Dechlorination of Excess Trichloroethene by Bimetallic and Sulfidated

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Dechlorination of Excess Trichloroethene by Bimetallic and Sulfidated Nanoscale Zero-Valent Iron Feng He, Zhenjie Li, Shasha Shi, Wenqiang Xu, Hanzhen Sheng, Yawei Gu, Yonghai Jiang, and Beidou Xi Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b01735 • Publication Date (Web): 28 Jun 2018 Downloaded from http://pubs.acs.org on June 29, 2018

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Environmental Science & Technology

1

Dechlorination of Excess Trichloroethene by Bimetallic and

2

Sulfidated Nanoscale Zero-Valent Iron

3

Feng He1, 2*, Zhenjie Li1, Shasha Shi1, Wenqiang Xu1, Hanzhen Sheng1,

4

Yawei Gu1, Yonghai Jiang3, 4, Beidou Xi3, 4 1

5

College of Environment, Zhejiang University of Technology, Hangzhou 310014, China

6 7

2

Key Laboratory of Microbial Technology for Industrial Pollution Control of Zhejiang Province, Zhejiang University of Technology, Hangzhou 310014, China

8 9

3

State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese Research Academy of Environmental Sciences, Beijing 100012, China

10 11

4

State Environmental Protection Key Laboratory of Simulation and Control of

12

Groundwater Pollution, Chinese Research Academy of Environmental Sciences,

13

Beijing 100012, China

14

*Corresponding author: Feng He

15

Email: [email protected], Phone: 86-571-88871509

16

Keywords: TCE, Dechlorination, Sulfidation, Bimetallic, Catalyst,

17

Electron Efficiency, Selectivity, Zero-Valent Iron, Nanoparticles

18

06/28/2018

19

1 ACS Paragon Plus Environment


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Graphical Abstract

21

22 23


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Abstract

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Nanoscale zero valent iron (nZVI) likely finds its application in source zone

26

remediation. Two approaches to modify nZVI have been reported: bimetal (Fe-Me)

27

and sulfidated nZVI (S-nZVI). However, previous researches have primarily focused

28

on enhancing particle reactivity with these two modifications under more plume-like

29

conditions. In this study, we systematically compared the TCE dechlorination

30

pathway, rate and electron selectivity of Fe-Me (Me: Pd, Ni, Cu and Ag), S-nZVI, and

31

nZVI with excess TCE simulating source zone conditions. TCE dechlorination on

32

Fe-Me was primarily via hydrogenolysis while that on S-nZVI and nZVI was mainly

33

via β-elimination. The surface-area normalized TCE reduction rate (k’SA) of Fe-Pd,

34

S-nZVI, Fe-Ni, Fe-Cu, and Fe-Ag were ~ 6,800, 190, 130, 20, and 8-fold greater than

35

nZVI. All bimetallic modification enhanced the competing hydrogen evolution

36

reaction (HER) while sulfidation inhibited HER. Fe-Cu and Fe-Ag negligibly

37

enhanced electron utilization efficiency (εe) while Fe-Pd, Fe-Ni and S-nZVI

38

dramatically increased εe from 2% to ~ 100%, 69% and 72%, respectively. Adsorbed

39

atomic hydrogen was identified to be responsible for the TCE dechlorination on

40

Fe-Me but not on S-nZVI. The enhanced dechlorination rate along with the reduced

41

HER of S-nZVI can be explained by that FeS conducting major electrons mediated

42

TCE dechlorination while Fe oxides conducting minor electrons mediated HER.



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43

Introduction

44

In the last two decades, nanoscale zero-valent iron (nZVI) has been demonstrated to

45

be an effective and promising alternative for in-situ remediation of contaminated

46

groundwater and soil due to its strong reductive capability (Fe2+ + 2e- → Fe(s), E0 =

47

-0.44 V).1-7 Furthermore, the nanoscale particle size allows nZVI to be injected into

48

the source zones (after modified with polymers or surfactants) and attack chlorinated

49

compounds or other pollutants before they spread out in the aquifer, which is crucial

50

for in situ remediation.5,

51

considerable large surface area compared with microscale or millimeter ZVI that are

52

typically

53

implementation of nZVI still faces challenges.6, 21 Particularly, the reactivity of bare

54

nZVI may be still unsatisfactory for source zone or “hot spot” remediation.21, 22 In

55

addition, nZVI injected into groundwater will be easily passivated by an iron oxide or

56

hydroxide film produced during iron corrosion process,19, 20, 23-25 thereby losing its

57

reactivity merit.

used

in

8-17

nZVI also showed enhanced reactivity due to its

permeable

reactive

barriers

(PRBs).18-20

However,

the

58

To date, a variety of approaches to enhance the reactivity of nZVI for pollutant

59

degradation have been explored.5, 26-31 The most used approach involves addition of a

60

secondary metal to form iron-based bimetallic nanoparticles especially Fe-Pd,22, 32-38

61

Fe-Ni,39-43 Fe-Cu,44,

62

nanoparticles especially Fe-Pd degraded organic halides more completely into

63

halogen-free products with reaction rates several orders of magnitude higher than bare

64

nZVI under anoxic and excess iron conditions.29,

45

and Fe-Ag.46,

47

Researches showed that bimetallic

32, 48


The enhanced reactivity of

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65

bimetallic particles was generally attributed to (1) the deposition of transition or noble

66

metals (e.g., Pd, Ni, Cu, and Ag) formed galvanic couple with nZVI, which

67

accelerates the electron transfer from Fe(0) core to metal additive on nZVI surface,39,

68

44, 49

and (2) the metal additive catalyzes hydrodechlorination reaction.37, 39

69

It should be noted that although numerous works have been done on

70

dechlorination of chlorinated compounds by different types of bimetallic particles, the

71

reaction pathway and mechanism remain elusive and sometimes contradictory.37-39,

72

49-51

73

proposed to control the hydrodechlorination reactivity of bimetallic ZVI particles

74

based on a nice correlation between the solubility of atomic hydrogen within each

75

additive and 1,1,1-TCA reduction rates.31 However, recent studies on electrocatalytic

76

hydrodechlorination of 2,4-dichlorophenol on Pd nanoparticles suggested that

77

adsorbed atomic hydrogen was the active species while absorbed atomic hydrogen

78

was inert.51 Besides, bimetallic modification accelerates electron transfer from Fe(0)

79

to not only contaminants but also H2O/H+ to promote molecular hydrogen evolution

80

reaction (HER) and would shorten the lifetime of bimetallic particles.21, 52, 53 The

81

selectivity of electrons from Fe(0) for contaminants over H2O/H+ can be defined as

82

efficiency of electron utilization, which has been growingly recognized as the most

83

challenging aspect of ZVI performance.21, 53-55 However, the HER rates along with the

84

efficiency of electron utilization (as well as dechlorination capacity) for iron-based

85

bimetallic particles have not been investigated/quantified.

For example, atomic hydrogen absorbed in the lattice of metal additives was



86

In recent years, another modification of nZVI, sulfidation of nZVI to form

87

secondary iron sulfide phases, has gained significant attention.21, 28, 56-65 The sulfidated

88

nZVI (S-nZVI) can not only enhance the rates of dechlorination but also improve the

89

electron selectivity of nZVI for contaminant reduction.28, 57, 61, 65, 66 These benefits

90

have been observed with S-nZVI produced through different approaches, for example

91

the so-called “one-pot” S-nZVI produced by reduction of ferrous iron with

92

borohydride in the presence of dithionite56 and the “two-step” S-nZVI produced by

93

adding the sulfidation agents (e.g., sodium sulfide and thiosulfate) after nZVI

94

synthesis.57, 60 After sulfidation, the rate of TCE dechlorination by nZVI increased by

95

~ 60-fold while H2 production was prohibited.60 However, the electron utilization

96

efficiency of nZVI was scarcely quantified except for that of carboxymethyl cellulose

97

stabilized S-nZVI, which was determined to be more than 95%.61 In addition,

98

although a few studies have explored the mechanism of the concurrent enhancement

99

of contaminant reduction and selectivity of S-nZVI,56, 57, 60 it remains elusive. One

100

explanation was that the iron sulfide on nZVI surface is more hydrophobic (i.e., it has

101

higher binding ability for organic contaminants than H2O/H+)56 and a more efficient

102

electron conductor than iron (hydr)oxides on bare nZVI.53, 57, 65, 67 However, some

103

researchers believe that sulfur in S-nZVI poisoned atomic hydrogen recombination,

104

which favored the contaminant reduction by atomic hydrogen.60 The pre-hypothesis of

105

this explanation is that atomic hydrogen mediates TCE dechlorination on S-nZVI,

106

which may not be applied.


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107

As both the bimetallic and sulfidation modifications are being or to be practiced

108

in field nZVI implementation, it is important to compare these two approaches

109

systematically under same conditions, which has not been done so far. In addition,

110

nZVI has relatively high cost and is likely to be used in source zone remediation (i.e.,

111

excess TCE conditions). Such scenario was not typically considered in previous

112

studies. The overall objective of this study was to comprehensively compare not only

113

the reaction rate and pathway of TCE dechlorination by bimetallic and sulfidated

114

nZVI but also the selectivity and dechlorination capacity of both particles under

115

source zone conditions. The mechanisms of how bimetallic and sulfidated

116

modifications cause the change of dechlorination rate and selectivity were also

117

explored and compared. To the best of our knowledge, this represents one of the first

118

trials of looking at all typical nZVI modifications together. The “excess TCE” used in

119

this study also facilitated the investigation of reaction pathway as they competed

120

reaction sites and allowed the accumulation of short-life reaction intermediates.

121

Materials and Methods

122

Chemicals. Details of the used chemicals are provided in Supporting Information

123

(SI).

124

Particle Preparation and Characterization. The nZVI particles were synthesized by

125

borohydride method adapted from Liu et al.19 The iron-based bimetallic and sulfidated

126

nanoparticles were synthesized by adding salts solution of each additive or Na2S into

127

nZVI suspension and allowed for 15-min reaction under sonication and anoxic



128

conditions. The metal additive/Fe and S/Fe molar ratio was typically 0.27 mol% and

129

20 mol%, respectively. The use of high S/Fe ratio in S-nZVI was due to that S is

130

environmentally safe and such ratio is likely to be used in practical application to

131

ensure high reactivity of S-nZVI.60 More details regarding the particle synthesis is

132

provided in SI.

133

The physical properties of nZVI, bimetals (Fe-Me), and S-nZVI particle were

134

characterized using a set of crystallographic, microscopic, and surface chemistry

135

methods including X-ray diffraction (XRD), scanning electron microscopy (SEM),

136

high resolution transmission electron microscopy (HR-TEM) with selected area

137

electron diffraction (SAED) and X-ray photoelectron spectroscopy (XPS).

138

Reaction Systems. TCE dechlorination was tested under excess TCE conditions

139

(TCE: 2.28 mM, particle concentration: 0.25 g/L) unless otherwise mentioned to

140

represent low iron to TCE ratios that might arise during treatment of a source zone.

141

Carbon balance of the TCE dechlorination experiments ranged from 80% to 105%.

142

Details regarding the batch experiments and chemical analyses for TCE degradation

143

and H2 generation are provided in SI. TCE dechlorination by carboxymethyl cellulose

144

stabilized Pd (CMC-Pd) nanoparticles in the presence of H2 at varied concentrations

145

was also performed for mechanism study purpose and the details are provided in SI.

146

Particle Efficiencies. Two types of particle efficiencies for the anoxic ZVI-TCE-H2O

147

system, defined in our previous work,53 were quantified. One is Fe(0) efficiency

148

(εFe(0)), which is the molar fraction of Fe(0) accessible to TCE or H2O. The other is the

149

efficiency of electron utilization (εe), which is the fraction of electron equivalents 8 ACS Paragon Plus Environment

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from Fe(0) that are used for TCE reduction. Details regarding their calculation are

151

provided in SI.

152

Results and Discussion

153

Particle Characterization. SEM images (Figure S1A, B) show spherical shape for

154

both nZVI and S-nZVI particles. HR-TEM images (Figure S2) further demonstrate

155

the core-shell structure of the spherical nZVI, Fe-Me, and S-nZVI with shell thickness

156

of 3~6 nm. The absence of lattice fringes and diffuse rings in SAED patterns of nZVI

157

(Figure 1A) suggests that nZVI particles were highly disordered, while the clear spots

158

and sharp rings in SAED patterns of S-nZVI (Figure 1A) indicate multiple

159

crystallinity of S-nZVI particles. The crystallinity of nZVI and S-nZVI was further

160

demonstrated by the XRD patterns (Figure 1A). No diffraction peaks of iron oxides

161

and only broad peak of α-Fe0 were detected for nZVI, suggesting its poorly ordered

162

crystallinity, while the presence of mackinawite (FeS), greigite (Fe3S4), and pyrrhotite

163

(Fe7S8) peaks in S-nZVI conformed to the multiple crystallinity of S-nZVI particles.

164

For Fe-Me, bimetallic modification did not change the crystallinity nature of nZVI

165

and second metals (i.e., Pd, Ni, Cu, and Ag) were not detected by XRD (also not

166

observed by TEM) due to their low loading (Figure S3).

167

The high resolution Fe2p XPS spectra for nZVI and S-nZVI are shown in Figure

168

S4A and S4B, respectively. Fitting these peaks gave the distribution of iron oxidation

169

states summarized in Figure 1B. The results show that Fe0 was only present on the

170

surface of S-nZVI, suggesting that sodium sulfide can etch the iron (hydr)oxide film



171

on nZVI to form iron sulfides and expose the Fe0 core that was otherwise completely

172

covered in pristine nZVI. This is also supported in this case by the decrease of oxygen

173

content in S-nZVI compared to nZVI (Figure 1C). The S 2p spectrum of S-nZVI in

174

Figure 1D fitted with doublets representing 2p1/2 and 2p3/2 suggests that the surface of

175

S-nZVI consisted predominately S2- (56%) and Sn2- (44%), which is consistent with

176

previous studies56,

177

characterization of S-nZVI. For Fe-Me, bimetallic modification did not cause obvious

178

change of nZVI surface in terms of Fe0 exposure and Fe(II)/Fe(III) ratio (Figure S5A).

179

The detected Me/Fe molar ratio on the surface was ~ 4.5 mol% for all four bimetals

180

(Figure S5B) and Me valences were zero (data not shown), which indicates that Me

181

deposited on nZVI surface in elemental form for all Fe-Me. The BET surface area of

182

nZVI was 16 m2/g. Bimetallic modification and sulfidation increased the BET surface

183

area to 20±1 and 23 m2/g (Table S2), respectively, likely due to the formation of more

184

dispersed second metal or Fe sulfide phases on the surface. Lines

57, 60

and also validates the results from SAED and XRD


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185 186 187 188 189 190 191 192

Figure 1. (A) XRD patterns of nZVI and S-nZVI (S/Fe = 20 mol%). The insets are the corresponding SAED images; (B) Molar fraction of Fe(0), Fe(II), and Fe(III) in nZVI and S-nZVI derived from fitting of XPS Fe 2p spectra; (C) Molar fraction of Fe, O, and S content in nZVI and S-nZVI. There was about 10% of B in the particles; (D) S 2p XPS spectra of S-nZVI. The particles were aged by naturally diffused air for 12 h after dried in argon before characterization.

193

Reaction Products and Pathways. The formation and distribution of major products

194

at the end of TCE reduction (8 days) by Fe-Me, S-nZVI, and nZVI are summarized in

195

Figure 2A. Chlorinated intermediates (particularly 1,1-DCE and cis-DCE) were

196

observed with all Fe-Me particles while they were almost absent in nZVI and S-nZVI

197

systems. In contrast, significant amount of acetylene was observed with nZVI (20% of

198

the final products) and S-nZVI (70% of the final products) while it was absent in all

199

Fe-Me systems. It suggests that hydrogenolysis was the major dechlorination pathway

200

of TCE on Fe-Me while β-elimination was the dominant TCE dechlorination pathway 11 ACS Paragon Plus Environment


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on nZVI and S-nZVI. The detection of ethane and coupling products in all iron

202

systems further suggests that hydrogenation and coupling reactions also occurred. The

203

more significant accumulation of chlorinated intermediates with Fe-Cu and Fe-Ag

204

than Fe-Pd and Fe-Ni is a result of their lower dechlorination reactivity as shown in

205

Figure 3. The larger accumulation of acetylene with S-nZVI than nZVI suggests that

206

the particle reactivity of S-nZVI is controlled by sulfidated iron phases, consistent

207

with our recent report on TCE reduction by mechanochemically sulfidated

208

micro-scale ZVI.53

A

as % of products

100 80 60

CPs C 2H 2

40

C 2H 4 C 2H 6

20

C4,C6 C3,C5

0 Fe-Pd Fe-Ni Fe-Cu Fe-Ag S-nZVI nZVI

209

210 211 212 213 214 215

Figure 2. Product distribution at the end of TCE dechlorination (8 days) by Fe-Me, S-nZVI, and nZVI (CPs represents chlorinated products) (A) and proposed reaction pathways of TCE dechlorination on Fe-Pd or Fe-Ni (B) and S-nZVI (C). Only stable reaction intermediates are shown. Numbers correspond to rate constants (k1 to k4) in Table S3.


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The final steps of the TCE dechlorination was further characterized with batch

217

experiments by either adding ethene to Fe-Me or adding acetylene and ethene to nZVI

218

and S-nZVI. Both Fe-Pd and Fe-Ni rapidly and completely transformed ethene to

219

ethane (Figure S6A, B) while Fe-Cu and Fe-Ag produced trace ethane (Figure S6C,

220

D). The results suggest a reaction pathway of TCE hydrogenolysis to chlorinated

221

intermediates, then to ethene followed by its hydrogenation to ethane on Fe-Me

222

especially on Fe-Pd and Fe-Ni. Yet, the generation of ethane and coupling products

223

was concurrent with chlorinated intermediates and ethene in Fe-Me (Figure 3), which

224

suggests another pathway of direct transformation of TCE to ethane and coupling

225

products before their desorption from particle surface. TCE dechlorination on Fe-Me

226

involving sorption, reduction (hydrogenolysis, hydrogenation and coupling processes)

227

and desorption processes is summarized in Figure S7. The sorption of TCE on Me

228

was first via C=C through π bond (physical adsorption) followed by forming di-δ

229

bond with Me surface (chemical adsorption).30, 68-73 It is noteworthy that the strength

230

of TCE sorption on Me somehow controls the product distribution.39 For example, Ni

231

facilitates the breaking of C=C bond while Pd does not,39,

232

formation of higher fraction of ethane and odd-numbered coupling products during

233

TCE dechlorinaton by Fe-Ni compared to Fe-Pd. For the convenience of kinetic

234

modeling, the TCE dechlorination pathway on Fe-Pd and Fe-Ni is simplified in

235

Figure 2B without including the negligible chlorinated intermediates.

72

which explains the

236

Both S-nZVI and nZVI transformed acetylene to ethene, ethane and longer chain

237

hydrocarbons (C3-C6) (Figure S8A, B). However, both particles were not able to 13 ACS Paragon Plus Environment


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transform ethene to ethane (Figure S8C, D). Therefore, the dechlorination of TCE on

239

S-nZVI and nZVI was via β-elimination to form acetylene followed by its

240

hydrogenation to form ethene, ethane and coupling products in parallel, as

241

summarized in Figure 2C. The results are consistent with previous studies on the

242

pathway of TCE dechlorination by sulfidated nanoscale or microscale ZVI.53, 60 It

243

should be noted that some previous studies also reported reduction of ethene to ethane

244

by ZVI.69 The underlying mechanism for this difference is unclear but may relate to

245

different types/sources of ZVI used in these studies.

246 247

Reaction Kinetics. The kinetics of TCE dechlorination by Fe-Ni (Figure 3B) and

248

S-nZVI (Figure 3E) were well described by a model containing zero-order terms for

249

reaction steps shown in Figure 2. The fitted rate constants of each reaction step (k1-k4)

250

and TCE disappearance (kTCE) are given in Table S3. Zero-order kinetics with high

251

dose of contaminants has been observed previously with ZVIs, rationalized as the

252

effect of site saturation, and modeled with Langmuir-Hinshelwood-Hougen-Watson

253

(or apparent zero-order) kinetics.19, 53, 69 However, the TCE dechlorination by Fe-Pd

254

was more consistent with first-order kinetics (Figure 3A). This may be because Fe-Pd

255

degraded the adsorbed TCE fast enough to leave the sites unsaturated. The kinetics of

256

TCE reduction by Fe-Cu, Fe-Ag, and nZVI were not modeled because there was no

257

sufficient degradation of TCE (< 3%) in the tested time period (8 days) (Figure 3C, D,

258

and F). In this case, initial TCE reduction rates (kTCE) were determined by fitting the


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product generation data to zero-order kinetics. The resultant rate constants normalized

260

by BET surface area of all nZVI-based particles are summarized in Table S3.

261 262 263 264 265 266 267

Figure 3. Kinetics of TCE dechlorination and formation of major products by Fe-Pd (A), Fe-Ni (B), Fe-Cu (C), Fe-Ag (D), S-nZVI (E) and nZVI (F) under excess TCE conditions (Me/Fe = 0.27 mol%, S/Fe = 20 mol%). CPs represents chlorinated products. Smooth curves on (A), (B) and (E) are calculated from the kinetic model with rate constants in Table S3. Connecting lines on (C), (D) and (F) are only interpolated. All experiments at pH 8 buffered by 25 mM HEPES and under anoxic conditions.

268



269

Pristine nZVI barely transformed TCE due to that the as-synthesized particles

270

(without acid pre-treatment) did not expose any Fe0 according to XPS results. Both

271

the bimetallic and sulfidated modification increased the reaction rate (kSA) of TCE

272

dechlorination for at least one-order magnitude. If the initial TCE dechlorination rate

273

by Fe-Pd was also estimated using zero-order kinetics, the rates follow an order of

274

Fe-Pd > S-nZVI > Fe-Ni >> Fe-Cu > Fe-Ag (Table S3 and Table S4), which are ~

275

6,800, 190, 130, 20 and 8-fold higher than that of nZVI, respectively. Previous

276

research reported a 2-order magnitude faster TCE dechlorination rate by nZVI under

277

excess TCE conditions,19 which may be due to that their nZVI particles were

278

synthesized in methanol/water phase and had a much higher Fe0 content of 97%. It

279

should be noted that much higher S/Fe molar ratio was used than Me/Fe in bimetals.

280

At same S/Fe and Me/Fe ratio (0.27 mol%), the reaction rate of S-nZVI became lower

281

than that of Fe-Ni but still higher than those of Fe-Cu and Fe-Ag, as shown in Figure

282

S9. In addition, the reduction of TCE by Fe-Pd, Fe-Ni, and S-nZVI did not show any

283

indication of decreased reactivity before the end of reaction, suggesting iron corrosion

284

did not cause passivation of reactive sites.

285

H2 Evolution. Time series data for HER by Fe-Me, S-nZVI and nZVI in the absence

286

and presence of excess TCE are shown in Figure 4A and 4B, respectively. Without

287

TCE, all the deposited metal accelerated HER in varying degrees despite of same

288

additive loading, and the Fe0 in all bimetals exhausted in 7 days (longevity of Fe-Pd,

289

Fe-Ni ~ 5 days, Fe-Cu, Fe-Ag ~ 7 days) while only ~ 50% Fe0 in nZVI was consumed.

290

The initial rate of H2 evolution for bimetals, defined as the molar quantity of H2 16 ACS Paragon Plus Environment

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generated at 24 h divided by the time, the particle mass concentration and the surface

292

area, ranged from 18 µmol·L·m-2·h-1 (Fe-Ag) to 49 µmol·L·m-2·h-1 (Fe-Pd), which

293

were ~3 to 9-fold higher than that of nZVI. In addition, the corrosion of nZVI was

294

nearly completely passivated after ~1 month (Figure S10). These results indicate that

295

bimetallic modification shortened the longevity of nZVI particles.

296 297 298 299 300 301 302

Figure 4. Hydrogen evolution in the absence (A) and presence (B) of excess TCE (Fe: 0.25 g/L). The relationship of corrosion currents of Fe-Me v.s. initial HER rates (C) and the standard Gibbs energy of adsorption of hydrogen on Me surfaces (∆GH ) v.s. initial HER rates (D). The flat line on (C) is a base line according to kH2 of nZVI. Connecting lines on (D) are only interpolated. All experiments at pH 8 buffered by 25 mM HEPES and under anoxic conditions.

303

In contrast, S-nZVI significantly inhibited iron corrosion. The initial HER rate of

304

S-nZVI (4.7 µmol·L·m-2·h-1) was comparable to that of nZVI, however, it slowed

305

down significantly thereafter and H2 evolution completely stopped after 7 days. The

306

total H2 produced from S-nZVI corrosion only accounted for ~35% of the initial Fe0 17 ACS Paragon Plus Environment


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in the particles, which suggests that the majority of Fe0 was preserved and therefore

308

the particle longevity was extended.

309

With excess TCE present, hydrogen evolution with Fe-Pd exhibited a volcano

310

curve, where H2 evolution reached a peak of 3% as initial Fe0 at 2 h, and then

311

gradually vanished along with TCE dechlorination (Figure 4B and Figure S11). This

312

suggests that the generated H2 was utilized for further reduction of TCE, in agreement

313

with the known ability of Pd in catalyzing H2 for hydrodechlorination.74-76 For Fe-Ni

314

and S-nZVI, the H2 evolution rate decreased with time and H2 generation leveled off

315

after 24 h at 19% and 9% of the added Fe0, respectively. The absence of a volcano

316

pattern for Fe-Ni and S-nZVI as well as Fe-Cu and Fe-Ag suggests all these particles

317

didn’t catalyze TCE reduction by H2. This is verified with experiments performed

318

with headspace purged with pure H2, which did not result in increased TCE

319

dechlorination (Figure S12). However, for Fe-Ni, when the Ni/Fe ratio was increased

320

to 20 mol%, the increased TCE dechlorination was significant with H2 headspace

321

(Figure S13), which suggests the ability of Ni in catalyzing H2 at high concentration.

322

The much lower H2 evolution with Fe-Pd, Fe-Ni and S-nZVI in the presence of excess

323

TCE suggests that a larger portion of the reducing equivalents of Fe0 were consumed

324

by TCE due to either TCE competing electrons with H+ (e.g., Fe-Pd, Fe-Ni, and

325

S-nZVI) or the utilization of the generated H2 (e.g. Fe-Pd).

326

It is widely agreed that the hydrogen evolution reaction (HER) occurred in two

327

key steps.77-81 First is the Volmer reaction (H+ + e- + Me → Me-Hads), which involved

328

the electron transfer to proton on the surface of electron conductors such as Me to 18 ACS Paragon Plus Environment

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329

form adsorbed atomic hydrogen, then followed by the second step of atomic hydrogen

330

combination and desorption: (2a) Tafel reaction (Me-Hads + Me-Hads → 2Me + H2) or

331

(2b) Heyrovsky reaction (Me-Hads + H+ + e- → Me + H2). For corrosion of pristine and

332

all modified nZVI particles, the corrosion currents of nZVI and each modified nZVI

333

particles (Table S5) can be obtained by fitting the Tafel curves of electrodes modified

334

by each particle type (Figure S14). Both doping nZVI with a second metal and

335

sulfidation increased corrosion current and therefore enhanced electron transfer from

336

Fe0 to the surface. The cause of this increase for Fe-Me is the galvanic effect while

337

that for S-nZVI is the increased electron conductivity of FeS compared to iron

338

oxides.53, 82 However, the initial HER rates of Fe-Me are not linear to their corrosion

339

currents (Figure 4C). For example, the corrosion current of Fe-Ni was about 2-fold

340

higher than that of Fe-Cu, however, their initial HER rates were comparable. This

341

suggests that HER is not just controlled by electron transfer. Interestingly, the plot of

342

HER rates v.s. standard Gibbs energy of adsorption of hydrogen on each material

343

surface (∆GH, Table S6) shows a volcano curve that is typical for electrochemical

344

HER (Figure 4D).83-85 This pattern is explained by that too weak binding of atomic

345

hydrogen with Cu and Ag made the proton/electron-transfer step thermodynamically

346

unfavorable while too strong binding of atomic hydrogen with Ni restricted their

347

combination to release hydrogen.86

348 349

Efficiencies and Dechlorination Capacity. In the presence of excess TCE, almost no

350

Fe0 remained in all residual particles except pristine nZVI after reaction, suggesting



351

that Fe0 in all modified nZVIs was accessible (i.e, εFe(0) was ~100%). This also

352

suggests that the iron (oxyhydr)oxide layer on Fe-Me and S-nZVI surfaces did not

353

become passivating enough to prevent electron transport across this layer to either

354

TCE or H2O (at pH 8 and other conditions of these experiments). In contrast, the εe of

355

each particle type varied significantly. While εe of nZVI, Fe-Cu, and Fe-Ag were < 4%,

356

εe of Fe-Pd, Fe-Ni, and S-nZVI were at least one order magnitude higher. The εe of

357

Fe-Pd was ~100%, which suggests that all the electrons were utilized by TCE either

358

directly or indirectly (i.e., via forming ·H first and then being utilized). However,

359

under field conditions the generated H2 would diffuse away and the εe may become

360

lower. The εe of Fe-Ni lying at 69% was lower than that of S-nZVI (72%), consistent

361

with the lower H2 evolution of S-nZVI. Nonetheless, the εe difference between Fe-Ni

362

and S-nZVI was not as dramatic as their difference in HER rates in the absence of

363

TCE but comparable to their rate difference in degrading TCE. Therefore, the most

364

important cause of the high εe of Fe-Ni, Fe-Pd and S-nZVI compared to nZVI under

365

excess TCE conditions was that they reacted with TCE much faster, which created

366

selectivity.

367

The overall capacity of each particle type to reduce TCE depends on not only

368

particle efficiencies but also the products formed from TCE dechlorination. The molar

369

numbers of electrons required to convert 1 mole TCE to corresponding products (n)

370

for all particles are provided in Table 1. The n of nZVI was higher than those of

371

Fe-Cu and Fe-Ag but lower than those of Fe-Pd and Fe-Ni. This is due to that Cu and

372

Ag modification caused the accumulation of chlorinated products while Pd and Ni 20 ACS Paragon Plus Environment

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modification led to formation of more reduced products (i.e., ethene and ethane). In

374

particular, Fe-Ni had the highest n as it converted most TCE to ethane. The n of

375

S-nZVI was lower than bimetals due to the formation of less saturated products (i.e.,

376

acetylene) by S-nZVI (Table 1). These results imply that if particle efficiencies are

377

the same, S-nZVI can dechlorinate about 20%~40% more TCE than bimetallic

378

particles. If we assume that the εe obtained from the batch experiments with excess

379

TCE is applicable to a source zone condition in field implementation, the mass of

380

TCE dechlorinated per unit mass of Fe0 in S-nZVI can be estimated to be 5.7 mmol/g

381

Fe0 (Table 1). The higher dechlorination capacity of S-nZVI over bimetals (5.2, 3.2,

382

0.2, 0.15 mmol/g Fe0 for Fe-Pd, Fe-Ni, Fe-Cu, and Fe-Ag, respectively) and nZVI

383

(0.13 mmol/g Fe0, Table 1) is a significant advantage of S-nZVI in groundwater

384

remediation. The calculated dechlorination capacity may be overestimated due to that

385

the contact between particles and contaminants may become a limiting factor for

386

dechlorination under field conditions. Nonetheless, polymers such as CMC coated on

387

nZVI to enhance particle transport would also stimulate microbial activity in field

388

applications, which may provide significantly more electrons.13, 87, 88

389

390

391

392

393



394 395

Page 22 of 34

Table 1. Efficiencies and Capacities of Fe-Me, S-nZVI, and Pristine nZVI Particles under Excess TCE Conditions.

n

Fe0 utilization

Electron efficiency

efficiency (εp, %)

(εe, %)

mmol of TCE dechlorinated per g of Fe0

Fe-Pd

6.5

99.8

99.9

5.2

Fe-Ni

7.6

99.9

69

3.2

Fe-Cu

6

99.8

3.3

0.2

Fe-Ag

5.6

99.7

2.3

0.15

S-nZVI

4.5

99.9

72

5.7

nZVI

6.5

-

2.4

0.13

396

Mechanisms of Enhanced Reactivity and Selectivity of Fe-Me and S-nZVI. On

397

the basis of corrosion currents obtained from Tafel characterization, bimetallic

398

modification enhanced electron transfer particularly to Me surface due to galvanic

399

effect. Yet, this increase did not result in proportional increase of HER rate due to that

400

HER was also affected by the binding energy of Me-H. For TCE dechlorination, the

401

rate could be further affected by the binding strength of Me-TCE. In order to link the

402

HER rate and TCE dechlorination rate, we investigated the TCE dechlorination by

403

modified nZVIs under excess iron conditions, where HER and TCE dechlorination

404

processes did not affect each other (Figure S15). We found a strong correlation

405

between HER rates and TCE dechlorination rates for the investigated Fe-Me (linear

406

fitting, R2 = 0.97, Figure 5A). This suggests that HER and TCE dechlorination on the

407

tested Fe-Me are controlled by same reaction steps, and Me-TCE binding is not an

408

important factor for TCE dechlorination. Since HER of Fe-Me is mainly a result of

409

combination of adsorbed atomic hydrogen (·Hads) on Me surface due to galvanic


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410

effect, TCE dechlorination should also be mediated by ·Hads on Me surface. Schrick et

411

al.39 also proposed that atomic hydrogen was responsible for TCE dechlorination by

412

Fe-Ni but no attempts were made to distinguish whether it was ·Hads or absorbed

413

atomic hydrogen (·Habs). Some later studies suggested that ·Habs into lattice of metal

414

additives controlled the hydrodechlorination reactivity of bimetallic ZVI particles.31, 37,

415

48

416

To elucidate whether ·Hads or ·Habs is the reactive species, we conducted

417

additional experiments of TCE hydrodechlorination with CMC-Pd nanoparticles at

418

varied H2 concentrations. We observed that TCE dechlorination only started to occur

419

after H/Pd molar ratio was over 1.4 (Figure 5B), at which the stoichiometric H/Pd

420

ratio is consistent with that of the palladium hydride (Pd38H55).89 It suggests that the

421

dechlorination process occurred only after the saturation of CMC-Pd by ·H, which

422

confirms that ·Hads on Me surface is the reactive species responsible for TCE

423

dechlorination by Fe-Me. This conclusion is consistent with previous understanding

424

of electrocatalytic dechlorination of 2,4-dichlorophenol on Pd nanoparticles

425

predominantly mediated by ·Hads.51 Since the amount of ·Hads on Fe-Me is controlled

426

by both the corrosion rate of Fe-Me and the binding strength of Me-H, TCE

427

dechlorination is also determined by both factors. Palladium with the optimal

428

properties of interacting with hydrogen possesses the highest reactivity for both HER

429

and TCE dechlorination. A mechanistic basis for the HER and TCE dechlorination on

430

Fe-Me is proposed in Figure 6A.



431 432 433 434 435 436

Figure 5. (A) The linear correlation of initial HER rates with initial kTCE of Fe-Me under excess iron conditions (S-nZVI not included; Iron: 1 g/L, TCE: 10 mg L-1; All experiments at pH 8 buffered by 25 mM HEPES and under anoxic conditions). (B) the 4h TCE removal by CMC-Pd at varied H2 concentrations represented by H/Pd atomic ratio (B); Lines are only interpolated.

437

For S-nZVI, sulfidation also enhanced electron transfer. However, this increase

438

of electron transfer resulted in reduced HER along with enhanced TCE dechlorination.

439

As expected, the point of S-nZVI in Figure 5A deviates significantly from the linear

440

line, which suggests that TCE dechlorination on S-nZVI may not be mediated by ·H

441

as the case for Fe-Me and a different mechanism was responsible for the enhanced

442

particle reactivity and selectivity. The non-atomic hydrogen mechanism is also

443

supported by the differed primary TCE dechlorination pathway on S-nZVI

444

(β-elimination) and Fe-Me (hydrogenolysis). We note that TCE dechlorination on

445

bare FeS was via β-elimination to form mainly acetylene and not mediated by atomic

446

hydrogen.68,

447

S-mZVI mainly occurred on FeS sites.53 It should be the similar case for S-nZVI, as

448

indicated by similar more accumulation of acetylene than nZVI during TCE

449

dechlorination. The observation that HER of S-nZVI ceased after 2 days while TCE

450

dechlorination continued and at an unchanged rate (Figure 3E and 4B) further

90

Our previous research has suggested that TCE dechlorination on


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451

indicates that the HER sits were mainly on iron oxides. This also suggests that the

452

atomic hydrogen on iron sulfides should be insignificant and not be responsible for

453

the TCE dechlorination on FeSx sites.

454

On the basis of the redox ladder constructed in our previous study,53 electron

455

flow is favorable through the Fe/FeS interface compared to Fe/Fe oxides. Therefore,

456

FeSx conducting major electrons (~72% based on electron utilization efficiency

457

calculation) mediated TCE dechlorination while Fe oxides conducting minor electrons

458

mediated HER, which explains the enhanced TCE reduction but reduced HER of

459

S-nZVI. Based on the discussion above, a mechanistic basis for the HER and TCE

460

dechlorination on S-nZVI was also proposed (Figure 6B).

461 462 463 464

Figure 6. Schematics of the proposed TCE dechlorination and HER on Fe-Me (A) and S-nZVI (B). The chemisorption of H2 on Me surface in (A) is only significant for Fe-Pd or Fe-Ni at high Ni loading. The portions of electron flow in (A) vary with Me.

465

We note that some researchers56 proposed a different explanation on the

466

enhanced TCE dechlorination but inhibited HER by S-nZVI that sulfur in S-nZVI

467

poisoned ·H recombination, which favored TCE reduction by ·H. Although the

468

underlying mechanism of this poisoning effect is not clear, it may be explained by the

469

strong binding of FeS-H that inhibited ·H recombination. However, the strong binding

470

of FeS-H should also inhibit the reaction of ·H with TCE to cause retarded 25 ACS Paragon Plus Environment


471

dechlorination. In addition, a pre-hypothesis of this explanation is that ·H mediates

472

TCE dechlorination on FeS surface. However, for bare FeS, dechlorination must be by

473

electron transfer.28, 67, 91 For S-nZVI, as TCE dechlorination was also mainly on FeS

474

surface, the process was more likely via electron transfer than via ·H. Further,

475

previous studies suggested that the pH dependence of TCE dechlorination was

476

opposite to HER or the availability of ·H on particle surface,28, 57, 58 which is also

477

inconsistent with this explanation.

478 479

Environmental Implications

480

Nanoscale ZVI with its flexibility in field injection, higher reactivity compared to

481

larger scale ZVI yet with high cost is likely to be used in source zone remediation.

482

Both bimetallic modification and sulfidation can enhance the reactivity of nZVI to

483

varying degrees. Sulfidation stands out as a promising nZVI modification approach as

484

it inhibits HER (i.e., longer lifetime), stays reactive for extended period of time, and

485

possesses the highest dechlorination capacity. The information of TCE dechlorination

486

rates, efficiencies, and capacities by modified nZVIs obtained from this study can

487

instruct practitioners to estimate the injection frequency and quantities of these

488

materials in field injection. The identification of electron transfer as the dominant

489

mechanism of TCE dechlorination on S-nZVI will further help predict the

490

dechlorination efficiency of other chlorinated contaminants in groundwater by

491

S-nZVI.


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492

Acknowledgements

493

The majority of this work was supported by the Natural Science Foundation of

494

Zhejiang Province (LR16E080003).

495

Supporting Information

496

Additional information on materials and methods, SEM and TEM images, Fe 2p XPS

497

spectra and fitting, reduction of additional compounds and electrochemical

498

characterization results.

499 500



501

References

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Dechlorination of Excess Trichloroethene by Bimetallic and Sulfidated

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