Science of the Total Environment 647 (2019) 47–56

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Enhanced hexavalent chromium removal by activated carbon modifiedwith micro-sized goethite using a facile impregnation method

Mei Su a, Yili Fang a,b,c, Bing Li a,b,c, Weizhao Yin d,⁎, Jingjing Gu e, Hao Liang c, Ping Li a,b,c, Jinhua Wu a,b,c,⁎⁎a The Key Lab of Pollution Control and Ecosystem Restoration in Industry Clusters, Ministry of Education, South China University of Technology, Guangzhou 510006, PR Chinab The Key Laboratory of Environmental Protection and Eco-Remediation of Guangdong Regular Higher Education Institutions, South China University of Technology, Guangzhou 510006, PR Chinac School of Environment and Energy, South China University of Technology, Guangzhou 510006, PR Chinad School of Environment, Jinan University, Guangzhou 510632, PR Chinae Water Purification Institute of Logistics Department of Guangzhou Military Region, Guangzhou 510500, PR China

H I G H L I G H T S G R A P H I C A L A B S T R A C T

• Micro-sized FeOOH was homogeneouslydistributed on activated carbon.

• Cr(VI) was removed via adsorption, re-duction and co-precipitationmechanism.

• The mFeOOH@AC composite has goodstability and reusability.

⁎ Correspondence to: W, Yin, School of Environment, Ji⁎⁎ Correspondence to: J. Wu, School of Environment and

E-mail addresses: weizhaoyin@jnu.edu.cn (W. Yin), jin

https://doi.org/10.1016/j.scitotenv.2018.07.3720048-9697/© 2018 Elsevier B.V. All rights reserved.

a b s t r a c t

a r t i c l e i n f o

Article history:Received 18 April 2018Received in revised form 2 July 2018Accepted 26 July 2018Available online 27 July 2018

Editor: Baoliang Chen

In this study, activated carbon (AC) was modified with micro-sized geothite (mFeOOH) using a facile and cost-effective impregnation method for enhanced Cr(VI) removal from aqueous solutions. X-ray diffraction (XRD)and scanning electron microscope (SEM) analysis showed that FeOOH particles with a diameter of 0.1–1 μmwere dispersed homogeneously on the surfaces and pores of the AC. Fourier transform infrared spectrum(FTIR) and X-ray photoelectron spectra (XPS) analysis indicated that Cr(VI) was easily adsorbed onto themFeOOH and reduced to Cr(III) by the AC, eventually deposited as Cr(III)-Fe(III) hydroxides (e.g., (CrxFe1−x)(OH)3). Hence, the mFeOOH@AC achieved a significantly higher Cr(VI) removal efficiency of 90.4%, 4.5 times ofthat the AC. The adsorption of Cr(VI) onto the mFeOOH@AC agreed well with the Langmuir adsorption model,demonstrating that the adsorption process was controlled by monolayer adsorption. This adsorption processalso followed the pseudo second-order kinetics and the adsorption rate constant K2 was determined to be0.013 g/mg·min. The Cr(VI) removal efficiency decreased with pH values as the adsorption process was highlypH-dependent. After the desorption-adsorption process by 0.1MHCl solution for 4 cycles, the removal efficiencyof Cr(VI) was still kept up to 75.1%, indicating that themFeOOH@AC has a good stability and can be easily regen-erated. In addition, themFeOOH@AC also exhibited a promising potential for reutilization since a Cr(VI) removal

Keywords:Iron-based materialHexavalent chromiumFacile modificationActivated carbon

nan University, Guangzhou 510632, PR China.Energy, South China University of Technology, Guangzhou 510006, PR China.huawu@scut.edu.cn (J. Wu).

48 M. Su et al. / Science of the Total Environment 647 (2019) 47–56

efficiency of 85.4%was achieved after stripping all themFeOOHand Cr(III)-Fe(III) hydroxides by 1MHCl solutionand regeneration with mFeOOH. We demonstrate that the modified AC with micro-sized goethite can remark-ably enhance its ability for Cr(VI) removal in water treatment.

© 2018 Elsevier B.V. All rights reserved.

1. Introduction

Chromium is widely used in various industries such as steelmanufacturing, leather tanning, metal finishing, wood preserving, andtextile productions (Jayakumar et al., 2014; Wu et al., 2012; Yoonet al., 2011). During the process of production, due to the leakage of in-dustrial effluents and improper practices of industrial sludge disposal, alarge number of chromium components have been released into thesurface and groundwater. In aqueous solutions, chromium usually ex-ists in the oxidative states of hexavalent chromium (Cr(VI)) and triva-lent chromium (Cr(III)) (Shi et al., 2011b). Compared to relativelyinnocuous Cr(III), Cr(VI) ismore soluble, mobile, toxic, and carcinogenic(Chai et al., 2017; Yirsaw et al., 2016; Zhuang et al., 2014). It is henceeasy to spread throughout the water body, contaminating the groundwater and endangering the human health (Min et al., 2017). Whenthe concentration of Cr(VI) reaches 0.1 mg/g body weight, Cr(VI) be-comes lethal to human (Richard and Bourg, 1991). Hence, Cr(VI) hasbeen considered as a priority pollutant by the United States Environ-mental Protection Agency (He et al., 2013) and it shall be removed ordecontaminated from the environment.

A variety of technologies have been developed to remove Cr(VI)from aqueous solutions, including chemical reduction (Vlyssides andIsrailides, 1997), membrane filtration (Fabiani et al., 1997), solvent ex-traction (Macchi et al., 1991), biological processes (Kapoor et al.,1999) and ion exchange (Tiravanti et al., 1997). However, most ofthese methods have technical or economic limits, such as high opera-tional cost, incomplete removal, and toxic metal sludge generation.Compared to these methods, adsorption is a promising and efficienttechnology to remove Cr(VI) from aqueous solutions (Barakat, 2011;Sahu et al., 2009). By now, activated carbon (AC) is one of the mostcost-effective adsorbents for various pollutants because of its abundantmicropores, huge specific surface areas and functional groups such ashydroxyl and carboxyl (Park and Jang, 2002). Moreover, ACmay reduceCr(VI) to Cr(III) since Cr(VI) is a strong oxidant (Espinoza-Quiñoneset al., 2010; Módenes et al., 2010), and the reduction of Cr(VI) benefitsthe immobilization of Cr via Cr(III) precipitation (Gheju and Balcu,2011). However, the functional groups of AC are negatively chargedand relatively nonpolar (Adhoum and Monser, 2002; Yin et al., 2007),resulting in a weak adsorption capacity for Cr(VI), generally below10 mg/g (Monser and Greenway, 1996). This limits the application ofAC for the removal of Cr(VI).

Recently, several studies have been proposed to increase Cr(VI) re-moval efficiency bymodifying AC with acids, oxidants andmetal oxides(Adhoum and Monser, 2004; Zhao et al., 2005). In particular, AC modi-fiedwith iron or iron oxides showed improved removal capacities for Cr(VI). Iron oxides are widely distributed in the environment and havebeen applied to remove heavy metal ions from aqueous solutions dueto their low toxicity, fast kinetics and strong affinity toward heavymetals (Zhou et al., 2015). However, iron and iron oxides are prone toaggregation and subsequent rapid sedimentation in aqueous solutionsbecause of their high surface energy (Xu et al., 2012), leading to thedecrease of their reactivity in the adsorption of heavy metal ions.Besides, powdery iron oxides are difficult to be separated from aqueoussolutions after their usage. It is hence needed to enhance the dispersityof iron oxide particles by supporting them onto physical materials suchas bentonite,multiwalled carbon nanotube, zeolite, smectite and pillaredclay (Lv et al., 2013; Ruan et al., 2015; Shi et al., 2011b). For instance,Ruan et al. (2015) prepared a bentonite-biochar-Fe2O3 composite by fac-ile pyrolysis that significantly enhanced the adsorption capacity of Fe2O3

for Cr(VI). Wu et al. (2012) used montmorillonite to support Fe0, whichgreatly reduced the aggregation of Fe0 and improved its adsorption andreduction of Cr(VI).

In the present study, we chose AC as a supporting material for an-other promising iron oxide, goethite (FeOOH). It is one of themost com-mon iron oxides which widely spread in water, soil, sediment and rockin the environment (Guo et al., 2016). Compared to iron or other ironoxides (e.g., Fe0, Fe3O4, and Fe2O3), FeOOH is more stable under aerobicconditions and can be easily synthesized by chemical processes. Besides,FeOOH generally exhibits strong affinity to oxyanions such as Cr(VI)(Qin et al., 2014) via ion exchange and complexation owing to theexisting of hydroxyl and ironoxy groups on its surface. However, it hasbeen found that FeOOH with a size less than 5 μm will be adverse toits practical application such as separation, recovery and reuse owingto its easily aggregation (Zuo et al., 2016). It is hence necessary to sup-port FeOOH onto physical materials in order to prevent it from aggre-gating. Among various physical materials, carbonaceous materialssuch as AC, biochar, graphene and graphene oxide sheets have beenwidely selected as supporters for metal oxides (Cong et al., 2012; Guoet al., 2016; Koduru et al., 2016; Xu et al., 2013). Here we chose AC asaneffective supportingmatrix for FeOOHbecause ferric ion canbe easilyadsorbed onto AC by electrostatic adsorption.We propose this modifiedFeOOH@AC approach to investigate the dispersity of FeOOH and the en-hancement of Cr(VI) removal. In this study, we will focus primarily onthe performance of the FeOOH@AC for the Cr(VI) removal capacity.Solid phase characterizationwas performed to illustrate themechanismof the Cr(VI) removal by the FeOOH@AC. Fourier transform infraredspectrum (FTIR) and X-ray photoelectron spectroscopy (XPS) wereemployed to measure the reduction of Cr(VI) and the oxidation of ACduring the adsorption process. Desorption-adsorption experimentswere also carried out to evaluate the stability and reutilization of theFeOOH@AC.

2. Materials and methods

2.1. Materials

The activated carbon (AC) was received (Chaoyang Senyuan Acti-vated Carbon Limited Company) with an average diameter of 0.1 mmand a surface area of 770 m2/g, respectively. Chromium stock solution(1000 mg/L) was prepared by dissolving 2.829 g of potassium dichro-mate (K2Cr2O7) into 1000 mL of deionized water. The required concen-trations of the chromium standard solution were made by diluting thestock solution. Nine hydrated ferric nitrate (Fe(NO3)3·9H2O) and otherchemicalswere analytical grade (GuangzhouChemical Reagent Factory).

2.2. Adsorbents preparation

Before use, the AC was pretreated with 0.1 M HCl for 24 h and ultra-sonic treatment for 15 min and rinsed with deionized water until it be-came neutral. It was then dried at 103 °C for 2 h and subsequentlypacked into zip lock bags and stored in a desiccator. The unsupportedmFeOOH was used as a control and was synthesized using a standardmethod (Schwertmann and Cornell, 2000). In brief, 100mL of 1M ferricnitrate solution was prepared in a polyethylene bottle and 180 mL of5 M KOHwas quickly poured into the polyethylene bottle. The solutionwas mixed at a stirring rate of 130 rpm. Immediately, the suspensionwas diluted to 2 L with deionized water and then kept in a sealed

Fig. 1. The XRD patterns of the AC, mFeOOH and mFeOOH@AC.

49M. Su et al. / Science of the Total Environment 647 (2019) 47–56

polyethylene bottle at 70 °C for 60 h. Finally, the suspensionwasfilteredand dried at 70 °C for 12 h. The FeOOH was left as residue.

The mFeOOH@AC was prepared with a facile impregnation method(Wu et al., 2013). An amount of 4.48 g pretreated AC was added into100 mL 0.1 M ferric nitrate solution. This mixture was stirred for 24 hwith a magnetic stirrer at room temperature (30 °C), followed by im-pregnating through drying at 70 °C for 12 h. Then the residuewas rinsedwith deionizedwater and centrifuged several times until it becameneu-tral. Eventually, the mFeOOH@AC composite was obtained after dryingthe solid at 70 °C for 12 h in an oven. The amount of FeOOH loaded onthe mFeOOH@AC was determined to be 2.3% by stripping all themFeOOH on the surfaces of the ACwith 1MHCl andmeasuring the fer-ric ion concentration in the eluent.

2.3. Experimental procedures

Batch experiments were carried out in 150 mL conical flask reactorswhich contained 100mLCr(VI) solution (50mg/L, pH=5.6). The initialpH value of the Cr(VI) solution was adjusted by titrating with a 0.1 MH2SO4 or NaOH solution. After the addition of mFeOOH@AC (4 g/L),the flask was sealed with a rubber plug and then shaken at a speed of120 rpm at room temperature (30 °C). At regular intervals, 2 mL mix-ture was withdrawn with a syringe and then filtered with a 0.45 μmmembrane, followed by an immediate analysis of Cr(VI). After the reac-tion, the pH value and the iron leaching concentrations of the solutionwere measured. The effects of initial pH and Cr(VI) concentrationswere evaluated on the basis of the above benchmark experiments.

Stability and reusability were conducted with HCl solutions as elu-ents. Initially, 0.4 gmFeOOH@ACwas added into 100mL Cr(VI) solutionof 50 mg/L to reach adsorption equilibrium. Then the resultingmFeOOH@AC was filtered and desorbed by adding 100 mL 0.1 M HClin a conical flask reactor at a shaking rate of 120 rpm for 4 h, followedby an immediate analysis of Cr(VI) and ferric ion concentrations in thesolution. This desorbed mFeOOH@AC was reused for Cr(VI) adsorptionat an initial concentration of 50 mg/L. After 4 cycles of desorption-adsorption process, its reusability was measured by adding 1 M HCl tostrip all the mFeOOH on the surfaces of the AC and remodifying theAC with mFeOOH following the above impregnated method. Prior tonext cycle, the mFeOOH@AC was washed repeatedly with deionizedwater to remove the adhered solution. Each experiment was repeatedthree times and an average value was used.

2.4. Analytical methods

The pH value was measured using a pH meter (PHS-3C, Sanxin,China). The concentration of ferric ion in the solution was determinedby the o-phenanthroline colorimetric method. The concentration of Cr(VI) was measured by the 1,5-diphenylcarbazide colorimetric methodwith an UV–Vis spectrophotometer (Shimazu 2450) at a wavelengthof 540 nm. The pHpzc of the mFeOOH@AC was obtained using the pHdrift method (Jr and Malik, 2002).

The Cr(VI) removal efficiency (R, %), adsorption capacity at the reac-tion time (qt, mg/g) and the equilibrium time (qe,mg/g)were calculatedaccording to Eqs. (1), (2), and (3), respectively.

R ¼ C0−Ctð Þ=Ct � 100% ð1Þ

qt ¼ C0−Ctð Þ � V=m ð2Þ

qe ¼ C0−Ceð Þ � V=m ð3Þ

where C0 (mg/L) is the initial concentration of Cr(VI), Ct (mg/L) and Ce(mg/L) are the concentrations of Cr(VI) at the reaction time (t) andthe equilibrium time, respectively, V (L) is the total volume of the solu-tion, and m (g) is the dosage of the adsorbent.

The leaching rate (L, %) of iron from the mFeOOH@AC and the de-sorption efficiency (D, %) of Cr(VI) was calculated according toEqs. (4), (5) respectively.

L ¼ the amount of leached ironð Þ= the amount of iron loaded onto the ACð Þ � 100%

ð4Þ

D ¼ the amount of desorbed Cr VIð Þð Þ= the amount of adsorbed Cr VIð Þð Þ � 100%

ð5Þ

The X-ray diffraction (XRD) patterns of the AC, mFeOOH andmFeOOH@ACweremeasuredwith a D8 ADVANCE X-ray diffractometer(Bruker, Germany). During the XRD analysis, the samples were scannedfrom 10 to 70° at a speed of 4°/min with Cu-Kα radiation at 40 kV and40 mA. The microscopic features of these materials before and after Cr(VI) adsorption were analyzed with a scanning electron microscope(SEM, Hitachi S-4500) equipped with an energy-dispersive X-ray(EDS) for elemental analysis. The Fourier transform infrared spectrum(FTIR) of the AC before and after Cr(VI) adsorption was measuredusing a Nicolet 6700 FTIR spectrometer (Thermo Fisher Scientific,USA). The X-ray photoelectron spectra (XPS) of the mFeOOH and themFeOOH@AC after Cr(VI) adsorption was measured with a photoelec-tron spectrometer (K-Alpha, Thermo Scientific, USA) using Al Kα radia-tion. The Brunauer emmett teller (BET) surface areas of the adsorbentswere obtained by N2 adsorption at 77 K with an ASAP 2020 surfacearea analyzer (Micromeritics, USA).

3. Results and discussion

3.1. Solid phase characterization

TheXRDpatterns (Fig. 1)were used to identify the crystalline phasesformed upon the modified AC and indicate the presence of FeOOH. Thestrong peak of the AC at 2θ=25°was attributed to C(200) plane, whilethe weak one at 2θ = 44° was C(101) plane (Li et al., 2012). ThemFeOOH synthesized in this study was considered to be goethite be-cause the diffraction peaks observed around 21°, 33°, 36°, 41°, 48°, 61°and 63° agreed well with the phase of goethite (PDF No. 29-0713), cor-responding to (110), (130), (111), (140), (041), (002) and (061) planes,respectively, similar to the results reported by Wang et al. (2015). The(061) plane was not observed in the mFeOOH@AC, but there was anew diffraction peak around 59° that corresponds to FeOOH (151)plane (PDF No. 29-0713), indicating the formation of FeOOH on theAC (mFeOOH@AC).

The surface morphology and composition of the mFeOOH@AC be-fore and after Cr(VI) adsorption were examined by SEM and EDS. As

Table 1The elemental composition and surface area of theAC and themFeOOH@AC before and af-ter Cr(VI) adsorption.

Sample Elementalcomposition (wt%)

BET surfacearea (m2/g)

Pore volume(cm3/g)

Pore size(nm)

C O Fe Cr

AC 72.8 27.2 NA NA 777 0.49 2.5mFeOOH NA 63.6 36.4 NA 29.6 0.07 8.9mFeOOH@AC 65.5 31.4 3.1 NA 714 0.39 2.3mFeOOH@AC⁎ 62.1 33.8 2.9 1.2 705 0.36 2.1

mFeOOH@AC⁎: mFeOOH@AC after Cr(VI) adsorption; NA: not detected.

50 M. Su et al. / Science of the Total Environment 647 (2019) 47–56

shown in Fig. 2, the surface of the fresh ACwas glossy, while some smalland bright aggregates with a diameter of 0.1–1 μm were distributedhomogenously on the surfaces and pores of the mFeOOH@AC andthese aggregates were primarily attributed to FeOOH. Elemental analy-sis (Table 1) showed the existence of Fe with a content of 3.1% in themFeOOH@AC while there was no Fe on the fresh AC, further demon-strating the formation of well dispersed FeOOH on the surfaces of theAC. After the adsorption of Cr(VI), the surfaces of the adsorbedmFeOOH@AC were coarse and many near-spherical particles with a di-ameter of 0.5–1.5 μm were formed on the pores of the AC. Those parti-clesweremost likely to be Fe(III)-Cr(III) co-precipitation such as (CrxFe1−x)(OH)3 (Mondragon et al., 2012; Shi et al., 2011a), since Cr(III) re-duced from Cr(VI) (Espinoza-Quiñones et al., 2010) tended to beembed in the iron corrosion products to form amixed Fe(III)-Cr(III) hy-droxide film due to the similar ionic radius of Cr(III) (0.063 nm) and Fe(III) (0.064 nm) (Gheju and Balcu, 2011; Lo et al., 2006). Besides, chro-mium with a content of 1.2% was determined on the surfaces of thenear-spherical particles of the mFeOOH@AC, suggesting the adsorptionof Cr(VI) onto themFeOOH@AC and the production of Fe(III)-Cr(III) hy-droxides (Mondragon et al., 2012; Shi et al., 2011a).

As expected (Table 1), the specific surface area, pore volume andpore size of the AC slightly decreased from 777 to 714 m2/g, 0.49 to0.39 cm3/g and 2.5 to 2.3 nm, respectively when mFeOOH was loadedonto the AC, which covered the surfaces and blocked the pores of theAC. The specific surface area, pore volume and pore size of themFeOOH@AC after Cr(VI) adsorption further decreased to 705 m2/g,0.36 cm3/g and 2.1 nm, attributed to the increasing concentration ofchromium on the surface of the mFeOOH@AC and the aggregation of(CrxFe1−x)(OH)3 particles in the pores of the AC.

3.2. Comparison experiments

The adsorption experiments were conducted with the AC, mFeOOHand mFeOOH@AC, respectively. As shown in Fig. 3, the removal effi-ciency of Cr(VI) in the AC systemwas only 20.0% after 480min reactiontime, indicating that the AC had weak interactions with Cr(VI), largelydue to the weakly negative charged and relatively nonpolar surfaces

Fig. 2. The SEM images of (a) AC; (b) mFeOOH; (c) mFeOOH@AC and (d) mFeOOH@AC afterconcentration: 50 mg/L, initial pH: 5.6).

of the AC, leading to electrostatic repulsion between the AC and theelectronegative Cr(VI). In the mFeOOH system, only 13.9% of Cr(VI)was removed due to a decrease of active sites by the easy aggregationof mFeOOH in the solution. However, in the mFeOOH@AC system, theCr(VI) removal efficiency was determined to be 90.4%, about 4.5 and6.5 times of that using AC and mFeOOH alone, respectively. The en-hanced performance of the mFeOOH@AC can be attributed to the in-crease of stabilization, dispersion, and separation properties of themFeOOH particles by supporting them onto the AC. The enhanced Cr(VI) removal in the mFeOOH@AC is also consistent with the results ofSEM which show that the mFeOOH particles were well dispersed onthe AC, leading to the full adsorption and co-precipitation of the Cr(VI) anion onto the surfaces of the mFeOOH@AC (Bradl, 2004). Besides,the mFeOOH@AC could be easily separated from the solution with ascreen mesh after its usage.

Kinetic analysis was carried out to describe the adsorption process ofCr(VI) onto the AC and the mFeOOH@AC. Two classic kinetic models(pseudo first-order model and pseudo second-order model) wereused to fit the data and the results are shown in Table 2. The correlationcoefficients (R2) from the pseudo second-order kinetic model werehigher than those from the pseudo first-order one for these adsorbents,implying that the pseudo second-order model fitted the kinetic databetter than the pseudo first-order for these adsorbents. It is concludedthat chemical adsorption controls the Cr(VI) adsorption onto the ACand the mFeOOH@AC. Our results are consistent with several other

Cr(VI) adsorption. (AC and mFeOOH@AC dosage: 4 g/L, FeOOH dosage: 0.09 g/L, Cr(VI)

Fig. 3. Different adsorbents for Cr(VI) removal. (AC and mFeOOH@AC dosage: 4 g/L,mFeOOH dosage: 0.09 g/L, Cr(VI) concentration: 50 mg/L, initial pH: 5.6, temperature:30 °C).

Fig. 4. The FTIR of the AC before and after Cr(VI) adsorption. (AC dosage: 4 g/L, Cr(VI)concentration: 50 mg/L, initial pH: 5.6).

51M. Su et al. / Science of the Total Environment 647 (2019) 47–56

studies (Adhoum and Monser, 2004; Al-Othman et al., 2012; Wu et al.,2012), showing that Cr(VI) adsorption onto the mFeOOH@AC followedchemisorption mechanism (e.g., ion exchange).

As shown by Espinoza-Quiñones et al. (2010), Cr(VI)was reduced toCr(III) by AC during the adsorption process with XPS analysis. Here wespeculated that Cr(VI) may oxidize AC to form new oxygen-containingfunctional groups (e.g., carboxyl and/or hydroxyl groups) and productscontaining Cr(III) species (Yue et al., 2009). This newly generated Cr(III)tends to be precipitated as Cr(OH)3 due to its low solubility (Shen et al.,2016). Consequently, the reduction of Cr(VI) to Cr(III) by the AC waslikely one of the main reasons for the enhanced Cr(VI) removal by themFeOOH@AC.We hence performed FTIR andXPS analysis to investigatethe reduction of Cr(VI) by the AC. Before analysis, the binding energy ofgraphitic carbon C 1 s peak was calibrated at 284.50 eV as an internalreference for charging of the samples. As shown in Fig. 4, there was asignificant decrease at the peak of C-OH (3420 cm−1)while a significantincrease at the peakof C\\O (1610 cm−1) after Cr(VI) adsorption,whichwas attributed to the oxidation of carbon (`C\\OH→_C_O) becauseof the reduction of Cr(VI) to Cr(III) by the AC (Yue et al., 2009). The XPSanalysis (Fig. 5) showed the existence of Cr(III) on the surfaces of themFeOOH@AC after Cr(VI) adsorption with the binding energies at586.5 and 577.2 eV (Yang et al., 2017), while no Cr(III) (only Cr(VI))was detected on the mFeOOH after Cr(VI) adsorption. Based on theabove analysis, we proposed a possible model to describe the mecha-nism of Cr(VI) removal by themFeOOH@AC (Scheme 1). After the addi-tion of mFeOOH@AC, Cr(VI) was instantaneously adsorbed onto thesurfaces of the mFeOOH@AC by electrostatic attraction, ion exchangeand complexation because of the positively charged surfaces, hydroxyland ironoxy groups of the mFeOOH. As mentioned above, the surface-bounded Cr(VI) could be reduced to Cr(III) by the AC directly(Espinoza-Quiñones et al., 2010) or via themFeOOHas electron shuttles(Wang et al., 2017). Since the mFeOOH has more negative conductionband position (0.58 V) than the reduction potential (1.33 V) of Cr2O7

2

−/Cr3+, the mFeOOH acting as a semiconductor could mediate electrontransfer from the AC to Cr(VI) and reduce the surface-bound Cr(VI) intoCr(III) (Sun et al., 2016). Subsequently, the resultant Cr(III) was partially

Table 2The pseudo first and second order kinetics parameters for the adsorption of Cr(VI).

Adsorbent qe, exp(mg/g)

Pseudo first-order kinetics

K1

(min−1)qe(mg/g)

AC 2.50 4.34 2.24mFeOOH@AC 11.3 2.31 10.8

precipitated as chromium hydroxides (Shen et al., 2016) or embed inthe AC in the form of Cr(III)-Fe(III) hydroxides such as (CrxFe1-x)(OH)3for their similar ionic radius and identical charge (Gheju and Balcu,2011; Lo et al., 2006). Eventually, Cr(VI) was firmly immobilized onthe mFeOOH@AC for it was favored by the growth and accumulationof spherical phase particles as a form of chromium hydroxides and Cr(III)-Fe(III) hydroxides (Lo et al., 2006).

In addition, as shown in Fig. 3, the rapid adsorption of Cr(VI) ontothe AC, mFeOOH and mFeOOH@AC was observed at the first 30 minand then the adsorption proceeded slowly, eventually reached the equi-librium at 480 min. This could be attributed to the more abundance ofadsorption sites at the initial stage than the later stage (Wu et al.,2009), leading to the more affordable adsorption sites for the Cr(VI)ions at the first 30 min than the later stage. In addition, the diffusionof Cr(VI) into the interior of the adsorbents became slower with timeon account of the increase of diffusion resistance and the decrease of ad-sorption driving force due to the decrease of Cr(VI) concentration in so-lution. Here we chose 480 min as the equilibrium time.

3.3. Effect of initial pH

The influence of initial pH on the Cr(VI) adsorption onto the AC andthe mFeOOH@AC was investigated by increasing the initial pH valuesfrom 3 to 11. Fig. 6 shows that themaximumequilibrium adsorption ca-pacity of Cr(VI) onto the AC and themFeOOH@ACwas reached at initialpH 3, with a value of about 4.5 and 12.4 mg/g, respectively. It was thendecreased to 0.8 and 2.1 mg/g, respectively for the adsorption onto theAC and the mFeOOH@AC with an initial pH of 11. It is obvious that theadsorption of Cr(VI) is a highly pH-dependent process, especially inthe mFeOOH@AC system. Depending on the pH, Cr(VI) exists in severalforms (e.g., H2CrO4, HCrO4

−, CrO42−,Cr2O7

2−) in aqueous solutions.H2CrO4 exists at a pH value of less than 1, while HCrO4

− predominatesat pH between 1 and 6.5 and CrO4

2− is most favorable at pH 8. Whenthe concentration of Cr(VI) reaches up to 1 g/L, Cr2O7

2− will be formed(Al-Othman et al., 2012). The surface adsorption and reduction of Cr(VI) is higher with the increasing number of H+ ions on the adsorbent

Pseudo second-order kinetics

R2 K2

(g/mg·min)qe(mg/g)

R2

0.880 0.023 2.55 0.9960.902 0.013 11.4 0.999

Fig. 5. The XPS survey of (a) mFeOOH, (b) mFeOOH@AC, high-resolution XPS survey of Cr 2p of (c) mFeOOH and (d) mFeOOH@AC after Cr(VI) adsorption. (mFeOOH@AC dosage: 4 g/L,mFeOOH dosage: 0.09 g/L, Cr(VI) concentration: 50 mg/L, initial pH: 5.6).

52 M. Su et al. / Science of the Total Environment 647 (2019) 47–56

surfaces at lower initial pH values due to the electronegative Cr(VI) andthemain participants areHCrO4

− ions. The surface of themFeOOH@AC ispositively charged or negatively charged depending on the pH of the so-lution (Zhuang et al., 2014). In this study, the pHpzc of themFeOOH@ACwas calculated to be 7.3 (Fig. S2). Below the pHpzc, the surface of themFeOOH@ACwas found to be positively charged due to the protonationprocess of the functional groups such as the carboxyl and phenol groups(Lazaridis and Charalambous, 2005), facilitating the effective uptake ofCr(VI) by electrostatically attractive force between the mFeOOH@ACand the Cr(VI) anion. As the pH values increased, the positively chargedsites were neutralized and the functional groups were deprotonated,even became negatively charged,which led to the inhibition orweaken-ing of the Cr(VI) adsorption due to electrostatic repulsion and hence a

Scheme 1.A proposedmodel to describe themechanism of Cr(VI) removal in themFeOOH@ACof Cr(III) with Fe(III), ④ electron transfer from the AC to the mFeOOH.

significant decrease of Cr(VI) adsorption capacity. In addition, wefound that the competition between the excess OH– and CrO4

2− for theadsorption sites also reduced the Cr(VI) adsorption onto the adsorbentsunder alkaline conditions, consistent with several previous studies(Yirsaw et al., 2016; Yoon et al., 2011; Zhuang et al., 2014). Our resultssuggest that initial pH is one of the most important parameters for theCr(VI) removal from aqueous solutions which can be attributed to theeffects of the initial pH on the following several factors during the ad-sorption process: the redox properties, the Cr(VI) ionization, and thesurface charge of the adsorbents.

Additional experiments were performed to investigate the ironleaching process and its leaching rate after the reactions under weaklyacidic environments. The results showed that the iron leaching rate

system:① Cr(VI) adsorption,② reduction of Cr(VI) to Cr(III) by the AC,③ coprecipitation

Fig. 6. The effect of initial pH on Cr(VI) removal. (Adsorbent dosage: 4 g/L, Cr(VI)concentration: 50 mg/L, contact time: 480 min, temperature: 30 °C).

53M. Su et al. / Science of the Total Environment 647 (2019) 47–56

decreased from 11.7% to 1.1% in the mFeOOH@AC system as the initialpH values increased from 3 to 11. The pH values after the reactions in-creased 0.1–0.5 unit more than the initial pH values. We also found thatthe mFeOOH particles were still loaded on the AC steadily after the ad-sorption. An optimal and relatively high adsorption capacity was found,while a low iron leaching rate was still maintained at initial pH value of5.6. We hence selected this initial pH value for further experiments.

3.4. Adsorption isotherms

The Cr(VI) concentration was set from10 to 500 mg/L to investigatethe adsorption isotherm for the Cr(VI) adsorption respectively onto the

Fig. 7. (a) The effect of Cr(VI) concentration; (b) Langmuir model; (c) Freundlic model. (

AC and themFeOOH@AC. As shown in Fig. 7 a, the adsorption isothermsof these adsorbents rose at low Ce and thenwere kept steadily at high Cedue to the availability of a vast number of active sites for the Cr(VI) ad-sorption at the low Cr(VI) concentrations and the adsorption becamegradually saturated in the high Cr(VI) concentrations due to less avail-ability of the active sites (Al-Othman et al., 2012; Bradl, 2004).

Langmuir and Freundlich adsorption models were employed to fur-ther investigate the interaction between the Cr(VI) and the adsorbentsby linear regression of the above adsorption processes (Eqs. 6 and 7), re-spectively.

Ce=qe ¼ 1= qm � KLð Þ þ Ce=qm ð6Þ

lgqe ¼ lgK F þ lgCe=n ð7Þ

where Ce (mg/L) and qe (mg/g) are the Cr(VI) concentration and ad-sorption capacity at the equilibrium time, respectively; qm (mg/g) isthe maximum adsorption capacity; KL (L/mg) is the Langmuir constantrelated to the affinity of binding sites. The KL and qm values are respec-tively calculated from the intercept and slope of the plot of Ce/qe versusCe (Fig. 7 b). KF (mg/g) and n are the Freundlich physical constants thatindicate the adsorption capacity and adsorption intensity, correspond-ing to the intercept and slope of the plot of lgqe versus lgCe respectively(Fig. 7c). Table 3 lists the parameters used in the Langmuir andFreundlich adsorption models.

Both Langmuir and Freundlich models fitted well the Cr(VI) adsorp-tion onto the AC and the mFeOOH@AC. However, the R2 values of theLangmuir model were higher than those of the Freundlich model forthese adsorbents, indicating that the Langmuirmodelwasmore suitablethan the Freundlichmodel to the experimental data. Based on the Lang-muir model, themaximumCr(VI) adsorption capacity of the AC and themFeOOH@ACwere 4.8 and 28.1mg/g, close to the experimental data of

Adsorbent dosage: 4 g/L, initial pH: 5.6, contact time: 480 min, temperature: 30 °C).

Table 3The Langmuir and Freundlich model parameters for the adsorption of Cr(VI).

Adsorbent Langmuir Freundlich

KL (L/mg) qm (mg/g) R2 KF (mg/g) n R2

AC 0.027 4.8 0.997 0.69 3.1 0.970mFeOOH@AC 0.082 28.1 0.995 7.03 4.1 0.981

Fig. 8. The removal efficiency of Cr(VI) with the recycled and regenerated mFeOOH@AC(mFeOOH@AC dosage: 4 g/L, Cr(VI) concentration: 50 mg/L, initial pH: 5.6, temperature:30 °C).

54 M. Su et al. / Science of the Total Environment 647 (2019) 47–56

4.5 and 27.2 mg/g, respectively.We conclude that the Cr(VI) adsorptionby the AC and the mFeOOH@AC is a homogeneous monolayer process.The adsorption capacities of Cr(VI) on other adsorbents reported in pre-vious literatures are showed in Table 4. In general, both AC and iron ox-ides (e.g., goethite and hematite) alone showed low Cr(VI) removalefficiencies due to electrostatic repulsion and agglomeration (Ajouyedet al., 2010). Although the experimental conditions were differentfrom each study, the mFeOOH@AC in this study possessed a reasonableadsorption capacity in comparison with other adsorbents. Furthermore,even though the adsorption capacity of the AC-Fe was a little higherthan the mFeOOH@AC, there was no information about the stabilityand regeneration of the AC-Fe and its preparation was complicated.Therefore, here we showed that the synthesized mFeOOH@AC can effi-ciently enhance the Cr(VI) adsorption capacities as the mFeOOH wasdispersed homogeneously on the AC.

3.5. Stability and reusability

Stability and reusability are essential in the adsorbent application.The adsorbents were recycled for the reduction of the processing cost(Aigbe et al., 2018; Liang et al., 2017). Here 0.1 M and 1MHCl solutionswere respectively used for the desorption (leaching) test and the regen-eration study. As shown in Fig. 8, the iron leaching rate was 16.9% afterthe first cycle and subsequently decreased to 5.2%, 2.9% and 1.9% afterthe next 3 cycles. The reduction of the leaching rates is attributed tothe facts that the surface-bounded iron on the adsorbent was easily dis-solved in acidic conditionswhile the other iron distributed on the poresof the adsorbent was relatively difficult to be dissolved. The Cr(VI) re-moval efficiencies were 87.6%, 84.1%, 79.9% and 75.1% after the firstfour cycles. The decrease of the Cr(VI) removal efficiencywas attributedto the less available active sites for Cr(VI) adsorption since total 17.3% ofthe adsorbed Cr(VI) was desorbed by 0.1 M HCl after four desorptionprocesses as a result of the chemisorption mechanism between the Cr(VI) and the mFeOOH@AC (Nityanandi and Subbhuraam, 2009). How-ever, compared to its first removal efficiency of 90.1%, the Cr(VI) re-moval efficiency of the desorbed mFeOOH@AC after four cycles onlydecreased by 14%. Thiswasmainly attributed to the increase of the acid-ity of themFeOOH@ACwith the production of oxygen-containing acidicgroups (such as carboxylic group) after acid treatment, which directlyinteracted with Cr(VI) through complexation that enhanced the Cr(VI)adsorption. This point of view is supported by the fact that the pHpzc

of the mFeOOH@AC decreased to 4.4 (Fig. S3) after acid activation(Huang et al., 2009; Wang and Zhu, 2007). Moreover, the dredging of

Table 4The comparison of adsorption capacities for Cr(VI) removal with different adsorbents.

Adsorbent T (°C) InitialpH

M

mFeOOH@AC 30 5.6 2AC 30 5.6 4Goethite 25 8.0 2Hematite 25 8.0 2SAC (coconut shell activated carbon) 25 2.0 9ATSAC (acid treated activated carbon) 25 2.0 1Carbon-encapsulated iron 25 7.0 1Fe(III)-impregnation sorbent 30 7.0 1AC-Fe 20 5.8 3

micro pores and the expansion of specific surface area after the acid ac-tivation may also play a role.

After four desorption cycles, 1 M HCl solution was used to regener-ated the used mFeOOH@AC by stripping all the mFeOOH and Fe(III)-Cr(III) hydroxides on the surfaces and pores, and subsequentlymFeOOHwas reloaded on the composite. In this way, a Cr(VI) removal efficiencyof 85.4% was obtained. We conclude that the mFeOOH@AC has a goodstability and it is promising for reutilization in the Cr(VI) removalprocess.

4. Conclusion

In this study, a mFeOOH@AC composite was synthesized to achieveenhanced adsorption capacity for Cr(VI). The preparationmethodwas afacile and cost-saving impregnation process. The micro size (0.1–1 μm)FeOOH particles were homogeneously dispersed on the surfaces andpores of the AC. The FTIR and XPS analysis revealed that the Cr(VI) re-moval by mFeOOH@AC was not an ordinary adsorption mechanism,but a adsorption-reduction process where Cr(VI) would be reduced toCr(III) by AC and eventually precipitated as a form of Fe(III)-Cr(III) hy-droxides. Therefore, themFeOOH@AC composite obtained a remarkablyhigher Cr(VI) removal efficiency of 90.4%, 4.5 times that of the AC and6.5 times that of the mFeOOH. The adsorption of Cr(VI) onto themFeOOH@AC agreed well with the Langmuir adsorption model, dem-onstrating that the adsorption processwas controlled bymonolayer ad-sorption. This adsorption process also followed the pseudo second-order kinetics which was chemically adsorbed for the Cr(VI). The Cr(VI) removal was highly pH-dependent and higher adsorption was ob-tained at lower initial pH. The recycling experiment demonstrated that

aximal adsorption capacity (mg/g) Reference

7.2 This study.5 This study.0 (Ajouyed et al., 2010).3 (Ajouyed et al., 2010).5 (Lazaridis and Charalambous, 2005)1.5 (Lazaridis and Charalambous, 2005)0.1 (Zhuang et al., 2014)2.6 (Zhu et al., 2012)4.4 (Sun et al., 2014)

55M. Su et al. / Science of the Total Environment 647 (2019) 47–56

the mFeOOH@AC composite has a sufficient life-span and stabilityunder acidic conditions. Furthermore, this composite could be regener-ated after stripping all the iron oxides on the AC surface by 1M HCl andremodifying with mFeOOH, making this technique cost-effective andsustainable. With these results, we recommend the mFeOOH@AC tobe an efficient and promising material that significantly enhances theCr(VI) removal in water treatment processes.

Acknowledgments

The authors thank the support from the National Natural ScienceFoundation of China (41571302), the Science and Technology Projectof Guangzhou (201803030036), the Fundamental Research Funds forthe Central Universities (2017ZD025 and 21618343), the Water Re-source Science and Technology Innovation Program of GuangdongProvince (2016-26), the Science and Technology Project of GuangdongProvince (2016A010103002) and the Guangzhou Elite Plan (JY201305).

Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2018.07.372.

References

Adhoum, N., Monser, L., 2002. Removal of cyanide from aqueous solution using impreg-nated activated carbon. Chem. Eng. Process. 41, 17–21.

Adhoum, N., Monser, L., 2004. Removal of phthalate on modified activated carbon: appli-cation to the treatment of industrial wastewater. Sep. Purif. Technol. 38, 233–239.

Aigbe, U.O., Das, R., Ho, W.H., Srinivasu, V., Maity, A., 2018. A novel method for removal ofCr(VI) using polypyrrole magnetic nanocomposite in the presence of unsteady mag-netic fields. Sep. Purif. Technol. 194, 377–387.

Ajouyed, O., Hurel, C., Ammari, M., Ben, A.L., Marmier, N., 2010. Sorption of Cr(VI) ontonatural iron and aluminum (oxy)hydroxides: effects of pH, ionic strength and initialconcentration. J. Hazard. Mater. 174, 616–622.

Al-Othman, Z.A., Ali, R., Naushad, M., 2012. Hexavalent chromium removal from aqueousmedium by activated carbon prepared from peanut shell: adsorption kinetics, equi-librium and thermodynamic studies. Chem. Eng. J. 184, 238–247.

Barakat, M.A., 2011. New trends in removing heavy metals from industrial wastewater.Arab. J. Chem. 4, 361–377.

Bradl, H.B., 2004. Adsorption of heavy metal ions on soils and soils constituents. J. ColloidInterface Sci. 277, 1–18.

Chai, L.Y., Wang, X., Wang, H.Y., Yang, W.C., Liao, Q., Wu, Y.J., 2017. Formation of one-dimensional composites of poly(m-phenylenediamine)s based on Streptomyces foradsorption of hexavalent chromium. Int. J. Environ. Sci. Technol. 1–12.

Cong, H.P., Ren, X.C., Wang, P., Yu, S.H., 2012. Macroscopic multifunctional graphene-based hydrogels and aerogels by a metal ion induced self-assembly process. ACSNano 6, 2693–2703.

Espinoza-Quiñones, F.R., Módenes, A.N., Câmera, A.S., Stutz, G., Tirao, G., Palácio, S.M.,Kroumov, A.D., Oliveira, A.P., Alflen, V.L., 2010. Application of high resolution X-rayemission spectroscopy on the study of Cr ion adsorption by activated carbon. Appl.Radiat. Isot. 68, 2208–2213.

Fabiani, C., Ruscio, F., Spadoni, M., Pizzichini, M., 1997. Chromium(III) salts recovery pro-cess from tannery wastewaters. Desalination 108, 183–191.

Gheju, M., Balcu, I., 2011. Removal of chromium from Cr(VI) polluted wastewaters by re-duction with scrap iron and subsequent precipitation of resulted cations. J. Hazard.Mater. 196, 131–138.

Guo, X., Dong, H., Yang, C., Zhang, Q., Liao, C., Zha, F., Gao, L., 2016. Application of goethitemodified biochar for tylosin removal from aqueous solution. Colloids Surf.Physicochem. Eng. Asp. 502, 81–88.

He, B., Yun, Z., Shi, J., Jiang, G., 2013. Research progress of heavy metal pollution in China:sources, analytical methods, status, and toxicity. Chin. Sci. Bull. 58, 134–140.

Huang, G.L., Shi, J.X., Langrish, T.A.G., 2009. Removal of Cr(VI) from aqueous solutionusing activated carbon modified with nitric acid. Chem. Eng. J. 152, 434–439.

Jayakumar, R., Rajasimman, M., Karthikeyan, C., 2014. Sorption of hexavalent chromiumfrom aqueous solution using marine green algae Halimeda gracilis: optimization,equilibrium, kinetic, thermodynamic and desorption studies. J. Environ. Chem. Eng.2, 1261–1274.

Jr, S.V., Malik, D.J., 2002. Characterization and metal sorptive properties of oxidized activecarbon. J. Colloid Interface Sci. 250, 213–220.

Kapoor, A., Viraraghavan, T., Cullimore, D.R., 1999. Removal of heavymetals using the fun-gus Aspergillus niger. Bioresour. Technol. 70, 95–104.

Koduru, J.R., Lingamdinne, L.P., Jiwan-Singh, Choo, K.H., 2016. Effective removal ofbisphenol-A (BPA) from water using a goethite/activated carbon composite. Process.Saf. Environ. Prot. 103, 87–96.

Lazaridis, N.K., Charalambous, C., 2005. Sorptive removal of trivalent and hexavalent chro-mium from binary aqueous solutions by composite alginate–goethite beads. WaterRes. 39, 4385–4396.

Li, Q., Zhuang, L., Chen, S., Xu, J., Li, H., 2012. Preparation of carbon-supported zinc ferriteand its performance in the catalytic degradation of mercaptan. Energy Fuel 26,7092–7098.

Liang, X., Fan, X., Li, R., Li, S., Shen, S., Hu, D., 2017. Efficient removal of Cr(VI) from waterby quaternized chitin/branched polyethylenimine biosorbent with hierarchical porestructure. Bioresour. Technol. 250, 178–184.

Lo, I.M.C., Lam, C.S.C., Lai, K.C.K., 2006. Hardness and carbonate effects on the reactivity ofzero-valent iron for Cr(VI) removal. Water Res. 40, 595–605.

Lv, X., Hu, Y., Tang, J., Sheng, T., Jiang, G., Xu, X., 2013. Effects of co-existing ions and nat-ural organic matter on removal of chromium (VI) from aqueous solution by nano-scale zero valent iron (nZVI)-Fe3O4 nanocomposites. Chem. Eng. J. 218, 55–64.

Macchi, G., Pagano, M., Pettine, M., Santori, M., Tiravanti, G., 1991. A bench study on chro-mium recovery from tannery sludge. Water Res. 25, 1019–1026.

Min, X., Wang, Y., Chai, L., Yang, Z., Liao, Q., 2017. High-resolution analyses reveal struc-tural diversity patterns of microbial communities in Chromite Ore Processing Residue(COPR) contaminated soils. Chemosphere 183, 266–276.

Módenes, A.N., Espinoza-Quiñones, F.R., Palácio, S.M., Kroumov, A.D., Stutz, G., Tirao, G.,Camera, A.S., 2010. Cr(VI) reduction by activated carbon and non-living macrophytesroots as assessed by Kβ spectroscopy. Chem. Eng. J. 162, 266–272.

Mondragon, R., Julia, J.E., Barba, A., Jarque, J.C., 2012. Characterization of silica–waternanofluids dispersed with an ultrasound probe: a study of their physical propertiesand stability. Powder Technol. 224, 138–146.

Monser, L.I., Greenway, G.M., 1996. Liquid chromatographic determination of methyl-amines. Determination of trimethylamine in fish samples using a porous graphiticcarbon stationary phase. Anal Chim Acta 322, 63–68.

Nityanandi, D., Subbhuraam, C.V., 2009. Kinetics and thermodynamic of adsorption ofchromium(VI) from aqueous solution using puresorbe. J. Hazard. Mater. 170,876–882.

Park, S.J., Jang, Y.S., 2002. Pore structure and surface properties of chemically modified ac-tivated carbons for adsorption mechanism and rate of Cr(VI). J. Colloid Interface Sci.249, 458–463.

Qin, X., Liu, F., Wang, G., Weng, L., Li, L., 2014. Adsorption of levofloxacin onto goethite:Effects of pH, calcium and phosphate. Colloids Surf. B. 116, 591–596.

Richard, F.C., Bourg, A.C.M., 1991. Aqueous geochemistry of chromium: a review. WaterRes. 25, 807–816.

Ruan, Z.H., Wu, J.H., Huang, J.F., Lin, Z.T., Li, Y.F., Liu, Y.L., Cao, P.Y., Fang, Y.P., Xie, J., Jiang,G.-B., 2015. Facile preparation of rosin-based biochar coated bentonite for supportingalpha-Fe2O3 nanoparticles and its application for Cr(VI) adsorption. J. Mater. Chem. A3, 4595–4603.

Sahu, J.N., Agarwal, S., Meikap, B.C., Biswas, M.N., 2009. Performance of a modified multi-stage bubble column reactor for lead(II) and biological oxygen demand removal fromwastewater using activated rice husk. J. Hazard. Mater. 161, 317–324.

Schwertmann, H.C.U., Cornell, R.M., 2000. Iron oxides in the laboratory: preparation andcharacterization. Clay Miner. 27, 393.

Shen, W., Mu, Y., Xiao, T., Ai, Z., 2016. Magnetic Fe3O4–FeB nanocomposites with pro-moted Cr(VI) removal performance. Chem. Eng. J. 285, 57–68.

Shi, L.N., Lin, Y.M., Zhang, X., Chen, Z.l., 2011a. Synthesis, characterization and kinetics ofbentonite supported nZVI for the removal of Cr(VI) from aqueous solution. Chem.Eng. J. 171, 612–617.

Shi, L.N., Zhang, X., Chen, Z.L., 2011b. Removal of chromium (VI) from wastewater usingbentonite-supported nanoscale zero-valent iron. Water Res. 45, 886–892.

Sun, Y., Yue, Q., Mao, Y., Gao, B., Yuan, G., Huang, L., 2014. Enhanced adsorption of chro-mium onto activated carbon by microwave-assisted H3PO4 mixed with Fe/Al/Mn ac-tivation. J. Hazard. Mater. 265, 191.

Sun, Y., Li, J., Huang, T., Guan, X., 2016. The influences of iron characteristics, operatingconditions and solution chemistry on contaminants removal by zero-valent iron: areview. Water Res. 100, 277–295.

Tiravanti, G., Petruzzelli, D., Passino, R., 1997. Pretreatment of tannery wastewaters by anion exchange process for Cr(III) removal and recovery. Water Sci. Technol. 36,197–207.

Vlyssides, A.G., Israilides, C.J., 1997. Detoxification of tannery waste liquors with an elec-trolysis system. Environ. Pollut. 97, 147–152.

Wang, S., Zhu, Z.H., 2007. Effects of acidic treatment of activated carbons on dye adsorp-tion. Dyes Pigments 75, 306–314.

Wang, L., Su, L., Chen, H., Yin, T., Lin, Z., Lin, X., Yuan, C., Fu, D., 2015. Carbon paper elec-trode modified by goethite nanowhiskers promotes bacterial extracellular electrontransfer. Mater. Lett. 141, 311–314.

Wang, Q., Huang, L., Pan, Y., Quan, X., Li, P.G., 2017. Impact of Fe(III) as an effectiveelectron-shuttle mediator for enhanced Cr(VI) reduction in microbial fuel cells: Re-duction of diffusional resistances and cathode overpotentials. J. Hazard. Mater. 321,896–906.

Wu, P., Wu,W., Li, S., Xing, N., Zhu, N., Li, P., Wu, J., Yang, C., Dang, Z., 2009. Removal of Cd2+ from aqueous solution by adsorption using Fe-montmorillonite. J. Hazard. Mater.169, 824–830.

Wu, P., Li, S., Ju, L., Zhu, N., Wu, J., Li, P., Dang, Z., 2012. Mechanism of the reduction ofhexavalent chromium by organo-montmorillonite supported iron nanoparticles.J. Hazard. Mater. 219, 283–288.

Wu, J., Lin, G., Li, P., Yin,W., Wang, X., Yang, B., 2013. Heterogeneous Fenton-like degrada-tion of an azo dye reactive brilliant orange by the combination of activated carbon-FeOOH catalyst and H2O2. Water Sci. Technol. 67, 572–578.

Xu, P., Zeng, G.M., Huang, D.L., Feng, C.L., Hu, S., Zhao, M.H., Lai, C., Wei, Z., Huang, C., Xie,G.X., Liu, Z.F., 2012. Use of iron oxide nanomaterials in wastewater treatment: a re-view. Sci. Total Environ. 424, 1–10.

Xu, H., Hu, Z., Lu, A., Hu, Y., Li, L., Yang, Y., Zhang, Z., Wu, H., 2013. Synthesis and super ca-pacitance of goethite/reduced graphene oxide for supercapacitors. Mater. Chem.Phys. 141, 310–317.

56 M. Su et al. / Science of the Total Environment 647 (2019) 47–56

Yang, T., Meng, L., Han, S., Hou, J., Wang, S., Wang, X., 2017. Simultaneous reductive andsorptive removal of Cr(VI) by activated carbon supported beta-FeOOH. RSC Adv. 7,34687–34693.

Yin, C.Y., Aroua, M.K., Daud, W.M.A.W., 2007. Review of modifications of activated carbonfor enhancing contaminant uptakes from aqueous solutions. Sep. Purif. Technol. 52,403–415.

Yirsaw, B.D., Megharaj, M., Chen, Z., Naidu, R., 2016. Reduction of hexavalent chromiumby green synthesized nano zero valent iron and process optimization using responsesurface methodology. Environ. Technol. Innov. 5, 136–147.

Yoon, I.H., Bang, S., Chang, J.S., Kim, M.G., Kim, K.W., 2011. Effects of pH and dissolved ox-ygen on Cr(VI) removal in Fe(0)/H2O systems. J. Hazard. Mater. 186, 855–862.

Yue, Z., Bender, S.E., Wang, J., Economy, J., 2009. Removal of chromium Cr(VI) by low-costchemically activated carbon materials from water. J. Hazard. Mater. 166, 74–78.

Zhao, N., Wei, N., Li, J., Qiao, Z., Cui, J., He, F., 2005. Surface properties of chemically mod-ified activated carbons for adsorption rate of Cr (VI). Chem. Eng. J. 115, 133–138.

Zhou, X., Lv, B., Zhou, Z., Li, W., Jing, G., 2015. Evaluation of highly active nanoscale zero-valent iron coupled with ultrasound for chromium(VI) removal. Chem. Eng. J. 281,155–163.

Zhu, Y., Zhang, H., Zeng, H., Liang, M., Lu, R., 2012. Adsorption of chromium(VI) fromaqueous solution by the iron(III)-impregnated sorbent prepared from sugarcane ba-gasse. Int. J. Environ. Sci. Technol. 9, 463–472.

Zhuang, L., Li, Q., Chen, J., Ma, B., Chen, S., 2014. Carbothermal preparation of porouscarbon-encapsulated iron composite for the removal of trace hexavalent chromium.Chem. Eng. J. 253, 24–33.

Zuo, X., Chen, M., Fu, D., Li, H., 2016. The formation of alpha-FeOOH onto hydrothermalbiochar through H2O2 and its photocatalytic disinfection. Chem. Eng. J. 294, 202–209.