Transformation of Polycyclic Aromatic Hydrocarbons and Formation of

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Environmental Processes

Transformation of PAHs and formation of environmentally persistent free radicals on modified montmorillonite: Role of surface metal ions and PAH molecular properties Hanzhong Jia, song Zhao, Yafang Shi, Lingyan Zhu, Chuanyi Wang, and Virender K. Sharma Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b00425 • Publication Date (Web): 16 Apr 2018 Downloaded from http://pubs.acs.org on April 16, 2018

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Environmental Science & Technology

1

Transformation of PAHs and Formation of Environmentally

2

Persistent Free Radicals on Modified Montmorillonite: Role of

3

Surface Metal Ions and PAH Molecular Properties

4 5 6

Hanzhong Jiaa,b, Song Zhaob, Yafang Shia, Lingyan Zhua, Chuanyi Wangb, and

7

Virender K. Sharmac*

8 9 10 a

11

Key Laboratory of Plant Nutrition and the Agri-environment in Northwest China,

12

Ministry of Agriculture, College of Natural Resources and Environment,

13

Northwest A & F University, Yangling 712100, China.

14

b

Xinjiang Technical Institute of Physics & Chemistry, Chinese Academy of Sciences, Urumqi 830011, China.

15 16

c

Program for the Environment and Sustainability, Department of Occupational and

17

Environmental Health, School of Public Health, Texas A&M University,

18

College Station, TX 77843, USA.

19 20 21 22 23 24

E-mails: [email protected] (HZJ); [email protected] (VKS).

1

ACS Paragon Plus Environment


25

ABSTRACT

26

This paper presents the transformation of PAHs (phenanthrene (PHE), anthracene

27

(ANT), benzo[a]anthracene (B[a]A), pyrene (PYR), and benzo[a]pyrene (B[a]P)) on

28

montmorillonite clays that are modified by transition-metal ions (Fe(III), Cu(II),

29

Ni(II), Co(II), or Zn(II)), at room temperature (~ 23 oC). Decay of these PAHs follows

30

first-order kinetics, and the dependence of the observed rate constants (kobs, d-1) on the

31

presence of metal ions follows the order Fe(III) > Cu(II) > Ni(II) > Co(II) > Zn(II).

32

The values of kobs show reasonable linear relationships with the oxidation potentials of

33

the PAHs and the redox potentials of the metal ions. Notably, transformation of these

34

PAHs results in the formation of environmentally persistent free radicals (EPFRs),

35

which are of major concern due to their adverse effects on human health. The

36

potential energy surface (PES) calculations using density functional theory were

37

performed to understand (a) the trends in kobs, and (b) the plausible mechanisms for

38

radical formation from the PAHs on modified clays. The yields and stability of these

39

EPFRs from ANT and B[a]P on clay surfaces varies with both the parent PAH and

40

metal ion. The results demonstrated the potential role of metals in the formation and

41

fate of PAH-induced EPFR at co-contaminated sites.

42 43 44 45 46 2


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TOC Art

48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67 68 69 70 71 72 73 74 75 76 77 78 3



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INTRODUCTION

80

Atmospheric particles, soil, and sediments that are co-contaminated with toxic

81

metals and polycyclic aromatic hydrocarbons (PAHs) have raised concerns due to

82

their potential to cause combined adverse effects on human and ecological health.1-4

83

The co-presence of toxic metals, especially some transition metals, may also change

84

the particle properties, which, in turn, affects the transport, fate, and toxicity of PAHs

85

and other organic pollutants.5,6 For example, particulate matter containing

86

chlorophenol and transition metal ions emitted from combustion sources in the

87

atmosphere produces environmentally persistent free radicals (EPFRs) that may

88

increase the human health risk of developing respiratory and cardiopulmonary

89

diseases.7-10 These types of EPFRs could also be observed during the oxidative

90

decomposition of aromatic compounds (e.g., catechol and dibenzofuran).11,12 The

91

formation of EPFRs and their ecotoxicological effects in the natural environment have

92

attracted increasing attention from scientists and public health decision makers.13,14

93

Fly ash and particulate matter contain organic contaminants and transition metals. As

94

such, several studies have been conducted to understand combustion related EPFRs

95

on

96

organic-contaminated

97

combustion-generated EPFRs has been explored.17-24 Progress has been made in these

98

systems to characterize the electron transfer from organic molecules, especially

99

chloro- and hydroxyl-substituted benzenes, to the metal/silica surfaces, resulting in

100

the formation of EPFRs.25-28 Comparatively, only limited studies have been performed

metal/mineral

surfaces.14-18 surfaces

The in

role the

of

transition

formation

4


and

metal

oxides

on

stabilization

of

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101

examining the formation of EPFRs and their stabilization on contaminated soils at

102

room temperature and also under environmental conditions.28-32

103

Recently, we carried out a study on the formation and stabilization of EPFRs

104

on soil samples obtained from former coking sites.33 The coexistence both of PAHs

105

and heavy metals was detected in the sampled soils, and what's more, the levels of

106

PAHs and heavy metals in these soil samples correlated with the concentration of

107

EPFRs. Smectite clay, a representative inorganic component of soil, acts as a sorbent

108

of metal and organic contaminants, and plays an important role in the generation of

109

EPFRs.29 In addition, there was a strong correlation between the presence of iron and

110

the formation of EPFRs, which is in agreement with investigations of the interaction

111

of anthracene with Fe(III)-modified clays.29 Interestingly, the presence of ZnO

112

nanoparticles, which are not so commonly involved in electron-transfer processes,

113

resulted in the formation of EPFRs when exposed to phenol at room temperature.30

114

Based on these observations, multiple transition metals may play a role in generating

115

EPFRs on organic-contaminated soils. In addition, the structural properties of

116

precursors also influence the type either carbon or oxygen centered of EPFRs.11,12

117

However, only a few studies have been performed on this topic, and they have been

118

typically conducted using only a single organic compound and a single type of

119

transition metal on clay mineral as a representative of contaminated soil.29 More

120

information is needed regarding the influence of the electronic properties and

121

molecular structures of the PAHs, and the redox properties of the metal ions, on the

122

generation and stabilization of EPFRs. Because transition or/and toxic metals and 5



123

PAHs are commonly found in contaminated soils and wastes, the focus of the present

124

study is to elucidate the role of surface metal ions and PAH molecular properties on

125

EPFR formation and fate on soil mineral surfaces.

126

The objectives of the current paper are: (i) to understand the transformations

127

of selected 3-, 4-, and 5-membered ring linear and branched PAHs (phenanthrene

128

(PHE), anthracene (ANT), benzo[a]anthracene (B[a]A), pyrene (PYR), and

129

benzo[a]pyrene (B[a]P)) on montmorillonite clays modified by transition metal ions

130

(Fe(III), Cu(II), Ni(II), Co(II), and Zn(II)); (ii) to elucidate the underlying mechanism

131

of the formation of EPFRs produced over time on clays containing selected PAHs and

132

transition metal ions by applying the electron paramagnetic resonance (EPR)

133

technique and density functional theory (DFT) calculations, and (iii) to evaluate the

134

stability of the formed EPFRs in order to identify potential risks associated with the

135

interactions of PAHs and metal ion-contaminated soils. These results have

136

implications for human health due to possible diseases associated with long term

137

exposure to metal ions and PAH-contaminated atmospheric and soil particles.

138 139

EXPERIMENTAL SECTION

140

Chemicals and materials. Detailed information on the chemicals used in this

141

study is provided in the Text S1 (Supporting Information). Reference montmorillonite

142

as clay was obtained from Zhejiang Feng-Hong Clay Chemicals Co., Ltd (ZheJiang,

143

China). The cation exchange capacity (CEC) and specific surface area were 82.0 cmol

144

kg-1 and 82.1 m2 g-1, respectively.29 6


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Preparation of PAH-contaminated clays. The preparation of polycyclic

146

aromatic hydrocarbon (PAH)-contaminated clays, modified by transition metal ions,

147

involved two steps: (i) saturating the cation exchangeable sites of the montmorillonite

148

clay with the desired metal ions according to a previously described method,34 and (ii)

149

spiking the modified clay samples with various PAHs to prepare contaminated clay

150

samples. Initially, manufactured acquired clay (Zhejiang Feng-Hong Clay Chemicals

151

Co., Ltd., ZheJiang, China) was suspended in Milli-Q water at a ratio of 1:20 (w/w)

152

(clay : water) and stirred for 12 h. This step allowed complete hydration of the clay.

153

This hydrated clay was centrifuged for 6 min at 60 ×g speed. In this procedure, the

154

impurities larger than 2 µm settled down and the clay supernatant was decanted into a

155

beaker. The obtained clay suspensions were further purified to remove carbonate by

156

titrating with 0.5 M sodium acetate buffer (pH 5.0) until the pH of the suspension

157

could be maintained at < 6.8 for 2 h. Following pH adjustment, the suspension was

158

centrifuged for 20 min at 3295 ×g speed, and the supernatant was discarded. The clay

159

samples were then replenished with 0.1 mol L-1 NaCl solution and the solution was

160

stirred for 8 h, followed by centrifugation for 20 min at 3295 ×g speed. The

161

supernatant liquid was discarded, and the clay samples were resuspended in 0.1 mol

162

L-1 NaCl solution. This procedure was repeated four times to ensure complete

163

saturation of the cation exchange sites of the clay with Na+ ions. The obtained Na+-ion

164

containing clay samples were washed using Milli-Q water until the supernatant liquid

165

was free of chloride ion. The absence of Cl- ions was confirmed by a negative test

166

using AgNO3. 7



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Modified clays containing metal ions (Fe(III), Cu(II), Ni(II), Co(II), or Zn(II))

168

were prepared by using the same procedures as described above for obtaining Na+

169

ion-containing clays. The only difference was the replacement of 0.1 mol L-1 NaCl

170

solution with Fe(III) (0.033 mol L-1), Cu(II) (0.05 mol L-1), Ni(II) (0.05 mol L-1),

171

Co(II) (0.05 mol L-1), or Zn(II) (0.05 mol L-1) solutions. After washing the modified

172

clay by Milli-Q water, the pH of the suspended clay was adjusted to 5.5-6.0 by adding

173

either 0.01 M H2SO4 or 0.01 M NaOH. After preparation, all of the metal

174

ion-modified clay samples were quickly frozen, at -40 oC, followed by freeze drying

175

and storage in polyethylene bottles. Metal levels in original montmorillonite clays and

176

metal ions loading clays were determined by an inductively coupled plasma-atomic

177

emission spectroscopy (ICP-AES) method (see Text S2, Supplementary Information).

178

Next, the reaction mixtures of PAH-contaminated metal ion-modified clays were

179

prepared by mixing 1 g of these different metal ion-modified clays with 5 mL of a

180

solution containing individual PAHs of 0.02 mg mL-1 (phenanthrene (PHE),

181

anthracene (ANT), benzo[a]anthracene (B[a]A), pyrene (PYR), and benzo[a]pyrene

182

(B[a]P)) in acetone solvent. The clay-PAH solutions were rapidly stirred for 1 h to

183

promote the swelling of the clay interlayers and interaction between PAH molecules

184

and clay surfaces. The PAH-containing clay samples were stored under ambient

185

conditions (~ 23 oC) without light irradiation until the acetone was completely

186

evaporated. This step resulted in an individual PAH concentration of 0.1 mg g-1 in

187

clays.

188

The

prepared

clay

samples

(metal

ion-modified

8


clays

and

metal

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ion-modified/PAH-contaminated clays) were characterized by X-ray diffraction

190

(XRD) to determine the d(001) basal spacings (see Text S2, Supplementary

191

Information). As shown in Figure S1 (Supporting Information), the basal spacings

192

were 13.45, 13.39, 13.45, 13.55, and 13.63 Å for Fe(III), Cu(II), Ni(II), Co(II), and

193

Zn(II) modified montmorillonite clays, respectively. The results indicated that a lower

194

diffraction angle and higher basal spacing was observed when Na+ ions were replaced

195

with transition metal ions. During the preparation of the PAH-containing modified

196

clays, acetone was used as the solvent for mixing PAHs with modified clays. The

197

mixed suspensions were rapidly stirred to promote the swelling of clay interlayers and

198

intercalating of PAH molecules into the clay surfaces. The intercalating of PAHs into

199

the spaces between layers further induced lowering of diffraction angle and increasing

200

the basal spacing of the dehydrated clays to ~ 13.89-15.45 Å (see Figure S1,

201

supporting Information). In addition, sorption of the PAHs broadened the peaks with a

202

decrease in intensity, suggesting that the interactions of the PAHs with the clay

203

interlayers disordered somewhat the crystalline structure. Those results indicated that

204

the PAHs was intercalated into the clay interlayers successfully.

205

PAH transformation and products analysis. After the preparation of

206

PAH-contaminated metal ion-modified clays, each sample was laid onto a Petri dish

207

and immediately placed inside a desiccator at room temperature (23 oC). The

208

desiccator had a relative humidity of ~ 7%. At a preselected time, a certain amount of

209

PAH-metal ion-modified clay was taken from the sample in the Petri dish to quench

210

the reaction and extraction. No influence of possible humidity change, during the time 9



211

period of less than 1 minute for sample taken and quenching the reaction, on the

212

decay of PAH and the formation of EPFRs was seen (see Text S3 and Figure S2,

213

Supplementary Information).

214

A total of 10 mL of solvent mixture (1:1 (v/v) acetone and dichloromethane)

215

was added to each sample immediately to quench the reaction, and the sample was

216

placed in an ultrasonic bath for 30 min to extract the PAHs and their products. After

217

this step, each suspension was centrifuged at 23300 ×g speed for 5 min to separate the

218

supernatants from the solids. This procedure was repeated twice for each sample to

219

ensure that the organic compounds were fully extracted. The extraction efficiency of

220

the newly prepared PAHs-contaminated clay samples was ~ 98% (see Table S1,

221

Supplementary Information). This suggests that the interaction between PAHs and

222

clay surfaces had no significant influence on the extractability. The ultrasonication

223

process also had no influence on the transformation of PAHs. The obtained

224

supernatants were collected and filtered using a 0.22 µm Nylon organic membrane

225

syringe filter (Titan, China). Control experiments were conducted using the original

226

montmorillonite clay, saturated mainly with Na+ on the negative sites of the clay

227

layers in the PAH-contaminated Na+-montmorillonite clay. The filtrates were stored

228

in amber vials for analyzing by either high performance liquid chromatography

229

(HPLC) or gas chromatography-mass spectrometry (GC-MS) technique. More details

230

for HPLC and GC-MS techniques are provided in Text S4 (Supporting Information).

231

In the EPR study, 0.2 g of solid samples were placed in a high purity quartz EPR tube

232

and measurements were performed at room temperature. Instrument and operating 10


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parameters are detailed in Text S5 (Supporting Information).

234

The interaction between PAHs and metal ions-modified clays was studied by

235

UV-vis spectroscopy (Text S5, Supporting Information). X-ray photoelectron

236

spectroscopy (XPS) technique was applied to detect the change of valence states of

237

metal ions after 10 d of reaction time on modified clays, and PAHs-contaminated clay

238

samples (Text S5, Supporting Information).

239

Density function theory (DFT) calculations. Density functional theory (DFT)

240

calculations were carried out to evaluate the reaction energies associated with the

241

proposed reaction pathway using the Materials Studio 6.0 of Dassault Systèmes

242

Biovia Corp (San Diego, California, United States). Based on the observation in

243

previous work, the interlayer exchangeable cation species in clay interlayers existed

244

as hydrated species coordinated by a shell of water molecules.35,36 In our study,

245

therefore, gas-phase [M(H2O)6]n+ was used as the model to represent the interlayer

246

metal ion species responsible for the catalytic oxidation of organic contaminants to

247

intermediate radicals, which is also applied in earlier study37. The structural

248

optimization and formation energies were calculated using DMol3 code.38 The

249

generalized gradient approximation (GGA) with the Perdew–Burke–Ernzerhof (PBE)

250

functional and all-electron double numerical basis set at polarized function (DNP)

251

were employed.39 The convergence tolerance of energy was set at 1.0×10-5 Ha (1

252

Ha=27.21 eV) and maximum force was 2.0×10-3 Ha/Å. Each structure was allowed to

253

fully relax to the minimum in the enthalpy without any constraints. Each atom in the

254

storage models is allowed to relax to the minimum in the enthalpy without any 11



255

constraints. The transition states of all systems were determined.40

256 257 258

RESULTS AND DISCUSSION Decay of PAHs.

The decay of various PAHs by metal ion-modified clays is

259

mostly occurred spontaneously over a period of days (Figure 1a-e). The extent of the

260

decay depends on the type of PAH and the type of metal ion present on the clay. The

261

degradation of PHE is insignificant in all the tested clay samples. The concentrations

262

of ANT, B[a]A, and PYR gradually decrease with time in all samples except the

263

Zn(II)-containing clay. The concentration of B[a]P decreases in all the modified clays.

264

The B[a]P decay rates on Fe(III)-montmorillonite are the highest among the tested

265

reaction systems, while overall, the lowest transformation rates are seen with the

266

Zn(II)-containing clay (Figure 1a versus Figure 1e). In general, the transformation

267

rate of individual PAH by clays modified with metal ions follows a decreasing trend

268

of Fe(III) > Cu(II) > Ni(II) > Co(II) > Zn(II). To evaluate the possible effect of lattice

269

(or structural) transition metal ions on the transformation of PAHs, control

270

experiments were conducted using the original montmorillonite clay, saturated mainly

271

by Na+ on the negative sites of the clay layer. No significant decay of PAHs (< 3%)

272

was observed in these control experiments during a 60 d time period (data not shown).

273

These observations suggest that the transformation of PAHs occurred mainly due to

274

the surface metal ions on the clays.

275

PAHs transformation on clay surfaces could be attributed to the single electron

276

transfer (SET) reaction between arenes molecules and surface cations.32,41 Generally, 12


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the adsorption of PAHs with clay minerals is accompanied by the formation of

278

“cation-π” interaction at the active sites. The coordination of PAHs to

279

electron-accepting sites, i.e., the Lewis acid, increases their electrophilicity, leading to

280

a charge-transfer (CT) (electron donor-acceptor) complex. This complex induces the

281

electron transfer from PAHs to surface cations to cause oxidation, which ultimately

282

leads to the formation of intermediate radicals or/and final products, such as PAH–

283

quinones.32,42,43 Significantly, our results demonstrate that properties of metal ions and

284

PAHs affect the binding strength of CT complex, which influences the electron

285

transfer reaction and the formation of EPFRs. For example, electron-deficient cations

286

(e.g., Fe(III) and Cu(II)) and electron-rich PAHs have relatively strong cation–π

287

interactions, which result in an initial single electron transfer (SET) step to cause

288

transformation of PAH readily.44,45

289

The concentration of PAHs versus reaction time could be fitted reasonably well

290

using first-order decay on clays containing different types of metal ions (Figure 1).

291

The observed first-order rate constants (kobs, d-1) are presented in Table S2

292

(Supporting Information). The values of kobs associated with Fe(III)- and

293

Cu(II)-modified clays follows the order B[a]P > ANT ~ PYR > B[a]A > PHE.

294

However, the Ni(II)- and Co(II)-clays show similar reactivity for the tested PAHs

295

(Table S2).

296

The variation in values of kobs with the types of PAHs is quantitatively

297

analyzed by seeking relationships with oxidation potential, ionization potential (IP),

298

and half-wave potential (E1/2). Because the electron-donating capacity of PAH 13



299

molecules can be characterized by one-electron oxidation potential, the values of kobs

300

can be plotted against oxidation potential for Fe(III)- and Cu(II)-modified clays. As

301

displayed in Figure S3 (Supporting Information), a linear relationship exists between

302

kobs and the oxidation potential of various PAHs. The electron-rich PAHs such as

303

B[a]P, ANT, and PYR have oxidation potential values in the range of from 1.16 to

304

1.37 (versus standard calomel electrode (SCE) (Table S2), and are more easily

305

transformed (or oxidized). PHE, which has an oxidation potential of 1.67 V (versus

306

SCE), is less easily transformed by modified clays. A value of IP represents a crucial

307

index of the electron-donating capacity of organic compounds that can be utilized to

308

elucidate the transformation of PAHs on clay surfaces.32 PAH molecules with low IP

309

values show favorability towards SET oxidation reactions. B[a]P, ANT, and PYR have

310

higher tendency (IP < 7.5 eV) to transfer one electron to surface cations than that of

311

PHE (IP = 7.9 eV). Therefore, higher decays of B[a]P and ANT were observed in

312

metal ions-modified clays than PHE (see Table S2, Supporting Information). Because

313

E1/2ox correlates linearly with the IP of aromatic hydrocarbons, the transformation of

314

PAHs is also dependent on E1/2ox.

315

The different slopes of the linear lines in Figure S3 suggest that the

316

transformation of PAHs also depends on the redox property of the metal ions. The

317

slope obtained for Fe(III) ion is higher than the slope obtained for Cu(II)-modified

318

clay. In case of Ni(II)- and Co(II)-modified clays, no significant relationships between

319

the values of kobs and oxidation potential are observed (see Table S2). However,

320

Ni(II)- and Co(II)-containing clays were involved in the transformation of PAHs. This 14


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321

indicates that metal ions with relatively higher redox potentials may have more

322

capability to influence the decay rates of PAHs. This can also be seen in the redox

323

potentials of Ni(II) and Co(II) (Ni(II) + 2e- ⇌ Ni(s); E0 = -0.257 V (vs NHE) and

324

Co(II) + 2e- ⇌ Co(s); E0 = -0.280 V (vs NHE)), which are much lower than the redox

325

potentials of Fe(III) and Cu(II) (Fe(III) + 2e- ⇌ Fe(II)(s); E0 = 0.771 V (vs NHE) and

326

Cu(II) + e- ⇌ Cu(I); E0 = 0.153 V (vs NHE)). The redox potential of Zn(II) ion is the

327

lowest among the metal ions (Zn(II) + 2e- ⇌ Zn(s); E0 = -0.7618 V (vs NHE)) and

328

hence it has the lowest capability to influence the transformation of PAHs. When

329

testing the transformation using Zn(II) ion-modified clay, decay of PAHs was only

330

seen with B[a]P, which is the most easily oxidized molecule of the tested PAHs (see

331

Figure 1 and Table S2). The results suggest that surface metal ions with higher redox

332

potential or electron deficiency could induce a stronger “cation-π” interaction within a

333

CT (electron donor-acceptor) complex, which would ultimately lead to the

334

transformation of the PAH molecules. In our study, B[a]P and Fe(III) may be

335

considered to be the strongest complex, which is supported by the highest values of

336

kobs (Table S2, Supporting Information).

337

The electron transfer process during PAH transformation on modified clays

338

was investigated by monitoring the valence states of surface metal ions after 10 d of

339

the interactions between ANT or B[a]P and clays. The results of the high resolution

340

X-ray photoelectron spectroscopy (XPS) measurements are presented in Figure S4

341

(Supporting Information). The peaks at binding energies of Fe 2p3/2 and Fe 2p1/2 are

342

presented in Figures S4a-c. The de-convolution of the Fe 2p3/2 peak implies two 15



343

different Fe(III) states therein, i.e., structural and intercalated Fe(III) species at Fe

344

2p3/2 714.8 and Fe 2p3/2 712 eV, respectively.46 After 10 d of transformation, the peak

345

at ~ 712 eV becomes relatively weak, while a new peak at ~ Fe 2p3/2 710.3 eV

346

appears, corresponding to Fe(II) (Figures S4b,c). The obtained results suggest the

347

reduction of Fe(III) to Fe(II) on the modified clay surfaces.

348

As shown in Figure S4 (Supplementary Information), the presence of Cu(II)

349

on clay surfaces could be confirmed by observing peaks of Cu 2p3/2 (934.0 eV) and

350

Cu 2p1/2 (953.8 eV). The satellite peaks at 940–945 eV also suggest Cu(II) on the

351

surface.47 The peaks seen at ~ 932 eV and ~ 952 eV can be ascribed to the existence

352

of Cu(I).48 The results of the peaks indicate that the Cu(II) accepts an electron from

353

adsorbates (i.e., PAHs) to yield Cu(I) on the clay surfaces (Figure S3d-f).

354

In case of Ni(II)- and Co(II)-modified clays, the peaks at ~ 857.1 eV and

355

784.2 eV become weak with concurrent growth of new peaks at ~ 855.5 eV and ~

356

782.5 for Ni(II) and Co(II), respectively (Figure S3g-l). This slight shift towards a

357

lower energy during PAH transformation indicates that some type of electron transfer

358

process has occurred without total conversion of these metal ions to zero-valent states.

359

An insignificant change in the XPS spectra was observed during the transformation of

360

ANT and B[a]P on Zn(II) ion-modified clay (Figure S3m-o). This was expected due

361

to the low electron-accepting ability of Zn(II), which results in relatively less electron

362

transfer from the PAHs.30

363 364

Formation of radicals.

Initially, the possible formation of a CT complex to 16


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365

produce a radical was explored by measuring the diffuse reflectance UV-visible

366

spectra of the interactions of ANT with Fe(III)-modified clay and B[a]P with

367

Cu(II)-modified clay (Figure S5, Supporting Information). The broad absorption band

368

at ~ 500 nm could be attributed to the CT complex formed between PAHs and active

369

sites of the clay surfaces.49 This CT complex might undergo ion-pair or radical-pair

370

collapse, followed by electron and proton loss, resulting in the formation of a radical

371

cation.50 The weak absorbance at ~ 750 nm in the UV-vis diffuse reflectance spectrum

372

affirms the formation of a radical cation.49 The produced radical gradually increases

373

to the point of the highest yield, then disappears as the reaction time progresses

374

(Figure S5). A water molecule which was sorbed on the clay surface may have acted

375

as a nucleophile to attack the radical cation, causing its disappearance with time.29

376

This possible reaction may have contributed to the formation of the final quinonyl

377

products.32

378

Formation of EPFRs from the interaction of PAHs with metal ion-modified

379

clays was directly observed by carrying out EPR measurements (Figure 2). No EPR

380

signals from PHE on any of the modified surfaces were seen (Figures 2a-e). This is in

381

agreement with the fact that no transformation of PHE was observed on these clay

382

surfaces (see Figures 1a-e). B[a]A and PYR also produce no significant EPR signals

383

(Figures 2a-e). In other words, the signals are too weak to be accurately identified.

384

However, both B[a]A and PYR are decayed on the metal-ion-modified clay surfaces,

385

except by Zn(II) (see Figures 1a-e). This suggests that free organic radicals may have

386

been formed in the presence of B[a]A and PYR, but if they were formed, they were 17



387

not stable on the clay surfaces. Comparatively, the transformation of ANT and B[a]P

388

results in EPFR formation on almost all the surfaces (Figures 2a-e), indicating the

389

stability of these free organic radicals formed in the presence of ANT and B[a]P. The

390

degradation rate of B[a]P was the highest on the Fe(III)-clay surface (see Figure 1a),

391

but no EPR signal from this transformation was observed (Figure 2a). It seems that

392

the radical formed from this transformation is not sufficiently stable to produce an

393

EPR signal from the B[a]P-Fe(III)-clay system. Comparatively, ANT on

394

Fe(III)-modified clay produces a strong EPR signal, which indicates the stability of

395

EPFRs from ANT. Figure 2 implies that the molecular structures of the PAHs play a

396

crucial role in stabilizing the EPFRs on the clay surfaces.

397

Formation of the EPFRs from the interactions of PAHs and metal

398

ion-modified clays may be understood by applying potential energy surface (PES)

399

calculations. The energy before the formation of the complex between the reactants

400

was set as zero. A typical example of the calculation is demonstrated for ANT (Figure

401

3), and corresponding optimized structures for the interaction of ANT and Fe species

402

are provided in Figure S6. All the results of PES under various systems are given in

403

Table S3. The interactions of PAHs with Fe(III) processes pass through the transition

404

state (TS), which has activation barriers of 41.16, 19.76, 22.91, 15.34, and 14.52 kcal

405

mol-1 for the S1→S2 step of the reaction systems associated with PHE, ANT, B[a]A,

406

PYR, and B[a]P, respectively (Table S3). In the case of PHE, the activation barriers

407

are higher than those for other PAHs (41.16 – 48.65 kcal mol-1); hence, no

408

transformations of PHE on the metal ion-modified clay surfaces were observed. 18


Page 18 of 36

Page 19 of 36


409

Comparatively, the activation barriers for other PAHs on the metal ion-modified clays

410

are very low and thus transformation could proceed easily, except in the case of the

411

Zn(II) ion. Significantly, the Zn(II)-clay is able to transform B[a]P because this step

412

has a relatively low activation barrier among the tested PAHs (28.26 for B[a]P, versus

413

38.00 – 48.65 kcal mol-1 for other PAHs) (Table S3).

414

The next step, S2→S3, corresponds to the CT complexes between PAHs and

415

metal ions, which have lower energy than the reactants expect the PHE-Zn(II)-clay

416

system (i.e., these reactions are exothermic, with barriers ranging from -31.57 to 1.07

417

kcal mol-1). The lowest barriers for S2→S3 are for B[a]P, which have the fastest

418

transformation first-order rate constants (Table S3). The next step is the SET within

419

the CT complex which results in the formation of radical cations and reduction of the

420

transition metal ions (Figure 3). The formed radical cations might be stabilized either

421

on the clay interlayer surface,27 or oxidized/hydrolyzed by H2O molecules,29 resulting

422

in the formation of other intermediate products. The stability of the intermediate

423

radicals may be correlated with their PES from S3 to S4 (S3→S4) (Table S3). This

424

step involves positive activation barriers, which varies from 5.85 to 67.39 kcal mol-1.

425

The activation barrier for the stability of the EPFR from the B[a]P-Fe(III)-clay system

426

is lower than that on the clays containing other metal ions (5.83 kcal mol-1 for Fe(III)

427

versus 22.63 – 31.53 kcal mol-1 for the other metal ions). The results suggest that the

428

B[a]P-type EPFR is less stable on the Fe(III)-clay surface than on other metal

429

ion-containing clay surfaces. In the case of ANT, the activation barrier for S3→S4 is

430

slightly higher for Fe(III) than for other metal ions (i.e., 28.36 kcal mol-1 for Fe(III) 19



431

versus 23.48 – 25.77 kcal mol-1 for Cu(II), Ni(II) and Co(II)), suggesting that the

432

ANT-type radical produced on the Fe(III)-clay surface is more stable than the

433

ANT-type radicals produced on other metal ion-modified clays. Finally, the

434

decomposition of the EPFRs to oxidized products (i.e., step S4→S5) occurs through

435

exothermic reactions with large negative barriers (-2.46 – -38.00 kcal mol-1) (Table

436

S3). The last step would therefore have been spontaneous.

437 438

Fate of radicals.

The peak areas of single EPR signals of EPFRs are presented

439

at different time intervals during the transformation of ANT and B[a]P on metal

440

ion-modified clays (Figure 4). Generally, the EPFR yields increase in the beginning,

441

and then gradually decrease with time. The highest yields of EPFRs from the

442

transformation of ANT on modified clays were observed at 8 d, 12 d, 21 d, and 23 d

443

for Fe(III), Cu(II), Ni(II), and Co(II) ions, respectively (Figure 4a). The amount of the

444

formed ANT-type EPFRs follow the order of Fe(III) > Cu(II) > Ni(II) > Co(II)). The

445

highest yields of the EPFRs formed on Fe(III)-clay were similar to 2 × higher than on

446

Cu(II)-clay and more than 5 × higher than on Ni(II)-clay. The highest EPFR yields

447

were at 9 d, 18 d, 31 d, and 45 d from B[a]P-contaminated clays of Cu(II), Ni(II),

448

Co(II), and Zn(II) ions, respectively (Figure 4b). The rates of the formation of EPRFs

449

to the highest yields have the similar order (i.e., Cu(II) > Ni(II) > Co(II) > Zn(II)).

450

This coincides with the transformation rates of the PAHs (see Figure 1). However, the

451

EPFR yields associated with B[a]P exhibit a reverse trend compared to

452

ANT-contaminated clays. The maximum yields of EPFRs, derived from the 20


Page 20 of 36

Page 21 of 36


453

B[a]P-contaminated clays, follow the order of Co(II) > Ni(II) > Cu(II) > Fe(III)

454

(Figure 4b). In addition, the yields of EPFRs from B[a]P on clays are generally

455

greater than the yields of free radicals from ANT (Figure 4a versus 4b).

456

The results of the peak areas of EPFRs on different clay surfaces also agree with

457

the g-factor values of the EPR signals (Figures 4c and 4d). The integrated g-factors

458

range from 2.0028 to 2.0039 initially, followed by a decrease with time for both

459

ANT- and B[a]P-contaminated metal ion-modified clays; similar to the trends seen in

460

the yields of the radicals. Those results suggest that the produced free organic radicals

461

with relatively high g-factor, such as benzoquinonyl radical, might have longer

462

lifetime (or persistence) and thus readily accumulated on clay surfaces compared to

463

the radicals with low g-factor, such as PAHs-type radical cations.29

464

Overall, the formation of free organic radical from PAHs, such as ANT and

465

B[a]P, proceeds more rapidly on clays saturated by Fe(III), followed by Cu(II), Ni(II),

466

Co(II), and Zn(II). This order correlates with the redox potential of metal ions. Higher

467

oxidation potential of transition metal ions such as Fe(III) and Cu(II) on mineral

468

surfaces potentially induce the accumulation to the maximum yields of EPFRs with

469

shorter reaction time compared to those produced over the Ni(II), Co(II), and Zn(II).

470

On the other hand, the relationship between EPFRs yields and redox potential of

471

active sites might be more complicated. As reported previously, the higher oxidation

472

potential of Fe2O3 result in greater decomposition of the adsorbates and lower EPFR

473

yields.23 Similar phenomenon was observed in reaction systems associated with B[a]P.

474

For ANT-type EPFRs, however, the bounding to relatively high oxidation potential 21



475

ions such as Fe(III) and Cu(II) enhance their yields compared to Ni(III) and Co(II).

476

The difference between two tested PAHs molecules might be due to their molecular

477

properties of PAHs and/or the different EPFRs being formed on clay surfaces.

478

Next, the fates of the radicals derived from ANT and B[a]P, shown in Figure 4,

479

were evaluated by the decays of various radicals after their formation of maximum

480

yields. Decay of various radicals on metal ion-containing clays fits well to first-order

481

kinetics (Figure S7, Supporting Information). The calculated values of first-order rate

482

constants (k, d-1) and life-times (t1/e) are given in Table S4. In the ANT-contaminated

483

clays, the 1/e life-times of the produced free radicals are 22.73 d, 21.28 d, 18.52 d,

484

and 11.76 d for Fe(III)-, Cu(II)-, Co(II)-, and Ni(II)-modified clays, respectively

485

(Table S4). Overall, the formed radicals have varied stability following the order as

486

Fe(III) > Cu(II) > Ni(II) > Co(II). These results suggest that ANT-type radicals

487

produced on Fe(III)-clay are more stable and hardly react with molecular species (i.e.,

488

H2O) compared to the same radicals interacting with other metal ions. On the other

489

hand, the relatively weak electron transfer between PAHs and surface metal ions with

490

low oxidation potential (e.g., Co(II)) may induce the formation of a relatively weak

491

CT complex, thus producing the formed radicals with relatively short lifetime. This

492

finding from the experimental observation of the EPR signals agrees well with the

493

PES calculation (see Table S3).

494

Significantly, the lifetimes of the of the B[a]P-type radicals are 13.70 d, 43.48

495

d, and 58.82 d for Cu(II)-, Ni(II)-, and Co(II)-modified clays, respectively (Table S4).

496

The presence of surface cations with higher oxidation potential accelerates the 22


Page 22 of 36

Page 23 of 36


497

transformation from B[a]P-type radicals to final products. Therefore, the stability of

498

the produced radicals of B[a]P-metal ions follows the order as Co(II) > Ni(II) > Cu(II),

499

which is opposite of the trend seen with the ANT-type radicals. This suggests that the

500

stability of radicals produced on modified clays depends not only on the type of metal

501

ions, but also on the molecular structure of the PAH. The dependence of the stability

502

of EPFRs on the structure of the parent organic molecule was also observed by other

503

researchers in combustion-related studies using phenol, chlorophenol, hydroquinone,

504

and catechol as organic contaminants on metal oxide surfaces.22-24 Overall, the

505

persistence of the EPFRs is determined by the properties of the precursor molecules

506

and/or the formed radicals.17,24 Also, other factors affecting radical persistence of the

507

EPFRs include the environmental conditions and reactivity of the formed radicals

508

toward molecular species such as H2O or/and oxygen.33

509 510

Environmental significance

511

Sites that are generally co-contaminated by PAHs and toxic metals include

512

coking plants, manufactured gas plants, and petrol stations. Atmospheric particulate

513

matter may also exhibit similar co-contamination. Clay minerals, represented here by

514

montmorillonite clay, are important components of soil and act as a major metal

515

repository and a sorbent of organic contaminants. On clay surfaces, the presence of

516

toxic metals, especially some transition metal ions, may affect the fate and toxicity of

517

organic contaminants, including PAHs, under various environment conditions. For

518

example, this study demonstrates the significant role of transition metal ions, enriched

23



519

on clay mineral surfaces, in the formation and fate of PAH-induced EPFR

520

intermediates under environmentally relevant conditions. This study suggests that the

521

interactions of PAHs and transition metal ion-modified clays generate radicals of

522

different stabilities and fate, depending on the molecular configuration of the organic

523

molecule and type of metal ions. It is well known that these EPFR-containing

524

minerals might cause adverse health effects, including increased susceptibility to

525

respiratory diseases, cardiopulmonary disease, and influenza virus infection, via

526

oxidative stress to humans.51,52 Similarly, the ecotoxicological effects in the soil

527

environment would also change with type of organic contamination, redox property of

528

metal ions, and time of exposure.

529 530

ASSOCIATED CONTENT

531

Supporting Information

532

The supporting information is available free of charge on the ACS Publications

533

website.

534

Chemicals and materials, concentration of metal ions in original and modified clays,

535

XRD analysis of clay samples, influence of sampling on PAHs transformation, HPLC

536

and GC-MS analyses, measurements of EPR, UV-visible spectroscopy, and XPS,

537

extraction efficiencies of PAHs in contaminated clay samples, first-order rate

538

constants (kobs, d-1) for PAHs decay, values of PES for DFT, 1/e life-times of EPFR

539

signals, correlation between rate constants and oxidation potential, and decay of EPR

540

signals from ANT and B[a]P. 24


Page 24 of 36

Page 25 of 36


541 542

ACKNOWLEDGMENTS

543

Financial support by the National Natural Science Foundation of China (41571446),

544

and the CAS Youth Innovation Promotion Association (2016380) are gratefully

545

acknowledged. We thank Dr. Leslie Cizmas for her comments to improve the paper.

546

We also thank anonymous reviewers for their comments, which improved the paper

547

greatly.

548 549

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699 700

42. Eastman, M.P. Reaction of Benzene with Cu(II)- and Fe(III)-exchanged hectorites. Clay Clay Miner. 1984, 32 (4), 327-333; 10.1346/ccmn.1984.0320411

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705 706 707

44. Tanimoto, I.; Kushioka, K.; Kitagawa, T. Maruyama, K. Binary phase chlorination of aromatic hydrocarbons with solid copper(II) chloride: reaction mechanism. Bull. Chem. Soc. Jpn. 1979, 52 (12), 3586-3591; 10.1246/bcsj.52.3586.

708 709 710

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711 712 713 714

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722 723 724 725

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736 737 738 739 740 741

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Page 30 of 36

Page 31 of 36


742

Caption of Figures

743 744 745 746 747 748 749 750 751 752 753 754 755 756 757 758 759 760 761 762 763

Figure 1. The degradation of selected polycyclic aromatic hydrocarbons (PAHs) on transition metal ion-modified montmorillonite clays as a function of time. (a) Fe(III); (b) Cu(II); (c) Ni(II); (d) Co(II); and (e) Zn(II). (Phenanthrene (PHE), anthracene (ANT), benzo[a]anthracene (B[a]A), pyrene (PYR), and benzo[a]pyrene (B[a]P). Figure 2. Electron paramagnetic resonance (EPR) spectra obtained from the interactions of polycyclic aromatic hydrocarbons (PAHs) with transition metal ion-modified montmorillonite clays after a 18 d reaction period for ANT, PHE, B[a]A, and PYR on various clays, and after 18 d, 3 d, 18 d, and 50 d reaction period for B[a]P-Cu(II)-clay, B[a]P-Ni(II)-clay, B[a]P-Co(II)-clay, and B[a]P-Zn(II)-clay system, respectively. (a) Fe(III); (b) Cu(II); (c) Ni(II); (d) Co(II); and (e) Zn(II). (Phenanthrene (PHE), anthracene (ANT), benzo[a]anthracene (B[a]A), pyrene (PYR), and benzo[a]pyrene (B[a]P)). Figure 3. Profile of the interaction of anthracene (ANT) with hydrated metal ions. The energies of ANT complexes with hydrated metal ions were set to zero. Figure 4. The evolution of electron paramagnetic resonance (EPR) peak area as a function of reaction time on metal ion-modified clay surfaces contaminated by (a) anthracene (ANT) and (b) benzo[a]pyrene (B[a]P). Variation of g-factor with reaction time on metal ion-modified clay surfaces contaminated by (c) ANT and (d) B[a]P.

764

765

31



Page 32 of 36

Figure 1

766 Cu(II)

[PAH]/[PAH]0

Fe(III) 1.0

1.0

0.8

0.8

0.6

0.6

0.4

0.4 0.2

0.2

(a)

0.0 0

2

4

6

8

(b) (b)

0.0 0

10

2

Time, d

6

8

10

[PAH]/[PAH]0

1.0

0.8 0.6 0.4 0.2

(c)

0.0 0

0.8 0.6 0.4 0.2

(d)

0.0 2

4

6

8

10

0

Time, d

2

4

6

Time, d

Zn(II) 1.0

[PAH]/[PAH]0

10

Co(II)

1.0

0.8

PHE ANT B[a]A PYR B[a]P

0.6 0.4 0.2

(e)

0.0 0

767

8

Time, d

Ni(II)

[PAH]/[PAH]0

4

2

4

6

8

10

Time, d

768

769

770 32


Page 33 of 36


Figure 2

771

Fe(III)

Cu(II) (b)

Intensity

Intensity

(a)

3250

3300

772

3350

3400

3450

3250

3300

3350

Field, G

Field, G

Ni(II)

Co(II)

3400

(d)

Intensity

Intensity

(c)

3250

3300

3350

3400

3250

3450

3300

Field, G

Field, G

773

3350

Zn(II) (e)

Intensity

PHE ANT B[a]A PYR B[a]P

3250

3300

3350

3400

3450

Field, G

774

775

776 33


3400

3450


Page 34 of 36

Figure 3

777

S4

Relative energy (kcal/mol)

S2

(H2O)6

n+

+ M (H2O)6

M

H OH

n+

CT complex

(n-1)+

M

+ OH-/H2O

(H2O)5

S1

n+

M (H2O)6

S3 TS

778

779

780

781

782

783

784

785

786

787

788

34


S5 M(n-1)+(H2O)6

Page 35 of 36


Figure 4

789

18

12 (a)

(b)

16

10

Spin(× 1016)/g

14 Fe(III) Cu(II) Ni(II) Co(II)

8 6

Cu(II) Ni(II) Co(II) Zn(II)

12 10 8

4

6 4

2

2 0

0 0

10

20

30

40

50

0

60

20

40

2.0040

g-Factor

2.0038

Fe(III) Cu(II) Ni(II) Co(II)

(c)

2.0040 2.0038

2.0036

2.0036

2.0034

2.0034

2.0032

2.0032

2.0030

2.0030

2.0028

2.0028

2.0026

2.0026 1

790

8

13

22

60

80

100

Time, d

Time, d

25

Cu(II) Ni(II) Co(II) Zn(II)

1

45

Time, d

(d)

9

18

31

Time, d

791 792 793 794 795 796 797

798 799

35


50

90


84x38mm (290 x 290 DPI)


Page 36 of 36

Transformation of Polycyclic Aromatic Hydrocarbons and Formation of

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