Experimental and theoretical insight into hydroxyl and sulfate radicals-mediated degradation of carbamazepine

Journal Pre-proof Experimental and theoretical insight into hydroxyl and sulfate radicals-mediated degradation of carbamazepine Ruiyang Xiao, Junye Ma, Zonghao Luo, Zongsu Wei, Richard Spinney, Wei‒Ping Hu, Dionysios D. Dionysiou, Weizhi Zeng PII:

S0269-7491(19)34007-2

DOI:

https://doi.org/10.1016/j.envpol.2019.113498

Reference:

ENPO 113498

To appear in:

Environmental Pollution

Received Date: 21 July 2019 Revised Date:

13 October 2019

Accepted Date: 25 October 2019

Please cite this article as: Xiao, R., Ma, J., Luo, Z., Wei, Z., Spinney, R., Hu, Wei‒Ping., Dionysiou, D.D., Zeng, W., Experimental and theoretical insight into hydroxyl and sulfate radicals-mediated degradation of carbamazepine, Environmental Pollution (2019), doi: https://doi.org/10.1016/j.envpol.2019.113498. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.

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Experimental and theoretical insight into hydroxyl and sulfate radicals-mediated degradation of carbamazepine

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Ruiyang Xiao†,⊥, Junye Ma†,⊥, Zonghao Luo†,⊥, Zongsu Wei§, Richard Spinney‡, Wei‒

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Ping Hu∆, Dionysios D. Dionysiouǁ , and Weizhi Zeng†,⊥,*

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†

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Central South University, Changsha, 410083, China

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⊥Chinese

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Metal Pollution, Changsha, 410083, China

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§

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Aarhus University, Hangøvej 2, DK-8200, Aarhus N, Denmark

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‡

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Ohio, 43210, U.S.A.

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∆

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Yi 62102, Taiwan

20

ǁ

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Cincinnati, Ohio, 45221, U.S.A.

Institute of Environmental Engineering, School of Metallurgy and Environment,

National Engineering Research Center for Control & Treatment of Heavy

Section for Biological and Chemical Engineering, Department of Engineering,

Department of Chemistry and Biochemistry, the Ohio State University, Columbus,

Department of Chemistry and Biochemistry, National Chung Cheng University, Chia‒

Environmental Engineering and Science Program, University of Cincinnati,

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*To whom correspondence should be addressed. W. Zeng. Phone: +86‒731‒

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88830875; fax: +86‒731‒88710171; Email address: [email protected] 1

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Abstract

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Carbamazepine (CBZ), a widely detected pharmaceutical in wastewaters, cannot

28

currently be treated by conventional activated sludge technologies, as it is highly

29

resistant to biodegradation. In this study, the degradation kineitcs and reaction

30

mechanisms of CBZ by hydroxyl radical (•OH) and sulfate radical (SO• 4 )–based

31

advanced oxidation processes (AOPs) were investigated with a combined

32

experimental/theoretical approach. We first measured the UV absorption spectrum of

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CBZ and compared it to the theoretical spectrum. The agreement of two spectra

34

reveals an extended π–conjugation system on CBZ molecular structure. The second–

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order rate constants of •OH and SO• 4 with CBZ, measured by competition kinetics

36

method, were (4.63 ± 0.01) × 109 M−1 s−1 and (8.27 ± 0.01) × 108 M−1 s−1, respectively

37

at pH 3. The energetics of the initial steps of CBZ reaction with •OH and SO• 4 were

38

also calculated by density functional theory (DFT) at SMD/M05–2X/6–

39

311++G**//SMD/M05–2X/6–31+G**level. Our results reveal that radical addition is

40

the dominant pathway for both •OH and SO• 4 . Further, compared to the positive

41

∆G0R value for the single electron transfer (SET) reaction pathway between CBZ and OH, the ∆G0R value for SET reaction between CBZ and SO• 4 is negative, showing

42

•

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that this reaction route is thermodynamically favorable. Our results demonstrated the

44

remarkable advantages of AOPs for the removal of refractory organic contaminants

45

during wastewater treatment processes. The elucidation of the pathways for the

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reaction of •OH and SO• 4 with CBZ are beneficial to predict byproducts formation

47

and assess associated ecotoxicity, providing an evaluation mean for the feasibility of

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AOPs application.

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Capsule: 2

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9 The k values of •OH and SO• 4 reacting with CBZ were measured to be 4.63 × 10

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and 8.27× 108 M−1 s−1, respectively. Radical addition is dominant reaction pathway.

53 54

Keywords: carbamazepine; hydroxyl radical; sulfate radical; advanced oxidation

55

processes; wastewater treatment; DFT

3

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1. Introduction

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Carbamazepine (CBZ), a widely consumed anticonvulsant pharmaceutical, is one

58

of the most commonly detected emerging organic pollutants in different waters

59

(Ensano et al., 2017; Fekadu et al., 2019; Vernouillet et al., 2010b). For example, in

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the Southwestern U.S., CBZ concentration up to 610 ng L−1 was detected in the

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groundwater, and 18 ng L−1 in drinking water (Benotti et al., 2009; Drewes et al.,

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2002). Although the detected concentrations of pharmaceuticals in water bodies are

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typically in the range of ng L−1 to µg L−1, such relatively low levels of

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pharmaceuticals were not safe to humans and aquatic organisms (Nassef et al., 2010;

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Tomas et al., 2014; Vernouillet et al., 2010b). CBZ was reported to be toxic to various

66

aquatic organisms such as green algae, crustacean, cnidarian, and Hydra attenuate

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(Vernouillet et al., 2010a).

68

Yet, this anthropogenic contaminant shows a strong resistance to conventional

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wastewater treatment technologies, such as biological filtration and activated sludge

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(Clara et al., 2004; Lam and Mabury, 2005; Roberto et al., 2002; Wei et al., 2019).

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Many studies reported that concentrations of CBZ in influent wastewaters ranged

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from 54 ng L‒1 to 1694  ng L‒1, but the removal efficiency was typically below 10%

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(Kong et al., 2009; Nakada et al., 2006; Zhang et al., 2008). Therefore, alternative

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wastewater treatment technologies such as advanced oxidation processes (AOPs) are

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employed to further remove CBZ in wastewaters.

76

AOPs have been proven to be a promising method to convert recalcitrant organic

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contaminants to less harmful compounds or even completely mineralize to CO2 and

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H2O (De la Cruz et al., 2012; Liu et al., 2013; Xiao et al., 2019). AOPs involve the

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generation of radical species such as hydroxyl radicals (•OH), which induce the

80

degradation of contaminants with nearly diffusion controlled rates (i.e., 109 to 1010

4

81

M−1 s−1) (Haag and Yao, 1992). For example, Ali et al. investigated the degradation

82

efficiency of 21 µM CBZ in both UV and UV/H2O2/Fe2+ systems (Ali et al., 2018).

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They reported that at the UV fluence of 3600 mJ cm−2, 7.5% and 90.6% of CBZ

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removal were achieved in the UV and UV/H2O2/Fe2+ system ([H2O2] = 17.9 µM,

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[Fe2+] = 1.06 mM, pH =3), respectively. On the other hand, sulfate radical anion

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(SO• 4 ) generated by persulfate (PS) or peroxymonosulfate (PMS) has drawn a great

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deal of attention over the last two decades. It was considered to be an excellent

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alternative to •OH due to its longer half‒life in waters (t1/2 = 30~40 µs) and high redox

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potential (E0 = 2.5~3.1 V vs. the normal hydrogen electrode) (Devi et al., 2016;

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Ghanbari and Moradi, 2017).

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The degradation kinetics of organic contaminants based on •OH and SO• 4 have

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been intensively investigated (Gao et al., 2014; Gao et al., 2016; Khan et al., 2014;

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Khan et al., 2017; Wang et al., 2018; Xiao et al., 2017). For example, Mercado et al.

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determined the rate constants of •OH and SO• 4 in reacting with flusilazole, an

95

organosilicon fungicide, using laser flash photolysis. They concluded that •OH and

96

SO• exhibited similar reactivity in the degradation of this kind of fungicide 4

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(Mercado et al., 2018). Similarly, Matta et al. compared the degradation efficiency of

98

II • CBZ by SO• 4 generated from a PMS/Co system and OH generated from the

99

Fenton’s reagent (H2O2/FeII). The results revealed that SO• 4 was more selective than

100

•

101

et al., 2011).

102 103

OH for the degradation of organic contaminant in the urban wastewater matrix (Matta

Both •OH and SO• 4 co-exist in a UV/PS system. The mechanism of the coexistence can be explained as follows: hv/heat

104 105

• • S2 O2 8  SO4 + SO4 2 2 • SO• 4 + S2 O8 → SO4 + S2 O8

5

(1) (2)

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 • SO• 4 + H2O → HSO4 + OH

(3)

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− •  SO• 4 + HO → HSO4 + OH

(4)

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•  SO• 4 + RH → HSO4 + R

(5)

•

(6)

109

OH + RH → H2O + R•

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• The co–existence of SO• 4 and OH in this system was confirmed by various probe

111

compounds, such as benzene, benzoic acid, anisole and 4−nitroaniline in various

112

studies (Anipsitakis and Dionysiou, 2004; Lindsey et al., 2000; Zhang et al., 2015).

113

However, it is still particularly difficult to verify the specific mechanisms of •OH and

114

SO• 4 in oxidizing organic contaminants.

115

The possible pathways for the reactions of CBZ with •OH and SO• 4 are radical

116

addition, H−abstraction, and single electron transfer (SET). For the radical addition

117

mechanism, free radicals add onto the unsaturated moiety of CBZ, forming a transient

118

radical. For the H−abstraction mechanism, free radicals abstract a hydrogen atom

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from CBZ. For the SET mechanism, CBZ provides an electron to •OH/SO• 4 , forming

120

a radical cation. Density functional theory (DFT) is reckoned to be a compelling

121

means for studying radical and non‒radical bimolecular reaction mechanisms

122

(Daisuke and John, 2011; Villamena et al., 2007; Yang et al., 2017). For example,

123

Yang et al. investigated the thermodynamic and kinetic behaviors for reactions of

124

neutral ibuprofen (IBU) with •OH and SO• 4 using M06–2X functional with 6–

125

311 ++G** basis set (Yang et al., 2017). Their result revealed that H–atom abstraction

126

was the most favorable pathway for both •OH and SO• 4 , but due to the steric

127

hindrance, SO• 4 exhibited significantly higher energy barriers by, on average, 4.78

128

kcal mol‒1. Similarly, Galano and Alvarez‒Idaboy studied the mechanisms and

129

kinetics of free radicals (e.g. •OH, HO•2 , •OCH3, and •OOCH3) in degrading

130

glutathione using M05–2X functional and 6‒311++G** basis set with the Solvation 6

131

Model based on Density (SMD) (Galano and Alvarezidaboy, 2011). The values of the

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overall rate constant of glutathione with HO•2 was calculated to be 2.69 × 107 M−1 s−1,

133

which demonstrated that glutathione can be used as exceptionally good as HO•2

134

scavenger due to its strong H−bonding interactions with the radical species at the

135

transition state (TS). Although previous studies have investigated the reactions between CBZ and

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OH/SO• 4 , most of them focused on removal efficiencies in engineered waters and

137

•

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radical‒mediated kinetic degradation process of CBZ via experimental means (Ali et

139

al., 2018; Deng et al., 2013; Zhang et al., 2008). There is lack of theoretical evidence

140

to verify the reliability of their experimental observations. More importantly, they

141

cannot elucidate specific reaction mechanisms and the dominant pathways. In this

142

study, we combined an experimental approach and theoretical one to study the

143

kinetics and thermodynamics of CBZ degradation mechanisms by both •OH and

144

SO• 4 , which were generated by UV photolysis of H2O2 and PS, respectively. We

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tested the hypothesis that both •OH and SO• 4 exhibit the similar reactivity with CBZ.

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We experimentally measured k values of CBZ reacting with •OH and SO• 4 using the

147

relative rate technique. We also theoretically studied these reactions using a DFT

148

method. The relevant reactants, products, TS, and intermediate species on the

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potential energy surfaces for all reactions were analyzed. Overall, we aim to provide

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mechanistic insight at the molecular level for CBZ oxidation with •OH and SO• 4 ,

151

which could extend the application of •OH and SO• 4 ‒based AOPs for removing

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recalcitrant organic contaminants in wastewaters.

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2. Materials and methods

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2.1 Materials 7

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CBZ (99%), Na2S2O8 (99%), p–chlorobenzoic acid (p–CBA, 99%), H3PO4

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(85~90%), Na2HPO4 (99%), and NaH2PO4 (99%) and tert–butanol (TBA) (99.7%)

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were purchased from Sigma Aldrich. H2O2 (30% by weight), H2SO4 (guaranteed

159

reagent), KMnO4 (analytical grade), and Na2C2O4 (analytical grade) were purchased

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from Sinopharm Chemical Reagent, China. These chemicals were used as received

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without further purification. It should be noted that, TBA, as a •OH scavenger, was

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added to exclude the influence of •OH in the UV/PS system when measuring the k

163

value of the reaction between CBZ and SO• 4 (Liu et al., 2016; Shah et al., 2013).

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Sample solutions were prepared by deionized water from A Milli–Q water purification

165

system (Molecular, 1010A). Solution pH was measured by a S220 pH meter (Mettler,

166

Toledo).

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2.2 Photochemical experiments

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The photochemical reactor used in this study was described in a great detail in

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our previous studies (Gao et al., 2019a; Luo et al., 2018b; Luo et al., 2018c; Xiao et

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al., 2017; Yang et al., 2017). The low pressure UV lamp was warmed up for 30 min to

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ensure the stable UV emission. The reaction solution temperature and UV lamp

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temperature were maintained at 20 ± 0.1 °C via a water circulating system (SC150–

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A25B, Thermo Fisher Scientific). The chemical photometry with potassium ferric

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oxalate was used to measure the average light intensity per volume (I0) in the UV

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reactor, and I0 was determined to be 2.50 × 10–6 Einstein L–1 s–1 (Hatchard and Parker,

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1956; Parker, 1953). Meanwhile, the effective optical path length (b) was determined

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to be 1.42 cm by H2O2 actinometry method (Beltran et al., 1995; Xiao et al., 2015).

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The photochemical degradation tests with 10 µM CBZ and 10 µM p–CBA were

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conducted in a 50 mL quartz tube with a lid. To ensure homogenous reactions in the

8

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reactor, a Teflon stirrer was used. The solution pH was adjusted to 3.0 and maintained

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with 10 mM phosphate buffer. The reason for choosing this pH is that SO• 4 is the

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dominant radical species in a UV/PS system at pH 3 (Criquet et al., 2010; Zhang et al.,

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2016). To quantify the concentration during the degradation course, a 1 mL sample

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was withdrawn periodically (e.g., 0, 5, 10, 15, 20, and 25 min) from the irradiated

186

solutions. Control experiments (i.e., degradation kinetics with UV alone) were

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conducted in parallel to the UV/H2O2 and UV/PS experiments. The experiments were

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conducted in triplicate.

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2.3 Analytical method

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Concentrations of H2O2 and PS were determined by the KMnO4 titration method

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(Greenspan and Mackellar, 1948; Kiassen et al., 1994; Razmi and Mohammad-Rezaei,

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2010). An ultra–performance liquid chromatography (UPLC, Waters ACQUITY H–

194

Class) was used to quantify the concentration of the CBZ and p–CBA. The column

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used to separate CBZ was a reverse phase BEH C18 column (2.1 mm × 50 mm, 1.7

196

µm, Waters). The injection volume of sample was 10 µL and the column temperature

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was maintained 35 °C. The mixture of phosphoric buffer (20 mM at pH 3) and

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acetonitrile, at a ratio of 68:32, was used as mobile phase at a flow rate of 0.3 mL

199

min–1. Similarly, the concentrations of p–CBA were also detected using UPLC with a

200

UV detector. The mobile phase was a 77:23 mixture of phosphoric buffer (20 mM at

201

pH 3) and acetonitrile at the same flow rate. The UV wavelengths were set at 284 nm

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and 238 nm for CBZ and p–CBA, respectively. In addition, a UV–1800 spectrometer

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(Shimadzu, Japan) was used to measure the absorption spectrum of CBZ from 200 to

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400 nm.

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9

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2.4 Computational method

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The DFT approach was used to calculate the thermodynamic and kinetic

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parameters of different reaction pathways between CBZ and •OH/SO• 4 . Due to

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multiple accessible conformations of CBZ, Spartan’10 with the MMFF force field

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was used for searching the minimum-energy conformation (Halgren, 1996; Hosoi et

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al., 2015; Obot et al., 2013). Previous studies showed that the possible first steps of

212

•

213

abstraction, and SET, and these pathways occurred in parallel to different extents

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(Neta et al., 1977a; Norman et al., 1970; Ramirez-Arizmendi et al., 2001; Yang and

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Tanner, 1987). Thus, to elucidate the mechanisms of •OH and SO• 4 reacting with

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CBZ, we compared the energy profiles of these pathways.

OH and SO• 4 oxidation with organic contaminants were via radical addition, H‒

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The geometries of the reactants, products and transition states (TS) were then

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optimized at M05‒2X/6‒31+G** level of theory using Gaussian 09 (Revision A.01)

219

(Frisch et al., 2009; Luo et al., 2018a; Zhao and Truhlar, 2008). Single point energy

220

calculation were performed at the M05‒2X/6‒311++G** level of theory with the

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SMD continuum solvation model (Galano and Alvarezidaboy, 2011; Wu et al., 2017).

222

The absorption spectra of CBZ were calculated using the time‒dependent DFT

223

(TDDFT). Considering the influence of specific solvent interactions with CBZ, one

224

explicit water molecule as part of the CBZ structure was included (Campillo et al.,

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2004; Huber et al., 2005). First, the B3LYP/6‒31+G** method was applied to

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optimize structures of the ground state CBZ molecule (Becke, 1988; Calais, 1993).

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Next, the UV absorption spectra of CBZ, specifically, the excitation energies and

228

oscillator strengths, were calculated at the SMD/TD-B3LYP/6‒311++G** level of

229

theory.

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10

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3. Results

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3.1 Comparison of experimental and calculated UV spectrum of CBZ

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Figure 1 depicts the experimental and calculated UV molar absorption

234

coefficients (ε) of CBZ as a function of wavelength (λ). The ε was determined by

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measuring the absorbance (A) of 10 µM CBZ solution at pH 3 with a 1 cm path length

236

(z) quartz cuvette via: ε = A / (z × [CBZ])

237

(7)

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It should be noted that the pKa value of CBZ is reported to be 13.9 at 25 °C, showing

239

that CBZ does not deprotonate in environmental relevant pH (Jones et al., 2002). As is

240

shown in Figure 1, the theoretical UV absorption spectra agree with experimental data

241

reasonably well. For the experimental spectra (blue), two absorption maxima appear

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at 210 nm and 284 nm with ε value of (3.06 ± 0.01) × 104 and (1.25 ± 0.01) × 104 M–1

243

cm–1, respectively. The ε was determined to be (6.00 ± 0.06) × 103 M–1 cm–1 for CBZ

244

at the wavelength of 254 nm, which was in a good agreement with values of 6025 M–1

245

cm–1 and 6072 M–1 cm–1 reported in previous study (Kim et al., 2009; Vogna et al.,

246

2004). Thus, it is expected that direct photolysis of CBZ by UV light at 254 nm is not

247

significant.

248

For the calculated UV spectra (red), the first absorption peak is present at 223

249

nm with ε value of 2.33 × 104 M–1 cm–1. An additional smaller peak is present at 225

250

nm. The other major absorption peak is at 309 nm with ε value of 1.56 × 104 M–1 cm–1.

251

Compared with the experimental spectra, the absorption peaks on the calculated UV

252

spectrum are red‒shifted by about 10~20 nm. The two absorption peaks can be

253

attributed to π–π∗ electronic transitions in benzene ring and n–π∗ electronic transitions

254

(O and N) based on conjugated structure of CBZ molecular. The absorption peak at

255

309 nm corresponds to the π–π∗ electronic transition from conjugated phenyl ring in 11

256

CBZ. This corresponds to a transition from the HOMO (Highest Occupied Molecular

257

Orbital) to the LUMO (Lowest Unoccupied Molecular Orbital) (Figure 2). There is

258

one lone pair electrons in –NH2 and a two more lone pairs available on the carbonyl

259

oxygen atom (C=O). The peaks at 223 and 225 both correspond to π=π* transitions,

260

but due to limitations of the basis sets used it is difficult to tell which peak

261

corresponds to which specific lone pair. The two peaks together do however match

262

with nicely the experimental UV spectra which show a large peak at 210 nm and a

263

shoulder at ~230 nm.

264 265

3.2 Degradation of CBZ in different systems

266

Figure 3 illustrates the degradation kinetics of CBZ in UV, UV/H2O2 and UV/PS

267

systems. The degradation data in these systems were fit to a pseudo–first–order

268

kinetic model. CBZ was highly resistant to direct UV photolysis (purple), and only 8%

269

of CBZ degradation was achieved after 25 min irradiation. This is expected since the

270

two absorption maxima peaks for CBZ appear at 210 nm and 284 nm (Figure 1). The

271

initial direct UV photolysis degradation rate (d[C]/dt|0) of CBZ under UV irradiation

272

was 0.031 µM min−1 under conditions of UV intensity of 2.50 × 10−6 Einstein L−1 s−1.

273

We also calculated the quantum yield (φ) of CBZ, which represents the ability of

274

a compound to utilize photons (Bolton and Stefan, 2002). The φ value can be

275

determined by the d[C]/dt|0 under condition of direct UV photochemical degradation:

276

φCBZ =

d[C]/dt| I0 × (1 – 10–εCBZ b[CBZ] )

(8)

277

where φCBZ (mol Einstein–1) is the quantum yield of CBZ at 254 nm, εCBZ is the molar

278

absorption coefficient of CBZ (5995 M–1 cm–1), and I0 is the incident UV intensity

279

(2.50 × 10–6 Einstein L–1 s–1). The φ value of CBZ was then calculated to be 1.16 ×

280

10–3 mol Einstein–1. The result supported our claim that, direct UV photolysis at 254 12

281

nm exhibited little effect on the degradation of CBZ.

282

As comparison, it was found that the degradation kinetics of CBZ was

283

significantly enhanced in the presence of 100 µM H2O2 (green) and PS (blue) (Figure

284

3). Thus, it can be concluded that the degradation of CBZ was largely due to

285

•

286

systems were determined to be 7.33 × 10−2 and 7.50 × 10−2 µM min−1, respectively.

287

Oxidation mediated by •OH/SO• 4 is the dominant mechanism of CBZ degradation.

OH/SO• 4 in the system. The d[C]/dt|0 values of CBZ in the UV/H2O2 and UV/PS

288 289

3.3 Measurements of k values of CBZ with •OH and SO• 4

290

The relative rate technique was used to measure the experimental k values

291

between •OH/SO• 4 and CBZ with p–CBA as a reference substance (Ali et al., 2018;

292

Baeza and Knappe, 2011). The reasons for selecting p–CBA as a reference substance

293

have been fully discussed in our previous study (Gao et al., 2019a; Xiao et al., 2017).

294

Briefly, the direct UV photolysis of p–CBA is reckoned to be weak. Benitez et al.

295

reported the quantum yields of p–CBA at the wavelength of 254 nm to be 2.1 × 10−3

296

and 3.0 × 10−3 mol Einstein−1 at pH 2 and 7, respectively (Benitez et al., 2004). Then,

297

the k values for p–CBA reacting with •OH and SO• 4 were on the same order of

298

magnitude as compared to those of CBZ ensuring the accuracy of our measured k

299

values (Kwon et al., 2015). The k values of •OH/SO• 4 oxidizing CBZ are calculated

300

as follows (Haag and Yao, 1992; Packer et al., 2003):

301

k•OH/SO• ,CBZ 4

k•OH/SO• ,p–CBA 4

[CBZ]t [CBZ]t –(ln ) [CBZ]0 [CBZ]0 UV [p–CBA]t [p–CBA]t

ln

= ln

[p–CBA]0

–(ln

)

[p–CBA]0 UV

[CBZ]t –k [CBZ]0 CBZ,UV [p–CBA]t ln –k [p–CBA]0 p–CBA,UV

ln

=

(9)

302

where kCBZ,UV and kp–CBA,UV are the first–order rate constants for the direct UV

303

photolysis of CBZ and p–CBA, respectively. According to eqn. 9 and kinetics data in

304

Figure 3, the k•OH,CBZ and kSO• values were determined to be (4.63 ± 0.01) × 4 ,CBZ 13

305

109 M−1 s−1 and (8.27 ± 0.01)× 108 M−1 s−1, respectively at pH 3.

306 307

3.4 Oxidation pathways of CBZ by •OH and SO• 4

308

We theoretically studied the energetics of the first step of •OH/SO• 4 oxidation of

309

CBZ via radical addition, H–abstraction and SET routes. It should be noted that the

310

abstraction of all possible hydrogen atoms at sites of the benzene rings were not taken

311

into account, as the electrophilic •OH/SO• 4 is prone to react with aromatic ring by

312

radical addition (Buxton et al., 1988; Ding et al., 2019). For example, Ding et al.

313

calculated the gas phase reaction between fluorine, a compound resembling the

314

structure of CBZ, and •OH at the M06‒2X/6-311++G(3df,2p)//M06‒2X/6‒311+G**

315

level. They compared the •OH addition and H‒abstraction pathways on the same

316

reaction site of aromatic ring, and found that the k values for the H‒abstraction on

317

aromatic ring and the •OH addition pathways were reported to be 2.26 × 10−14 and

318

4.25 × 10−11 cm3 molecule−1 s−1, respectively, suggesting that •OH addition pathway

319

can be considered dominant (Ding et al., 2019).The calculated enthalpy change ∆H0R

320

(kcal mol−1), Gibbs free energy change, ∆G0R (kcal mol−1), and the activation free

321

energy ∆≠ G0 (kcal mol−1) are tabulated in Table 1. The Cartesian coordinates of all

322

the structures of the TSs were tabulated in Table S1‒S16, and TS structures were

323

depicted in Figure S1 in the Supplementary data.

324

For •OH, it was found that all reactions were thermodynamically favorable (∆G0R

325

< 0) except for the SET pathway and two addition sites (C4 and C5, see Table 1 and

326

Figure 4). For the radical addition pathway, the ∆H0R of the reaction ranges from –

327

33.8 to –9.42 kcal mol−1, and the ∆G0R ranges from –23.4 to 24.1 kcal mol−1. The

328

∆H0R in the H–abstraction (H30) is –7.74 kcal mol−1, while the ∆G0R is –8.98 kcal

329

mol−1. The reaction energy barrier of H–abstraction was predicted as high as 18.8 kcal 14

330

mol−1, which demonstrated that the oxidation of CBZ on H30 site via •OH should be

331

extremely slow. This observation is supported by the study by Vogna et al. since the

332

loss of the H atom would lead to a rearrangement and loss of isocyanic acid, a

333

compound not observed in their GC-MS analysis (Vogna et al., 2004). Further, the

334

SET pathway of CBZ reacted with •OH is thermodynamically unfavorable with ∆G0R

335

= 9.64 kcal mol−1 and exothermic with ∆H0R =16.4 kcal mol−1. In conclusion,

336

compared with H–abstraction, radical addition on the unsaturated carbon bonds of the

337

heterocyclic and bilateral benzene rings was prone to be dominant pathway due to its

338

low energy barriers.

339

0 For SO• 4 , addition reaction with CBZ are exothermic with ∆HR ranging from –

340

29.0 to –4.19 kcal mol−1. The addition reactions of C3 and C4 sites by SO• 4 are

341

thermodynamically unfavorable with the ∆G0R value of 0.76 and 9.53 kcal mol−1,

342

respectively. The H–abstraction reactions of CBZ by SO• 4 are thermodynamically

343

favorable with negative ∆G0R (–3.05 kcal mol−1). Compared with the positive ∆G0R

344

value for the SET between CBZ and •OH, the ∆G0R value for SET reaction between

345

CBZ and SO• is –14.1 kcal mol−1, demonstrating that this reaction route is 4

346

thermodynamically favorable. The SET pathway of SO• 4 oxidizing CBZ should be

347

dominant due to the very low barrier (0.60 kcal mol−1).

348 349

4. Discussion

350

4.1 Kinetic aspects of CBZ

351

CBZ was highly resistant to direct UV photolysis. There are two factors

352

accounting for the UV resistance. First, CBZ exhibited a poor response under UV

353

irradiation at 254 nm (Figure 1). Second, the highly resistant nature of CBZ to UV

354

irradiation was attributed to the structural rigidity and low reactivity towards 15

355

hydrolysis of amide moiety (RCONH2) (Deng et al., 2013; Ilho and Hiroaki, 2009).

356

Kim and Tanaka compared the degradation ability of 30 kinds of pharmaceutical and

357

personal care products (PPCPs) commonly detected in surface water by UV

358

irradiation at the wavelength of 254 nm. Among them, cyclophosphamide, N,N–

359

diethyl–m–toluamide (DEET), and CBZ were classified as slowly–degrading PPCPs

360

for UV treatment. All these tested PPCPs have RCONR2 group which made them

361

highly resistant to UV treatment. RCONR2 are reported to undergo photodegradation

362

by breaking R–CO or CO–N bonds. However, amides were reported to be stable by

363

the carbonyl couplings due to their resonance structure between the N–C and C=O

364

bonds (Ilho and Hiroaki, 2009). This high resonance is due to the low

365

electronegativity of nitrogen making it a good lone pair donor (Gao et al., 2019b).

366

Compare this to the highly reactive acyl chloride where the chlorine atom is highly

367

electronegative and is not prone to share lone pair electrons, the compound is highly

368

reactive to nucleophilic acyl substitution reactions.

369

The degradation of CBZ was improved considerably in the UV/H2O2 or UV/PS

370

system as compared to the UV system. This is consistent with results performed by

371

Vogna et al (Vogna et al., 2004). They compared direct photolysis of CBZ during

372

UV/H2O2 and UV treatment and found that direct photolysis of CBZ was negligible in

373

the absence of H2O2. Consequently, they suggested that UV/H2O2 treatment could

374

cause efficient reduction of CBZ partly via a series of acridine intermediates (Vogna

375

et al., 2004).

376

The measured k values in this study were in a good agreement with the reported

377

k values (see Table 2). With benzoic acid as a reference compound, Vogna et al.

378

determined the k value for •OH oxidizing CBZ to be (3.07 ± 0.33) × 109 M−1 s−1 at pH

379

3, which is slightly lower than that determined in our study (Vogna et al., 2004). But, 16

380

Huber et al. reported the k value for •OH oxidizing CBZ at a pilot study using p–CBA

381

as a reference to be (8.8 ± 1.2) × 109 M−1 s−1, which is nearly two times higher than

382

our measured value (Huber et al., 2005). The difference could be attributed to higher

383

reaction temperature (T = 25 °C) and pH (pH=7). For the reaction between SO• 4 and

384

CBZ, our measured value is slightly lower than the reported k value of (1.92 ± 0.01) ×

385

109 M−1 s−1 by Matta et al. in a PMS/CoII system using benzene as a reference at pH 3

386

(Matta et al., 2011). The higher k value in their system was due to the existence of

387

CoII which not only activated PMS but also acted as a catalyzer of the intermediate

388

reactions.

389 390

4.2 Mechanistic aspect of CBZ

391

Many studies investigated the degradation of CBZ using AOP techniques and

392

identified degradation products by GC-MS (Huber et al., 2005; Rao et al., 2014;

393

Vogna et al., 2004). Primary degradation products included hydroxylation of the

394

benzene rings and cycloheptene ring double bond. This leads to further degradation

395

including breaking the cycloheptane ring and loss of catechol and benzoic acid

396

derivatives such as 2-hydroxybenzoic acid, 2-aminobenzoic acid. Alternatively, the

397

hydroxylated cycloheptane ring undergoes a reduction in the ring size to create a

398

carbaldehyde product, leading to the formation of acridine or its hydroxylated

399

products 1- or 2-hydroxyacridine. Smaller degradation products result from the

400

cleavage of the ring system and include hydroxyacetic, oxaloacetic, oxalic, malonic,

401

maleic, tartaric acids, succinic, malic, and fumaric acids.

402

For •OH, the reaction pathways including H‒abstraction and radical addition

403

were thermodynamically favorable, but radical addition routes were the kinetically

404

dominant. The result was in good agreement with that reported by Wu et al (Wu et al.,

17

405

2019). Based on their byproduct identification, the reactivity on the unsaturated

406

carbon bonds on CBZ molecule was relatively high, and these bonds were susceptible

407

to be added by •OH. The low energy barrier and negative Gibb’s free energy of the

408

reaction at C7 is also consistent with the studies by Wu et al. and Vogna et al. where

409

both found numerous products from the cleavage of the polycyclic ring system

410

(Vogna et al., 2004; Wu et al., 2019).

411

For SO• 4 , the multiple reaction mechanisms were confirmed with the

412

dominance by SET pathway. As reported by Beitz et al. (Beitz et al., 1998), the

413

oxidation mechanism of binuclear azaarenes reacting with SO• 4 were compared

414

according to differences in the products. It was found that SO• 4 reacted in SET

415

pathway with the heterocycles. Similarly, Luo et al. calculated the energetics of the

416

first step of reaction between SO• 4 and 76 compounds via the SET pathway at

417

SMD/M06–2X/6–311++G** level of theory (Luo et al., 2017). The results showed

418

that SET pathway is thermodynamically favorable to those contained typically

419

electron donating groups, such as –NH2, –NH–, –NH<, –O–, –O–CH3, and –OH. CBZ

420

has electron donating amide group, thus SET is preferred. Note that the average ∆≠ G0

421

values of radical addition pathway (∆≠ G0 = 8.22 kcal mol−1) and SET (∆≠ G0 =0.60 kcal

422

mol−1) are lower than H–abstraction pathways ( ∆≠ G0 = 8.69 kcal mol−1). The

423

calculated result is in good agreement with those reported by Neta et al (Neta et al.,

424

1977b). They reported that SO• 4 did not react by addition to the aromatic ring, but

425

most likely by SET pathway to initially produce the radical cation according to the

426

correlation of the rate constants with the Hammett substituent constant.

427

In order to understand the essential distinction between the mechanisms for •OH

428

and SO• 4 reactions with CBZ, we compared their energy profiles in Figure 4. As it

429

illustrates, the average barrier height for the radical addition reactions for •OH 18

430

−1 oxidation of CBZ is slightly lower than that of SO• 4 by a mean of 1.19 kcal mol .

431

Thus, it can be concluded that the degradation kinetics of CBZ by SO• 4 is slower

432

than that of •OH, which was corroborated by the fact that experimental determined

433

k•OH (4.63 ± 0.01 × 109 M−1 s−1) is larger than SO• (8.27 ± 0.01 × 108 M−1 s−1). As 4

434

discussed in our previous studies, this could be attributed to greater steric hindrance of

435

• SO• 4 than OH (Yang et al., 2017). Although the H–abstraction pathways on amide

436

• 0 group for •OH and SO• 4 are thermodynamically favorable, the ∆GR value of OH

437

−1 was found to be double (18.8 kcal mol−1) than that of SO• 4 (8.69 kcal mol ). It

438

could be attributed to the bond dissociation energy and covalence energy of the

439

• • byproducts (H2O vs. HSO 4 ) (Yang et al., 2017). Interestingly, for OH and SO4 ,

440

both radicals have the lowest ∆≠ G0 values on the radical addition site of C7, the

441

olefinic double bond on the central heterocyclic ring, which exhibited a high degree of

442

similarity in reactivity and reaction pathways.

443 444

5. Conclusion In this study, we experimentally and theoretically investigated the initial step of

445

OH and SO• 4 oxidation of CBZ. First, the experiments for degradation kinetics of

446

•

447

the two radicals oxidizing CBZ have been conducted by the relative rate technique to

448

determine experimental k value. The k values of •OH and SO• 4 were measured to be

449

(4.63 ± 0.01) × 109 M−1 s−1 and (8.27 ± 0.01) × 108 M−1 s−1, respectively at pH 3. We

450

also conducted a DFT study to elucidate the mechanism of the first step of the

451

oxidation reactions between CBZ and •OH/SO• 4 . The calculated results at SMD/M05–

452

2X/6–31+G**//SMD/M05–2X/6–311++G** level of theory were used to verify the

453

mechanisms (i.e., radical addition, H–abstraction, and SET), and were compared with

454

the experimental observations. Specifically, the calculated results showed that the 19

455

∆≠ G0 values for the H–abstraction pathway were higher than those of radical addition.

456

The olefinic double bond on the central heterocyclic ring (C7) was the major additive

457

≠ 0 site of CBZ reaction with both •OH and SO• 4 due to lowest ∆ G value. Further,

458

• SET pathway is thermodynamically feasible for SO• 4 but not for OH. The biggest

459

challenge of AOPs applications during wastewater treatments is the formation of

460

byproducts with unknown (eco)toxicological consequences to aquatic biota and

461

human. This work elucidated mechanistic insight of reaction pathways of •OH and

462

SO• 4 with CBZ, which could quantitatively differentiate various degradation process

463

for target contaminants. Consequently, it could potentially evaluate the treatability and

464

feasibility of AOPs applications.

465 466 467 468

6. Acknowledgments Funding from National Nature Science Foundation of China (No. 21976212 and No. 21507167) is gratefully acknowledged.

20

469 470 471 472

Table 1: Enthalpy change ∆H0R (kcal mol−1), Gibbs free to change, ∆G0R (kcal mol−1), activation energy ∆≠ G0 (kcal mol−1), and imaginary frequency for the reactions of CBZ with •OH and SO• 4 calculated at SMD/M05–2X/6–311++G**//M05–2X/6–31+G** level of theory (the position # see Figure 3).

reaction pathway

radical addition

H–abstraction SET

Position of • OH/SO• 4 C1 C2 C3 C4 C5 C6 C7 H30

•

∆H0R –20.3 –16.6 –15.9 –23.5 –9.42 –21.0 –33.8 –7.74 16.4

OH ∆G0R –10.5 –6.67 –5.83 24.1 1.26 –11.1 –23.4 –8.98 9.64

≠

0

∆ G 8.11 9.34 10.3 10.2 13.0 8.44 6.53 18.8 9.93

∆H0R –16.7 –14.5 –11.6 –4.19 –14.8 –18.6 –29.0 –3.27 7.12

473

21

SO• 4 ∆G0R –4.06 –1.23 0.76 9.53 –0.98 –6.33 –15.0 –3.05 –14.1

≠

0

∆ G 5.70 6.51 9.11 14.3 11.1 6.62 4.22 8.69 0.60

imaginary frequency • OH SO• 4 –512 –605 –530 –600 –553 –623 –410 –647 –555 –563 –526 –626 –338 –431 –1.59 × 103 –2.39 × 103

474

Table 2: The second–order rate constants (k, M−1 s−1) of CBZ with •OH and SO• 4 measured in this study and literature. k (M−1 s−1) measured in this study reported

•

OH SO• 4 (4.63 ± 0.01) × 109 (6.83 ± 0.01) × 108 (8.02 ± 1.90) × 109 (Wols and Hofman-Caris, 2012) (1.92 ± 0.01) × 109 (Matta et al., 2011) (3.07 ± 0.33) × 109 (Vogna et al., 2004) (1.82 ± 0.06) × 109 (Cvetnic et al., 2019) (8.83 ± 0.27) × 109 (Ali et al., 2018)

22

23

3.5x104

experimental calculated

4

ε (M-1 cm-1)

3.0x10

2.5x104 2.0x104

N O

1.5x104

NH2

1.0x104 5.0x103 0.0

200

250

300

350

400

wavelength (nm) Figure 1: Comparison of experimental and calculated UV absorption spectra of CBZ (The measured one was conducted under condition of [CBZ] = 10 µM and pH = 3, while the calculated one was done at the SMD/B3LYP/6‒311++G** level of theory and one explicit water molecule).

24

Figure 2: Highest occupied molecular orbital (HOMO) and lowest unoccupied molecular orbital (LUMO) level frontier orbital contour of CBZ. Oxygen atom, hydrogen atom, carbon atom, and nitrogen atom were denoted by red, white, grey, and blue, respectively.

25

0.0

ln(C/C0)

-0.5 -1.0 -1.5 UV: y= - 0.0031x + 0.0006 UV/H2O2: y= - 0.0733x - 0.0384 UV/PS: y= - 0.0750x - 0.2596

-2.0 -2.5

0

5

10

15

20

25

time(min)

Figure 3: Time–dependent degradation kinetics of CBZ in the UV, UV/H2O2, and UV/PS systems ([CBZ] = 10 µM, [H2O2] = [PS] = 100 µM, pH = 3, and 2.50 × 10–6 Einstein L–1 s–1). The degradation was fitted to a first–order kinetic model (lines).

26

C1add C4add C7add

30

C2add C5add H30abs

C3add C6add SET

∆GοR (kcal mol-1)

20 10 0 -10 -20

SO•− 4

•

OH

-30

products

TSs

reactant

TSs

products

Figure 4: Comparison of profiles of the potential energy surface of •OH and SO• 4 reactions with CBZ at SMD/M05–2X/6–311++G**//SMD/M05–2X/6–31+G** level of theory.

27

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9 8 −1 −1 The k values of •OH and SO•ି 4 with CBZ were 4.63×10 and 8.27×10 M s .




• DFT result showed energy barriers of SO•ି 4 reacting with CBZ was higher than OH.




Radical addition was the dominant pathway for both •OH and SO•ି 4 reacting with CBZ.




Remarkable advantages of AOPs for removal of organic contaminants were exhibited.

Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: