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Radical Chemistry and Structural Relationships of PPCP Degradation by UV/Chlorine Treatment in Simulated Drinking Water Kaiheng Guo, Zihao Wu, Chii Shang, Bo Yao, Shaodong Hou, Xin Yang, Weihua Song, and Jingyun Fang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b02059 • Publication Date (Web): 15 Aug 2017 Downloaded from http://pubs.acs.org on August 16, 2017

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Environmental Science & Technology

Radical Chemistry and Structural Relationships of PPCP Degradation by

1 2

UV/Chlorine Treatment in Simulated Drinking Water

3

Kaiheng Guo†, Zihao Wu†, Chii Shang‡, Bo Yaoǁ, Shaodong Hou†, Xin Yang†, Weihua Song*,ǁ,

4

Jingyun Fang*,†

5 6

†

7

Technology, School of Environmental Science and Engineering, Sun Yat-Sen University, Guangzhou

8

510275, P. R. China

9

‡

Guangdong Provincial Key Laboratory of Environmental Pollution Control and Remediation

Department of Civil and Environmental Engineering, the Hong Kong University of Science and

10

Technology, Clear Water Bay, Kowloon, Hong Kong

11

ǁ

12

China

Department of Environmental Science & Engineering, Fudan University, Shanghai 200433, P. R.

1

ACS Paragon Plus Environment


13

TOC Art

14 15

2


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16


ABSTRACT

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The UV/chlorine process is an emerging advanced oxidation process (AOP) used for the

18

degradation of micropollutants. However, the radical chemistry of this AOP is largely unknown for

19

the degradation of numerous structurally diverse micropollutants in water matrices of varying quality.

20

These issues were addressed by grouping 34 pharmaceuticals and personal care products (PPCPs)

21

according to the radical chemistry of their degradation in the UV/chlorine process at practical PPCP

22

concentrations (1 µg L-1) and in different water matrices. The contributions of HO• and reactive

23

chlorine species (RCS), including Cl•, Cl2•– and ClO•, to the degradation of different PPCPs were

24

compound specific. RCS showed considerable reactivity with olefins and benzene derivatives, such

25

as phenols, anilines and alkyl-/alkoxybenzenes. A good linear relationship was found between the

26

RCS reactivity and negative values of the Hammett ∑σp+ constant for aromatic PPCPs, indicating

27

that electron-donating groups promote the attack of benzene derivatives by RCS.The contribution of

28

HO•, but not necessarily RCS, to PPCP removal decreased with increasing pH. ClO• showed high

29

reactivity with some PPCPs, such as carbamazepine, caffeine, and gemfibrozil, with second-oreder

30

rate constants of 9.2 × 107, 1.03 × 108 and 4.16 × 108 M-1 s-1, respectively, which contributed to their

31

degradation. Natural organic matter (NOM) induced significant scavenging of ClO• and greatly

32

decreased the degradation of PPCPs that was attributable to ClO•, with a second-order rate constant

33

of 4.5 × 104 (mg L-1)-1 s-1. Alkalinity inhibited the degradation of PPCPs that was primarily attacked

34

by HO• and Cl• but had negligible effects on the degradation of PPCPs by ClO•. This is the first study

35

on the reactivity of RCS, particularly ClO•, with structurally diverse PPCPs under simulated drinking

36

water condition.

3



37

INTRODUCTION

38

Pharmaceuticals and personal care products (PPCPs) are a group of emerging contaminants that

39

have gained much attention due to the potential risk of inducing adverse ecological and health effects

40

in wildlife and humans.1 PPCPs have been identified worldwide in varied aquatic environments at

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ng-µg L-1 concentrations.1, 2 The most frequently detected pharmaceuticals are anti-inflammatory

42

drugs, anticonvulsants, antibiotics, β-blockers, and lipid regulators.3 Recent studies have shown that

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some PPCPs resist biological wastewater treatment and conventional drinking water treatment,

44

including coagulation, sedimentation and filtration.2, 4 Some of these compounds also resist UV

45

photolysis and oxidation by chlorine or ozone, while others are resistant to adsorption by activated

46

carbon.5-7 Alternative treatment technologies thus need to be sought.

47

Advanced oxidation processes (AOPs) are effective technologies for the removal of refractory

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micropollutants via the generation of reactive radicals, such as hydroxyl radical (HO•) (E0 = 2.8 V).8

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The UV/H2O2 process has become a common AOP in water treatment, which generates HO• through

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photolysis of H2O2. There are a series of studies focused on the UV/H2O2 degradation of PPCPs,

51

such as various antibiotics (tetracycline, nalidixic acid, macrolides, chloramphenicol, sulfonamides),

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naproxen, carbamazepine, gemfibrozil, ibuprofen, caffeine, metoprolol, propranolol, clenbuterol,

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primidone, bisphenol A, and so on.9-14 However, the efficiency of PPCP degradation is highly

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dependent on the water matrix, as HO• is a nonselective oxidant and components such as dissolved

55

organic matter (DOM) and alkalinity in the matrix can scavenge HO•.15 Additionally, this treatment

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requires high dosages of H2O2.16 The chemical cost of quenching the residual H2O2 further increases

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the overall cost of the process.17, 18

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The UV/chlorine process is an emerging AOP and an alternative to the UV/H2O2 process. The

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UV/chlorine process has been reported to be effective at degrading a variety of micropollutants

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including trichloroethylene,19 benzoic acid,20 atrazine,21 taste and odor compounds (e.g.,

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2-methylisoborneol and geosmin),22 certain PPCPs (e.g., ibuprofen, trimethoprim, sulfamethoxazole,

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carbamazepine, diclofenac, nitroimidazoles, and metoprolol), and some hormones (e.g.,

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17a-ethinylestradiol).23-28 Compared to the UV/H2O2 process, the UV/chlorine process more

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efficiently degrades some PPCPs such as carbamazepine, sulfamethoxazole, and trimethoprim.24, 28, 29

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Overall, the UV/chlorine process has been suggested to be a possible alternative to the UV/H2O2

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process in real water treatment because of the more efficient radical production and the use of the

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residual chlorine for disinfection in the former process.22, 30, 31

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UV photolysis of chlorine produces primary radicals of both HO• and Cl• (eq 1).32-34 The secondary radical species, such as Cl2•- and ClO•, are formed as shown in eqs 2–6.20, 24, 35

70

HOCl/OCl- HO•/O•- + Cl•

71

Cl• + Cl-

72

HO• + HOCl → ClO• + H2O

k = 2.0 × 109 M−1 s−1

(3)

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HO• + OCl- → ClO• + OH-

k = 8.8 × 109 M−1 s−1

(4)

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Cl• + HOCl → ClO• + H+ + Cl-

k = 3.0 × 109 M−1 s−1

(5)

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Cl• + OCl- → ClO• + Cl-

→( ) ←( )

Cl2•-

(1) k1 = 6.5 × 109 M−1 s−1, k2 = 1.1 × 105 s−1

k = 8.2 × 109 M−1 s−1

(2)

(6)

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HO• is a broad-spectrum strong oxidant, while reactive chlorine species (RCS) such as Cl•, Cl2•- and

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ClO• are selective.20, 36, 37 Cl•, Cl2•- and ClO• are oxidants with redox potentials of 2.4 V, 2.0 V and

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1.5-1.8 V, respectively.20, 38 Cl• is selective and has higher reactivity than HO• toward certain organics,

79

such as benzoic acid, chlorobenzene and phenol.36, 39 Cl2•- is selective for olefinic compounds and 5



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aromatics when the ring is substituted with hydroxy, methoxy and amino groups.40 ClO• reacts

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rapidly with aromatics possessing methoxy groups.38 As the above structural features can be found in

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many PPCPs and considering the different reactivities of the diverse radicals of HO• and RCS in the

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UV/chlorine process, this process may be a complementary solution for the degradation of many

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structurally diverse PPCPs. For example, degradation of trimethoprim and caffeine is primarily

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achieved with RCS but not HO•,24, 35 carbamazepine can react significantly with both HO• and Cl•,25

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and ibuprofen elimination is primarily attributed to HO• and not RCS.23 However, the radical

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mechanisms of the degradation of only a few PPCPs have been studied at much higher

88

concentrations (µM level) than those found in real water samples. Therefore, the contributions of

89

HO• and RCS to the degradation of structurally diverse PPCPs in the UV/chlorine process should be

90

comprehensively studied under practically relevant conditions. With understanding the roles of HO•

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and RCS, it would be helpful for predicting the transformation of PPCPs in the UV/chlorine process,

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especially for potential drinking water treatment.

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Quantitative structure-activity relationships (QSARs) are often determined between the k-values

94

for an oxidation reaction of closely related compounds and substituent descriptor variables such as

95

Hammett or Taft sigma constants. Many QSARs have been developed for the reactions of chlorine,

96

chlorine dioxide, ferrate, ozone, HO• and sulfate radicals with micropollutants.41, 42 The development

97

of QSARs for the reaction between RCS and structurally diverse PPCPs will help explore the

98

reactivity property of RCS and thus deserves critical insights.

99

The main objectives of this study were 1) to investigate the degradation kinetics of a number of

100

structurally diverse PPCPs at practically relevant concentrations in the UV/chlorine process, 2) to

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elucidate the radical chemistry for the degradation of different PPCPs in different water matrices, and

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3) to explore the relationships of RCS reactivity with PPCP structures. Simulated water samples were

103

used as the real water matrices. Thirty-four PPCPs, including β-adrenergic agonists, β-blockers,

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nitroimidazoles, psychiatric drugs, sulfonamides, lipid regulators, macrolides, cephalosporins and

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anti-inflammatory agents, were selected in this study because of their frequent occurrence in aquatic

106

environments.

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MATERIALS AND METHODS

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Materials. All PPCPs and isotopic internal standards were purchased from Sigma-Aldrich or

109

Toronto Research Chemicals (TRC) unless otherwise noted. Individual stock solutions of the PPCPs

110

were prepared in methanol at 1000 mg L-1. Then a series of mixed PPCP solutions were prepared in

111

methanol with concentrations ranging from 1.0 to 2000 ng L-1 for calibration. The stock working

112

solution for treatment was prepared in ultrapure water (18.2 MΩ cm) with each compound at 1.0 mg

113

L-1. Individual stock solutions of the isotopic internal standards were prepared in methanol at 100 mg

114

L-1. A mixture stock solution containing all the isotopic internal standards was prepared in methanol,

115

with each compound at 0.55 mg L-1, including atenolol-D7, carbamazepine-D10, dimetridazole-D3,

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flumequine-C3,

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venlafaxine-D6 and ibuprofen-D3.

metronidazole-D4,

primidone-D5,

roxithromycin-D7,

trimethoprim-C3,

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Sodium hypochlorite (NaOCl) solution (available chlorine, 4.00-4.99%), hydrogen peroxide

119

(H2O2) and nitrobenzene (NB) (99%) were obtained from Sigma-Aldrich. High performance liquid

120

chromatography (HPLC)-grade methanol, acetonitrile, hexane, o-phosphoric acid and t-butanol were

121

purchased from Fisher Scientific, and NaOH and NaHCO3 were purchased from Sinopharm

7



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Chemical Reagent Co., Ltd. (Shanghai, China). Suwannee River natural organic matter (NOM)

123

(Cat.# 2R101N) obtained from the International Humic Substances Society was dissolved in pure

124

water and filtered through a 0.45-µm membrane to prepare a stock NOM solution with the total

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organic carbon (TOC) of 123.2 mg L-1.

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Experimental procedure. The photochemical experiments were conducted in a 750 mL,

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magnetically stirred, cylindrical, borosilicate glass reactor with a quartz tube in the center, in which a

128

low-pressure mercury lamp (Heraeus GPH 212T5L/4, 10 W) was placed. The temperature was

129

maintained at 25 ± 0.2 °C. The UV photon flux (I0) entering the solution was determined to be 0.477

130

µE L-1 s-1 using iodide/iodate chemical actinometry.43 The effective path length (L) was determined

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to be 2.43 cm by measuring the photolysis kinetics of dilute H2O2 (Figure S1).44 The corresponding

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average UV fluence rate (Ep0) was estimated to be 0.78 mW cm-2.

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A 750 mL test solution containing 1 µg L-1 of each PPCP and 2 mM phosphate buffer was dosed

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with the NaOCl stock solution and simultaneously exposed to UV irradiation. The chlorine dosages

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were 10 µM and 20 µM. Samples were collected at different time intervals and quenched with

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ascorbic acid at a molar ratio of ascorbic acid to chlorine of 1.5:1. Control tests of PPCP degradation

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by direct UV photolysis and dark chlorination were carried out in a similar manner but in the absence

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of chlorine and UV light, respectively. All tests were conducted at least twice. All data plots represent

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the average of the experimental data of the duplicated test results.

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Analytical methods. The PPCPs were quantified on an Agilent 1290 ultra high performance

141

liquid chromatography (UHPLC) online solid phase extraction (SPE) system coupled with an Agilent

142

6430 triple quadrupole mass system with an electrospray ionization source. An Agilent ZORBAX

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SB-C18 HD column (2.1 mm × 50 mm × 1.8 µm) thermostatted at 30 °C was used to separate the

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PPCPs. The mobile phase consisted of 0.1% acetic acid water solution and 0.1% acetic acid

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acetonitrile (ACN) solution at a flow rate of 0.25 mL/min. The gradient elution of ACN/water (v/v)

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kept at 5/95 for 2.6 min, then increased to 95/5 from 2.6 to 6.6 min and held for 3 min. The mass

147

conditions can be found in Supporting Information (SI) Tables S1.

148

An Agilent 1290 infinity flexible cube was employed for online SPE. First, a 900 µL sample

149

was withdrawn and delivered to one of the online SPE columns with 3.0 mL of H2O by the flush

150

pump. Then, the valve switched to the position that coupled the SPE column with the analysis

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column, and the analytes were eluted from the SPE column in back-flush mode. Simultaneously,

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another SPE column began the regeneration process with 5.0 mL of acetonitrile, followed by 10.0

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mL of H2O. The flow rate was set at 1.5 mL min-1.

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NB was determined by GC-MS (Shimadzu, QP2010 Ultra) with an SH-Rxi-5Sil MS column

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(30 m × 0.25 mm × 0.25 µm) (Details in Text S1). The concentration of free chlorine was determined

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using DPD/FAS titration.45 The pH was measured with a pH meter (Thermo Scientific, Orion Star

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A111).

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Determination of the relative contributions of HO• and RCS in the UV/chlorine process.

159

To quantitatively describe the roles of HO• and RCS, NB was used as a probe compound that

160

selectively reacts with HO• (k = 3.9×109 M-1 s-1) and resists RCS.20 The steady-state concentration

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of HO• ([HO•]ss) can be calculated using eq 7. The pseudo first-order rate constants (k′) of the

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reactions of a specific PPCP with HO• (kHO•-PPCP′) and RCS (kRCS-PPCP′) can be calculated using eqs 8

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and 9, respectively.

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kNB′ = kHO•-NB[HO•]ss

(7)

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kHO•-PPCP′ = kHO•-PPCP[HO•]ss

(8)

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kRCS-PPCP′ = kPPCP′ - (kHO•-PPCP′ + kUV-PPCP′ + kchlorine-PPCP′)

(9)

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where kPPCP′ and kNB′ are defined as the k′ of PPCP and NB, respectively. kNB′ is depicted in Figure

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S2 of the SI. kHO•-PPCP and kHO•-NB represent the second-order rate constants of HO• reacting with

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PPCP and NB, respectively. The kHO•-PPCP was obtained in this study (Details in Text S2 and Figure

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S3). kRCS-PPCP′, kUV-PPCP′ and kchlorine-PPCP′ represent the k′ of PPCP attenuation by RCS oxidation, UV

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photolysis and chlorination, respectively. Note that the residual chlorine after 10 min UV/chlorine

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treatment was similar to that after 10 min chlorination under the test conditions (difference less than

173

10%, data not shown), which demonstrates that the chlorine exposure was similar in both processes.

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Determination of the second-order rate constants for ClO•. The second-order rate constants

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for the reaction of ClO• with selected PPCPs were determined from competition kinetics using 1,4-

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dimethoxybenzene (DMOB) or 2,5-dimethoxybenzoate ion (DNBA) as reference compounds (R),

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which react with ClO• at second-order rate constants of 2.1 × 109 M-1 s-1 and 7.0 × 108 M-1 s-1,

178

respectively.38 To create effective conditions for ClO•, the UV/chlorine process was carried out at a

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chlorine dosage of 50 µM and pH 8.4, and the solution was spiked with 100 mM sodium bicarbonate

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to scavenge the HO•, Cl• and Cl2•- in the system. A reference compound (R) and selected PPCPs were

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simultaneously spiked into the system at concentrations of 5 µM. Note that CO3•- was also presented

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in the created ClO• system. However, the reactivity of CO3•- with selected PPCPs and reference

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compounds could be ignored, due to their negligible degradation in a CO3•- system (shown in Text S3

184

and Figure S4). The second-order rate constants of the PPCPs with ClO• were calculated from eq 10.

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ln

[] []

=

•, •,

× ln

[] []

10


(10)

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where kClO•,PPCP and kClO•,R represent the second-order rate constants of ClO• reacting with the

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selected PPCP and the reference compound, respectively (Figure S5).

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The second-order rate constant for the reaction between ClO• and NOM was determined using

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DMOB as a reference compound with varying NOM concentrations. The k′ of DMOB varied in the

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presence or absence of NOM, as shown in eq 11.

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′ ′ !

=1+

•,$% •,

×

[&'(] []

(11)

192

where [NOM]0 and [R]0 represent the initial concentrations of NOM and the reference compound

193

respectively. kR-0′ and kR-i′ represent the k′ of the reference compound in the absence and presence of

194

NOM, respectively. kClO•,NOM and kClO•,R represent the second-order rate constants of NOM and the

195

reference compound, respectively (Figure S6).

196

QSAR analysis. To explore the relationship between the structures of aromatic PPCPs and their

197

reactivity toward different radicals, QSAR was employed. Hammett σ+ constants were used in the

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QSAR analysis because they are the most common substituent descriptors in physical organic

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chemistry and are relatively easily accessible and applicable.41,

200

quantitatively express the electron-donating (large negative value) or electron-withdrawing (large

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positive value) properties of substituents on aromatic compounds, including the inductive and

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resonance effects of substituents. To account for the effects of multiple substituents on the aromatic

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ring, Σσ+ was used.41 In the case of mono- or di-substituted benzenes, multiple sites for radical attack

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are possible. Each scenario of radical attack results in a different value of Σσ+. To solve this problem,

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radical attack was assumed to occur at the site where the Σσ+ value was the lowest (i.e., σp+ was used

206

instead of σm+ and σo+).41 Also, to simplify the complex structures of PPCPs, a structural

11


46

Hammett σ+ constants


207

approximation was used, in which only one or two neighboring atoms from an aromatic ring

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determined the inductive/resonance effects of whole substituents.41

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Kinetic model simulation. The model contained 100 reactions and was implemented with

210

Kintecus v4.55 47 using rate constants obtained from the literature when available or estimated based

211

on analogy to similar reactions (Table S2). The modeling considered only the reactions involving

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inorganic compounds and components of the water matrix to evaluate the concentrations of HO• and

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RCS because the rate constants of RCS with many of the target PPCPs were not available. Although

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the PPCPs and their products influenced the absolute radical concentrations via scavenging, the

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kinetic model served to clarify the effect of pH and components of the water matrix on the radical

216

concentrations.

217

RESULTS AND DISCUSSION

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Degradation kinetics of PPCPs in UV/chlorine treatment in pure water. Figure 1 compares

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the observed first-order rate constants (kobs′) of the degradation of PPCPs by UV, chlorination, and

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UV/chlorine processes at the same molar oxidant dosage and pH 7 in pure water. Direct photolysis at

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254 nm was not effective for most PPCPs at a UV fluence of 468 mJ cm-2, except for

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sulfamethoxazole (kobs′ = 0.072 min-1) and diclofenac (kobs′ = 0.159 min-1), which is consistent with

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the literature.5 The degradation of PPCPs containing amine groups by chlorination was too rapid to

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obtain kobs′, including cephalexin, ciprofloxacin, tetracycline, terbutaline, famotidine and cimetidine,

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which were removed within 0.5 min (shown in Figure S7). PPCPs containing phenol and aniline

226

moieties, such as salbutamol, ractopamine, bisphenol A and sulfamethoxazole, were also removed

227

efficiently by chlorination, with kobs′ > 0.1 min-1. Degradation by the UV/chlorine process was much

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more efficient for all the PPCPs, with kobs′ broadly ranging from 0.022 to 4.135 min-1, indicating the

229

dominant roles of radicals of HO• and RCS.

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The specific pseudo first-order rate constants (k′) of HO• (kHO•′) and RCS (kRCS′) were

231

significantly different for reactions with the different PPCPs (Figure 1). kRCS′ clearly covered a much

232

broader range than kHO•′ (as shown in the boxplot in Figure 1 inset), indicating that PPCPs react more

233

selectively with RCS than with HO•. The PPCPs were grouped according to their reactivity with HO•,

234

RCS and chlorine, as shown in Table S3. (1) Compounds that were primarily degraded by HO• were

235

defined as group I (kHO•′/ kRCS′ > 1). These PPCPs included dimetridazole, tinidazole, ornidazole,

236

metronidazole, ibuprofen, primidone, venlafaxine, nalidixic acid and flumequine. (2) Compounds

237

that were degraded by both HO• and RCS were designated group II (0.2 < kHO•′/ kRCS′ < 1). Group II

238

contained atenolol, metoprolol, and macrolides such as erythromycin, azithromycin and

239

roxithromycin. (3) Compounds that were primarily degraded by RCS were designated group III

240

(kHO•′/ kRCS′ < 0.05). Group III contained carbamazepine, caffeine, gemfibrozil, naproxen,

241

propranolol, clenbuterol and diclofenac. (4) Compounds degraded by RCS and HOCl/OCl- were in

242

group IV (kHO•′/ kRCS′ < 0.05, kchlorination′ > 0.02 min-1), which included trimethoprim, salbutamol,

243

ractopamine, bisphenol A and sulfamethoxazole. The UV/chlorine process had much higher kobs′ for

244

groups III and IV. (5) Compounds that were degraded extremely rapidly by chlorination, were

245

defined as group V. Group V contained cephalexin, ciprofloxacin, tetracycline, terbutaline,

246

famotidine and cimetidine, for which the kobs′ by chlorination and UV/chlorine was too high to be

247

determined and thus not included in the comparison. In addition, chloramphenicol was primarily

248

degraded by direct UV photolysis, and it was not included in the comparison.

13



249

Obviously, kRCS′ of the PPCPs containing benzene ring attached with electron-rich functional

250

groups in groups III and IV was extremely high. To better explain the reactivity of HO• and RCS

251

toward aromatic PPCPs, a primary QSAR was established, based on the assumption that the benzene

252

ring was the main attacking site. Several PPCPs in each group were selected in the QSAR: primidone

253

and ibuprofen in group I, atenolol in group II, gemfibrozil in group III, and trimethoprim, salbutamol,

254

ractopamine & bisphenol A in group IV. These PPCPs were selected due to their available Hammett

255

constants. As shown in Figure 2, the reactivity of HO• kept stable as the Hammett constants

256

increased, while that of RCS increased significantly, further proving that RCS is more selective to

257

electron-rich compounds than HO•. The reactivity of RCS with aromatic PPCPs relies on the

258

electron-donating properties of substituents on aromatic ring.

259

Effects of pH on the radical chemistry of PPCP degradation in pure water. pH has been

260

reported to have a profound effect on the degradation efficiency of the UV/chlorine process. The

261

degradation of micropollutants such as benzoic acid, ibuprofen, para-chlorobenzoic acid, NB,

262

ronidazole and trichloroethylene was reported to be greatly reduced by increasing pH.19, 20, 27, 34

263

Interestingly, pH-dependent trends in kobs′ for different PPCP groups were not always consistent with

264

the literatures (Figure 3). kobs′ decreased significantly with increasing pH from 6 to 8 for PPCPs in

265

group I (expect for primidone and venlafaxine), and atenolol & metoprolol in group II, but showed

266

different pH-dependent trends for the PPCPs in other groups. The kobs′ of macrolides (i.e.,

267

erythromycin, azithromycin, and roxithromycin) in group II, bisphenol A and ractopamine in group

268

IV increased significantly with increasing pH, while the kobs′ for gemfibrozil, naproxen, propranolol,

269

diclofenac, trimethoprim and sulfamethoxazole in groups III and IV reached maximum at pH 7.

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Figure 3 also shows the specific k′ of UV photolysis, chlorination, HO• and RCS in PPCP

271

degradation via UV/chlorine treatment. The contributions of HO• and RCS varied with pH, and the

272

trend was different for different groups of PPCPs. The kHO•′, but not necessarily the kRCS′, for all the

273

PPCPs decreased significantly as the pH increased from 6 to 8. The degradation of PPCPs in group I

274

was attributed to HO•, and thus, kHO•′ decreased significantly with increasing pH. For group II, both

275

HO• and RCS were important oxidants. The contribution of HO• was almost equal to or even higher

276

than that of RCS in the degradation of PPCPs at pH 6. As the pH increased, kHO•′ decreased

277

significantly, while kRCS′ decreased for atenolol and metoprolol slightly. For groups III and IV, kRCS′

278

dominated the PPCP degradation (ratio of kRCS′ to kobs′ > 50%), particularly at alkaline pH.

279

Meanwhile, UV photolysis and chlorination played minor roles in the degradation of PPCPs in

280

groups I–IV. The k′ for UV photolysis (kUV′) remained constant with increasing pH. Chlorination was

281

effective at degrading the PPCPs in group IV; the kchlorine′ of sulfamethoxazole (pKa1 = 1.83; pKa2 =

282

5.57) decreased with increasing pH, while those of bisphenol A, ractopamine and salbutamol

283

increased.

284

The steady-state radical concentrations at different pH values depend on their formation and

285

scavenging rates. Increasing pH affects the abundance of HOCl and OCl- (pKa = 7.5), which are also

286

radical scavengers in addition to being precursors.20 The radical scavenging effects of OCl- on HO•

287

and Cl• are higher than those of HOCl.20 Parallelly, HO• and Cl• are consumed as ClO• is produced

288

(eqs 3-6). The steady-state concentrations of different radicals were calculated using the kinetic

289

model with the reactions shown in Table S4. The concentration of HO• decreased significantly with

290

increasing pH while Cl• decreased gradually and ClO• kept constant at different pHs. Additionally,

15



291

the ClO• concentration was 3-4 orders of magnitude higher than those of HO•, Cl• and Cl2•- (Table 1).

292

As ClO• reacts with methoxybenzene and phenolic compounds,38 it may contribute to the

293

degradation of PPCPs in groups III and IV, such as gemfibrozil, naproxen, trimethoprim, and

294

bisphenol A. The second-order rate constants of ClO• with 3 PPCPs in group III, i.e., carbamazepine,

295

caffeine and gemfibrozil, were determined to be 9.2 × 107, 1.03 × 108 and 4.16 × 108 M-1s-1,

296

respectively (Figure S5), demonstrating that ClO• played a significant role in the degradation of the

297

PPCPs in group III.

298

pH can also affect the proton transfer reactions of PPCPs with acid/base characteristics. Some

299

PPCPs can undergo proton transfer reactions to yield their protonated or deprotonated forms at

300

varying pH, which may also affect their reactivity in the UV/chlorine process. Increasing pH from 6

301

to 8 promotes the transformation of PPCPs from neutral to anionic forms (i.e., phenolic compounds)

302

or from cationic to neutral forms (i.e., amines). The latter form in each pair is richer in electrons. The

303

degradation of salbutamol, ractopamine and bisphenol A was promoted when these compounds were

304

transformed to the corresponding phenoxides with increasing pH, which contributed to the increase

305

in their kRCS′ and kchlorine′. The transformation of PPCPs with increasing pH may also account for the

306

increasing kRCS′ of the three macrolides (group II) containing tertiary amino groups (pKa = 8.5–8.9).

307

Radical chemistry of PPCP degradation in simulated NOM-containing water. The presence

308

of 1 mg L-1 NOM exerted different effects on the degradation of PPCPs in the UV/chlorine process

309

(Figure 4). The presence of NOM significantly reduced the kobs′ of PPCPs in groups III and IV.

310

However, the kobs′ of the PPCPs in group I as well as of atenolol and metoprolol in group II were less

311

affected by NOM. kHO•′ changed slightly by ~10% in the presence of 1 mg L-1 NOM, while kRCS′

16


Page 16 of 34

Page 17 of 34


312

decreased by 19.3~>90% (Figure 4). Thus, RCS is more susceptible to NOM than HO•. The rate

313

constants of the reaction of NOM with HO• and Cl• were reported to be 2.5 × 104 (mg L-1)−1 s−1 and

314

1.3 × 104 (mg L-1)−1 s−1, respectively.20 As the rate constant of the reaction of NOM with Cl• is lower

315

than that with HO•, ClO• might be more susceptible to NOM. The second-order rate constant

316

between ClO• and NOM was determined to be 4.5 × 104 (mg L-1)-1 s-1 (Figure S6). Thus, NOM

317

scavenges ClO• significantly faster than either HO• or Cl•, resulting in the significant reduction of

318

kRCS′, particularly for the PPCPs in groups III and IV. This result suggests that the presence of NOM

319

at realistic concentrations (1 mg L-1) can significantly inhibit the degradation of PPCPs and that this

320

degradation is dominated by ClO•. The modeling result further demonstrates the effect of NOM on

321

radicals (Table 1). The steady-state concentration of ClO• decreased by > 90% in the presence of

322

NOM, while that of HO•, Cl• and Cl2•- decreased by < 20%. Additionally, NOM can filter UV light

323

and consume HOCl/OCl- (k = 0.7 to 5 M−1 s−1).20, 48 However, these two effects were relatively not

324

significant.

325

Some PPCPs in group I, such as ornidazole, metronidazole, ibuprofen and venlafaxine, had kobs′

326

values that were even higher in the presence of NOM, which may have been due to their sorption to

327

NOM. Nitroimidazoles has been reported to be absorbed onto effluent organic matter (EfOM)

328

through dispersive interactions between the π electrons in the aromatic rings of the nitroimidazoles

329

and EfOM.49

330

Radical chemistry of PPCP degradation in water with simulated alkalinity. The proportions

331

of H2CO3*, HCO3- and CO32- were 17.96%, 82% and 0.04%, respectively, at pH 7.21 HCO3-

332

scavenges HO•, Cl• and Cl2•- with rate constants of 8.5 × 106, 2.2 × 108 and 8.0 × 107 M-1 s-1,

17



Page 18 of 34

333

respectively (eqs 12-14).20 These scavenging reactions concurrently form CO3•-. ClO• is not

334

scavenged by bicarbonate (k < 600 M-1 s-1).36 Note that the reactions of H2CO3* with radicals were

335

negligible.20

336

HO• + HCO3- → H2O + CO3•-

k = 8.5 × 106 M−1 s−1

(12)

337

Cl• + HCO3- → H2O + CO3•-

k = 2.2 × 108 M−1 s−1

(13)

338

Cl2•- + HCO3- → 2Cl- + H+ + CO3•-

k = 8.0 × 107 M−1 s−1

(14)

339

The kobs′ values of most PPCPs were decreased by ~20-52% in the presence of 1 mM HCO3-,

340

whereas the kobs′ of some PPCPs slightly changed or even increased. kHO•′ decreased by 15% in the

341

presence of 1 mM HCO3-, while kRCS+CO3•-′ decreased by ≤ 72% or increased by ≤ 36% (Figure 5).

342

The decrease in the kobs′ of some PPCPs was primarily due to the scavenging of HO• or Cl• by

343

bicarbonate. Meanwhile, the kobs′ of certain PPCPs slightly changed or even increased, which could

344

be ascribed to the slight change in ClO• and the increased concentration of CO3•-. ClO• likely

345

accounted for the slightly changed kobs′ values of some PPCPs, including the macrolides in group II

346

as well as trimethoprim, salbutamol, clenbuterol and ractopamine in groups III and IV. CO3•- can

347

contribute to the degradation of PPCPs such as erythromycin, roxithromycin, clenbuterol, salbutamol,

348

ractopamine and trimethoprim, whose second-order rate constants with CO3•- range from 107 to 108

349

M-1 s-1, especially that of clenbuterol (kCO3•- = 5.2×108 M-1 s-1) (Table S4).9 The modeling result

350

further demonstrates the effect of alkalinity on radicals (Table 1), showing that the steady-state

351

concentrations of HO•, Cl• and Cl2•- decreased in the presence of HCO3-, while the concentration of

352

ClO• slightly changed. The concentration of CO3•- was approximately 10-9 M (Table 1), 3–6 orders of

353

magnitude higher than those of HO• and Cl•.

18


Page 19 of 34


354

Engineering Implications. The radical chemistry and structural relationships of the

355

degradation of 34 structurally diverse PPCPs by the UV/chlorine treatment were demonstrated under

356

environmentally relevant conditions. The degradation efficiency depends on the PPCP structure and

357

water matrix. The functional groups in PPCPs affect their reactivity toward HO• and RCS. The

358

degradation of PPCPs in group I possessing electron-withdrawing groups, such as –NO2, –Cl or –F,

359

largely involves HO•, while the degradation of PPCPs in groups III and IV containing

360

electron-donating groups, such as phenol, aniline, alkyl-/alkoxy aromatics or olefin, is largely

361

attributable to RCS such as Cl•, Cl2•- and ClO•. ClO• was found to significantly oxidize PPCPs such

362

as carbamazepine, caffeine, gemfibrozil and naproxen with the rate constants of 108 to 109 M-1 s-1.

363

Thus, the roles of RCS, particularly ClO•, in PPCP degradation should be considered in the

364

UV/chlorine process.

365

Water matrix characteristics such as pH, NOM and alkalinity can greatly affect the radical

366

chemistry of the UV/chlorine process. pH can interfere with the chlorine (photo)chemistry, radical

367

transformation reactions and acid-base properties of micropollutants. The contribution of HO•

368

decreased with increasing pH, but the same trend did not occur for all RCS. For the degradation of

369

micropollutants primarily by HO• (Group I), the efficiency decreased significantly with increasing

370

pH in the UV/chlorine process. However, for the degradation primarily by RCS, the efficiency

371

depended on the speciation of RCS. The efficiency decreased if Cl• and Cl2•- were the major RCS,

372

but little or no change was observed if ClO• was the major RCS. Additionally, the transformation of

373

micropollutants from their neutral/cationic form to anionic/neutral form with increasing pH enhanced

374

the electrophilic reaction with RCS.

19



375

The concentrations of NOM and alkalinity (mg L-1) are several orders of magnitude higher than

376

that of micropollutants (ng L-1),50 which can significantly affect the radical chemistry. NOM can

377

scavenge HO• and Cl•,20 and the scavenging effect on ClO• is even much greater. Thus, the

378

degradation of micropollutants that are readily degradable by ClO• can be greatly inhibited by the

379

presence of NOM. Alkalinity components, including bicarbonate and carbonate ions, can scavenge

380

HO• and Cl•, with rate constants of 8.5 × 106 and 2.2 × 108 M-1 s-1, respectively, but not ClO• (< 600

381

M-1 s-1).36 Thus, the degradation of micropollutants primarily by ClO• is slightly affected by the

382

presence of alkalinity, while that by HO• and Cl• is inhibited by alkalinity.

383

The PCPP degradation mechanisms and the factors affecting it can also be extrapolated to other

384

micropollutants sharing similar structures. In addition, the formation of halogenated disinfection by

385

products (DBPs) through the interactions between RCS and DBP precursors is a general concern for

386

the UV/chlorine process, careful evaluation is required prior to large scale application.

387 388

ASSOCIATED CONTENT

389

Supporting Information.

390 391

Texts S1-2, Figures S1-11 and Tables S1-5. This material is available free of charge via the Internet at http://pubs.acs.org.

392 393

AUTHOR INFORMATION

394

*Corresponding author: J. Fang, Phone: + 86-18680581522; e-mail: [email protected].

395

*Co-corresponding author: W. Song, Phone: +86-15821951698; e-mail: [email protected].

20


Page 20 of 34

Page 21 of 34

396 397

398


Notes The authors declare no competing financial interest.

ACKNOWLEDGMENTS

399

This research was supported by the Natural Science Foundation of China (21677181, 51378515,

400

21422702), the National Key Research Development Program of China (2016YFC0502803), the

401

Tip-top Scientific and Technical Innovative Youth Talents of Guangdong Special Support Program

402

(2015TQ01Z552), and the research fund program of the Guangdong Provincial Key Laboratory of

403

Environmental Pollution Control and Remediation Technology (2016K0004).

404 405 406

REFERENCES

407

(1) Richardson, S. D.; Ternes, T. A. Water analysis: Emerging contaminants and current issues. Anal.

408

Chem. 2014, 86, 2813-2848.

409

(2) Daughton, C. G.; Ternes, T. A. Pharmaceuticals and personal care products in the environment:

410

Agents of subtle change? Environ. Health Persp. 1999, 107, 907-938.

411

(3) Yan, S.; Song, W. Photo-transformation of pharmaceutically active compounds in the aqueous

412

environment: A review. Environ. Sci. Proc. Impacts. 2014, 16, 697-720.

413

(4) Huerta-Fontela, M.; Galceran, M. T.; Ventura, F. Occurrence and removal of pharmaceuticals

414

and hormones through drinking water treatment. Water Res. 2011, 45, 1432-1442.

415

(5) Kim, I.; Tanaka, H. Photodegradation characteristics of PPCPs in water with UV treatment.

416

Environ. Int. 2009, 35, 793-802.

417

(6) Westerhoff, P.; Yoon, Y.; Snyder, S.; Wert, E. Fate of Endocrine-Disruptor, pharmaceutical, and

21



418

personal care product chemicals during simulated drinking water treatment processes. Environ. Sci.

419

Technol. 2005, 39, 6649-6663.

420

(7) Yang, X.; Flowers, R. C.; Weinberg, H. S.; Singer, P. C. Occurrence and removal of

421

pharmaceuticals and personal care products (PPCPs) in an advanced wastewater reclamation plant.

422

Water Res. 2011, 45, 5218-5228.

423

(8) Sharma, V. K.; Triantis, T. M.; Antoniou, M. G.; He, X.; Pelaez, M.; Han, C.; Song, W.; O Shea,

424

K. E.; de la Cruz, A. A.; Kaloudis, T.; Hiskia, A.; Dionysiou, D. D. Destruction of microcystins by

425

conventional and advanced oxidation processes: A review. Sep. Purif. Technol. 2012, 91, 3-17.

426

(9) Wols, B. A.; Harmsen, D. J. H.; Wanders-Dijk, J.; Beerendonk, E. F.; Hofman-Caris, C. H. M.

427

Degradation of pharmaceuticals in UV (LP)/H2O2 reactors simulated by means of kinetic modeling

428

and computational fluid dynamics (CFD). Water Res. 2015, 75, 11-24.

429

(10) Kim, I.; Yamashita, N.; Tanaka, H. Performance of UV and UV/H2O2 processes for the removal

430

of pharmaceuticals detected in secondary effluent of a sewage treatment plant in Japan. J. Hazard.

431

Mater. 2009, 166, 1134-1140.

432

(11) Shu, Z.; Bolton, J. R.; Belosevic, M.; Gamal El Din, M. Photodegradation of emerging

433

micropollutants using the medium-pressure UV/H2O2 Advanced Oxidation Process. Water Res.

434

2013, 47, 2881-2889.

435

(12) Kim, I.; Yamashita, N.; Tanaka, H. Photodegradation of pharmaceuticals and personal care

436

products during UV and UV/H2O2 treatments. Chemosphere. 2009, 77, 518-525.

437

(13) Rosenfeldt, E. J.; Linden, K. G. Degradation of endocrine disrupting chemicals bisphenol a,

438

ethinyl estradiol, and estradiol during UV photolysis and advanced oxidation processes. Environ. Sci.

22


Page 22 of 34

Page 23 of 34


439

Technol. 2004, 38, 5476-5483.

440

(14) Zuorro, A.; Fidaleo, M.; Fidaleo, M.; Lavecchia, R. Degradation and antibiotic activity

441

reduction of chloramphenicol in aqueous solution by UV/H2O2 process. J. Environ. Manage. 2014,

442

133, 302-308.

443

(15) Wols, B. A.; Hofman-Caris, C. H. M.; Harmsen, D. J. H.; Beerendonk, E. F. Degradation of 40

444

selected pharmaceuticals by UV/H2O2. Water Res. 2013, 47, 5876-5888.

445

(16) Rosario-Ortiz, F. L.; Wert, E. C.; Snyder, S. A. Evaluation of UV/H2O2 treatment for the

446

oxidation of pharmaceuticals in wastewater. Water Res. 2010, 44, 1440-1448.

447

(17) Lee, Y.; Gerrity, D.; Lee, M.; Gamage, S.; Pisarenko, A.; Trenholm, R. A.; Canonica, S.;

448

Snyder, S. A.; von Gunten, U. Organic contaminant abatement in reclaimed water by UV/H2O2 and a

449

combined process consisting of O3/H2O2 followed by UV/H2O2: Prediction of abatement efficiency,

450

energy consumption, and byproduct formation. Environ. Sci. Technol. 2016, 50, 3809-3819.

451

(18) Keen, O. S.; Dotson, A. D.; Linden, K. G. Evaluation of hydrogen peroxide chemical quenching

452

agents following an advanced oxidation process. J. Environ. Eng. 2013, 139, 137-140.

453

(19) Wang, D.; Bolton, J. R.; Hofmann, R. Medium pressure UV combined with chlorine advanced

454

oxidation for trichloroethylene destruction in a model water. Water Res. 2012, 46, 4677-4686.

455

(20) Fang, J.; Fu, Y.; Shang, C. The roles of reactive species in micropollutant degradation in the

456

UV/Free chlorine system. Environ. Sci. Technol. 2014, 48, 1859-1868.

457

(21) Kong, X.; Jiang, J.; Ma, J.; Yang, Y.; Liu, W.; Liu, Y. Degradation of atrazine by UV/chlorine:

458

Efficiency, influencing factors, and products. Water Res. 2016, 90, 15-23.

459

(22) Wang, D.; Bolton, J. R.; Andrews, S. A.; Hofmann, R. UV/chlorine control of drinking water

23



460

taste and odour at pilot and full-scale. Chemosphere. 2015, 136, 239-244.

461

(23) Xiang, Y.; Fang, J.; Shang, C. Kinetics and pathways of ibuprofen degradation by the

462

UV/chlorine advanced oxidation process. Water Res. 2016, 90, 301-308.

463

(24) Wu, Z.; Fang, J.; Xiang, Y.; Shang, C.; Li, X.; Meng, F.; Yang, X. Roles of reactive chlorine

464

species in trimethoprim degradation in the UV/chlorine process: Kinetics and transformation

465

pathways. Water Res. 2016, 104, 272-282.

466

(25) Wang, W.; Wu, Q.; Huang, N.; Wang, T.; Hu, H. Synergistic effect between UV and chlorine

467

(UV/chlorine) on the degradation of carbamazepine: Influence factors and radical species. Water Res.

468

2016, 98, 190-198.

469

(26) Nam, S.; Yoon, Y.; Choi, D.; Zoh, K. Degradation characteristics of metoprolol during

470

UV/chlorination reaction and a factorial design optimization. J. Hazard. Mater. 2015, 285, 453-463.

471

(27) Qin, L.; Lin, Y.; Xu, B.; Hu, C.; Tian, F.; Zhang, T.; Zhu, W.; Huang, H.; Gao, N. Kinetic

472

models and pathways of ronidazole degradation by chlorination, UV irradiation and UV/chlorine

473

processes. Water Res. 2014, 65, 271-281.

474

(28) Sichel, C.; Garcia, C.; Andre, K. Feasibility studies: UV/chlorine advanced oxidation treatment

475

for the removal of emerging contaminants. Water Res. 2011, 45, 6371-6380.

476

(29) Yang, X.; Sun, J.; Fu, W.; Shang, C.; Li, Y.; Chen, Y.; Gan, W.; Fang, J. PPCP degradation by

477

UV/chlorine treatment and its impact on DBP formation potential in real waters. Water Res. 2016, 98,

478

309-318.

479

(30) Watts, M. J.; Hofmann, R.; Rosenfeldt, E. J. Low-pressure UV/Cl2 for advanced oxidation of

480

taste and odor. J. Am. Water Works Ass. 2012, 104, 47-48.

24


Page 24 of 34

Page 25 of 34


481

(31) Feng, Y.; Smith, D. W.; Bolton, J. R. Photolysis of aqueous free chlorine species (HOCl and

482

OCl-) with 254 nm ultraviolet light. J. Environ. Eng. Sci. 2007, 6, 277-284.

483

(32) Nowell, L. H.; Hoigné, J. Photolysis of aqueous chlorine at sunlight and ultraviolet wavelengths

484

—I. Degradation rates. Water Res. 1992, 26, 593-598.

485

(33) Nowell, L. H.; Hoigné, J. Photolysis of aqueous chlorine at sunlight and ultraviolet wavelengths

486

—II. Hydroxyl radical production. Water Res. 1992, 26, 599-605.

487

(34) Watts, M. J.; Linden, K. G. Chlorine photolysis and subsequent OH radical production during

488

UV treatment of chlorinated water. Water Res. 2007, 41, 2871-2878.

489

(35) Sun, P.; Lee, W.; Zhang, R.; Huang, C. Degradation of DEET and caffeine under UV/Chlorine

490

and simulated Sunlight/Chlorine conditions. Environ. Sci. Technol. 2016, 50, 13265-13273.

491

(36) NIST, National Institute of Standards and Technology. NDRL/NIST Solution Kinetcis

492

Database on the Web. Http://www3.nd.edu/~ndrlrcdc/index.html (Accessed February 20, 2017).

493

2017.

494

(37) Grebel, J. E.; Pignatello, J. J.; Mitch, W. A. Impact of halide ions on natural organic

495

Matter-Sensitized photolysis of 17β-Estradiol in saline waters. Environ. Sci. Technol. 2012, 46,

496

7128-7134.

497

(38) Alfassi, Z. B.; Huie, R. E.; Mosseri, S.; Neta, P. Kinetics of one-electron oxidation by the ClO

498

radical. J.phys.chem. 1987, 91, 85–88.

499

(39) Martire, D. O.; Rosso, J. A.; Bertolotti, S.; Le Roux, G. C.; Braun, A. M.; Gonzalez, M. C.

500

Kinetic study of the reactions of chlorine atoms and Cl2•- radical anions in aqueous solutions. II.

501

Toluene, benzoic acid, and chlorobenzene. J. Phys. Chem. A. 2001, 105, 5385-5392.

25



502

(40) Hasegawa, K.; Neta, P. Rate constants and mechanisms of reaction of chloride (Cl2•-) radicals.

503

J.phys.chem. 1978, 82, 854-857.

504

(41) Lee, Y.; von Gunten, U. Quantitative structure – activity relationships (QSARs) for the

505

transformation of organic micropollutants during oxidative water treatment. Water Res. 2012, 46,

506

6177-6195.

507

(42) Xiao, R.; Ye, T.; Wei, Z.; Luo, S.; Yang, Z.; Spinney, R. Quantitative structure‒activity

508

relationship (QSAR) for the oxidation of trace organic contaminants by sulfate radical. Environ. Sci.

509

Technol. 2015, 49, 13394-402.

510

(43) Bolton, J. R.; Linden, K. G. Standardization of methods for fluence (UV Dose) determination in

511

bench-scale UV experiments. J. Environ. Eng. 2003, 129, 209-215.

512

(44) Garoma, T.; Gurol, M. D. Modeling aqueous Ozone/UV process using oxalic acid as probe

513

chemical. Environ. Sci. Technol. 2005, 39, 7964-7969.

514

(45) APHA-AWWA-WEF, Standard methods for the examination of water and wastewater. 22nd

515

ed.; American Public Health Association/American Water Works Association/Water Environment

516

Federation: Washington DC, 2012.

517

(46) Hansch, C.; Leo, A.; Taft, R. W. A survey of Hammett substituent constants and resonance and

518

field parameters. Chem. Rev. 1991, 91, 165-195.

519

(47) Ianni, J. C., Kintecus Version 4.5. http://kintecus.com/. 2012.

520

(48) Westerhoff, P.; Chao, P.; Mash, H. Reactivity of natural organic matter with aqueous chlorine

521

and bromine. Water Res. 2004, 38, 1502-1513.

522

(49) Lian, L.; Yao, B.; Hou, S.; Fang, J.; Yan, S.; Song, W. Kinetic study of hydroxyl and sulfate

26


Page 26 of 34

Page 27 of 34


523

Radical-Mediated oxidation of pharmaceuticals in wastewater effluents. Environ. Sci. Technol. 2017,

524

51, 2954-2962.

525

(50) Crittenden, J. C.; Trussell, R. R.; Hand, D. W.; Howe, K. J.; Tchobanoglous, G., Water

526

treatment: Principles and design. John Wiley & Sons, Inc.: New Jersey, USA, 2005.

527 528

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Page 28 of 34

Captions of Tables and Figures Table 1. The estimated steady-state concentrations of radicals in the UV/chlorine process..................... 2 Figure 1. Comparison of the first-order rate constants (kobs′) of 27 PPCPs by direct UV photolysis, chlorination and the UV/chlorine processes in pure water at pH 7. The inset shows the ranges of the k′ attributable to OH• (kOH•′) and that attributable to RCS (kRCS′) of 27 PPCPs in the UV/chlorine process, the dash line and solid line show the medium value and mean value, respectively. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, average UV fluence rate = 0.78 mW cm-2............................................................................................................................................. 3 Figure 2. Correlations between k′ for the reactions of RCS and HO• with 8 aromatic PPCPs versus ∑σp+ at pH 7. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, average UV fluence rate = 0.78 mW cm-2..................................................................................................................................... 4 Figure 3. Effect of pH on the kobs′ of 27 PPCPs and the specific k′ (stacked bar) by UV, chlorine, OH• and RCS in the UV/chlorine process. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, average UV fluence rate = 0.78 mW cm-2. ........................................................................................ 5 Figure 4. Comparison of the kobs′ of 27 PPCPs and the specific k′ (stacked bar) by UV, chlorine, OH• and RCS with or without the presence of NOM in the UV/chlorine process at pH 7. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 20 μM, average UV fluence rate = 0.78 mW cm-2, [NOM]0 = 1 mg L-1. ................................................................................................................................................ 6 Figure 5. Comparison of the kobs′ of 27 PPCPs and the specific k′ (stacked bar) by UV, chlorine, OH• and RCS with or without the presence of bicarbonate in the UV/chlorine process at pH 7. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, [HCO3-]0 = 1 mM, average UV fluence rate = 0.78 mW cm-2. ................................................................................................................................ 7

ACS Paragon Plus Environment 1

Page 29 of 34


Table 1. The estimated steady-state concentrations of radicals in the UV/chlorine process HO• (M)

Cl• (M)

Cl2•- (M)

ClO• (M)

CO3•- (M)

pH 6a

2.55×10-13

1.05×10-14

1.46×10-14

1.24×10-10

-

pH 7a

1.52×10-13

1.01×10-14

1.40×10-14

1.24×10-10

-

pH 8a

7.48×10-14

8.78×10-15

1.22×10-14

1.24×10-10

-

pH 7a w. HCO3- (1 mM)

9.61×10-14

5.32×10-15

3.22×10-15

1.23×10-10

1.53×10-9

pH 7b

1.43×10-13

1.75×10-14

4.87×10-14

1.72×10-10

-

pH 7b w. NOM (1 mg L-1)

1.05×10-13

1.68×10-14

4.68×10-14

2.01×10-13

-

Note: PPCPs were not included in the model simulation. a. Conditions: average UV fluence rate = 0.78 mW cm-2, [chlorine]0 = 10 μM. b. Conditions: average UV fluence rate = 0.78 mW cm-2, [chlorine]0 = 20 μM.



Group I

.10

k' (min-1)

.15

5

Group IV

Group III

4

Left bar: UV Medium bar: Chlorine Right bar: UV/chlorine

1 .1

3 .01 2

.001 OH·

RCS

Sulfamethoxazole

Bisphenol A

Salbutamol

Ractopamine

Trimethoprim

Diclofenac

Clenbuterol

Propranolol

Naproxen

Gemfibrozil

Caffeine

Carbamazepine

Roxithromycin

Azithromycin

Erythromycin

Metoprolol

Atenolol

Flumequine

Nalidixic acid

Venlafaxine

Primidone

Ibuprofen

Ronidazole

0

Metronidazole

0.00

Ornidazole

1

Tinidazole

.05

Dimetridazole

kobs' (min-1)

.20

Group II

kobs' (min-1)

.25

Page 30 of 34

Figure 1. Comparison of the first-order rate constants (kobs′) of 27 PPCPs by direct UV photolysis, chlorination and the UV/chlorine processes in pure water at pH 7. The inset shows the ranges of the k′ attributable to OH• (kOH•′) and that attributable to RCS (kRCS′) of 27 PPCPs in the UV/chlorine process, the dash line and solid line show the medium value and mean value, respectively. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, average UV fluence rate = 0.78 mW cm-2.


Page 31 of 34


2.5 RCS HO·

Trimethoprim 2.0 y = -0.8103x - 0.2586 R2 = 0.8879

k' (min-1)

1.5 Ractopamine

Bisphenol A

1.0 Gemfibrozil .5

Bisphenol A

Trimethoprim 0.0

-3.0

Salbutamol

y = -0.0117x + 0.0256 R2 = 0.6695

Ractopamine

-2.5

-2.0

Gemfibrozil

-1.5 ∑ σ

Atenolol

Primidone Ibuprofen

Atenolol Primidone Salbutamol Ibuprofen -1.0

-.5

0.0

+ p

Figure 2. Correlations between k′ for the reactions of RCS and HO• with 8 aromatic PPCPs versus ∑σp+ at pH 7. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, average UV fluence rate = 0.78 mW cm-2.



Page 32 of 34

.4

5 Group II

Group I

3 .2 2

-1 kobs' (min )

4

Left bar: pH 6 Medium bar: pH 7 Right bar: pH 8

.3

kobs' (min-1)

Group IV

Group III

.1 1

Sulfamethoxazole

Bisphenol A

Ractopamine

Salbutamol

Trimethoprim

Diclofenac

Clenbuterol

Propranolol

Naproxen

Gemfibrozil

Caffeine

Carbamazepine

Roxithromycin

Azithromycin

Erythromycin

Atenolol

Metoprolol

Flumequine

Nalidixic acid

Venlafaxine

Primidone

Ibuprofen

Ronidazole

Metronidazole

Ornidazole

Tinidazole

0

Dimetridazole

0.0

Figure 3. Effect of pH on the kobs′ of 27 PPCPs and the specific k′ (stacked bar) by UV, chlorine, OH• and RCS in the UV/chlorine process. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, average UV fluence rate = 0.78 mW cm-2.


Page 33 of 34


.28

Group I

7

Group IV

Group III

Group II

.24

6 UV Chlorine OH· RCS

.12

5 4

Left bar: control Right bar: w. 1-mg L-1 NOM

3

Sulfamethoxazole

Bisphenol A

Ractopamine

Salbutamol

Trimethoprim

Diclofenac

Clenbuterol

Propranolol

Naproxen

Gemfibrozil

Caffeine

Carbamazepine

Roxithromycin

Azithromycin

Erythromycin

Metoprolol

Atenolol

Flumequine

Nalidixic acid

Venlafaxine

Primidone

0

Ibuprofen

0.00

Ronidazole

1

Metronidazole

.04

Ornidazole

2

Tinidazole

.08

kobs' (min-1)

.16

Dimetridazole

kobs' (min-1)

.20

Figure 4. Comparison of the kobs′ of 27 PPCPs and the specific k′ (stacked bar) by UV, chlorine, OH• and RCS with or without the presence of NOM in the UV/chlorine process at pH 7. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 20 μM, average UV fluence rate = 0.78 mW cm-2, [NOM]0 = 1 mg L-1.



Group I

kobs' (min-1)

4

UV Chlorine OH· RCS

.15

.10

5

Group IV

Group III

Group II

3

Left bar: control Right bar: w. 1-mM HCO3-

2

kobs' (min-1)

.20

Page 34 of 34

.05 1

Sulfamethoxazole

Bisphenol A

Ractopamine

Salbutamol

Trimethoprim

Diclofenac

Clenbuterol

Propranolol

Naproxen

Gemfibrozil

Caffeine

Carbamazepine

Roxithromycin

Azithromycin

Erythromycin

Metoprolol

Atenolol

Flumequine

Nalidixic acid

Venlafaxine

Primidone

Ibuprofen

Ronidazole

Metronidazole

Ornidazole

Tinidazole

0

Dimetridazole

0.00

Figure 5. Comparison of the kobs′ of 27 PPCPs and the specific k′ (stacked bar) by UV, chlorine, OH• and RCS with or without the presence of bicarbonate in the UV/chlorine process at pH 7. Conditions: [PPCP]0 = 1 μg L-1, [chlorine]0 = 10 μM, [HCO3-]0 = 1 mM, average UV fluence rate = 0.78 mW cm-2.


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