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Water Research 118 (2017) 196e207

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Water Research

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Degradation of sulfamethoxazole by UV, UV/H2O2 and UV/persulfate(PDS): Formation of oxidation products and effect of bicarbonate

Yi Yang a, Xinglin Lu a, Jin Jiang a, *, Jun Ma a, **, Guanqi Liu a, Ying Cao a, Weili Liu a,Juan Li a, Suyan Pang b, Xiujuan Kong a, Congwei Luo a

a State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology, Harbin, 150090, Chinab Key Laboratory of Green Chemical Engineering and Technology of College of Heilongjiang Province, College of Chemical and Environmental Engineering,Harbin University of Science and Technology, Harbin, Heilongjiang, 150040, China

a r t i c l e i n f o

Article history:Received 7 September 2016Received in revised form20 March 2017Accepted 25 March 2017Available online 28 March 2017

Keywords:SulfamethoxazoleHydroxyl radicalSulfate radicalCarbonate radicalTransformation products

* Corresponding author.** Corresponding author.

E-mail addresses: jiangjinhit@126.com (J. Jiang), m

http://dx.doi.org/10.1016/j.watres.2017.03.0540043-1354/© 2017 Elsevier Ltd. All rights reserved.

a b s t r a c t

The frequent detection of sulfamethoxazole (SMX) in wastewater and surface waters gives rise of con-cerns about their ecotoxicological effects and potential risks to induce antibacterial resistant genes. UV/hydrogen peroxide (UV/H2O2) and UV/persulfate (UV/PDS) advanced oxidation processes have beendemonstrated to be effective for the elimination of SMX, but there is still a need for a deeper under-standing of product formations. In this study, we identified and compared the transformation products ofSMX in UV, UV/H2O2 and UV/PDS processes. Because of the electrophilic nature of SO4

�-, the second-orderrate constant for the reaction of sulfate radical (SO4

�-) with the anionic form of SMX was higher than thatwith the neutral form, while hydroxyl radical (�OH) exhibited comparable reactivity to both forms. Thedirect photolysis of SMX predominately occurred through cleavage of the NeS bond, rearrangement ofthe isoxazole ring, and hydroxylation mechanisms. Hydroxylation was the dominant pathway for thereaction of �OH with SMX. SO4

�- favored attack on eNH2 group of SMX to generate a nitro derivative anddimeric products. The presence of bicarbonate in UV/H2O2 inhibited the formation of hydroxylatedproducts, but promoted the formation of the nitro derivative and the dimeric products. In UV/PDS, bi-carbonate increased the formation of the nitro derivative and the dimeric products, but decreased theformation of the hydroxylated dimeric products. The different effect of bicarbonate on transformationproducts in UV/H2O2 vs. UV/PDS suggested that carbonate radical (CO3

�-) oxidized SMX through theelectron transfer mechanism similar to SO4

�- but with less oxidation capacity. Additionally, SO4�- and

CO3�- exhibited higher reactivity to the oxazole ring than the isoxazole ring of SMX. Ecotoxicity of

transformation products was estimated by ECOSAR program based on the quantitative structure-activityrelationship analysis as well as by experiments using Vibrio fischeri, and these results indicated that theoxidation of SO4

�- or CO3�- with SMX generated more toxic products than those of �OH.

© 2017 Elsevier Ltd. All rights reserved.

1. Introduction

Sulfamethoxazole (SMX) is an important antibiotic and has beenwidely used in human and veterinary medicines to treat diseasesand infections. A large portion of SMX is excreted unchanged intothe sewage system (Lienert et al., 2007; Ternes and Joss, 2006). Dueto the less efficient elimination of SMX in conventional wastewatertreatment plants, it eventually enters into the aquatic environment.

ajun@hit.edu.cn (J. Ma).

SMX has been regularly detected in wastewater effluent in a con-centration range of 100e2500 ng/L, in surface water in a concen-tration range of 60e150 ng/L, and even in drinking water at aconcentration of 12 ng/L (Al Aukidy et al., 2012; Kolpin et al., 2002;Morasch et al., 2010; Padhye et al., 2014; Ratola et al., 2012). Thechronic exposure of bacteria to trace levels of antibiotics in theaquatic environment raises concerns about their ecotoxicologicaleffects and potential risks to induce antibacterial resistant genesamong native bacterial populations. Many studies have previouslyinvestigated the occurrence and fate of SMX in the aquatic envi-ronment. Compared with biodegradation, photodegradation is apredominant pathway for SMX elimination in surface waters(Boreen et al., 2003). However, the efficiency of SMX degradation

Y. Yang et al. / Water Research 118 (2017) 196e207 197

by natural sunlight varied extensively with solar exposure by sea-sons (Bonvin et al. 2011, 2012). The frequent occurrence of SMX inthe aquatic environment indicates that more efficient treatment isneeded to destruct SMX in order to meet water demands.

Advanced oxidation processes (AOPs) are promising technolo-gies to destruct organic contaminants. Hydroxyl radical-basedAOPs (e.g., UV photolysis of hydrogen peroxide (UV/H2O2)) havebeen used for the destruction of recalcitrant organic contaminantsfrom drinking water and wastewater, as hydroxyl radicals (�OH)react with many organic chemicals at near diffusion-controlledrates (Buxton et al., 1988). Pharmaceuticals could be effectivelyremoved by �OH-based AOPs in various conditions (Baeza andKnappe, 2011; Keen and Linden, 2013; Wols et al., 2013). Sulfateradical-based AOPs (e.g., UV photolysis of peroxodisulfate (UV/PDS)) have attracted more interests, because sulfate radical (SO4�

-)also reacts with a wide range of organic contaminants with neardiffusion-limited rate constants (Neta et al., 1988). Although SO4�

- isa strong oxidant with a redox potential (2.5e3.1 V) (Neta et al.,1988) comparable to that of �OH (1.8e2.7 V) (Buxton et al., 1988),SO4�

- is a more selective oxidant than �OH.The rate constant for the reaction of SMX with SO4�

- wasdetermined to be (11.7e16.1) � 109 M�1 s�1 in the pH range from 6to 9, which was slightly higher than that with �OH at(7.02e7.89) � 109 M�1 s�1 (Zhang et al., 2015). Meanwhile, directphotolysis at 254 nm was identified as an effective pathway forSMX degradation with a relatively high quantum yield (Canonicaet al., 2008; Wols et al., 2013). These results suggest a potentialapplication of UV-based AOPs in the removal of SMX from variouswaters. The apparent destruction efficiency of SMX by UV at254 nm, UV/H2O2 and UV/PDS in reverse osmosis (RO) brines frommunicipal wastewater reuse facilities has been investigated in ourprevious study (Yang et al., 2016). The average UV fluence for 50%removal of SMX was less than 180 mJ/cm2 in two RO brine samples.Anions at high concentration in RO brines acted as �OH and SO4�

-

scavengers to generate reactive species. Bicarbonate has beensuggested to be an important scavenger of �OH, SO4�

- and halogenradicals to form carbonate radical (CO3�

-) (Anastasio and Matthew,2006; Grebel et al., 2010; Yang et al., 2014). Modeling resultsindicated that CO3�

- concentration exceeded those of other radicalsby several orders of magnitude in RO brines, especially in UV/PDSprocess (Yang et al., 2014). CO3�

- is a more selective oxidant than�OH and SO4�

-, but it is sufficient to oxidize some pharmaceuticals(Canonica et al., 2005). The presence of bicarbonate promoted thedegradation of SMX by UV/PDS, but slightly inhibited that by UV/H2O2 (Yang et al., 2016). Zhang (Zhang et al., 2015) found similarresults in synthetic human urine samples and explained this dif-ference by the lower rate constants for the reaction of CO3�

- withPDS than that with H2O2. The rate constant for the reaction of CO3�

-

with SMXwas determined to be (2.68 ± 0.71)� 108 M�1 s�1 (Zhanget al., 2015).

Understanding the transformation products (TPs) of SMX andcharacterizing the reaction pathways in different oxidation pro-cesses are critical to evaluating the risk associated with the pres-ence of SMX in the aquatic environment. The TPs of SMX byphotolysis have been well investigated in solar and UV 254 nmphotolysis (Bonvin et al., 2012; Trov�o et al., 2009; Zhou and Moore,1994), and many efforts have also been put on investigating TPsformation of SMX in �OH- and SO4�

--based processes includingelectrochemistry, thermo-activated persulfate and perox-ymonosulfate/cobalt(II) systems (Ji et al., 2015; Lin et al., 2013;Mahdi Ahmed et al., 2012). Nevertheless, the formation mecha-nisms of the TPs of SMX by direct photolysis combined with �OH orSO4�

- oxidation (i.e., UV/H2O2 and UV/PDS), as well as the relativeyields of TPs, are not well documented. Also, little work has beenconducted to address the TPs of SMX by CO3�

- oxidation.

The objective of this study is to identify and compare the TPs ofSMX in UV, UV/H2O2 and UV/PDS processes. The contributions of�OH and SO4�

- on the TPs formation were investigated in UV-basedAOPs. The relative yields of TPs by �OH and SO4�

- were evaluated,revealing the different reactivity of �OH and SO4�

- to the specificgroups of SMX. This study is the first time to investigate the TPs ofSMX by CO3�

- and assess the impact of bicarbonate on ecotoxicityduring UV/H2O2 and UV/PDS processes.

2. Materials and methods

2.1. Materials

Sulfamethoxazole (SMX), potassium perdisulfate (PDS), ammo-nium acetate were purchased from Sigma-Aldrich. Sulfamethoxa-zole-d4 (SMX-d4) was purchased from Santa Cruz Biotech.Hydrogen peroxide (H2O2) solution (35% w/w), tert-butanol (t-BuOH) and sodium bicarbonate were purchased from SinopharmChemical Reagent Co. Ltd., China. HPLC grade methanol and aceticacid were received from Thermo Fisher Scientific, and acetonitrilewas received fromMerck. All solutions were prepared in deionized(DI) water (18.2 MU/cm) from a Milli-Q purification system (Milli-pore, Billerica, MA).

2.2. Experimental procedures

UV apparatus was applied using a semi-collimated beam systemcomprising four 10 W low pressure mercury lamps emitting at254 nm shining down onto a 100 mL crystallization dish(pathlength ¼ 4.1 cm), as described previously (Yang et al., 2015).Samples were withdrawn periodically and supplemented with20 mL methanol per mL sample to quench any radicals formed bythermolysis of PDS or H2O2, and kept in a 4 �C refrigerator forfurther analysis within 12 h. The surface irradiance(8.69 � 10�7 Einsteins L�1 s�1) was determined by iodide-iodateactinometry (Rahn et al., 2003). Experiments were conducted in10 mM phosphate buffer at 20 ± 2 �C.

Some experiments were conducted with filtrated waters fromtwo drinking water plants using surface water and ground water aswater sources. These water samples were filtered through 0.7 mmnominal pore size glass fiber filters (Whatman) and stored at 4 �C.The surface water had a low DOC and low alkalinity (DOC 4.24 mg/L, alkalinity 0.2 mMHCO3

�, UV254¼ 0.045, pH 7.2), while the groundwater showed a relatively high DOC and high alkalinity (DOC6.94 mg/L, alkalinity 8 mM HCO3

�, UV254 ¼ 0.034, pH 8.25).

2.3. Kinetics of the reactions of SMX with �OH and SO4�-

Since pKa,1 (1.7) of SMX is not relevant to natural water condi-tions, only pKa,2 (5.7) was considered herein. The pH-dependentkinetics of direct photolysis of SMX were performed in the pHrange of 3e8 (see Text S1 for details). To evaluate the specific rateconstants for the reactions of �OH and SO4�

- with the neutral andanionic forms of SMX, the apparent pH-dependent rate constants(kapp;�OH and kapp;SO��

4) were determined by competition kinetics

(see Text S2 for details). kapp;�OH and kapp;SO��4

are expressed asequation (1) and equation (2), respectively.

kapp;�OH ¼ aSHkSH;�OH þ aS�kS�;�OH (1)

kapp;SO��4¼ aSHkSH;SO��

4þ aS�kS�;SO��

4(2)

where kSH;�OH , kS�;�OH , kSH;SO��4and kS�;SO��

4are the specific rate con-

stants for the reactions of �OH and SO4�- with neutral and anionic

Fig. 1. Apparent second order rate constants for reactions of SMX with �OH (A) andSO4

�- (B) as a function of pH.

Y. Yang et al. / Water Research 118 (2017) 196e207198

forms, respectively; aSH and aS� are the proportions of neutral (SH)and anionic (S�) forms at a given pH and can be calculated byequation (3) and equation (4), respectively.

aHS ¼½Hþ�

Ka;2 þ ½Hþ� (3)

aS� ¼ Ka;2

Ka;2 þ�Hþ� (4)

Therefore, the second-order rate constants of �OH or SO4�- with

specific forms of SMX were calculated by nonlinear regression ofpH-dependent kapp;�OH and kapp;SO��

4data.

2.4. Identification of oxidation products

SMX was analyzed using a Waters 1525 HPLC with a Waters2487 dual l detector. Chromatographic separations were per-formed using a Waters symmetry C18 column (150 mm � 4.6 mm,5 mm). The concentrations of SMX were quantified at l ¼ 260 nm,with an eluent consisting of 0.1% acetic acid and methanol with aratio of 40:60 (v/v) at a flow rate of 1 mL/min.

To identify the oxidation products of SMX, a triple quadrupoleTOF mass spectrometer (Triple TOF 5600, AB Sciex) coupled withthe Ekspert ultralLC110 was used. Chromatographic separationswere performed using a Poroshell 120 EC-C18 column(50 mm � 3.0 mm, 2.7 mm). Accurate MS and MS/MS patterns ofSMX and its oxidation products were analyzed in a molecular ionscanning mode (m/z 50 to 1000) in both positive and negativeelectrospray ionization (ESI) modes. Analyst (version 1.6, AB Sciex)and PeakView (version 1.2.0.3, AB Sciex) were employed to identifyand analyze peaks. The obtained fragmentation patterns of prod-ucts were incorporated into a multiple reaction mode (MRM)method to quantify the formation of products by a triple quadru-pole mass spectrometer (QTrap 5500 MS, ABSciex) coupled withthe Agilent 1260 HPLC. The LC program was identical to thatdescribed above. Peak areas of TPs were normalized using the in-ternal standard (SMX-d4). Further information on the appliedsetup, the optimized chromatographic conditions and the condi-tions of data acquisition parameter is provided in Text S3.

2.5. Toxicity analysis

The acute toxicity assay of samples treated by UV, UV/H2O2, andUV/PDS was carried out by measuring the decrease in the biolu-minescence of Vibrio fischeri (Hamamatsu Photonics, Beijing,China). Residue H2O2 or PDS in samples were quenched by bovinecatalase and ascorbic acid, respectively. Both bovine catalase andascorbic acid had a negligible impact on luminescence. As 500 mMbicarbonate exceeded the osmolality of Vibrio fischeri, only 50 mMbicarbonatewas tested for the acute toxicity. The luminescencewasdetected by a luminometer (Promega GloMax®-Multi Jr). The acutetoxicity of SMX and its TPs were also estimated using the EcologicalStructure Activity Relationships (ECOSAR) program for fish, daph-nid and green algae.

3. Results and discussion

3.1. Oxidation kinetics of the reactions of SMX with �OH and SO4�-

The pH dependence of SMX degradation by direct photolysis hasbeen investigated by Canonica et al., (2008). The quantum yields ofneutral and anionic forms of SMX determined in this study were0.187 ± 0.002 mol Einstein�1 and 0.028 ± 0.002 mol Einstein�1

(Text S1), respectively, which were consistent with their results.

As pH varied from 3 to 8, no significant difference of kapp;�OH forreaction of SMX with �OH was observed (Fig. 1 A). This resultindicated that �OH exhibited similar reactivity to the neutral formand anionic form of SMX. Therefore, the same value of kSH;�OH andkS�;�OH was suggested herein and determined to be(7.63 ± 0.85) � 109 M�1 s�1. Interestingly, kapp;SO��

4increased grad-

ually with increasing pH from 3 to 8 (Fig. 1 B). This phenomenonwas ascribed to deprotonation of the secondary amine group inSMX, and SO4�

- was expected to have high reactivity as a result ofits electrophilic nature.kS�;SO��

4was measured to be

(1.34 ± 0.02) � 1010 M�1 s�1, which was 14-fold higher than kSH;SO��4

((9.29 ± 1.5) � 108 M�1 s�1). The reaction kinetics of SO4�- with

substructure model substrates of SMX demonstrated that anilinewas the most reactive group to SO4�

- (Ji et al., 2015). Substitution bythe sulfonic group or the sulfonamide group at the para position ofaniline reduced the rate constants significantly, while 3-amino-5-methyl-isoxazole was also reported to be less reactive than ani-line (Ji et al., 2015). When eNH2 group of SMX was substituted byeNHCOCH3, the second-order rate constants of acetyl-SMX wasonly one fifth of that of SMX (Zhang et al., 2015). These resultsindicate that the reaction of SO4�

- with the SMX structure takesplace primarily at the p-sulfonylaniline group. The pH dependenceof kapp;SO��

4likely reflects enhancement of the electron-donating

effect of SMX's aniline group by sulfonamide nitrogen deprotona-tion, which may favor the electrophilic attack by SO4�

-.

Y. Yang et al. / Water Research 118 (2017) 196e207 199

3.2. Product identification

Several studies have examined transformation products (TPs)arising from direct photolysis, �OH- and SO4�

--based oxidations ofSMX. The major photoproducts were formed from cleavage of theNeS bond (Boreen et al., 2004; Trov�o et al., 2009). Hydroxyl de-rivatives of SMX were produced during �OH- or SO4�

--dominantprocess (Hu et al., 2007; Zhang et al., 2016). In addition, SO4�

- wasreported to favor oxidation of eNH2 group of SMX (Mahdi Ahmedet al., 2012). However, only few studies quantified the formationproducts among UV, UV/H2O2 and UV/PDS (Zhang et al., 2016).

TPs produced by direct photolysis, �OH, SO4�- and CO3�

- wereidentified by ESI-TOF-MS. The accurate m/z values are provided inTable S2. For structural elucidation, the fragmentation pathways ofTPs were studied by performing product ion scans. The production spectra and the fragmentation pathways of SMX([m þ H]þ ¼ 254.0591) were illustrated in Fig. S3, where m/z 156,m/z 108, m/z 99 and m/z 92 were the characteristic fragment ionsof SMX. The chemical structures of TPs could be elucidated basedon the fragment ions of SMX. Normalized peak areas were used tocompare the relative yields of TPs of SMX.

3.2.1. Direct photolysisThe formation of TPs by direct photolysis of SMX was conducted

at pH 3, where more efficient degradation of SMX could result indetectable TPs. The normalized peak areas of main TPs at pH 3wereshown in Fig. 2. TP 98 ([m þ H]þ ¼ 99.0554) and TP 173 ([m -H]-¼ 172.0074) were confirmed as 3-amino-5-methylisoxazole andsulfanilic acid (Fig. S4 and Fig. S5), respectively, resulting fromcleavage of the NeS bond of SMX. The peak area of TP 98 increasedin the first 30 min and then reached a plateau, which was consis-tent with the degradation trend of SMX. This result indicated that3-amino-5-methylisoxazole was not susceptible to bephotodegraded.

TP 253 ([m þ H]þ ¼ 254.0588, Fig. S7) reported by Zhou andMoore, (1994) as an isomer of SMX was produced through photo-isomerization of the isoxazole ring. TP 253 was accumulated duringthe initial 20 min and then gradually decreased, suggesting that theisomer of SMX underwent photodegradation. Two peaks with anominal mass [m þ H]þ ¼ 270.0542 were found (see Fig. S8 forexample). Their MS2 spectra exhibited similar fragment ions. TheMS2 spectra of TP 269 showed characteristic fragment ions at m/z172,124 and 108, suggesting the addition of one oxygen atom to thesulfanilic group. The occurrence of the m/z 99 fragment ion indi-cated that the isoxazole ring remained unchanged. Although theretention time of TP 269 at 8.5 min matched that of an authenticstandard of N4-hydroxy-sulfamethoxazole, the relative intensity offragment ions was different. This indicated the addition of hydroxylgroup of the benzene ring, which was also suggested by previousstudies (Gao et al., 2014; Trov�o et al., 2009). Due to the insufficientseparation of isomers during chromatography, a total peak area ofTP 269 was calculated. The peak area of TP 269 reached themaximum value at 20 min and then gradually decreased. TP 189([m - H]- ¼ 188.0013) might result from cleavage of the NeS bondby the photodegradation of TP 269 or the hydroxylation of TP 173.The product ion scan of both TP 271 ([m þ H]þ ¼ 272.0700) and TP287 ([m þ H]þ ¼ 288.0635) showing the fragment ions at m/z 156,108 and 92 signified an unmodified sulfanilic group (Fig. S9 andFig. S11). The absence of fragment ion at m/z 99 suggested the highreactivity of the isoxazole ring during direct photolysis. TP 271 andTP 287 were considered to result from the addition of one and twooxygen atoms to the double bond in the isoxazole ring to formcorresponding alcohols, respectively. Two peaks of TP 287 wereseparated at distinct retention times. The peak of TP 287 at 3.6 minwas identified to derive from TP 253, while the peak of TP 287 at

6.8 min was identified to derive from SMX (see details in Section3.2.4 below). The peak area of TP 271 decreased after 20 min, whilethose of TP 287-1 and TP 287-2 remained constant. Similar TPs ofSMX by direct photolysis were observed at pH 8. Note that themaximum peak value of TP 269 at pH 8 was eight times higher thanthat at pH 3. This indicated that the photolysis efficiency of TP 269was different as pH varied.

3.2.2. Identification of reactive species during direct photolysisBesides the degradation of SMX by direct photolysis, reactive

species (e.g., 1O2 and �OH) produced during SMX photolysis couldalso contribute to the degradation of SMX (Zhou and Moore, 1997).�OH reaction can be scavenged by t-BuOH with the second-orderrate constant of 6.0 � 108 M�1 s�1 (Buxton et al., 1988), but 1O2 isinert to t-BuOH (Rodgers, 1983). Therefore, the oxidation pathwayby �OH would be completely inhibited in a large excess of t-BuOH(10 mM). The degradation rate of SMX was not affected in thepresence of t-BuOH (Fig. 2). The generation of TP 98, TP 253, TP 271and TP 287-1 also exhibited similar trends as that without t-BuOH.However, the maximum peak area of TP 269 was reduced by 64%,while no TP 287-2 was observed in the presence of t-BuOH. Thedecrease of these hydroxylated products suggested that �OHcontributed to the formation of secondary products. Specifically, TP287-2 could derive from SMX by �OH, while TP 253 might be oneprecursor of TP 269.

NaN3 was an effective scavenger of 1O2 with the second-orderrate constant of 1 � 109 M�l s�l (Gsponer et al., 1987), while italso effectively quenched �OH with the second rate constant of1.2 � 1010 M�l s�l (Buxton et al., 1988). The presence of NaN3inhibited the degradation of SMX, due to an inner filter effect ofNaN3 for direct photolysis (Fig. 2). The quantum yield of SMX wasequal to that without NaN3, indicating that the degradation of SMXby 1O2 was negligible, which was consistent with previous studies(Boreen et al., 2004; Zhou and Moore, 1997). The peak area of TP254 increased, as the peak area of TP 271 decreased. These resultssuggested that 1O2 might contribute to the degradation of TP 253leading to the formation of TP 271. Neither TP 269 nor TP 287-2 wasdetected, because NaN3 scavenged �OH as well.

3.2.3. UV/H2O2

3.2.3.1. pH 3. The apparent degradation rate of SMX in UV/H2O2was comparable to that by direct photolysis (Fig. 2). Because of thehigher phototransformation quantum yield at pH 3, the directphotolysis was a major pathway for SMX degradation. Although�OH contributed to only 13% of SMX degradation, TPs were signif-icantly affected by the involvement of �OH. The major differencesbetween UV/H2O2 and direct photolysis were the formation ofhydroxylated products. For UV/H2O2 process, the maximum peakarea of TP 269 increased by around 15 times, and the subsequentdegradation of TP 269 was attributed to further oxidation by �OH.Interestingly, �OH promoted the formation of TP 287-2 (derivedfrom SMX), while it inhibited the formation of TP 287-1 (derivedfrom TP 253, an isomer of SMX). These results suggested that �OHfavored the addition reaction on the isoxazole ring over the oxazolering, which was consistent with t-BuOH quenching experimentabove. In contrast to direct photolysis, the peak area of TP 98decreased after 25 min, implying further oxidation of TP 98 by �OH.

3.2.3.2. pH 8. Fig. 3 depicted the degradation of SMX and the for-mation of TPs at pH 8. The contribution of �OH to SMX degradationincreased to 54% at pH 8. Maximum peak areas of both TP 253 andTP 271 in UV/H2O2 were only half of those formed in directphotolysis. Meanwhile, the maximum peak area of TP 269 wasseven times higher in UV/H2O2 than that in direct photolysis. Theformation of TP 287-1 and TP 287-2 was not observed. This result

Fig. 2. Degradation of SMX and evolution of main TPs during UV, UV/H2O2 and UV/PDS at pH 3. Peak areas were normalized using the peak area of the internal standard.Experimental condition: [SMX]0 ¼ 20 mM, [H2O2]0 ¼ 1 mM, and [PDS]0 ¼ 1 mM.

Y. Yang et al. / Water Research 118 (2017) 196e207200

indicated that the further oxidation of TP 287-1 and TP 287-2 by�OH, leading to their peak areas below the detection limit.

3.2.4. UV/PDS3.2.4.1. pH 3. The apparent degradation rate of SMX by UV/PDSwas close to that by direct photolysis (Fig. 2), where SO4�

-

contributed to only 7% of SMX degradation. However, the genera-tion of TPs in UV/PDS was significantly different from that in directphotolysis. The peak of TP 253 was hardly detected in UV/PDS,while the maximum peak area of TP 271 was as half as that in directphotolysis. Meanwhile, SO4�

- resulted in an increase in the forma-tion of TP 287-1, and themaximum peak area of TP 287-1 was three

times greater than that in direct photolysis, suggesting a correlationbetween TP 253 and TP 287-1. The peak area of TP 287-2 wascomparable to that in direct photolysis. Considering that theretention time of SMX was longer than TP 253 by the separation ofchromatography, a longer retention time of TP 287 derived fromSMX was expected. This evidence confirmed that TP 287-1 and TP287-2 contained an oxazole ring and an isoxazole ring, respectively.Therefore, SO4�

- showed higher reactivity to the oxazole ring thanthe isoxazole ring. These results also indicated that SO4�

- favorablydestructed the TPs formed in direct photolysis (i.e., TP 253).

Furthermore, four new TPs were produced in UV/PDS. Theretention time of TP 269was different from those observed in direct

Fig. 3. Effect of bicarbonate on degradation of SMX and evolution of main TPs during UV/H2O2 at pH 8. Peak areas were normalized using the peak area of the internal standard.Experimental condition: [SMX]0 ¼ 20 mM, and [H2O2]0 ¼ 1 mM.

Y. Yang et al. / Water Research 118 (2017) 196e207 201

photolysis and UV/H2O2 (Fig. S14). Their similar MS2 spectra indi-cated that �OH and SO4�

- exhibited different reactivity to differentpositions on the sulfanilic group. Due to different structures of TP269, the peak area of TP 269 in UV/PDS was not compared withthose in direct photolysis and UV/H2O2. TP 283([m þ H]þ ¼ 284.0328, Fig. S10) was identified as nitro-SMX. A lossof 46 Da corresponding to the nitro group resulted in a m/z 238fragment ion. The fragment ion m/z 122 resulted from cleavage ofthe NeS bond to generate nitrobenzene group. Two dimericproducts were also detected during the degradation of SMX in UV/PDS. TP 502 ([m þ H]þ ¼ 503.0813, Fig. S12) was proposed as

azosulfamethoxazole, and TP 518 ([m þ H]þ ¼ 519.0755, Fig. S13)was a hydroxyl substitution of TP 502.

3.2.4.2. pH 8. Since the rate constant for the reaction of SO4�- and

SMX increased at a higher pH, SO4�- and direct photolysis contrib-

uted to 72% and 28% of SMX degradation in UV/PDS, respectively.No TP 253 was observed in Fig. 4. A higher relative yield of TP 98was found in UV/PDS over both direct photolysis and UV/H2O2. Thepeak area of TP 284 reached the maximum value as completeremoval of SMX. The decrease of TP 502 peak area after 60 minsuggested further oxidation by SO4�

- to form other TPs, like TP 518.

Fig. 4. Effect of bicarbonate on degradation of SMX and evolution of main TPs during UV/PDS at pH 8. Peak areas were normalized using the peak area of the internal standard.Experimental condition: [SMX]0 ¼ 20 mM, and [PDS]0 ¼ 1 mM.

Y. Yang et al. / Water Research 118 (2017) 196e207202

3.2.5. Effect of bicarbonateThe effect of bicarbonate was investigated at pH 8. Relatively

high concentrations of bicarbonate to generate a CO3�- dominant

condition were chosen as 50 and 500 mM, which scavenged 74%and 97% of �OH, and 34% and 84% of SO4�

-, respectively. In thepresence of bicarbonate (Fig. 3), the degradation of SMX wasinhibited in UV/H2O2. The addition of 50 mM bicarbonate in UV/PDS slightly enhanced the degradation of SMX, while the effect of500 mM bicarbonate on SMX degradation was negligible (Fig. 4).These results might be explained by the lower rate constant of CO3�

-

with PDS than that with H2O2 (Zhang et al., 2015).The presence of bicarbonate at 50 mM in UV/H2O2 inhibited the

formation of TP 253, TP 269 and TP 271 (Fig. 3). Neither TP 253 norTP 269 was observed with 500 mM bicarbonate. Consistent withthe relatively low yield of TP 271, TP 287-1 was detected in thepresence of bicarbonate. Meanwhile, the peak area of TP 98 waspromoted with 500 mM bicarbonate. The formations of TP 283, TP502 and TP 518, derived from the oxidation of aniline group, werealso enhanced in the presence of bicarbonate. Since the rate con-stant of �OH with SMX was 30 times higher than that of CO3�

-, thedegradation of SMXwas still predominant by �OH in the presence of50 mM bicarbonate, and thus producing only small amounts ofthese TPs. As bicarbonate concentration increased to 500 mM, the

relative yields of these TPs were promoted substantially. The for-mation trends of these TPs were similar to those in UV/PDS, sug-gesting that the reaction pathway of SMX degradation by SO4�

-

likely also applied to CO3�-.

UV/PDS produced similar TPs in the presence or absence of bi-carbonate (Fig. 4). The generation of CO3�

- increased the relativeformation yield of TP 98. The accumulation of TP 283 was slower inthe presence of bicarbonate than that in the absence of bicarbonatein 60 min, and continued increasing after that. This result impliedthe stronger oxidation capacity of SO4�

- than CO3�- to further oxidize

TP 284. This analysis also explained a greater peak area of TP 502,but a smaller peak area of TP 518 obtained in the presence ofbicarbonate.

3.3. Proposed transformation pathways

3.3.1. Direct photolysis pathwayThe formation of the identified TPs occurred through cleavage

reaction by photolysis or oxidation by reactive species (i.e., 1O2, �OH,SO4�

- and CO3�-). For direct photolysis (Scheme 1), TP 98 and TP 173

were formed through cleavage of the NeS bond of SMX. Previousstudies also found the formation of aniline from cleavage of the CeSbond by solar irradiation (Bonvin et al., 2012; Zhou and Moore,

Scheme 1. Proposed reaction pathway for SMX degradation by UV and UV/H2O2. The solid arrows represent major pathways, and dash arrows represent minor pathways.

Y. Yang et al. / Water Research 118 (2017) 196e207 203

1994). However, no peak matched an authentic standard of anilinein this study. TP 253 was generated by rearrangement of the iso-xazole ring, which can be further degraded in direct photolysis. Theaccumulation of TP 253 and the smaller peak area of TP 271 withthe presence of NaN3 indicated that TP 253 was one of the pre-cursors for TP 271. The reaction of 1O2 involved a cycloaddition tothe electron-rich olefinic double bond. The exact mechanism ofhydroxylated products was not yet known. Possible mechanismsmay include a rearrangement of addition product to form epoxide(Frimer, 1979), which could be further reduced by photosensitizedelectron transfer reactions (Zhou and Moore, 1997). Additionally,the formation of TP 271 and TP 287-1 in the presence of 1O2 and�OH scavengers indicated that an oxidation pathway still occurred.This was likely attributed to the generation of intermediate radicalsby t-BuOH and NaN3, as indicated by previous studies that�CH2C(CH3)2OH and N3� possibly reacted with some organic com-pounds (NIST, 2016).

3.3.2. �OH pathway�OH exhibited high reactivity with olefinic double bonds and

anilines (Buxton et al., 1988). One of the main reaction pathways of�OH was an addition reaction to the double bond located on theisoxazole (oxazole) ring (Scheme 1). The formation of TP 287-1 andTP 287-2 were somewhat surprising since they were saturated al-cohols. Hu et al., (2007). studied the transformation products ofSMX by TiO2 photolysis and also found the formation of TP 287 (noisomers). They proposed that the addition of �OH to C]C of theisoxazole ring lead to the formation of a tertiary carbon-centeredradical (Hu et al., 2007). This radical reacted with oxygen to givea peroxy radical, which may abstract a hydrogen from a donor andform the corresponding hydroperoxide (Hu et al., 2007). The ho-molytic cleavage of hydroperoxide produced a hydroxyl radical andan alkoxy radical (Larson and Weber, 1994), which can further ab-stract a hydrogen atom to form the corresponding saturated alcoholproduct (Hu et al., 2007). This mechanism provided a possibleexplanation for the degradation of TP 98 by �OH.

In the reaction of �OH with aniline, the abstraction of hydrogen

is a preferred reaction pathway. �OH was expected to attack anilineon various positions. Solar et al., (1986). investigated the absor-bance characteristics of formed aniline radicals by �OH attack. Theyfound that the main primary radical (54%) resulted from the reac-tion of �OH with the ortho-position of aniline, while 36% of �OHreacted directly with eNH2 group, and the remaining 10% of �OHprobably attacked the para-position (Solar et al., 1986). Hydrolysisof aniline radicals formed hydroxylated radical products, whichfurther reacted with O2 to produce hydroxylated anilines. Thismechanism explained the formation of TP 269 through the reactionon the ortho- or para-positions of aniline. TP 502 was subsequentlyformed through coupling of the N-centered radical derived fromeNH2 group.

3.3.3. SO4�- pathway

One important reaction pathway of SO4�- was initiated by the

electrophilic attack at the olefinic double bond in the isoxazole(oxazole) ring to generate olefinic radical (Scheme 2). Hydrolysis ofthe olefinic radical and subsequent reactions are similar to that of�OH, which gave rise to alcohol products (TP 287-1 and TP 287-2).The relative yield of TP 287-1 by SO4�

- was much higher than thatby �OH, whereas the relative yield of TP 287-2 by �OH was muchhigher than that by SO4�

-. Because SO4�- was an electrophilic and

more selective oxidant than �OH, electron-donating groups couldenhance the reactivity towards SO4�

-. In particular, the nitrogenatom in the oxazole ring could enhance electron density of theolefinic double bond and exhibited stronger electron-donatingproperties than the olefinic double bond in the isoxazole ring. Itindicated that SO4�

- preferred to react with the oxazole ring thanwith the isoxazole ring, resulting in the higher relative yield of TP287-1. The higher relative yield of TP 287-2 by �OH was ascribed tothe relatively higher concentration of SMX than TP 253.

Furthermore, the reaction of SO4�- with eNH2 group was pro-

posed to produce N-centered radical (Scheme 2). The N-centeredradical readily reacted with water, and this reaction was followedby a fast elimination of Hþ to form anilino radical. The addition ofoxygen to the formed anilino radical resulted in the formation of

Scheme 2. Proposed reaction pathway for SMX degradation by UV/PDS in the absence and presence of bicarbonate. The solid arrows represent major pathways, and dash arrowsrepresent minor pathways.

Fig. 5. Impact on Vibrio fischeri luminescence by SMX after different treatments. Thedash line represents the luminescence induced by the remaining SMX. Errors representthe standard deviation (n ¼ 4).

Y. Yang et al. / Water Research 118 (2017) 196e207204

hydroxylamine with the release of HO2�. Further oxidation of hy-droxylamine led to the formation of nitroso derivative (MahdiAhmed et al., 2012). Hydrolysis of the nitroso derivative resultedin the formation of nitro derivative (i.e., TP 284). Meanwhile, thecoupling of the N-centered radical generated TP 502, which wasfurther oxidized to generate TP 518. The different behaviors of TPsbetween �OH and SO4�

- depended on their reactivity to specificgroups of SMX. SO4�

- would attackeNH2 groupmore efficiently andthus favor the formation of nitrogen oxidized products.

3.3.4. CO3�- pathway

CO3�- was a more selective oxidant than �OH and SO4�

-(Buxtonet al., 1988; Neta et al. 1977, 1988), and thus reacted with organiccompounds through electron transfer as SO4�

-. The conversion of�OH to less reactive CO3�

- was responsible for the decrease of TP 269(Fig. 3). CO3�

- was more selective to attack eNH2 group, increasingthe formations of TP 283 and TP 502 in both UV/H2O2 and UV/PDS.Interestingly, the formation of TP 518 was slow in UV/H2O2 at initial7 min, and then the accumulation of TP 518 was accelerated (Fig. 3).This result indicated that TP 518 was the secondary product andmight derive from TP 502 or the coupling reaction between SMXand TP 269. However, this lag phase was not obviously observed inUV/PDS (Fig. 4), which was ascribed to the higher formation yield ofSO4�

- than �OH to shorten the initial lag phase. Additionally, theformation of TP 518 increased in UV/H2O2, but decreased in UV/PDS, indicating that SO4�

- was a stronger oxidant than CO3�- to form

the further oxidized products. CO3�- attacked the secondary amine

Y. Yang et al. / Water Research 118 (2017) 196e207 205

leading to cleavage of the NeS bond, which might explain the highyield of TP 98.

3.4. Toxicological implications of transformation products

3.4.1. Acute toxicity to Vibrio fischeriThe luminescence inhibition of Vibrio fischeri by SMX after

different treatments was shown in Fig. 5. L0 is the luminescence ofeach sample without treatment, while L is the luminescence of thesample with corresponding treatment. The ratio of L/L0 was used asan indicator of acute toxicity; namely, the lower value of L/L0

Fig. 6. SMX degradation and main TPs formation in authentic waters. SW and GW representthe peak area of the internal standard. Experimental condition: [SMX]0 ¼ 1 mM, [H2O2]0 or

indicated a higher acute toxicity, and vice versa. The dash linerepresents the luminescence induced by the remaining SMX in thesample.

With the degradation of SMX in UV and UV/H2O2, the lumi-nescence slightly increased. Comparatively, the luminescence ofsamples treated by UV/PDS reduced by 65%, indicating a hightoxicity of TPs. The higher toxicity of the samples treated by UV/PDSwas also observed by Vibrio qinghaiensis in freshwater (Zhang et al.,2016). The presence of 50 mM bicarbonate had no significant effecton luminescence in both UV/H2O2 and UV/PDS. This result wasascribed to the fact that bicarbonate at 50 mM did not provide a

the surface water and the ground water, respectively. Peak areas were normalized using[PDS]0 ¼ 1 mM.

Y. Yang et al. / Water Research 118 (2017) 196e207206

CO3�- dominant condition to destruct SMX under the conditions

investigated. Unfortunately, due to the limitation of the osmolalityof Vibrio fischeri, a higher concentration of bicarbonate cannot beapplied in this experiment. Nevertheless, the similar productsgenerated by SO4�

�and CO3�- implied that these products were

more toxic than those generated by �OH.

3.4.2. Toxicity assessment of TPs by ECOSARThe antibacterial activity of SMX is originated from its compe-

tition with p-aminobenzoic acid for enzyme inhibition and meta-bolic interference, which is essential for bacterial folic acidsynthesis (Majewsky et al., 2014). Three types of TPs were observedin this study: (i) TPs retaining the sulfonamide toxicophore; (ii)breakdown TPs not exhibiting the toxicophore; (iii) dimeric prod-ucts. To demonstrate different responses of SMX and TPs on variousspecies (i.e., fish, daphnid and green algae), QSAR analysis byECOSAR program was conducted (Table S4). According to the pre-vious result of 48-h half effective concentration (EC50) value for D.magna (Kim et al., 2007), the value of class aniline (unhindered)was selected for prediction.

Except for TP 253, the acute and chronic toxicity of four TPsretaining sulfonamide group exhibited lower toxicity than those ofSMX. This modeling result showed a similar trend with theexperimental result using Vibrio fischeri (Majewsky et al., 2014).Additionally, Lienert et al., (2007). suggested that the toxicity ofparent compound and its metabolite was positively related to theiroctanol-water partition coefficients (Kow). The hydroxylated prod-ucts of SMX increased hydrophilicity, which reduced the toxicity ofTPs. TP 283was expected to possess higher toxicity. Majewsky et al.,(2014). investigated the antibacterial activity of selected SMX TPs inthe aquatic environment. Their results indicated that the trans-formation of SMX at the para-position still remained or evenincreased its toxicity. The increasing toxicity of TP 283 wasexplained by a negative inductive effect of the nitro group, whichmight either favor release of the proton in the amide group toenhance the affinity of the key enzyme for folic acid synthesis, orreplace p-hydroxybenzoic acid to form faulty products (Majewskyet al., 2014).

For breakdown products, TP 173 and TP 189 were less toxic thanSMX. In contrast, higher toxicity of TP 98 was predicted by ECOSAR,which was opposite to the experimental data (Majewsky et al.,2014). The prediction of the isoxazole ring by the class of aniline(unhindered) might lead to a large bias to the real value.

For dimeric products, both TP 502 and TP 518 exhibited highertoxicity towards fish. The mechanism of azo carcinogenicity wasproposed including: (i) cleavage of the azo bond to generate ani-lines during metabolic reactions; (ii) oxidation of azo to highlyreactive electrophilic diazonium salts (Brown and De Vito, 1993). Itwas likely that the oxidation of SO4�

- or CO3�- with SMX generated

more toxic products than that of �OH.

3.5. Oxidation in authentic water

Most of the TPs identified in the synthetic solutions could alsobe found in the authentic waters (Fig. 6). No dimeric products wereobserved in both water samples treated by UV/H2O2. This resultmight be ascribed to the relatively low concentration of SMX in theauthentic waters than that in the synthetic water, resulting in theirpeak areas below the detection limit. Interestingly, TP 283 wasdetected in both water samples treated by UV/PDS, while TP 502and TP 518were only identified in the groundwater. The absorptionat 254 nm represented the content of aromatic moieties of DOC.The ground water contained higher DOC with lower UV254 inrelative to those of the surface water. As selective radicals, SO4�

- andCO3�

- might be more reactive to DOC in the surface water than that

in the groundwater. Amine functional group on the aromaticstructure also increased its reactivity towards SO4�

- and CO3�-.

Therefore, the N-centered radical derived from SMX might couplewith another N-centered radical derived from DOC, and lead to thelower formation of TP 502 and TP 518 in the surface water. Theimpact of natural organic matter properties on TPs formation andfollowing toxicological potential need further investigation.

4. Conclusions

SO4�- showed the higher reactivity towards the anionic form of

SMX than the neutral form due to the electrophilic nature of SO4�-,

while �OH exhibited the comparable reactivity to SMX species. Thegeneration of the TPs in direct photolysis of SMX was derived fromcleavage of the NeS bond, rearrangement of the isoxazole ring, andhydroxylation of SMX. Although direct photolysis of SMX was thepredominant pathway at pH 3, the combination of �OH with directphotolysis enhanced the formation of hydroxylated products, whileSO4�

- further oxidized the TPs generated by direct photolysis. At pH8, �OH was confirmed to produce hydroxylated products, whileSO4�

- favored attack on eNH2 group, and thus generating the nitroderivative. Additionally, SO4�

- promoted the formation of oxazolering oxidized product rather than that of the isoxazole ring, while�OH reacted in the opposite pathway. More dimeric products weredetermined by SO4�

-, resulting from recombination of the N-centered radical.

The high concentration of bicarbonate scavenged �OH and SO4�-

to generate CO3�-, which was a more selective oxidant than �OH and

SO4�-. CO3�

- exhibited higher reactivity towards eNH2 group, andthus increasing the formation of the TPs related to the N-centeredradical, but inhibiting the formation of hydroxylated products onthe benzene ring. Both the acute toxicity experiments and ECOSARassessments suggested the TPs generated by SO4�

- or CO3�- possess

higher toxicity than those by �OH. Nevertheless, further studies areneeded to evaluate the impact of natural organic matter propertieson TPs formation and their toxicological potential when UV-basedoxidation processes are applied in different background waters.

Acknowledgments

This research was financially supported by the National NaturalScience Foundation of China (No. 51578203), the Funds of State KeyLaboratory of Urban Water Resource and Environment (HIT,2016DX13), and the Foundation for the Author of National ExcellentDoctoral Dissertation of China (201346). We are also thankful to theInternational Postdoctoral Exchange Fellowship Program (No.20160074) to Y.Y. and the Excellent Graduate Student Scholarshipfrom the Shanghai Tongji Gao Tingyao Environmental Science andTechnology Development Foundation awarded to X.L.

Appendix A. Supplementary data

Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.watres.2017.03.054.

References

Al Aukidy, M., Verlicchi, P., Jelic, A., Petrovic, M., Barcel�o, D., 2012. Monitoringrelease of pharmaceutical compounds: occurrence and environmental riskassessment of two WWTP effluents and their receiving bodies in the Po Valley,Italy. Sci. Total Environ. 438, 15e25.

Anastasio, C., Matthew, B.M., 2006. A chemical probe technique for the determi-nation of reactive halogen species in aqueous solution: Part 2-chloride solutionsand mixed bromide/chloride solutions. Atmos. Chem. Phys. 6, 2439e2451.

Baeza, C., Knappe, D.R.U., 2011. Transformation kinetics of biochemically activecompounds in low-pressure UV Photolysis and UV/H2O2 advanced oxidationprocesses. Water Res. 45 (15), 4531e4543.

Y. Yang et al. / Water Research 118 (2017) 196e207 207

Bonvin, F., Omlin, J., Rutler, R., Schweizer, W.B., Alaimo, P.J., Strathmann, T.J.,McNeill, K., Kohn, T., 2012. Direct photolysis of human metabolites of theantibiotic sulfamethoxazole: evidence for abiotic back-transformation. Environ.Sci. Technol. 47 (13), 6746e6755.

Bonvin, F., Rutler, R., Ch�evre, N., Halder, J., Kohn, T., 2011. Spatial and temporalpresence of a wastewater-derived micropollutant plume in lake Geneva. Envi-ron. Sci. Technol. 45 (11), 4702e4709.

Boreen, A., Arnold, W., McNeill, K., 2003. Photodegradation of pharmaceuticals inthe aquatic environment: a review. Aquat. Sci. 65 (4), 320e341.

Boreen, A.L., Arnold, W.A., McNeill, K., 2004. Photochemical fate of sulfa drugs in theaquatic environment: sulfa drugs containing five-membered heterocyclicgroups. Environ. Sci. Technol. 38 (14), 3933e3940.

Brown, M.A., De Vito, S.C., 1993. Predicting azo dye toxicity. Crit. Rev. Environ. Sci.Technol. 23 (3), 249e324.

Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., 1988. Critical review of rateconstants for reactions of hydrated electrons, hydrogen atoms and hydroxylradicals. J. Phys. Chem. Ref. Data 17 (2), 513e886.

Canonica, S., Kohn, T., Mac, M., Real, F.J., Wirz, J., von Gunten, U., 2005. Photosen-sitizer method to determine rate constants for the reaction of carbonate radicalwith organic compounds. Environ. Sci. Technol. 39 (23), 9182e9188.

Canonica, S., Meunier, L., von Gunten, U., 2008. Phototransformation of selectedpharmaceuticals during UV treatment of drinking water. Water Res. 42 (1e2),121e128.

Frimer, A.A., 1979. The reaction of singlet oxygen with olefins: the question ofmechanism. Chem. Rev. 79 (5), 359e387.

Gao, S., Zhao, Z., Xu, Y., Tian, J., Qi, H., Lin, W., Cui, F., 2014. Oxidation of sulfa-methoxazole (SMX) by chlorine, ozone and permanganateda comparativestudy. J. Hazard. Mater. 274, 258e269.

Grebel, J.E., Pignatello, J.J., Mitch, W.A., 2010. Effect of halide ions and carbonates onorganic contaminant degradation by hydroxyl radical-based advanced oxida-tion processes in saline waters. Environ. Sci. Technol. 44 (17), 6822e6828.

Gsponer, H.E., Previtali, C.M., García, N.A., 1987. Kinetics of the photosensitizedoxidation of polychlorophenols in alkaline aqueous solution. Toxicol. Environ.Chem. 16 (1), 23e37.

Hu, L., Flanders, P.M., Miller, P.L., Strathmann, T.J., 2007. Oxidation of sulfameth-oxazole and related antimicrobial agents by TiO2 photocatalysis. Water Res. 41(12), 2612e2626.

Ji, Y., Fan, Y., Liu, K., Kong, D., Lu, J., 2015. Thermo activated persulfate oxidation ofantibiotic sulfamethoxazole and structurally related compounds. Water Res. 87,1e9.

Keen, O.S., Linden, K.G., 2013. Degradation of antibiotic activity during UV/H2O2advanced oxidation and photolysis in wastewater effluent. Environ. Sci. Technol.47 (22), 13020e13030.

Kim, Y., Choi, K., Jung, J., Park, S., Kim, P.-G., Park, J., 2007. Aquatic toxicity of acet-aminophen, carbamazepine, cimetidine, diltiazem and six major sulfonamides,and their potential ecological risks in Korea. Environ. Int. 33 (3), 370e375.

Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber, L.B.,Buxton, H.T., 2002. Pharmaceuticals, hormones, and other organic wastewatercontaminants in U.S. streams, 1999�2000: a national reconnaissance. Environ.Sci. Technol. 36 (6), 1202e1211.

Larson, R.A., Weber, E.J., 1994. Reaction Mechanisms in Environmental OrganicChemistry. CRC press.

Lienert, J., Gudel, K., Escher, B.I., 2007. Screening method for ecotoxicological hazardassessment of 42 pharmaceuticals considering human metabolism and excre-tory routes. Environ. Sci. Technol. 41 (12), 4471e4478.

Lin, H., Niu, J., Xu, J., Li, Y., Pan, Y., 2013. Electrochemical mineralization of sulfa-methoxazole by Ti/SnO2-Sb/Ce-PbO2 anode: kinetics, reaction pathways, andenergy cost evolution. Electrochimica Acta 97, 167e174.

Mahdi Ahmed, M., Barbati, S., Doumenq, P., Chiron, S., 2012. Sulfate radical anionoxidation of diclofenac and sulfamethoxazole for water decontamination.Chem. Eng. J. 197, 440e447.

Majewsky, M., Wagner, D., Delay, M., Br€ase, S., Yargeau, V., Horn, H., 2014. Anti-bacterial Activity of Sulfamethoxazole Transformation Products (TPs): generalRelevance for Sulfonamide TPs Modified at the para Position. Chem. Res. Tox-icol. 27 (10), 1821e1828.

Morasch, B., Bonvin, F., Reiser, H., Grandjean, D., De Alencastro, L.F., Perazzolo, C.,Ch�evre, N., Kohn, T., 2010. Occurrence and fate of micropollutants in the VidyBay of Lake Geneva, Switzerland. Part II: micropollutant removal betweenwastewater and raw drinking water. Environ. Toxicol. Chem. 29 (8), 1658e1668.

Neta, P., Huie, R.E., Ross, A.B., 1988. Rate constants for reactions of inorganic radicalsin aqueous solution. J. Phys. Chem. Ref. Data 17 (3), 1027e1284.

Neta, P., Madhavan, V., Zemel, H., Fessenden, R.W., 1977. Rate constants andmechanism of reaction of sulfate radical anionwith aromatic compounds. J. Am.Chem. Soc. 99 (1), 163e164.

NIST, 2016. National Institute of Standards and Technology. NDRL/NIST SolutionKinetics Database on the Web. http://kinetics.nist.gov/solution/ (Accessed 13May 2016).

Padhye, L.P., Yao, H., Kung'u, F.T., Huang, C.-H., 2014. Year-long evaluation on theoccurrence and fate of pharmaceuticals, personal care products, and endocrinedisrupting chemicals in an urban drinking water treatment plant. Water Res. 51,266e276.

Rahn, R.O., Stefan, M.I., Bolton, J.R., Goren, E., Shaw, P.S., Lykke, K.R., 2003. Quantumyield of the iodide-iodate chemical actinometer: dependence on wavelengthand concentration. Photochem. Photobiol. 78 (2), 146e152.

Ratola, N., Cincinelli, A., Alves, A., Katsoyiannis, A., 2012. Occurrence of organicmicrocontaminants in the wastewater treatment process. A mini review.J. Hazard. Mater. 239e240, 1e18.

Rodgers, M.A.J., 1983. Solvent-induced deactivation of singlet oxygen: additivityrelationships in nonaromatic solvents. J. Am. Chem. Soc. 105 (20), 6201e6205.

Solar, S., Solar, W., Getoff, N., 1986. Resolved multisite OH-attack on aqueous anilinestudied by pulse radiolysis. Int. J. Radiat. Appl. Instrum. Part C. Radiat. Phys.Chem. 28 (2), 229e234.

Ternes, T.A., Joss, A., 2006. Human Pharmaceuticals, Hormones and Fragrances: theChallenge of Micropollutants in Urban Water. IWA Publishing, London.

Trov�o, A.G., Nogueira, R.F.P., Agüera, A., Sirtori, C., Fern�andez-Alba, A.R., 2009.Photodegradation of sulfamethoxazole in various aqueous media: persistence,toxicity and photoproducts assessment. Chemosphere 77 (10), 1292e1298.

Wols, B., Hofman-Caris, C., Harmsen, D., Beerendonk, E., 2013. Degradation of 40selected pharmaceuticals by UV/H 2 O 2. Water Res. 47 (15), 5876e5888.

Yang, Y., Jiang, J., Lu, X., Ma, J., Liu, Y., 2015. Production of sulfate radical and hy-droxyl radical by reaction of ozone with peroxymonosulfate: a novel advancedoxidation process. Environ. Sci. Technol. 49 (12), 7330e7339.

Yang, Y., Pignatello, J.J., Ma, J., Mitch, W.A., 2014. Comparison of halide impacts onthe efficiency of contaminant degradation by sulfate and hydroxyl radical-basedadvanced oxidation processes (AOPs). Environ. Sci. Technol. 48 (4), 2344e2351.

Yang, Y., Pignatello, J.J., Ma, J., Mitch, W.A., 2016. Effect of matrix components on UV/H2O2 and UV/S2O82� advanced oxidation processes for trace organic degra-dation in reverse osmosis brines from municipal wastewater reuse facilities.Water Res. 89, 192e200.

Zhang, R., Sun, P., Boyer, T.H., Zhao, L., Huang, C.-H., 2015. Degradation of phar-maceuticals and metabolite in synthetic human urine by UV, UV/H2O2, and UV/PDS. Environ. Sci. Technol. 49 (5), 3056e3066.

Zhang, R., Yang, Y., Huang, C.-H., Li, N., Liu, H., Zhao, L., Sun, P., 2016. UV/H2O2 andUV/PDS treatment of trimethoprim and sulfamethoxazole in synthetic humanurine: transformation products and toxicity. Environ. Sci. Technol. 50 (5),2573e2583.

Zhou, W., Moore, D.E., 1994. Photochemical decomposition of sulfamethoxazole. Int.J. Pharm. 110 (1), 55e63.

Zhou, W., Moore, D.E., 1997. Photosensitizing activity of the anti-bacterial drugssulfamethoxazole and trimethoprim. J. Photochem. Photobiol. B Biol. 39 (1),63e72.

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