Chromium(VI) removal by mechanochemically sulfidated zero valent iron and its effect on dechlorination of trichloroethene as a co-contaminant

Science of the Total Environment 650 (2019) 419–426

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Chromium(VI) removal by mechanochemically sulfidated zero valent iron and its effect on dechlorination of trichloroethene as a co-contaminant Haowen Zou a , Erdan Hu a , Shangyuan Yang b , Li Gong a , Feng He a ,c ,⁎ a b c

College of Environment, Zhejiang University of Technology, Hangzhou 310014, China R&D Center of Zhejiang Zone-King Environment Co., Ltd, Hangzhou, 310014, China Key Laboratory of Microbial Technology for Industrial Pollution Control of Zhejiang Province, Zhejiang University of Technology, Hangzhou 310014, 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

• Cr(VI) removal by S-mZVIbm was mainly a surface reduction and precipitation process. • Mechanochemical sulfidation enhanced Cr(VI) removal by increasing surface area. • Electron efficiency of Cr(VI) removal was 100% but particle efficiency was b1%. • Cr(VI) was an electron sink for TCE dechlorination by S-mZVIbm .

a r t i c l e

i n f o

Article history: Received 13 July 2018 Received in revised form 26 August 2018 Accepted 1 September 2018 Available online 03 September 2018 Editor: Jay Gan Keywords: Chromium Electron efficiency Passivation Selectivity Sulfidation Zero-valent iron

a b s t r a c t Mechanochemically sulfidated microscale zero valent iron (S-mZVIbm ) is a promising groundwater remediation material as it has been proven to be not only efficient in dechlorinating chlorinated compounds but also amenable to up-scaling. Yet, its efficiency in treating metal contaminants remains barely studied. In this study, we investigated the mechanism and efficiencies of Cr(VI) removal by S-mZVIbm and its effect on TCE dechlorination as a co-contaminant. The Cr(VI) removal by S-mZVIbm was mainly a chemisorption process and its kinetics was well fitted by a pseudo-second-order model. Alkaline pH inhibited Cr(VI) removal while dissolved oxygen slightly depressed the Cr(VI) removal. The Cr(VI) removal rapidly formed a non-conductive layer on S-mZVIbm surface to hinder further electron transfer from Fe0 core before H+ was able to accept any electrons to produce H2, which resulted in 100% electron efficiencies of Cr(VI) removal but b1% of Fe0 utilization efficiency. The presence of Cr(VI) also dramatically inhibited the dechlorination of TCE and its electron efficiency as a co-contaminant by passivating the FeS surface. Therefore, Cr(VI) is likely to be an electron sink if present for remediation of other contaminants in groundwater. © 2018 Elsevier B.V. All rights reserved.

1. Introduction ⁎ Corresponding author at: College of Environment, Zhejiang University of Technology, Hangzhou 310014, China. E-mail address: [email protected] (F. He).

https://doi.org/10.1016/j.scitotenv.2018.09.003 0048-9697/© 2018 Elsevier B.V. All rights reserved.

Chromium is a priority pollutant and is often found in groundwater due to its widely use in industrial processes such as leather tanning, electroplating and stainless-steel production (Palmer and Wittbrodt,

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1991; Sarin et al., 2006; Yoon et al., 2011). Cr(VI) and Cr(III) are two oxidation states of chromium. Cr(VI) species, such as Cr2O72− and HCr2O7− , are harmful to both animals and plants due to their strong toxicity and mobility (Jin et al., 2016a; Jin et al., 2015; Jin et al., 2016b; Myers et al., 2000; Qian et al., 2013; Zhitkovich et al., 2002), whereas Cr(III) species are less hazardous and have low solubility in water (i.e., generally precipitate as hydroxides, Ksp(Cr(OH)3) = 6.7 × 10−31 ) (Ai et al., 2008; He et al., 2013; Hoch et al., 2008; Manning et al., 2007; Shen et al., 2013). Therefore, converting Cr(VI) to Cr(III) is a common strategy in remediating Cr-contaminated groundwater. Zero valent iron (ZVI), considered as an efficient and low-cost reductant, has been widely investigated for the dechlorination of chlorinated compounds and reduction of redox-sensitive heavy metals and anions in groundwater (Feng et al., 2018; Fu et al., 2014; Gao et al., 2015; Guan et al., 2015; He et al., 2010; Lv et al., 2014; Mu et al., 2017; O'Carroll et al., 2013; Tang et al., 2017). Previous studies have proven that ZVI could easily transform Cr(VI) to Cr(III) via coupled oxidation with Fe0 and Fe2+ , and then Cr(III) species precipitated with Fe(III) generated by forming (CrxFe1-x)(OH)3 (Li et al., 2008; Manning et al., 2007; Ponder et al., 2000). One of the most challenging aspects of applying ZVI in groundwater remediation is its low selectivity (or electron utilization efficiency) (Fan et al., 2016; Gu et al., 2017; Li et al., 2018). This is due to that the concurrent corrosion of ZVI competes electrons from Fe0 with contaminants. Our previous research has shown that when microscale ZVI was used for trichloroethene (TCE) dechlorination under simulated plume conditions, only 0.19% of electrons from Fe0 was used for TCE reduction while the rest was consumed by water corrosion (Gu et al., 2017). Yet, the reaction of Cr(VI) with ZVI is more favorable and the precipitation of Cr (III) species that are poor in conducting electrons would cause the passivation of ZVI surface. Both factors could significantly affect the selectivity of ZVI toward Cr(VI), which remains to be explored. In recent years, one modification of ZVI called sulfidation—defined as ZVI modified by sulfur compounds—has emerged as a promising approach in enhancing both the particle reactivity but also the selectivity in dechlorinating TCE, one of the most detected contaminants in groundwater (Fan et al., 2016; Han and Yan, 2016; Li et al., 2017; Rajajayavel and Ghoshal, 2015). Until now, researches on sulfidated ZVI have been primarily focused on nanoscale ZVI (Cao et al., 2017; Dong et al., 2018; Fan et al., 2017; He et al., 2018; Li et al., 2016; Tang et al., 2016; Wu et al., 2018). However, nanoscale ZVI (nZVI) is expensive and only likely to be used at modest quantity for targeted application (Crane and Scott, 2012). In addition, the reported nZVI sulfidation processes were typically carried out in aqueous phase using dissolved, low-valent sulfur species that are readily oxidized by oxygen, which makes the process hard to be scaled up (Fan et al., 2016; Li et al., 2017). To solve these problems, our group recently reported the synthesis of sulfidated microscale ZVI (S-mZVIbm ) with ZVI and elemental sulfur using mechanochemical method (Gu et al., 2017; Huang et al., 2017). The obtained S-mZVIbm could dechlorinate TCE with at least an order magnitude faster rates and greater electron efficiencies than the unsulfidated counterparts (mZVIbm ) and the process is highly amenable to up-scaling. Yet, its efficiencies in treating heavy metal contaminants such as Cr have not been fully determined. Several studies on Cr(VI) removal by S-ZVI have been reported (Du et al., 2016; Gong et al., 2017; Li et al., 2018; Shao et al., 2018). Du et al. (2016) prepared a nanoscale FeS@Fe0 material using a two-step method. Nanoscale ZVI was firstly synthesized through reduction of ferrous salts by borohydride and then dispersed in ferrous iron solution followed by adding sulfide. The FeS@Fe0 showed much higher reactivity than the unsulfidated nZVI toward Cr(VI) sequestration by forming FeCr hydroxides. This was attributed to the coverage of FeS accelerating the electron transfer from Fe0 core to the sorbed Cr(VI). Dissolved oxygen (DO) was found to negatively affect Cr(VI) removal. Gong et al. (2017) fabricated S-nZVI particles though “one-pot” synthesis by adding dithionite during nZVI synthesis. The FeS coating on the Fe0

surface effectively inhibited the aggregation of Fe0 and resulted in larger specific surface area and consequently higher Cr(VI) removal efficiency than nZVI. More recently, Shao et al. (2018) showed that oxygenation, however, significantly enhanced Cr(VI) removal by sulfidated microscale ZVI, which was attributed to enhanced Fe(II) production that resulted from accelerated corrosion of Fe(0), in contrast to that reported by Du et al. (2016). Nonetheless, the electron selectivity of S-ZVI toward Cr(VI) was not determined and how sulfidation affects particle selectivity was not studied under typical anoxic conditions in groundwater. In addition, as a strong passivation agent, Cr(VI) can significantly affect the interaction of ZVI or S-ZVI with its typical cocontaminants such as TCE in groundwater. Yet, this effect has not been explored in previous studies. The overall objective of this study is to understand the interfacial reaction between S-mZVIbm and Cr(VI) and its effects on dechlorination of TCE as a co-contaminant. The specific objectives are to (i) quantify and compare the removal rate of Cr(VI) and the electron selectivity by S-mZVIbm and mZVIbm under various conditions, (ii) gain insights into the mechanism of Cr sequestration by S-mZVIbm , and (iii) determine the effect of Cr passivation on TCE dechlorination by S-mZVIbm . 2. Materials and methods 2.1. Chemicals All chemicals used in this study were of analytical grade or better unless otherwise mentioned. Microscale ZVI (reagent grade), elemental S powder, FeS, NaCl, NaOH and HCl (reagent grade) were provided by Aladdin (Shanghai, China). Potassium dichromate (K2Cr2O7, 99.8%) was purchased from Wuxi Haishuo Biological Co., LTD (Jiangsu, China). TCE (99%, GC grade), methanol (N99%, GC) and liquid chlorinated ethene standards in methanol including TCE (1000 ppm), cisDCE (1000 ppm), trans-DCE (1000 ppm), 1,1-DCE (1000 ppm), and VC (1000 ppm) were obtained from Aladdin (Shanghai, China). Gaseous alkane, alkene, and alkyne standards (1000 ppm of methane, ethene, ethane, acetylene, propene, propane, butene, butane, pentene, pentane, hexene, and hexane) were purchased from Dalian Special Gases Co. (Dalian, China). Ultra-high purity nitrogen, hydrogen and air were supplied by Hangzhou Special Gas Co. (Hangzhou, China) for GC measurements. Solutions were typically prepared with deionized (DI) water, which was deoxygenated by bubbling N2 for 0.5 h before use except for experiments under aerobic conditions. 2.2. Particle preparation and characterization S-mZVIbm particles were prepared following the reported method (Gu et al., 2017). Briefly, ZVI and sulfur powders (S/Fe molar ratio of 0, 0.1, and 0.2) were placed in stainless steel jars with zirconia balls and milled at 400 rpm under argon conditions in a planetary ball mill. After 20 h of milling, S-mZVIbm particles were collected in a N2-filled glovebag and stored in a glovebox before use. The specific surface area of S-mZVIbm was measured by N2 adsorption-desorption tests at 77 K using a Micromeritics ASAP2020 (USA). The oxidation states of elements on S-mZVIbm surface before and after Cr(VI) removal were obtained by a Kratos AXIS Ultra DLD XPS and the details are provided in SI. 2.3. Batch experiments of Cr(VI) removal and particle efficiencies Batch experiments of Cr(VI) removal were performed in a 250 mL, three-necked round-bottom flask with continuous mechanical stirring at 400 rpm. Typically, 200 mL Cr(VI) solution (10 mg/L) with initial pH of 6 was added to the flask followed by S-mZVIbm or mZVIbm (1 g/L). At selected time intervals, 1 mL of sample was withdrawn and filtered through a 0.22 μm nylon filter membrane (ANPEL Laboratory Technologies, Shanghai, China) for measurements of Cr(VI) and

H. Zou et al. / Science of the Total Environment 650 (2019) 419–426

dissolved iron. Control tests suggested that the filter did not remove Cr (VI). The Cr(VI) removal was tested under different initial Cr(VI) concentration (5, 10, 15 mg/L), pH (5, 6, 7, 8) and S/Fe molar ratio (0, 0.1, 0.2). In addition, to investigate the effect of dissolved oxygen, experiments were conducted under three conditions: (i) Continuously purging vessels with nitrogen, (ii) continuously purging vessels with oxygen, and (iii) exposed to air without aeration. All the experiments were at least duplicated. To determine the electron utilization efficiency of Cr(VI) removal (εe, Cr) by ZVI particles, experiments were conducted in 52 mL glass vials. The reactors were prepared in an argon-filled glove-box and contained 26 mL of deoxygenated Cr(VI) solution, 26 mL of headspace, and 1 g/L of ZVI particles. The reactors were mixed on a rotary shaker (30 rpm) at 25 ± 0.5 °C in an incubator for 3 h and sampled for determination of Cr(VI) and H2 generation. Then, the εe, Cr can be determined using Eq. (1). εe;Cr ¼

3MCr 3MCr þ 2MH2

ð1Þ

where MCr and MH2 are the molar quantity of Cr reduced and H2 generated, respectively, at the end of batch experiments. The utilization efficiency of Fe(0) (εFe(0)) can be determined using Eq. (2). εFeð0Þ ¼

MI;Feð0Þ −M F;Feð0Þ MI;Feð0Þ

ð2Þ

where MI, Fe(0) and MF, Fe(0) are the molar quantity of Fe(0) at the start and end of reaction in the particles, respectively. 2.4. Batch experiments of TCE dechlorination and electron efficiency of dechlorination TCE dechlorination experiments were conducted in 52 mL glass vials sealed with PTFE septa lined caps. The reactors were prepared in an argon-filled glove-box and contained 26 mL of deoxygenated solution with 10 mg/L TCE and 0–50 mg/L Cr(VI), 26 mL of headspace, and 10 g/L S-ZVIbm (S/Fe = 0.1). The vials were mixed on a rotary shaker (30 rpm) at 25 ± 0.5 °C in an incubator. At selected intervals, 100 μL of headspace sample was withdrawn for measurement of TCE and its reaction products. The H2 concentration in the parallel reactors was also measured. The εe, TCE of TCE dechlorination can be determined using Eq. (3). X εe;TCE ¼ X

ni p i

i

ni pi þ 3MCr þ 2MH2

ð3Þ

i

The values of n depends on the degradation products formed (Gu et al., 2017), ni is stoichiometry for product i, and pi is the molar quantity of that product. 2.5. Chemical analysis Concentrations of Cr(VI) and total Cr in solution were determined using the 1, 5- diphenylcarbazide colorimetric method by UV–visible spectrophotometer at the wavelength of 540 nm while the concentration of Fe2+ was measured using the o-phenanthroline method at the wavelength of 510 nm. The total dissolved Fe concentrations were determined using a Persi TAS-990 atomic absorption spectrophotometer (AAS). Dissolved sulfide was measured using the methylene blue method. TCE and its degradation products were measured by gas chromatography (GC) coupled with FID detector and H2 was measured using GC-TCD. The details can be found in SI.

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3. Results and discussion 3.1. Cr(VI) removal kinetics As shown in Fig. 1A, an increase of initial Cr(VI) concentration from 5 to 15 mg/L decreased Cr(VI) removal efficiency from 52% to 24% by SmZVIbm within 180 min, which indicated an adsorptive saturation at tested aqueous Cr(VI) concentrations. The removal of Cr(VI) by SmZVIbm could involve the adsorption of Cr(VI) onto S-mZVIbm via physical and chemical interactions and then reduction of adsorbed Cr(VI) to Cr(III) by Fe(0) or by solid Fe(II) phase such as FeS (Du et al., 2016; Gong et al., 2017). The corrosion of ZVI or dissolution of FeO or FeS would give out Fe2+ , which can also reduce Cr(VI) as a secondary reductant. The increase of Cr(VI) concentration could promote the formation of (CrxFe1bm surface, which inhibited electron x)(OH)3 precipitates on S-mZVI 0 transfer from Fe core to Cr(VI), and therefore retarded the reduction of Cr(VI) (Lu et al., 2012). The kinetics of Cr(VI) removal at different Cr(VI) concentration was analyzed using pseudo-first-order and pseudo-second-order models (Wan et al., 2018a; Wan et al., 2018b) and the results are summarized in Table S1. It is clear that pseudo-second-order kinetic modeling provides better fitting of the adsorption data, which is consistent with many previous findings (Ai et al., 2008; Liang et al., 2014; Lv et al., 2012). The results illustrate that Cr(VI) removal by S-mZVIbm was predominantly a chemisorption process (Gerente et al., 2007).

3.2. Effects of S/Fe molar ratio on Cr(VI) removal Cr(VI) removal by S-mZVIbm with different S/Fe molar ratios was studied and the results are shown in Fig. 1B. S-mZVIbm removed Cr (VI) more efficiently than mZVIbm , and increasing S/Fe molar ratio promoted Cr(VI) sequestration, which was 5.2%, 37.4% and 55.8% for mZVIbm (BET surface area: 0.21 m2 /g), S-mZVIbm (S/Fe = 0.1, 1.46 m2 /g) and S-mZVIbm (S/Fe = 0.2, 2.08 m2 /g), respectively. It is interesting that the increase of Cr(VI) removal was almost linear to the increase of specific surface area of ZVI particles, and the surface-area normalized Cr(VI) removal was comparable (2.48, 2.56 and 2.68 mg/m2 for mZVIbm , S-mZVIbm (S/Fe = 0.1), and S-mZVIbm (S/Fe = 0.2), respectively). This suggests that the Cr(VI) removal was mainly determined by the number of surface reaction sites and the enhanced removal of Cr(VI) by S-mZVIbm was likely not due to the formation of FeS but rather the increase of surface area by mechanochemical sulfidation. In addition, the surface-area normalized Cr(VI) removal by S-mZVIbm is about twice of that reported for sulfidated nanoscale ZVI (~1.1–1.3 mg/m2 ) (Du et al., 2016; Gong et al., 2017), suggesting the higher density of reactive sites on S-mZVIbm . However, the exact underlying mechanism for this difference is not clear although it may relate to the different surface properties of milled and chemically sulfidated ZVI. Despite the formation of FeS on S-mZVIbm played a minor role in the total Cr(VI) removal, the kinetics of Cr(VI) removal by S-mZVIbm showed a different pattern from that by mZVIbm . The removal of Cr (VI) by mZVIbm was rapid and completed in 10 min while that by S-mZVIbm also had an initial rapid stage but followed by a second slow stage (up to hours). As a comparison, the kinetics of Cr(VI) removal by the milled FeS (FeSbm , BET surface area: 1.4 m2 /g) showed a similar pattern as S-mZVIbm ,which suggested that the removal of Cr(VI) by S-mZVIbm was controlled by the FeS phases at the second stage. In addition, S-mZVIbm (S/Fe = 0.1) outperformed FeS bm for Cr(VI) removal despite of their similar BET surface area. Therefore, the initial stage of rapid and major Cr(VI) removal by S-mZVIbm was mainly determined by the ZVI phases. This notion is further demonstrated by the detection of 0.3 mg/L SO 42− after Cr(VI) (10 mg/L) sequestration by S-mZVIbm (S/Fe = 0.1), which was produced from the oxidation of FeS phases by Cr(VI) and accounted for only 16% of the total Cr(VI) removal.

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Fig. 1. (A) The kinetics of Cr(VI) removal by S-mZVIbm (S/Fe = 0.1) at initial Cr(VI) concentrations of 5, 10, and 15 mg/L. (B) Comparison of 10 mg/L Cr(VI) removal by S-mZVIbm (S/Fe = 0.1, 0.2), mZVIbm , and FeSbm particles (particle dosage: 1 g/L; initial pH: 6; exposed to N2).

3.3. Effects of dissolved oxygen and pH on Cr(VI) removal

3.4. Evolution of aqueous Cr(VI), Fe(II), and Fe(III) during sorption

Figs. 2A and S2 show the dependence of Cr(VI) removal by SmZVIbm with S/Fe of 0.1 and 0.2 on dissolved oxygen (DO), respectively. DO only slightly depressed the Cr(VI) removal by both S-mZVIbm , likely due to it weakly competed surface sites/electrons with Cr(VI). The inhibitory factor of DO on Cr(VI) removal by bare or sulfidated nanoscale ZVI was commonly reported in previous studies (Du et al., 2016; Zhang et al., 2018). It is noteworthy that although the Cr(VI) removal by SmZVIbm in the presence of air or N2 was comparable, the resulting solution pH after reaction was quite different, with pH of 7.8 for N2 and 6.8 for air atmosphere. The low solution pH with air demonstrates that dissolved oxygen can compete with Cr(VI) for Fe2+ , which generated H+ , as shown in Eq. (4).

The aqueous Cr(VI), total Cr (TCr), Fe(II), and total Fe (TFe) concentrations during the removal of Cr(VI) by S-mZVIbm were monitored and the results are shown in Figs. 3 and S3. The difference between the concentrations of TCr and Cr(VI) was the concentration of aqueous Cr(III). The decrease of TCr was ascribed to the precipitation of Cr(III) on the surface of S-mZVIbm . The concentrations of TCr and Cr(VI) were nearly the same for S-mZVIbm (S/Fe = 0.1, 0.2), which suggested that all Cr (III) were precipitated on particle surface. Furthermore, both TFe and Fe(II) were not detected, which can be explained by the fact that Fe (II) reacted with Cr(VI) instantly and transformed to Fe(III) and then precipitated on the S-mZVIbm surface in forms of (CrxFe1-x)(OH)3. The precipitations formed a passivation layer on particle surface, which inhibited the further corrosion of ZVI to give out Fe(II). A control experiment without Cr(VI) showed that the corrosion of S-mZVIbm produced 0.05 mg/L Fe(II) and 0.035 mg/L S2− (Fig. S4). This suggests that dissolution of S-mZVIbm in water was negligible and aqueous reduction should not contribute much to the Cr(VI) removal.

4Fe2þ þ O2 þ 10H2 O→4FeðOHÞ3 þ 8Hþ

ð4Þ

With this reaction, the removal of Cr(VI) should be depressed as Fe2+ , an important reductant for Cr(VI), was consumed. However, the protons generated in Eq. (4) could alleviate the surface passivation, which promoted the Cr(VI) removal on the other hand. Since the final Cr(VI) removal was comparable, the two effects seem balanced. Removal of Cr(VI) by S-mZVIbm was strongly pH-dependent. The Cr (VI) removal at 180 min decreased from 41.7% to 17.0% as pH increased from 5 to 8 under N2 atmosphere (Fig. 2B). Decreasing pH alleviated surface passivation and caused faster iron corrosion, which favored Cr (VI) removal. Moreover, under alkaline conditions, the surface bound and free ferrous ions are more easily oxidized, forming passivation layer on the surface and further inhibiting Cr(VI) removal (Powell et al., 1995).

3.5. XPS analysis Fig. 4 shows the XPS spectra of S-mZVIbm before and after Cr(VI) sequestration. The Fe 2p peaks at 709.2 and 711.2 eV corresponded to Fe 2p3/2 for Fe(II) and Fe(III), respectively, while the peaks at 719.8, 722.8, and 724.8 eV were assigned to Fe 2p1/2 for Fe(0), Fe(II) and Fe (III), respectively (Bhargava et al., 2007; Graat and Somers, 1996; Zhang et al., 2013). Moreover, the satellite peak positions for Fe(II) were 714.6 and 729.5 eV and those for Fe(III) were 718.9 and

Fig. 2. Effects of oxygen (A) and initial pH (B) on Cr(VI) removal by S-mZVIbm (S/Fe = 0.1) (Cr(VI): 10 mg/L; particle dosage: 1 g/L).

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Fig. 3. (A) Aqueous TCr and Cr(VI) concentrations during removal of Cr(VI) by S-mZVIbm (S/Fe = 0.1) and (B) aqueous TFe and Fe2+ concentrations during Cr(VI) removal by S-mZVIbm (Cr (VI): 10 mg/L; particle dosage: 1 g/L; initial pH: 6; with N2).

733.4 eV (Yamashita and Hayes, 2008). Fitting these peaks gave the distribution of iron oxidation states. After Cr(VI) uptake, the surface Fe (0) molar content decreased from 1% to 0 and the Fe(II) molar content decreased from 52.2% to 31.8%. Meanwhile, the Fe(III) molar content increased from 46.8% to 68.2%. This suggested that all Fe(0) and part of the Fe(II) on particle surface were transformed to Fe(III) after Cr(VI) sequestration. The S 2p peak at 161.3, 162.6, and 168.1 eV corresponded to FeS (Stypula and Stoch, 1994), FeS2 (Dedonato et al., 1993), and surface-bounded SO42− (Thomas et al., 1998), respectively. After reaction with Cr(VI), the FeS molar content decreased from 100% to 20.2% accompany with the increase of S22− from 0% to 48.8%. Moreover, SO 42− increased from 0% to 31%, which indicated that FeS

reacted with Cr(VI) resulting the oxidation of sulfur (Eqs. (5)–(6)) (Demoisson et al., 2005). Clearly, most of the surface FeS had been oxidized by Cr(VI) to form FeS2 and SO42− . þ 6FeS þ 3HCrO− 4 þ 21H →3FeðIIIÞ þ 3FeS2 þ 3CrðIIIÞ þ 12H2 O

ð5Þ

2− þ 3FeS2 þ 15HCrO− 4 þ 57H →3FeðIIIÞ þ 6SO4 þ 15CrðIIIÞ þ 36H2 O ð6Þ

The Cr 2p peaks at 577.6 and 587.3 eV, 579.8 and 589.1 eV corresponded to Cr(III) and Cr(VI) (Ikemoto et al., 1976), respectively. The fitting of these peaks suggests that the Cr(III) and Cr(VI) molar contents were 74.8% and 25.2%, respectively. This indicates that Cr(VI) removal by S-mZVIbm was predominately a chemisorption process with

Fig. 4. (A) Fe 2p and (B) S 2p XPS spectra of S-mZVIbm (S/Fe = 0.1) before and after Cr(VI) adsorption and (C) Cr 2p and (D) O 1 s XPS spectra of Cr-laden S-mZVIbm (S/Fe = 0.1).

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caused by the adsorption of partial Cr(VI). Therefore, despite electron utilization efficiencies were 100%, Fe(0) utilization efficiencies of all particles were fairly low for Cr(VI) sequestration, as shown in Table 1. The b1% utilization efficiency of Fe(0) in all cases suggests that Cr(VI) is a strong passivator for ZVI and sulfidated ZVI and would be detrimental for these particles' efficiency in degrading other contaminants, which will be elaborated in Section 3.8.

Table 1 Electron and Fe0 utilization efficiency of Cr(VI) sequestration under different conditions (MH2 was zero for all conditions). Reaction variables Initial Cr(VI) concentration

S/Fe ratio

Initial pH

5 mg/L 10 mg/L 15 mg/L 0 0.1 0.2 5 6 7 8

MCr μmol

εe %

εFe(0) %

9.60 13.50 13.08 1.52 13.50 20.19 15.72 13.50 9.41 5.98

100 100 100 100 100 100 100 100 100 100

0.41 0.15 0.22 0.06 0.15 0.47 0.23 0.15 0.54 0.18

3.7. Proposed mechanism of Cr(VI) removal by S-mZVIbm On the basis of above results, the mechanism of Cr(VI) removal by SmZVIbm is proposed and summarized in Scheme 1. Since Fe2+ was absent during Cr(VI) removal by S-mZVIbm and only negligible Fe2+ (b0.05 mg/L) was detected during S-mZVIbm corrosion without Cr (VI), the Cr(VI) removal by Fe2+ was negligible. To further verify this point, phenanthroline was added into the Cr(VI)/S-mZVIbm system to trap the free Fe2+ and negligible difference in Cr(VI) removal was observed (Fig. S5). Therefore, Cr(VI) removal by S-mZVIbm was mainly a surface chemisorption process. The adsorbed Cr(VI) can be reduced to Cr(III) by electrons from Fe0 via conductive FexOy or FeS while it can also be reduced directly by FeS and Fe(II) oxides. The produced Cr(III) was present as the forms of Cr0.33Fe0.67(OH)3, which resulted in a passivation layer inhibiting further removal of contaminants.

reduction of Cr(VI) to Cr(III). Further analysis suggests that the Cr(III) compound formed at S-mZVIbm surface was Cr(OH)3 not Cr2O3 because the chromium 2p peaks cannot be appropriately fitted with the Cr2O3 2p peaks proposed by Pratt and McIntyre (1996). The O 1 s peak at 530, 531.4, and 533.6 eV correspond to O2− ,\\OH and H2O, respectively (Manning et al., 2007). The O 1 s spectra can be reasonably fitted with the Cr(OH)3 rather than Cr2O3 1 s spectra (Biesinger et al., 2004), which also suggested the presence of Cr(OH)3. Previous researches further suggested that Cr(III) was incorporated onto the iron oxyhydroxide structure and generated (CrxFe1-x)(OH)3 (0 b x b 1) when Cr(VI) reacted with both zero valent iron (x is typically 0.66) and iron sulfide surfaces (x is typically 0.75) (Li et al., 2008; Mullet et al., 2004). Therefore, the Cr(III) endproduct here is expected to be (CrxFe1-x)(OH)3 and the x can be determined to be 0.33 by the XPS analysis, which showed the surface Cr/Fe molar ratio was 1/2.

3.8. Effect of Cr(VI) on TCE dechlorination by S-mZVIbm Fig. 5A shows the effect of Cr(VI) concentration on TCE dechlorination by S-mZVIbm . The S-mZVIbm loading in the current experiments was increased to 10 g/L because the Cr(VI) removal experiments have suggested that 10 mg/L of Cr(VI) can completely passivate the surface of 1 g/L S-mZVIbm . When the Cr(VI) concentration was increased from 0 to 10 mg/L, the TCE degradation at 24 h was decreased from 90% to 53%. During the process, the Cr(VI) was completely removed within 2 h while the solution pH was rapidly increased from 6.0 to 10.3 due to the consumption of H+ caused by Cr(VI) removal and corrosion of S-mZVIbm (Fig. 5C), which produced about 4.1 μmol H2 at the end of reaction (Fig. 5B). Therefore, the chemisorption of Cr(VI) apparently passivated the surface of S-mZVIbm for TCE dechlorination. Our previous study suggested that TCE dechlorination sites were mainly on FeS phase. The inhibition of TCE dechlorination by Cr(VI) suggested that the surface FeS sites were passivated by reacting with Cr(VI). This is consistent with the XPS results that FeS was oxidized by Cr(VI) to

3.6. Particle efficiencies of Cr(VI) removal It is noteworthy that H2 generation was not detected in any of the Cr (VI) removal experiments by either mZVIbm or S-mZVIbm under any tested conditions. Therefore, the electron utilization efficiencies were 100% for all Cr(VI) removal experiments. This is likely because Cr(VI) is a much more preferred electron acceptor than H+ and completely overcompeted H+ for the surface reactions on either mZVIbm or SmZVIbm . However, the removal of Cr(VI) (10 mg/L) by all particles was not complete, which suggested that electrons from Fe(0) were not accessible by both H+ and Cr(VI) after the surface passivation

Cr(III)

Cr(VI)

(Cr0.33Fe0.67)(OH)3

adsorption precipitation

reduction

reduction

FeS

precipitation

FexOy

ee-

Fe0

e-

Scheme 1. Proposed mechanism for Cr(VI) removal by S-mZVIbm .

H. Zou et al. / Science of the Total Environment 650 (2019) 419–426

425

Fig. 5. Effects of Cr(VI) concentration on (A) TCE dechlorination and (B) H2 production by S-mZVIbm . The change of Cr(VI) concentration and solution pH during TCE dechlorination by SmZVIbm at initial Cr(VI) concentration of (C) 10 mg/L and (D) 50 mg/L. All experiments were carried out under anoxic conditions with 10 g/L of S-mZVIbm (S/Fe = 0.1) and the solution was not buffered.

form higher valent sulfur species and Cr(III). When the Cr(VI) was further increased from 10 mg/L to 50 mg/L, the TCE dechlorination was completely inhibited. At the same time, the Cr(VI) removal was 35% in the first 2 h while only 5% was removed in the next 22 h and hydrogen generation was not observed thorough the experiment (Fig. 5B). Therefore, the high content of Cr(VI) completely overcompeted TCE and protons and passivated the surface before TCE dechlorination even started. The εe, TCE of TCE degradation in the absence and presence of Cr(VI) was calculated and the results are shown in Table 2. The εe, TCE decreased from 26.8% to 0 when the Cr(VI) concentration was increased from 0 to 50 mg/L. The decrease of εe, TCE in the presence of Cr(VI) again demonstrated the strong electron competing ability of Cr(VI) over TCE. Therefore, Cr(VI) is going to be an electron sink if present during the remediation of chlorinated solvents such as TCE in groundwater.

mZVIbm was mainly due to its enhanced surface area. As Cr(VI) removal was a rapid process, it completely overcompeted H+ , which resulted in 100% electron efficiencies of Cr(VI) removal. However, the removal of Cr (VI) formed a passivated layer composed of Fe/chromium hydr(oxide)s on the surface of S-mZVIbm . This caused the electrons from remaining Fe0 core inaccessible, which resulted in particle efficiency b1%. The passivation of S-mZVIbm by Cr(VI) also dramatically inhibited the rate and the electron efficiency of TCE dechlorination as a co-contaminant.

4. Conclusions

Appendix A. Supplementary data

The present study investigated the interfacial reaction mechanism and electron efficiencies of Cr(VI) removal by S-mZVIbm as well as its effect on TCE degradation as a co-contaminant. The Cr(VI) removal by SmZVIbm was mainly a surface reduction and precipitation process and aqueous reduction by dissolved Fe2+ and S2− was negligible. The higher Cr(VI) removal by S-mZVIbm than its unsulfidated counterparts

Acknowledgements The majority of this work was supported by the Natural Science Foundation of Zhejiang Province (LR16E080003).

Additional information on methods, adsorption fitting results, Cr(VI) removal by S-mZVIbm (S/Fe = 0.2), corrosion results of S-mZVIbm (S/Fe = 0.1) and removal of Cr(VI) by S-mZVIbm in the presence of phenanthroline. Supplementary data to this article can be found online at doi:https://doi.org/10.1016/j.scitotenv.2018.09.003.

References Table 2 Electron utilization efficiency of TCE dechlorination with and without Cr(VI). Cr(VI) concentration mg/L

MCr μmol

MTCE μmol

MH2 μmol

εe, TCE %

0 10 50

0 3.7 7.6

7.3 3.5 0

10.0 4.1 0

26.8 15.2 0

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