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16 Apr 2018 - (e.g., catechol and dibenzofuran).11,12 The formation of EPFRs and their ecotoxicological effects in the natural environment have attrac...
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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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Transformation of PAHs and Formation of Environmentally

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Persistent Free Radicals on Modified Montmorillonite: Role of

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Surface Metal Ions and PAH Molecular Properties

4 5 6

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

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Virender K. Sharmac*

8 9 10 a

11

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

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Ministry of Agriculture, College of Natural Resources and Environment,

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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,

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College Station, TX 77843, USA.

19 20 21 22 23 24

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

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ABSTRACT

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

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

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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

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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

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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

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

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

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

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

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TOC Art

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INTRODUCTION

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

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

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

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

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

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

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chlorophenol and transition metal ions emitted from combustion sources in the

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atmosphere produces environmentally persistent free radicals (EPFRs) that may

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increase the human health risk of developing respiratory and cardiopulmonary

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

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

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

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

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

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

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on

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organic-contaminated

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

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systems to characterize the electron transfer from organic molecules, especially

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

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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

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and

metal

oxides

on

stabilization

of

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

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room temperature and also under environmental conditions.28-32

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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information is needed regarding the influence of the electronic properties and

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

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

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

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

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EPFR formation and fate on soil mineral surfaces.

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

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

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

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

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

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

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

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

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

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

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

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

138 139

EXPERIMENTAL SECTION

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

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

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

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

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

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Preparation of PAH-contaminated clays. The preparation of polycyclic

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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using AgNO3. 7

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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clays.

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The

prepared

clay

samples

(metal

ion-modified

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clays

and

metal

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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interlayers disordered somewhat the crystalline structure. Those results indicated that

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the PAHs was intercalated into the clay interlayers successfully.

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PAH transformation and products analysis. After the preparation of

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

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

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

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

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

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

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

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Supplementary Information).

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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in amber vials for analyzing by either high performance liquid chromatography

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

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

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

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

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parameters are detailed in Text S5 (Supporting Information).

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

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

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

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

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samples (Text S5, Supporting Information).

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

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

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proposed reaction pathway using the Materials Studio 6.0 of Dassault Systèmes

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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RESULTS AND DISCUSSION Decay of PAHs.

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

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

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

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

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

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

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

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

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

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

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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

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experiments were conducted using the original montmorillonite clay, saturated mainly

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

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

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

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the surface metal ions on the clays.

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

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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

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

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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

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

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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

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

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

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

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

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transformation of PAH readily.44,45

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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

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

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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

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

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

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

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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

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

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

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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

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

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

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

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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

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

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

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

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

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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

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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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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

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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

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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

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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

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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

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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

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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

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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

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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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35. Helsen J A. Characterization of Iron(II)- and Iron(III)-exchanged montmorillonite and hectorite using the mössbauer effect. Clay Miner 1983, 18 (2), 117-125; 10.1180/claymin.1983.018.2.01.

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36. Wasserman S R, Soderholm L, Staub U. Effect of surface modification on the interlayer chemistry of iron in a smectite clay. Chem. Mater. 1998, 10 (2), 559-566; 10.1021/cm9705597.

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37. Gu, C., Liu, C., Johnston, C. T., Teppen, B. J., Li, H., Boyd, S. A. Pentachlorophenol radical cations generated on Fe(III)-montmorillonite initiate octachlorodibenzo-p-dioxin formation in clays: density functional theory and fourier transform infrared studies. Environ. Sci. Technol. 2011, 45 (4), 1399-1406; 10.1021/es103324z.

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38. Becke, A.D. Density-functional thermochemistry. III. The role of exact exchange. J. Chem. Phys. 1993, 98 (7), 5648-5652; 10.1063/1.464913.

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39. Perdew, J.P.; Burke, K. Ernzerhof, M. Generalized gradient approximation made simple. Phys. Rev. Lett. 1996, 77 (18), 3865-3868; 10.1103/PhysRevLett.77.3865.

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40. Henkelman, G.; Uberuaga, B.P. Jonsson, H. A climbing image nudged elastic band method for finding saddle points and minimum energy paths. J. Chem. Phys. 2000, 113 (22), 9901-9904; 10.1063/1.1329672.

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41. Svanholm, U.; Hammerich, O. Parker, V.D. Kinetics and mechanisms of the reactions of organic cation radicals and dications. II. anisylation of thianthrene cation radical. J. Am. Chem. Soc. 1975, 97 (1), 101-106; 10.1021/ja00834a018.

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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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43. Li, H.; Pan, B.; Liao, S.; Zhang, D. Xing, B. Formation of environmentally persistent free radicals as the mechanism for reduced catechol degradation on hematite-silica surface under UV irradiation. Environ. Pollut. 2014, 188, 153-158; 10.1016/j.envpol.2014.02.012.

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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.

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45. Li, L.; Jia, H.; Li, X. Wang, C. Transformation of anthracene on various cation-modified clay minerals. Environ. Sci. Pollut. R. 2015, 22 (2), 1261-1269; 10.1007/s11356-014-3424-4.

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46. Huang, Z.; Wu, P.; Li, H.; Li, W.; Zhu, Y. Zhu, N. Synthesis and catalytic properties of La or Ce doped hydroxy-FeAl intercalated montmorillonite used as heterogeneous photo Fenton catalysts under sunlight irradiation. RSC Adv. 2014, 4 (13), 6500-6507; 10.1039/c3ra46729e.

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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

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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

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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

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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

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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

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