peer reviewed: in-situ treatment of chromium-contaminated groundwater

Although not as common as solvent or fuelproducts, chromium contamination ofgroundwater is relatively widespread. TheU.S. EPA has estimated that as many as1300 sites in the United States may have

groundwater contaminated with chromium—someof which date to World War II.

As with most groundwater contaminants, chromi-um contamination is traditionally remediated with anextraction and treatment system—“pump and treat”.Groundwater is extracted from the aquifer through a

well network and conveyed to an aboveground treat-ment plant, where the chromium is removed usinganion exchange or precipitated with various treatments.

In many ways, chromium contamination has beenan ideal candidate for pump and treat, because themost common chemical form, hexavalent chromium[Cr(VI)], is highly soluble and not readily adsorbedonto sediment surfaces. However, because of well-known limitations inherent to methods that requiregroundwater extraction, such as exponentially de-creasing response to treatment and diffusion-limited

464 A ■ ENVIRONMENTAL SCIENCE & TECHNOLOGY / DECEMBER 1, 2002 © 2002 American Chemical Society

DECEMBER 1, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY ■ 465 A

rates of extraction, pump-and-treat remediation ofchromium-contaminated groundwater is not alwayssatisfactory (1–3). Thus, alternative treatment meth-ods have been developed. Most are in situ methodsthat treat the impacted groundwater in the aquiferand eliminate the extraction step. This article pro-vides an overview of this technology.

Chromium primer Chromium has three relatively stable valence states:Cr(0) (the metallic state), Cr(III), and Cr(VI). Other va-

lence states are possible, but are not very stable. Cr(III)is relatively insoluble in water under common envi-ronmental conditions (pH 6–9), forming hydroxidesand oxyhydroxides alone, and a solid solution withiron (4, 5). However, Cr(VI) is quite soluble and mo-bile in the environment.

Chromium has various industrial uses, includingchrome plating, steel making, corrosion inhibition,wood preservation, well drilling (as a fluid additive),biocides, and paint and primer pigments. Smeltersthat produce chromium can also be a source of con-

In Situ Treatment of

CHROMIUM-CONTAMINATED

GROUNDWATER

New technologies

show promise for

removing chromium(VI)

pollution at lower cost.

J O N A T H A N F R U C H T E R

U.S.

DEP

ARTM

ENT

OF E

NER

GY

INSE

TS: P

ACIF

IC N

ORTH

WES

T N

ATIO

NAL

LAB

ORAT

ORY

tamination, because the recovery of chromium fromchromite ores requires oxidation to Cr(VI) before con-version to other forms.

Cr(VI) is usually shipped commercially in thedichromate form (Cr2O7

2–), because the chromium ismore concentrated. However, in dilute, near-neutralpH aqueous solution, Cr(VI) commonly forms hy-drochromate (HCrO4

–) and chromate (CrO42–) anions.

These anions are generally poorly adsorbed onsoils and sediments because they already have nu-merous uncharged or negatively charged surface sitesat near-neutral pHs. Because they are sparingly sol-uble, barium chromate precipitates may form undersome circumstances. However, the hydrochromateand chromate ions generally are not retarded but flowunimpeded in groundwater aquifers (6).

Cr(VI) concentrations in groundwater in theUnited States are federally regulated under the SafeDrinking Water Act, with a maximum contaminantlevel of 0.100 mg/L as total chromium, and under theClean Water Act with an ambient water quality crite-ria of 0.011 mg/L as chromium. Some U.S. states andother countries have implemented more stringentstandards than the U.S. government for chromiumin groundwater. There is continued uncertainty as tothe appropriate drinking water standard for bothhexavalent and total chromium.

The in situ approach In situ remediation of chromium-contaminatedgroundwater involves chemically or biologically re-ducing Cr(VI) to Cr(III), which is less toxic, less solu-ble, and less mobile than Cr(VI). In addition, Cr(III)can then precipitate as a hydroxide, usually as a solidsolution with ferric iron hydroxide, and will be effec-tively immobilized (7, 8). This reduction is usually apermanent solution, because Cr(III) is not easily re-oxidized to Cr(VI) under conditions that occur in mostnatural groundwater environments.

Cr(III) oxides and hydroxides occur naturally insmall concentrations in the sediment and soil in manygroundwater aquifers (crustal average concentration= 102 ppm) (9). Although manganese oxides could re-oxidize Cr(III), Cr(VI) is rarely detected in theseaquifers. In many instances, the chromium oxidationby manganese oxides in soils and sediments appearsto be limited by surface alteration effects (10).

Various approaches to reducing Cr(VI) in situ havebeen developed and tested. These methods usually in-volve adding some already reduced compound to actas a source of electrons.

System selection and designThe goal of any in situ aquifer treatment method isto deliver an appropriate reagent to the contamina-tion in the aquifer. Therefore, one way to classify thedifferent treatment options is by reagent and deliv-ery system. The correct option is then a matter ofchoosing the right combination for the site condi-tions. To chemically reduce Cr(VI), reagents usuallyconsist of reduced forms of three elements: carbon formost biological remediation systems and iron or sul-fur for abiotic approaches and certain specialized bi-ological systems. Delivery systems include trenches,

infiltration galleries, groundwater wells, and direct-push injection.

Choosing the correct approach begins with char-acterizing the site by geochemical, hydrological, andgeological means, along with describing the natureand distribution of the contaminants. These charac-teristics are tied together with a site conceptualmodel. Frequently, a numerical model is used tomake the site description more quantitative.Treatability studies in the lab and sometimes at the

field scale also help facilitate the selection process.The following are examples of successful remediationsystems.

In situ treatment methodsThese methods involve abiotic approaches, usuallyinvolving reduced iron or sulfur compounds as theelectron donor, and sometimes both.

Permeable reactive barriers. Permeable reactivebarriers, or treatment walls, treat groundwater as itflows away from the source and through the aquifer.The permeable barrier cuts off only the flow of con-taminants, but not the groundwater. In the trench-and-fill barrier configuration, a trench is excavatedand filled with a chemically reactive medium. Caremust be taken to ensure that the hydraulic conduc-tivity is equal to or higher than that of the surround-ing aquifer to allow groundwater to flow through thetreatment zone. As the water flows through the zone,the chromium is reduced by the reactive medium andsubsequently precipitates as a Cr(III) hydroxide.

Because permeable reactive barriers are passive,their operation and maintenance costs are low.Another benefit is that the contaminants that requiretreatment and disposal are not brought to the sur-face. Barriers can be used even if the contaminantsource has not been identified or well characterized,and the natural groundwater flow pattern is unaltered.

However, this approach cannot treat contaminantsources and is not suitable for all geologic or hydro-logic regimes. For example, hydraulic conductivities

466 A ■ ENVIRONMENTAL SCIENCE & TECHNOLOGY / DECEMBER 1, 2002

FIGURE 1

Cross-sectional schematic of a trench-and-fill permeable reactive barrierThe barrier, which uses metallic iron to reducecontaminants, has found some use in treatingchromium.

Fill

Reactive cell

Treatedgroundwater

Aquitard

Contaminantplume


that are low may not permit sufficient groundwaterflow to treat significant volumes of contaminatedwater. Groundwaters that are very high in dissolvedconstituents, such as calcium, may form precipitatesthat clog the barrier’s pores. Another limitation is thatthe barrier must eventually be replaced. If this hap-pens too often, the cost may be higher than alterna-tive approaches.

Iron particle barriers. These barriers are the mostcommon form of permeable systems. They rely onmetallic iron [Fe(0)] for chemically reducing contam-inants (Figure 1). Although originally developed totreat chlorinated organic solvents, they have foundsome use in treating chromium and other metalliccontaminants (11, 12). Particulate iron has the ad-vantage that each atom can donate up to three elec-trons, giving the barrier significant redox (reduction/oxidation) capacity.

When iron is present, Cr(III) can precipitate as amixed iron–chromium hydroxide, which has a lowersolubility than pure chromium hydroxide. This type ofbarrier is most frequently emplaced using trenchingtechniques, although other methods have been used.Often sheet piles are used to facilitate the installationof the barrier. Barriers emplaced by trenching havemost commonly been restricted to depths of 10 m orless below the surface (13). Various methods for in-stallation at deeper depths have been investigated, in-cluding vibrating beams and jet grouting, but these

become increasingly difficult at greater depths. In ad-dition, the high pHs that form in these barriers maylead to precipitation of various minerals, with subse-quent plugging, so Fe(0) barriers rarely use their fullredox capacity (14). Nevertheless, iron particle barri-ers for Cr(VI) reduction appear to be operating suc-cessfully in the United States (15). Some difficultieswith passivation of the iron particles in a Cr(VI) bar-rier have been reported by Danish scientists (16).

Other permeable reactive barriers. Other forms ofpermeable reactive barriers have been used to treatchromium contamination. One particularly inexpen-sive alternative uses sawdust, compost, and limestone(17 ). This barrier is actually a cross between chemi-cal and biological methods, because the compost actsas a carbon source for microbial populations, whichare active in the reduction of the chromium. Naturallyoccurring zeolite coated with cationic surfactants hasalso been investigated as an adsorbant for chromiumin permeable reactive barriers (18).

In situ redox manipulation (ISRM). ISRM tech-nology creates a permeable subsurface treatmentzone to reduce mobile chromium in groundwater toan insoluble form. The permeable treatment zone iscreated by reducing Fe(III), which is present as sur-face oxides, to Fe(II) within the aquifer sediments(19). Some of the Fe(III) in 2:1 smectite clays is alsoreduced by injecting sodium dithionite (Na2S2O4) intothe aquifer (Figure 2) (20). Sodium dithionite is a

FIGURE 2

In situ redox manipulation (ISRM) processThis technology creates a permeable subsurface treatment zone in aquifer sediments, where mobile chromium ingroundwater is reduced to a less soluble and mobile form. RM-X are monitoring wells.

Mobile field labInjection solution

RM-2RM-5

RM-7RM-6

RM-8 RM-1a

RM-1b RM-4RM-3

RM-9Injection

well

Static viewer level

Vadose zone

Office/storage/trailer

Low-permeability unit

Permeabletreatment zone

Groundwater flow

Contaminantplume fromupgradientsource

High-permeability unit

Low-permeability unit

strong reducing agent that has several desirable char-acteristics for this type of application, including in-stability in the natural environment and reaction anddegradation products that ultimately oxidize to sul-fate. This instability is beneficial because it meansthat the reaction period is rapid, and that after aperiod of several days, no dithionite remains in the

aquifer. Potassium carbonate/bicarbonate is addedto the injection solution as a pH buffer to enhancethe stability of dithionite during the reduction of avail-able iron (by buffering H+ generated during ironreduction).

As with permeable reactive barriers, an ISRM treat-ment zone is placed perpendicular to the groundwa-ter flow to intercept the contaminant plume. Thisgeometry is created by a series of overlapping injec-tion/withdrawal wells. Advantages include the use ofconventional groundwater wells, which leads to eas-ier installation at greater depths, and the ability of asingle injection of dithionite to create a treatmentzone that will last for many years. The technology islimited to clastic (silt, sand, and gravel) aquifers thathave sufficient hydraulic conductivity to allow thereagent injection (>10–2 cm/s). The aquifer materialsmust also contain at least small amounts of reactiveiron compounds (0.01–0.1%), although this is not usu-ally a concern. Most aquifers contain considerablymore total iron than this, but only a fraction of it isreactive. The longevity of the barrier depends on sev-eral factors, including the amount of reactive iron,the concentration of oxygen and contaminant—bothof which will reoxidize the barrier—in the ground-water, and the groundwater flow velocity.

Chemically enhanced pump and treat. Also calledgeochemical fixation, chemically enhanced pumpand treat adds a chemical reducing agent to the treat-ed groundwater before it is reinjected into the aquifer.In this way, the residual Cr(VI) that is not actually re-moved during the groundwater extraction phase canbe treated in situ, alleviating some of the problemsof conventional pump and treat (21).

The reagent of choice, usually sodium metabisul-fite, Na2S2O5 , or calcium polysulfide, CaSx, is a func-tion of site geochemistry. Ferrous sulfate and sodiumbisulfide have also been used. However, ferrous sul-fate injection can lower pH as a result of reactions

that are similar to acid mine drainage, and sulfidescan cause precipitation, which may clog the aquifer.

Chemically enhanced pump and treat can be usedto treat source areas. Another advantage is that treat-ed water does not need to be discharged to the sur-face. However, the approach needs a sufficientlypermeable aquifer, removal of reduction reactionproducts if they become too concentrated, and reg-ulatory approval to reinject treated water.

Electrochemical methodsElectrochemical or electokinetic remediation placesa series of electrodes into the contaminated zone, towhich a low-voltage (50–150 V), direct current chargeis applied (22). Contaminant ions in the water willmigrate toward the electrode of opposite charge,which is called electromigration. Because hydrogenions will migrate, the pH will decrease at the anodeand increase at the cathode.

For groundwater remediation, the electrodes cansimply be placed in slotted nonmetallic wells, suchas those made of polyvinyl chloride. The drift veloc-ities of the contaminant ions are relatively slow,around 1 cm per day, so that the electrokineticmethod is not applicable to fast-moving groundwa-ters. The slow drift velocity also requires relativelyclose well spacing, another potential limitation.Attempts to increase the drift velocity using highervoltages can lead to problems with soil heating.However, it can be useful for treating unsaturatedsoils and slow-moving groundwater in tight aquifersin which the permeability is too low to permit othertypes of in situ remediation. Highly mobile anionssuch as chromium are good candidates for electroki-netic remediation, because they drift through theaquifer with little or no adsorption. A successful pilot-scale demonstration of the electrokinetics technolo-gy for vadose-zone Cr(VI) contamination has beenconducted, but there have been no reported deploy-ments for groundwater chromium contamination todate (23).

Biological in situ methodsReduction of Cr(VI) by living organisms either canoccur inside the cell or can be mediated in solutionby extracellular enzymes. It can involve direct reduc-tion of Cr(VI) or biological reduction of another metalspecies, such as iron, followed by abiotic reduction ofchromium by the reduced metal.

Microbial reduction. Microbial reduction of Cr(VI)has been known for over two decades, with early stud-ies showing that facultatively anaerobic Pseudomonasspecies are capable of catalyzing direct metabolic re-ductions of Cr(VI) to Cr(III) (24). Since that time, nu-merous investigators have shown that bacterialreduction of Cr(VI) is a widespread trait across sev-eral chemotrophic and phototrophic bacterial genera.

In situ reduction of chromium by bacteria can beachieved by the introduction of nutrients (electrondonors), microbes (bioaugmentation), or both.Nutrients may be sugars (e.g., molasses) or organicacids such as acetate, which can be used by manymicroorganisms (25), or lactate, which is metabolizedby a restricted number of organisms. Injection of nu-


PACI

FIC

NOR

THW

EST

NAT

ION

AL L

ABOR

ATOR

Y


trients alone assumes that a suitable population of in-digenous metal-reducing or metal-accumulating or-ganisms exists at the site. Recent research has shownthat such organisms occur widely, so bioaugmenta-tion should not be needed (26).

Several species of bacteria, yeast, and algae cul-tured in the laboratory are capable of sequesteringmetals or changing the redox status of the aquifer sothat the metals precipitate or are more easily adsorbed(27). The unicellular yeast Saccaromyces cerevisiaedemonstrates the most favorable results for metalaccumulators.

Reduction of Cr(VI) to Cr(III) by microorganismscan be direct or indirect. Direct enzymatic reductioncan be achieved by two types of bacteria: dissimila-tory metal-reducing bacteria that can use metals aselectron acceptors for growth, or the fermentativeand other anerobic metabolic groups that reduce met-als, especially relatively easy-to-reduce metals likeCr(VI), as a byproduct of their primary metabolic ac-tivity. An example of a dissimilatory metal-reducingbacterium is Shewanella oneidensis, strain MR-1 (28).A fermentative bacterium that has been shown to re-duce Cr(VI) is Enterobactor cloacae, strain HO1 (29).Cr(VI) reduction by mixed cultures enriched from soilsamples has also been demonstrated in the labora-tory (26). Numerous other bacterial genera have alsobeen shown to reduce Cr(VI).

The indirect approach uses iron-reducing bacte-ria, such as Shewanella alga, strain BrY, to reduce ironoxides and iron-containing clay minerals in aquifermaterials to the ferrous state (30). In this way, a re-ducing barrier of ferrous iron, similar to that describedin the ISRM barrier section, can be created.

Microbial remediation offers relatively low costsand only uses environmentally benign, carbon-basedreducing agents rather than sulfur or iron. The ulti-mate result of carbon metabolism is usually CO2 asopposed to sulfates or ferric iron salts, which maycause secondary problems. However, there are fourconcerns: Nutrients must be injected periodically overthe entire remediation period, which may be years;sufficient formation permeability is needed to allowinjection of nutrients; achieving microbial growthwhere it is needed can be difficult; and unwanted for-mation plugging can occur as a result of excessivegrowth near injection points.

Phytoremediation. Phytoremediation uses plantsto remediate contaminated soil and groundwaterthrough uptake, accumulation/sequestration, or bio-chemical degradation. All vascular plants take up met-als through their root systems, and some canaccumulate and store large amounts. Laboratory stud-ies and small-scale field studies of the uptake/sequestration of several metal contaminants in plants,including chromium, have been conducted. Phyto-remediation is most effective for relatively shallow(3 m or less) groundwater contamination, where theroots can easily reach. Certain trees, such as poplars,can be planted so that their roots reach greater depths,perhaps as much as 10 m. However, because roots donot extend very far into the aquifer, phytoremediationis most effective for chromium contamination in thetop meter or so of the aquifer.

Phytoremediation is most applicable to widely dis-persed, dilute contaminant plumes in which toxicityis not an issue. The approach is relatively inexpensive.Most field trials of phytoremediation to date havebeen aimed at organic contaminants and certain met-als. There has also been some experience with usingcrops such as sunflowers for remediating radioactivecontamination in the former Soviet Union. No fieldtrials aimed specifically at chromium contaminationhave been reported in the literature. Accumulationof chromium in the plants could be a problem withherbivores, unless there were controls prior to harvest.

Natural attenuationNatural attenuation relies just on natural processes inthe environment to achieve the remediation goals ata contaminated site in a reasonable amount of time.This process is also referred to in the literature byother names, including intrinsic remediation, passiveremediation, natural recovery, natural assimilation,and natural flushing. In the form accepted by the U.S.EPA, it is called “monitored natural attenuation” andmust follow a strict protocol, which includes com-prehensive site characterization, monitoring, andclear proof that natural processes are decreasing themass and mobility of the contaminant (31). Examplesof natural attenuation processes that apply to Cr(VI)include dispersion, sorption, dilution, reduction bynaturally occurring reducing agents, and chemicalprecipitation.

Natural attenuation processes typically occur atmost sites, but frequently are not rapid enough toprevent the migration of contaminants past the siteboundaries, or allow remediation of the site in thedesired timeframe (32). It is most applicable to siteswith relatively dilute contamination (<100 times thecleanup goal) and is most frequently mentioned in theliterature as a follow-up to active remediation.Although a study documented that natural attenua-tion of chromium can occur in the field under ap-propriate circumstances, the author was unable tofind any site records of decision that included natu-ral attenuation as a remedy for chromium contami-nation (33).

Applications at two groundwater sitesHanford,Wash. As a legacy of weapons production atthe U.S. Department of Energy (DOE) Hanford Site insouth-central Washington state, several Cr(VI)groundwater plumes are currently impinging on thenearby Columbia River. These plumes may poten-tially impact aquatic receptors in the river. Pump-and-treat systems are operating at two of the plumes.Because permeable reactive barriers appear to offerseveral advantages, the DOE decided to test this con-cept on a third plume. Because of the depth of con-taminant plumes in the contaminated Hanfordareas—an average of 25 m below the ground surface—ISRM was selected.

The groundwater plume requiring treatment is~700 m in width and has maximum concentrationsof over 4000 ppb Cr(VI). A 700-m long ISRM barrieris being installed parallel to the Columbia River’sbank, ~150 m from the river. Design and planning

were completed in 1999, construction began in 2000,and completion is planned for June 2003. The con-struction is proceeding in three phases, with evalua-tions being conducted at the end of each phase(Figure 3). At the time of this writing, phases one andtwo had been completed.

Because groundwater travel times to the ColumbiaRiver from the barrier are approximately three years,it is too soon to assess the success of the ISRM bar-rier. However, early signs are encouraging. Cr(VI) con-centrations in 55 out of 61 wells in the barrier arebelow the 8 µg/L detection limit of the analyticalmethod being used. Cr(VI) concentrations also con-tinue to decline in two monitoring wells that are 50and 75 m downgradient of the barrier section thathas been in place the longest. Some operational dif-ficulties have been encountered with the use of air-rotary drilling to place some of the injection wells.Studies have shown that the air in some cases createdhigh-permeability channels that allowed the contam-inated water to bypass the treatment. It is expectedthat this problem will be resolved by reinjection ofdithionite or the emplacement of new wells (34).

Turlock, Calif. In this central California location,Cr(VI) was released to the soil and groundwater froma wood-treating plant. A groundwater pump-and-treat system handled ~34 million gallons of water overseven years and removed about half of the estimat-ed mass of chromium contamination.

There was concern, however, that because the ef-ficiency of the extraction decreases exponentially withfalling contaminant concentration, final cleanupwould require a considerably longer period of time.Therefore, it was decided that a chemical reducingagent would be used to enhance the pump-and-treatsystem. After a push–pull pilot test in which thereagent was injected into a well and then extractedafter a suitable reaction period, CaSx was chosen asthe reducing agent. A full-scale treatment systembegan in January 1998. Treatment through October1999 reduced the size and mass of the plume by ~98%,according to the investigators (Figure 4).

Chemical additions were discontinued in October1999 because sufficient reductant had been injected toreact with the remaining chromium. Pumping is con-tinuing to move the reductant to the remaining contam-inated area. Sampling during October 2001 showed thatchromium in excess of the 50-ppb cleanup standardremained in only 5 of 43 wells. Groundwater monitor-ing data also have shown that sulfate concentrations inseveral wells have increased to greater than the 250-ppm secondary drinking water standard and that man-ganese concentrations have increased slightly in severalwells. These increases were not unexpected, given thelarge amounts of sulfur compounds added and the in-duction of reducing conditions in the aquifer, and areprobably transient reactions as the aquifer adjusts to thenew reduced conditions.


FIGURE 3

Phased installation of the ISRM barrier for treating chromium-contaminatedgroundwater at the Hanford sitePhases one and two have been completed; phase three will be finished by June 2003. Key: Blue, treatability study(barrier length of ~45 m); green, phase I barrier emplacement (~150 m); pink, phase II (~300 m); and yellow,phase III (~215 m).

1100110

440

110

20

440

20

20

Compliance wellsBarrier performance wellsPlume monitoring wellsRiver pore water tubesAquifer sampling tubesChromium isopleth (µg/L)ISRM barrier

20

Columbia River


CostsReliable cost figures are difficult to obtain from mostof the in situ Cr(VI) demonstrations that have beenconducted. Because these tests have been largely ex-perimental, there are various research and one-timecosts embedded in the totals. The longevity of the re-mediation generally has also not been demonstrated

for most of the processes. It is assumed that becauselong-term operation and maintenance costs will beless for a passive system than an active one like pumpand treat, the long-term costs will also be less.Attempts to quantify the difference are usually sen-sitive to the assumed lifetime of the in situ method,the overall time period selected for the comparison,

FIGURE 4

Chromium plumes in central California before and after enhanced pump-and-treat remediationFull-scale treatment system began in (a) January 1998 and by (b) October 1999, the plume had decreased in sizeand mass by ~98%. Map shows property boundaries and the Cr(VI) plume in milligrams-per-liter concentrations.

Golf

Road

PondGolden State Boulevard (Northbound)

Feet 2500

Scale

Pond

Golden State Boulevard (Southbound)

Youn

gsto

wn

Road

1.0

0.01

1.0

0.10

(a)

Golf

Road

Pond

Golden State Boulevard (Northbound)

Pond

Golden State Boulevard (Southbound)

Youn

gsto

wn

Road

0.01

(b)

0.01


and the net discount rate assumed. Also, there is noaccepted method for factoring the effectiveness ofthe treatment method into the overall comparison.Better and more reliable cost data will help in situchromium treatment move forward (35).

A look aheadOn the basis of some promising results to date, in situtreatment of chromium contamination in ground-water would appear to have a promising future. Insitu treatment could reduce contaminant concentra-tions in aquifers much more quickly than pump andtreat, and there is the promise of lower operation andmaintenance costs. However, it is difficult to gener-alize because the remedy for groundwater cleanup atany given site depends on various site-specific factors.In addition, the current system for selecting reme-dies under the U.S. Superfund law does not encour-age application of innovative technologies. Thecurrent baseline technology for groundwater conta-mination, pump and treat, is considered by the EPAto be a presumptive remedy. Thus, its approval is rel-atively easy and risk-free, even though it can gener-ally be shown that site closure will not be achieved.In addition, despite its shortcomings, there is a lot ofexperience with designing and determining costs ofpump-and-treat systems.

In contrast, experience with in situ technologies islimited, and they have their own problems. Their un-certain performance and costs are difficult to evalu-ate during the feasibility study and proposed planningsteps required by the Superfund law. Thus, many in-novative treatment concepts enter a difficult stage—the data to validate them can only be acquired by afield deployment, but a field deployment is not pos-sible without the validation data. In many cases, insitu remedies have been tried only after pump-and-treat systems have failed. Because much of the re-mediation community supports the concept of in situtreatment of groundwater contamination, the situa-tion is slowly improving. It is hoped that as more per-formance and cost data are acquired from theongoing demonstrations and continuing research in-creases our understanding of subsurface processes,in situ treatment methods for Cr(VI) contaminationin groundwater will gain wider acceptance.

AcknowledgmentsThanks to Bob Fellows, Yuri Gorby, and Ken Krupka ofPacific Northwest National Laboratory; Eric Lindgrenof Sandia National Laboratory; Bill Apel of IdahoNational Engineering and Environmental Laboratory;and Jim Rouse of Montgomery Watson Harza for reviewand additions to several sections of this paper.

Jonathan Fruchter is a staff scientist with the Environ-mental Technology Directorate at Pacific NorthwestNational Laboratory in Richland,Wash. ([email protected]).

References(1) Gilham, R. W.; Sudicky, E. A.; Cherry, J. A.; Frind, E. O.

Water Resour. Res. 1984, 20, 369–378.(2) Mackay, D. M.; Cherry, J. A. Environ. Sci. Technol. 1989, 23,

630–636.

(3) MacDonald, J. A.; Kavanaugh, M. C. Environ. Sci. Technol.1994, 28, 362A–368A.

(4) Buerge, I.; Hug, S. Environ. Sci. Technol. 1999, 33, 4285–4291.(5) Pratt, A.; Blowes, D.; Ptacek, C. Environ. Sci. Technol. 1997,

31, 2492–2498.(6) Davis, A.; Olsen, R. L. Ground Water 1995, 33, 759–768.(7) Sass, B. M.; Rai, D. Inorg. Chem. 1987, 26, 2228–2232.(8) Loyaux-Lawniczak, S.; Lecomte, P.; Ehrhardt, J. Environ.

Sci. Technol. 2001, 35, 1350–1357.(9) Lide, D.R., Ed. Handbook of Chemistry and Physics, 79th

ed.; CRC Press: Boca Raton, FL, 1998.(10) Fendorf, S. E.; Zasoski, R. J. Environ. Sci. Technol. 1992, 26,

79–85.(11) Blowes, D. W.; Ptacek, C. J.; Jambor, J. L. Environ. Sci.

Technol. 1997, 38, 3348–3357.(12) Puls, R. W.; Paul, C. J.; Powell, R. M. Appl. Geochem. 1999,

14, 989–1000.(13) Vidic, R. D. Permeable Reactive Barriers: Case Study Review,

GWTRAC Technology Evaluation Report, E-Series, TE-01-01, Ground-Water Remediation Technologies AnalysisCenter; Pittsburgh, PA, 2001.

(14) Yabusaki, S. B.; Cantrell, K. J.; Sass, B.; Steefel, C. I. Environ.Sci. Technol. 2001, 35, 1493–1503.

(15) In Situ Treatment of Soil and Groundwater Contaminatedwith Chromium; EPA/625/R-00/005; U.S. EnvironmentalProtection Agency, U.S. Government Printing Office:Washington, DC, 2000.

(16) Kjeldsen, P.; Locht, T.; Karvonen, A. P. Removal of TCE andChromate in a Permeable Reactive Barrier Using Zero-Valent Iron. In Proceedings of the 2001 InternationalContainment and Remediation Technology Conference,June 10–13, 2001; Florida State University: Tallahasse, FL,2001; 121–124.

(17) Benner, S. G.; Blowes, D. W.; Ptacek, C. J. Ground WaterMon. Remed. 1997, 17, 99–107.

(18) Li, Z; Bowman, R. S. Environ. Eng. Sci. 1998, 15, 237–245.(19) Fruchter, J. S.; et al. Ground Water Mon. Remed. 2000, 20,

66–77.(20) Nzegung, V. A.; Castillo, R. M.; Gates, W. P.; Mills, G. L.

Environ. Sci. Technol. 2001, 35, 2244–2251.(21) Rouse, J. V.; Leahy, M. C.; Brown, R. A. Environ. Eng.World

1996, 2, May–June 6–11.(22) Electrokinetic Laboratory and Field Processes Applicable to

Radioactive and Hazardous Mixed Waste in Soil andGroundwater; EPA/402/R-97/006; U.S. EnvironmentalProtection Agency, U.S. Government Printing Office:Washington, DC, 1997.

(23) Lindgren, E. R.; Hankins, M.; Mattson, E. D.; Duda, P. M.Electrokinetic Demonstration at the Unlined Chromic AcidPit. SAND97-2592 Sandia National Laboratory: Albuquer-que, NM, 1998.

(24) Lebedeva, E. V.; Lyalikova, N. N. Mikrobiolgiya 1979, 54,517–522.

(25) Suthersan, S. Remediation Engineering Design Concepts;Lewis Publishers: Boca Raton, FL, 1997.

(26) Schmieman, E. S.; et al. Appl. Biochem. Biotechnol. 1997,63, 855–864.

(27) Krauter, P.; Martinelli, R.; Williams, K.; Martins, S. Biodeg-radation 1996, 7, 277–286.

(28) Viamajala, S.; Peyton, B. M.; Apel, W. A.; Petersen, J. N.Biotechnol. Prog. 1997, 18, 290–295.

(29) Rege, M. A.; et al. Biotechnol. Lett. 1997, 19, 691–694.(30) Caccavo, F.; Niels, R.; Costerton, W. Appl. Environ. Micro-

biol. 1996, 62, 4678.(31) Fed. Regist. 1999, 64 (89), 25,039–25,040.(32) Brady, P.; Brady, M.; Borns, D. Natural Attenuation: CER-

CLA, RBCAs, and the Future of Environmental Reme-diation, Lewis Publishers: Boca Raton, FL, 1998.

(33) Henderson, T. Ground Water 1994, 32, 477–486.(34) Fiscal Year 2001 Annual Summary Report for the In Situ

Redox Manipulation Operations; DOE/RL-2002-01; U.S.Department of Energy, U.S. Government Printing Office:Washington, DC, 2002.

(35) Schmelling, S. G. Overview of Abiotic In Situ GroundwaterApproaches (EPA-R-99-012). In Proceedings of the U.S.EPA Conference on Abiotic In Situ Technologies for Ground-water Remediation Conference, Dallas, TX, August 31–September 2, 1999, U.S. Government Printing Office:Washington D.C., 2000; 7–9.

peer reviewed: in-situ treatment of chromium-contaminated groundwater

Documents