Effects of HCO3â on Degradation of Toxic Contaminants of Emerging

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Remediation and Control Technologies 3-

Effects of HCO on Degradation of Toxic Contaminants of Emerging Concern by UV/NO

3-

Ying Huang, Minghao Kong, Danielle Westerman, Elvis Genbo Xu, Scott Coffin, Kristin H. Cochran, Yiqing Liu, Susan D. Richardson, Daniel Schlenk, and Dionysios D. Dionysiou Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b04383 • Publication Date (Web): 04 Oct 2018 Downloaded from http://pubs.acs.org on October 5, 2018

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Environmental Science & Technology

Effects of HCO3− on Degradation of Toxic Contaminants of

1

Emerging Concern by UV/NO3−

2 3

Ying Huang1, Minghao Kong1, Danielle Westerman2, Elvis Genbo Xu3, Scott Coffin3,

4

Kristin H. Cochran2, Yiqing Liu1,4, Susan D. Richardson2, Daniel Schlenk3, and Dionysios D.

5

Dionysiou1*

6

1Environmental

7

Engineering, University of Cincinnati, Cincinnati, Ohio 45221, USA

8

2Department

9

Carolina 29208, USA

Engineering and Science, Department of Chemical and Environmental

of Chemistry and Biochemistry, University of South Carolina, Columbia, South

10

3Department

11

USA

12

4Faculty

13

Chengdu 611756, China

of Environmental Sciences, University of California, Riverside, California 92521,

of Geosciences and Environmental Engineering, Southwest Jiaotong University,

14 15 16 17

*Correspondence to: Dionysios D. Dionysiou ([email protected])

ACS Paragon Plus Environment


19

Abstract

20

This study investigated the significant influence of HCO3− on the degradation of

21

contaminants of emerging concern (CECs) during nitrate photolysis at 254 nm for water reuse

22

applications. The second-order rate constants for the reactions between selected contaminants with

23

carbonate radical (CO3•−) were determined at pH = 8.8 and T = 20 °C: estrone ((5.3 ± 1.1) × 108

24

M−1 s−1), bisphenol A ((2.8 ± 0.2) × 108 M−1 s−1), 17α-ethynylestradiol ((1.6 ± 0.3) × 108 M−1 s−1),

25

triclosan ((4.2 ± 1.4) × 107 M−1 s−1), diclofenac ((2.7 ± 0.7) × 107 M−1 s−1), atrazine ((5.7 ± 0.1) ×

26

106 M−1 s−1), carbamazepine ((4.2 ± 0.01) × 106 M−1 s−1), and ibuprofen ((1.2 ± 1.1) × 106 M−1 s−1).

27

Contributions from UV, reactive nitrogen species (RNS), hydroxyl radical (•OH), and CO3•− to the

28

CEC decomposition in UV/NO3− in the presence and absence of HCO3− were investigated. In

29

addition, possible transformation products and degradation pathways of triclosan, diclofenac,

30

bisphenol A, and estrone in UV/NO3−/HCO3− were proposed based on the mass (MS) and MS2

31

spectra. Significant reduction in the cytotoxicity of bisphenol A was observed after the treatment

32

with UV/NO3−/HCO3−.

33

34

Keywords

35

Carbonate radical, UV/NO3−, Contaminants of emerging concern, Cytotoxicity, Water reuse


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Page 3 of 40

36


TOC art ONOO h

CO2

NO3

h

h O2

+ NO2

NO

+O

2

h

37

NO2

h

or +

3

/O

+ NO + OH

H

NO2

OH +HCO3

H

/O

+O

H

OH

h +C O

ONOO

h


CO3


38

Page 4 of 40

Introduction

39

Nitrate (NO3−) photolysis causes the formation of reactive oxygen species (ROS) and

40

nitrogen species (RNS), nitrite (NO2−) and peroxynitrite (ONOO−), which may pose a health threat

41

in water undergoing treatment1, 2. Irradiation of water containing NO3− under low-pressure (LP)-

42

UV at 254 nm (UV/NO3−) primarily leads to the formation of ONOO−, nitrogen dioxide radical

43

(•NO2), and hydroxyl radical (•OH)3-6. Nitrite formation is mainly ascribed to the decomposition

44

of peroxynitrite when the irradiation wavelength is lower than 280 nm2, 4. Nitrogen oxide radical

45

(•NO) is subsequently produced via the photolysis of nitrate, nitrite, peroxynitrite, and the

46

substantial oxidation of peroxynitrite by •OH7-9. Nitrate photolysis is highly dependent on the

47

reaction pH and the irradiation wavelength6. The primary reactions of nitrate photolysis at 254 nm

48

and high pH (7 ≤ pH < 9) in solutions containing dissolved O2 are summarized in Figure 1.

49

Carbonate radical (CO3•−) can also be formed with low yield during the NO3− photolysis in surface

50

water due to the presence of CO2 via the decomposition of an adduct ONOOC(O)O− produced2.

51

As prevalent anions in natural waters, HCO3−/CO32− play a significant role in NO3− photolysis,

52

affecting the RNS speciation via enhancing the ONOOC(O)O− formation and quenching •OH with

53

the generation of CO3•− (eqs. 1-2)5, 6, 10, 11.

54

•

OH + HCO3― → CO•3 ― + H2O

k = 8.5 × 106 M ―1s ―1

(1)

55

•

OH + CO23 ― → CO•3 ― + OH ―

k = 3.9 × 108 M ―1s ―1

(2)

56

Among the reactive species in UV/NO3− in the presence of HCO3− (UV/NO3−/HCO3−), •OH

57

(E0 = 2.0 V12) is a non-selective oxidant, which could react with various contaminants at high rate

58

constants (> 109 M−1 s−1) through electron transfer, H-abstraction, and radical addition13. These

59

three routes are favorable for •NO2 (E0 = 1.03 V) as well, but it selectively oxidizes anilines,

60

phenolic moieties, phenothiazines, thiols, and ascorbate at lower rate constants12. •NO2-addition


Page 5 of 40


61

was observed on the ortho or para-position of -OH group on the aromatic ring due to the

62

electrophilic character2,

63

derivatives stimulate high environmental concern11. Nucleophilic addition on the carbonyl

64

moieties is a well-established pathway for ONOO− (E0 = 1.03 V)2, and it could selectively react

65

with phenolic moieties at relatively high rates14. ROS other than •OH, such as O2•− (consumed

66

rapidly by •NO2 and •NO), O•− (pKa = 11.9), and O(3P) (Φ > 0.1%), are not considered in this study

67

due to their low concentration and weak oxidation ability at high pH under UV254nm 4, 6. The RNS

68

and •OH produced can be utilized to decompose CECs from wastewater during the UV irradiation

69

in the presence of NO3− 11, 15-16. Nevertheless, studies about the effects of HCO3− on various CEC

70

degradation in LP-UV/NO3−, and the resulting environmental impacts of treated solutions are

71

limited11, 16, 17.

11.

The toxicity and potential mutagenicity of the produced nitro-

72

As an important one-electron oxidant (E0 = 1.23 V18), the contribution from CO3•− to

73

contaminant removal is nonnegligible, which was demonstrated with the treatment of

74

oxytetracycline19, 20 and cylindrospermopsin21. Electron-rich aromatic compounds are preferred

75

for CO3•− attack at relatively high reaction rates, especially phenolic and aromatic amine

76

moieties19, 22, 23. Electron transfer is the dominant route for CO3•−-oxidation, removing an electron

77

from the basic nitrogen atom of an aniline group or from the oxygen atom of a phenolic group to

78

produce a carbonate anion and an aniline/phenol radical cation22-2, 23. H-abstraction for CO3•− is

79

slow24, and radical-addition is not favored due to steric inhibition22. Limited information could be

80

found about the second-order rate constants of CO3•− with CECs20, 21, 25-30. However, few studies

81

evaluated the contribution of CO3•− in removing CECs for water reuse. Moreover, estrone, 17α-

82

ethynylestradiol, diclofenac, triclosan, bisphenol A, and ibuprofen are ranked of high concern for

83

water reuse by a State of California expert panel due to their lack of removal in wastewater



84

treatment and their presence at levels that could pose an ecological or human health risk31, 32. For

85

this reason, degradation of these compounds is investigated in this research.

86

The goals of this research are to investigate the influence of alkalinity on UV/NO3−

87

treatment and to provide a fundamental understanding of the selective oxidation of CECs with

88

RNS and CO3•−. The degradation kinetics of selected CECs in the UV/NO3− process with and

89

without HCO3− were evaluated. The second-order rate constants of the selected CECs with CO3•−

90

were determined. Contributions to CEC degradation from UV photolysis, •OH oxidation, CO3•−

91

oxidation, and RNS oxidation were quantified to evaluate the impacts of HCO3−. Transformation

92

products of diclofenac, triclosan, bisphenol A, and estrone in UV/NO3−/HCO3− treatment were

93

detected to elucidate the resulting cytotoxicity of these treated solutions. The RO permeate from

94

the Orange County Water District’ Groundwater Replenishment System (GWRS) was used as a

95

reaction matrix with spiked addition of chemicals (CECs, NO3−, and HCO3−) to assess the

96

applications of UV/NO3−/HCO3− to remove CECs in nitrate/carbonate-rich water reuse scenarios.

97

Experiment Section

98

Chemicals.

99

Eight selected CECs were 17α-ethynylestradiol (EE2), estrone (E1), diclofenac (DCF),

100

triclosan (TCS), bisphenol A (BPA), atrazine (ATZ), carbamazepine (CBZ), and ibuprofen (IBP).

101

Analytical standards were purchased from Sigma-Aldrich at the highest available purity. Their

102

structures and properties are listed in Table 1. RO permeate was collected from the GWRS indirect

103

potable reuse project (April 25, 2017). The GWRS purifies secondary-treated wastewater effluent

104

via chloramination, microfiltration, RO, UV/H2O2, followed by post-treatment stabilization; the

105

general water quality parameters of the supplied RO permeate are summarized in Table S1. The

106

other chemicals and reagents used are depicted in Text S1.


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107

Analysis Methods.

108

All contaminants and probe compounds were determined with an 1100 Series HPLC

109

(Agilent) equipped with a diode array detector. A Supelco Discovery C18 HS column (2.1 mm

110

×150 mm, 5 µm) was used before a Poroshell 120 EC-C18 column (2.1 mm × 150 mm, 4 µm) was

111

obtained and utilized. Detailed HPLC conditions may be found in Text S2 and Table S2. The

112

concentration of nitrate and chloride ions were determined using a Dionex ion chromatograph with

113

IonPac AS18 column (2 × 250 mm). The mobile phase was comprised of 1.0 mM NaHCO3 and 3.5

114

mM Na2CO3. Concentrations of free chlorine, total chlorine, and monochloramine were

115

determined following the standard analysis methods33.

116

Transformation products of DCF, TCS, BPA, and E1 were detected using an Agilent 1290

117

infinity HPLC with an Agilent 6540 quadrupole time-of-flight mass spectrometer (LC-Q-TOF-

118

MS) at University of Cincinnati, equipped with an Eclipse XDB-C18 column (2.1 mm × 50 mm ×

119

3.5 µm, Agilent ZORBAX). Transformation products of DCF, TCS, BPA were also studied suing

120

an Agilent 6545 LC-Q-TOF-MS with a Poroshell C18 column (2.1 mm × 150 mm × 2.7 µm, Agilent

121

InfinityLab) at University of South Carolina. Transformation products were not measured for EE2

122

due to the similar structure to E1, and were not analyzed for ATZ, CBZ, and IBP either due to the

123

inefficient removal in UV/NO3−/HCO3−. Detailed LC/MS/MS conditions are found in Text S3 and

124

Table S3. Mass spectra were analyzed by Agilent Mass Hunter B.04.00 software.

125

Photochemical Experiments.

126

Photolysis experiments were carried out in a bench scale photochemical apparatus installed

127

with two 15 W low-pressure mercury UV lamps (Cole-Parmer) with monochromatic UV at 254

128

nm. The average UV fluence rate was measured as 0.1 mW cm−2 34. A round Petri dish (60 × 15

129

mm) was used as the reactor, which was covered by a quartz cover (Quartz Scientific Inc., OH) to



130

minimized evaporation during the reactions34, 35. A typical experiment was as follows: each tested

131

CEC was spiked into Milli-Q water with an initial concentration of 1 µM and a total volume of 10

132

mL. The initial concentrations of NO3− and HCO3− were added as 10 mM and 3 mM unless

133

specified elsewhere. At different time intervals, 200 µL of the reaction solutions were sampled,

134

quenched with 50 µL of methanol (MeOH), and analyzed by HPLC. No loss of the selected CECs

135

was observed in dark. For detecting the transformation products of DCF, TCS, BPA, and E1, a

136

higher initial concentration was applied. The pH of UV/NO3−/HCO3− process was maintained at

137

8.8 due to the presence of 3 mM HCO3−; borate-boric acid solution (10 mM) was used to stabilize

138

the pH at 8.8 when no HCO3− was added36. The RO permeate water sample was filtered once

139

received and stored at 4 °C if not used immediately. CECs were spiked into RO permeate to test

140

the performance of UV/NO3−/HCO3− within one week of receiving samples. All experiments were

141

performed in triplicate.

142

Determining the Rate Constants for the CECs with CO3•−.

143

Competition kinetic studies were conducted with UV/H2O2 in the presence of 3 mM HCO3−

144

using 1 µM 4-chlorophenol (4-CP) as the reference substance for CO3•− (kCO•3 ― ,4 - CP = 1.9 × 108

145

M−1 s−1)24 and 10 mM tert-butanol (t-BuOH) as the quenching agent for •OH (k•OH,t - BuOH = 6 ×

146

108 M−1 s−1, kCO•3 ― ,t - BuOH < 1.6 × 102 M−1 s−1)12, 37. UV/H2O2 was utilized since the contributions

147

from RNS to the CEC degradation in UV/NO3− and UV/NO3−/HCO3− might be different. The

148

kCO•3 ― ,CEC was calculated using eq. 3, where k'CEC and k'4-CP are the observed degradation rate

149

constants of the selected CEC and 4-CP. The k'CEC and k'4-CP in the UV/H2O2 (1 mM)/t-BuOH (10

150

mM) and UV/H2O2 (1 mM)/HCO3− (3 mM)/t-BuOH (10 mM) system were summarized in Table

151

S4.


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Page 9 of 40

152


k

CO•3 ― ,CEC

= k

CO•3 ― ,4

- CP

×

k'CEC,UV '

k 4 - CP,UV

― H2O2 HCO3 t - BuOH ― H2O2 HCO3 t - BuOH

― k'CEC,UV

H2O2 pH 8.8 buffer t - BuOH

'

― k 4 - CP,UV

(3)

H2O2 pH 8.8 buffer t - BuOH

153

Determining the Steady-State Concentrations of •OH and CO3•−.

154

The concentrations of NO3− were not reduced significantly after the complete removal of

155

CECs under the current conditions (Table S5), therefore, steady-state radical concentrations could

156

be assumed. The steady-state concentration of •OH ([•OH]ss) in the UV/NO3− systems was

157

quantified indirectly by monitoring the degradation of 50 µM of nitrobenzene (NB) as the •OH

158

probe. The depletion of NB by UV only, UV/NO3−, and UV/NO3−/HCO3− was measured by HPLC.

159

Moreover, the degradation of NB with UV/NO3−/10 mM MeOH was almost the same as with UV

160

alone at pH 8.8 (Figure S1), confirming the negligible oxidation of NB by RNS. Therefore, the

161

[•OH]ss in the UV/NO3− system was calculated by using eq. 4.

162

―

163

where kobs is the observed degradation rate constant of NB in the UV/NO3− system, kphotolysis is the

164

observed first-order rate constant of NB under direct UV irradiation, and k•OH,NB (3.9 × 109 M−1

165

s−1)12 is the second-order rate constant of NB with •OH, respectively.

d[NB] dt

= k•OH,NB[•OH]ss[NB] + 0.1 mW cm ―2 × kphotolysis[NB] = kobs[NB]

(4)

166

[•OH]ss was quantified using NB in the UV/NO3−/HCO3− system through eq. 5, since the

167

second-order rate constant of NB with CO3•− (kCO•3 ― ,NB) is lower than 1.3  102 M−1 s−1 23, which

168

is much lower than k•OH, NB. The calculated [•OH]ss was subsequently used to determine [CO•3 ― ]

169

via another probe compound, 4-chlorophenol (4-CP) through eq. 6, which has high rate constants

170

with both •OH (k•OH,4 - CP = 7.6  109 M−1 s−1)12 and CO3•− (kCO•3 ― ,4 - CP = 1.9  108 M−1 s−1)24.

171

Although 4-CP could also react with peroxynitrite, the extremely low second-order rate constant (

172

kPN,4 - CP = 5.1 M−1 s−1)38 made the oxidation by peroxynitrite (PN) negligible in the presence of


ss


173

•OH

174

―

175

(5) ―

176

Page 10 of 40

and CO3•−.

d[NB] dt

'

k•OH,NB[•OH]ss [NB] + kCO•3 ― ,NB[CO•3 ― ]ss[NB] + 0.1 mW cm ―2 × kphotolysis[NB]

=

d[4 - CP] dt

'

= k•OH,4 - CP[•OH]ss [4 - CP] + kCO•3 ― ,4 - CP[CO•3 ― ]ss[4 - CP] +

0.1 mW cm ―2 kphotolysis

(6)

177

[4 - CP] + kPN,4 - CP[PN][4 - CP]

178

where kphotolysis is the observed first-order rate constant of NB/4-CP under direct UV irradiation.

179

Detailed information about calculating [•OH]ss and [CO3•−]ss is shown in Text S4, and degradation

180

of NB and 4-CP are shown in Figure S1 and Figure S2.

181

Calculating the Contributions from UV photolysis, •OH, CO3•−, and RNS.

182

The degradation of CECs by UV/NO3− with or without the addition of HCO3− can be

183

attributed to the direct UV photolysis, •OH oxidation, CO3•− oxidation, and RNS oxidation as

184

shown in eq. 7. ―

185

d[CEC] dt

= 0.1 mW cm ―2 × kphotolysis[CEC] + k•OH,CEC[•OH][CEC] + kCO•3 ― ,CEC[CO•3 ― ] [CEC] (7)

186

+ kRNS,CEC[RNS][CEC]

187

where kphotolysis is the observed first-order rate constant of CEC under direct UV irradiation; and

188

k•OH,CEC, kCO•3 ― ,CEC, and kRNS, CEC represents the second-order rate constant of CEC with •OH,

189

CO3•−, and RNS, respectively. The contributions of each process to CEC degradation was

190

calculated through eq. 8 following the methods previously reported by Fang et al. 39, 40.

191

R=

[CEC]0 ― [CEC]t [CEC]0

t

=

∫0kphotolysis[CEC]dt [CEC]0

t

t

+

∫0k•OH,CEC[•OH][CEC]dt [CEC]0

+

∫0kCO• ― ,CEC[CO•3 ― ][CEC]dt


3

[CEC]0

+

Page 11 of 40


t

192

∫0kRNS,CEC[RNS][CEC]dt [CEC]0

= Rphotolysis + R•OH + RCO•3 ― + RRNS

(8)

193

where R is the fractional removal of the CEC, [CEC]0 is the initial concentration of the CEC, and

194

[CEC]t is the concentration of the CEC at a specific reaction time. The degradation of CECs in

195

UV/NO3− and UV/NO3−/HCO3− can be found in Figure S3. Detailed information about the

196

calculation could be found in Text S5. Table S6 is an example to show the calculations on the

197

contributions from UV, •OH, CO3•−, and RNS to the degradation of DCF in the UV/NO3−/HCO3−

198

process. Results for selected CECs are shown in Figure S4-S5.

199

Cytotoxicity Analysis.

200

Respective cytotoxicity studies for DCF, TCS, E1, and BPA treated by UV/NO3−/HCO3−

201

were carried out using the 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide (MTT)

202

assay in GeneBLAzer CYP1A1-bla LS-180 cells (Life Technologies, Carlsbad, CA)41. Solid phase

203

extraction (SPE) was used to extract and concentrate samples for cytotoxicity analyses. Details

204

can be found in Text S6. Green (650 nm) and blue (595 nm) absorbance of analytes were measured

205

on a SpectraMax+ 384 plate reader (Molecular Devices, San Jose, CA). The resulting blue: green

206

ratio provides a normalized reporter response, with the higher value indicating lower cytotoxicity.

207

All sample groups were analyzed in triplicate at the concentration of 2.5% MeOH.

208

Results and Discussion

209

CEC Degradation Kinetics in UV/NO3−/HCO3−.

210

The UV fluence-based pseudo-first-order reaction rate constants (kobs) of the eight CECs

211

in different systems at pH 8.8 are compared in Figure 2, namely UV only, UV/HCO3−, UV/NO3−

212

and UV/NO3−/HCO3−. In the absence of HCO3−, only DCF and TCS could be removed by UV

213

only, with a relatively high kobs of 6.0 × 10−3 and 8.0 × 10−3 cm2 mJ−1, respectively. The addition



214

of 3 mM HCO3− did not significantly affect the kobs for most of the CECs by UV only. However,

215

the kobs in the presence of 3 mM HCO3− in the UV/NO3− system increased by 61%-177% for CECs

216

with either secondary amine (DCF) or phenolic groups (TCS, E1, EE2, and BPA). For other CECs

217

with weakly electron-donating moieties (ATZ, CBZ, and IBP), kobs decreased by 33%-56%.

218

Through comparing the structure of these eight CECs (Table 1), it was evident that the CECs with

219

phenolic and aniline groups were degraded at a higher rate than other types of CECs with

220

UV/NO3−/HCO3− treatment.

221

Contributions of UV, •OH, CO3•−, and RNS to the CEC Degradation.

222

With the addition of HCO3−, more CO3•− would be formed due to the reaction with •OH5,

223

24.

On the other hand, the scavenging of •OH would lower the Ф(NO2−)3 and enhanced the

224

concentrations of RNS. To further explore the roles of UV, •OH, CO3•−, and RNS on the

225

degradation of E1, EE2, DCF, TCS, BPA, ATZ, CBZ, and IBP in UV/NO3− with or without

226

HCO3−, calculations were conducted following methods reported by Fang et al.39, 40. Steady-state

227

concentrations of •OH and CO3•− in UV/NO3−(10 mM)/HCO3− (3 mM) were assumed and

228

measured as 6.25  10−15 M and 6.91  10−14 M, respectively.

229

UV photolysis. For the selected CECs, only DCF and TCS could be removed at a high rate

230

by direct UV irradiation. The contributions of UV to the removal of DCF and TCS were significant

231

in UV/NO3− and were reduced by 10% and 22% in the presence of HCO3−. UV alone was not able

232

to efficiently decompose other compounds, and thus it made little contribution to their removal,

233

except for ATZ (9%). For these less photo-degradable CECs, the addition of HCO3− did not

234

significantly change the contributions of direct UV photolysis in UV/NO3− at pH 8.8.

235 236

•OH

oxidation. In the absence of HCO3−, ROS (•OH and O2•−) were generated, with a •OH

yield of 9% at 254 nm42. Contributions of O2•− were ignored since O2•− is less reactive than •OH2.


Page 12 of 40

Page 13 of 40


237

The removal percentage of the selected CECs ascribed to •OH oxidation was in the range of 5% to

238

41% in UV/NO3− at 172 mJ cm−2 UV fluence (Figure 3). The contribution from •OH-oxidation

239

was directly related to the second-order rate constants of the selected CECs with •OH. E1 has the

240

highest second-order rate constant with •OH, leading to highest removal percentage by •OH

241

oxidation (41%). ATZ and TCS have the low rate constants with •OH, where removal percentages

242

attributable to •OH-oxidation were only 6% and 5%, respectively. For other CECs (BPA, EE2,

243

DCF, IBP, and CBZ) with k•OH,CEC in the range of 7.4 - 10  109 M−1 s−1, the removal percentages

244

attributed to •OH-oxidation were in the range of 12%-20%. Although the second-order rate

245

constants of selected CECs are higher than 109 M−1 s−1, the steady-state concentration of •OH was

246

measured as 1.56  10−14 M with 10 mM NO3−. Thus, the removal percentage of selected CECs

247

due to •OH oxidation was not higher than 41% in UV/NO3− (10 mM).

248

In the presence of HCO3−, the removal percentage of selected CECs due to •OH-oxidation

249

was eliminated to the range of 1% to 9% in UV/NO3− at 172 mJ cm−2 UV fluence (Figure 3). The

250

dominant reason is that the steady-state concentration of •OH was decreased to 6.25  10−15 M in

251

presence of 3 mM HCO3−. However, the overall degradation rate was enhanced by HCO3− for

252

BPA, E1, EE2, TCS, and DCF in the UV/NO3− system. Therefore, the elimination of •OH by

253

adding HCO3− might be the reason of decreased degradation rate for ATZ, CBZ, and IBP, but not

254

the dominant reason for others.

255

CO3•− oxidation. In the absence of HCO3−, the CO3•− can be generated via the reaction of

256

CO2 with ONOO− that is formed during the photolysis of NO3− (eq. 9-10) in alkaline solutions2.

257

Nevertheless, the kCO2,ONOO ― is as low as (2.9 ± 0.3) × 104 M−1 s−1, and the formed adduct,

258

ONOOC(O)O−, decomposes to only 33% NO2• and CO3•− 2. The insignificant amount of CO3•−

259

was not measured and the contributions of CO3•− were negligible for selected CECs in absence of



260

HCO3−.

261

ONOO ― + CO2 → ONOOC(O)O ―

262

ONOOC(O)O ― → CO•3 ― + •NO2

Page 14 of 40

(9) (10)

263

Oxidation reactions with CO3•− are highly selective reactions, as the rate constants range

264

over about 5 orders of magnitude23. Among these investigated CECs, the compounds with phenolic

265

groups exhibited higher rate constants at pH 8.8, including BPA, E1, EE2, and TCS. Noticeably,

266

the kCO•3 ― ,BPA is much higher than the previous reported one, which was determined in UV/Na2CO3

267

using ATZ probe compound27. The difference might be resulted from the extremely low yield of

268

CO3•− in UV/Na2CO3. DCF, which has an aniline group, reacted with CO3•− at a moderate rate of

269

2.7  107 M−1 s−1. For other CECs with weakly electron-donating moieties, such as ATZ, CBZ,

270

and IBP, the second-order rate constants are lower than 107 M−1 s−1. Although the steady-state

271

concentration of CO3•− (6.91  10−14 M) was 10 times higher than that of •OH with the addition of

272

HCO3−, kCO•3 ― ,CEC (106-108 M−1 s−1) of selected CECs are much lower than k•OH,CEC (109-1010 M−1

273

s−1) so that only the removal percentages of BPA, E1, and EE2 ascribable to CO3•− oxidation are

274

visible in Figure 3, while the contributions of CO3•− to remove other CECs are too low to be

275

depicted in the Figure 3. Thus, the elevated concentration of CO3•− in the presence of HCO3− likely

276

contributes to the enhanced degradation rate of BPA, E1, and EE2, rather than for DCF and TCS.

277

RNS oxidation. During the UV photolysis of NO3− at 254 nm, reactive nitrogen species

278

(RNS) were produced, namely ONOO−, •NO2, •NO, NO2−, and peroxynitrite (O2NOO−). Without

279

the addition of HCO3−, EE2 had the largest removal percentage ascribed to RNS (85%), somewhat

280

higher than that of BPA (66%). E1, TCS, and DCF had moderate removal percentages of 21%,

281

29%, and 15%, respectively. RNS contributed to only 5% ATZ degradation, 4% CBZ degradation,

282

and 3% IBP degradation. RNS-based oxidation has preference for CECs with electron-rich


Page 15 of 40


283

moieties3, 24, 38. For example, the second-order rate constants with peroxynitrite range from 103 to

284

106 M−1 s−1 1, and the phenolic group is the reactive site for peroxynitrite5, 14, 43. •NO2 can oxidize

285

electron-rich moieties with moderate rate constants, such as phenolic groups, anilines,

286

phenothiazines, and thiols24. The reaction rate constants between most of the RNS and selected

287

CECs are not available due to the complex constituents of NO3− related reactions and the low redox

288

potential (e.g., •NO has an E0 = 0.39 V44). Therefore, the contributions of RNS could not be isolated

289

for each reactive species.

290

In the presence of HCO3−, the removal percentages attributable to RNS enhanced

291

dramatically by 29%, 64%, 37%, and 31% for BPA, E1, TCS, and DCF, respectively. However,

292

the contributions from RNS did not change significantly with the addition of HCO3− for ATZ (-

293

3%), CBZ (-2%), and IPB (3%), respectively. HCO3− affected the RNS species and concentrations

294

via quenching •OH that could minimize the concentration of RNS through eqs. 11-14. In the

295

presence of •OH scavengers, NO2− yields were lowered at pH 83, and ONOO− yields were

296

increased at pH 134, which generates more reactive species, such as •NO2, •NO, and O2•−.

297

Therefore, the change of RNS concentrations should be responsible for the enhancement or

298

reduction of the degradation efficiency of CECs tested when HCO3− was present in the UV/NO3−

299

reaction.

300

•

OH + NO2― → OH ― + •NO2

301

•

OH + •NO2 → ONOOH

302

•

OH + •NO2 → NO3― + H +

(13)

303

•

OH + ONOO ― → OH ― + O2 + •NO

(14)

(11) (12)

304

The Effects of HCO3− in UV/NO3− Treatment.

305

According to the calculations and discussions above, the CECs are classified into three



306

groups based on the contributions of UV photolysis, •OH, CO3•−, and RNS in UV/NO3− treatment

307

with the addition of HCO3−. Group I, as less photo-degradable CECs, were primarily decomposed

308

by •OH, CO3•−, and RNS even though they might not have the highest overall removal rate

309

constants. This group include the CECs with phenolic groups, namely BPA, E1, and EE2. They

310

could react rapidly with CO3•− (> 108 M−1 s−1) and RNS5, 14, 43. This group probably contains CECs

311

with aniline moieties, which deserves further exploration.

312

Group II, as photo-degradable CECs, could be efficiently decomposed by direct UV

313

photolysis at 254 nm and RNS. DCF and TCS in this group have electron-rich moieties such as

314

phenolic and aniline groups, which lead to the moderate second-order reaction rate constants (107

315

- 108 M−1 s−1) with CO3•−. However, the removal percentages ascribed to CO3•− oxidation were

316

negligible compared with the significant contributions from RNS and UV photolysis.

317

Group III CECs with weakly electron-donating moieties could not be efficiently removed

318

by UV/NO3− and the degradation rates declined in the presence of HCO3−. CBZ and IBP in this

319

group were primarily degraded by •OH, while the dominant contribution to ATZ removal was UV

320

photolysis.

321

The effects of HCO3− on the degradation of three groups CECs were ascribed to the change

322

of •OH concentrations, RNS speciation, and the formation of CO3•−. On one hand, the •OH

323

concentrations were reduced 10-fold with the continuous generation of CO3•−. The lowered •OH

324

lever affected all kinds of CECs. Since CO3•− reacts more selectively towards phenolic and aniline

325

moieties22,23, the elevated CO3•− concentration contributed to the degradation of the Group I and

326

Group II CECs. On the other hand, the concentrations of ONOO−, •NO2, •NO, and O2•− would be

327

subsequently increased due to the scavenging of •OH. RNS have the inclination to react with CECs

328

with electron-rich moieties at higher reaction rate24-5, 14, 43, significantly affecting the removal of


Page 16 of 40

Page 17 of 40


329

Group I and II CECs. The selectivity properties of CO3•− and RNS lead to the inefficient removal

330

of the total organic carbon (TOC) during the degradation of selected CECs.

331 332

Degradation Mechanisms of Diclofenac, Triclosan, Estrone, and Bisphenol A in UV/NO3−/HCO3−.

333

Degradation products detected for TCS, DCF, BPA, and E1 during treatments are

334

summarized in Table S7-S10. The chemical structures of the primary transformation products of

335

TCS, DCF, BPA, and E1 were further supported by MS and MS2 fragmentation using Q-TOF-

336

LC/MS/MS (Figure S6-S12). Isomeric structures for certain products were formed due to the

337

complex oxidation processes, which were observed at different retention time and denoted with a,

338

b, and c. The UV fluence-dependent evolution of transformation products during the degradation

339

of TCS, DCF, and BPA in UV/NO3−/HCO3− can be found in Table S11 and Figure S13-16.

340

Dechlorination-hydrogenation products were observed during TCS degradation in the

341

UV/NO3−/HCO3− process. The transformation product with relatively largest volume was one

342

isomer of dechlorination-hydrogenation products, T253c, (Figure S13), which also was detected at

343

a relatively high volume in the UV-alone system at pH 8.8. Thus, the formation of T253a-c and T264

344

can be attributable to UV photolysis at short wavelength45. Dechlorination-hydroxylation that led

345

to the generation of T235a-c could be initiated by •OH through ipso-attack at the carbon attached to

346

the chlorine (Figure S17a), which has been well-established in the reaction of •OH with

347

halobenzenes46, trimethoprim40, and ATZ47,

348

through the •OH addition, followed by a heterolytic cleavage of the C-Cl bond. Ether-bond

349

breakage was also observed in the process at relatively high UV fluence (> 960 mJ cm−2), leading

350

to T127a, T127b, T161, and T143. Electron transfer through the oxygen atom to •OH/CO3•−/•NO2 led to

351

the formation of a carbon centered radical, undergoing further reaction to break the C1−O bond of

48.

A carbon centered radical was firstly formed



352

ether group, which is the weakest site based on the calculation by Gaussian 09 program at

353

B3LYP/6-311+G* level49. The ether-bond can also be broken by •OH attacking on C1 atom through

354

•OH addition-elimination considering the lowest reaction energy barrier of •OH-addition on TCS50.

355

In addition, 2,4-diclophenol (T161) has been detected and identified when TCS was decomposed

356

under irradiation at 316 nm45, confirming the role of UV photolysis to ether-bond cleavage.

357

Interestingly, NO2-adducts, such as T264 and T280, were detected during TCS decomposition

358

which could be ascribed to the contributions of RNS. This route has been proven by the formation

359

of 3-nitrotyrosine through the reaction of tyrosine with •NO2 at a high reaction rate constant (3 ×

360

109 M−1 s−1) (Figure S17b)2. An oxygen-centered phenolic radical was firstly generated via

361

transferring an electron to •OH, CO3•−, or •NO2, followed by the addition of •NO2 radical.

362

Additionally, quinone derivatives, T249, T283, and T317, were observed during the TCS degradation.

363

T303, p-hydroquinone of triclosan, is a common transformation product in •OH-based oxidation

364

process of triclosan, due to the •OH attack at para-position49, 51. Further oxidation of T303 by •OH

365

would lead to the generation of T301, p-quinone of triclosan49, 51. T303 and T301 were not observed

366

during the degradation of TCS in UV/NO3−/HCO3−, which could be ascribed to subsequent

367

formation of T249, T283, and T317 through the dechlorination-hydrogenation, dechlorination-

368

hydroxylation, and hydroxylation routes. Dioxin derivative, T267, was generated through the UV

369

photolysis and •OH oxidation, which raised environmental concern about the toxicity of treated

370

solutions.

371

Cyclization product D260 was detected and identified during DCF decomposition in

372

UV/NO3−/HCO3−. This route was initiated by •OH-induced H-abstraction at C8 as illustrated in

373

Figure S17c45, 52. The formed radical anion subsequently removed a chloride anion at C6 via the

374

dechlorination route to form a biradical under UV irradiation, which leads to a quick recombination


Page 18 of 40

Page 19 of 40


375

with the generation of a five-membered cyclic product. Decarboxylation that led to the generation

376

of D216, could be initiated with •OH-involved electron transfer on the carboxylic acid of the D260,

377

which firstly formed a carboxylic radical, followed by removal of CO2 and electron53. No

378

accumulation of D216 was observed possibly due to the further oxidation by •OH or CO3•−,

379

producing a quinine-like product, D214 19, 22, 53-55. Details of decarboxylation are shown in Figure

380

S17d.

381

Hydroxylation was also observed during the BPA degradation with the formation of

382

hydroxylated BPA, B243. Hydroxylation could be initiated with •OH addition as shown in Figure

383

S17e, forming a carbon centered radical, sequentially subjected to oxygen addition, and the

384

removal of perhydroxyl radicals (HOO•)40,

385

undergo CO3•−-initiated addition-elimination route21, the steric inhibition limits the CO3•− addition

386

to the carbon centered radical that formed in the first step22. Considering the activation energy and

387

Gibbs free energy, ortho-hydroxylation was preferred for BPA compared to para-hydroxylation

388

and meta-hydroxylation58. B243 could be further oxidized by •OH to generate the corresponding

389

quinone derivative, B241. B287 is a quinone-like transformation product that has not been reported,

390

which might be generated via further hydroxylation on formed quinone derivatives59. Bond

391

breakage also occurred adjacent to the methyl bridge due to the •OH oxidation with the formation

392

of phenol60, which could react rapidly with •NO2 to produce the product B13857. Interestingly, the

393

concentration of B138 continuously increased with the UV photolysis on NO3− in the presence of

394

HCO3− (Figure S15), even after the completely removal of parent compound, indicating B138 as

395

one of the final oxidation products under current reaction conditions.

56, 57.

Although the hydroxylation was proposed to

396

Hydroxylation played a crucial role during the decomposition of E1 with the formation of

397

transformation product E285a-b, due to the electrophilic attack on the carbon in aromatic ring by



398

•OH61.

A quinone product E283 was formed through sequential oxidation of E285 by •OH. E299 was

399

produced as the primary transformation products, with highest amount via the subsequent

400

hydroxylation of E283. Noticeably, NO2-adducts, E314a-b, were also generated during E1

401

degradation in the UV/NO3−/HCO3− process, which was attributable to the contributions of •NO2.

402

Cytotoxicity Studies.

403

The nitro-aromatics that formed during NO3− photolysis might pose a risk on human health

404

and the ecological environment11, 62. Therefore, it is critical to evaluate the cytotoxicity of the

405

resulting treated solution even though Group I and II CECs could be effectively removed by

406

UV/NO3−/HCO3−. In general, the complex reactions of direct photolysis, •OH, CO3•−, and RNS

407

result in complex mixtures of transformation products at low concentration which are difficult to

408

be isolated. Therefore, cytotoxicity of the treated solutions in the UV/NO3−/HCO3− system was

409

evaluated.

410

As shown in Figure 5a, at the beginning stages of DCF degradation, the cytotoxicity

411

declined with decreased DCF (0-160 mJ cm−2 UV fluence). However, the cytotoxicity significantly

412

increased (p = 0.001) after the complete degradation of DCF (> 160 mJ cm−2 UV fluence). The

413

elevated cytotoxicity might be attributed to the accumulation of D260 rather than D214 that was

414

subsequently removed along with the depletion of DCF (Figure S14).

415

Dioxin derivatives observed during the TCS decomposition in the UV/NO3−/HCO3−

416

process, such as T267, possibly have a risk on the ecological and human health11, 50, however, the

417

resulting cytotoxicity analyzed with AhR cells and MTT assays did not significantly change (p =

418

0.264) (Figure 5b). This phenomenon could be ascribed to the rapid degradation of the

419

transformation products (T253, T264, and T314), the low concentrations of the formed transformation

420

products, and the incomplete UV photochemical oxidation of transformation products (T267, and


Page 20 of 40

Page 21 of 40


421

T143) within the UV fluence of 640 mJ cm−2 as shown in Figure S13. Similarly, under current

422

reaction conditions (10 mM NO3− and 3 mM HCO3−), no obvious cytotoxicity change was

423

observed (p = 0.426) during the removal of E1 at the UV fluence of 640 mJ cm−2 as shown in

424

Figure 5c.

425

Remarkably, the cytotoxicity of treated BPA decreased (p = 0.067) along with the

426

degradation of BPA (Figure 5d). The generated B243 and B241 were simultaneously removed with

427

BPA, while the formation of nitrophenol (B138) continued to increase in the UV/NO3−/HCO3−

428

system (Figure S15). This result indicates that nitrophenol at low concentration has little effect on

429

the cytotoxicity.

430

Environmental Implications.

431

To evaluate the CEC degradation when NO3− and HCO3− exposed to UV irradiation, RO

432

permeate (UV influent) from GWRS was used as a reaction matrix with spiked additions of

433

chemicals (CECs, NO3−, and HCO3−) to compare the UV/NO3−/HCO3− process with UV only and

434

UV/NO3− methods. Due to the presence of •OH quenching agents such as NO3−, HCO3−, and

435

chloramines63 in the RO permeate (Table S1), the removal of certain CECs with UV/H2O2 was

436

similar with UV only (Figure S18). As shown in Figure 6, the efficiency of three different AOPs

437

was in the order of: UV/NO3−/HCO3− > UV/NO3− > UV only for DCF, TCS, and BPA; and

438

UV/NO3− > UV/NO3−/HCO3− > UV only for ATZ, CBZ, and IBP. These results suggest that higher

439

removal of Group I and Group II CECs by UV/NO3−/HCO3− in RO permeate might be ascribed to

440

the contributions of RNS and CO3•−.

441

This research demonstrates the important role of HCO3− in UV/NO3− treatment,

442

accelerating the removal of Group I and II CECs with electron-rich moieties such as phenolic and

443

aniline groups but inhibiting the destruction of Group III CECs with weakly electron-donating



444

moieties. The combined effects of UV photolysis, RNS, •OH, and CO3•− contributed to the

445

degradation of CECs in UV/NO3−/HCO3−. Among the selected CECs, the cytotoxicity was reduced

446

during the BPA degradation in the UV/NO3−/HCO3− treatment. Moreover, the selectivity of RNS

447

and CO3•− made them less affected by NOM and other constituents such as free chlorine and

448

chloramine compared to UV photolysis and UV/H2O2 technologies. Therefore, the residual NO3−

449

and HCO3− in the wastewater have the potential be utilized under UV irradiation for CECs removal

450

in carbonate-rich water reuse scenarios.

451

452

Associated Content

453

Supporting Information

454 455

The Supporting Information is available free of charge on the ACS publications website at http://pubs.acs.org/.

456

Information on chemical used; CECs and transformation products analysis; methodology

457

to determine steady-state concentration of •OH and CO3•−; methodology to calculation

458

contributions on CEC degradation; cytotoxicity analysis methods; water parameters of RO

459

permeate; CO3•− reactions rate constants; LC/MS conditions for transformation products analysis;

460

MS and MS2 spectra for transformation products; UV fluence-based evolution for transformation

461

products during CEC degradation; degradation of NB, 4-CP, and selected CECs (BPA, E1, EE2,

462

TCS, DCF, ATZ, CBZ, and IBP); and additional references. (Text S1-S6, Table S1-S11, and

463

Figure S1-S18).

464

Author Information

465

Corresponding Author


Page 22 of 40

Page 23 of 40


466

*Phone: +001- (513) 556-0724; email: [email protected].

467

Current Address

468

1Department

469

Ohio 45221, USA

470

Notes

471

The authors declare no conflicts of financial interest.

472

Acknowledgements

of Chemical and Environmental Engineering, University of Cincinnati, Cincinnati,

473

The authors acknowledge financial support from the U.S. Geological Survey (USGS)-

474

Water Resources Research Institute (WRRI) (2015SC101G) for this research. Ying Huang

475

acknowledges support from the China Scholarship Council (CSC) scholarship (201306270057).

476

Minghao Kong acknowledges support from the CSC scholarship (201608110134). We are

477

thankful to Orange County Water District for collecting and sending water samples used as a real-

478

world reaction matrix.

479



480

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677


Table 1. Second-order rate constants for the reaction of CO3•− with the investigated CECs.

pK CEC

Structure a

kCO•3 ― ,CEC at

kCO•3 ― ,CEC

pH = 8.8

(M−1 s−1)

(M−1 s−1)

in ref

k•OH,CEC (M−1 s−1)

Group I: CECs with phenolic moieties O

10. Estrone

(5.3 ± 1.1) 

2.6 × 1010

108

64

H

7

H

H

HO

(2.8 ± 0.2)  Bisphenol A

1.0 × 1010 0.9) ×

9.6 HO

17α-

(3.89 ±

OH

OH

10.

108

106 65

27

(1.6 ± 0.3) 

(9.8 ± 1.2)

108

× 109 66

H

Ethynylestradiol

7

H

H

HO

Group II: photo-degradable CECs OH

Cl

Triclosan

(4.2 ± 1.4) 

O

7.9

4.43 × 109 Cl

Cl

COOH Cl

Diclofenac

H N

4.2

107 (2.7 ± 0.7)  8.67 × 109 107

Cl

Group III: CECs with weakly electron-donating moieties Cl

Atrazine

1.6

N

N H

(5.7 ± 0.1) 

(3.7-4) ×

106

106 29, 30

2.4 × 109 66

N

N

N H



13. Carbamazepine

(4.2 ± 0.01) 

(2.3-2.5)

(8.8 ± 1.2)

106

× 106 25, 28

× 109 66

(1.2 ± 1.1) 

7.89 

(7.4 ± 1.2)

106

105 25

× 109 66

N

9

Ibuprofen

Page 34 of 40

O

NH2

OH

4.9 O

678

679


Page 35 of 40

680


Figure Legends

681

ONOO h

CO2

NO3

h

h O2

+ NO2

NO

h +O

2

h 682 683

NO2

or

h +C O

3

/O

H

+ NO + OH

OH

NO2

+HCO3

H

/O H

+O H

+O

ONOO

CO3

h

Figure 1. Primary reactions during nitrate photolysis at 254 nm in alkalinity solutions.

684 685



UV

30

UV/HCO3UV/NO3UV/NO3-/HCO3-

25 kobs (x 10-2 cm2 mJ-1)

Page 36 of 40

20 15 10 5 0

BPA

E1

EE2

TCS DCF ATZ

CBZ

IBP

686 687

Figure 2. The UV fluence-based pseudo-first-order rate constant (kobs) of CECs in the

688

UV/NO3−/HCO3− processes. [CEC]0 = 1 µM, [NO3−]0 = 10 mM, [HCO3−]0 = 3 mM, pH = 8.8.

689 690


Page 37 of 40


100

UV •OH CO3•− RNS

Removal (%)

80

60

40

20

0 BPA

E1

EE2

TCS

DCF

ATZ

CBZ

IBP

691 692

Figure 3. Removal percentage of CECs by •OH, CO3•−, RNS, and UV during UV/NO3− treatment

693

with (left bar without pattern) and without (right bar with pattern) the addition of HCO3− after the

694

172 mJ cm−2 UV fluence.

695 696



Cl

O

OH

Cl

O

OH

UV

Cl 3

15

Cl

7 8

O

4

5

OH

OH CO3

HO

OH

OH

OH

NO2

OH Cl

Cl

Cl

UV

O HO

Cl

T317, m/z 316.9180 O

OH O

Cl

Cl

Cl O

O

T143, m/z 142.9905

T127a-b, m/z 126.9964

OH

OH

a n UV d

O OH

OH

Cl O

T301

OH

OH

Cl

Cl

Cl O

T303

Cl O

OH

O

OH

UV

O

Cl

OH

Cl

CO3

T161, m/z 160.9597

T280, m/z 280.0005 O

Cl

NO2

Cl

T235,m/z 235.0190

O

Cl

O

NO2

Cl

Cl

T253a-c, m/z 252.9850

OH

OH

O

OH

Cl 17

11

m/z 286.943

OH

T264, m/z 264.0095

NO2

OH Cl

10

6 12

UV

O

UV

9

1

T267, m/z 266.9645

OH

OH

13

2

NO2

Cl

O

Triclosan 14 16

Page 38 of 40

T249, m/z 248.9960

T283, m/z 282.9589

Diclofenac 14

15 6

1

H N

5 4

2

7

12

16

Cl

OH

11

8

Cl

3

COOH

13

17

Cl

H N

COOH

OH

Cl

Cl

H N

N

OH

10

CO3

9

m/z 296.0242

D260, m/z 260.0475

D214, m/z 214.0416

D216

Bisphenol A OH HO

HO

OH

m/z 227.1103

HO

697 698

OH

Estrone

OH

HO

E314a-b, m/z 314.1398

NO2

CO3

O

OH

B138, m/z 138.0211

B287, m/z 287.0558

O

OH

HO

m/z 227.1103

O O

B241, m/z 241.0895

HO

NO2

OH

O

O

OH HO

O O

O

NO2 CO3

HO

OH

B243, m/z 243.1053

O O 2N

OH

O

O

OH

O

E285a-b, m/z 285.1495

E283, m/z 283.1338

O

OH

O

E299, m/z 299.1288

Figure 4. Possible degradation pathways of TCS, DCF, BPA, and E1 in UV/NO3−/HCO3−.

699


Page 39 of 40


(a) 1.4

0.6

0.2

0.4 0.1

0.0

Normalized absorbance


0.3

0.8

100

200

300

400

500

600

0.3

0.8 0.6

0.2

0.4 0.1

0.2 0.0

0.0 0

0.4

1.0

0.0 0

700

100

1.2

0.3

0.8 0.6

0.2

0.4 0.1

0.2 0.0

0.0 200

300

400

500

400

500

600

700

600

700

0.5

Bisphenol A

0.4

1.0 0.3

0.8 0.6

0.2

0.4 0.1

0.2 0.0

Concentration (mg L -1)

1.0

100

300

1.2 Concentration (mg L -1)

0.4

0

(d) 1.4

0.5

Estrone


(c) 1.4

200

UV fluence (mJ cm-2 )


700

Concentration (mg L-1)-1)

1.0

Concentration (mg L -1)

0.4

0.2


0.5

Triclosan

1.2

1.2

701

(b) 1.4

0.5

Diclofenac

0.0 0

100


200

300

400

500

600

700


702

Figure 5. Cytotoxicity of DCF (a), TCS (b), E1 (c) and BPA (d) treated with UV/NO3−/HCO3−.

703

Left axis represents cytotoxicity by bars; right axis represents concentration by lines. The higher

704

the bar, the lower the toxicity. [CEC]0 = 1 µM, [NO3−]0 = 10 mM, [HCO3−]0 = 3 mM, pH = 8.8.

705



UV

30 k obs (x 10 -3 cm2 mJ -1 )

Page 40 of 40

UV/NO3 UV/NO3 -/HCO3 -

25 20 15 10 5 0 DCF

TCS

BPA

IBP

ATZ

CBZ

706 707

Figure 6. Degradation of spiked CECs by UV/NO3−/HCO3− in the RO permeate from GWRS.

708

[CEC]0 = 1 µM, [NO3−]0 = 10 mM, [HCO3−]0 = 1 mM, pH = 8.0 after the addition of 1 mM HCO3−.

709 710


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