Science of the Total Environment 412-413 (2011) 366–374

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Science of the Total Environment

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Quantifying the environmental impact of As and Cr in stabilized/solidified materials

Milena Dalmacija a,⁎, Miljana Prica b, Bozo Dalmacija a, Srdjan Roncevic a, Mile Klasnja c

a Faculty of Sciences and Mathematics, Department of Chemistry, Biochemistry and Environmental Protection, University of Novi Sad, Trg Dositeja Obradovica 3, Novi Sad, Serbiab Faculty of Technical Sciences, University of Novi Sad, Trg Dositeja Obradovica 6, Novi Sad, Serbiac Faculty of Technology, University of Novi Sad, Bulevar cara Lazara 1, Novi Sad, Serbia

⁎ Corresponding author at: Trg Dositeja ObradovicTel.: +381 214852734; fax: +381 21454065.

E-mail address: milena.dalmacija@dh.uns.ac.rs (M. D

0048-9697/$ – see front matter © 2011 Elsevier B.V. Alldoi:10.1016/j.scitotenv.2011.10.006

a b s t r a c t

a r t i c l e i n f o

Article history:Received 2 May 2011Received in revised form 3 October 2011Accepted 5 October 2011Available online 1 November 2011

Keywords:S/S treatmentLeaching testsBioavailabilityCr and As

The assessment of the quality of sediment from the Great Backi Canal (Serbia) based on the pseudo-total Asand Cr content according to the corresponding Dutch standards and Canadian guidelines showed its severecontamination with these two elements. Microwave assisted BCR sequential extraction procedure wasemployed to assess their potential mobility and risk to the aquatic environment. Comparison of the resultsof sequential extraction and different criteria for sediment quality assessment has led to somewhat contra-dictory conclusions. While the results of sequential extraction showed that Cr comes under the mediumrisk category, As shows no risk to the environment, despite of its high pseudo-total content.The contaminated sediment, irrespective of the different distribution of As and Cr, was subjected to the sameimmobilization, stabilization/solidification (S/S) treatment. Semi-dynamic leaching test was conducted for Asand Cr contaminated sediment in order to assess the long-term leaching behavior of these elements. In orderto simulate “worst case” leaching conditions, the test was modified using acetic acid and humic acid solutionas leachants instead of deionized water. The effectiveness of S/S treatment was evaluated by determining dif-fusion coefficients. Four different single-step leaching tests were applied to evaluate the extraction potentialof As and Cr. A diffusion-based model was used to elucidate the controlling leaching mechanisms. The testresults indicated that all applied S/S treatments were effective in immobilizing As and Cr, irrespective oftheir different availabilities in the untreated samples. In most treated samples, the controlling leaching mech-anism appeared to be diffusion, which indicates that a slow leaching of As and Cr could be expected.

© 2011 Elsevier B.V. All rights reserved.

1. Introduction

Arsenic (As) and chromium (Cr) from various sources (industrial,mining, municipal sewage, agricultural and other activities) have en-tered waterways over time. Chromium often occurs in the trivalent orhexavalent states under natural environmental conditions. The hexa-valent state of chromium, Cr(VI), is highly toxic and a human muta-gen. The trivalent state of chromium, Cr(III), is much less toxic thanCr(VI). Arsenic (As) is a commonly occurring toxic and carcinogenicelement. Even though As occurs in various oxidation states (+5,+3, 0 and −3), arsenite (As(III)) and arsenate (As(V)) are the mostcommon species in natural environments. As (III) is reported to bemore mobile than As (V), and 25–60 times more toxic than As (V)(Jing et al., 2006; Moon and Dermatas, 2007).

Metal contaminants, depending on their physicochemical behav-ior, may adsorb onto suspended solids and subsequently accumulatein sediments, where they may pose a significant contamination prob-lem. In some conditions, more than 99% of heavy metal entering intoriver can be stored in river sediments in various forms. However,

a 3, 21000 Novi Sad, Serbia.

almacija).

rights reserved.

heavy metals cannot fix in sediment forever. With the variation ofthe physical–chemical characteristics of water conditions, part ofthese fixed metals will re-enter the overlying water and becomeavailable to living organisms (Peng et al., 2009). In some cases, dueto the different factors, sediment may need dredging and remediationtreatments.

While various remediation technologies are available for metals,feasible treatment options for As polluted soils and sediments arelimited. Moreover, technologies to simultaneously treat As andmetal contaminants are far from complete. The successful stabiliza-tion of multi-element contaminated sites depends on the combina-tion of critical elements and the choice of amendment (Lee et al.,2011). In this study, a stabilization/solidification (S/S) process wasused in order to remediate As and Cr sediment contamination, be-cause S/S technology has been widely applied to immobilize heavymetals in contaminated soils and sludges with various additives(Conner, 1990; Dermatas and Meng, 2003; Jing et al., 2004; Moonand Dermatas, 2007). At the moment there is not enough dataabout S/S treatments of contaminated sediments, especially thosecontaminated with As and Cr, and the conclusions obtained for soilsthat could be found in literature cannot be valid in full for sediments.

The study examined low-cost effective sorbents (clays, zeolites, flyash) for remediation of contaminated sediment. Fly ash, i.e. the by-

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product produced from fossil fuel thermal power plants, may containincreased amounts of calcium and magnesium oxides, dependingupon the coal sources and may present highly alkaline values similarto lime (Dermatas and Meng, 2003). Thus, fly ash can be used as analternative material for sludge stabilization with additional benefits,such as the reduced purchasing cost, and the minimization of flyash disposal cost. Besides, this method may be proved as an attractivemanagement option for the combined reuse of two solid “waste”materials/by-products, produced in vast amounts worldwide.

The use of cation exchange properties of natural zeolites as stabi-lizing agents has received considerable attraction over the past de-cade in waste treatments (Babel and Kurniawan, 2003; Erdem et al.,2004; Kosobucki et al., 2008; Kumpiene et al., 2008; Shi et al.,2009). Natural zeolites are crystalline microporous aluminosilicateswith very well defined structures and they possess permanent nega-tive charges in their crystal structures that can be balanced by ex-changeable cations.

The large surface area of the natural clays, helped by edges and facesof clay particles, accounts for the excellent capacity of the clay mineralsto adsorb heavy metals (Bhattacharyya and Gupta, 2008). Kaolinite andmontmorillonitewere chosen because they represent the two extremesof physicochemical clay behavior, based on their surface area and cationexchange capacity (CEC). Hence, the effects on As and Cr immobiliza-tion of a relatively non-reactive clay (kaolinite) were compared to ahighly reactive clay (montmorillonite).

Quantifying the environmental impact of S/S materials in real en-vironment scenarios is crucial for selecting proper disposal and reusealternatives and for certification of immobilization technologies. Theperformance of S/S treated wastes is generally measured in terms ofleaching tests (Jing et al., 2004). Batch leaching tests are the preferredchoice for regulatory assessment due to their simplicity, improved re-producibility, and shorter time requirements. However, as batchleaching tests are typically run over short time frames, it is debatablewhether the compounds of interest behave similarly in the long term.

The mechanisms governing heavy metals leachability of contami-nants from monolithic solidified waste forms and evaluation of thelong-term behaviors of S/S wastes can be effectively examined usingthe American Nuclear Society's (ANS) semi-dynamic leaching test(ANS, 1986). The ANS 16.1 provides substantially more informationregarding the “real time” rate at which heavy metals are releasedfrom the solidified product as compared to other leaching tests(Dermatas and Meng, 2003). The leaching results extend over a 90-day period instead of a single result at the end of the test. The mostoften used leaching test, recommended by the USEPA, but which pro-vides only one result for defining the waste toxicity, is the toxicitycharacteristic leaching procedure — TCLP test (USEPA, 2002b). How-ever, it may underestimate the leachability of some redox sensitiveelements, such as As, since redox reactions may happen during theextraction (Meng et al., 2001).The TCLP was specifically designed tomimic acidic conditions in a sanitary landfill and identify wastesthat have potential to contaminate ground water. The waste extrac-tion test — WET is used in California, US, in a similar manner as theTCLP (determination of whether a solid waste is a hazardouswaste), with the exception of the liquid-to-solid ratio (10:1) andthe leaching time, the remainder of the test was the same as theTCLP. Synthetic precipitation leaching procedure — SPLP tests weredeveloped by USEPA in order to assess metal mobility in wastes(USEPA, 2002a, 2002b). The SPLP test reproduces acid rain conditionsand estimates metal mobility when wastes are disposed in an openarea. The SPLP is conducted in a similar fashion as the TCLP with theexception of the leaching fluid.

In view of the above the objectives of this study were: 1) to defineAs and Cr distribution in sediment and evaluate their environmentalrisk; 2) to evaluate the effectiveness of S/S treatments (with the addi-tion of kaolinite, montmorillonite, fly ash and zeolite) 3) to evaluatethe effectiveness of S/S treatment in conditions that mimic the landfill

environment, 4) to determine the controlling leaching mechanisms ofAs and Cr from untreated and treated sediment samples, and 5) tocorrelate the results of sequential extraction procedure and S/S treat-ments, 6) to evaluate leaching potential and environmental impactbased on the different leaching procedures.

2. Methods and materials

2.1. Study area and sample preparation

Sediment samples were collected from the Great Backi Canal(Vojvodina, Serbia).

The canal began to be more intensively polluted at the end of the20th century. Most of the industries (two sugar refineries, a tannery, ametal works, an edible oil refinery, slaughterhouses, etc.) dischargeduntreated or partially treated wastewater into the canal. The total or-ganic pollution from industry is 36.6 tCOD/day or 17.9 tBOD5/day andfrom municipal wastewaters 1329 kg COD/day or 619 kg BOD5/day.Due to the high pressure from the wastewaters in the canal, along a6 km stretch about 400,000 m3 of sediment has been formed. Thecanal is about 3 m deep, with the sediment varying in depth from 1to 2.5 m, depending upon the point of wastewater discharge.

Sediment monitoring is designed in order to obtain informationabout the horizontal and vertical distribution of pollutants in the pol-luted section. The monitoring included seven profiles (profiles weresampled at approximately every 1 km of canal). Each profile was sam-pled with an Eijkelkamp core sampler from 3 to 9 individual samplesof undisturbed sediment, depending upon the thickness of the sedi-ment layer. Composite sample was made up from those individualsamples and as such was used in solidification/stabilization process.

The results of sediment metal pseudo-total concentrations pre-sented in the study are discussed in reference to Dutch RegulationStandards (Ministry of Housing, Spatial Planning and EnvironmentDirectorate-General for Environmental Protection, 2000), and Cana-dian Sediment Quality Guidelines (CCME,1995), since Serbia has nei-ther an established system of continual monitoring of sedimentquality nor regulations concerning the quality standards. Pseudo-total trace metal contents were assessed on sample triplicate afteraqua regia digestion (USEPA, 2007a, 2007b, 2007c) and mean valueswere used. The relative standard deviations (% RSD) obtained(n=3) were below 10%. The sediment pH was measured accordingto ASTM D4972-01 (ASTM, 2007) and it was 7.3±0.4.

S/S treatment was performed by mixing sediment with the follow-ing immobilization agents: fly ash class C, natural zeolite, kaoliniteand montmorillonite. Specimens were designated by the capital let-ters, i.e. K: kaolinite, M: montmorillonite, Z: zeolite, F: fly ash, fol-lowed by a number indicating the percent weight of the givenattribute. The S/S agent content was expressed as percentage of thetotal solids weight. Additionally, the amount of S/S agents presentwas varied in order to evaluate its relative contribution to As and Crimmobilization. During the leaching test, 17 types of specimenswere tested: K10, K20, K30, M10, M20, M30, Z10, Z20, Z30, F10,F20, F30, K5F10, K5F20, Z5F10, Z10F10, and Z5F20.

Samples were prepared in the form of monolithic cubes ((3±0.1)×(3±0.1)×(3±0.1)cm) by compaction at an optimum watercontent, defined as the water content at which the maximum drydensity is achieved for a given compactive effort. The compactionwas performed according to ASTM D1557-00 (ASTM, 2000), provid-ing a compactive effort of 2700 kN mm−3. Samples were cured at20 °C in sealed sample bags for 28 days and then subjected to the dif-ferent leaching tests.

Summary of material chemical and physical properties: class Ccoal fly ash was provided from the Kolubara thermal power plant(Serbia) and its composition was (% wt.): SiO2 (39.4), Al2O3 (20.1),Fe2O3 (4.95), MgO (4.01), CaO (23.2), K2O (0.64), Na2O (2.12),SO3 (1.88). The zeolite (clinoptilolite) composition was as follows

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(% wt.): SiO2 (66.9), Al2O3 (13.5), Fe2O3 (0.98), MgO (0.69), CaO(3.85), K2O (0.54), Na2O (0.37), SO3 (1.18) and ignition loss (11.4).Two types of clay minerals were used (%): kaolinite (SiO2 (45.9),Al2O3 (37.2), Fe2O3 (3.34), MgO (1.40), CaO (0.25), K2O (0.14), Na2O(0.10), ignition loss (13.3)) and montmorillonite SiO2 (58.9), Al2O3

(22.7), Fe2O3 (4.83), MgO (1.40), CaO (1.85), K2O (0.24), Na2O(0.12), ignition loss (10.6).

As noted above, kaolinite and montmorillonite were selected inorder to represent the two extremes of physicochemical clay behaviorbased on their surface area and cation exchange capacity (CEC). TheCEC (meq/100 g) was 5.4 for kaolinite and 90.6 for montmorillonite,while the respective surface areas (m2 g−1) were 66 and 720.

2.2. Microwave assisted sequential extraction procedure (MWSE)

MWSE was performed as described by Jamali et al. (2009), usingidentical operating conditions applied in each individual BCR fraction.Mean values were used and the RSDs (n=3) were below 5%. The ex-tractions were performed at ambient temperature and the optimiza-tion of the MW power and extraction time was carried out bycarefully controlling the temperature of the extracting solutions,which did not exceed 50 °C, and the solutions were never broughtto boiling. Blanks (containing reagent but no samples) were alsotaken through each complete procedure. Mileston, Stare E microwave(MW) was used for MW extraction and digestion.

2.3. Leaching tests

Five leaching methods were used to evaluate the leaching poten-tial of Cr and As in untreated and treated sediment samples asfollows.

2.3.1. ANS 16.1 testThe long-term leachability of As and Cr from the S/S treated mate-

rials was evaluated using the ANS method 16.1 (ANS, 1986). By ap-plying this test we get the cumulative fraction of metals leachedversus time. Mathematical diffusion model based on Fick's secondlaw is used to evaluate the leaching rate as a function of time.

The ANS has standardized the Fick's law-based mathematical dif-fusion model as follows (Godbee and Joy, 1974):

De ¼ πanA0

Δtð Þn

" #2VS

� �2Tn ð1Þ

where an is the contaminant loss (mg) during the particular leachingperiod with subscript n; A0 is the initial amount of contaminant pre-sent in the specimen (mg); V is the specimen volume (cm3); S isthe surface area of specimen (cm2); Δtn is the duration of the leach-ing period in seconds; Tn is the time that elapsed to the middle ofthe leaching period n (s), and De is the effective diffusion coefficient(cm2 s−1).

Due to the slow diffusion rate of contaminants, it can be assumedthat kaolinite, montmorillonite, zeolites and fly-ash based wasteforms are a semi-infinite media (Moon and Dermatas, 2007). This im-plies that the release of the contaminant from the waste form is neg-ligible when compared to the contaminant's total mass. As a result ofthis implication, diffusion is expected to be the controlling leachingmechanism in sediments treated with these S/S agents.

According to the ANS 16.1 test the liquid/solid ratio was 10/1(l kg−1). The leachate was collected and replaced at defined time in-tervals (2, 7, 24, 48, 72, 96, 120, 456, 1128 and 2160 h). The leachateswere filtered through a 0.45 μm pore size membrane filter.

Only for the mixture K10 ANS 16.1 semi-dynamic test was per-formed in triplicate and mean values are presented, the RSD (n=3)being below 5%.

In this study, the ANS 16.1 method was modified by including0.014 M acetic acid (AA) pH 3.25 (similar to Toxicity CharacteristicsLeaching Procedure) and humic acid (HA) solutions (20 mg TOC l−1,pH 6.55) as leachants. The objective was to mimic the worst possibleconditions of the S/S waste disposed in the landfill environment.

2.3.1.1. Determination of the controlling leaching mechanism. The typeof leaching mechanism that controls the release of metals can be de-termined based on the values of the slope of the logarithm of cumu-lative fraction release, log(Bt), versus the logarithm of time, log(t),line (de Groot and van der Sloot, 1992).

If diffusion is the dominant mechanism, then theory suggests thefollowing relationship:

log Btð Þ ¼ 12log tð Þ þ log Umaxd

ffiffiffiffiffiffiffiffiffiffiffiffiffiDe

π

� �s" #ð2Þ

where De is the effective diffusion coefficient in m2 s−1 for compo-nent x; Bt is the cumulative maximum release of the component inmg m−2; t is the contact time in seconds; Umax is the maximum leach-able quantity in mg kg−1, and d is the bulk density of the product inkg m−3.

When the slope is close to 1 (0.60–1.00), according to de Grootand van der Sloot (1992) the process is defined as dissolution. Inthat case, the dissolution of the material proceeds faster than the dif-fusion. If the slope is around 0.50 (0.40–0.60), the release of heavymetals will be slow and diffusion will be the controlling mechanism.If the slope is less than 0.40 the release of metal will be probablydue to surface wash-off. Typically, the long-term leaching characteris-tics of S/S treated wastes are controlled by diffusion. However, thereare cases when the other processes, dissolution and wash-off, mayalso occur (Moon and Dermatas, 2007).

2.3.2. TCLP testAccording to the USEPA protocol (USEPA, 2002b) a 0.l M acetic

acid solution with a pH of 2.88 was used to extract control sampleand S/S-treated samples since the pH was above 5. The sedimentsamples were extracted at a liquid to solid (L/S) ratio of 20 in cappedpolypropylene bottles on a rotary tumbler at 30 rpm for 18 h. Afterthe extraction, the final pH of the leachate was measured and the liq-uid was separated from the solids by filtration through a 0.45 μm poresize membrane filter. This test was applied on every sample in tripli-cate and the RSDs were below 5%.

2.3.3. Synthetic precipitation leaching procedure — SPLP testThe SPLP test is performed in the same manner as the TCLP. The

extraction fluid is made of two inorganic acids (nitric and sulfuricacid) to simulate acidic rainwater (pH 4.2). In a similar fashion asthe TCLP, a 100-g sample of waste material is placed in a 2-liter ex-traction vessel and mixed with the extraction fluid. The mixture is ro-tated for 18± 2 h at 30 rpm. The leachate is then filtered through a0.45 μm pore size membrane filter (USEPA, 2002a). This test was ap-plied on every sample in triplicate and the RSDs were below 5%.

2.3.4. Waste extraction test — WETThe WET (CCR, 1998) is similar to the TCLP in that it uses a buff-

ered organic acid solution as the extraction fluid. This test uses a pHbuffered citrate acid solution with sodium hydroxide, a 10:1 L/Sratio, and a 48-h testing period. The WET extraction solution isprepared with a combination of 0.2 M citric acid solution and 4.0 NNaOH to pH 5.0±0.1. One liter of this solution is added to a 100-gsample and rotated for 48 h. After rotation, the final pH is measured,and the samples were filtered through a 0.45 μm pore size membranefilter. This test was applied on every sample in triplicate and the RSDswere below 5%.

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2.3.5. Bioavailability testSamples were subjected to a single dilute HCl chemical digestion.

This involved mixing 10 ml of 0.5 M HCl with 0.5 g of sediment andshaking at room temperature for 1 h (Sutherland, 2002). After the ex-traction, the final the liquid was separated from the solids by filtrationthrough a 0.45 μm pore size membrane filter. This test was applied onevery sample in triplicate and the RSDs were below 5%.

Pseudo-total sediment metal content, metal content in sequentialextraction steps and metal content in leachates were determined byAAS (Perkin Elmer AAnalyst™ 700) or ICP-MS (Perkin Elmer SciexElan 5000) according to the standard procedures (USEPA, 1994,2007a, 2007b, 2007c). All the results are expressed with respect tosediment dry matter. All materials in contact with the leachantwere pre-cleaned with HNO3 and subsequently rinsed with deionizedwater.

3. Results and discussion

3.1. Pseudo-total metal concentrations and sequential extraction ofuntreated sample

The pseudo-total contents of Cr and As in the untreated sedimentsample were 1335 mg kg−1 and 63.9 mg kg−1, respectively, whichaccording to theDutch regulation standards (Ministry of Housing. SpatialPlanning and Environment Directorate-General for EnvironmentalProtection. Circular on target values and intervention values for soilremediation. Netherlands Government Gazette, 2000), is above inter-vention value and such sediment is considered severely polluted withCr and As (class 4) and needs dredging, disposal in special reservoirsand, if possible, sediment clean-up measures. Compared with CanadianSediment Quality Guidelines (CCME, 1995) for aquatic life protection,contents of Cr and As were above the probable effect level (PEL). Sedi-ment concentrations above PEL values are expected to be frequently as-sociatedwith adverse biological effects. Although PEL is considered to beapplicable to a variety of sediment types, it cannot define uniform valuesof sediment pollution, as the bioavailability (and hence toxicity) of con-taminants can differ (CCME, 1995).

The distribution of Cr and As in different phases extracted by MWassisted sequential extraction procedure is shown in Fig. 1.

Sequential extraction procedure partitions the metals into fourfractions: exchangeable and carbonate bound, iron and manganeseoxides bound, organic matter bound and residual metal (Tessier etal., 1979; Quevauviller, 2002).

The results of this method are not in full agreement with the re-sults of pseudo-total metal concentration in the sediment, whichonly confirms the opinion that the total metal concentration is notsufficient to define the real danger to the environment. Judgingfrom the results of sequential extraction (Fig. 1) chromium is at thehighest percentage present in the fourth phase (65%) which is the

Fig. 1. Distribution of As and Cr in sequentially extracted fractions of untreated sedi-ment sample.

environmentally least dangerous phase, since this fraction representsthe stable metal forms associated with anthropogenic or geogeniccomponents (Quevauviller, 2002). The first fraction is the most dan-gerous for environment and significant proportion of Cr (18%) is pre-sent in the exchangeable forms. Metals in exchangeable fraction canbe exchanged and are in equilibrium with the ionic content inwater. This fraction is sensitive to pH variations.

On the other hand, As was mainly bound to Fe–Mn oxides (46%)and organic matter (47%). The Fe–Mn oxides fraction includes the sol-uble metal oxides/hydroxides under slightly acidic pH as well as themetal associated with reducible amorphous Fe–Mn oxyhydroxides.The low level of As in step 1 of the sequential extraction could beexplained by the high percentage of As associated with the iron-oxides (Fitz and Wenzel, 2002; Kumpiene et al., 2008; Beesley et al.,2010).

As to study metal retention, individual contamination factors (Cf)were calculated. These factors are defined as the sum of heavy metalconcentration in the mobile phases (non-residual phases) of the sam-ple divided by the residual phase content. Then, the lower the Cfvalue the higher the relative metal retention (Barona et al., 1999).The obtained results showed that the relative metal retention wasfor As 15.1 and for Cr 0.54 indicating larger mobility of As.

According to the risk assessment code (RAC), 18% of chromium incarbonate phases comes under the medium risk category (Jain, 2004).On the other hand, As shows low risk to the environment with 0.8% incarbonate fraction.

A potential method to determine if the heavy metals can be re-moved by remediation techniques or predict removal efficiency is todetermine metal distribution in sediment with selective extractiontechniques (Mulligan et al., 2001). Although As and Cr distributionin sediment was different, we applied the same remediation treat-ments because there are not enough data about the behavior ofmetals differently distributed in their mixture in a sediment duringtheir S/S treatment, and about the treatment efficiency in general(Lee et al., 2011). The main objective of every remediation proceduredealing with several contaminants is to carry out the treatmentwith the same agents, and thus achieve economic and environmentalbenefits.

3.2. ANS 16.1 test

The cumulative values of As and Cr leachability from the speci-mens treated with kaolinite, montmorillonite, zeolites, fly ash, andzeolites-fly ash combinations are presented in Table 1.

The amount of As and Cr released during the ANS 16.1 tests (ANS,1986) for the untreated samples did not exceed 20% of the total massof the contaminant in the waste (Table 1), which is the upper limit forthe diffusion model to be still applicable.

The clays employed exhibited good sorption of As and Cr, reducingsignificantly their leachability compared to untreated sample. This isin a good correlation with literature data (Dermatas and Meng,2003; Saada et al., 2003; Dermatas et al., 2004; Moon et al., 2004;Moon and Dermatas, 2006; Turan et al., 2007; Giacomino et al.,2010). Based on the results (Table 1), the As and Cr leachability isinfluenced by the clay type. Montmorillonite treated samples showedslightly higher reduction in As and Cr leachability compared to thosetreated with kaolinite. This is most probably due to the larger surfacearea and the greater CEC of montmorillonite (Dermatas et al., 2004;Moon and Dermatas, 2006, 2007). Amount of metal leached in somecases increases with high sorbent loading (Table 1). Similar resultshave been reported by other authors (e.g. Cu(II) by sawdust, Yu etal., 2000; Cr (III) by ion exchange resins, Rengaraj et al., 2002; Bhatta-charyya and Gupta, 2008). This may be attributed to two reasons: (i)a large adsorbent amount effectively reduces the unsaturation of theadsorption sites and correspondingly, the number of such sites perunit mass comes down resulting in comparatively less adsorption at

Table 1Total cumulative fraction of Cr and As leached (%) after ANS 16.1 test completion usingdeionized water (DI), acetic acid (AA) and humic acid (HA) solutions as leachants.

Cr As

DI AA HA DI AA HA

Untreated sediment sample 11.9 15.8 12.4 15.4 18.4 17.7K10 0.12 0.21 0.11 1.87 3.87 3.47K20 0.19 0.25 0.14 2.14 4.93 3.68K30 0.21 0.29 0.17 2.83 5.28 4.78M10 0.10 0.17 0.12 2.19 4.12 3.53M20 0.08 0.14 0.09 1.73 3.86 3.15M30 0.06 0.12 0.07 1.51 3.27 3.01Z10 0.13 0.37 0.26 1.85 4.20 3.33Z20 0.10 0.30 0.22 1.42 3.91 2.67Z30 0.04 0.24 0.19 1.21 3.21 2.32F10 0.09 0.18 0.14 1.18 4.64 2.91F20 0.06 0.14 0.10 0.70 3.56 1.87F30 0.12 0.25 0.09 2.20 6.39 5.58K5F10 0.09 0.26 0.16 7.96 8.29 8.06K5F20 0.02 0.22 0.12 3.89 7.81 5.91Z5F10 0.06 0.28 0.12 6.27 9.18 7.66Z10F10 0.02 0.19 0.08 5.32 8.83 7.93Z5F20 0.03 0.22 0.05 5.84 6.78 6.35

Table 2Slope and R2 values obtained from the diffusion model.

Cr As

Slope R2 Slope R2

Untreated SS 0.76 0.98 0.18 0.99K10 0.36 0.96 0.21 0.96K20 0.38 0.95 0.27 0.99K30 0.41 0.86 0.59 0.95M10 0.40 0.98 0.49 0.98M20 0.50 0.95 0.60 0.93M30 0.48 0.99 0.58 0.94Z10 0.50 0.85 0.56 0.84Z20 0.51 0.88 0.41 0.90Z30 0.58 0.86 0.46 0.91F10 0.40 0.85 0.46 0.91F20 0.43 0.92 0.44 0.95F30 0.47 0.94 0.45 0.98K5F10 0.41 0.92 0.38 0.86K5F20 0.40 0.93 0.39 0.88Z5F10 0.56 0.96 0.62 0.92Z10F10 0.50 0.99 0.63 0.85Z5F20 0.51 0.99 0.65 0.95

370 M. Dalmacija et al. / Science of the Total Environment 412-413 (2011) 366–374

higher adsorbent amount, and (ii) higher adsorbent amount createsparticle aggregation, resulting in a decrease in the total surface areaand an increase in diffusional path length both of which contributeto decrease in amount adsorbed per unit mass (Shukla et al., 2002).

Arsenic and chromium leachability was even lower (1.21% and0.04% respectively) when the samples were treated with zeolites.With the increasing of zeolite percentage in mixtures there was a de-crease in the percentage of cumulative leached As. Shanableh andKharabsheh (1996) concluded the same when they examined thepossibility of using zeolite in the remediation of soil contaminatedwith cadmium, nickel and lead, as there is no sufficient data on the re-mediation of sediments contaminated with Cr and As using zeolite.

Fly ash treatment was effective for the sediment samples, espe-cially F20 with 0.06% released of Cr and 0.70% released of As. Themechanisms of reduced metal availability might be attributed tohigh pH values, as well as to physical adsorption (Chlopecka andAdriano, 1996). Moreover, increasing the amount of fly ash in somemixtures (K5F20, Z5F10, Z5F20) led to a reduction in leached Asand Cr, most probably as a result of the calcium present in fly ash.

Humic acid leachant releases more As than deionized water in allsamples. That is also true for chromium in most samples. Humic acidsare very important in the formation of stable organo-mineral com-plexes due to their physicochemical and biological stability. Due totheir amphipathic nature and structural features, HAs play an impor-tant role in environmental processes governing the fate and transportof organic and inorganic pollutants in natural systems (Stevenson,1994). They include binding sites with different complexion strength,able to form inert and labile complexes with inorganic cations and or-ganic compounds (Yamamoto and Ishiwatari, 1992). Based on someprevious studies (Kartal et al., 2007; Kim et al., 2007; Zaccone et al.,2009), As shows pronounced tendency for complexing with humicacids. Also, leachant might be richer in As because of the greater oc-currence in Fe and Al, acting as bridge between HA molecules andAs oxianions (Zaccone et al., 2009). This could be also supported bythe results of sequential extraction procedures because As is presentin more available phases compared to chromium (high percentagein residual phase). Also previous studies have reported that soil OMcan affect Cr mobility due to the formation of high molecular weightinsoluble complexes with HA (Zaccone et al., 2009).

Overall all the treatments were efficient having in mind that sed-iment sample had high initial As and Cr concentrations. This may bedifficult to readily explain due to the degree of complexity of the nat-ural sediment samples. In nature there are many constituents that

could participate in and influence As and Cr leachability (de Grootand van der Sloot, 1992). Further research is required to obtainsome of this information in order to evaluate the As and Cr oxidationspeciation and subsequent mechanisms of incorporation and releasein these samples.

3.2.1. The controlling leaching mechanismThe controlling leaching mechanisms were evaluated using a dif-

fusion model developed by de Groot and van der Sloot (de Grootand van der Sloot, 1992) for the case of leaching with deionizedwater. The slope and R2 values obtained from the diffusion modelare presented in Table 2.

For untreated samples, slope values were 0.76 for Cr and 0.18 forAs which indicates that the dominant leaching mechanism for As issurface wash-off and dissolution for chromium. For the most treatedsediment samples slope values were in the range of 0.40 to 0.60which indicates that the dominant leaching mechanism is diffusion.Exceptions are the mixtures K10 and K20 in case of chromiumwhere slope values were 0.36 and 0.38 respectively and for mixturesK10, K20, K5F10, and K5F20 in case of As slope values were in therange of 0.21 to 0.39. On the basis of these values it can be concludedthat the release from these treated samples was controlled by surfacewash-off. Also, based on the slope values for mixtures ZF in case of As(0.62–0.65) it can be concluded that the dominant leaching mecha-nism is dissolution.

Further research should be focused on a more detailed analysisaiming at the elucidation of the encapsulation of metals into thestructure and their leaching mechanism, relying on the studiesof mineralogy (qualitative and quantitative X-ray diffraction) aswell as micromorphology (scanning electron microscopy and opticalmicroscopy).

3.2.2. Effectiveness of S/S treatmentDiffusion coefficients De for untreated and treated samples are

listed in Table 3.The diffusion coefficients generally vary from 10−5 cm2 s−1 (very

mobile) to 10−15 cm2 s−1 (immobile) (Nathwani and Phillips, 1980).The diffusion coefficients for Cr in treated samples ranged from10−13 cm2 s−1 (low mobility) to 10−15 cm2 s−1 (immobile) and forAs from 10−10 cm2 s−1 to 10−13 cm2 s−1 (low mobility). Diffusioncoefficients are in agreement with the results of cumulative releaseof the two metal ions from the S/S treated mixtures.

Table 3Mean diffusion coefficients, De (cm2 s−1) for untreated and treated sediment sampleswhen deionized water (DI), acetic acid (AA) and humic acid (HA) solutions were usedas a leachants.

Cr As

DI AA HA DI AA HA

Mean De (cm2 s−1)

Untreated SS 2.67E-09 1.46E-08 7.11E-09 1.54E-08 1.29E-07 6.83E-08K10 1.58E-13 1.96E-13 1.26E-13 1.58E-12 6.38E-12 3.98E-12K20 2.51E-13 3.15E-13 1.57E-13 1.94E-12 7.94E-12 5.04E-12K30 5.01E-13 3.96E-13 3.16E-13 3.16E-12 1.26E-12 6.26E-12M10 1.26E-13 3.17E-13 1.51E-13 1.89E-12 1.28E-11 3.98E-12M20 7.94E-14 1.94E-13 1.24E-13 1.54E-12 7.94E-12 1.94E-12M30 6.31E-14 1.56E-13 1.02E-13 7.96E-12 5.06E-12 1.16E-12Z10 3.96E-14 1.26E-13 7.96E-13 1.91E-12 1.26E-11 3.15E-12Z20 3.16E-14 1.06E-13 6.36E-13 1.26E-12 1.04E-11 2.54E-12Z30 1.51E-14 5.02E-14 3.12E-14 6.36E-13 6.28E-12 1.63E-12F10 3.90E-14 1.03E-13 5.06E-14 1.58E-13 6.26E-12 3.89E-13F20 3.23E-14 7.93E-14 3.93E-14 1.54E-12 1.54E-12 1.54E-12F30 1.21E-13 3.20E-13 1.60E-13 7.96E-12 3.96E-12 6.26E-12K5F10 2.51E-14 6.31E-14 3.27E-13 2.48E-10 4.68E-10 3.31E-10K5F20 6.23E-15 1.53E-14 1.03E-14 1.54E-11 1.34E-10 3.14E-11Z5F10 2.51E-14 6.29E-14 3.21E-13 1.06E-10 7.26E-10 1.62E-10Z10F10 7.94E-15 1.54E-14 1.24E-14 7.98E-11 1.88E-10 1.48E-10Z5F20 1.23E-14 2.53E-14 1.93E-14 6.34E-11 5.34E-10 1.21E-10

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3.3. Comparison of three-stage sequential extraction and leaching tests toevaluate metal mobility in treated sediment samples

The results of TCLP, SPLP, WET and step 1 BCR on treated samplesare presented in Fig. 2.

For untreated sediment sample extracted concentrations forchromium were: 244.2 mg kg−1 (TCLP), 202.5 mg kg−1 (SPLP),345.5 mg kg−1 (WET) and 239.2 mg kg−1 (Step 1 BCR). Arsenic

Fig. 2. Comparison of the results of SPLP, WET,

extracted concentrations were: 28.5 mg kg−1 (TCLP), 25.8 mg kg−1

(SPLP), 31.3 mg kg−1 (WET) and 3.19 mg kg−1 (Step 1 BCR).The TCLP metal concentrations in all samples are lower than those

obtained by Step 1 BCR due to the smaller solid:liquid ratio and aceticacid solution used during the extraction. This is in agreement with theliterature data (Marguí et al., 2004). The regulations require the con-centrations of specific compounds in the TCLP leachate to be com-pared to TC concentrations in the regulations (USEPA, 2002b). BothAs and Cr have TC concentration limits of 5 mg/L. None of the samplesexceeded 5 mg/L for Cr and As.

Lower concentrations of As and Cr were leached in SPLP thanthe TCLP test in most samples. This is consistent with literature(Townsend et al., 2004; Janin et al., 2009). Differences in metal leach-ability between TCLP and SPLP might result from several factors.Depending on the alkalinity of the waste tested, changes in the solu-tion pH that occur during the 18 h of leaching may differ betweenSPLP and TCLP, and thus result in different amounts of metal leaching.Another factor is the complexation ability of the acid used in theleaching fluid. The anions resulting from organic acids such as citricor acetic acid can complex metals causing them to leach in greaterconcentrations (Townsend et al., 2004).

The WET extracted higher concentrations of As and Cr than boththe TCLP and the SPLP in most samples, but below WET limit values(CCR, 1998). This is consistent with literature (Townsend et al.,2004, 2005). TCLP and SPLP are conducted at a 20:1 liquid-to-solidratio and WET is carried out at 10:1 liquid to solid ratio; the TCLPand SPLP are twice diluted compared to WET and in general higherleachate concentrations are observed at lower liquid-to-solid ratio(Townsend et al., 2004). The greater element concentrations ob-served in the WET leachates relative to the TCLP and SPLP leachatesmost likely result, however, from citrate's propensity to chelateAs and Cr. Other leaching studies support this conclusion (Hooper

and TCLP leaching test and step 1 of BCR.

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et al., 1998; Townsend et al., 2004). Hooper et al. (1998) reported thatacetic acid failed to complex oxyanion-forming elements such as Asand Cr. Citric acid has multidentate ligands while acetic acid hasmonodentate ligands, and in general, complexes with monodentateligands are less stable than those with multidentate ligands (Stummand Morgan, 1996).

While the TCLP, SPLP, andWET leaching tests are rapid they do notsimulate more complex environmental settings. SEP provides usefulinformation for risk assessment since the amount of metal availableunder different environmental conditions can be estimated. Sequen-tial extractions provide semi-quantitative information on elementdistribution between operationally defined geochemical fractions.Therefore, the fractions obtained from sequential extractions do notnecessarily reflect true chemical distribution.

The physico-chemical conditions in sequential extraction experi-ments (strong reagents and rapid reactions) differ from natural con-ditions (weak reagents and slow reactions). Although leachingtechniques such as column leaching tests are probably more realisticto field conditions, sequential extractions and single step extractionscan give an indication of the ‘pools’ or ‘sinks’ of heavy metals thatare potentially available under changing environmental conditions.

However, the direct comparison between methods is difficult tocarry out, especially when different reagents are applied to extracta specific phase or when reagents with different concentrations areused in the methods to be compared.

3.4. Bioavailability test

Fig. 3 shows 3-step sequential extraction procedure and a single0.5 M HCl leach comparison. As can be seen from the figure, the single

Fig. 3. Element concentrations liberated by a 3-step sequential extraction

0.5 M HCl leach was slightly more aggressive than the 3 phases of thesequential procedure, with greater concentrations of As and Cr. Thereis not enough literature data related to this data for As and Cr, butthere are for some other metals that showed similar behavior whencomparing single-step extraction with HCl and the sum of three step(Sutherland, 2002). The single 0.5 M HCl leach of urban media hasbeen shown to be an effective, inexpensive and rapid approach for in-organic contaminant assessment for routine pollution monitoring, forinitial reconnaissance surveys, or for geochemical anomaly identifica-tion. However, once highly contaminated (anomalous) sites havebeen identified by the dilute HCl leach, they may be further examinedin a cost-effective manner by sequential extraction procedures.

4. Conclusion

The assessment of the sediment quality based on the pseudo-totalAs and Cr content according to the corresponding Dutch standards andCanadian guidelines showed its severe contamination. The modifiedBCR fractionation scheme was employed to determine exchangeable(soluble) oxidizable, reducible and residualmetal fractions in contam-inated sediment from the Great Backi canal (Serbia). Metals exhibiteddifferent fractionation profiles.While Aswas foundmost in the oxidiz-able and reducible fraction, Cr was present mainly in the residual frac-tion. In this study, based on the RAC, chromium comes under themedium risk category and As shows no risk.

The immobilization treatment applied appeared to be efficient inthe remediation of sediment contaminated with As and Cr. Thesame treatment was applied despite of the fact that the results atthe beginning of the experiment gave different distribution and avail-ability of metals. The TCLP, SPLP, WET and bioavailability test with

procedure (‘sequential’) and a single 0.5 M HCl leach (‘single leach’).

373M. Dalmacija et al. / Science of the Total Environment 412-413 (2011) 366–374

0.5 N HCl were applied to evaluate the extraction potential of As andCr. The results showed that all S/S samples can be considered as non-hazardous. The single 0.5 M HCl leach of urban media has been shownto be an effective, inexpensive and rapid approach for inorganic con-taminant assessment.

Based on the diffusion coefficient and results of some other leach-ing tests it appeared that the distribution of metals in the startingsediment did not influence the efficiency of treatments. This is advan-tageous from an economic point of view, and may justify the applica-tion of the already expensive remediation procedure, especially whenit comes to treat a material containing a mixture of pollutants. In mostsamples the controlling leaching mechanism of metals upon the S/Streatment appeared to be diffusion. Hence, only small amounts ofmetals could be expected to leach into the environment over time.

Acknowledgment

Special thanks to Ministry of Science and Technological Develop-ment (grant nos. III43005 and TR3700)

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