Removal of Chromium (VI) byAcid-Washed Zero-Valent Iron underVarious Groundwater GeochemistryConditionsK E I T H C . K . L A I , † , § A N DI R E N E M . C . L O * , ‡

Civil, Architectural and Environmental EngineeringDepartment, University of Texas at Austin, Austin, Texas 78712,and Department of Civil Engineering, The Hong KongUniversity of Science and Technology, Clear Water Bay,Kowloon, Hong Kong, China

Received June 27, 2007. Revised manuscript receivedOctober 22, 2007. Accepted November 23, 2007.

The hexavalent chromium (Cr(VI)) removal capacity of acid-washed zerovalent iron (AW-Fe0) was evaluated under differentgroundwater geochemistry conditions through columnexperiments. It was found that each gram of the AW-Fe0

could remove 0.65–1.76 mg of Cr(VI) from synthetic groundwaterin the absence of bicarbonate (HCO3

-), magnesium and/orcalcium ions. Groundwater geochemistry was found to exertvarious degrees of impact on Cr(VI) removal by the AW-Fe0, inwhich HCO3

- alone gave the mildest impact, whereas thecopresence of calcium and HCO3

- exerted the greatest impact.In comparison with the unwashed Fe0, the AW-Fe0 showeda poorer Cr(VI) removal capacity and was also more susceptibleto the influence of the dissolved groundwater constituentson Cr(VI) removal, thereby indicating the unsuitability of using AW-Fe0 in permeable reactive barriers for remediation of Cr(VI)-contaminated groundwater. On the AW-Fe0 surface, where theindigenous iron precipitates were almost erased, trivalentchromium including chromium (III) oxides, hydroxides, andoxyhydroxides in irregular strip, chick footmark-liked or boulder-liked forms as well as Cr(III)-Cr(VI) mixed oxides weredetected.

IntroductionChromium is a common groundwater contaminant athazardous sites because of its widespread application inmetallurgy, organic chemical syntheses, leather tanning, andwood preserving industries (1, 2) In aqueous environments,chromium usually exists in hexavalent (Cr(VI)) and trivalentforms (Cr(III)). Cr(VI) is acutely toxic and carcinogenic. It isalso highly mobile in groundwater since it does not sorbstrongly onto most soils (1). Contrarily, Cr(III) is relativelynontoxic and an essential human nutrient (3). It does notreadily migrate in groundwater since it usually precipitatesas hydroxides, oxides or oxyhydroxides (4).

Zero-valent iron (Fe0)-based permeable reactive barriers(PRBs) exploit the marked contrast of the toxicity and mobility

between Cr(VI) and Cr(III) to remediate Cr(VI)-contaminatedgroundwater (5). Inside the reactive barriers, both Fe0 andthe Fe2+ released from the anaerobic Fe0 corrosion act asreductants to first chemically reduce Cr(VI) to Cr(III) followedby precipitation as chromium or chromium-iron oxides/hydroxides/oxyhydroxides on the Fe0 surface (6, 7). AlthoughFe0 has been proven to be an effective material for thereductive precipitation of Cr(VI) (4, 5), its Cr(VI) reductionkinetics are prone to be affected by the initial Cr(VI)concentration, groundwater pH, background electrolyteconcentration, iron surface area and type, and copresenceof other dissolved contaminants. It was reported that theCr(VI) reduction kinetics were inversely proportional tothe initial Cr(VI) concentration between 10 and 80 mg/L andto the groundwater pH between 3 and 10, but directlyproportional to the iron surface area (6–8). Increasing theionic strength of the sodium sulfate electrolyte from 3600 to11 000 mg/L decreased the Cr(VI) removal rate by more thana factor of 2 (9). Waste iron metals were reported showinga higher Cr(VI) reduction kinetics than the commercial Fe0

(6). Moreover, the copresence of trichloroethylene (TCE) orcupric ion (Cu2+) with Cr(VI) could lower the Cr(VI) reductionkinetics or Cr(VI) removal capacity of Fe0 probably due tothe competition with the Cr(VI) for the released electrons orthe adsorption sites on Fe0 surface (4, 8).

Many previous studies also focused on the reactivityenhancement methods. Currently, acid-washing of Fe0 (10–12)and deposition of secondary metals such as palladium (13)and nickel (14) on Fe0 surfaces are the methods commonlyapplied. The enhancement from the acid-washing probablystems from the breakdown of the indigenous passivatingoxide layers from the Fe0 surface and the increase of the Fe0

surface area by acid etching and pitting (15, 16). Althoughacid-washing is a recognized approach for enhancing Fe0

reactivity, its influence on the performance of Fe0 PRBs atdifferent subsurface settings is not clear.

The main objective of this study is to evaluate the efficiencyof acid-washed (AW) Fe0 on Cr(VI) removal under variousgroundwater geochemical conditions commonly observedat contaminated sites (17, 18). Calcium (Ca2+), magnesium(Mg2+), and/or bicarbonate (HCO3

-) ions were used toprepare the groundwater at specific geochemistry since theyare the principal ions contributing to the hardness andalkalinity in natural water (17, 18). Column experiments wererun for about 850–1840 h (∼130–280 pore volumes) and wereconducted to simulate the operation of Fe0 PRBs. The Cr(VI)removal capacity of AW-Fe0 obtained was then compared tothat of unwashed Fe0 so as to investigate the appropriatenessof using AW-Fe0 in Fe0 PRBs. The other objective is to identifythe newly formed surface precipitates on the AW-Fe0 on whichthe indigenous surface precipitates have been mostly erased.

Experimental SectionAcid-Washed Fe0. Fe0 filings (ETC-CC-1004) used in this studywere obtained from Connelly GPM Inc. in which their grainsize, specific surface area, and particle density were 0.25 to2.0 mm, 1.8 m2/g, and 6.43 g/cm3, respectively (20). AW-Fe0

was prepared by soaking the Fe0 filings with 1 N argon-sparged hydrochloric acid (HCl, Fisher Scientific H/120 PC17)solution for 3 days followed by sonication in argon-spargedMillipore water and acetone (Aldrich <99% 179973). Thespecific surface area of the AW-Fe0 measured by a Brunauer–Emmett–Teller (BET) surface area analyzer (Coulter SA-3100)and its particle density were 1.94 ( 0.2 m2/g and 6.29 ( 0.1g/cm3, respectively.

* Corresponding author phone: (852) 2358-7157; fax: (852) 2358-1534; e-mail: cemclo@ust.hk.

† University of Texas at Austin.§ Former address: Department of Civil Engineering, The Hong

Kong University of Science and Technology.‡ The Hong Kong University of Science and Technology.

Environ. Sci. Technol. 2008, 42, 1238–1244

1238 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 42, NO. 4, 2008 10.1021/es071572n CCC: $40.75 2008 American Chemical SocietyPublished on Web 01/15/2008

Column Experiments. The column setup used in thisstudy was similar to those mentioned in Lo et al. (4) andAndrea et al. (21) in which the solution was fed into rigid PVCcolumns in an up-flow manner at a flow rate of 400 ( 10m/yr (Supporting Information Figure S1). Columns of 30 cmlong with two sampling ports at both ends and six samplingports located at distances of 2.5, 5, 10, 15, 20, and 25 cm fromthe influent end were packed with 100% AW-Fe0 inside ananaerobic chamber, which resulted in a porosity of 0.6, bulkdensity of 2.5 g/cm3, and an iron surface area to solutionvolume ratio (λSA) of 8083 m2/L. Groundwater samplings weretaken from each sampling port every 5-15 pore volumes(PVs) to collect 3-4 mL samples for pH, total dissolved iron(FeT), total chromium (CrT), Cr(VI), Ca2+, and Mg2+ analyses.A further 30-40 mL sample uniquely from the effluentsampling port was collected for total alkalinity (TAL)measurement.

In total, six columns were operated to evaluate the Cr(VI)removal capacity and identify the surface precipitates formed(Table 1). Columns 1 (C1) and 2 (C2) were only fed by aerobicand argon-sparged Millipore water, respectively, so as todetermine the type and morphology of iron oxides/hydroxides/oxyhydroxides formed on the AW-Fe0 underaerobic and anaerobic conditions. Since these iron precipi-tates were also expected to form in the other AW-Fe0 packedcolumns, the results from C1 and C2 provided the baselineinformation for distinguishing both the chromium precipi-tates and other precipitates formed under specific ground-water geochemistry conditions from the iron precipitates.Columns 3 (C3) to 6 (C6) were fed with argon-spargedsynthetic groundwater spiked with 25 mg/L of Cr(VI).Columns 4 (C4) and 5 (C5) were fed with the above solutioncontaining 200 mg/L of HCO3

- and Mg2+ as CaCO3, respec-tively. The synthetic Cr(VI)-contaminated groundwater spikedwith 200 mg/L of Ca2+ and HCO3- as CaCO3 was the feedsolution of C6. The reason why 25 mg/L of the Cr(VI)concentration was chosen as the feed concentration is thatit is close to the historical record of the Cr(VI) groundwaterconcentration in the U.S. Coast Guard Support Center atElizabeth City, North Carolina at which a Fe0 PRB has beeninstalled for the remediation of Cr(VI)-contaminated ground-water (23) The synthetic groundwater was prepared bydissolving reagent grade potassium dichromate (K2Cr2O7,Riedel-deHaën 12255), calcium chloride (CaCl2 ·2H2O, 31307RDH), magnesium chloride (MgCl2 ·6H2O, RDH 13135), and/or sodium bicarbonate (NaHCO3 BDH30151–5V) into Ar-sparged Millipore water. After complete dissolution, the pHwas adjusted to 7.0 ( 0.1 using 0.5 N HCl and sodiumhydroxide (NaOH, AnalaR102525P).

Analytical Methods. Concentrations of CrT, FeT, Ca2+, andMg2+ were measured by atomic absorption spectrometer(Hitachi Ltd. Z - 8200), and 1,5-diphenylcarbohydrazide

spectrophotometric test (Hach DR/2000) was applied for themeasurement of Cr(VI) concentration. The solution pH wasmeasured using a milli-voltmeter (Orion model 420A) withpH electrode (Orion 9107BN). A titration method using 0.02N sulfuric acid (H2SO4, General Chemical UN 1830) as a titrantand 0.1 wt.% bromcresol green (Exaxol Chemical CorporationB-0443–100)asanindicatorwasutilizedforTALmeasurement.

Surface Characterization. Immediately after the columnexperiments, AW-Fe0 samples located at distances of 5 and15 cm from the column influent end were collected insidethe anaerobic chamber. The samples were freeze-dried(Girovac Super Modulvo) and then stored under argon gasbefore surface characterization. Scanning electron micro-scope (SEM) (model JSM 6300) equipped with energydispersive X-ray analyzer (EDX) was utilized to observe themorphology and elemental composition of the precipitatesformed on the AW-Fe0. The type of the surface precipitatesformed was evaluated by using Raman spectroscope (RM3000) and X-ray photoelectron spectroscope (XPS) (PHI-5600). This was achieved by comparing the measured Ramanspectra with those of the pure compounds reported in theliterature (Supporting Information Table S1) and by findingthe binding energy of the peaks of the XPS spectra (23). Thestructure of the surface precipitates was determined by aX-ray diffractometer (XRD) (PW-1830 Philips) using Cu KRradiation (wavelength ) 1.540562 Å).

Results and DiscussionCr(VI) Removal by the Acid-Washed Fe0. Measurement ofthe CrT and Cr(VI) concentration in the solution collectedfrom the sampling ports indicated that chromium wasremoved from the solution after passing through the AW-Fe0

packed columns. Detection of the Cr(III) peaks (at bindingenergy 576.9–577.5 eV) and the absence of the K2Cr2O7 peak(at binding energy 580 eV) in the XPS spectra of the AW-Fe0

samples of C3 to C6 collected after the column experimentsconclusively showed that reductive precipitation is the mainCr(VI) removal mechanism by the AW-Fe0 (SupportingInformation Figure S2). Since no Cr(III) was detected in thesample solution, the Cr(III) was believed to be completelyprecipitated on the AW-Fe0 surface. These Cr(III) precipitatesacted as insulators inhibiting the AW-Fe0 from furtherreduction of Cr(VI), thereby leading to the subsequentincrease in the Cr(VI) concentration in the collected samplesolution and the formation of the Cr(VI) front. As illustratedin Figure 1 and Supporting Information Figure S3, the Cr(VI)fronts migrated progressively along the columns at constantrates throughout the experiments.

Quantification of the Cr(VI) removal in terms of thereduction rate constant is not appropriate for engineeringdesigns of Fe0 PRBs for Cr(VI) remediation because of theexhaustion of Fe0. Hence the Cr(VI) removal capacity of the

TABLE 1. Column Settings and Composition of the Feed Solutions for the Column Experimentsa

[Cr(VI)] [Ca2+] [Mg2+] [HCO3-]

column Fe0 type mg/L mg/L as CaCO3 mg/L mg/L as CaCO3 mg/L mg/L as CaCO3 mg/L oxygen level

C1b AW-Fe0 0 0 0 0 0 0 0 aerobicC2b AW-Fe0 0 0 0 0 0 0 0 anaerobicC3b AW-Fe0 25 0 0 0 0 0 0 anaerobicC4b AW-Fe0 25 0 0 0 0 200 244 anaerobicC5b AW-Fe0 25 0 0 200 48 0 0 anaerobicC6b AW-Fe0 25 200 80 0 0 200 244 anaerobicC7c,d UW-Fe0 25 0 0 0 0 0 0 aerobicC8b,d UW-Fe0 25 0 0 0 0 200 244 aerobicC9c,d UW-Fe0 25 0 0 200 48 0 0 aerobicC10b,d UW-Fe0 25 200 80 0 0 200 244 aerobic

a Note: AW-Fe0 refers to acid-washed Fe0, whereas UW-Fe0 means unwashed Fe0. b 30 cm long column - 1 pore volumeis equivalent to 6.6 ( 0.2 h. c 50 cm long column - 1 pore volume is equivalent to 11.0 ( 0.3 hrs. d Lo et al. (24).

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AW-Fe0 was calculated in this study using eq 1 based on theCr(VI) front migration rates (i.e., the slope of the lines inFigure 1b) (4, 7, 24).

Removal capacity (mg ⁄ g)) [Cr(VI)]M × A × Fb

(1)

where [Cr(VI)] is the initial Cr(VI) concentration (mg/L), Mis the normalized migration rate of the Cr(VI) front (cm/L)at a relative Cr(VI) concentration of 0.3, which is calculatedby dividing the migration rate (cm/PV) by the pore volumeof the columns (L/PV), A is the cross section area of thecolumns (cm2) and Fb is the bulk density of the AW-Fe0 inthe columns (g/cm3).

As can be seen in Table 2, each gram of the AW-Fe0 couldremove 1.76 mg of Cr(VI) for C3 without containing hardnessand HCO3

- in the feed solution. It is important to note thatthe Cr(VI) removal capacities reported in this paper are onlyvalid at the studied Cr(VI) concentration (25 mg/L), and willbe different from those observed at other influent Cr(VI)concentrations. This is because the higher the influent Cr(VI)concentration, the greater the passivation of the Fe0 surface(25). According to Table 2, the average solution pH inside C3before the Cr(VI) breakthrough was about 10.1, but it droppedto approximately 8.6 when the AW-Fe0 was completelyexhausted. A similar change of pH was also observed in C4and C6. Stoichiometrically, a reduction of 1 mol of dichromateion (Cr2O7

2-) by Fe0 results in the release of 2 moles of Cr3+

and Fe3+, and 14 moles of hydroxide ion (OH-) (eq 2).However, complete precipitation of the released Cr3+ andFe3+ as oxides, hydroxides, and/or oxyhydroxides (SCHEME1 and SCHEME 2) can only consume 12 moles of the OH-.Thus, the accumulation of the remaining 2 moles of the OH-

may be the culprit for an initial increase in pH and a highsolution pH in C3, C4, and C6 before the Cr(VI) breakthrough,as illustrated in Figure 2c. Another possible reason for thesephenomena may be due to a higher rate of the Cr(VI)reduction than the Cr(III) precipitation, thereby resulting inan accumulation of OH- before the Cr(VI) breakthrough.Certainly, the contribution of the OH- released from the waterreduction by the AW-Fe0 cannot be excluded. As shown inFigure 2b, the solution pH started to drop when Cr(VI) beganto break through. This phenomenon may be ascribed to the

partial exhaustion of the AW-Fe0, which might lead to adecrease in Cr(VI) reduction rate and OH- production. Whenthe AW-Fe0 was totally exhausted (Figure 2a), no OH- wasreleased from eq 2, but precipitation of metal hydroxidesmight still continue. Therefore, the solution pH droppedsubstantially and tended to revert to the original value (i.e.,7.0).

Cr2O72-+ 2Fe0(s)+ 7H2O(l)S 2Cr3++ 14OH-+ 2Fe3+ (2)

Effects of Bicarbonate. In the presence of HCO3- in the

feed solution of C4, each gram of the AW-Fe0 could remove1.47 mg of Cr(VI) (Table 2). Comparison of the results betweenC3 and C4 showed that an addition of HCO3

- caused a 16.5%drop in the Cr(VI) removal capacity. Since the solution pHof C4 before the Cr(VI) breakthrough reached 9.55, bothHCO3

- and carbonate (CO32-) ion were the dominated

carbonate species before the Cr(VI) breakthrough in whichthe [HCO3

-]/[CO32-] ratio was about 6 (19). The impact of

HCO3- on the AW-Fe0 for the Cr(VI) removal may be due to

the passivation from carbonate precipitates, such as ironhydroxyl carbonate (Fe2(OH)2CO3) (26) and siderite (FeCO3)(27) formed on the iron surface since there was about a10–50% drop of TAL along C4 in the first 15 PVs (SupportingInformation Figure S4). Another possible reason is thecompetition between HCO3

- and Cr(VI) for the reactive siteson the AW-Fe0 and/or its released electrons, because HCO3

-

is known to be adsorbed onto Fe0 and then reacts as anoxidant (27). Because of the domination of both HCO3

- andCO3

2- in C4, it is believed that a substantial portion of theremoved HCO3

- formed passivated carbonate precipitatesand competed with Cr(VI) to inhibit the AW-Fe0 for the Cr(VI)removal.

Effects of Magnesium. The AW-Fe0 in C5 fed by thesynthetic groundwater containing magnesium showed 1.03mg Cr/g Fe0 of the Cr(VI) removal capacity (Table 2). Incomparison to C3, there was approximately a 41.5% decreasein the removal capacity after the addition of magnesium.The lower removal capacity in C5 in comparison to C4indicated that magnesium exerted a greater impact thanHCO3

- on the deterioration of the AW-Fe0 for Cr(VI) removal.Analogous to C4, the deterioration from magnesium wasprobably attributable to iron passivation caused by themagnesium precipitates such as magnesium oxides (MgO)(eq 3) or hydroxides (Mg(OH)2) (eq 4) formed on the AW-Fe0

during the column experiment (28), since there was amaximum drop of 25% in the magnesium concentrationalong C5 (Supporting Information Figure S5a).

Mg2++ 2OH-SMgO(s) +H2O(l) (3)

Mg2++ 2OH-SMg(OH)2(s) (4)

It is interesting to note that the magnesium breakthroughcurves obtained from the sampling points along C5 (Sup-porting Information Figure S5a) highly correlated with theCr(VI) breakthrough curves (Supporting Information FigureS3b) in which the magnesium front also moved along C5 ata nearly constant rate. Overlapping of the magnesium andCr(VI) breakthrough curves in Figure 3 indicated thatmagnesium and Cr(VI) broke through at the same time. Henceit is believed that factors associated with the reductiveprecipitation of Cr(VI) by the AW-Fe0, such as the reducedchromium generated or the OH- released, may be crucial tothe magnesium precipitation/coprecipitation.

In marked contrast to C3, C4, and C6, the average solutionpH in C5 before starting the Cr(VI) breakthrough was lowerthan that after the complete Cr(VI) breakthrough (Table 2).As shown in Figure 3b and c, the solution pH before startingthe Cr(VI) breakthrough was even lower than the initial pHvalue of 7.0, probably due to the consumption of OH- for the

FIGURE 1. (a) Cr(VI) breakthrough curves of C3 in which 0, 9,4, 2, and O refer to the sampling ports located at distances of2.5, 5, 10, 15, and 20 cm from the column influent end,respectively, and (b) migration distances of the Cr(VI) frontsmeasured at C/C0 ) 0.3.

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magnesium precipitation (eqs 3 and 4). When magnesiumand Cr(VI) began to breakthrough, the drop in the amountof OH- consumed by the magnesium precipitation wasbelieved to be larger than the decrease in the amount of OH-

released from eq 2, resulting in the accumulation of thereleased OH- and subsequent increase in the pH. However,in case the AW-Fe0 was completely exhausted with respectto Cr(VI) and magnesium, no OH- was consumed by themagnesium precipitation and produced from eq 2. Thus, thesolution pH finally decreased again and tended to revert toits original pH value as illustrated in Figure 3a.

Effects of Calcium and Bicarbonate. In the syntheticCr(VI)-contaminated groundwater containing both calciumand HCO3

-, each gram of the AW-Fe0 in C6 could only remove0.65 mg of Cr(VI), which was the smallest in comparison tothe Cr(VI) removal capacity of C3, C4, and C5. This impliesthat a combination of calcium and HCO3

- exerted a strongerimpact than the magnesium and HCO3

- alone on thedeterioration of the AW-Fe0 for the Cr(VI) removal. The

noticeable decrease in the TAL level (Supporting InformationFigure S4) and calcium concentration (Supporting Informa-tion Figure S5b) along C6 showed that AW-Fe0 passivationby calcium carbonate (CaCO3) precipitates may be the culpritleading to the deterioration in the Cr(VI) removal capacityin C6. Unlike magnesium, calcium breakthrough did notcorrelate with the Cr(VI) breakthrough. As illustrated inSupporting Information Figure S5b, the calcium concentra-tion in the various sampling ports increased and reachedplateaus at nearly the same time. Since TAL was still beingremoved from the solution, even though the calciumcompletely broke through, it is believed that HCO3

- in theC6 feed solution not only formed precipitates with calcium,but also precipitated with other cations such as iron.

Acid-Washed Fe0 vs Unwashed Fe0. Comparison of theCr(VI) removal capacity between the AW-Fe0 and unwashedFe0 (i.e., C3 vs C7, C4 vs C8, C5 vs C9, and C6 vs C10) illustratedthat under various groundwater geochemistry conditionsfocused on in this study, the AW-Fe0 possessed lower Cr(VI)

TABLE 2. Cr(VI) Removal Capacity of the Acid-Washed Fe0 at Different Groundwater Geochemistry Conditions and the Comparisonwith Unwashed Fe0 a

column Fe0 type

migrationrate of Cr(VI)

front at C/C0 ) 0.3(cm/PV)

normalizedmigration rate

of Cr(VI)front at C/C0 ) 0.3

(cm/L) r2

Cr(VI) removalcapacity of the

Fe0 at C/C0 ) 0.3(mg Cr/g Fe0)

average pHbefore starting

Cr(VI) breakthrough

average pHafter complete Cr(VI)

breakthrough

C3b AW-Fe0 0.094 5.111 × 10-1 0.99 1.76 10.11 ((0.36) 8.64 ((0.71)C4b AW-Fe0 0.113 6.182 × 10-1 0.98 1.47 9.55 ((0.17) 7.95 ((0.27)C5b AW-Fe0 0.161 8.792 × 10-1 0.98 1.03 6.70 ((1.21) 7.97 ((0.49)C6b AW-Fe0 0.258 14.09 × 10-1 0.97 0.65 8.51 ((0.54) 8.15 ((0.27)C7c,d UW-Fe0 0.070 2.3 × 10-1 0.98 4.1 10.08 ((0.59)e 10.40 ((0.14)C8b,d UW-Fe0 0.040 2.2 × 10-1 0.98 4.2 10.15 ((0.32)e 10.10 ((0.0)C9c,d UW-Fe0 0.081 2.7 × 10-1 0.99 3.5 5.42 ((0.35)e 9.32 ((0.16)C10b,d UW-Fe0 0.060 3.3 × 10-1 0.96 2.8 8.85 ((0.29)e 8.73 ((0.33)

a Note: AW-Fe0 refers to acid-washed Fe0, whereas UW-Fe0 means unwashed Fe0. Figures in parentheses are thestandard deviation. C3, C4, C5, and C6 are under the same groundwater geochemistry conditions as C7, C8, C9, and C10,respectively. b 30 cm long column - 1 pore volume is equivalent to 6.6 ( 0.2 h. c 50 cm long column - 1 pore volume isequivalent to 11.0 ( 0.3 hrs. d Lo et al. (24). e The average pH value for C7, C8, C9, and C10 was obtained when the Cr(VI)relative concentration was 0.72–0.77, 0.71, 0.83–0.88, and 0.74–0.83, respectively.

SCHEME 1

SCHEME 2

FIGURE 2. Relative Cr(VI) concentration (0) and pH profiles (9) measured from the sampling ports at distances of (a) 2.5 cm, (b) 15cm, and (c) 30 cm from the C3 influent end.

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removal capacity than the unwashed Fe0 (Table 2). Therewas approximately a 57-77% drop in the Cr(VI) removalcapacity after acid-washing. This phenomenon is oppositeto the findings of many batch studies in which acid-washingwas reported to be able to enhance the degradation kineticsof chlorinated organics, nitroaromatic compounds, andnitrate by Fe0 (10–12, 15, 29). However, a similar phenomenonwas observed from the column studies conducted by Lai etal. (30) in which unwashed palladized-Fe0 (Pd-Fe0) alsoshowed a faster TCE degradation kinetics than AW Pd-Fe0.

It is generally accepted that acid-washing can increasethe initial Fe0 reactivity significantly by breaking down nearlyall the indigenous iron oxides on the Fe0 surface (SupportingInformation Figure S6) and/or increasing the Fe0 surfacearea (15, 16). However, this might subsequently result inmore severe mineral precipitation on the Fe0 surface incomparison to the unwashed Fe0, thereby leading to moresevere and rapid loss of the AW-Fe0 reactivity afterward andpoorer long-term efficiency of the AW-Fe0 on contaminantremoval. This hypothesis may be the reason why positiveinfluences from acid-washing were observed from the relativeshort-term batch studies, whereas long-term column studyresults showed the deterioration of the Fe0 efficiency oncontaminant removal by acid-washing. The Auger electronspectra of the AW Pd-Fe0 and unwashed Pd-Fe0 reportedby Lai et al. (30) further supported this hypothesis. Thepalladium on the AW Pd-Fe0 was found to be buried underiron oxide layers, whereas the palladium on the unwashedPd-Fe0 was still located on the Fe0 surface, without beingburied, after 3000 PVs of the column experiment.

Both the AW-Fe0 and the unwashed Fe0 suffered similarimpacts from the groundwater geochemistry on the Cr(VI)removal but the Cr(VI) removal capacity of the AW-Fe0 wasmore susceptible to the deterioration by the dissolvedgroundwater constituents. The presence of magnesium, andboth calcium and carbonate in the feed solution only caused14.6 and 31.7% drops in the Cr(VI) removal capacity of theunwashed Fe0, respectively, whereas there were 41.5 and63.1% decreases for the AW-Fe0. In the light of the poor Cr(VI)removal capacity and the susceptibility to the impact fromthe dissolved groundwater constituents on the Cr(VI) re-moval, the AW-Fe0 was relatively unsuitable in comparisonto the unwashed Fe0 for application in PRBs for Cr(VI)remediation.

In addition, it is important to note that the high solutionpH, which has been proven to be able to deteriorate theCr(VI) reduction kinetics of Fe0 (7, 16) did not deteriorate theCr(VI) removal capacity of the AW-Fe0 determined in thisstudy. As seen in Table 2, C3 possessed the highest solutionpH (i.e., 10.1) before the Cr(VI) breakthrough in comparisonto C4-C6 but showed the highest Cr(VI) removal capacity(1.76 mg Cr/g Fe0). On the other hand, the lowest Cr(VI)removal capacity observed in C6 (i.e., 0.65 mg Cr/g Fe0)

appeared at a relatively low solution pH (i.e., 8.51). It isbelieved that the Cr(VI) removal capacity determined in thisstudy can only reflect the groundwater geochemical influ-ences on the amount of Cr(VI) being removed by the AW-Fe0

but is not able to reflect the Cr(VI) reduction kinetics and therelevant pH effects. This is because a large amount (or totalsurface area) of the acid-washed Fe0 was applied to thecolumns in which the Cr(VI) reduction kinetics in the columnswas too fast to be determined by the obtained Cr(VI) sampledata and consequently reflected by the calculated removalcapacities. In C3-C6, the λSA was about 8083 m2/L, which ismuch higher than those in the systems for studying the pHeffect on the Cr(VI) reduction kinetics of Fe0 (e.g., 19 m2/Lin Alowitz and Scherer (16). Based on the Cr(VI) reductionkinetics-pH relationship reported by Alowitz and Scherer(16) the Cr(VI) reduction half-life in C3-C6 between pH 6.7and pH 10.1 should be less than a minute (SupportingInformation Figure S7). However, the sample solutioncollected from the sampling ports along our columns, at least,had stayed in the columns for half-an hour. Thus, it is stronglybelieved that the obtained Cr(VI) sample data and thecalculated Cr(VI) removal capacities cannot reflect the changeof the Cr(VI) reduction kinetics of the AW-Fe0 with varyingpH.

Precipitates Formed on the Acid-Washed Fe0 Surface.Iron Precipitates. During the acid-washing, nearly all theindigenous iron precipitates such as magnetite (Fe3O4) andgoethite (R-FeOOH)/hematite (R-Fe2O3) on the as-receivedFe0 surface were erased (Supporting Information Figure S6).Thus, the precipitates newly formed on the AW-Fe0 surfaceafter the column experiments were completely related to thegroundwater geochemistry of the feed solutions. As sum-marized in Supporting Information Table S2, crystal formsof Fe3O4, ferric hydroxides (Fe(OH)3), and ferric oxyhydroxides(R, γ-FeOOH)/R-Fe2O3 were the major iron precipitatesformed on the AW-Fe0 in pristine water (C1 and C2). Ofparticular concern was that similar types of the ironprecipitates were formed in aerobic and anaerobic feedsolutions. Analogous XRD patterns and Raman spectra wereobtained from the AW-Fe0 samples of C1 and C2 (SupportingInformation Figure S8). The insignificant effect of dissolvedoxygen (DO) on the types of iron precipitates formed on theAW-Fe0 is most likely due to the rapid consumption of nearlyall DO at the C1 influent end so that the bulk AW-Fe0 mediaof both C1 and C2 are believed to be anaerobic.

As illustrated in Supporting Information Figure S9 andSupporting Information Table S3, botryoidal clusters, eu-hedral tabular structures, platy pseudohexagonal forms, andamorphous forms were the main morphology of the ironprecipitates observed in C1 and C2. This observation was ingood agreement with the literature (31) in which the firsttwo morphologies were reported to be the most commoncoating morphologies on Fe0 surface. More importantly, the

FIGURE 3. Relative Cr(VI) concentration (0), pH (9) and relative magnesium concentration (—) profiles measured from the samplingports at distances of (a) 2.5 cm, (b) 15 cm, and (c) 30 cm from the C5 influent end.

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platy pseudohexagonal form of the iron precipitates, whichdoes not belong to green rust widely recognized by itshexagonal shape, was first observed on Fe0. In addition toC1 and C2, the same types and morphologies of the ironprecipitates were also identified in C3-C6. Redox (Eh)-pHdiagrams illustrated in Supporting Information Figure S10ato S10c further showed the stability of Fe3O4, R,γ-FeOOH,Fe(OH)3, and R,γ-Fe2O3 under the pH conditions of C3-C6(6.7-10.11) and the prevailing Eh conditions of Fe0 media(0.3 to-0.45 V) (5, 17, 20, 32–34). Hence the morphology of thenewly formed iron precipitates observed in C1 and C2 was usefulto distinguish the chromium, magnesium, calcium, and car-bonate precipitates in C3-C6 from the iron precipitates.

Chromium Precipitates. After the column experiments,chromium precipitates detected on the AW-Fe0 samples weremainly in trivalent forms. XRD patterns and Raman spectraof the AW-Fe0 samples from C3 shown in SupportingInformation Figure S11, indicated the precipitation ofchromium hydroxides (Cr(OH)3), chromium oxides (Cr2O3),and crystal form of chromium oxyhydroxides (CrOOH).Similar types of the chromium precipitates were also formedon the AW-Fe0 in C4 fed with the HCO3

--rich solution(Supporting Information Table S2). Although XRD, Ramanspectroscopic, and XPS results only identified Cr(OH)3 in C5,the observation of three different shapes of the chromiumprecipitates, which were also observed in C3 and C4(Supporting Information Table S3), showed the high pos-sibility of the existence of CrOOH and Cr2O3 in C5. In C6,Cr(OH)3 and/or CrOOH was identified on the AW-Fe0

samples. The coexistence of Cr2O3, CrOOH, and Cr(OH)3 onthe AW-Fe0 samples may be attributed to the variation of thedegree of hydration of the chromium precipitates undervarious solution pH values. An increase in solution pH couldincrease the degree of hydration on the order of Cr2O3 fCrOOH f Cr(OH)3 (35) SEM micrographs shown in Sup-porting Information Figure S9d to S9f indicated that theidentified chromium precipitates were most likely in irregularstrip, chick footmark-like or boulder-like forms. Besides, itis interesting to note that a Cr(OH)3-Cr2O7 mixed oxide peakat a frequency 854 cm-1 was observed in the Raman spectrafrom C3 (Supporting Information Figure S11b). Probablybecause of the release of OH-, as indicated in eq 2, in C3,some of the Cr(OH)3 might undergo hydrolytic polymeri-zation. These polymeric Cr(III) hydroxides consist of Cr(III)octahedra with hydroxide groups bridging between Cr3+

centers, and H2O or OH- completing the octahedral coor-dination around each Cr3+ (36). It is expected that Cr2O7

2-

is bounded to these polymeric Cr(III) hydroxides by formingCr(III)-O-Cr(VI) linkages (eq 5) and/or through electrostaticattraction by the positive charges on the polymeric Cr(III)hydroxides.

Carbonate, Magnesium, and Calcium Carbonate Precipi-tates. Raman spectra of the AW-Fe0 sample from C4 showeda peak at frequencies between 504 and 507 cm-1 (SupportingInformation Figure S12a). This peak was attributed to theFe3+-OH- stretching mode in green rust precipitates (37).The accompanied Fe2+-OH- stretching mode at frequency420 cm-1 unfortunately was covered by the strong peak ofR-Fe2O3 or R-FeOOH at frequency 415 cm-1 (Supporting

Information Table S1). Since there was no SO42- and only a

trace quantity of Cl- in the C4 feed solution, the green rustidentified is believed to be carbonate-containing green rust(GR(CO3

2-)). The Eh-pH diagram shown in SupportingInformation Figure S10d illustrated the stability of theGR(CO3

2-) in the pH range in C4 (7.95–9.55) and commonEh conditions of Fe0 media (0.3 to -0.45 V). Although FeCO3

and Fe2(OH)2CO3 were not identified by XPS, XRD, and Ramanspectroscopy in this study, their existence in C4 cannot beruled out.

Surface characterization of the AW-Fe0 samples from C5could not identify the type of magnesium precipitates formed.However, its SEM micrograph showed flat, rounded formsof the magnesium precipitates on the AW-Fe0 (SupportingInformation Figure S9g). Besides, in case there were bothcalcium and HCO3

- in the feed solution (C6), calcite crystals(Supporting Information Figure S12b) in the shape ofdistorbed rhombs were formed on the AW-Fe0 (SupportingInformation Figure S9h). To provide more information forstudying the mineral precipitates newly formed on Fe0,precipitate morphologies from the literature and this studyare summarized in Supporting Information Table S4.

AcknowledgmentsWe appreciate the support of the Research Grants Councilfor granting a RGC Competitive Earmarked Research Grant(project no. 617006).

Supporting Information AvailableAdditional 4 tables and 12 figures. This material is availablefree of charge via the Internet at http://pubs.acs.org.

Literature Cited(1) Palmer, C. D.; Wittbrodt, P. R. Processes affecting the reme-

diation of chromium-contaminated sites. Environ. HealthPerspect. 1991, 92, 25–40.

(2) National Research Council. Alternatives for Ground WaterCleanup; National Academy Press: Washington, DC, 1994.

(3) Buerge, I. J.; Hug, S. J. Kinetics and pH dependence ofchromium(VI) reduction by iron(II). Environ. Sci. Technol. 1997,31, 1426–1432.

(4) Lo, I. M. C.; Lam, C. S. C.; Lai, K. C. K. Competitive effects ofTCE on Cr(VI) removal by zero-valent iron. J. Environ. Eng.2005, 13 (11), 1598–1606.

(5) U.S. Environmental Protection Agency (EPA). An In SituPermeable Reactive Barrier for the Treatment of HexavalentChromium and Trichloroethylene in Ground Water Volume 1Design and Installation; EPA: Washington, DC, 1999.

(6) Lee, T. Y.; Lim, H. J.; Lee, Y. H.; Park, J. W. Use of waste ironmetal for removal of Cr(VI) from water. Chemosphere 2003, 53,479–485.

(7) Karvonen, A. J. Cation effects on chromium removal inpermeable reactive walls. Environ. Eng. 2004, 130 (8), 863–866.

(8) Chang, L. Y. Chromate reduction in wastewater at different pHlevels using thin iron wires-A laboratory study. Environ. Prog.2005, 243, 305–316.

(9) Gould, J. P. The kinetics of hexavalent chromium reduction bymetallic iron. Water Res. 1982, 16, 871–877.

(10) Matheson, L. J.; Tratnyek, P. G. Reductive dehalogenation ofchlorinated methanes by iron metal. Environ. Sci. Technol. 1994,28 (12), 2045–2053.

(11) Weber, E. J. Iron-mediated reductive transformations: Inves-tigation of reactive mechanism. Environ. Sci. Technol. 1996, 30(2), 716–719.

(12) Su, C.; Puls, R. W. Kinetics of trichloroethene reduction by zero-valent iron and tin: Pretreatment effect, apparent activationenergy, and intermediate products. Environ. Sci. Technol. 1999,33 (9), 163–168.

(13) Grittini, C.; Malcomson, M.; Fernando, Q.; Korte, N. Rapiddechlorination of polychlorinated biphenyls on the surface ofa Pd/Fe bimetallic system. Environ. Sci. Technol. 1995, 29 (11),2898–2900.

(14) Lai, G.; Gillham, R. W.; Odziemkowski, M. S. Reduction ofN-nitrosodimethylamine with granular iron and nickel-en-hanced iron. 1. Pathways and kinetics. Environ. Sci. Technol.2000, 34 (16), 3489–3494.

(5)

VOL. 42, NO. 4, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 1243

(15) Agrawal, A.; Tratnyek, P. G. Reduction by nitro aromaticcompounds by zero-valent iron metal. Environ. Sci. Technol.2002, 30 (1), 153–160.

(16) Alowitz, M. J.; Scherer, M. M. Kinetics of nitrate, nitrite, andCr(VI) reduction by iron metal. Environ. Sci. Technol. 2002, 36(3), 299–306.

(17) O’Hannesin, S. F.; Gillham, G. W. Long-term performance of anin situ “iron wall” for remediation of VOCs. Ground Water 2002,36 (1), 164–170.

(18) Vogan, J. L.; Focht, R. M.; Clark, D. K. Performance evaluationof a permeable reactive barrier for remediation of dissolvedchlorinated solvents in groundwater. J. Hazard. Mater. 1999,68, 97–10.

(19) Master, G. M. Introduction to Environmental Engineering andScience; Prentice Hall: Upper Saddle River, NJ, 1998.

(20) Lai, K. C. K.; Lo, I. M. C.; Birkelund, V.; Kjeldsen, P. Fieldmonitoring of a permeable reactive barrier for removal ofchlorinated organics. J. Environ. Engrg. 2006, 132 (2), 199–210.

(21) Andrea, P. D.; Lai, K. C. K.; Kjeldsen, P.; Lo, I. M. C. Effect ofgroundwater inorganics on the reductive dechlorination of TCEby zero-valent iron. J. Water, Air, Soil Pollut 2005, 162, 401–420.

(22) Naftz, D. L.; Morrison, S. J.; Davis, J. A.; Fuller, C. C. Handbookof Groundwater Remediation Using Permeable Reactive Barriers-Applications to Radionuclides, Trace Metals, and Nutrients;Academic Press: New York, 2002.

(23) Moulder. J. F.; Chastain, J.; King, R. C. Handbook of X-rayPhotoelectron Spectroscopy: A Reference Book of Standard Spectrafor Identification and Interpretation of XPS Data; PhysicalElectronics: Eden Prairie, MN. 1995.

(24) Lo, I. M. C.; Lam, C. S. C.; Lai, K. C. K. Hardness and carbonateeffects on the reactivity of zero-valent iron for Cr(VI) removal.Wat. Res. 2006, 40, 595–605.

(25) Melitas, N.; Chuffe, O.; Farrell, J. Kinetics of soluble chromiumremoval from contaminated water by zerovalent iron media:Corrosion inhibition and passive oxide effects. Environ. Sci.Technol. 2001, 35, 3948–3953.

(26) Jeen, S. W.; Gillham, R. W.; Blowes, D. W. Effects of carbonateprecipitates on long-term performance of granular iron forreductive dechlorination of TCE. Environ. Sci. Technol. 2006,40 (20), 6432–6437.

(27) Agrawal, A.; Ferguson, W. J.; Gardner, B. O.; Chrsit, J. A.; Bandstra,J. Z.; Tratnyek, P. G. Effects of carbonate species on the kineticsof dechlorination of 1,1,1-trichloroethane by zero-valent iron.Environ. Sci. Technol. 2002, 36, 4326–4333.

(28) Lo, I. M. C.; Lai, K. C. K. Preliminary design of permeable reactivewall based on column studies. In Proceedings of the 4thInternational Congress on Environmental Geotechnics; ISEG: Riode Janeiro, 2002.

(29) Westerhoff, P. Reduction of nitrate, bromate, and chlorate byzero-valent iron (Fe0). J. Environ. Eng. 2003, 129 (1), 10–16.

(30) Lai, G.; Gillham, R. W.; McGuire, M. M.; Fairbrother, H. Chapter10 the performance of palladized granular iron: Enhancementand deactivation. In Zero-Valent Iron Reactive Materials forHazardous Waste and Inorganics Removal; Lo, I. M. C., Suram-palli, R., Lai, K. C. K., Eds.; ASCE: VA, 2006.

(31) Pratt, A. R.; Blowes, D. W.; Ptacek, C. J. Products of chromatereduction on proposed subsurface remediation material. En-viron. Sci. Technol. 1997, 31, 2492–2498.

(32) Gillham, R. W.; O’Hannesin, S. F. Enhanced degradation ofhalogenated aliphatics by zero-valent iron. Ground Water. 1994,32 (6), 958–967.

(33) U.S. Environmental Protection Agency (EPA). Permeable ReactiveBarrier Technologies for Contaminant Remediation; EPA: Wash-ington, DC, 1998.

(34) Roh, Y.; Lee, S. Y.; Elless, M. P. Characterization of corrosionproducts in the permeable reactive barriers. Environ. Geol. 2000,40 (1–2), 184–194.

(35) Oblonsky, L. J.; Devine, T. M. A surface enhanced Ramanspectroscopic study of the passive films formed in borate bufferon iron nickel, chromium and stainless steel. Corros. Sci. 1995,37 (1), 41–168.

(36) Stünzi, H.; Marty, W. Early stages of the hydrolysis of chromium(III) in aqueous solution. 1. characterization of a tetramericspecies. Inorg. Chem. 1983, 22, 2145–2150.

(37) Boucherit, A.; Hugot-Le Goff, A.; Joiret, S. Raman studies ofcorrosion films grown on Fe and Fe-6Mo in pitting conditions.Corros. Sci. 1991, 32 (5/6), 497–507.

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