Opposite pH-dependent roles of hydroxyl radicals in ozonation and UV photolysis of genistein

Science of the Total Environment 709 (2020) 136243

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Opposite pH-dependent roles of hydroxyl radicals in ozonation and UV photolysis of genistein Yang Huang a,b,c, Lihao Su a,b,d, Siyu Zhang a,⁎, Qing Zhao a, Xuejiao Zhang a, Xuehua Li b,d, Haibo Li c, Lifen Liu b,d, Jingwen Chen b,d, Xiaoxuan Wei e a

Key Laboratory of Pollution Ecology and Environmental Engineering, Institute of Applied Ecology, Chinese Academy of Sciences, Shenyang 110016, China Key Laboratory of Industrial Ecology and Environmental Engineering, School of Environmental Science and Technology, Dalian University of Technology, Dalian 116024, China School of Resources and Civil Engineering, Northeastern University, Shenyang 110004, China d Key Laboratory of Industrial Ecology and Environmental Engineering, Ministry of Education, School of Food and Environment, Dalian University of Technology, Panjin, 124221, China e College of Geography and Environmental Sciences, Zhejiang Normal University, Jinhua 321004, China b c

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• pH increases slowed ozonation but accelerated photodegradation of genistein. • Neutral and anionic genistein showed similar reactivity to O3. • ∙OH competed with genistein to consume O3 at alkaline pH. • Gs photodegraded fast at alkaline pH following direct and •OH sensitizing mechanism. • Ozonation and photolytic products were similar at low pH but different at high pH.

a r t i c l e

i n f o

Article history: Received 11 October 2019 Received in revised form 2 December 2019 Accepted 18 December 2019 Available online 20 December 2019 Editor: Yifeng Zhang Keywords: Phytoestrogens Genistein Ozonation Photodegradation pH

a b s t r a c t Phytoestrogens were frequently detected in municipal or industrial wastewater, and raised great attentions due to potential risks to humans or organisms. Until now, transformation mechanisms of phytoestrogens in advanced wastewater treatments were largely unknown. Here, pH influence mechanisms on transformations of phytoestrogens during two typical advanced wastewater treatments (ozonation and photolysis) were investigated, employing genistein (Gs) as a case. Removal efficiencies of Gs decreased significantly with increases of pH during ozonation, while photolytic rates increased by 44 or 200 times from pH 4.9 to 11.6 under irradiations without or with UVC. pH increases caused both dissociation of Gs and formation of hydroxyl radicals (•OH) in ozonation or photolysis, however, led to opposite changes to degradation rates. This was because that •OH played negatively as a competitor for O3 in ozonation, but acted as an accelerating species inducing self-sensitized photooxidation of Gs under UV light. Ozonation and photolytic products of Gs were similar at pH 4.9 or 8.6, but were totally different at pH 11.6. Most of the transformation products maintained isoflavone structures, and might possess phytoestrogenic effects. This study provided a deep insight into the pH influencing mechanism on typical advanced wastewater treatment processes of phytoestrogens.

⁎ Corresponding author at: Key Laboratory of Pollution Ecology and Environmental Engineering, Institute of Applied Ecology, Chinese Academy of Sciences, Shenyang 110016, China. E-mail address: [email protected] (S. Zhang).

https://doi.org/10.1016/j.scitotenv.2019.136243 0048-9697/© 2019 Elsevier B.V. All rights reserved.

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Y. Huang et al. / Science of the Total Environment 709 (2020) 136243

Main finding of the work: Opposite pH-dependent degradation mechanisms caused by hydroxyl radicals (•OH) were elucidated for ozonation and UV photolysis of phytoestrogens, taking genistein as a case. © 2019 Elsevier B.V. All rights reserved.

1. Introduction Phytoestrogens, a typical class of endocrine disrupting chemicals (EDCs), have attracted serious concerns due to potential ecological risks to aquatic organisms and humans (Dogan and Simsek, 2016; Jefferson et al., 2007; Omar et al., 2016; Sood et al., 2017; Yiannis et al., 2003). Phytoestrogens can cause a series of physiological impacts on aquatic organisms (Latonnelle et al., 2002; Thorpe et al., 2003), including decreasing the aggressive behavior, causing development of male-like physical characteristics to females, and competing for estrogen receptors, etc. (Ardia and Clotfelter, 2010; Clotfelter and Rodriguez, 2006; Zhang et al., 2002). Genistein (Gs) was the phytoestrogen most frequently presented in our daily diets (e.g. soy, peas, and beans), and widely used in pharmaceutical industries. Gs was observed in industrial wastewater at concentrations above a theoretical threshold (≥1000 ng/L) potentially inducing negative environmental impacts (Latonnelle et al., 2002; Thorpe et al., 2003). The concentration of Gs ranged from 98.9 to 108.0 μg/L in effluents of soy-processing industries (Kiparissis et al., 2001; Lundgren and Novak, 2010), and was even higher (13.1 mg/L) in discharges of pulp or paper mills (Kiparissis et al., 2001). Fish in rivers impacted by mill effluents showed startling developmental and reproductive impairments (Hewitt et al., 2010; Hewitt et al., 2006; Larsson et al., 2010; Parks et al., 2001; Sandström and Neuman, 2003). Thus, in order to improve the safety and quality of wastewater for discharging into environments or for reuses to alleviate the global water scarcities (Maier et al., 2016), it was crucial to develop effective approaches to remove phytoestrogens from wastewater. Traditional water treatment technologies exhibited low removal efficiencies to EDCs (Hewitt et al., 2010; Hewitt et al., 2006; Juliane et al., 2009; Kiparissis et al., 2001; Larsson et al., 2010; Maier et al., 2016; Parks et al., 2001; Sandström and Neuman, 2003). After biological treatments, only 41–53% of phytoestrogens can be removed (Balest et al., 2008; Han et al., 2011). Thus, a series of advanced water treatment technologies was developed in order to improve removal efficiencies of EDCs. Among them, ozonation and UV photolysis was found to be highly efficient (Lubomira et al., 2013), and was expected to reach ca. 93% decomposition rate to estrogens (Xu et al., 2006). However, most of organic pollutants were not completely mineralized in ozonation or UV photolysis, and produced diverse oxidized products (OPs) (Laera et al., 2012; Maier et al., 2016). Some OPs showed higher endocrine disruption potency or persistence than parent compounds (Laera et al., 2012; Maier et al., 2016). Thus, it was necessary to unveil OPs of phytoestrogens formed after ozonation or UV treatments before large-scale applications of these technologies. Generally, pH values of municipal wastewater were around 7.50 (Garcia-Ivars et al., 2017; Giannakopoulos et al., 2016), while that of soy-processing or the pulp/paper mill water containing phytoestrogens was 4.89 (Tu et al., 2019) or 6.83–11.0 (Asaithambi et al., 2017; Veluchamy and Kalamdhad, 2017). Treatment efficiencies of phytoestrogens by ozonation or UV photolysis may vary with pH. In one aspect, most phytoestrogens are polyhydroxy phenols and can dissociate in water. Anionic forms usually show higher reactivity in electrophilic reactions or absorb more long-wavelength UV lights than neutral forms (Zhang and Li, 2014). In another aspect, pH influences removals of EDCs not only by changing dissociation species, but also through other complex pathways. Previous studies showed that low pH favored the photolysis of estrone, 17β-estradiol, estriol, and 17αethinylestradiol (Zhang and Li, 2014), but Gs photodegraded quickly at high pH (Kelly and Arnold, 2012). The reason was not understood until now. The hydrolysis of O3 produced a stronger oxidant, i.e.

hydroxyl radical (•OH), in a yield increased at alkaline pH (Zhang and Li, 2014). Despite of this, removal efficiencies of estrone, 17β-estradiol and 17α-estradiol decreased from 65 to 95% to 15–55% when pH was increased from 6 to 12 during ozonation (Kelly and Arnold, 2012). The pH influence on ozonation or photolysis of phytoestrogens was not clear. Understanding underlying mechanisms was crucial to the development of suitable technologies for phytoestrogen-containing wastewaters of a wide pH range. Here, Gs was selected as a representative phytoestrogen to investigate pH dependence of ozonation or photolytic process. There were 3 hydroxyl functional groups in chemical structure of Gs, corresponding to 3 dissociation equilibrium constants (pKa1 ~ pKa3), and 4 dissociation species (Fig. 1). Reactivity of different dissociation states of Gs in ozoantion or photolysis was unveiled. Reactivity differences were theoretically elucidated by employing density functional theory (DFT) calculations. Roles of •OH in the two treatment processes were elucidated. OPs generated at different pH were identified. Transformation pathways of Gs at different pH were proposed to further elucidate the pH-dependent degradation mechanism. This work improved insight understandings to ozonation and UV photolytic mechanisms of phytoestrogens. 2. Materials and methods 2.1. Chemicals Gs (purity N98%) was purchased from TCI (Shanghai, China). Methanol (CH3OH) was supplied by J. T. Baker (America). Humic acid (HA) was from Sigma-Aldrich. Na2HPO4, NaH2PO4, HCl, and NaOH (analytical purity) were purchased from Tianjin Fuyu Fine Chemicals Co., Ltd. (China). Ultrapure water (18.2 MΩ∙cm) used in experiments was prepared by a Milli-Q Academic A10 system (Milli-Q, Millipore). 2.2. Ozonation and photolytic experiments Ozone was generated by an ozone generator (ENALY 1000BT-12, Shenzhen) with extra-dry and high-purity O2 as a source. Ozone stock solutions were freshly prepared before experiments by purging an ozone/oxygen gas mixture into ice-cooled ultrapure water to a concentration of 3.5 mmol/L measured by a UV spectrophotometer (UV-2700, Shimadzu, Japan). Ozonation experiments were performed in 100 mL conical flasks. Gs stock solution was diluted with ultrapure water to obtain an initial experimental concentration (c0) of 5 μmol/L. Then, the ozone stock solutions were added into Gs solutions to reach different O3:Gs ratios. The flasks were sealed immediately, mixed gently to start the reactions, and then placed statically in dark for over 24 h to ensure that the reactions completed finished. Photolytic experiments were performed in a photochemical reactor (XPA-7, Xujiang Electromechanical Plant, Nanjing, China) equipped with a low temperature thermostatic bath (YRDC-2020, Yarong Biochemical Instrument Factory, Shanghai, China). A 300 W mercury lamp was employed as a light source. Light intensity of the lamp filtered by Pyrex glasses (Fig. S1) was measured by a TriOS spectroradiometer (TriOS GmbH, Germany). UV–visible absorption spectra of Gs at different pH were measured with a Hitachi U-2900 spectrophotometer (Fig. S2). Quantum yield (Φ) was determined according to methods reported by Leal et al., 2013, and calculated following reaction (1): k½Gs  Φ¼P  Iλ 1−10−ελ ½Gsl  Δλ

ð1Þ

Y. Huang et al. / Science of the Total Environment 709 (2020) 136243

Gs0

pKa1 = 7.2

Gs-

3

pKa2 = 10.0 2- pKa3 = 13.1 3Gs Gs

Percentage

1.0 0.8

0

Gs Gs 2Gs 3Gs

0.6 0.4 0.2 0.0

0

2

4

C15H10O5 (Gs) At pH 4.9, Gs0 = 99.5%, pH 8.6, Gs- = 92.6%, 6

8

10

12

14

pH 11.6, Gs2- = 94.6%

pH Fig. 1. Dissociation species distributions of Gs at pH 1–14 (Gs0: neutral form; Gs−: monovalent anion; Gs2−: divalent anion; Gs3−: trivalent anion. Dissociation sequence of hydroxyl functional groups was 7-OH, 4′-OH, and 5-OH (Zielonka et al., 2003)).

where k (s−1) was the first-order photolytic rate constant of Gs; λ (nm−1) was wavelength; [Gs] (mol·L−1) was the initial concentration of Gs; Iλ (einsteins·m−2·sec−1·nm−1) was the light intensity at centers of Pyrex tubes; Δλ standed for wavelength intervals of the spectral irradiance; ελ (L·mol−1·cm−1) was molar absorptivity of Gs at λ; l (cm) was the calculated light path length (Zhang et al., 2018). Quartz or Pyrex tubes were used as containers to mimic different light conditions, i.e. with or without UVC (λ b 290 nm) irradiations, respectively. No remarkable loss of Gs was observed in the dark. Phosphate buffer solutions were used to maintain constant pH conditions (4.9, 8.6, or 11.6) in the experiments. The pH values were selected according to pKa (7.2, 10.0, and 11.6 (Zielonka et al., 2003)) of Gs, to ensure only one dominant dissociation species (N 90%) at each pH (Fig. 1). HA (5 mg/L) were employed as a source of photoreactive species. To explore photodegradation mechanisms of Gs in HA, 25% (v: v) methanol or 3 mmol/L sodium azide (NaN3) was used as a quencher of •OH or a quencher of both •OH and singlet oxygen (1O2) (Zhou et al., 2015; Li et al., 2015). Photogeneration of •OH from Gs was verified by using benzene (3 mmol/L) as a probe, by detecting phenols formed through •OH oxidizing benzene (Dong and Rosario-Ortiz, 2012). At each set time interval, a small aliquot of the samples was withdrawn and stored at 4 °C before analysis. All experiments were performed in triplicate, and data reported were average values and standard errors.

OPs were analyzed using a XTERRA MS C18 (2.1 × 100 mm, 3.5 μm) chromatographic column by a high performance liquid chromatography-tandem mass spectrometry (RRLC/6410B, QQQ). Column temperature was 25 °C. Flow rate was 0.25 mL/min, and injection volume was 2 μL. Mobile phase was methanol and water. Samples were eluted by 20% water for 8.0 min, followed by 40% water until 10.0 min, and then by 80% water at 10.1 to 18.0 min. A full scan (60–300 m/z) was used in a negative electrospray ionization (ESI−) mode.

2.3. Analytical methods

3. Results and discussion

Gs was analyzed by a high performance liquid chromatography (Agilent 1260 Infinity II), equipped with a quaternary pump, an autosampler and a UV detector. A reverse-phase Wonda Sil C18 (150 × 4.6 mm, 5 μm) was used for separation. Column temperature was 25 °C. Injection volume was 10 μL. Mobile phase was methanol: water (60:40) and eluted at 1 mL/min. UV detector was operated at 256 nm. A Zorbax Eclipse XDB-C18 column (3.0 × 150 mm, 3.5 μm) was employed for the quantification of phenols. Column temperature was 25 °C. Injection volume was 40 μL. Mobile phase was methanol: water (45:55) and eluted at 0.4 mL/min. UV detector was operated at 269 nm. To identify OPs, samples were extracted by a solid phase extraction (SPE) method. SPE columns (Waters Oasis HLB column, 6 mL, 500 mg) were preconditioned with 5 mL methanol and 10 mL water in sequence, and then were loaded with samples. After the samples flowed through gradually by gravity, the columns were washed with ultrapure water, dried in vacuum for 30 min and then eluted with 8 mL methanol. Eluents were concentrated by a rotary evaporator, and transferred to glass vials. Final volume of the eluents was adjusted to 1.0 mL with methanol.

3.1. pH-dependence of ozonation of Gs

2.4. Calculation methods All calculations were performed using Gaussian 09 (Gaussian, Inc., USA) software package (Frisch et al., 2009). Geometry optimizations and frequency calculations were performed based on density functional theory (DFT) at a B3LYP/6–31 + G(d,p) level. Optimized geometries of Gs2−, Gs−, and Gs0 without any imaginary frequencies were shown in Fig. S3 and Table S3. Reactivity of different species of Gs towards O3 was compared based on the highest occupied molecular orbital (HOMO) and lowest unoccupied molecular orbital (LUMO) energy values (Asghar et al., 2019; Ge et al., 2019). HOMO and LUMO of Gs were calculated at a B3LYP/6–311++g(d,p) level. Solvent effects were mimicked by employing an integral equation formalism of the polarized continuum model (IEFPCM) (Ottonello and Zuccolini, 2008).

Removal efficiencies of Gs varied within pH 4.9–11.6 and were almost linearly related with O3 doses at each pH (Fig. 2a). High removal efficiencies were obtained at low pH. The removal efficiency of Gs at pH 4.9 or neutral was higher than that at pH 8.6 or 11.6 under the same dose of O3. A 100% removal was achieved at O3:Gs of 50:1, 70:1, 400:1, and 500:1 at pH 4.9, neutral, 8.6 and 11.6, respectively. Thus, the removal efficiency of Gs at pH 4.9 or neutral was 6–10 times of that at pH 8.6 or 11.6. Generally, ozone preferred to attack the HOMO of organic pollutants following an electrophilic pattern (Minju and Zimmermann-Steffens, 2015). Reactivity of pollutants to O3 was dependent on HOMO energies (Minju and Zimmermann-Steffens, 2015). As inferred from the HOMO energies of Gs (Fig. 3), the reactivity to O3 would follow an order: Gs2−N Gs− N Gs0. Theoretically, the removal of Gs should be faster at neutral pH (close to pKa1) than at pH 4.9 (Gs0 N 99%, Fig. 1), as 50% of Gs0 dissociated to form the more reactive Gs− at pKa1. Interestingly, removal efficiencies of Gs were close at pH 4.9 and at neutral. The most probable reason was that the reactivity of Gs0 and Gs− to ozone was close, as

4

Y. Huang et al. / Science of the Total Environment 709 (2020) 136243

(a)

75 50

pH 4.9 No pH adjustment pH 8.6 pH 11.6 200 300 400 500 600 O3:Gs

25 0

0

100

100 Removal (%)

Removal (%)

100

No CH3OH

CH3OH (b)

75 50 25 0

pH 8.6

pH 11.6

Fig. 2. Removals of Gs at different O3:Gs ratios (a), and influence of CH3OH at pH 8.6 (O3:Gs = 60:1) and 11.6 (O3:Gs = 200:1) (b).

phenols usually showed high reactivity to ozone even in neutral forms (Lee and Gunten, 2010). By adding CH3OH to scavenge •OH (Liu, 2011), removals of Gs were decreased by up to 21% at pH 11.6 (Fig. 2b). Roles of •OH were less significant at pH 8.6 compared with that at pH 11.6, because hydrolysis of ozone to form •OH was decreased exponentially with pH decreases (Kelly and Arnold, 2012; Vandersmissen et al., 2008), and more Gs was oxidized by •OH at pH 11.6 than at pH 8.6. The effect of CH3OH was negligible at pH 4.9, because hydrolysis of ozone to form •OH was slow in acidic solutions, due to scavenges of protons (Zhang and Li, 2014). Although the oxidation potential of •OH (2.80 eV) (Ku et al., 1996) was a bit higher than that of ozone (2.07 eV) (Harris, 1999), removals of Gs were not enhanced in alkaline solutions containing more •OH. On the contrary, Gs degraded faster at neutral or pH 4.9, where formation of •OH was negligible. This was mainly attributed to a high reaction rate (7.25 × 1011 L mol−1 s−1 (Atkinson et al., 2004)) of •OH and O3, which was higher than reaction rates of most phenols with •OH (0.1 × 109–20 × 109 L mol−1 s−1 (Wang et al., 2009)). Consumptions of •OH by O3 resulted in great decreases of the two oxidants at alkaline pH, and then removal efficiencies of Gs were largely decreased. This phenomenon was opposite to typical EDCs, such as estrone, 17βestradiol and 17α-estradiol, of which ozonation was usually enhanced by pH increases (Kelly and Arnold, 2012), because these pollutants were almost inert to O3, and can only be oxidized by •OH. 3.2. pH-dependence of photodegradation of Gs Effects of pH on photodegradation of Gs were contrary to that observed in ozonation. Photolytic rate constants (kobs) of Gs were increased sixfold from pH 4.9 to 8.6, and sharply increased by almost 44 times at pH 11.6 (Fig. 4a). In quartz tubes, rates increased more significantly by over 200 times from pH 4.9 to 11.6 due to the presence of UVC

0.0 0

Energy (a.u.)

Gs

-

Gs

2-

Gs

-0.1 O3

-0.2 -0.3

HOMO LUMO

-0.4 Fig. 3. Frontier molecular orbital energies of Gs. HOMO: the highest occupied molecular orbital; LUMO: the lowest unoccupied molecular orbital.

(Fig. 4b–d). Apparent photolytic quantum yields (Φ) were determined as (8.03 ± 2.78) × 10−7, (4.70 ± 0.011) × 10−6, and (3.10 ± 0.202) × 10−5 at pH 4.9, 8.6 and 11.6 (Table 1), respectively. The cumulative light absorption of Gs were 6.93 × 10−6, 8.85 × 10−6 and 9.48 × 10−6 Einstein·cm−3·s−1 at pH 4.9, 8.6 and 11.6, respectively. Thus, the photolytic rate increases at alkaline pH were mainly attributed to the increases of Φ with pH. According to the inhibition effects of methanol or NaN3 (Fig. 4a), •OH was proposed to be involved in the photodegradation of Gs especially at alkaline pH. kobs was decreased by 39% and 42% at pH 8.6 and 11.6, respectively, after adding methanol to scavenge •OH (Table 1). Similar rate decreases were observed after adding NaN3, a quencher of both •OH and 1O2 (40% and 41%, Table 1). The photolytic rates of Gs were similar in the presence of methanol and NaN3, implying that the role of 1O2 was negligible. The kobs values observed in the presence of methanol were used as direct photolytic rate constants, and to calculate contributions of direct photolysis. Thus, proportions of direct photolysis were 61% and 58% at pH 8.6 and 11.6, respectively. Quantum yields of direct photolysis were (3.13 ± 0.543) × 10−6 and (1.67 ± 0.175) × 10−5 at pH 8.6 and 11.6, respectively (Table 1). This meant that anionic Gs (Gs− and Gs2−) produced •OH under irradiation and photodegraded through direct photolysis combined with •OH induced selfsensitization process, which contributed to approximate 40% of the total photodegradation. Interestingly, kobs was increased by 43% after adding methanol at pH 4.9. This completely different effect of methanol at acidic pH was also observed in experiments performed in quartz tubes. This indicated that •OH induced self-sensitization was less important to Gs0. Otherwise, kobs of Gs should be largely decreased by adding methanol at pH 4.9. The observed rate increase was likely due to an increase of pH (0.6 unit) because of adding methanol. As photolytic rates of Gs were greatly dependent on pH, a slight pH increase was expected to lead to significant accelerations to the photodegradation. To further confirm •OH was photoproduced by Gs, benzene was added into the photolytic solutions as a probe. After irradiated for 6 h, phenols were observed at the three pH values. HPLC peak areas of phenols were 24.1 ± 0.65 (pH 4.9), 173.6 ± 2.01 (pH 8.6), and 162.0 ± 1.30 (pH 11.6). Assuming reaction rates of benzenes with •OH to form phenols were the same at each pH, the concentrations of photogenerated •OH followed pH 8.6or 11.6 N pH 4.9. This result not only verified the formation of •OH during Gs photolysis, but also indicated that the formation yields of •OH at pH 4.9 was much lower than that at pH 8.6 or 11.6. To mimic natural water, HA was added to increase total organic carbon (TOC) of photolytic solutions. HA accelerated the photodegradation of Gs significantly (Fig. 4a and Table 1). This acceleration could be partly attributed to photooxidation induced by •OH photogenerated by HA, as indicated by a negative effect of methanol. Besides, triplet excited states of HA (3HA⁎) were inferred to be responsible for inducing indirect photodegradation of Gs (Chen et al., 2013). The effect of TOC in photodegradation of Gs was different from that in ozonation, where TOC usually caused a reduction in degradation rates by consuming

Y. Huang et al. / Science of the Total Environment 709 (2020) 136243

0.7

0.5

kobs (h-1)

(a)

4.9 8.6 11.6

0.6

5

0.4 0.3 0.2 0.1 0.0 Pyrex

CH3OH

NaN3

HA

HA+CH3OH

Fig. 4. Photolytic rate constants (kobs, h−1) of Gs at pH 4.9, 8.6 and 11.6 in the presence of CH3OH, NaN3 or HA in Pyrex tubes (a); Photolytic kinetic curves of Gs in Pyrex or quartz tubes at pH 4.9 (b), 8.6 (c) and 11.6 (d). The error bars represent 95% confidence interval, n = 3.

ozone or •OH (Lim et al., 2015). It was worth to notice that kobs of Gs in the presence of HA decreased by 60% after adding methanol at pH 4.9, but by 46–55% at pH 8.6 or 11.6. This implied that the reactivity of Gs0 to •OH was not much lower than that of anionic Gs. Then, the low involvement of •OH in direct photolysis of Gs at pH 4.9 (Fig. 4a, b) could only be attributed to a lower yield of •OH photogenerated by Gs0, in accordance with the low formation yields of •OH indicated by benzene experiments at pH 4.9. 3.3. Ozonation and photolytic products of Gs at different pH Transformation products of Gs during ozonation or photodegradation were identified by HPLC-MS. According to total ion chromatogram (TIC) and extraction ion chromatograms (XICs), a product of m/z 285 (P285) was generated in either ozonation (Fig. S4, S5) or photodegradation (Figs. S9–S11) at pH 4.9 or 8.6 (Table S1). Based on m/ z (Fig. S6), P285 was a hydroxylation product of Gs (Fig. 5). Besides, a product of m/z 187 (P187) was formed in ozonation but only observed at pH 4.9 (Table S1). As no •OH was generated at pH 4.9 in ozonation, P187 was generated through a direct reaction between Gs and O3 (Fig. 5) probably following Criegee addition mechanism (Zhang et al., 2017). Transformation products detected at pH 11.6 were completely different from those at pH 8.6 or 4.9. Ozonation products included P242

and P299 at pH 11.6 (Fig. S7, S8). As ozone was fast hydrolyzed or consumed by HO· at alkaline pH, these products were likely generated through HO· reacting with Gs−. Neither P242 nor P299 was detected in photolysis, only one photolytic product was detected at pH 11.6, i.e. P271 (Fig. S12, S13), indicating that photoreduction occurred to Gs, which existed mainly as Gs2− (Fig. 1). It was interesting that no common transformation products were detected at pH 11.6 unlike that at pH 8.6, though HO· was involved in both ozonation and photodegradation of Gs. This was most likely because only products of moderate to high molecular weights were concerned in this study. These products kept an isoflavone structure, and might possess phytoestrogenic effects. 4. Conclusion In this study, effects of pH on ozonation and photolytic kinetics or products of a representative phytoestrogen Gs were investigated. The influence of pH on the two processes was opposite. Increasing pH decreased ozonation rates of Gs, but accelerated photodegradation. Although •OH was generated in both systems, but showed different significance. In ozonation, •OH generated at alkaline pH resulted in a fast consumption of ozone, and slowed down the decomposition of Gs. In photodegradation, photogenerated •OH initiated selfsensitization of Gs and became an important accelerator at alkaline

Table 1 Photolytic rate constants (kobs, h−1) and apparent photolytic quantum yields (Φ) of Gs under different conditions (in Pyrex tubes) Φ

kobs Photolytic conditions Gs Gs + CH3OH Gs + NaN3 Gs + HA Gs + CH3OH + HA a

Not determined.

pH 4.9

pH 8.6 -3

(4.60 ± 2.00) × 10 (6.60 ± 1.00) × 10-3 (6.33 ± 1.00) × 10-3 (7.47 ± 0.500) × 10-2 (2.96 ± 0.4) × 10-2

pH 11.6 -2

(3.21 ± 1.00) × 10 (1.97 ± 0.200) × 10-2 (1.94 ± 0.300) × 10-2 0.127 ± 0.012 (6.85 ± 0.3) × 10-2

0.204 ± 0.012 0.119 ± 0.012 0.121 ± 0.010 0.607 ± 0.067 0.271 ± 0.015

pH 4.9

pH 8.6 -7

(8.03 ± 2.78) ×10 (1.16 ± 0.700) × 10-6 (1.47 ± 0.297) × 10-6 -a -a

pH 11.6 -6

(4.70 ±0.011) × 10 (3.13 ± 0.543) × 10-6 (5.25 ± 3.02) × 10-6 -a -a

(3.10 ±0.202) × 10-5 (1.67 ± 0.175) × 10-5 (3.39 ± 0.152) × 10-5 -a -a

6

Y. Huang et al. / Science of the Total Environment 709 (2020) 136243

H

Gs (m/z = 269) pH 4.9 (99.5% Gs 0 : 0.5% Gs -) pH 11.6 (2.4% Gs -: 94.6% Gs 2- : 3.0% Gs 3-) O3 /HO·O3 pH 8.6 (3.7% Gs 0 : 92.6% Gs -: 3.7% Gs 2-)

+ O3 + O3 or + hν

+ hν/HO·pg

+ O3 or + hν/HO·pg

P187 (C11 H8 O3 )

P242 (C14 H9 O4 )

or

P299 (C15 H8 O7 )

P271 (C15 H8 O5 )

or

or

P285 (C15 H10 O6 )

Fig. 5. Ozonation or photolytic pathways and products of Gs at pH 4.9, 8.6 and 11.6 (Products circled by blue dash line were generated in ozonation; Products circled by black dash line were generated in photolysis; P285 was detected in either process; hν: directly photolyzed by absorbing photons; O3/HO·O3: degraded by reacting with O3 or HO· generated from hydrolysis of O3; hν/HO·pg: degraded by absorbing photons or photogenerated HO·). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

pH. Degradation products of Gs in the two processes were similar at pH 4.9 or 8.6, but completely different at pH 11.6. This study indicated opposite influencing mechanisms of pH on ozonation and photodegradation of Gs, suggesting different treatment strategies for phytoestrogen containing wastewater of varied pH. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgement This work was supported by the National Natural Science Foundation of China (21607165), the National Key Research and Development Program of China (2016YFD0800305). The authors also acknowledge Youth Innovation Promotion Association CAS awarded to S.-Y. Z. (2017–2020), CAS Pioneer Hundred Talents Program awarded to X.-J. Z. (2015–2020), Hundreds Talents Program of CAS awarded to Q. Z. (2014–2019), Shenyang Pioneer Scientific and Technological Innovation Talents Team (RC170020) and Zhejiang Provincial Natural Science Foundation of China (LQ18B070003). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.136243. References Ardia, D.R., Clotfelter, E.D., 2010. The novel application of an immunological technique reveals the immunosuppressive effect of phytoestrogens in Betta splendens. J. Fish Biol. 68, 144–149. Asaithambi, P., Aziz, A.R.A., Sajjadi, B., Wan, M.A.B.W.D., 2017. Sono assisted electrocoagulation process for the removal of pollutant from pulp and paper industry effluent. Environmental Science & Pollution Research International 24, 5168–5178.

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