batch studies on arsenic adsorption onto lignite...

J. SE Asian Appl. Geol., Jul–Dec 2012, Vol. 4(2), pp. 53–67

BATCH STUDIES ON ARSENIC ADSORPTIONONTO LIGNITE, BENTONITE, SHALE ANDIRON SAND: EFFECTS OF pH, TIME, PARTICLESIZE AND SULFATE CONCENTRATION

Kyu Kyu Mar∗1,3, Dwikorita Karnawati1, Doni Prakasa Eka Putra1, Sarto2, ToshifumiIgarashi3, and Carlito Baltazar Tabelin3

1Geological Engineering Department, Faculty of Engineering, Gadjah Mada University, Yogyakarta, Indonesia2Chemical Engineering Department, Faculty of Engineering, Gadjah Mada University, Yogyakarta, Indonesia3Division of Sustainable Resources Engineering, Faculty of Engineering, Hokkaido University, Kita 13 Nishi 8,Kita-Ku, Sapporo 060-8628, Japan

Abstract

Arsenic (As) is a toxic element found in both naturaland anthropogenic sources. High concentration ofthis element was recently uncovered in the ground-water of Sumbawa Island, Indonesia. To mitigatethis problem, As adsorption potential of natural ge-ological materials like lignite, bentonite, shale, andiron sand obtained in Indonesia were evaluated bybatch experiments. Arsenic adsorption onto thesematerials was investigated as a function of solu-tion pH, particle sizes of adsorbents and coexistingsulfate concentration. In addition, batch leachingexperiments were performed to elucidate the stabil-ity of geogenic As present in all adsorbents at dif-ferent pHs. The results showed that among thesenatural materials tested, lignite was the most ef-fective adsorbent of As(V) followed by bentonite,shale and then iron sand, and that the amountsof As(III) adsorbed onto all adsorbents were lowerthan those of As(V).This indicates that As(III) ismore mobile in comparison to As(V). The adsorptionisotherms of As(III) and As(V) conformed to non-linear types, either Langmuir or Freundlich. It wasfound that adsorption of As onto these natural adsor-bents was pH-dependent. This could be attributed to

∗Corresponding author: K.K. MAR, Department ofGeological Engineering, Faculty of Engineering, GadjahMada University, Jl. Grafika 2 Yogyakarta, 55281, Indone-sia. E-mail:[email protected]

the changes in the surface charges of the adsorbentswith pH. With respect to the adsorbent particle size,the amount adsorbed somewhat increased with de-creasing particle size, which could be explained bythe larger surface area of the smaller particles. Acidic(pH < 6) and alkaline (pH >10) conditions destabi-lized the geogenic As content of the adsorbents, in-dicating that the effectiveness of these natural mate-rials as adsorbents is greatly limited by the pH of thecontaminated system.Keywords: Adsorption, arsenic, natural geologicalmaterials, particle size, pH

1 Introduction

Arsenic (As), which is toxic to most living or-ganisms, creates potentially serious problemsthroughout the world. Arsenic contaminationof the groundwater has been reported in manycountries in different parts of the world. For ex-ample, high concentration of As in the ground-water of many countries like India (Das et al,1994; Chattarjee and Mandal, 1995), Bangladesh(Karim, 2000; Smedley and Kinniburgh, 2002),Taiwan (Chen et al, 1994; Hricko, 1994), Mongo-lia (Fujinoa et al, 2004), and China (Hricko, 1994;Niu et al, 1995) used for drinking and cookinghas caused symptoms of chronic As poisoninglike arsenicosis and keratosis to the local hu-

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man population. Recent advances in the anal-ysis of As as well as more detailed studies of itshuman health effects show that As is carcino-genic even at minute concentrations; so that theWorld Health Organization (WHO) lowered theprovisional guideline for As concentration indrinking water from 50 to 10 µg/L that wasimmediately adopted by developed countrieslike Japan and USA (WHO, 1993; USEPA, 2002).Arsenic can be released from natural sourcesas well as a variety of anthropogenic activitiesincluding mining, agriculture and combustionof coal contribute to the As contamination ofthe surrounding soil and groundwater (Fergu-son and Gavis, 1972). In addition, dependingon the redox conditions, inorganic As occurs ingroundwater mostly as arsenate [As(V)] or ar-senite [As(III)]. Both oxidation states are oftenobserved due to their relatively slow arsenic re-dox transformations (Smedley and Kinniburgh,2002; Lai et al, 2004). These redox transfor-mations of As are important because As(III) ismore mobile and toxic than As(V).

Removal of hazardous elements like As fromaqueous solutions can be achieved by differenttechnological methods which include chemicalprecipitation, ion exchange, adsorption, reverseosmosis, modified coagulation/filtration, mod-ified lime softening, electro deposition and oxi-dation/filtration (USEPA, 2002). Among them,significant attention has centered on adsorp-tion because of its simplicity, cheaper pollu-tion control, ease of operation and handling,sludge-free operation, and regeneration capa-bility. Several solid materials have been em-ployed as adsorbents for As like activated alu-mina, fly ash, pyrite fines, manganese green-sand, amino-functionalized mesoporous silicas,clinoptilolite and other zeolites, iron oxides, ac-tivated carbon and zero-valent iron (Subrama-nian et al, 1997; Yoshitake et al, 2003; Elizalde-Gonzalez et al, 2001; Bang et al, 2005; Lackovicet al, 2000). Although these adsorbent materialsare effective, most of them are expensive. It isnecessary to substitute naturally occurring ma-terials for effective but expensive adsorbents.However, investigations involving these natu-rally occurring materials are few and most ofthese studies have been conducted using min-

erals synthesized in the laboratory (Baskan andPala, 2011; Shevade and Ford, 2004; Pena et al,2005).

Although most environmental As problemsare the result of mobilization under natural con-ditions, contamination due to mining activitiesis also significant. For example, high concen-tration of As in the groundwater of SumbawaIsland, Indonesia has recently been uncovered.To mitigate this problem, the authors envisionthe use of a permeable reactive barrier (PRB)for the immobilization of As. However, a suit-able adsorbent for these PRBs is still not avail-able. Therefore, this study aims to characterizeseveral natural geological materials and to eval-uate their As(V) and As(III) adsorption prop-erties. In the present study, lignite, bentonite,shale and iron sand are used as adsorbents ofAs removal from aqueous solution because theyare cheap and readily available and can be ap-plied to the remediation of contaminated sitesin Indonesia. The effects of solution pH andshaking time on As were also evaluated be-cause these parameters are known to be sensi-tive. In addition, the effects of particle sizes ofthe adsorbents and SO2−

4 concentration on theadsorption of As were also elucidated. Specialconsideration was given to SO2−

4 because it hasbeen reported as having a moderate negative ef-fect on the adsorption of As onto Fe and Al oxy-hydroxides and the contaminated groundwaterin Sumbawa Island has substantial SO2−

4 con-centration (ca. 1500 mg/L) (LAPI-ITB, 2009).Moreover, equilibrium isotherm models wereapplied to predict the adsorption capacity ofeach adsorbent. Finally, we elucidated the sta-bility of geogenic As found in all adsorbents atdifferent pH values.

2 Materials and Methods

2.1 Sample collection

The materials used as adsorbents in this re-search were natural geological materials col-lected in Indonesia: lignite from Samigaluh,Kulon Progo area; bentonite from Punung, Pac-itan, East Java; shale from, Bogor, West Java;and iron sand from Glagah beach, Kulon Progo

54 © 2012 Department of Geological Engineering, Gadjah Mada University

BATCH STUDIES ON ARSENIC ADSORPTION ONTO LIGNITE, BENTONITE, SHALE AND IRON SAND

Figure 1: Location of the sampled places of adsorbents.

area. The location of the sampled places isshown in Figure 1.

2.2 Sample preparation and characterization

Prior to experiments, the geological materialswere air-dried, crushed and sieved through0.125, 0.25, 0.5 and 1 mm aperture screensto segregate into different particle sizes. Forthe chemical and mineralogical analysis, partsof these materials were further ground into<50 µm. To identify the mineralogical com-position, an X-ray diffractometer (XRD), Mul-tiflex (Rigaku Corporation, Japan), was em-ployed while the chemical composition wasdetermined using an X-ray fluorescence spec-trometer, Spectro Xepos (Rigaku Corporation,Japan). After drying the samples (<2 mm) inan oven at 110◦C for 24 hr, loss on ignition (LOI)was determined through gravimetry by heatingthe sample inside a furnace at 750◦C for 1 hr.Total organic carbon (TOC) content of all ad-sorbents was analyzed using TOC-L Total Or-ganic Carbon Analyzer coupled with a SolidSample Module SSM-5000A (Shimadzu Corpo-ration, Japan). Specific surface area and totalpore volume of the adsorbents were measuredby N2 adsorption using the BET method. Afterdegassing under vacuum at 120◦C, the amount

of adsorbed N2 was determined at constanttemperature (77 K) by using Beckman CoulterSA 3100 (Beckman Coulter Inc., CA). Variationsof the zeta potential with pH were obtained us-ing NanoZS 90 (Malvern Instruments Ltd., UK).The amount of sample used in this analysis was10 mg with particle size of <50 µm. The pH wasadjusted using 0.1 M of hydrochloric acid (HCl)or sodium hydroxide (NaOH) solutions.

2.3 Adsorption experiments

Batch adsorption tests were performed to de-termine the adsorption performance of the ge-ological materials to remove As. Both As(V)and As(III) solutions were prepared by dissolv-ing reagent grade disodium hydrogen arsenateheptahydrate [Na2HAsO4· 7H2O], and dilutingarsenic standard solution for atomic adsorp-tion spectrometry (As2O3 and NaOH in wa-ter, pH 5.0 with HCl, Wako Pure ChemistryIndustry Ltd., Japan), respectively. The ad-sorption experiments were carried out in 250mL flasks. One gram of each material wasadded to 100 mL of As(V) or As(III) solutionwith concentrations of 1–50 mg/L, and then themixture was shaken at room temperature for24 hr at 120 rpm until adsorption equilibriumwas attained. After mixing, the pH and redox

© 2012 Department of Geological Engineering, Gadjah Mada University 55

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potential (ORP) of suspensions were immedi-ately measured and the suspensions were fil-tered using 0.45µm Millex® sterile membranefilters ((Merck Millipore, USA). The suspen-sions of bentonite and shale samples were cen-trifuged at 3,000 rpm for 20–25 min before fil-tration. Arsenic concentrations before and af-ter mixing were analyzed using an inductivelycouples plasma atomic emission spectrometer(ICP-AES) (ICPE-9000, Shimadzu Corporation,Japan).

The effect of shaking time on the adsorbedamount of As(V) and As(III) were investigatedusing the best three adsorbents of As: lignite,bentonite and shale. This was done by using aconstant As initial concentration and amount ofadsorbent added. For As(V), 0.5 g of lignite, 1g of bentonite or 1 g of shale were added to 2mg/L of As(V) solution. On the other hand, 2g of each adsorbent were mixed with 2 mg/Lof As(III) solution. The mixing period rangedfrom 1 to 48 hr.

The effect of pH on As adsorption was ex-amined by setting initial As concentration rang-ing 1 to 5 mg/L and adsorbent added rang-ing 1 to 2 g. The contact time was fixed at24 hr. The initial pH of the solution was ad-justed by dropping hydrochloric acid (HCl) orsodium hydroxide (NaOH) solution resultingin suspension pH values from 2–10.

To investigate the effects of particle size, thebest two adsorbents of As(V) (i.e., lignite andbentonite) were used. In addition to these twoadsorbents, shale was also used in the As(III)adsorption experiments because there were notso significant differences in the amounts ofAs(III) onto on these three adsorbents. The ad-sorption experiments were carried out with ad-sorbents of variable sizes (0.125–1mm) by usingthe same initial As concentration and adsorbentdosage. For the adsorption of As(V), 1 g of lig-nite or bentonite were added to 100 mL of Assolution with concentrations of 5 and 3 mg/L;respectively. For As(III) adsorption, 100 mL ofAs(III) solution with concentration of 2 mg/Land 2 g of adsorbent were mixed.

The effects of SO2−4 concentration were inves-

tigated using solutions with variable concentra-tions of SO2−

4 (500, 1000, 1500 and 2000 mg/L)

and 10 mg/L of As(V). For this experiment, 1g of lignite and 5 g of bentonite, respectively,were mixed with 10 mg/L As(V) solution con-taining SO2−

4 .

2.4 Adsorption isotherms

Equilibrium isotherms were obtained frombatch adsorption experiments. The adsorbedamount of As (q) per unit absorbent mass wascalculated using the follow equation:

q =(Co − Ce) V

m(1)

where, q is the adsorbed amount (mg/g), Co

is the initial As concentration (mg/L), Ce is theconcentration of As at equilibrium (mg/L), m isthe amount of adsorbent added (g), and V is thevolume of solution (L).

To investigate the adsorption isotherms,the equilibrium data obtained in the differ-ent adsorbent-solute systems were fitted withthe linear, Langmuir and Freundlich models.The linear isotherm was calculated as follows:

q = KdCe (2)

where, Kd is the distribution coefficient (L/g).The Langmuir isotherm was fitted using the fol-lowing equation:

q =qmKLCe

1 + KLCe(3)

A linear form of this equation is rewritten asfollowing:

Ce

q=

1

qmKL+

Ce

qm(4)

where, KL (L/mg) is the Langmuir constantrelated to adsorption energy expressing howstrong the adsorbate is attached onto sur-faces, qm is the maximum adsorption capacity(mg/g). The Langmuir model assumes mono-layer adsorption onto a surface with a finitenumber of identical sites.

The Freundlich isotherm is an empiricalequation based on the adsorption onto hetero-geneous surfaces that describes the relationshipbetween the adsorbed amount (mg/g) andthese remaining in solution (mg/L) (Limousinet al, 2007).



q = KFCne (5)

KF and n are Freundlich constants that are re-lated to the adsorption capacity and adsorptionintensity, respectively. According to the Fre-undlich equation, the isotherm does not reacha plateau as Ce increases (Limousin et al, 2007).

2.5 Leaching experiments

The leaching experiments were conducted bymixing 15 g of sample and 150 mL of preparedleachants. Variable concentrations of HCl andNaOH solutions were used as leachants to yieldfinal suspension pH in the range of 2 to 13. Thesuspensions were shaken under the same con-ditions as those of the batch adsorption exper-iments. After 24 h, the suspensions were fil-tered and the filtrates were analyzed for As us-ing ICP-AES. For As concentrations less than0.1 mg/L, the filtrates were pretreated and an-alyzed using the hydride generation processcoupled with the ICP-AES. For the pretreat-ment of filtrates with low As concentrations, 30ml of the leachate sample was mixed with 15 mlof 12 M HCl, 2 ml of 20% potassium iodide (KI)solution, and 1 ml of 10% ascorbic acid solutionand then diluted with deionized water to 50 ml.

3 Results and Discussion

3.1 Characterization of the adsorbents

The chemical compositions of geological mate-rials used as adsorbent in this study are shownin Table 1. Although lignite contains 13.6 wt.%Fe, 9.37 wt.% is associated with pyrite and theremainder (4.18 wt.%) with Fe2O3 by consider-ing its mineralogical properties (Figure 2). TheAs content of lignite is 1.8 mg/kg with highsulfur content of 10.8 %. Total organic carboncontent (TOC) of lignite is 32.3 %, which meansthat it has appreciable amounts of volatile mat-ter. The As contents of bentonite, shale and ironsand are 5, 14.5, 2 mg/kg, respectively, withlow sulfur contents. The As content of shale isslightly higher than the average As concentra-tion of sedimentary rock (around 5-10 mg/kg)(Webster, 1999).

Table 2: The BET surface area and total pore vol-ume of adsorbents.

The mineralogical compositions of the ad-sorbents are shown in Figure 2. The promi-nent mineral phase observed in lignite is pyrite[FeS2] as shown in Figure 2a. However, thereare several smaller, but well defined peaks thatcorrespond to orpiment [As2S3]. The mineralsdetected in bentonite are primarily silicate min-erals such as, montmorillonite, halloysite, kaoli-nite and quartz (Figure 2b). Prominent peaksof quartz and kaolinite were also detected inshale (Figure 2c). The main minerals observedin iron sand are magnetite [Fe3O4] and albite[(Na,Ca)Al(Si,Al)3O8] (Figure 2d).

The specific surface areas and total pore vol-umes of the adsorbents are shown in Table 2.Among the adsorbents, bentonite exhibited thehighest specific surface area (126 m2/g), fol-lowed by shale (27.3 m2/g), lignite (7.53 m2/g)and then iron sand (0.905 m2/g). Total pore vol-ume of bentonite was the highest (0.175 mL/g),followed by shale (0.045 mL/g), lignite (0.016mL/g) and iron sand (0.003 mL/g).

The zeta potentials of the adsorbents as afunction of pH are shown in Figure 3. Lignitehad two PZCs at pH 2.4 and 7.0. This meansthat lignite has a positively charged surface be-tween pH 2.4 and 7.0, and has a negativelycharged surface at pH less than 2.4 and greaterthan 7.0. In contrast, iron sand had a single PZCat pH 5.5, indicating that this adsorbent has apositive charged surface at pH less than 5.5 anda negatively charged surface at pH greater than5.5 (Figure 3b). Bentonite and shale had nega-tively charged surfaces in the entire pH rangewithout PZCs (Figures 3c-d).


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Table 1: Chemical composition of adsorbents.

Figure 2: X-ray diffractograms of (a) lignite, (b) bentonite, (c) shale and (d) iron sand.



Figure 3: Zeta potential analysis of (a) lignite, (b) iron sand, (c) bentonite and (d) shale.

3.2 Effect of shaking time

Figure 4 shows the effect of shaking time onthe adsorption of As(V) and As(III) onto lignite,bentonite and shale. Apparent equilibrium wasreached after several hours for lignite whereas24 to 48 hours was required to reach equilib-rium for bentonite and shale.

3.3 Effect of pH

The effect of pH on the adsorption of As(V) andAs(III) is shown in Figure 5. From these results,As(V) adsorption was most favorable in the pHrange of 3–8 for lignite and 2–7 for bentoniteand shale. However, the amount of As ad-sorbed decreased dramatically at pH>8. In con-trast, adsorption of As(V) onto iron sand wasnegligible in the entire pH range (Figure 5a).These indicate that arsenate adsorption was fa-vored from pH 4 up to circumneutral pH andthen gradually decreases with pH.

Thermodynamics dictates that As(V) exist inan aqueous solution in the pH range of 3 to

12 as negatively charged oxyanionic species(H2AsO−

4 , HAsO2−4 , AsO3−

4 ) (Figure 6). On theother hand, the surface charge of adsorbents,which is controlled by the transfer reactions ofproton between the solution and the mineralsurface, is positive, negative or equal to zero de-pending on the pH (Kundu et al, 2004). In thepH range corresponding to the predominanceof H2AsO−

4 (pH 2–7), the surface charge of lig-nite was positive, thus the electrostatic attrac-tion between the anionic species and the posi-tively charged surface sites was promoted. Onthe other hand, at pH values greater than 7.0,the surface of charge becomes more electroneg-ative, resulting in the decrease of As adsorptiononto lignite.

The amount of adsorbed As(V) onto ben-tonite and shale decreased above neutral pHvalue (Figure 5a). Although the surface chargesof bentonite and shale were negative in the en-tire pH range, positive charges could be cre-ated due to edge defects, that is, protonation ofbroken Al-OH bonds exposed at particle edges


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Figure 4: Effect of shaking time on the adsorption of (a) As(V) and (b) As(III).

Figure 5: Effect of pH on the adsorption of (a) As(V) and (b) As(III).



Figure 6: Eh – pH predominance diagram of As at 25◦C, 1.013 bars, activity As = 10−6 and activitySO2−

4 = 10−2.

leading to anion adsorption at these sites (Linand Puls, 2000). Aluminum silicate mineralswith edge defects like kaolinite and halloysitewere found in substantial amounts in bentoniteand shale (Figures 2b-c). Consequently, the ad-sorption of As may also occur due to the brokenAl-OH bonds at particle edges of minerals (Linand Puls, 2000).

In case of As(III) adsorption, the maximumamount adsorbed was observed at pH>4 forlignite, whereas favorable adsorption tookplace at pH 8 for bentonite and at 6–9 forshale (Figure 5b). Arsenite mainly occur as anuncharged oxyanion [H3AsO3] in the pH rangeof 2–9 (Figure 6). However, this unchargedAs(III) oxyanion [H3AsO3] deprotonates to thecharged H2AsO−

3 oxyanion at pH>9. As a re-sult, greater As(III) removal was observed atlow pH value (pH 4–10) for lignite due to itspositively charged surface (Figures 3 and 5).

Consequently, the amounts of As(III) ad-sorbed onto bentonite and shale were less thanthat of As(V) because of the less favorable elec-

trostatic interactions between the positivelycharged particle edges and the unchargedH3AsO3 oxyanion. In addition, the adsorbedamount on bentonite was higher than shale,which could be attributed to the higher surfacearea of bentonite.

3.4 Effects of particle size

The effects of the particle size of adsorbents onthe adsorption of As(V) and As(III) are shownin Figure 7. It can be seen that decreasingthe particle size of the adsorbents somewhatincreased the amount of As adsorbed, whichcould be attributed to the increasing surfacearea of the particle with decreasing particle size.A bigger surface area means higher concentra-tion of active sites for adsorption on the surfaceof the adsorbents that enhance the adsorptionof As. However, the effects of the particle sizewere not so obvious. This means that ligniteand shale are the promising adsorbents used inPRBs by considering the permeability.


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Figure 7: Effects of particle size on the adsorption of (a) As(V) and (b) As(III).

3.5 Effects of sulfate concentration

The effects of concentrations of coexisting SO2−4

on the adsorption of As(V) onto lignite and ben-tonite are shown in Figure 8. Although SO2−

4

has been reported to have moderate negative ef-fects on As adsorption at concentrations greaterthan 250 mg/L onto Al and Fe-oxyhydroxides(Sun et al, 2006; Asta et al, 2009), it was foundthat SO2−

4 did not have any noticeable effecton the adsorption of As onto lignite. However,SO2−

4 appears to have a slight positive effect onAs(V) adsorption onto bentonite especially atlow adsorbent dosage (1 g/100 mL) (Figure 8b).These results indicate that the effects of SO2−

4

on As adsorption is largely dependent on majorfunctional groups or minerals available in theadsorbent. This could be attributed to inabilityof SO2−

4 ions to form any insoluble layers withthese natural adsorbents, which could prohibitthe diffusion of dissolved As species (Kundu etal, 2004).

3.6 Adsorption isotherms

The adsorption isotherms of As(V) are shownin Figure 9. Linear, Langmuir and Freundlichequations are applied to the obtained re-sults. From this figure, nonlinear adsorptionisotherms like Langmuir and Freundlich typesfitted better with the observed results than thelinear one for all adsorbents.

The linear, Langmuir and Feundlich modelparameters and their corresponding statisticalfits to the adsorption data are given in Table 3.

The values of R in the Langmuir and Freundlichtypes were higher than that of the linear type,which indicates that the adsorption of As(V)cannot be expressed by the linear adsorptionisotherm. In addition, the fitted curves showedthat lignite had the highest As(V) adsorptioncapacity at 10.9 mg As/g, followed by bentonite(0.33 mg As/g), shale (0.14 mg As/g) and ironsand (0.024 mg As/g). This means that the bestadsorbent for As(V) in this study is lignite andthe order of adsorption capacity is as follows:lignite > bentonite > shale > iron sand. The ad-sorption capacity (qm) results mentioned abovecorrespond to the pH interval of 2.8–3 for lig-nite, 6.5–7 for bentonite, 7.5–7.8 for shale and6.7–6.8 for iron sand.

For As(III) adsorption, the amounts of ad-sorbed onto all adsorbents were lower thanthose of As(V) (Figure 10). According to Rvalues in Table 3, nonlinear type adsorptionisotherms fitted better than the linear isothermfor lignite, bentonite and shale whereas the lin-ear type is more applicable to iron sand. Lig-nite had the highest adsorption capacity, fol-lowed by bentonite and shale similar to thoseobserved in the adsorption of As (V). The max-imum As(III) adsorption capacity (qm) shownin Table 3 is limited to the pH interval of 2.5–2.7for lignite, 6.2–6.7 for bentonite and 7.5–7.8 forshale. The better fit of the Langmuir isothermto the experimental data indicates monolayeradsorption of As(V) and As(III) onto the adsor-bents except As(III) adsorption onto iron sand.Non-linear isotherms also suggest progressive



Figure 8: Effects of SO2−4 on the adsorption of As(V) onto (a) lignite and (b) bentonite.

Figure 9: Adsorption isotherms of As(V) onto (a) lignite, (b) bentonite, (c) shale and (d) iron sand.


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Table 3: Fitted parameters and correlation coefficients of linear, Langmuir and Freundlich isothermscalculated based on the least squares method.

Figure 11: Leaching behaviors As from lignite,bentonite, shale and iron sand with pH.

saturation of the adsorbent, leading to the pro-portional decrease of adsorption with increas-ing As concentration (Limousin et al, 2007).

3.7 Leaching experiments

The leaching behavior with pH of geogenic Ascontained in all adsorbents is shown in Figure11. The results show that the leaching of ge-ogenic As was strongly pH dependent. En-hanced release of geogenic As occurred underboth acidic and alkaline conditions while a min-imum was observed around the circumneutralpH range.

The release of As was minimal in the pH

range of 6 to 10 for lignite, shale and bentoniteand at pH 3 to 7 for iron sand. On the otherhand, at pH values above 10 and below 6, theconcentration of As in the leachate significantlyincreased. The maximum amount of As re-leased from lignite was observed at pH 2 and 13at 0.08 and 0.09 mg/L, respectively. For shale,the highest concentration of As was observedat pH 1 and 11 amounting to 0.03 mg/L. Simi-larly, the amount of As released from bentonitewas highest at very acidic pH (pH = 1). For ironsand, As concentration increased at pH valuesbelow 3 and above 7 reaching a maximum at pH1 and 12 (0.05 mg/L). The experimental resultsindicate that the effectiveness of these naturalmaterials as adsorbents is greatly limited by thepH of the contaminated system.

4 Conclusion

The preliminary study was conducted to de-termine the As adsorption properties of sev-eral naturally occurring adsorbents. Lignitewas the most effective adsorbent for As(V)and followed by bentonite, shale and then ironsand. On the other hand, the amounts ofAs(III) adsorbed onto all adsorbents were lowerthan those of As(V). This indicates that As(III)is more difficult to immobilize through ad-sorption in comparison to As(V). The adsorp-



Figure 10: Adsorption isotherms of As(III) onto (a) lignite, (b) bentonite (c) shale and (d) iron sand.


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tion equilibrium data fitted well with nonlinearmodels for all adsorbent regardless of the chem-ical forms of As. The amount of adsorptionwas pH-dependent, and the maximum amountof adsorption was observed at around circum-neutral pH regardless of the adsorbent and thechemical forms of As. The zeta potential mea-surements helped to explain the changes in theamounts of adsorbed As based on the electro-static attraction/repulsion of As species and thesurface of the adsorbents. The effects of par-ticle sizes of adsorbents were not so obvious.Although SO2−

4 did not have any noticeable ef-fect on the adsorption of As onto lignite, it ap-peared to have a slight positive effect on As(V)adsorption onto bentonite. From the resultsof batch leaching tests, acidic (pH<6) and al-kaline (pH>10) conditions destabilized the ge-ogenic As of the adsorbents, indicating that theeffectiveness of these natural materials as ad-sorbents is greatly limited by the pH of the con-taminated system.

References

Asta, M.P., J. Cama, M. Martnez, J. Gimnez (2009)Arsenic removal by goethite and jarosite in acidicconditions and itsenvironmental implications.Journal of Hazardous Materials 171: 965–972.

Bang, S., G.P. Korfiatis and X. Meng (2005) Removalof arsenic from water by zero-valent iron. J. Haz-ard. Mater. 121: 61–67.

Baskan, M.B. and A. Pala (2011) Removal of arsenicfrom drinking water using modified natural zeo-lite. Desalination 281: 396–403.

Chatterjee, A., D. Das, and B.K. Mandal (1995) Ar-senic contamination in groundwater in six dis-tricts of West Bengal, India: The biggest arseniccalamity in the world. Part I. Arsenic species indrinking water and urine of affected people. An-alyst 120: 643–650.

Chen, S.L., S.R. Dzeng, M.H. Yang, K.H. Chiu, G.G.Shieh, C.M. Wai (1994) Arsenic species in ground-waters of the blackfoot disease area, Taiwan. En-viron. Sci. Technol. 33: 877–881.

Das, D., A. Chatterjee, and B.K. Madal (1994) Ar-senic contamination in groundwater in six dis-tricts of West Bengal, India: The biggest arseniccalamity in the world. Analyst 119: 168–170.

Elizalde-Gonzalez, M.P., J. Mattusch, W.D. Einickeand R. Wennrich (2001) Sorption on natural solidsfor arsenic removal. Chem. Eng. J. 81: 187–195.

Ferguson, F. and J.A. Gavis (1972) Review of the Ar-senic Cycle in Nature Waters. Water Res. 6: 1259–1274.

Fujinoa, Y., X. Guoc, J. Liuc, L. Youc, M. Miyataked,and T. Yoshimura, and Japan Inner MongoliaArsenic Pollution (JIAMP) (2004) Study groupmental health burden amongst inhabitants of anarsenic-affected area in Inner Mongolia, China.Social Science & Medicine 59: 1969–1973.

Hricko, A. (1994) Environmental problems behindthe Great Wall. Environ. Health Presp. 102: 154–159.

Karim, M.M. (2000) Arsenic in groundwater andhealth problems in Bangladesh. Water Res. 34:304–310.

Kundu, S., S.S. Kavalakatt, A. Pal, S.K. Ghosh, M.Mandal, T. Pal (2004) Removal of arsenic usinghardened paste of Portland cement: batch adsorp-tion and column study. Water Research 38: 3780–3790.

Lackovic, J.A., N.P. Nikolaidis and G.M. Dobbs(2000) Inorganic arsenic removal by zero valentiron. Environ. Eng. Sci. 17: 29–39.

Lai, V.W.W., Y. Sun, Y, E. Ting, W.R. Cullen and K.J.Reimer, K J (2004) Arsenic speciation in humanurine: are we all the same? Toxicology and Ap-plied Pharmacology 198: 297–306.

LAPI-ITB (2009) Review Laporan tentang konsen-trasi arsenik terlarut dalam air tanah di bagianhilir daerah timbunan batuan penutup (wasterock dump). Laporan Akhir.

Limousin, G., J.P. Gaudet, L. Charlet, S. Szenknect,V. Barthes and M. Krimissa (2007) Sorptionisotherms: A review on physical bases, mod-eling and measurement. Applied Geochemistry22: 249–275.

Lin, Z. and R. W. Puls (2000) Adsorption, desorp-tion and oxidation of arsenic affected by clay min-erals and aging process. Environmental Geology39: 753–759.

Niu, S., S. Cao, and E. Shen (1995) The status ofarsenic poisoning in China. SEGH Second In-ternational Conferrence on Arsenic Exposure andHealth Effects, San Diego, CA.

Pena, M.E., G.P. Korfiatis, M. Patel, L. Lippincottand X. Meng (2005) Adsorption of As(V) andAs(III) by nanocrystalline titanium dioxide. Wa-ter Research 39: 2327–2337.

Smedley, P.L., D.G. Kinniburgh (2002) A review ofthe source, behavior and distribution of arsenic innatural waters. J. Appl. Geochem. 17: 517–568.

Shevade, S. and R.G. Ford (2004) Use of synthetic ze-olites for arsenate removal from pollutant water.Water Research 38: 3197–3204.

Subramanian, K.S., T. Viraraghavan, T. Phom-mavong and S. Tanjore (1997) Manganese green-



sand for removal of arsenic in drinking water.Wat. Qual. Res. J. Can. 32: 551–561.

Sun, H., L. Wang, R. Zhang, J. Sui, G. Xu (2006)Treatment of groundwater polluted by arseniccompounds by zero valent iron. Journal of Haz-ardous Materials 129: 297–303.

USEPA (2002) Implementation Guidance for the Ar-senic Rule, EPA 816-D-02-005, The U.S. Environ-mental Protection Agency, Washington, DC.

Webster, G. (1999) in: C.P. Marshall, R.W. Fairbridge(Eds) Arsenic. Encyclopedia of Geochemistry,Chapman Hall, London, pp. 21–22. WHO (1993)Guidelines for Drinking Water Quality. WorldHealth Organization, Geneva.

Yoshitake, H., T. Yokoi and T. Tatsumi (2003) Ad-

sorption behavior of arsenate at transition metal

cations captured by amino-functionalized meso-

porous silicas. Chem. Mater. 15: 1713–1721.


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