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Electrochimica Acta 86 (2012) 115– 123

Contents lists available at SciVerse ScienceDirect

Electrochimica Acta

jou rn al hom epa ge: www.elsev ier .com/ locate /e lec tac ta

lectrodialytic versus acid extraction of heavy metals from soil washing residue

ernille E. Jensena,∗, Lisbeth M. Ottosena, Bert Allardb

Department of Civil Engineering, Kemitorvet, Building 204, Technical University of Denmark, 2800 Lyngby, DenmarkMTM Research Center, Örebro University, 70182 Örebro, Sweden

r t i c l e i n f o

rticle history:eceived 6 January 2012eceived in revised form 2 July 2012ccepted 2 July 2012vailable online 9 July 2012

eywords:lectrodialysis

a b s t r a c t

The feasibility of electrodialytic remediation (EDR) for treatment of suspended sludge after soil washingis in focus in the present paper. Five industrially contaminated soils were treated in laboratory scaleremediation experiments, and the toxic elements of the investigation were: As, Cd, Cu, Cr, Ni, Zn andPb. The results showed that all investigated elements could be removed from all soils to some extentby EDR. For all anthropogenic contaminants, higher extraction can be obtained under the influence ofthe direct current during EDR than by washing. During EDR most elements were transferred primarilyto the cathode, where Cu and Pb precipitated at the cathode, while Cd, Cr, Ni and Zn primarily, or solely

oil washingeavy metalsemediationlectrokinetics

(for Ni), were dissolved in the catholyte, showing how cationic species dominated the chemistry of theseelements. Despite the differences between the soils, the remediation results were explained well by thehydrolytic chemistry of the elements, with a difference between the soils. The only element transportedprimarily towards the anode, was arsenic suggesting that As(V) is the dominating species, and showingthat As(III) is oxidized during the remediation process. In contrast the kinetic stability of Cr(III) hindersoxidation of this element, and leaves this as the least removable of the seven.

. Introduction

Soil washing is a clean-up technology for soils contaminatedith heavy metals and organics built upon mineral process-

ng techniques. The processes employed are water washing andensity/grain size separation combined with more advanced tech-iques designed for the individual remediation case, such asagnetic or hydrophobic separators and washing with chemical

eagents. During the washing process a contaminated clay and siltludge is separated from the cleaned sand and oversize materials.he sludge is subsequently dewatered and pressed into a filter-ake. This residual sludge is loaded with contaminants at higherevels than the original soil, and requires landfill deposition. Theolume of residual, dewatered sludge (including bound washingiquid) may approach the volume of originally contaminated soil,nd the soil washing technology is commonly not economicallyeasible if the clay/silt content is above 30–50% [1]. Economical han-ling of the residue is, thus, a main issue in full scale application ofoil washing.

Although many commercial suppliers suggest further treatment

f this fraction, few studies on treatment of this type of sludgeesidual exists. A suggested treatment method is electrodialysis,hich was at first developed for remediation of contaminated soil

∗ Corresponding author. Tel.: +45 4525 2255; fax: +45 4588 5935.E-mail address: pej@byg.dtu.dk (P.E. Jensen).

013-4686/$ – see front matter © 2012 Elsevier Ltd. All rights reserved.ttp://dx.doi.org/10.1016/j.electacta.2012.07.002

© 2012 Elsevier Ltd. All rights reserved.

in bulk [2], and later developed towards treatment of fine particu-late materials in suspension [3,4]. The first works on electrodialyticremediation (EDR) of soil washing sludge residue, showed thatPb, Cu, Cr and partly As can be efficiently removed in the acidicenvironment produced during the EDR process [5–8]. The positiveresults were obtained despite the fact that Pb and Cr are some of themost recalcitrant toxic elements towards electrokinetic remedia-tion (EKR) and EDR in bulk [9–11]. By direct comparison it was showthat remediation of soil fines in suspension is more efficient thanremediation of the whole soil in suspension [8]. Furthermore, As,Cu and Cr could be simultaneously extracted from the soil washingsludge residue [7,8], while they could not be removed simultane-ously from the bulk soil without chemical enhancement [12]. Aprocess diagram visualizing the suggested combined process of soilwashing succeeded by EDR of the fine fraction with material-flowsis shown in Fig. 1.

We showed that EDR of fine particulate materials in suspensionin general is most efficient if water is used as suspension liquid,as opposed to when mixed with various enhancement solutions.E.g. water suspension resulted in efficient extraction of Cu, Zn, Pband Cd from harbor sediments, while neither HCL, NaCL, citric acid,lactic acid or ammonium citrate improved removal [13]. Likewise,11 different organic acids did not improve Pb removal from soil

washing sludge residue [14]; and I2 did not improve Cr removalfrom soil washing sludge residue [7]. In the same work, improvedremoval of Cr and Cu was, however, observed when suspension pHwas maintained at 1 by addition of HNO3 [7]. The reason water

116 P.E. Jensen et al. / Electrochimica Acta 86 (2012) 115– 123

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acidic liquid samples were preserved with one part of concentratedHNO3 to four parts of liquid prior to analysis. Prior to analysis,sludge samples were dried at 105 ◦C and digested according to theDanish standard method DS259 [15], which includes acid digestion

Fig. 1. Process diagram of suggested combined soil washing

uspension is sufficient, is a pH-drop down to 1–3 that occurs dur-ng the EDR process [5–7], which is responsible for mobilizationnd thus successful removal of the toxic elements.

It could be speculated whether simple acid-washing down toimilar pH-values could provide a similar extraction of the toxic ele-ents from the sludge without current application. In the presentork we therefore investigate the feasibility and general applica-

ility of the EDR treatment of soil washing sludge residues fromve different soils contaminated industrially by As, Cd, Cu, Cr, Ni,n and Pb, and compare to acid- and alkali-leaching. Mechanismsnd influence of pH, redox-conditions, soil-type and origin of con-amination are discussed.

. Experimental

.1. Soils

Five soils, all industrially contaminated by at least three differentoxic elements, were chosen for the investigation. Soils 1–4 wereollected in Denmark and soil 5 in Sweden. The soil washing sludgeesidue fraction was obtained by wet sieving (soils 3 and 4) or dryieving (soils 1, 2 and 5) through a 0.063 mm sieve. The carbonateontent of the sludge residue was determined volumetrically by thecheibler-method when reacting 3 g of soil with 20 mL of 10% HClnd calculated assuming that all carbonate is present as calcium-arbonate. Organic matter was determined by loss of ignition in

heating furnace at 550 ◦C for 1 h. pH was measure after shakingf 5 g of soil with 25 mL distilled water for 1 week, and allowinghe particles to settle before measurement with Radiometer pH-lectrode.

.2. Electrodialysis experiments

Electrodialysis experiments were made in cylindrical poly-ethyl methacrylate cells with three compartments (Fig. 2).

ompartment II, which contained the soil-slurry was 10 cm longnd 8 cm in internal diameter. The slurry was kept in suspen-ion by constant stirring with plastic-flaps attached to a glass-sticknd connected to an overhead stirrer (RW11 basic from IKA). The

nolyte was separated from the soil slurry by an anion-exchangeembrane, and the catholyte was separated from the soil slurry

y a cation-exchange membrane. Fig. 2 shows a schematic draw-ng of the setup. Both membranes were obtained from Ionics (types

lectrodialytic remediation of heavy metal contaminated soil.

AR204SZRA and CR67 HVY HMR427). Electrolytes were circulatedby mechanical pumps (Totton Pumps Class E BS5000 Pt 11) betweenelectrolyte compartments and reservoirs. Platinum-coated elec-trodes (length 74 mm diameter 2.9 mm) from Permascand wereused as working electrodes, and the power supply was a HewlettPackard E3612A. The anolyte and the catholyte consisted of 300 and500 mL of 0.01 M NaNO3 adjusted to pH 2 with HNO3, respectively.

The conductivity of the sludge suspension in compartment II, pHin all compartments, and the voltage between working electrodeswere monitored daily. pH in the electrolytes was kept between 1and 2 by manual addition of HNO3/NaOH. After termination of theexperiments, the soil solution was separated from the soil finesby dripping off through filter paper (particle retention 5–13 �m)overnight. Electrodes and membranes were rinsed in 5 M and 1 MHNO3 respectively, tubes and pumps were rinsed by pumpingthrough 1 M HNO3. All liquid volumes were recorded and samplesstored for subsequent element analysis. Two identical experimentswere made with each soil, designated e.g. experiments 1A and 1Bfor soils 1. All experiments ran for 240 h with a current density of20 mA (0.4 mA/cm2). The liquid to solid ratio (L/S) was 10 (40 g soiland 400 mL distilled water). Except for soil 5, the identical experi-ments were made at different times such that any influence fromconditions or behavior of persons maintaining the experiments willbe revealed. The two experiments with soil 5 were made at the sametime and place, and maintained at exactly same way.

Analysis of the selected elements in sludge and process liquidswas made prior to and after the experimental remediation. Non-

Fig. 2. Schematic view of a cell used for experimental EDR remediation of soil-finesin suspension. AN = anion-exchange membrane, CAT = cation-exchange membrane.

P.E. Jensen et al. / Electrochimica Acta 86 (2012) 115– 123 117

Table 1Characteristics and total metal concentrations of experimental sludge residue. Values exceeding limiting values are bold. *Valid for Cr(III); Cr(VI) maximum 5 mg/kg, **validfor Cr(III); Cr(VI) maximum 20 mg/kg, *** for soils to be applied for sensitive land use.

Sludge Contamination origin pH CaCO3 [%] Org. Mat. [%] As Cd Cr Cu Ni Pb Zn

[mg/kg]

1 Unknown 7.7 13.5 7.3 178 43 80 504 55 383 72102 Metal Foundry 7.5 14.4 7.9 24 6 97 1520 75 418 12703 Wood preservation 7.7 0.0 2.9 9260 0.5 2310 6820 <20 5 2414 Wood preservation 7.3 0.0 5.0 3030 0.5 1680 2780 <20 5 2885 Chlor-alkali processing 6.4 0.5 4.8 <0.1 3.7 196 163 57 281 496

Limiting value in Sweden*** 15 0.4 120* 100 35 80 350

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Figs. 3 and 4 show the development of pH and conductivity inthe sludge during EDR treatment. In all experiments pH dropped toaround 2 units. pH dropped immediately in the carbonate deficientsludge (3, 4 and 5), while the carbonate in sludge 1 and 2 functioned

Limiting value in Denmark***

f 1 g soil with 20.00 mL of 7 M HNO3 in autoclave at 200 kPa and20 ◦C for 30 min followed by vacuum filtration through a 0.45 �mlter. All elements except As were analyzed by flame-AAS. Sam-les with Pb and Cd concentrations below the detection limit of theame-AAS were analyzed by GF-AAS. Arsenic was analyzed by ICP-S. Initial element concentrations and selected soil characteristics

re shown in Table 1.After experimental remediation and analysis of samples, the

mounts of the individual elements remaining in the soil (com-artment II), dissolved in the suspension solution (compartment

I), and transported to the cathode (compartment III) and the anodecompartment I) were calculated. The amount transported to theathode end was calculated as the sum of the element mass foundn the cation-exchange membrane, in the catholyte and precipi-ated at the cathode. Correspondingly, the amount transported tohe anode end was calculated as the sum of the element mass foundn the anion-exchange membrane, in the anolyte and precipitatedt the anode.

.3. Acid/alkali-leaching experiments

The effect of acid/alkali-leaching of toxic elements from theludge was studied by reaction of 5.00 g dry residual with acid andase reagents at L/S 5 at 200 rpm for 7 days. Literature on desorp-ion kinetics of heavy metals from historically as well as laboratoryontaminated soil and clay suggests that equilibrium is obtainedn the time span of hours [16–18], and that maximum desorptions reached faster when soil is historically contaminated or adsorp-ion has been allowed during long-term (6 months) as opposedo short term adsorption (24 h) [17]. It can thus be anticipatedhat maximum extraction is obtained during the 7 days of reac-ion time in our experiments. The reagents were as follows: 1.0 MaOH, 0.5 M NaOH, 0.1 M NaOH, 0.05 M NaOH, 0.01 M NaOH, dis-

illed water, 0.01 M HNO3, 0.05 M HNO3, 0.1 M HNO3, 0.5 M HNO3,.0 M HNO3. pH was measured after 10 min settling, and the liquidas filtered through a 0.45 �m filter for subsequent element anal-

sis by ICP-OES. Non-acidic samples were preserved with one partf concentrated HNO3 to four parts of liquid prior to analysis.

. Results

.1. Electrodialysis experiments

.1.1. Mass balancesMass balances, understood as the post-treatment mass of the

ontaminating elements encountered in the whole system (com-

artment I, II, III, electrodes and membranes) given in percent ofhe initial mass of contaminant anticipated from the initial con-entration in the sludge, are given for each investigated elementnd sludge in Table 2. Also, the mass balance for the sludge residue

20 0.5 500** 500 30 40 500

itself is given, i.e. the fraction of the sludge, which was recoveredafter the experiments and thus not dissolved during the process.

The sludge mass balances varied between the soils, while thetwo identical experiments (A and B) showed quite similar sludgemass balances. During EDR the total mass of sludge decreased within average 25% for soil 1, 22% for soil 2, 17% for soil 5, 12% for soil 3and 11% for soil 4. The loss of sludge mass during EDR in suspensionwas earlier suggested to occur primarily due to dissolution of car-bonates [5]. The present work supports this hypothesis by showinga larger mass reduction for sludge with high carbonate content;but the high mass reductions observed illustrate that also othersoil constituents must be partly dissolved during remediation ofthe present sludge.

The mass balances for the contaminants varied between 71%and 136% with an average of 105%. Compared to other works onEDR and EKR of industrially contaminated soil, the mass balancesare good. Because identical experiments (A and B) in most casesshowed similar mass balances for the same elements, it is likelythat deviations from the ideal 100% is primarily due to inexact esti-mation of the initial contaminant concentrations in the soil finesdue to the inhomogeneous nature of industrial contaminationsrather than loss during experiment or inaccurate analysis. The sim-ilar mass-balances for identical experiments further support therepeatability of the experiments. In the following sections, resultsare therefore given with respect to the final amount of contaminantencountered.

3.1.2. pH and conductivity

Fig. 3. pH development with time during electrodialysis experiments (average oftwo experiments with standard deviation error bars – no standard deviation after6 days for sludge 1–4, because pH was measured in only one of the experiments atthis time).

118 P.E. Jensen et al. / Electrochimica Acta 86 (2012) 115– 123

Table 2Mass balances for each sludge residue and element in the 10 experiments (%).

Sludge Experiment Dry mass As Cd Cr Cu Ni Pb Zn pH start pH end

1 A 76 115 129 103 92 105 117 6.9 1.8B 74 112 118 96 92 104 106 7.1 1.9

2 A 78 72 114 101 98 100 106 6.6 1.6B 77 71 110 101 93 100 107 7.1 1.6

3 A 90 94 93 97 5.6 3.6B 86 98 84 102 6.7 3.8

4 A 89 101 82 102 4.1 2.1B 88 94 83 95 4.7 2.2

5 A 83 129 101 111 127 107 123 3.9 1.40

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s a buffer during the first 6 days. In opposition to sludge 5, pHncreased almost 2 points again after the first decrease in sludge 3nd 4. For sludge 4 it decreased finally again to approximately 2,hile it remained at almost 4 for sludge 3, as the only one.

Conductivity increased in all experiments, however significantlyore in the experiments with sludge 5, which also maintained

he lowest pH throughout the experiments. Thus hydrogen-ionsave probably been a primary contributor to the conductivity,lthough dissolution of soluble chloride salts could be an addi-ional explanation as this soil was contaminated by chlor-alkalirocessing. This effect is supported by a higher initial conductiv-

ty in this sludge. For all soils sludge, the conductivity increaseccurred simultaneously with the pH-drop, and conductivity devel-pment was in general inversely related to pH. Hence conductivityf sludge 3 and 4 decreased as pH increased, and conductivity ofludge 4 increased again as pH decreased by the end of the exper-ments, highlighting the dominating effect of hydrogen-ions andcid-soluble salts.

For both pH and conductivity, the standard deviation betweenwo identical experiments was in general small, thus the exper-ments can be concluded to be repeatable on these parameters.eviation in the pH-development between identical experimentsas in general between 0.0 and 0.6 pH units, except during the

teep pH-decrease of carbonaceous sludge where it was up to 1.8H-units, which shows that the exact time point of this decreaseo happen is sensitive to small changes in experimental con-itions. The sensitivity of experimental conditions was further

bserved through the difference in deviation between the two iden-ical experiments of sludge 5, which ran simultaneously, and theemainder identical experiments, which ran non-simultaneously:

ig. 4. Conductivity development with time during electrodialysis experimentsaverage of two experiments with standard deviation error bars – no standard devi-tion after 6 days for sludge 1–4, because conductivity was only measured in one ofhe experiments at this time).

94 111 103 120 4.2 1.5

the deviation in pH and conductivity for sludge 5 was in generalone size order smaller.

3.1.3. EDR removalThe final concentrations of the target toxic elements are shown

in Table 3. In all instances, contaminant concentration was reduced,however, for most contaminants not enough to meet the lowestregulatory limits. The standard deviation of final concentrationsof identical experiments was <12%, except for Cd in soil 5, forwhich the standard deviation was 32% due to the very low ini-tial concentration. The low standard deviations demonstrated thatthe EDR process also is repeatable concerning removal efficiency,and no effect on the remediation result of the small differences inmanagement of identical experiments run at different times wasobserved.

Fig. 5 shows the post-treatment distribution of the elements inthe different parts of the electrodialytic cell. The fraction removedcalculated as the fraction moved either to the cathode- or anodeends varies between 7% and 96% depending on element and soil.In general good results were obtained for Cd, Zn, Pb, Cu, Ni and As;while in particular Cr was removed only to a small extent. Mostelements were transferred primarily to the cathode end, where Cuand Pb precipitated at the cathode, while Cd, Cr, Ni and Zn wereprimarily, or solely (for Ni), dissolved in the catholyte, showinghow cationic species dominated the chemistry of these elementsduring remediation. For Cd, Cu, Ni, Pb and Zn this is in accordancewith the expectation, because free, hydrated cations dominate thechemistry of those elements under the acidic conditions prevailingduring remediation. Only a minor fraction was found dissolved inthe liquid phase of the sludge, thus recirculation of the liquid posttreatment is an actual option.

3.2. Acid/alkali-washing

Fig. 6 shows the results of acid/alkali extraction and compareswith the extraction obtained during EDR. In general, extraction ofthe elements was obtained at pH below 4 and for As and Cu alsoat pH above 10. The highest extraction obtained was 81% Cu and49% Pb from sludge 2 at pH 0.4 and 57% Cd from sludge 1 at pH1.0. Extraction just around 40% was obtained for many elementsin several of the sediments at pH < 1. Less extractable were Ni andZn from sludge 1; Ni, Zn and Cd from sludge 2; Pb and Ni fromsludge 5; and, in particular, Cr from sludge 4 and 5. The extractionobtained by EDR at pH values between 1.4 and 1.9 was significantlyhigher than the extraction obtained by acid washing at pH below

1 for all elements except Cr from soil 1 and 2. The only excep-tion to this was for Cr in soils 1 and 2, which most likely concernsnon-anthropogenic Cr as the level did not exceed lowest regulatorylimits.

P.E. Jensen et al. / Electrochimica Acta 86 (2012) 115– 123 119

Fig. 5. Distribution of contaminating elements in the electrodialytic cell after 10 days of experimental remediation. Anode end (I) includes metal at the anode, in theanion-exchange membrane and in the anolyte. Cathode end (III) includes metal at the cathode, in the cation-exchange membrane and in the catholyte.

Fig. 6. Comparison of extraction obtained by washing with acid/alkali at different pH values (lines) and extraction obtained during EDR (single points). EDR results are plottedagainst the lowest pH-value recorded during the period of the experiments.

120 P.E. Jensen et al. / Electrochimica Acta 86 (2012) 115– 123

Table 3Final concentrations of the target elements in each soil washing sludge residual (mg/kg DW ± standard deviation).

Sludge Contamination origin As Cd Cr Cu Ni Pb Zn

[mg/kg]

1 Unknown 3.1 ± 0.4 116 ± 6 157 ± 2 39 ± 0.5 211 ± 2 572 ± 332 Metal Foundry 0.6 ± 0.0 127 ± 1 317 ± 3 38 ± 0.5 229 ± 3 154 ± 23 Wood preservation 2139 ± 188 1039 ± 144 368 ± 394 Wood preservation 1902 ± 45 1271 ± 9 122 ± 7

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5 Chlor-alkali processing 0.8 ± 0.2

. Discussion

.1. Overall removal

Earlier works [5,7] showed that longer EDR remediation timesan reduce the Pb and Cu content substantially not only during,ut also after acidification. Pb was reduced down to below reg-latory limits after approximately 25 days [5], while it took 22ays at substantially lower current density and L/S for Cu andr [7]. It was, however, also indicated earlier, that longer reme-iation time is not always improving remediation, as for As andr [7] in a CCA-contaminated soil. In this work, our experimentsan for only 10 days, thus the results represent intermediate val-es reached after the first acidification meant for comparison withcid-/alkali leaching. Those results are not directly comparable tohe results of this work even though the soil used is from theame site as soil 3 of this work, because the batch used in thisork contained much higher contaminant concentrations. Nev-

rtheless the reduction from 9260/6820/2310 mg/kg As/Cu/Cr to139/368/139 mg/kg in 10 days at 20 mA and L/S 10 obtained

n this work, compared to reduction from 3200/2170/710 mg/kgs/Cu/Cr to 1040/320/510 mg/kg in 14 days at 2.5 mA and L/S 3.6upport the studies that suggest a delicate balance between L/Snd current density must be maintained to obtain most efficientemoval efficiency [6,8]. Thus prior to large scale EDR remedia-ion, experiments should always be made to determine optimal/S and current density. In contrast longer reaction-times whenpplying acid/alkali-washing cannot be expected to reduce con-entrations further, and since the EDR results show higher removalor all anthropogenic elements in all soils, this method, in general,roves more potent for remediation of sludge of soil washing. Theact that higher extraction is obtained by EDR than by acid-leaching,ven when pH during the acid-leaching was lower than the pHbserved during EDR, shows that the desorption obtained duringDR cannot be driven by acidification only, but is somehow accel-rated under the influence of the current field. This could be partlyue to the concentration of the current near the particle surfaces,orcing desorption of contaminants within the Stern-layer directly.urther mechanisms are discussed below for the different elementategories.

.2. Removal of cationic elements

The removal-rates of various toxic metals by traditionalKR/EDR was observed to decrease according to the followingrders: Ni > Cd > Cr > Zn > Cu > Pb [10], Zn > Cu ≈ Pb and Cu > Cr [9],d ≈ Zn > Cu ≈ Pb [19], Zn > Cu > Pb [11], Zn > Cu > Pb > Cr [20] andi ≈ Zn > Cu > Cr [21]. Most of these observations support theypothesis that removal in general follows the order of the firstydrolysis constants for the elements [21], and is consistent with

he observation that the selectivity of mineral soils for adsorptionf heavy metals corresponds to the order of increasing pK’s of therst hydrolysis product of the various metals [22]. If this hypothesis

s valid for EDR in suspension as well, the removal order among the

182 ± 1 59 ± 0 40 ± 3 228 ± 4 146 ± 5

studied elements expected is Cd > Ni > Zn > Cu > Pb > Cr(III). Arsenic,which is a metalloid, does not behave as a cationic metal regardingits chemistry in soil; neither does Cr(VI), which explains why theydo not appear in the removal order. The observed removal ordersare, soil 1: Cd > Zn > Cu > Pb > Ni » Cr; soil 2: Cd > Zn > Cu > Ni ≈ Pb »Cr; soils 3 and 4: Cu > As > Cr; and soil 5: Zn > Cd > Cu > Ni > Pb > Cr.Coherence with the expected removal order exists apart from Ni,which consistently seems to be less mobile than suggested by thefirst hydrolysis constants of the elements. This is in contrast toresults of stationary EKR of soil [10,21], of which the latter workfound Ni to be more amenable to remediation than Zn and Cufrom the same soil referred to as soil 5 in this work. The incon-sistence suggests that either the mechanisms of the remediationprocesses in the stationary and suspended setup are non-identical,or the processes change as remediation proceeds and less mobilefractions of the contaminant is targeted: remediation proceededsubstantially further in this work than in the previous works withrespect to the fraction of each contaminant removed. Further-more, the removal order in soil 5 deviates in that Zn was moreamenable to remediation than Cd. This is likely to be due to theanionic heavy metal complexation prevailing in soil 5 as discussedbelow. When comparing the order of extraction by EDR with theorder of extraction by acid-leaching (Fig. 6), substantial differencesespecially between sludge 1 and 2 are observed. Acid leachingresults in the following orders: sludge 1: Cd > Cu = Pb = Cr > Zn ≈ Ni;sludge 2: Cu > Pb > Zn > Cd ≈ Ni, Cr; sludge 3 and 4: Cu > As > Cr; andsludge 5: Cu > Cd > Zn > Ni > Pb > Cr. These results illustrate largerdissimilarities in acid-solubility of elements between soils and lesspredictability and coherence of results with order of pK’s of thefirst hydrolysis product than observed during EDR-removal. Acidextraction is thus not directly predictive of EDR remediation results.Among the cationic elements, it seems that in particular Zn and Niremoval may be underestimated by acid-extraction results. Alsothere is no correlation between fraction extractable by acid andfraction removed during EDR. E.g. acid-extraction of Cu from sludge2 is approximately double the extraction from the remainder soils,while EDR removal of Cu is quite similar.

4.3. Arsenic speciation and removal

The only element, which was transported primarily towards theanode, was arsenic, suggesting that anionic arsenic-species domi-nated under the prevailing acidic conditions during remediation. Ingeneral, arsenic may be present as As(III) or As(V) in soil as well as insolution. Anionic species of As(V) prevail over a much wider rangeof pH and p� conditions than As(III), which is anionic only underalkaline conditions. As pH in the sludge decreased from 6–7 to 1.5–2during remediation, our good removal results indicate that As(V)was the dominating specie in both investigated soils. In contrast,earlier results with stationary EKR/EDR of arsenic contaminated

soils, showed that arsenic was immobile under acidic and neu-tral conditions [12,23], and could not be extracted together withCu and Cr at such conditions because they were not charged andthus mobile for electromigration by the current at same pH [12].

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P.E. Jensen et al. / Electroc

nstead good removal was obtained by addition of either ammo-ia [12] or hydroxide [23] to maintain alkaline conditions (pH > 9),uggesting that As(III) was the dominating species in those soils.ne of the soils used in this study (soil 4), was identical to thene used in [12], thus the transfer of arsenic to the anolyte in theresent study indicates that beneficial oxidation of As(III) to As(V)ook place during electrodialytic treatment in suspension, while itid not during traditional EKR/EDR in stationary setup. In contrasto stationary EKR/EDR, oxygen and carbon dioxide concentrationsn suspended EDR can, indeed, be assumed to be in equilibrium

ith the atmosphere, as the suspension is in direct contact with thetmosphere through the whole for the stirrer which should allowor oxidation of As(III) to As(V) during remediation. One impor-ant consideration, however, was the kinetics of the oxidation, sincehe rates of change do not always appear to be very rapid, and theroportion of various arsenic species present may not always corre-pond to equilibrium [24]. The results of this study demonstratedhat the kinetics, however, is fast enough to allow for oxidationf As(III) to As(V) either during the washing process or under thecidic and oxidizing conditions prevailing during suspended EDR.nother consideration was the lower mobility of As(V) compared

o As(III): in a previous investigation, EKR of As(III)-contaminatedoil was successfully enhanced by addition of an oxidizing agentNaClO) [23]. This enhancement was assumed to be a result ofhe arsenic-release induced by oxidation of organic soil compo-ents and/or ion-exchange between anionic arsenic species andlO− [23]. Oxidation of As(III) to As(V) was excluded as explanationue to the lower mobility of As(V). The lower mobility is, how-ver, observed under natural conditions [24], and may not applyo EDR/EKR, where the prevalence of charged species is of cru-ial importance to the remediation result. Indeed, mobilization andemoval of As(V) from CCA-impregnated waste wood by EDR wasemonstrated in several studies [25–27]. The observed transfer ofrsenic to the anolyte in the present study proves that althoughs(III) is considered more mobile than As(V) under natural condi-

ions, oxidation to As(V) facilitates a successful mobilization andemoval of As(V) under the influence of direct current. The acid-xtraction of As from the two As-contaminated residuals was morer less identical, while EDR removal was substantially higher fromludge 3, showing that the mobilization of As caused by oxidationuring EDR cannot be predicted by acid-leaching results.

.4. Chromium speciation and removal

In contrast to the encouraging removal of arsenic, mobilizationf chromium was low in most soils and transport occurred almostxclusively to the catholyte, which according to the Cr-Pourbaixiagram reveals that Cr(III) must be the dominating species in alloils. Removal of Cr(III) as a free, hydrated cation was expected to beore recalcitrant and require lower pH than removal of Pb accord-

ng to the pK’s of their first hydrolysis product [22]. In comparisonhe mobility of Cr(VI) is considerably higher, and anionic species ofr(VI), prevail in the full pH interval. Investigations of the influencef Cr-speciation on stationary EKR showed that removal of Cr(III)ccurred only under highly acidic conditions, while removal ofr(VI) was observed to increase at neutral/alkaline conditions, andhat Cr(VI) was observed to be faster remediated than both Cd andi even under acidic conditions [28]. In general, Cr was recovered in

he anolyte when soils were spiked with Cr(VI) [28–31], and in theatholyte when soils were spiked with Cr(III) [32–34], in accordanceith the high kinetic stability of chromium in soil. After stationary

KR of a contaminated soil from a military site, Cr was recovered

rom both electrolytes [35], while chromium was almost exclu-ively recovered in the catholyte after stationary EDR of soil 4 usedn this work [9], supporting the dominance of Cr(III) in this soil. Oxi-ation of Cr(III) to Cr(VI) corresponding to the oxidation of As(III)

Acta 86 (2012) 115– 123 121

to As(V) during remediation in suspension would be expected toimprove remediation accordingly, but the kinetic stability of Cr(III)renders this process less likely, in accordance with the results.Although removal of chromium from most of the present exper-imented soils was low, the possibility of mobilizing and removingCr(III) by EDR in suspension was established after removal of 53% ofthe chromium during 240 h from the severely contaminated sludge3. The high removal from this particular sludge may be due to spe-cific soil-characteristics and chromium speciation. This hypothesisis supported by the higher acid extraction from this sludge residuecompared to the other Cr contaminated sludge residues. It wasshown previously that remediation of Pb-contaminated soil is moreefficient from severely contaminated soils, while impeded by car-bonate and organic matter [36]. The fact that this soil is carbonatedeficient and low in organic matter suggests that removal of Cr(III)behaves similarly.

4.5. Influence of soil characteristics

The maximum removal percentages after 240 h for each ele-ment were 79% As (sludge 3), 92% Cd (sludge 1), 55% Cr (sludge3), 96% Cu (sludge 4), 52% Ni (sludge 2), 53% Pb (sludge 1) and 88%Zn (sludge 1). Among the sludge residuals, removals from residu-als 1 and 2 were similar. Slightly better results were obtained forPb, Cd, Cr and Zn from residual 1, and for Cu and Ni for residual2 and the removal order among the elements was identical forthe two residuals. The similar results suggest that not only theresidual characteristics but also the contaminant speciation is sim-ilar in these residuals, as it was earlier shown that the originalsource and thus speciation of the contaminant is a predominantparameter when determining the remediation feasibility [36]. Incomparison removal of As and Cr from residual 3 was substan-tially higher than from residual 4 although these residuals bothorigin from CCA-impregnation of wood. Reasons may be the highercontamination-level in residual 3, which may cause a higher frac-tion of the contaminants to be mobile and/or the higher organiccontent of residual 4. Cr and Cu were present as contaminants in allof the soils sludge, which allow for comparison of removal betweensludge of relatively high/low organic matter and carbonate con-tent. The conclusion is that removal is significantly more efficientfrom the non-carbonaceous and inorganic sludge, in accordancewith previous results for stationary EDR of Pb [36], while the leastsuccessful results were obtained for the intermediate soil 5 prob-ably highlighting the influence of contaminant origin and elementspeciation. Apart from binding the contaminants stronger than theremainder soils, soil 5 was also unique in that a fraction of all con-taminants was recovered in the anolyte. This is in consistence withthe results obtained for Pb-removal from this soil by stationary EKR[21], in which the observed transfer to the anolyte was suggestedto be a result of an extraordinary high sulfate content in the soil(up to 4%), which result in transfer of negatively charged lead sul-fate Pb(SO4)2

2− towards the anode. In the previous work, transporttowards the anode was however not observed for the remainder ofthe studied elements: Ni, Zn, Cd and Cu [21] as this study. Chloridecomplexes could be an alternative explanation.

Despite the differences between the soils, the remediationresults were explained well by the hydrolytic chemistry of the ele-ments, and to a minor extent by the relatively small differencebetween the soils, suggesting that pH and p� conditions are ofcrucial importance to EDR in suspension. Theoretical simulationsof these basic operational parameters could in that case be madebased on simple chemical equilibrium data. However it is well

known that the metal retention in soils also depend on a num-ber of factors like interaction with organic matter, precipitationof sparingly soluble compounds and in particular the presence ofextremely stable contaminating compounds [37]. Furthermore, it

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as earlier shown that the contaminant speciation influences theemediation results of stationary EDR seriously [36]. Our resultsuggest that EDR of soil washing residuals in suspension is lessnfluenced by those factors. One reason may be that in the sludgeesidual, the contaminant chemistry is dominated by adsorptionnd exchange adsorption to the mineral matrix and complexationith organic matter. Such mechanisms were also suggested by the

esults of [6], and may be one reason that EDR of soil fines in sus-ension appears more efficient than EDR and EKR of the whole soil

n stationary matrix.

.6. Combined soil washing and electrodialysis

Apart from the higher efficiency, the advantages of combiningDR and soil washing as opposed to only EDR are: (1) a smaller vol-me of particulate material is to be treated by EDR compared to theulk soil (saves energy and time); (2) absence of coarse materials isllowing for remediation in a more homogeneous slurry and possi-ly continuous remediation; (3) possibility for control of attractiveH- and redox conditions and good mixing of enhancement solu-ions into the soil when necessary. The advantages of succeedingoil washing by EDR of the sludge fraction are: (1) no landfill deposi-ion of the soil washing sludge residue is required; (2) the washingater, which otherwise requires water treatment [1], is left clean

fter EDR and can be recirculated, (3) reduced need of chemicalnhancement, (4) the toxic elements are concentrated in a smallerolume of solution, thus producing less, toxic wastewater to beandled. A disadvantage is that the soil must be excavated prior toreatment. This may however be the selected solution anyway ataluable former industrial sites in inner cities, that are re-developnto residential areas, as time is a governing economical parame-er. In addition, the sand and oversize residuals from the washingrocess are left in well defined size fractions which can be used

n the construction industry. E.g. reuse of the materials >63 �m asoncrete aggregates was shown to be feasible [38]. If remediations made on-site, the residual materials may also be mixed and filledack after conditioning with e.g. lime, or with humic material if theaterial is to be used as topsoil. This solution may be feasible for

bandoned countryside industrial areas. At grounds where build-ngs have to be preserved slower in situ remediation is the besthoice.

. Conclusions

This work shows that EDR is a superior method of removal ofoxic elements from sludge residue of soil washing compared tocid-extraction. Better extraction of As, Cd, Cu, Cr, Ni, Zn and Pbs obtained by EDR. The removal by EDR is not driven by acidi-cation only, but accelerated under the influence of the currenteld. Acid extraction results can thus not be used in general toredict EDR result. In contrast, the coherence of EDR-results withhe order of pK’s of the first hydrolysis product can be used as arst estimate of results, apart from Ni, which consistently seemo be less mobile than expected. In particular for As, Zn and NiDR-removal may be underestimated by acid-extraction results.t must be taken into account that remediation proceeds slower inludge with high content of organic matter and carbonate. Smallcale experiments should always be made to optimize L/S andurrent density, prior to large scale EDR remediation. Our studyurther shows that although As(III) is considered more mobilehan As(V) under natural conditions, oxidation to As(V) facili-

ates a successful mobilization and removal of As(V) under thenfluence of direct current. The results suggest that the oxidizingonditions prevailing in the suspended sludge enhance removalf some species, and that the EDR-method is particularly efficient

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Acta 86 (2012) 115– 123

for treatment of suspended clay and silt materials. Cr is the ele-ment least amenable to remediation: Cr(III) is not oxidized duringthe EDR-process due to its high kinetic stability. Only a minorfraction of the investigated elements was found dissolved in theliquid phase of the sludge, thus post treatment recirculation of theliquid is an actual option. The relatively small influence of con-taminant origin on remediation results suggests that in the sludgeresiduals investigated, the contaminant chemistry is dominatedby adsorption and exchange adsorption to the mineral matrix andcomplexation with organic matter. Our work support the hypoth-esis that the mass reduction of the sludge material during EDRis mainly due to dissolution of carbonate; but our work furthershows also other soil constituents are partly dissolved. Further,our work documents that laboratory scale EDR-experiments arerepeatable concerning removal efficiency, pH and conductivitydevelopment.

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