Accepted Manuscript

The importance of reactive oxygen species on the aqueous phototransformation of sulfonamide antibiotics: Kinetics, pathways, and comparisons with direct photolysis

Linke Ge, Peng Zhang, Crispin Halsall, Yanying Li, Chang-Er Chen, Jun Li, Helin Sun, Ziwei Yao

PII: S0043-1354(18)30930-8 DOI: https://doi.org/10.1016/j.watres.2018.11.009 Reference: WR 14216

To appear in: Water Research

Received Date: 4 July 2018 Revised Date: 2 October 2018 Accepted Date: 4 November 2018

Please cite this article as: Ge, L., Zhang, P., Halsall, C., Li, Y., Chen, C.-E., Li, J., Sun, H., Yao, Z., The importance of reactive oxygen species on the aqueous phototransformation of sulfonamide antibiotics: Kinetics, pathways, and comparisons with direct photolysis, Water Research, https://doi.org/10.1016/ j.watres.2018.11.009.

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+ 0 - H2SAs HSAs SAs m aj direct photolysis SA i + •OH ·OH o H2SAs+H SAs+ HSAs00 SAs-- r 16 2 HSAs SAs major

O

) 1

- 12 s

1 1 H

- M

9 H N S N 8 2 ·OH

10 R

×

( i 4 O r •OH, o k n HO 1 mi 0 O2 250450 SA i + 1O H SAs+ HSAs0 SAs- 2 H2SAs+2 HSAs0 SAs- direct photolysis

200250

) 1

- 1 s

1 1 major

- 150 O2 M 7 O m 10 100 N i × n ( H i or 1O2, 50 H N S N R k 2 0 O ·OH ·OH jor OH HO ma

Photooxidation of dissociated sulfonamides (SAs)

MANUSCRIPT

ACCEPTED ACCEPTED MANUSCRIPT

1 The importance of reactive oxygen species on the aqueous

2 phototransformation of sulfonamide antibiotics: Kinetics, pathways, and

3 comparisons with direct photolysis

4 Linke Ge a,b , Peng Zhang a,b , Crispin Halsall b,*, Yanying Li b, Chang-Er Chen c, Jun Li a, Helin

5 Sun a, Ziwei Yao a

6 a Key Laboratory for Ecological Environment in Coastal Areas (SOA), National Marine

7 Environmental Monitoring Center, Dalian 116023, P. R. China

8 b Lancaster Environment Centre, Lancaster University, Lancaster LA1 4YQ, United Kingdom

9 c Environmental Research Institute, MOE Key Laboratory of Environmental Theoretical

10 Chemistry, South China Normal University, Guangzhou 510006, P. R. China

11 12 ABSTRACT

13 Sulfonamide antibiotics (SAs) are increasingly detected as aquatic contaminants and exist as 14 different dissociated species depending on the pH MANUSCRIPTof the water. Their removal in sunlit surface 15 waters is governed by photochemical transformation. Here we report a detailed examination

1 16 of the hydroxyl radical (•OH) and singlet oxygen ( O2) mediated photooxidation of nine SAs:

17 sulfamethoxazole, sulfisoxazole, sulfamethizole, sulfathiazole, sulfamethazine, sulfamerazine,

1 18 sulfadiazine, sulfachloropyridazine and sulfadimethoxine. Both •OH and O2 oxidation

+ 19 kinetics varied depending on the dominant protonated states of the SA in question (H 2SAs ,

20 HSAs 0 and SAs -) as a function of pH. Based on competition kinetic experiments and matrix

0 - 21 deconvolution calculations,ACCEPTED HSAs or SAs (pH ~5–8) were observed to be more highly - 1 22 reactive towards •OH, while SAs (pH ~8) react the fastest with O2 for most of the SAs

23 tested. Using the empirically derived rates of reaction for the speciated forms at different pHs,

* Corresponding author. Tel./fax: +44 (0)1524 594330.

E-mail address: [email protected] (C. Halsall).

1 ACCEPTED MANUSCRIPT 1 24 the environmental half-lives were determined using typical O2 and •OH concentrations

1 25 observed in the environment. This approach suggests that photochemical O2 oxidation

26 contributes more than •OH oxidation and direct photolysis to the overall phototransformation

27 of SAs in sunlit waters. Based on the identification of key photointermediates using tandem

1 28 mass spectrometry, O2 oxidation generally occurred at the amino moiety on the molecule,

29 whereas •OH reaction experienced multi-site hydroxylation. Both these reactions preserve the

30 basic parent structure of the compounds and raise concerns that the routes of

31 phototransformation give rise to intermediates with similar antimicrobial potency as the

32 parent SAs. We therefore recommend that these phototransformation pathways are included

33 in risk assessments concerning the presence and fate of SAs in waste and surface waters.

34 Keywords: Sulfonamides; dissociated forms; photodegradation; oxidation kinetics;

35 transformation products

36 37 1. Introduction MANUSCRIPT 38 The occurrence, transformation and risks of antibiotics in aquatic systems are

39 well-acknowledged environmental issues and have raised particular concerns

40 (Baena-Nogueras et al., 2017; Luo et al., 2011). Sulfonamides (SAs) are a large class of

41 antibiotics that are widely used in human treatments, animal husbandry and aquaculture

42 (Managaki et al., 2007). Due to their incomplete elimination by metabolism (Campagnolo et

43 al., 2002) and wastewater treatment (Chen et al., 2013), SAs enter the environment 44 continually and ACCEPTEDare regularly detected as aqueous micropollutants. Their presence in surface 45 waters has been reported in many countries including the United States (Kolpin et al., 2002),

46 Europe (Loos et al., 2009), Vietnam (Managaki et al., 2007) and China (Luo et al., 2011).

47 Concentrations in surface waters can vary quite widely, with mean concentrations observed to

48 range from 0.1 to 150.8 ng L -1 in the coastal waters of China through a variety of recent

2 ACCEPTED MANUSCRIPT 49 monitoring studies (see Fig. S1 and associated references). There is evidence of

50 ecotoxicological effects of SAs on freshwater biota, with SAs also promoting bacterial

51 resistance (Białk-Bieli ńska et al., 2017; Białk-Bieli ńska et al., 2011; Zhang et al., 2015).

52 Interestingly, exposure to sunlight appears to enhance the toxicities of SAs towards several

53 aquatic organisms, such as D. magna (Jung et al., 2008; Trovó et al., 2009b), and Vibrio

54 fischeri (Xu et al., 2014) and this reflects the toxicity exhibited by other photo-transformed

55 antibiotics, such as fluoroquinolones (Ge et al., 2010; Ge et al., 2018).

56 SAs may undergo different transformation processes including photodegradation and

57 biodegradation. However, in sunlit surface waters, photodegradation has been identified as the

58 major transformation pathway (Carstens et al., 2013). Specifically, SAs can undergo direct

59 photodegradation as well as sensitized indirect photoprocesses such as oxidation reactions

1 60 with hydroxyl radicals (•OH), singlet oxygen ( O2), and other reactive oxygen species (ROS)

1 61 (Boreen et al., 2004, 2005; Li et al., 2018). In sunlit surface waters, •OH and O2 are 62 ubiquitous in waters and serve as the most importan MANUSCRIPTt transient photooxidants leading to the 63 oxidative degradation of aqueous organic micropollutants (Cooper et al., 1989; Vione et al.,

64 2014; Zhang et al., 2014). They are generated from sunlight-induced reactions involving

65 common water constituents, such as dissolved organic matter (DOM), nitrate ions, and Fe(III).

66 Furthermore, the environmental photochemical processes of SAs can also be affected by the

- 2+ 67 presence of other water constituents such as halide ions, HCO 3 and Cu ions (Guerard et al.,

68 2009; Li et al., 2018; Li et al., 2016; Niu et al., 2013a). As the molecular structures of SAs 69 contain two ionizableACCEPTED groups (i.e., −NH 2 and −NH−), their various dissociated forms are 70 expected to influence their photochemical behavior (Boreen et al., 2004; Wei et al., 2013),

71 although the behavior of the various dissociated forms is not well understood.

72 Previous studies have demonstrated separate phototransformation kinetics of different

73 dissociated species for specific compounds in various classes of antibiotics, such as

3 ACCEPTED MANUSCRIPT 74 tetracyclines (Niu et al., 2013b; Werner et al., 2006), fluoroquinolones (Ge et al., 2015; Wei et

75 al., 2013), and SAs (Boreen et al., 2004, 2005; Willach et al., 2018), although for SAs direct

76 photodegradation rather than ROS photooxidation has been the focus to date. For direct

77 photodegradation, the anionic, neutral and cationic species of SAs have different quantum

78 yields (Φ) and rate constants ( k), and the most photoreactive component varies depending on

79 the SA under investigation (Bonvin et al., 2013; Boreen et al., 2004, 2005; Niu et al., 2013a;

80 Xu et al., 2014). Without considering the protonation state, SAs react with •OH at a mean rate

9 -1 -1 1 4 8 81 (k) of 5.3×10 M s , and their O2 oxidation k ranges over 4 orders of magnitude (10 –10 )

82 (Boreen et al., 2004, 2005). The different protonation states are expected to react with ROS at

83 different reactivities. However, the ROS photooxidation kinetics of the different dissociation

84 species has received little attention in the literature. Furthermore, to apportion the relative

85 contributions of the photooxidation to the fate of SAs, the corresponding environmental

86 half-lives in surface waters need to be assessed based on the distribution of the speciated 87 forms and the ambient water conditions. MANUSCRIPT 88 A thorough understanding of the transformation routes via ROS photooxidation of SAs is

89 important for the overall assessment of their aquatic fate and to understand the relative

90 importance of this phototransfromation pathway compared to direct photodegradation. To date,

91 only several intermediates have been identified for selected SAs for the oxidation reaction

92 with •OH in advanced oxidation processes (Batista et al., 2014; Du et al., 2018; Hu et al.,

93 2007; Trovó et al., 2009a). The •OH can induce multihydroxylation of the aniline moiety, and 94 cleavage of heterocyclicACCEPTED rings (Batista et al., 2014; Hu et al., 2007; Shah and Hao, 2017). As 1 95 for O2 oxidation, some relevant mechanisms have been proposed. For example, a [2+2]

1 96 cycloaddition of O2 on isoxazole rings was speculated for SAs containing five-membered

1 97 heterocyclic groups (Boreen et al., 2004). Furthermore, O2 tends to oxidate electron-rich

98 compounds, such as anilines, dialkyl sulfides, pyrroles, and phenolate anions (Mill, 1999).

4 ACCEPTED MANUSCRIPT 99 These previous studies provide a useful rationale to further investigate ROS photooxidation of

100 SAs in an environmental context.

1 101 The purpose of this study was to investigate the ROS (i.e., •OH and O2) oxidation kinetics

102 of SAs and compare the oxidation products and kinetics with those formed from direct

103 photodegradation. The nine SAs selected in the study are widely used and commonly reported

104 in aquatic environments (Chen et al., 2013; Kolpin et al., 2002; Luo et al., 2011; Managaki et

105 al., 2007). Importantly, the oxidation kinetics and reactivities were examined for the various

106 dissociated forms which, in turn, are dependent on the ambient pH. ROS-induced

107 transformation pathways are currently under-represented in the literature and yet may account

108 for a signficant proportion of phototransformation of SAs in surface waters. This has

109 important implications for how we assess their fate and undertake environmental risk

110 assessments for this group of chemicals.

111 112 2. Materials and methods MANUSCRIPT 113 2.1. Reagents and materials

114 Four SAs containing the five-membered heterocyclic groups and five SAs containing the

115 six-membered heterocyclic groups were selected as model compounds for the study (Fig. 1),

116 including sulfamethoxazole (SMX), sulfisoxazole (SIX), sulfamethizole (SMT), sulfathiazole

117 (ST), sulfamethazine (SMZ), sulfamerazine (SM), sulfadiazine (SD), sulfachloropyridazine

118 (SCP), and sulfadimethoxine (SDM). As described in Table S1, the nine SAs were purchased 119 from different suppliers.ACCEPTED All the compounds, as received, were at least 98% purity. Furfuryl 120 alcohol (FFA, 98%) and perinaphthenone (97%) were obtained from J&K Technology Co.,

121 Ltd. and Sigma-Aldrich, respectively. Hydrogen peroxide (H 2O2, 30%) was purchased from

122 Acros Organics. Methanol, trifluoroacetic acid and acetophenone (AP) were of HPLC grade,

123 and other reagents such as phosphates were of guaranteed grade and used as received. All

5 ACCEPTED MANUSCRIPT 124 solutions were prepared in ultrapure water obtained from a Millipore-Milli Q system.

125

126 Fig. 1. Chemical structures and the different protonation states of sulfonamide antibiotics (SAs), where

+ 0 - 127 H2SAs , HSAs , and SAs represent the cationic form, the neutral form, and the anionic form, respectively.

128

129 2.2. Photochemical experiments

130 A merry-go-round reactor with quartz tubes (50 mL volume and 20 mm diameter)

131 containing reaction solutions was used to perform photochemical experiments under

132 irradiation of a water-refrigerated 500W high-pressure mercury lamp. The light intensities at

133 the solutions were 1.95 mW cm -2 at 365 nm and 1.83 mW cm -2 at 420 nm, measured by a UV

134 spectral irradiometer (Photoelectric Instrument Factory of Beijing Normal University). The

135 experimental setup and spectral irradiance are illustrated in Figs. S2 and S3, respectively. 136 Competition kinetics were employed to determineMANUSCRIPT the bimolecular rate constants kROS (i.e., 137 k•OH and k1O2 ) for the reactions between SAs and ·OH according to Eq. (1):

ln([S ]/[S ]) kS= t 0 k R ROSln([R ]/[R ]) ROS 138 t 0 (1)

139 where S is the substrate SAs, R is a reference compound, and the square brackets [] represent

1 140 the concentrations. The method has been used successfully to examine the •OH/ O2

141 reactivifties of pharmaceuticals and other environmental organic contaminants (Boreen et al.,

142 2004; Edhlund et al., 2006; Wei et al., 2015; Xie et al., 2013). 143 The initial concentrationACCEPTED of individual SAs was set at 10 µM. Hydroxyl radicals (•OH) 144 were generated through the addition of H2O2 (100 mM) with a 340 nm cut-off filter

145 surrounding the Hg-lamp light source. Acetophenone (10 µM) was spiked into the reaction

9 -1 -1 146 solution as the reference compound with k•OH = 5.9 × 10 M s (Edhlund et al., 2006). For

1 147 the O2 reaction experiments 380 nm cut-off filters were used for the light source, with

6 ACCEPTED MANUSCRIPT 148 perinaphthenone (20 µM) and furfuryl alcohol (20 µM) as the photosensitizer and reference

8 -1 -1 149 compound (k1O2 = 1.2 × 10 M s ), respectively. Perinaphthenone is often used to

1 150 photogenerate O2 with a quantum yield close to unity (Boreen et al., 2008; Han et al., 2008;

1 151 Latch et al., 2003; Oliveros et al., 1991; Schmidt et al., 1994). The O2 reaction solutions were

152 pre-aerated to be saturated with dissolved oxygen. The pH of the reaction solutions (pH = 2, 5

153 and 8) was adjusted to the desired value using phosphate buffers. All the competition kinetic

154 experiments with controls were conducted in triplicate.

155 2.3. Matrix deconvolution calculation

156 The ROS oxidation rate constants (kROS, i) for SA protonation states (i) were calculated

157 using matrix deconvolution methods, based on previous studies (Boreen et al., 2004, 2005;

158 Werner et al., 2006) and used successfully by our group in a previous study on ionisable

159 fluoroquinolones (Ge et al., 2018). For each SA, the individual k•OH,i or k1O2 ,i of the fully

+ 0 - 160 protonated (H 2SAs ), neutral (HSAs ), and anionic forms (SAs ) were fitted according to the 161 following Matrix (1): MANUSCRIPT pH 2 δ, δ , δ  k  pH 2 H SAs+0- HSAs SAs ROS,H SAs + k  2  2  ROS    =  M M  k 0  M    ROS,HSAs   M M     M    k -   pH 8 162 pH 8 ROS,SAs  (1)

163 where δi represents the fraction of each dissociated form and the values are referred to in Fig.

164 S4, kROS is k•OH or k1O2 of SAs at different pH, and kROS,H2SAs+ − kROS,SAs- are the k•OH, i or

165 k1O2 ,i of the three SA forms to be fitted. 166 2.4. Analytical DeterminationACCEPTED 167 A Waters UPLC with PDA detector, fitted with a BEH C18 column, was employed for the

168 separation and quantification of SAs, acetophenone and furfuryl alcohol in the irradiated

169 samples, which were collected from the reactor and spiked with methanol (•OH quencher) or

1 170 NaN 3 ( O2 quencher). To identify intermediates, irradiated solutions (2 L for each sample)

7 ACCEPTED MANUSCRIPT 171 was purified by solid phase extraction and then analyzed by an Agilent HPLC coupled to a

172 triple quadrupole mass spectrometer operating in both positive and negative modes. The

173 analytical details are provided in the Supplementary material.

174

175 3. Results and discussion

176 3.1. ROS oxidation kinetics under different pH

177 In the control experiments, the losses of SAs and reference compounds either under dark or

178 in the absence of the ROS photosensitizer were <2%. Thus, degradation by direct photolytic,

179 thermal and hydrolytic processes was considered negligible during the duration of the

180 photochemical experiments (2 hrs). However, when exposed to light irradiation in the

181 competition kinetic experiments, the SAs disappeared quickly, indicating that the SAs were

182 photooxidised efficiently by ROS. The oxidative degradation curve of SMZ as an example is

183 provided in Fig. S5. The values for the bimolecular reaction rate constants of the nine SAs

1 184 with •OH (k•OH ) and O2 (k1O2 ) in pure water are MANUSCRIPTlisted in Tables S2 and S3. The k•OH ranged 9 -1 -1 9 -1 -1 185 from (5.00 ± 0.77) × 10 M s for SM to (9.16 ± 1.66) × 10 M s for SMT, while the k1O2

186 ranged from (2.31 ± 0.06) × 10 7 M-1 s-1 for SMX to (1.34 ± 0.15) × 10 9 M-1 s-1 for SCP. The

187 k•OH values were larger than k1O2 , indictive of the higher reactivity of •OH with the various

1 188 SA compounds. However, considering that the steady-state concentrations of O2 can be

1 189 three-orders of magnitude higher than that of •OH (Cooper et al., 1989; Mill, 1999), O2 may

190 be more important than •OH in the attenuation of SAs in sunlit waters. 191 In comparisonACCEPTED with other antibiotics, the k•OH values of SAs fall in the same order of 192 magnitude as the fluoroquinolones (Ge et al., 2015; Santoke et al., 2009), tetracyclines (Jeong

193 et al., 2010), and nitrofurans (Edhlund et al., 2006). However, the k1O2 of SAs are 1 – 3 orders

194 of magnitude greater than those observed for the other antibiotics (Castillo et al., 2007;

195 Edhlund et al., 2006; Wei et al., 2015). This notable difference of k1O2 among antibiotics can

8 ACCEPTED MANUSCRIPT 1 196 be attributed to the higher selectivity of O2 than that of •OH to oxidize organic chemicals

197 (Keen and Linden, 2013; Larson and Weber, 1996; Mill, 1999; Yao et al., 2013). •OH can

198 oxidize almost all classes of organic chemicals (Keen and Linden, 2013; Li et al., 2014; Mill,

1 199 1999), whereas O2 is more sensitive to electron-rich chemicals, such as sulfides and phenols

200 (Larson and Weber, 1996; Mill, 1999).

201 Since aqueous SAs can undergo two acid-base dissociation processes (Fig. S4), the ROS

202 oxidation reactivity of each SA is expected to depend on its speciation at a given pH value. To

1 203 quantify the reactivity of each protonation state towards •OH and O2, the k•OH and k1O2 were

204 determined for three different pHs. The results are illustrated in Fig. 2, and the corresponding

205 values are provided in Tables S2 and S3. It is found that both k•OH and k1O2 are dependent on

206 pH (p < 0.05), with generally higher values of k•OH at pH 5 and pH 8 than at pH 2, although

207 SIX was the exception to this, displaying the fastest reactivity at pH 2. Similarly, the

1 208 reactivity of O2 oxidation was fastest for seven of the SAs at pH 8, while only two SAs, SIX 209 and SCP, had the fastest kinetics at pH 2 and pH 5,MANUSCRIPT respectively. The higher reactivity of SIX 210 at pH 2 suggests that the reactive site might be the dimethylisoxazole moiety towards •OH

1 211 and O2 at the lower pH. For SMX with its monomethylisoxazole moiety, dimethylation of

212 the isoxazole ring significantly increases the electron density of this ring (Dogan et al., 1990),

1 213 making SIX more susceptible to ROS attack, particularly to O2 oxidation (Fig. 2).

1 214 The effect of pH on the O2 oxidation kinetics was more significant than the reaction

215 with •OH. For example, the k1O2 values showed a 1.3- to 11.2-fold variation (mean = 4.5) 216 over the pH range,ACCEPTED while k•OH showed a 1.5- to 4.3-fold variation (mean = 2.2). The k•OH and 217 k1O2 values reported here compare favourably to an earlier study examining the photochemical

218 fate of SAs (Boreen et al., 2004, 2005). In this case, k1O2 was only measured at pH 10.2

219 (carbonate-buffered solution) and k•OH at pH 3 (Fenton system), with k1O2 for SMT reported as

7 -1 -1 9 -1 -1 220 (5.5 ± 0.4) × 10 M s and k•OH as (6.6 ± 0.2) × 10 M s (Boreen et al., 2004, 2005). For

9 ACCEPTED MANUSCRIPT 221 other chemicals, the pH effect resulting in different protonation states and their

222 corresponding ROS reactivities has been reported for fluoroquinolone antibiotics (Ge et al.,

223 2015; Wei et al., 2013) and hydroxylated polybrominated diphenyl ethers (Xie et al., 2013).

224

225 Fig. 2. ROS oxidation rate constants ( k•OH and k1O2 ) of 9 sulfonamide antibiotics (SAs) under different pH

226 conditions

227

228 3.2. ROS reactivities of different dissociated species

229 As shown in Fig. 3, ROS reactivities of SA dissociated species (i) are represented with the

230 calculated bimolecular reaction rate constants ( k•OH, i and k1O2, i). Their corresponding values

231 are listed in Tables S4 and S5. Each SA showed variable ROS reactivities for the different

+ - 232 dissociated forms, H2SAs to SAs . With the exception of SIX ，the other eight SAs in their

233 neutral (HSAs 0) or anionic (SAs -) forms were more highly reactive towards •OH whereas

1 234 most of the SAs (other than SIX and SCP) showedMANUSCRIPT the fastest reaction with O2 in their - 1 235 anionic form (SAs ). Compared with •OH oxidation, the O2 reaction activities varied by

236 several orders of magnitude across the various speciated forms of the different SAs (e.g.

+ - 237 H2SAs to SAs ). The proportion of each form is dependent on the pH (e.g. 2 to 8). Increasing

238 pH enhances the deprotonation and electron density of the reactive aniline moiety, which in

239 turn increases the electron donating ability and decreases the steric hindrance, thus making the

240 aniline more vulnerable toward ROS oxidant/radical attack (Dodd et al., 2006; Jiao et al.,

241 2008; Wei et al., 2013). For example, the aniline moiety in SMX makes this compound more

242 susceptible to electrophilicACCEPTED O 3 attack particularly following deprotonation experienced at

243 higher pHs (Dodd et al., 2006).

244 For emerging micropollutants including antibiotics, a database of kROS, i values was

245 constructed to describe the reactivities of their different protonated states towards •OH and

1 246 O2 (Table S6). This table includes the k•OH, i and k1O2, i for SAs from this study, as well as

10 ACCEPTED MANUSCRIPT 247 values for a selection of ionisable antibiotics, flame retardants and other relevant chemicals of

248 environmental concern. Similar to the SAs investigated here, anionic forms of all other

1 249 chemicals react faster with O2 than neutral or cationic forms. Tetracyclines in their neutral or

250 anionic forms were more highly reactive towards •OH. However, fluoroquinolones in their

251 neutral form showed higher reactivities of •OH than in their protonated/cationic and anionic

252 forms (Ge et al., 2018; Ge et al., 2015), while anionic HO-PBDEs were oxidised faster by

253 •OH than the neutral forms (Xie et al., 2013). These ROS reactive patterns, coupled to the

254 dominant speciated form of the chemical, are of importance for the fate assessment of the

1 255 ionisable pollutants that usually coexist with •OH and O2 in surface waters.

256

257 Fig. 3. Bimolecular rate constants ( k•OH,i and k1O2,i) for the reaction between dissociated SA species (SA i)

258 species and ROS

259 260 3.3. Photooxidation half-lives in surface waters MANUSCRIPT 1 261 In sunlit surface waters, the photoinduced •OH and O2 radicals are the most important

262 transient oxidants (Cooper et al., 1989; Vione et al., 2014; Zhang et al., 2014). Their

263 concentrations ([ROS]) range from 10 -17 to 10 -15 M and from 10 -15 to 10 -12 M, respectively

264 (Cooper et al., 1989; Mill, 1999). Based on ROS reactivities of dissociated SA species, the

265 corresponding environmental half-lives (t1/2,ROS,E , i.e., t1/2,•OH,E and t1/2,1O2,E ) at a specific pH

266 can be calculated according to Eqs. 2 – 3:

267 kROS,E = [ROS]∑(δikROS,i) (2) ACCEPTED = ln 2 t1/ 2,ROS,E 268 kROS,E (3)

1 269 where kROS,E is the environmental rate constant of SAs towards •OH or O2; kROS,i represents

270 k•OH, i or k1O2, i; and δi values (i.e. the fraction of each dissociated form) are taken from Fig. S4.

271 The calculation method takes into account the acid-base dissociation of SAs and the

11 ACCEPTED MANUSCRIPT 272 pH-dependent kinetics of ROS oxidation, and is also applicable for the calculation of the

273 t1/2,ROS,E of other dissolved ionizable organic pollutants. Previous studies have tended to assess

274 the t1/2,ROS,E without considering the pH difference between laboratory solutions and surface

275 waters (Boreen et al., 2004, 2005).

276 As shown in Table 1, the calculated values of t1/2,•OH,E and t1/2,1O2,E for the nine SAs range

277 from 14.62 h to 39.23 h and from 0.06 h to 2.22 h, respectively. To assess the importance of

278 the photooxidation on the fate of SAs, t1/2,ROS,E values were compared with direct photolytic

279 half-lives (t1/2,D,E ) of SAs in near surface waters (Boreen et al., 2004, 2005). For individual

280 SAs, t1/2,1O2,E are considerably shorter than t1/2,•OH,E and t1/2,D,E , indicating that oxidation of

1 281 SAs towards O2 can be a central factor in their environmental phototransformation. Even

282 though •OH is an unselective, highly reactive species, the environmental fate of SAs is likely

283 not controlled by the reaction with •OH, which can be attributed to such low concentrations of

284 •OH in surface waters. However, since the t1/2,•OH,E are comparable to t1/2,D,E in midsummer

285 and less than t1/2,D,E in midwinter, suggesting •OH MANUSCRIPT oxidation is also appreciable compared to 286 direct photolysis. Furthermore, the •OH is expected to be more effective at degrading SAs due

287 to its non-selectivity (Ge et al., 2015; Keen and Linden, 2013; Li et al., 2014; Mill, 1999).

288 The values in Table 1 represent the minimum expected half-lives, as these were calculated

289 based on the highest ROS levels and continual light irradiation, although the

290 phototransformation may be hindered by light scattering, or limited photic zone depth. Apart

291 from the photooxidation and direct photolysis, SAs may undergo solar photosensitized

3 292 degradation via ACCEPTEDDOM, which is of significance in most waters (Boreen et al., 2005; Li et al., 293 2015). In addition, concomitant metal ions can also influence the photochemical behavior of

294 SAs by complexation effects (Shah et al., 2015; Wei et al., 2015). Thus, to better understand

295 the environmental photochemical fate of SAs, further studies are needed to study specific

296 water bodies and examine the effects of non-continuous solar exposure, water depth and other

12 ACCEPTED MANUSCRIPT 297 aqueous constituents that may attenuate or absorb sunlight (Bodrato and Vione, 2014).

298

1 299 Table 1 Expected environmental half-lives (t1/2,E , h) for reactions with either •OH or O2, and direct

300 photodegradation of sulfonamides (SAs) in sunlit surface waters.

a 1 301 Calculated based on the •OH and O2 concentration levels in the euphotic zone of some surface waters, i.e.,

-15 1 -12 b 302 [•OH] = 1 × 10 M and [ O2] = 1 × 10 M (Cooper et al., 1989; Mill, 1999). Corresponding to direct

303 photolytic half-lives of SAs in near surface waters in midsummer and midwinter at 45° latitude (Boreen et

304 al., 2004, 2005). The half-lives were calculated with quantum yields (correcting for lens effect) and

305 noon-averaged tabulated solar intensities (assuming continuous irradiation) (Leifer, 1988).

306

307 3.4. ROS oxidation intermediates and pathways

308 Four SAs, comprising of SIX, ST, SMZ and SCP were selected to investigate their

309 phototransformation byproducts and deduce their main transformation pathways based on the

1 310 three different reactions including O2 and •OH MANUSCRIPToxidation, as well as direct photolysis. We 311 identified 21 significant intermediates generated from their photooxidation. All details on the

312 identification of these photoproducts are shown in Table S7, including chromatographic

313 retention times, molecular weight ( Mw), as well as mass spectral parent and fragment ions, m/z.

314 The corresponding MS and MS 2 mass spectra in positive ionisation mode are displayed in Fig.

315 S6. Referring to the Nitrogen rule and the analytical approaches in previous studies

316 (Baena-Nogueras et al., 2017; Sági et al., 2015), chemical structures were proposed and

317 tentative phototransformation pathways for the four SAs are presented in Fig. 4. 318 There are threeACCEPTED types of transformation pathways (Paths 1–3), corresponding to the three 1 319 different reactions of each SA (Fig. 4). Among the O2 oxidation reactions (Path 1),

320 complexes occur through –R group addition at the sulfonyl-N (N 1) site for SAs containing

321 five-membered heterocyclic substituents, whereas SAs with six-membered heterocyclic

322 groups experience an oxidation at the amino-N (N 11) site. As for transformation products with

13 ACCEPTED MANUSCRIPT 323 molecular weight of 352 (TP352a/TP352b), the –R groups involve one nitrogen atom. TP292

1 324 of SMZ generated from the addition of O2 to amino-N, and other moieties of SMZ were not

325 transformed due to the detection of main fragment ions with m/z = 186 and 124 (Fig. S6)

1 326 (Sági et al., 2015). A O2 oxidative product similar to TP292 was also reported for

327 sulfapyridine (Xu et al., 2014). Besides, Li et al., (2015) revealed that N1 and N11 are also the

328 reaction sites for the reactions of SD with excited triplet state of DOM proxies. Boreen et al.,

329 (2004) speculated that the electron-rich alkene in the isoxazole ring would be susceptible to

330 singlet oxygenation for SIX. However, the corresponding products were not detected.

331 In Path 2 (Fig. 4), the reaction of •OH radicals to each SA generated diverse hydroxylated

332 products, which involved multiple photooxidation pathways: phenyl hydroxylation,

333 heterocyclic hydroxylation and cleavage, as well as amino-N oxidation and removal. These

334 multi-pathways can be prioritized based on the by-product abundance in the total ion

335 chromatograms. TP283, TP271, TP294a and TP300b showed optimal responses respectively 336 for the four SAs, suggesting that phenyl hydroxylat MANUSCRIPTion is the most prevalent pathway. This 337 can be attributed to the electron-rich aniline moiety that favors the hydroxylation, which in

338 turn promotes further •OH addition, i.e., multihydroxylation (Batista et al., 2014; Shah and

339 Hao, 2017). In comparison, hydroxylated heterocyclic reactions can be considered as minor or

340 secondary pathways, which generated products with small responses (TP255, TP289, TP294b

341 and TP310). Hu et al. also detected similar hydroxy intermediates for UVA-TiO 2

342 photocatalyzed SMX degradation (Hu et al., 2007). Though some intermediates were not 343 identified, •OHACCEPTED could induce the cleavage of C−S and N−S bonds, thereby generating 344 low-molecular-mass photo-byproducts, and leading ultimately to the full mineralisation of

345 these pollutants into CO 2, H 2O and other inorganic compounds (Batista et al., 2014; Hu et al.,

346 2007).

347 In Path 3 (Fig. 4), direct photodegradation occurs through a cleavage at various positions

14 ACCEPTED MANUSCRIPT 348 for SAs containing five-membered heterocyclic groups, with sulfanilic acid (TP173) being

349 observed as a common byproduct (Boreen et al., 2004). However, SAs containing

350 six-membered heterocyclic substituents are primarily transformed to desulfonation products

351 (TP214 and TP220) with the loss of SO 2 (Boreen et al., 2005). The SO 2 extrusion can be

352 attributed to aromatic nucleophilic substitution of an aniline radical cation and subsequent

353 Smiles rearrangement (Tentscher et al., 2013). Except for the primary pathways, more direct

354 photolytic intermediates and processes have also been reported for several SAs, including ST

355 (Niu et al., 2017), SMX (Baena-Nogueras et al., 2017), SMZ and sulfapyridine (García-Galán

356 et al., 2012), as well as their human metabolites (Bonvin et al., 2013; García-Galán et al.,

357 2012).

358 Finally, the knowledge aquired through deducing the reaction pathways provides

359 fundamental insights into the photochemical fate of SAs. In turn this will allow the evaluation

360 of the phototransformation of SAs during wastewater treatment and the wider aquatic 361 environment as part of a detailed risk assessment MANUSCRIPT process. The distinct pathways of SAs 362 during reaction with •OH would support the understanding of relevant mechanisms in

363 advanced oxidation processes (where •OH is dominant), such as UV/Fenton (Trovó et al.,

0 364 2009a), TiO 2 photocatalysis (Yang et al., 2010) and Fe /bisulfite/O 2 systerms (Du et al., 2018).

365 Interestingly, compared to direct photolysis that generally leads to photoproducts of low

366 antimicrobial potency (Niu et al., 2017), ROS photooxidation of the SAs does not initially

367 destroy the core backbone structure of the molecule, indicating that the antibacterial activity 368 may persist in theseACCEPTED initial intermediates. Therefore, further studies are required to screen or 369 examine the antimicrobial activities of ROS oxidation products (see Ge et al., 2018).

370

371 Fig. 4. Proposed phototransformation pathways of four sulfonamide antibiotics, sulfisoxazole (SIX),

372 sulfathiazole (ST), sulfamethazine (SMZ), sulfachloropyridazine (SCP). The transformation products are

373 labeled “TP n”, with n standing for the molecular weight. Pathways of direct photolysis are referred to the

15 ACCEPTED MANUSCRIPT 374 previous study (Boreen et al., 2004, 2005).

375

376 4. Conclusion

377 This study provides a detailed examination of the aqueous photooxidation of nine widely

378 detected sulfonamide antibiotics (SAs), and shows that photochemical behavior can be greatly

379 influenced by the prevailing major dissociated forms of the SAs in question. The oxidation

1 380 kinetics of SAs towards •OH and O2 were found to be dependent on pH, which was

381 attributable to the disparate reactivities of individual SA protonated states. Among the three

382 dissociated forms, HSAs 0 or SAs - were more highly reactive towards •OH in most cases,

- 1 383 while SAs reacted the fastest with O2 for the majority of the SAs. The expected

1 384 environmental half-lives in sunlit surface waters suggested that O2 oxidation can be a central

385 factor in determining the fate of SAs (especially in waters with a pH > 7), with •OH oxidation

386 also contributing appreciably to phototransformation relative to direct photolysis. Moreover, 387 the three reactions involved different primary tranMANUSCRIPTsformation pathways and the creation of 1 388 multiple intermediates. Unlike direct photodegradation, O2 oxidation preferred to occur at the

389 sites of the amino moiety, whereas •OH reaction experienced multi-site oxidation with phenyl

390 hydroxylation prevailing as the major pathway. Many of these primary photo byproducts still

391 retained the core chemical structure of the parent SA, which suggests that these compounds

392 may present similar ecotoxicity and antibacterial activity as the parent chemicals.

393

394 Acknowledgements

395 This study wasACCEPTED supported by the National Natural Science Foundation of China (21577029

396 and 41476084), and China Scholarship Council (CSC) Scholarship (201704180014 and

397 201704180009). The authors are grateful to two anonymous reviewers for their useful insight

398 and additional information regarding the formation of phototransformation by-products.

399

16 ACCEPTED MANUSCRIPT 400 Appendix A. Supplementary data

401 Supplementary data to this article can be found online at

402 http://dx.doi.org/10.1016/j.watres.2018.08.026.

403

404 References

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22 ACCEPTED MANUSCRIPT

1 Table 1. Expected environmental half-lives (t1/2,E , h) for reactions with either •OH or O2, and direct photodegradation of sulfonamides (SAs) in sunlit surface waters.

a 1 a b •OH oxidation O2 oxidation direct photolysis SAs seawater, fresh water, seawater, fresh water, summer, winter ,

pH 7.9–8.4 pH 6.0–9.0 pH 7.9–8.4 pH 6.0–9.0 pH 7 pH 7 SMX 29.28–29.29 29.15–29.29 1.01–1.02 1.01–1.77 22 231 SIX 35.33–35.34 34.44–35.34 0.27 0.23–0.27 12 121 SMT 32.84–32.88 29.86–32.89 1.19 1.19–1.54 19 201 ST 14.95–15.85 14.62–27.92 0.82–0.84 0.81–1.01 3 31 SMZ 22.85–25.72 21.17–32.12 0.08–0.13 0.06–1.06 31 180 SM 25.30– 25.75 25.13–31.95 0.34–0.36 0.33–1.19 55 300 SD 24.44– 24.50 24.42–26.47 0.41–0.42 0.41–0.77 31 160 SCP 39.11–39.20 33.88–39.23 0.21 0.08–0.21 9 48 SDM 30.17–30.19 29.33–30.20 1.23–1.25 1.23–2.22 45 420 a 1 Calculated based on the •OH and O2 concentration levels in the euphotic zone of some surface waters, i.e.,

-15 1 -12 b [•OH] = 1 × 10 M and [ O2] = 1 × 10 M (Cooper et al., 1989; Mill, 1999). Corresponding to direct photolytic half-lives of SAs in near surface waters in midsummer and midwinter at 45° latitude (Boreen et al., 2004, 2005). The half-lives were calculated with MANUSCRIPT quantum yields (correcting for lens effect) and noon-averaged tabulated solar intensities (assuming continuous irradiation) (Leifer, 1988).

ACCEPTED

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O O O pKa1 pKa2 H3N S NH R H2N S NH R H2N S N R O O O + 0 - H2SAs HSAs SAs SAs, Abbr. −R SAs, Abbr. −R CH3 N O N Sulfamethoxazole, SMX CH3 Sulfamethazine, SMZ N CH3 O N CH3 N Sulfisoxazole, SIX CH3 Sulfamerazine, SM CH3 N S N Sulfamethizole, SMT CH3 Sulfadiazine, SD N N N S Sulfathiazole, ST Sulfachloropyridazine, SCP Cl N N N OCH3 Sulfadimethoxine, SDM N N OCH3 Fig. 1. Chemical structures and the different protonation states of sulfonamide antibiotics (SAs), where

+ 0 - H2SAs , HSAs , and SAs represent the cationic form, the neutral form, and the anionic form, respectively.

MANUSCRIPT

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SAs + •OH 12 pH = 2 pH = 5 pH = 8 ) -1

s 9 -1 -1 M 9 6 (× 10 •OH

k 3

0

250400 SAs + 1O 2 pH = 2 pH = 5 pH = 8 200350 ) -1 s

-1 -1 150 M 7 100 (× 10 1O2

k 50 0 MANUSCRIPT

Fig. 2. ROS oxidation rate constants ( k•OH and k1O2 ) of 9 sulfonamide antibiotics (SAs) under different pH conditions

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SA i + •OH + 0 - 16 H2SAs+H2SAs HSAs0HSAs SAsSAs- )

-1 12 s -1 -1 M 9 8 (× 10 i 4 •OH, k

0

250450 1 + 0 - SA i + O2 H2SAs+H2SAs HSAsHSAs0 SAs-SAs 200250 ) -1 s

-1 -1 150 M 7 100 MANUSCRIPT

(× 10 i

1O2, 50 k

0

Fig. 3. Bimolecular rate constants ( k•OH,i and k1O2,i) for the reaction between dissociated SA species (SA i) species and ROS

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1 O R O2 O H2N S N N Path 1 O TP352a/TP352b H3C CH3 OH OH O O H O H O H2N S N N HO S N N O O r jo TP283 H3C CH3 TP284 H3C CH3 a m OH OH O O O H O ·OH H O H O H2N S N N H2N S N N HO S N N O Path 2 O O m OH OH H3C CH3 in H3C CH3 H3C CH3 o TP299 SIX MW = 267 r TP300 O H OH H2N S N NH O TP255 H3C direct O O photolysis H2N N H2N S OH +

Path 3 O H3C CH major 3 TP173 TP112

1 O R O2 S H2N S N Path 1 O N TP283/TP297 OH O H S H2N S N O N r jo TP271 a m O O H S ·OH H S H2N S N N S N O N Path 2 O O N m in TP269 o ST MW = 255 r O H S MANUSCRIPT H2N S N O N TP289 HO OH direct photolysis O S H2N H2N S OH + Path 3 O N major TP173 TP100

CH3 1 O N O2 H N S N Path 1 O O N CH TP292 3 CH O 3 H N H2N S N O N r OH CH3 jo a m TP294a CH3 CH3 CH3 O O N O H N ·OH H H N H2N S N H2N S N H2N S N OH OH O N Path 2 O N O N ACCEPTEDm CH CH3 OH CH 3 in TP294b 3 SMZ MW = 278 o r CH TP310 O 3 H N HO S N O N CH3 TP279

direct H3C photolysis H2N N CH3 + SO2 Path 3 N HN TP214

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1 O O2 H HN S N Cl Path 1 OH O NN TP300a OH O O H H r jo H2N S N Cl H2N S N Cl a m O NN O NN OH OH O ·OH H TP300b TP316 H2N S N Cl O NN Path 2 m O in H o SCP MW = 284 r H2N S N OH O NH TP255 direct Cl photolysis N - Cl N H2N N H2N N Path 3 HN + SO2 HN TP220 TP186

Fig. 4. Proposed phototransformation pathways of four sulfonamide antibiotics, sulfisoxazole (SIX), sulfathiazole (ST), sulfamethazine (SMZ), sulfachloropyridazine (SCP). The transformation products are labeled “TP n”, with n standing for the molecular weight. Pathways of direct photolysis are referred to the previous study (Boreen et al., 2004, 2005).

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ACCEPTED MANUSCRIPT ‹ Aquatic photooxidation of sulfonamides (SAs) via reactive oxygen species is important

+ 0 - ‹ The kinetics depend on pH and separate reactivities of H 2SAs , HSAs and SAs

1 ‹ O2 plays a key role in aquatic fate of SAs compared with •OH and direct photolysis

‹ Different primary pathways and multiple intermediates occur for the three reactions

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ACCEPTED