etCosmochimica Acta, 49: 1263-1275.

Wittbrodt, P.R. and C.D. Palmer, 1992. Limitations to Pump-and-Treat Remediation of a Chromium Contaminated Site.Aquifer Restoration: Pump-and-Treat and the Alternatives.National Ground Water Association National Convention, LasVegas, NV, Sept. 30- Oct. 2, 1992.

Wittbrodt, P.R. and C.D. Palmer, 1994. Reduction of Cr(Vl) inthe Presence of Excess Soil Fulvic Acid. EnvironmentalScience and Technology, (accepted for publication).

Yassi, A. and E. Nieboer, 1988. Carcinogenicity of Chromiumcompounds. IN: Chromium in the Natural and HumanEnvironments (J.O. Nriagu and E. Nieboer, editors): 443-496.

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