Reductive Dechlorination of Trichloroethene by Zero-valent Iron

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Reductive Dechlorination of Trichloroethene by Zero-valent Iron Nanoparticles: Reactivity Enhancement through Sulfidation Treatment Yanlai Han, and Weile Yan Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b03997 • Publication Date (Web): 08 Nov 2016 Downloaded from http://pubs.acs.org on November 8, 2016

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Reductive Dechlorination of Trichloroethene by Zero-valent Iron

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Nanoparticles: Reactivity Enhancement through Sulfidation Treatment

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Yanlai Han1, Weile Yan1,*

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Department of Civil, Environmental, and Construction Engineering, Texas Tech University,

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Lubbock, TX, 79409, USA

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Corresponding author. Tel: +1 806 834 3478; Fax: +1 806 742 3449

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Email address: [email protected]

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Abstract

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Zero-valent iron nanoparticles (nZVI) synthesized in the presence of reduced sulfur compounds

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have been shown to degrade trichloroethene (TCE) at significantly higher rates. However, the

27

applicability of sulfidation as a general means to enhance nZVI reactivity under different particle

28

preparation conditions and the underlying cause for this enhancement effect are not well

29

understood. In this study, the effects of sulfidation reagent, time point of sulfidation, and sulfur

30

loading on the resultant particles were assessed through TCE degradation experiments. Up to 60-

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fold increase in TCE reaction rates was observed upon sulfidation treatment, with products being

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fully dechlorinated hydrocarbons. While the reactivity of these sulfur-treated nZVI (S-nZVI) was

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relatively unaffected by the sulfidation reagent (viz., sodium sulfide, dithionite, or thiosulfate) or

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the sequence of sulfidation relative to iron reduction, TCE reaction rates were found to depend

35

strongly on sulfur to iron ratio. At a low sulfur loading, TCE degradation was accelerated with

36

increasing sulfur dose. The rate constant reached a limiting value, however, as the sulfur to iron

37

mole ratio was greater than 0.025. Different from previous propositions that iron sulfidation

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leads to more efficient TCE or tetrachloroethene (PCE) degradation by enabling depassivation of

39

iron surface, affording catalytic pathways, or facilitating electron transfer, we show that the role

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of sulfur in nZVI lies essentially in its ability to poison hydrogen recombination, which drives

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surface reactions to favor reduction by atomic hydrogen. This implies that the reactivity of S-

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nZVI is contaminant-specific and is selective against the background reaction of water reduction.

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As the effect of sulfur manifests through surface processes, sulfidation represents a broadly

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applicable surface modification approach to modulate or increase the reactivity of nZVI for

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treating TCE and other related contaminants.

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1. Introduction

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Due to widespread historical applications in a broad range of industrial and commercial

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processes and their persistence again natural attenuation [1, 2], trichloroethene (TCE) and

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tetrachloroethene (PCE) in the form of dissolved chemicals or non-aqueous phase liquid (NAPL)

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are the most frequently encountered contaminants at the U.S. superfund sites [3,4, 5]. Compared

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to microbial-mediated reduction of PCE and TCE, which tends to produce toxic intermediates

53

such as dichloroethenes (DCEs) and vinyl chloride (VC), abiotic dechlorination undergoes

54

predominantly a reductive elimination pathway to yield completely dechlorinated products of

55

benign nature (e.g., acetylene, ethene, or ethane) [6,7,8]. With recently reported TCE and PCE

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transformation by reduced iron minerals under field conditions [9,10], there is a surge of interest

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in using indigenous or engineered abiotic materials to improve the remediation performance of

58

sites impacted by chlorinated ethenes. To this end, two broad categories of iron materials have

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been evaluated. Ferrous-containing minerals such as iron sulfides (e.g., mackinawite or pyrite),

60

magnetite, and green rust are able to reduce chlorinated ethenes to acetylene [7,11,12,13,14].

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These Fe(II)-containing minerals are naturally present in subsurface soils, or their formation can

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be stimulated under conditions favorable for biologically mediated sulfate or Fe(III) reduction in

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processes known as in situ biogeochemical treatment [3,9,10]. However, reductive

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dechlorination on ferrous minerals is relatively slow and the minerals appear to possess limited

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reduction capacities [12,15], thus it requires high mass loadings of the solids to outcompete the

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less desirable biological reduction pathways [16]. Another form of iron materials extensively

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studied for the degradation of chlorinated ethenes is zero-valent iron (ZVI) [17-21]. Various

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forms of ZVI, including iron granules and powder, colloidal iron nanoparticles (nZVI), and

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bimetallic iron carrying a small amount of catalyst metal (e.g., Pd-Fe and Ni-Fe) have been 3 ACS Paragon Plus Environment

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studied in the past. In spite of the intrinsic reactivity of ZVI materials, corrosion of iron in

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aqueous solutions causes spontaneous surface passivation and the catalyst additives on ZVI are

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prone to deactivation by common groundwater solutes [22, 23,24,25].

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Recent studies have reported enhanced reactivity of ZVI towards chlorinated contaminants in the

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presence of sulfur compounds [26,27,28,29,30,31]. Hassan [26] observed that iron filings

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containing sulfur impurities were more efficient at TCE degradation than high purity iron. Butler

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and Hayes noted enhanced reduction of chlorinated ethenes when the reaction mixture was

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amended with sulfide ion [11]. By adding sodium dithionite into the synthesis broth of nZVI,

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Kim et al. created a Fe/FeS nanocomposite material with up to 20-fold increases in TCE removal

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rates [27,28]. Similarly, iron nanoparticles that had been conditioned in dilute sulfide or

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dithionite solutions were found to degrade TCE [29,31] and 1,2-dichloroethane [30] more rapidly.

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While multiple explanations have been postulated on the origin of the enhanced reactivity caused

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by iron sulfidation, including a catalytic effect ascribed to the iron sulfides formed on the particle

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surface [26], more efficient charge transfer mediated by the sulfides [27, 29,32], and increased

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depassivation of iron surface [33,34], these views remain largely hypothetical awaiting

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experimental verification. Moreover, variations in experiment conditions, type of iron substrates

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used, and sulfidation procedures in these studies preclude the identification of critical factors

87

controlling the reactivity of sulfur-modified iron. As a result, the broader implications of

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sulfidation as a means to increase the performance of ZVI materials for the treatment of

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chlorinated contaminants are unclear.

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The objective of this study was to examine the effects of sulfidation on the physicochemical

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characteristics of nZVI and their reactivity in TCE dechlorination experiments. Our choice of

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nanosized ZVI stems from their consistent quality and an ability to manipulate particle synthesis 4 ACS Paragon Plus Environment


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conditions to accommodate different sulfidation procedures. While the focus of the present study

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was on nZVI, our ongoing investigations suggest that sulfidation is applicable to other forms of

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iron materials such as commercial ZVI products, thus the findings presented here will lend

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relevant insights into the general role of sulfur in modulating the reactivity of ZVI materials.

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2. Materials and methods

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2.1 Chemicals and Materials

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All chemicals used are listed in the Supporting Information. Deoxygenated deionized-distilled

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water (DDI), prepared by purging DDI with N2 for 30 min, was used in all procedures including

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material synthesis and TCE dechlorination experiments. nZVI was synthesized using the

102

borohydride reduction method [23]. Sulfur-treated nZVI particles (denoted as S-nZVI) were

103

prepared using two approaches. The first approach follows that of Kim et al 27, which involves

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amending an appropriate amount of sulfidation reagent to an Fe(III) solution prior to the addition

105

of borohydride and is referred to as pre-synthesis S-nZVI. Three common sulfur compounds

106

were evaluated as sulfidation reagents in this study, namely, sodium sulfide (Na2S), sodium

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dithionite (Na2S2O4), and sodium thiosulfate (Na2S2O3). The dose of the sulfur compound was

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varied such that the mole ratio of the sulfur reagent to the initial concentration of ferric salt in the

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synthesis mixture (denoted as S/Fe mole ratio) was the range of 1.25 x 10-3 to 0.75. In the

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second approach, a sulfidation reagent was dosed into the synthesis mixture at 20 min after the

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onset of Fe(III) reduction via the addition of borohydride. The resultant particles are denoted as

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post-synthesis S-nZVI. Amorphous iron sulfide (FeS) was synthesized in the lab following the

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method by Butler and Hayes [35]. All iron sulfide or sulfided iron particles were used

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immediately in subsequent experiments upon preparation.


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To investigate the mechanism of sulfur-induced reactivity improvement, a small amount of

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arsenic-modified nZVI was prepared following the same post-synthesis method as that of S-

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nZVI, except that the sulfur compound (sodium thiosulfate) was replaced with equivalent moles

118

of sodium arsenite (NaAsO2).

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2.2. TCE dechlorination experiments

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Batch TCE dechlorination experiments were performed to compare the reactivity of S-nZVI

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prepared under different conditions. All experiments were conducted in 45-mL EPA vials

122

containing 30 mL of aqueous solution and the balance as headspace. The initial pH of all

123

solutions was adjusted to between 7.8-8.2 using dilute NaOH or HCl to simulate the typical pH

124

in groundwater. The solutions were amended with 5 g/L of particles (dry weight). The vials were

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capped with PTFE-lined mininert valves. Experiments were started by injecting a small volume

126

of TCE stock solution in methanol to reach an initial TCE concentration of 25 mg/L. The

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reactors were placed on a wrist-action shaker at 250 rpm at 22 +/- 1 oC. Control experiments

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without iron materials or with unmodified nZVI were performed in parallel. Periodically, an

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aliquot (25 – 50 μL) of headspace gas was withdrawn using a gastight syringe. The samples were

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directly injected into a GC-FID system (Agilent 6890) equipped with an Agilent PoraPlot Q

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column (25 m x 0.32 mm) to analyze for the concentrations of TCE, chlorinated intermediates

132

(not detected in this study), acetylene, ethene, ethane, and longer chain hydrocarbons (up to C6).

133

The analysis conditions are described in SI. This method provides adequate separation between

134

TCE and the daughter products. TCE calibration line was constructed by headspace analysis of

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TCE aqueous standard solutions prepared in the same type of vials as the experimental reactors.

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Calibrations for C2-C6 hydrocarbons were performed using commercial gas standards as

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mentioned in the SI. The results were used to compute their total concentrations in the reaction 6 ACS Paragon Plus Environment


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vials after accounting for partition between headspace and aqueous phases using the respective

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Henry’s Law constants 36, 37 (Table S1). Details on H2 evolution measurements and isotope

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fractionation analysis of TCE during reactions with S-nZVI are available in the SI.

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The solids were subject to microscopic, crystallographic, and surface chemistry characterizations

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and the details are described in the SI.

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3. Results and discussion

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3.1. Effect of sulfidation conditions

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The effect of sulfidation reagent on the reactivity of nZVI for TCE dechlorination was evaluated.

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The particles shown in Figure 1 were prepared using different sulfur compounds at a constant

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S/Fe mole ratio of 0.05 following the post-synthesis sulfidation procedure. As a comparison, the

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inset of Figure 1 shows TCE degradation by fresh nZVI prepared under equivalent conditions but

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without exposure to any sulfur reagent. It is evident that all sulfur-amended nZVI displayed

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remarkable improvements in TCE dechlorination rates. The observed mass-normalized pseudo-

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first-order reaction rate constants (km) of various s-nZVI were approximately 60 folds higher

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than that of the untreated nZVI (Table 1). Distribution of products was qualitatively similar

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among different S-nZVI, with ethene being the dominant product, accompanied by lesser

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amounts of ethane, acetylene, and heavier hydrocarbons (mixture of C3-C6 alkanes and alkenes)

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(Table 1). Dichloroethene isomers (DCEs) and vinyl chloride (VC), common intermediates

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generated by hydrogenolysis reactions, were not detected in nZVI or S-nZVI systems during the

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course of experiments, which agrees with prior studies that reduction of TCE on abiotic surfaces

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occurs predominantly via a dichloro-elemination pathway bypassing the formation of chlorinated

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intermediates [21,38]. 7 ACS Paragon Plus Environment

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In prior studies, dithionite and sulfide ions have been employed to restore the reactivity of

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passivated ZVI [39] or to synthesize Fe(0)/FeS nanocomposite materials [27,29,31]. Aqueous

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sulfide (as H2S or HS- at near neutral pH) is a corrosive chemical and its attaking on iron results

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in deposition of a layer of iron sulfide (FeS) on the surface. Hydrolysis of dithionite in acidic

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solutions gives rise to thiosulfate and sulfite [27, 40] (R1). At an alkaline pH, dithionite may

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hydrolyze via another pathway producing sulfite and sulfide [39, 41] (R2). Disproportionation of

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thiosulfate leads to the formation of elemental sulfur and sulfite (R3) [42]. Elemental sulfur may

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react with iron directly (R4) or convert to sulfide that subsequently binds with iron to form FeS

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[43,44]. As FeS is cathodic to Fe(0), its formation propels further corrosion of Fe(0).

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2S2O42- + H2O  S2O32- + 2HSO3-

(R1)

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S2O42- + 6OH- 5SO32- + S2- + 3H2O

(R2)

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S2O32-  S0 + SO32-

(R3)

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Fe(0) + S0  FeS

(R4)

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In principle, sulfidation of the ZVI material can be achieved with the use of either thiosulfate,

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dithionite, or free sulfide. Dithionite is a fairly strong reductant, especially under alkaline pH

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[39,45], for which it has been proposed as a reductant to prepare nZVI [46]. Thiosulfate does not

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have an as strong reducing capability, but it readily decomposes to release elemental sulfur or

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sulfide (R3), and the former is reduced in the presence of Fe(0) to sulfide [42]. Thus thiosulfate

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effectively serves as a source of sulfide in aqueous nZVI suspension. The sulfide salt used, Na2S,

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is highly hygroscopic and tends to absorb moisture and CO2 in the air, posing material storage

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and handling difficulties. Furthermore, the rapid release of toxic fume upon addition of a sulfide

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chemical raises process safety concerns. In field applications, the above considerations are 8 ACS Paragon Plus Environment


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significant enough to justify the choice of dithionite or thiosulfate. Considering the availability of

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thiosulfate (both as a synthetic chemical and a naturally occuring sulfur compound) and the

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concern that excess dose of dithionite may consume Fe(0) [31], thiosulfate was chosen as the

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sulfidation reagent in all subsequent experiments.

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In the method proposed by Kim et al. [27], dithionite was introduced into the synthesis solutions

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prior to Fe(III) reduction by borohydride. More recently, post-synthesis sulfidation involving

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reacting pre-formed nZVI in sulfide solutions has been employed [29,47]. To assess whether the

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time point of suflidation exerts an effect on the nature of the particles formed, we prepared pre-

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and post-synthesis S-nZVI using thiosulfate at the same S/Fe mole ratio (0.05). The morphology

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of the solids emerging from the two preparations was considerably different. Under TEM, the

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pre-synthesis S-nZVI consists of a heterogeneous mixture of spherical particles that are typical

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of solution-derived iron nanoparticles (indicated by a white arrow) together with some cubic (red

194

arrow) and platy (blue arrow) structures that resemble iron sulfides or oxides (Figure 2b). In

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contrast, the post-synthesis S-nZVI sample in Figure 2c shows more uniform structure

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characterized by spherical particles aggregating in string-like clusters. The appearance was akin

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to that of the unmodified nZVI prepared in our earlier studies [48,49]. The surface chemistry of

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the pre- and post-synthesis S-nZVI was analyzed with X-ray photoelectron spectroscopy (XPS).

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The S 2p spectra of pre- and post-synthesis S-nZVI are shown in Figures 2d and 2e. The spectra

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were fitted with S 2p3/2 and S 2p1/2 spin-orbit doulets that are sperated by 1.2 eV with an intensity

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ratio of 2:1 [50]. Peak assignement was based on literature reported binding energies of sulfide

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minerals [50, 51,52] and the spectra of reference materials acquired under the same conditions as

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the samples. Pre-synthesis S-nZVI carried prodominantly monosulfide (S2-) and disulfide (S22-),

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accounting for 63 and 37 atomic percents (at.%) of total sulfur species, respectively. The surface 9 ACS Paragon Plus Environment

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of post-synthesis S-nZVI consists mainly of S2- (34 at.%) and S22- (46 at.%), with S22-

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contributing a higher portion than that in the pre-synthesis S-nZVI. The post-synthesis sample

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also features a group of low-rising peaks in the binding energy range of 163.3 – 164.3 eV,

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corresponding to polysulfides (Sn2-) and possibly elemental sulfur, and a sulfate (SO42-)

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component at the highest binding energy (167.6 eV). Comparison of the two S 2p spectra

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suggests that oxidation of the sulfur precursor has occcurred to a greater extent during the post-

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synthesis sulfidation process.

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X-ray diffraction analysis (Figure S2) detected the presence of Fe(0) and magnetite in both types

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of S-nZVI, with the post-synthesis S-nZVI exhibiting a lower degree of crystallinity due to

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broadening of diffraction peaks. No mono-, di-, or polysulfides of iron can be discerned in the

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diffraction spectra, which in conjunction with the XPS analysis confirms that sulfide formation

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on the surface is amorphous. This observation is consistent with the notion that rapid corrosion

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of iron in sulfidic water tends to produce poorly ordered iron sulfides [53].

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We noticed that post-synthesis S-nZVI can be efficiently separated from the aqueous phase

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through a filtration step, while the pre-synthesis S-nZVI had a significant portion of the solids

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passing through a 0.2 μM filter. This observation was likely caused by a tendency of post-

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synthesis nZVI to form aggregates and the presence of fine, loose iron sulfide or oxide particles

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in the pre-synthesis S-nZVI as suggested by TEM images.

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In spite of significant structrual differences, the two forms of S-nZVI exhibit similar reactive

224

behavior in TCE dechlorination experiments. km for pre-synthesis and post-synthesis S-nZVI

225

was 0.9 ± 0.1 x 10-3 and 0.8 ± 0.05 x 10-3 L/g-min, respectively, and the product composition

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matches closely with each other (Table 1). In view of the uniform texture of particles prepared



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via the post-synthesis method and their amenability to fast solid/liquid separation, this method

228

was adopted to produce S-nZVI in all subsequent experiments.

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3.2. Effect of sulfur to iron ratio

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A series of S-nZVI were prepared using different doses of thiosulfate such that the S/Fe mole

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ratio in the synthesis solutions varied in the range of 1.25 x 10-3 to 0.75. The concentration of

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thiosulfate after S-nZVI synthesis was measured, and the amount of sulfur deposited on nZVI,

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estimated based on thiosulfate consumption, correlates well with the initial S/Fe mole ratio

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(Table S2 in SI). It was found that the loading of sulfur on iron has a strong impact on TCE

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degradation rates. With particles prepared using the highest sulfur dose (S/Fe = 0.75, Figure 3a),

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ethene and ethane were the dominant products, accounting for 70% and 17%, respectively, of

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total products identified (product yields were determined at approximately 90% TCE conversion).

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Close inspection of Figure 3a indicates there was an accumulation of acetylene during the initial

239

phase of the reaction, nonetheless, its concentration declined over time accompanied by

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concurrent increases in ethene and ethane concentrations. Partially dechlorinated intermediates

241

(i.e., DCEs or VC) were not detected in the headspace mixture. C3-C6 hydrocarbons contributed

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to a minor fraction (11%) of the products formed. In comparison, particles treated with the

243

lowest sulfur dose, corresponding to a S/Fe mole ratio of 1.25 x 10-3, produced ethene, ethane,

244

and C3-C6 products (Figure 3b), and acetyelne was below detection limit at any sampling point.

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Overall, a carbon recovery (as C2 equivalent) of 50% to 90% was achieved for all S-nZVI used

246

in this study. Incomplete carbon recovery has been noticed in prior studies of TCE

247

dechlroination using nZVI [38] or iron sulfide materials [7,11]. As noted in later discussion, the

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missing carbon is likely products of acetylene polymerization reactions, which are affinitive to


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metal surfaces, thus their quantities cannot be reliably measured using the headspace sampling

250

method.

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The results of control experiments, namely, the reactions of TCE with unmodified nZVI and pure

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FeS prepared from aquoues precipitation, were shown in Figure 3c and 3d. Greater than 90%

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TCE degradtion was achieved within a time frame of 0.3 to 2 days when S-nZVI was the

254

reductant, whereas it required 21 days to attain a similar extent of TCE removal by nZVI.

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Degradation of TCE in FeS suspsension was even slower, with only 25% TCE being degraded in

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19 days. The apparent mass-normalized reaction rates of various solids vary by approximately

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three orders of magnitude (Table 1). The composition of product mixture is similar for the

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original nZVI and those receiving a low dose of sulfur (Figures 3b and 3c). At a high sulfur dose,

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considerable accumulation of acetylene during the intermeidate stage of the reaction was

260

observed and ethene was the dominant final product (Figure 3a). Contrary to nZVI or S-nZVI,

261

TCE degradation by FeS yielded exclusively acetylene. This slow transformation of TCE to

262

acetylene by FeS without further hydrogenation of aceytlene agrees with earlier findings by other

263

investigators [7,11].

264

As all TCE degradation data conform to a first-order rate model, the effect of sulfur loading on

265

TCE reduction kinetics was assessed by plotting the mass-normalized rate constant, km, against

266

the S/Fe mole ratio. The results, shown in Figure 4a, reveal a biphasic trend. When thiosulfate

267

was applied at a small dose (S/Fe < 0.025), more rapid TCE dechlorination occurred with

268

increasing S/Fe ratio. However, when the S/Fe mole ratio exceeds 0.025, the rate constant levels

269

out approaching a limiting value with increasing sulfur loading. The highest rate constant was

270

1.3 x 10-3 L/g-min, in comparison to 2.2 x 10-5 and 1.5 x 10-6 L/g-min achieved by unmodified

271

nZVI and FeS, respectively. The effect of S/Fe ratio on product distribution is depicted in Figure 12 ACS Paragon Plus Environment


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4b. The final yields of ethene or ethane, defined as the amount of product formed over TCE

273

consumed determined at the point of 90% TCE conversion, do not bear a strong correlation with

274

the S/Fe ratio. However, the maximum accumulation of acetylene during the course of TCE

275

transformation is strongly affected by the sulfur dose. For particles prepared under a low sulfur

276

loading condition (S/Fe < 0.025), no acetylene was detected in the product mixture, whereas

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particles containing a higher sulfur dose caused a substantial buildup of acetylene before its

278

gradual conversion to downstream products.

279

Acetylene is a reactive chemical and it readily undergoes polymerization reactions on metal

280

surfaces to form longer chain hydrocarbons [54], or in the presence of a hydrogen source,

281

hydrogenates to more saturated products. Our separate experiments reacting acetylene with nZVI

282

and S-nZVI confirm that both solids are able to rapidly transform acetylene into ethene, ethane,

283

and higher order hydrocarbons (Figure S3). Notably, significant gaps in carbon recovery were

284

also observed during these reactions, which was attributed to deposition of non-volatile

285

polymerized products (>C6) on the particle surface. FeS, as expected from previous studies, did

286

not show any appreciable reaction with acetyelene. The composition of TCE daughter products

287

and the reactivity of acetylene towards different iron materials suggest that TCE reduction on

288

nZVI, S-nZVI, or FeS shares an identical pathway of β-elimination leading to acetylene

289

formation [6, 7]. In the presence of nZVI or S-nZVI, acetylene is further converted to ethene,

290

ethane, and higher order hydrocarbons, while it remains intact on FeS. The reaction pathway and

291

its pertinent kinetic parameters are shown schematically in Figure 5. Although multiple steps are

292

invovled in TCE transformation to acetylene, including surface adsorpton of TCE and

293

conversion of chloroacetylene (i.e., the immediate product of TCE β-elimination) to acetylene,

294

the rate of these sequential steps can be captured by a single kinetic parameter (k1) that 13 ACS Paragon Plus Environment

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represents the rate-limiting step. Note that in Figure 5 we consider ethene and ethane as being

296

formed via two parallel pathways instead of going through sequential hydrogenation because

297

ethene hydrogenation by nZVI or S-nZVI was exceedingly slow (Figure S4) and the ratio of

298

ethene to ethane remained constant during TCE degradation. The values of rate constants in

299

Figure 5 were estimated from TCE reduction and acetylene hydrogenation data and are

300

summarized in Table 2. It is intersting to note that sulfiation of nZVI effectively increases the

301

value of k1, but the treatment has no enhancement effect on k2 or k3 value. Thus, the effect of

302

sulfidation is specific for TCE conversion to acetylene and it does not accelerate the subsequent

303

hydrogenation steps.

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3.3. Role of sulfidation in iron reactivity improvement

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A pertinent question then arises on why the incorporation of sulfur into nZVI substrate would

306

cast such a prominent effect on TCE dechlorination. Earlier studies suggest that reduced sulfur

307

compounds such as free sulfide and dithionite are able to depassivate iron surface by reducing

308

Fe(III) to Fe(II) leading to disintegration of the native oxide layer and/or the formation of Fe(II)-

309

containing oxides (e.g., magnetite) that have greater charge transfer abilities [39,53,45].

310

Depassivation effect alone is, however, unable to account for our findings here, since treating

311

nZVI with dilute acid or amending the nZVI suspension with ascobate (a reductant of Fe(III))

312

[55] did not bring about substantial improvements in TCE reduction rates compared to the

313

freshly synthesized particles (Figure S5). This suggets that the presence of sulfided iron is

314

necessary to enable the large increases in TCE degradation rates. It has been proposed that iron

315

sulfide may catalyze PCE or TCE reduction by ZVI [26]. Nevertheless, our analysis of carbon

316

isotope fractionation during TCE experiments did not record consistent shifts in TCE bulk

317

enrichment factors of the unmodified nZVI and its sulfur-treated counterparts (Table S3), nor did 14 ACS Paragon Plus Environment


Page 16 of 33

318

saturating the reaction mixture with H2 bring about more rapid TCE removal, refuting a possible

319

catalytic role played by the sulfide formation.

320

An alternative explanation points to enhanced electron transfer from the Fe(0) core to the

321

solution phase by FeS surface deposits owing to its good electron conducting ability [27,32]. The

322

proposition is supported by the obsevatin of increased anodic currents in recent electrochemical

323

investigations [53,56], however, more rapid iron oxidation may not beget higher rates of TCE

324

transformation since the latter reaction is not limited by electron transfer but the availability of

325

atomic hydrogen [57,58]. Furthermore, the rate enhancement effect caused by accelerated iron

326

corrison is expected to apply to other reducible contaminants, such as carbon tetrachloride (CT,

327

CCl4), whose reduction is governed by a direct electron transfer process [57,59]. To this end, we

328

evaluated reactions of CT with nZVI and sulfided nZVI. The results reveal nearly identical

329

performance by nZVI and those receiving varying levels of sulfur dose (Figure S6). The effect of

330

sulfidation is therefore specific for TCE dechlorination and cannot be ascribed to a general cause

331

related to increased iron corrosion.

332

In the catalysis literature, sulfur is a potent poison of hydrogen recombination reactions on metal

333

surfaces [60,61]. In the case of iron, corrosion in anaerobic water consists of two fundamental

334

processes, namely, the transfer of electrons to protons resulting in surface-adsorbed hydrogen

335

atoms, and the recombination of hydrogen adatoms to form molecular hydrogen that bubbles off

336

the surface. The addition of sulfur on metal surface inhibits hydrogen recombindation and as a

337

result, slows down H2 evolution. This forces more atomic hydrogen to remain on the surface or

338

penetrate into the bulk substrate. Such effect has been investigated extensively and is known to

339

cause hydrogen embrittlement and stress-induced cracking of steel [62,63,64]. Nonetheless,

340

when iron is used as a chemical reductant, inhibition of hydrogen recombination would favor 15 ACS Paragon Plus Environment

Page 17 of 33


341

reactions involving atomic hydrogen, and we believe this effect is the primary mechanism

342

responsible for the remarkable improvements in TCE degradtion rates, since the reduction of

343

chlorinated ethenes on iron are considered to be predominantly mediated by surface adsorbed

344

atomic hydrogen [58,59]. This action of sulfur is consistent with the highly specific effect of

345

sulfidation on TCE reduction relative to that of chlorinated compounds undertaking different

346

reduction mechanisms (e.g., carbon tetrachloride). It also corroborates with recent findings that

347

H2 production was surpressed in the presence of S-nZVI, in contrary to what would be expected

348

if the increase in reactivity is contributed by enhanced iron corrosion [29,31]. In this study,

349

H2 production by S-nZVI was also evaluated under conditions relevant to TCE dechlorination

350

experiments. The results (Figure 6a) clearly demonstrate that sulfur amendment exerts a strong

351

impact on the rate of H2 evolution. The trend of H2 production shown in Figure 6a, that there is a

352

notable decrease in H2 generation rate when S/Fe ratio increases from 0.01 to 0.05 while further

353

increase in sulfur loading does not give rise to significant reduction in H 2 generate rate, agrees

354

with the effect of S dosage on TCE dechlorination kinetics (Figure 4a). An additional argument

355

in support of the poisoning effect of sulfur comes from TCE reduction by nZVI loaded with

356

arsenic (As-nZVI), another potent deactivator of H recombination reactions [63]. The As-nZVI

357

was prepared using the same protocol as that of S-nZVI except that sodium arsenite was

358

employed in place of sodium thiosulfate. The results indicate that arsenic-modified nZVI

359

exhibited similar reactive behavior as the sulfur-modified particles towards TCE (Figure 6b),

360

releasing ethene as the dominant product. Thus, sulfur does not act as a direct facilitator of TCE

361

dechlorination reaction, but rather modifies iron surface chemistry to favor the production of a

362

key reactive species invovled in TCE reduction. When sulfur is deposited on the iron surface, it

363

induces dissolution of the native oxide and causes the surface to be more favorable for atomic



Page 18 of 33

364

hydrogen adsorption. Meantime, inhibtion of hydrogen recombination promotes the build-up of

365

atomic hydrogen on the iron surface and its reaction with TCE. The process is illustrated

366

schematically in Figure 7.

367

Finally, the dependence of the rate constant on S/Fe ratio in Figure 4a suggests the existence of

368

an optimal surface sulfur coverage. Exceeding this optimal loading, there is no further increase in

369

reactivity, possibly due to the formation of polysulfides, discrete FeS precipitates, or oxidized

370

sulfur species that do not contribute as effectively towards reactivity enhancement. The relatively

371

low concentration of sulfur required to achieve this optimal loading implies effective sulfidation

372

can be attained in dilute solutions of sulfur reagents.

373

3.4. Environmental implications

374

ZVI is one of the most frequently applied and environmentally benign reductants for treating a

375

broad vareity of water contaminants. The ZVI chemistry has been studied extensively for

376

chlorinated ethenes. Degradation of TCE by iron alone is relativley slow, and iron reactivity

377

tends to be short-lived due to rapid passivation in air or water. Attemps to improve the

378

performance of ZVI in the past has largely concentrated on a group of bimetallic ZVI particles.

379

The catalyst metals amended on iron surface, such as Pd or Ni, are able to improve the rates of

380

contaminant reduction by catalyzing the activation of H2 [65,66]. The sulfidation method

381

examined here represents a different approach to modify ZVI reactivity. Instead of serving as a

382

catalyst, sulfur poisons a parallel reaction that competes with TCE for the electron source (i.e.,

383

water reduction and H2 evolution), thereby increasing the accessibility of atomic hydrogen for

384

TCE reduction. Although under optimal conditions, the reactivity of these sulfur-modified ZVI

385

may fall short of that of the highly reactive bimetallic materials (for example, km of fresh Pd-


Page 19 of 33


386

nZVI prepared in our lab is approximately 1.4 x 10 -2 L/g-min [25]), iron sulfidation promises

387

some crucial advantages from an enviornmental chemistry perspective. It circumvents the use of

388

catalyst metals that are expensive or toxic to aquatic environment. Inhibition of hydrogen

389

evolution effectively favors the reduction of the target contaminant (e.g., TCE) against the

390

prevalent background reaction between iron and water. The last point will be of interest in large-

391

scale field implementations as material efficiency and longevity in environmental matrices are

392

important considerations in those circumstances. As the effects of sulfur manifest primarily

393

through surface processes, sulfdation can be applied as a surface treatment procedure to pre-

394

synthesized nZVI, and the method should in principle be extendable to other ZVI materials,

395

including bulk iron granules or filing that are frequently used in remediation applications.

396

Further investigations on sulfidation of other forms of ZVI material are underway.

397

Reduced sulfur compounds are ubiquitous in anoxic environment. Partially reduced sulfur anions

398

such as thiosulfate, polysulfides, and sulfite may arise as intermediates during sulfide oxidation

399

and their interconversion is strongly coupled with biogeochemical processes [67,68]. Interstingly,

400

these species are known to be strong inhibitor of catalytic systems including the bimetallic Pd-Fe

401

[25]. From this viewpoint, iron sulfidation not only offers a material that can sustain its reactivity

402

in underground matrix where reduced sulfur ligands are abundant, but also suggests possibilities

403

to optimize the outcome of iron-based remediation technologies via biogeochemical

404

manipulations.

405



406

Acknowlegement

407

The authors acknowledge the start-up fund by TTU and funding support from the National

408

Science Foundation (CHE-1611465). The authors appreciate assistance from Drs. Juske Horita

409

and Kaz Suroweic at TTU for carbon isotope and H2 analyses.

410

Supporting Information

411

Data of XRD characterization, carbon isotope analysis, TCE degradation products by S-nZVI

412

prepared under different synthesis conditions, acetylene and ethene hydrogenation, and carbon

413

tetrachloride reduction by S-nZVI are available in the supporting information. The Supporting

414

Information is available free of charge on the ACS Publications website.

415

Reference

416 417 418 419 420 421 422 423 424 425 426 427 428 429 430 431 432 433 434 435 436 437 438 439

1.

2.

3.

4.

5.

6.

7. 8.

Page 20 of 33

Doherty, R. E., A History of the Production and Use of Carbon Tetrachloride, Tetrachloroethylene, Trichloroethylene and 1,1,1-Trichloroethane in the United States: Part 1--Historical Background; Carbon Tetrachloride and Tetrachloroethylene. Environmental Forensics 2000, 1, (2), 69-81. Doherty, R. E., A History of the Production and Use of Carbon Tetrachloride, Tetrachloroethylene, Trichloroethylene and 1,1,1-Trichloroethane in the United States: Part 2--Trichloroethylene and 1,1,1-Trichloroethane. Environmental Forensics 2000, 1, (2), 83-93. He, Y. T.; Wilson, J. T.; Su, C.; Wilkin, R. T., Review of Abiotic Degradation of Chlorinated Solvents by Reactive Iron Minerals in Aquifers. Ground Water Monitoring and Remediation 2015, 35, (3), 57-75. Stroo, H. F.; Leeson, A.; Marqusee, J. A.; Johnson, P. C.; Ward, C. H.; Kavanaugh, M. C.; Sale, T. C.; Newell, C. J.; Pennell, K. D.; Lebron, C. A.; Unger, M., Chlorinated Ethene Source Remediation: Lessons Learned. Environmental Science & Technology 2012, 46, (12), 6438-6447. Stroo, H. F.; Unger, M.; Ward, C. H.; Kavanaugh, M. C.; Vogel, C.; Leeson, A.; Marqusee, J. A.; Smith, B. P., Remediating chlorinated solvent source zones. Environmental Science & Technology 2003, 37, (11), 224A-230A. Roberts, A. L.; Totten, L. A.; Arnold, W. A.; Burris, D. R.; Campbell, T. J., Reductive elimination of chlorinated ethylenes by zero valent metals. Environmental Science & Technology 1996, 30, (8), 2654-2659. Butler, E. C.; Hayes, K. F., Kinetics of the transformation of trichloroethylene and tetrachloroethylene by iron sulfide. Environmental Science & Technology 1999, 33, (12), 2021-2027. Lojkasek-Lima, P.; Aravena, R.; Shouakar-Stash, O.; Frape, S. K.; Marchesi, M.; Fiorenza, S.; Vogan, J., Evaluating TCE Abiotic and Biotic Degradation Pathways in a Permeable Reactive Barrier Using Compound Specific Isotope Analysis. Ground Water Monitoring and Remediation 2012, 32, (4), 53-62.


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29. Rajajayavel, S. R. C.; Ghoshal, S., Enhanced reductive dechlorination of trichloroethylene by sulfidated nanoscale zerovalent iron. Water Research 2015, 78, 144-153. 30. Garcia, A. N.; Boparai, H. K.; O'Carroll, D. M., Enhanced Dechlorination of 1,2-Dichloroethane by Coupled Nano Iron-Dithionite Treatment. Environmental Science & Technology 2016, 50, (10), 5243-5251. 31. Fan, D.; Johnson, G. O.; Tratnyek, P. G.; Johnson, R. L., Sulfidation of Nano Zerovalent Iron (nZVI) for Improved Selectivity during In- Situ Chemical Reduction (ISCR). Environmental Science & Technology 2016, (50), 9558-9565. 32. Turcio-Ortega, D.; Fan, D.; Tratnyek, P. G.; Kim, E.-J.; Chang, Y.-S., Reactivity of Fe/FeS Nanoparticles: Electrolyte Composition Effects on Corrosion Electrochemistry. Environmental Science & Technology 2012, 46, (22), 12484-12492. 33. Lipczynskakochany, E.; Harms, S.; Milburn, R.; Sprah, G.; Nadarajah, N., Degradation of carbon tetrachloride in the presence of iron and sulfur-containing compounds Chemosphere 1994, 29, (7), 1477-1489. 34. Hansson, E. B.; Odziemkowski, M. S.; Gillham, R. W., Influence of Na2S on the degradation kinetics of CCl4 in the presence of very pure iron. Journal of Contaminant Hydrology 2008, 98, (34), 128-134. 35. Butler, E. C.; Hayes, K. F., Effects of Solution Composition and pH on the Reductive Dechlorination of Hexachloroethane by Iron Sulfide. Environmental Science & Technology 1998, 32, (9), 12761284. 36. Yaws, C. L., Yaws' Handbook of Thermodynamic and Physical Properties of Chemical Compounds. In Knovel. 37. Williams, M. L., CRC Handbook of Chemistry and Physics, 76th edition. Occupational and Environmental Medicine 1996, 53, (7), 504-504. 38. Arnold, W. A.; Roberts, A. L., Pathways and kinetics of chlorinated ethylene and chlorinated acetylene reaction with Fe(O) particles. Environmental Science & Technology 2000, 34, (9), 17941805. 39. Xie, Y.; Cwiertny, D. M., Use of Dithionite to Extend the Reactive Lifetime of Nanoscale ZeroValent Iron Treatment Systems. Environmental Science & Technology 2010, 44, (22), 8649-8655. 40. Lister, M. W.; Garvie, R. C., Sodium dithionite, decomposition in aqueous solution and in the solid state. Canadian Journal of Chemistry-Revue Canadienne De Chimie 1959, 37, (9), 1567-1574. 41. Greenwood, N. N.; Earnshaw, A., Chemistry of the Elements. 2nd ed.; Elsevier Butterworth Heinemann 1997. 42. Kappes, M.; Frankel, G. S.; Sridhar, N.; Carranza, R. M., Reaction Paths of Thiosulfate during Corrosion of Carbon Steel in Acidified Brines. Journal of the Electrochemical Society 2012, 159, (4), C195-C204. 43. Macdonald, D. D.; Roberts, B.; Hyne, J. B., Corrosion of carbon-steel by wet elemental sulfur Corrosion Science 1978, 18, (5), 411-425. 44. Schmitt, G., Effect of elemental sulfur on corrosion in sour gas systems Corrosion 1991, 47, (4), 285-308. 45. Szecsody, J. E.; Fruchter, J. S.; Williams, M. D.; Vermeul, V. R.; Sklarew, D., In situ chemical reduction of aquifer sediments: Enhancement of reactive iron phases and TCE dechlorination. Environmental Science & Technology 2004, 38, (17), 4656-4663. 46. Sun, Q.; Feitz, A. J.; Guan, J.; Waite, T. D., Comparison of the reactivity of nanosized zero-valent iron (nZVI) particles produced by borohydride and dithionite reduction of iron salts Nano 2008, 3, (5), 341-349. 47. Fan, D.; Anitori, R. P.; Tebo, B. M.; Tratnyek, P. G.; Pacheco, J. S. L.; Kukkadapu, R. K.; Engelhard, M. H.; Bowden, M. E.; Kovarik, L.; Arey, B. W., Reductive Sequestration of Pertechnetate ((TcO4)-Tc-99) by Nano Zerovalent Iron (nZVI) Transformed by Abiotic Sulfide. Environmental Science & Technology 2013, 47, (10), 5302-5310.


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48. Martin, J. E.; Herzing, A. A.; Yan, W. L.; Li, X. Q.; Koel, B. E.; Kiely, C. J.; Zhang, W. X., Determination of the oxide layer thickness in core-shell zerovalent iron nanoparticles. Langmuir 2008, 24, (8), 4329-4334. 49. Yan, W.; Herzing, A. A.; Kiely, C. J.; Zhang, W.-x., Nanoscale zero-valent iron (nZVI): Aspects of the core-shell structure and reactions with inorganic species in water. Journal of Contaminant Hydrology 2010, 118, (3-4), 96-104. 50. Smart, R. S.; Skinner, W. M.; Gerson, A. R., XPS of sulphide mineral surfaces: Metal-deficient, polysulphides, defects and elemental sulphur. Surface and Interface Analysis 1999, 28, (1), 101-105. 51. Herbert, R. B.; Benner, S. G.; Pratt, A. R.; Blowes, D. W., Surface chemistry and morphology of poorly crystalline iron sulfides precipitated in media containing sulfate-reducing bacteria. Chemical Geology 1998, 144, (1-2), 87-97. 52. Mullet, M.; Boursiquot, S.; Abdelmoula, M.; Genin, J. M.; Ehrhardt, J. J., Surface chemistry and structural properties of mackinawite prepared by reaction of sulfide ions with metallic iron. Geochimica Et Cosmochimica Acta 2002, 66, (5), 829-836. 53. Hansson, E. B.; Odziemkowski, M. S.; Gillham, R. W., Formation of poorly crystalline iron monosulfides: Surface redox reactions on high purity iron, spectroelectrochemical studies. Corrosion Science 2006, 48, (11), 3767-3783. 54. Egloff, G.; Lowry, C. D.; Schaad, R. E., Polymerization and decomposition of acetylene hydrocarbons. J. Phys. Chem. 1932, 36, (5), 1457-1520. 55. Suter, D.; Banwart, S.; Stumm, W., Dissolution of hydrous iron(III) oxides by reductive mechanisms Langmuir 1991, 7, (4), 809-813. 56. Kim, E.-J.; Kim, J.-H.; Chang, Y.-S.; Turcio-Ortega, D.; Tratnyek, P. G., Effects of Metal Ions on the Reactivity and Corrosion Electrochemistry of Fe/FeS Nanoparticles. Environmental Science & Technology 2014, 48, (7), 4002-4011. 57. Li, T.; Farrell, J., Reductive dechlorination of trichloroethene and carbon tetrachloride using iron and palladized-iron cathodes. Environmental Science & Technology 2000, 34, (1), 173-179. 58. Elsner, M.; Hofstetter, T. B., Current Perspectives on the Mechanisms of Chlorohydrocarbon Degradation in Subsurface Environments: Insight from Kinetics, Product Formation, Probe Molecules, and Isotope Fractionation. In Aquatic Redox Chemistry, Tratnyek, P. G.; Grundl, T. J.; Haderlein, S. B., Eds. 2011; Vol. 1071, pp 407-439. 59. Li, T.; Farrell, J., Electrochemical investigation of the rate-limiting mechanisms for trichloroethylene and carbon tetrachloride reduction at iron surfaces. Environmental Science & Technology 2001, 35, (17), 3560-3565. 60. Oudar, J., Sulfur adsorption and poisoning of metallic catalysts Catalysis Reviews-Science and Engineering 1980, 22, (2), 171-195. 61. Burke, M. L.; Madix, R. J., Hydrogen on Pd(100)-S The effect of sulfur on precursor mediated adsorption and desorption. Surface Science 1990, 237, (1-3), 1-19. 62. Radhakrishnan, T.; Shreir, L., Permeation of hydrogen through steel by electrochemical transfer-I. Influence of catalytic poisons. Electrochimica Acta 1966, 11, 1007-1021. 63. Berkowitz, B. J.; Horowitz, H. H., The role of H2S in the corrosion and hydrogen embrittlement of steel Journal of the Electrochemical Society 1982, 129, (3), 468-474. 64. Berkowitz, B. J.; Heubaum, F. H., The role of hydrogen in sulfide stress cracking of low-alloy steels Corrosion 1984, 40, (5), 240-245. 65. Alonso, F.; Beletskaya, I. P.; Yus, M., Metal-mediated reductive hydrodehalogenation of organic halides. Chemical Reviews 2002, 102, (11), 4009-4091. 66. Urbano, F. J.; Marinas, J. M., Hydrogenolysis of organohalogen compounds over palladium supported catalysts. Journal of Molecular Catalysis a-Chemical 2001, 173, (1-2), 329-345. 67. Luther, G. W.; Church, T. M.; Scudlark, J. R.; Cosman, M., Inorganic and organic sulfur cycling in salt-marsh pore waters. Science 1986, 232, (4751), 746-749.



590 591 592

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68. Cantrell, K. J.; Yabusaki, S. B.; Engelhard, M. H.; Mitroshkov, A. V.; Thornton, E. C., Oxidation of H2S by iron oxides in unsaturated conditions. Environmental Science & Technology 2003, 37, (10), 2192-2199.

593


Page 25 of 33


594 595 596 597 598

Figure 1. TCE dechlorination by S-nZVI prepared using different sulfidation reagents. The initial mole ratio of sulfidation reagent to iron was fixed at 0.05. Initial TCE concentration was 25 mg/L. Inset shows TCE degradation by fresh nZVI without sulfidation treatment. The particle dose was 5 g/L in all experiments.

599 600 601 602 603 604 605 606 607 608 609



(a)

0.8

(b) Pre-synthesis S-nZVI

Pre-synthesis S-nZVI Post-synthesis S-nZVI

0.6

C/C0

(d) CPS (a.u.)

1.0

Page 26 of 33

S22- S2-

Pre-synthesis S-nZVI

100 nm

(e)

0.4

(c) Post-synthesis S-nZVI

0.2

S22-

Post-synthesis S-nZVI SO42-

Sn2-

S2-

0.0

0

200 Time (min)

400 50 nm

169 167 165 163 161 159 Binding Energy (eV)

610 611 612 613 614

Figure 2. (a) TCE dechlorination by S-nZVI that receives sulfidation treatment at different stages of particle synthesis. (b) and (c) TEM micrographs of the particles used in (a). (d) and (e) XPS S 2p3/2 spectra of the corresponding particles in (a). All materials were prepared with a S/Fe mole ratio of 0.05.

615 616 617 618 619 620


Page 27 of 33


621 622 623 624

Figure 3. TCE dechlorination and product formation in reactor of (a) S-nZVI with low sulfur dose, (b) S-nZVI with high sulfur dose, and (c) unmodified nZVI, and (d) FeS. The particle dose was 5 g/L.

625 626



Page 28 of 33

km (10-3 L/g-min)

(a) 1

0.1 nZVI 0.01 0.0001

0.01

1

S/Fe mole ratio 0.7

Product Yield

0.6 0.5

Acetylene (max yield) Ethene (final yield) ethane (final yield)

(b)

0.4 0.3 0.2 0.1

0.0 0.0001 627 628 629

0.01

1

S/Fe mole ratio Figure 4. Effect of S/Fe mole ratio on (a) TCE degradation rate and (b) product yields. The particle dose was 5 g/L.

630


Page 29 of 33


Cl

H C

C

Cl

Cl

k1 H

k2 H H 631 632 633 634

C

Polymerization products

H

k3 H

C

C

C H

H H H C C H H H

Figure 5. Proposed reaction pathways of TCE decomposition on S-nZVI. Only experimentally observed intermediates or products are shown. Dashed line indicates possible involvement of multiple reaction steps. Values of reaction rate constants (k1 – k3) are tabulated in Table 2.

635



H2 production, mmole

0.6

nZVI

0.5

Page 30 of 33

(a)

S-nZVI (S/Fe=0.01) S-nZVI (S/Fe=0.05)

0.4

S-nZVI (S/Fe=0.25)

0.3 0.2 0.1 0.0

0

24

48

72

Time, h

636

1.0

(b)

S-nZVI (S/Fe = 0.05)

As-nZVI (As/Fe = 0.05)

C/Co

0.8 0.6 0.4 0.2 0.0

0

200 Time (min)

400

637 638 639 640

Figure 6. (a) H2 production by nZVI and S-nZVI of varying S/Fe mole ratio. (b) TCE degradation by S-nZVI and arsenic-modified nZVI (As-nZVI) at an As or S to Fe mole ratio of 0.05. Particle dose was 5 g/L in all experiments.

641


Page 31 of 33


Unmodified nZVI H+

H• H•

H•

e-

Fe(0)

H2

H+

C2HCl3 C2H2

H+

H• H•

e-

Sulfur-treated nZVI H+

C2HCl3 C2H2

H• H• H• H• H• Fe Sulfide e-

Fe(0)

H+

H2

H+

slow

H• H•

e-

642 643 644

Figure 7. Schematics of reactions on nZVI and S-nZVI.

645



646 647

Table 1. Pseudo-first-order rate constants and product distribution of TCE dechlorination by nZVI and S-nZVI prepared under different conditions Particle type FeS

nZVI

S-nZVI (thiosulfate)

S-nZVI (dithionite)

S-nZVI (sulfide)

S-nZVI (thiosulfate)

648 649 650 651 652

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Sulfidation condition

Products

Yielda

Carbon recovery

km (10-3 L g-1 min-1 )c

(%)b Ethene Ethane Acetylene

N.D. N.D. 126

C3-C6

N.D.

Ethene Ethane

17 6.5 N.D. 5.7 44 10 4.5 7.2 33 8.2 10 9.6 39 9.4 7.2 7.5 45 10 1.3 8.1

Acetylene C3-C6 Ethene Post-synthesis sulfidation, S/Fe Ethane Acetylene = 0.05 C3-C6 Post-synthesis Ethene sulfidation, S/Fe Ethane = 0.05 Acetylene C3-C6 Post-synthesis Ethene sulfidation, S/Fe Ethane = 0.05 Acetylene C3-C6 Ethene Pre-synthesis sulfidation, S/Fe Ethane Acetylene = 0.05 C3-C6

107

0.0024 ± 0.0007

36

0.023± 0.008

68

0.80 ± 0.05

68

0.78 ± 0.21

67

1.04 ± 0.13

70

0.90 ± 0.08

a

calculated from product formation over TCE consumption. b sum of products (as C2 equivalents) and TCE remain. Both a and b were determined at the point of ca. 90% TCE conversion or, for slow reactions, the last sampling point. c mass-normalized pseudo-first-order rate constants, uncertainties represent 95% confidence intervals.

653


Page 33 of 33

654 655


Table 2. Mass-normalized rate constants of steps involved in TCE degradation by S-nZVI of varying S/Fe mole ratio S/Fe mole ratio

656

k1 (10-3 L g-1 min-1)a

k2 (10-3 L g-1 min-1)b

k3(10-3 L g-1 min-1)b

0.00125 0.18 1.03 0.22 0.05 1.16 0.68 0.15 0.25 1.11 0.60 0.11 a obtained from TCE degradation experiments. b estimated from acetylene hydrogenation experiments.

657


Reductive Dechlorination of Trichloroethene by Zero-valent Iron

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