Degradation of propranolol by UV-activated persulfate oxidation: Reaction kinetics, mechanisms, reactive sites, transformation pathways and Gaussian calculation

Science of the Total Environment 690 (2019) 878–890

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Degradation of propranolol by UV-activated persulfate oxidation: Reaction kinetics, mechanisms, reactive sites, transformation pathways and Gaussian calculation Tiansheng Chen a, Jingshuai Ma a, Qianxin Zhang a, Zhijie Xie a, Yongqin Zeng a, Ruobai Li a, Haijin Liu b, Yang Liu c, Wenying Lv a,⁎, Guoguang Liu a,⁎ a b c

School of Environmental Science and Engineering, Guangdong University of Technology, Guangzhou 510006, China School of Environment, Henan Normal University, Key Laboratory for Yellow River and Huaihe River Water Environment and Pollution Control, Xinxiang 453007, China Faculty of Environmental & Biological Engineering, Guangdong University of Petrochemical Technology, Maoming 525000, China

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

• Efficient oxidation of propranolol could be achieved by UV-activated persulfate. • UV/PS process was much more effective than the UV/H2O2 process in degrading propranolol. • Sulfate radical was identified as the primary oxidizing species. • Gaussian calculations were applied to the intermediate product path speculation and the mechanism was further analyzed. • Studies on the transformation mechanisms revealed three different reaction pathways.

a r t i c l e

i n f o

Article history: Received 26 April 2019 Received in revised form 21 June 2019 Accepted 3 July 2019 Available online 06 July 2019 Editor: Paola Verlicchi Keywords: Propranolol UV-activation persulfate Reaction kinetics Molecular orbital calculations Transformation products

a b s t r a c t Contamination with β-blockers such as propranolol (PRO) poses a potential threat to human health and ecological system. The present study investigated the kinetics and mechanisms of PRO degradation by UV-activated persulfate (UV/PS) oxidation. Here, the experimental results showed that the degradation of PRO followed pseudo-first-order reaction kinetics, the degradation rate constant (kobs) was increased dramatically with increasing PS dosage or decreasing initial PRO concentration. And increasing the initial solution pH could also enhance the degradation efficiency of PRO. Radical scavenging experiments demonstrated that the main radical species was sulfate radicals (SO•− 4 ), with hydroxyl radicals (HO·) playing a less important role. Meanwhile, the and HO· were determined to be 1.94 second-order rate constants of PRO degradation with SO•− 4 × 1010 M−1 s−1 and 6.77 × 109 M−1 s−1, respectively. In addition, the presence of natural organic matter (NOM) and nitrate anion (NO− 3 ) showed inhibitory effect on PRO degradation, whereas bicarbonate anion − (HCO− 3 ) and chlorine anion (Cl ) greatly enhanced the degradation of PRO. Moreover, the transformation products of PRO were identified by applying ultra performance liquid chromatography coupled with quadrupole time-of-flight mass spectrometry (UPLC/Q-TOF-MS) technique. Molecular orbital calculations were used to estimate the reaction site of PRO with radicals, simultaneously. Hence, the transformation pathways including

⁎ Corresponding authors at: School of Environmental Science and Engineering, Guangdong University of Technology, No. 100 Waihuan Xi Road, Guangzhou Higher Education Mega Center, Guangzhou 510006, China. E-mail addresses: [email protected] (W. Lv), [email protected] (G. Liu).

https://doi.org/10.1016/j.scitotenv.2019.07.034 0048-9697/© 2019 Elsevier B.V. All rights reserved.

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hydroxylation, dehydration, naphthalene ring opening, and the cleavage of aldehyde groups were proposed. This work enriches the mechanism of PRO degradation under UV/PS system on the basis of results obtained by experimental characterization and Gaussian theoretical calculation. © 2019 Elsevier B.V. All rights reserved.

1. Introduction The discharge of emerging contaminants such as pharmaceuticals into the aquatic environment has increased the concern of humans in recent decades (Kümmerer, 2009). The sources of these pharmaceuticals in surface waters include hospitals, pharmaceutical industries, and domestic wastewater containing human excreted or unused medicines (Deblonde et al., 2011). β-blockers are β-adrenergic receptor antagonists, and comprise a vitally significant class of pharmaceuticals, which are usually employed for the treatment of hypertension, angina and cardiac dysfunctions (Alder et al., 2010; Maurer et al., 2007). The occurrence of β-blockers has been reported in surface waters and in sewage treatment plant effluents across the United States (Conkle et al., 2008), in Germany (Sacher et al., 2001), and in Canada (Brun et al., 2006). In general, β-blockers cannot be completely decomposed during traditional sewage treatment due to their recalcitrant properties (Gabet-Giraud et al., 2010). Propranolol (PRO, Fig. S1 shows the chemical structure) is a sympatholytic non-selective β-blocker that is used to treat cardiovascular diseases (Khetan and Collins, 2007). This pharmaceutical has been found in the aquatic environment, including wastewater (Fernández et al., 2010). Several studies have also revealed the potential toxic effects of this compound to fish, green algae, and crustaceans (Robinson et al., 2007; Rivera-Utrilla et al., 2013). It has been clearly established that traditional sewage treatment methods do not have the capacity to remove such a emerging pollutant (Maurer et al., 2007). Therefore, it is of great importance to exploit new and efficient treatment technologies to removal this contaminant from aquatic environment. Advanced oxidation processes (AOPs) form a promising technology for the treatment of wastewater that contains pharmaceuticals by generating high-efficiency reactive species (Wols and Hofman-Caris, 2012; Khan et al., 2013). Traditional hydroxyl radical (HO·)-induced AOPs, such as simulated solar irradiated TiO2 (Santiago-Morales et al., 2013), UV/H2O2 (Andreozzi et al., 2004), UV/O3 (Kim et al., 2009), and O3/ H2O2 (Kim et al., 2009), have been frequently used to degrade and mineralize PRO. Nevertheless, the degradation efficiency of HO· is easily impacted by the presence of natural organic matter (NOM) and inorganic anions (He et al., 2013). More recently, sulfate radical (SO•− 4 )-induced AOPs considered as effective treatment methods for the decomposition of persistent organic contaminants. SO•− 4 possesses a similar standard redox potential (E0 = 2.5–3.1 V) than that of HO· (E0 = 1.9–2.7 V), thus it also has excellent ability to degrade organic matter. Not only can degrade most organic pollutants, but also degrade some organic substances that are difficult to degrade by hydroxyl radicals. (Buxton et al., 1988; Hori et al., 2005; Lee et al., 2010). However, there is lack of more in-depth research on the degradation mechanism and degradation pathway of propranolol. Due to its relatively high water solubility and stability, as well as moderate cost, persulfate (PS) is becoming the commonly used precur•− sor of SO•− 4 (Khan et al., 2014). Usually, SO4 is typically produced from the activation of PS by heat, transition metals, ultraviolet light, bases, and sonolysis (Tsitonaki et al., 2010). Because the radical quantum yield of PS under UV irradiation (λ = 254 nm) is much higher than that of H2O2 (Mark et al., 1990; Baxendale and Wilson, 1957), the UVactivated PS oxidation process has gained more and more attention and is proved to be a high-efficiency treatment technology for the degradation of all kinds of refractory organic contaminants, such as pesticides (Rasoulifard et al., 2015), pharmaceuticals (Zhang et al., 2015), and cyanotoxins (He et al., 2014). Compared with the report of Gao

et al. who only studied the UV/PS system for drug degradation and some factors affecting system degradation efficiency.(Gao et al., 2018), our study compared the effect of UV-activated H2O2 on the degradation of propranolol. At the same time, Gaussian calculation is used to deeply analyze the reaction sites and the choice of active free radical attack, enriching the degradation mechanism of propranolol. The underlying mechanism on the degradation of organics by UV-activated persulfate was also fully discussed, which is of great importance to better comprehending the key reactions in different AOP systems. In this study, we employed UV/PS oxidation system to remove PRO in an aqueous solution, also compared this results with UV/H2O2 system. Specifically, detailed studies on kinetics were conducted to better comprehend the effects of factors such as PS dosage, initial PRO concentration, initial pH, and natural water composition. Finally, the transformation products were identified via UPLC/Q-TOF-MS technique, as well as theoretical calculations frontier electron densities (FED) which can infer the degradation mechanism and degradation path of propranolol more deeply and accurately. The objectives of this work were to: (1) identify if SO•− 4 was the main radical species responsible for PRO removal through radical scavenging tests; (2) determine the secondorder rate constants for the reaction of PRO with SO•− 4 and HO·; (3) research the effects of pH, of the presence of Cl− and HCO− 3 , of the presence of natural water NOM and NO− 3 on the degradation of PRO; (4) explore the mineralization of PRO and PS consumption; (5) propose the detailed transformation mechanisms and pathways for PRO degradation in UV/PS oxidation process. 2. Materials and methods 2.1. Chemicals and materials Propranolol hydrochloride (99%) was purchased from Sigma (USA). HPLC-grade acetonitrile was purchased from U.S. ACS Enke chemistry Co. Ltd. (Guangzhou, China). Benzoic acid (BA) was obtained from Aladdin reagent Co. Ltd. (Shanghai, China). Sodium persulfate was purchased from Damao Chemical Reagent Co. Ltd. (Tianjin, China). Sodium thiosulfate, tertiary butanol, ethanol, and hydrogen peroxide were obtained from Zhiyuan Chemical Reagent Co. Ltd. (Tianjin, China). All chemicals above were of analytical grade and required without further purification. All aqueous solutions were freshly prepared with ultrapure water from a water purification system (TKA, Germany). 2.2. Experimental procedures UV irradiation experiments were conducted within an XPA-7 photochemical reactor (Nanjing XUJ Co. Ltd.), as shown in Fig. S2. A 5 W lowpressure mercury lamp emitting 254 nm UV light was used as the light source to provide the UV irradiation and the irradiation strength on reaction solution was measured to be 0.285 mW·cm−2 in average with a full-spectrum type optical power meter (CEL-NP2000, Beijing China Education Aulight Co., Ltd.). A 25 mL solution was placed into a 50 mL quartz tube with a stopper, which was subsequently introduced into the photochemical reactor for irradiation. The reaction temperature was maintained at 20 ± 2 °C by constant temperature liquidcirculating apparatus. At certain time intervals, the quartz tube was removed. And then, 2-mL aliquots were withdrawn and quenched immediately with an equivalent volume of 0.1 M sodium thiosulfate. The concentration of PRO was then immediately analyzed using HPLC. All

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of the experiments were conducted at least in triplicate, and the results were averaged. Error bars represent the standard deviation of the mean. 2.3. Identification of radical species In order to identify the main radical species that were mainly responsible for the degradation of PRO. Two radical scavengers (ethanol (EtOH) and tertiary butanol (TBA)) were added into systems to identify the species of principle oxidizing radicals, and to explore the underlying degradation mechanisms in the UV/PS systems. The results of the individual and overall contribution of SO•− 4 and HO· at 0 to 600 mM dosages of EtOH or TBA at pH 7.0, respectively. According to the distinct reaction rate constants, the TBA could consume HO· prior to the organic pollutants while methanol could scavenge HO· and SO•− 4 effectively.

1. For a nucleophilic reaction, where the reaction takes place the highest density of two electrons (2FED2LUMO) when they are in the LUMO at ground state. 2. For an electrophilic reaction, where the reaction takes place the highest density of two electrons (2FED2HOMO) when they are in the HOMO at ground state. 3. For a radical reaction, where the reaction takes place the highest density of the sum of each electrons (FED2LUMO + 2FED2HOMO) when they are in the LUMO and HOM, respectively. On the basis of the computation of Frontier Orbital Theory, HO· typically attacks the positions with higher FED2LUMO + 2FED2HOMO values. Besides, the sites with a more positive point charge were more easily attacked by O•− 2 (Ji et al., 2013; Wahab et al., 2009). 2.8. Analytical method

2.4. Determination of the second-order rate constants The second-order rate constants of PRO reacting with SO•− 4 and HO· were measured by competition kinetic methods, in which BA was selected as a reference compound. The pH of reaction solution was maintained at 7.0 through the use of a 10 mM phosphate buffer. TBA (100 mM) was introduced into the reaction solution that eliminate HO· in UV/PS oxidation process when determining the second-order rate constant of PRO with SO•− 4 . Experiments were also performed in UV/H2O2 process to detect the second-order rate constant of PRO with HO·. 2.5. Investigation of other influencing factors The solution pH was adjusted by using 0.1 M H2SO4 or NaOH, to a desired value to which PS was subsequently added. The pH of the solutions was measured using a pHS-3C pH meter (Leici., China) equipped with a – pH electrode. The inorganic anion effect of Cl−, NO− 3 , and HCO3 (from NaCl, NaNO3 and NaHCO3, respectively) at 1 mM level on the destruction of PRO was studied. Different dosages of NOM (0–10 mg/L) were added in the samples to elucidate their effects on the degradation of PRO. This study used humic acid (HA) as a typical NOM. 2.6. Determination of PS consumption The residual concentration of PS was determined spectrophotometrically with the iodometric titration method modified by Liang et al. (Liang et al., 2008). The principle is persulfate anion reacts with KI when NaHCO3 is present, thus forms an iodine yellow color that absorbs in the UV–Vis range with absorption peaks at 288 and 352 nm. A UV spectrophotometric method is as follows: (1) dilute the sample solution to about 30 mL; (2) remove the oxygen of solution by bubbling nitrogen; (3) add some sodium bicarbonate in the solution to avoid oxygen-oxidation of iodide; (4) add 4 g of potassium iodide, stir to dissolve the iodide, and then allow to stand for 15 min; (5) the subsequent experiments were conducted at 352 nm. 2.7. Theoretical calculation The PRO of molecular orbital calculations were obtained through the B3LYP/6-311G + (d, p) level, thereby the optimal conformation of propranolol having a minimum energy derived in the Gaussian 09 program (Frisch et al., 2009). The atomic numbering of propranolol and its optimized structure are showed in Fig. S3. The two-dimensional electron density color-filled maps of the highest occupied molecular orbital (HOMO) and the lowest unoccupied molecular orbital (LUMO) of PRO was obtained by the Multiwfn (Lu and Chen, 2012). The position of nucleophilic, electrophilic and radical reactions take place as follows (Lee et al., 2001):

The residual concentrations of PRO and BA were analyzed using a Shimadzu High Performance Liquid Chromatography (HPLC) system. The chromatographic conditions of PRO and BA were as follows. Column: VP-ODS C18 (4.6 × 150 mm, 5 μm); temperature: 40 °C; mobile phase: acetonitrile /10 mM KH2PO4 buffer solution with pH 3.0 (35:65 v/v); flow rates: 1.0 mL·min−1; injection volume: 10 μL. The detection wavelength of PRO and BA were 213 nm and 227 nm, respectively. An ultra performance liquid chromatography coupled with quadrupole time-of-flight mass spectrometry (UPLC/Q-TOF-MS) technique was employed to identify the transformation products of PRO. Separation was accomplished using an BEN-C18 column (1.7 μm, 2.1 × 50 mm). The mobile phase was ultrapure water with 0.1% formic acid (Solvent A) and acetonitrile (Solvent B) with a flow rate at 0.3 mL min−1. The gradients for PRO are described in Table S1. Waters micro mass Q-TOF micro system (Waters Co., USA) was equipped with the Lock Spray and electrospray ionization (ESI) interface operating in positive ion mode. The capillary voltage and sample cone voltage were 3000 V and 30 V, respectively. The source temperature was set at 100 °C, whereas the desolvation temperature was maintained at 300 °C. The cone and desolvation gas (nitrogen) flow rates were 50 L·h−1 and 600 L·h−1, respectively. Data were obtained in continuum mode in the mass scanning rage of m/z 50–600. Total organic carbon (TOC) was determined using a TOC analyser (TOC-VCPH, Shimadzu). 3. Results and discussion 3.1. Comparison of PRO removal by UV/H2O2 and UV/PS oxidation Fig. 1 exhibits the degradation of PRO in the H2O2, PS, UV, UV/H2O2 and UV/PS processes, respectively. As shown, the degradation of PRO by H2O2 or PS oxidation alone in the dark was negligible. Furthermore, the UV irradiation alone resulted in a PRO removal of only 5.9% within 30 min, which was determined by its own molar absorption coefficient and lower PRO quantum yield. In contrast, the combination of UV irradiation with H2O2 or PS significantly enhanced the degradation rate of PRO, from 5.9% to 78.3% and 91.8%, respectively, which indicated a considerable contribution from oxidizing radical species (HO· and SO•− 4 ). Moreover, the degradation of PRO in UV/H2O2 and UV/PS processes both obeyed pseudo-first-order reaction kinetics, and the degradation efficiency could be described as following:
ln ½PRO=½PRO0 ¼ −kobs t

ð1Þ

where kobs is the pseudo-first-order rate constant (min−1). [PRO]0 is the initial concentration (μM) of PRO, and [PRO] is the concentration at any time t (min). And the kobs value of PRO in UV/H2O2 and UV/PS processes can be obtained to be 0.0518 and 0.0817 min−1, respectively. It may be clearly seen that kobs of PRO in UV/PS oxidation process was higher than that in UV/H2O2 process when they were conducted under the same

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(a) 1.0

881

(b)

0.8 [PRO]/[PRO]0

UV UV/H2O2 UV/PS H2O2 PS

0.6 0.4

Kobs(min-1)

0.08

0.06

0.04

0.02

0.2 0.0

0.00 0

5

10

15 20 Time (min)

25

30

UV

UV/H2O2 UV/PS

2E-4

1E-4

H2O2

PS

Fig. 1. Degradation of PRO in H2O2, PS, UV, UV/H2O2 and UV/PS processes. Experimental conditions: [PRO]0 = 20 μM, [PS]0 = [H2O2]0 = 0.4 mM, pH = 7.0.

oxidant ratio. These results were consistent with other similar studies (Gao et al., 2012; Zhou et al., 2017). It is generally accepted that HO· and SO•− 4 are the primary oxidizing species that are responsible for removal of contaminants, in UV/H2O2 and UV/PS processes, respectively. Thus, it can be conducted that SO•− 4 was the better effective oxidizing specie for the degradation of PRO in the present study.

species than EtOH. Above results suggested that SO•− 4 were the primary radical species in reaction solutions at neutral pH, whereas HO· played a much less important role. These results were consistent with previous studies where SO•− 4 was identified as the main radical species during the degradation of 2,4-Di-tert-butylphenol in UV/PS process (Wang et al., 2016).

3.2. Identification of predominant radical species

3.3. Determination of the second-order rate constants

In order to identify the main radical species that were mainly responsible for the degradation of PRO and to explore the underlying degradation mechanisms in UV/PS process, experiments with radical scavengers (ethanol (EtOH) and tertiary butanol (TBA)) were conducted. The rate constant of EtOH reacting with HO· is about 50 times higher than that with SO•− 4 , on the contrary, the rate constant of TBA reacting with HO· is approximately 1000 times greater than that with SO•− 4 . The difference of reaction rates between EtOH or TBA with HO· or SO•− 4 may be used to distinguish the predominant oxidizing species (Anipsitakis and Dionysiou, 2004; Ji et al., 2015). The results presented in Fig. 2 show that the kobs decreased markedly with increasing EtOH dosages, which suggested that the radical species were efficiently quenched with the addition of EtOH. Nevertheless, when TBA is used, the decrease of kobs was comparatively slower, which indicated that TBA had less efficacy in quenching the radical

The radical scavenger experiments above displayed that both SO•− 4 and HO· contributed to the decomposition of PRO in UV/PS process. In the present study, competition kinetics experiments were conducted to indirectly determine the second-order rate constants of PRO reacting with SO•− 4 and HO·. BA was selected as reference compound, since its reaction rate constants with SO•− 4 and HO· is known (kSO4•−/BA = 1.2 × 109 M−1 s−1; kHO•/BA = 5.9 × 109 M−1 s−1, 41]. The reaction rate constants of PRO reacting with SO•− 4 and HO· were determined by Eq. (2).

0.09 0.08 EtOH TBA

Kobs(min-1)

0.07 0.06 0.05 0.04 0.03 0.02

ln

½PRO0 kPRO ½BA0 ¼ ln ½PROt ½BAt kBA

ð2Þ

where [PRO]t and [BA]t are the concentrations of PRO and BA at reaction time t, respectively; [PRO]0 and [BA]0 are the initial concentrations of PRO and BA, respectively; kPRO and kBA are the second-order rate constants of PRO and BA for SO•− 4 or HO·. Fig. 3 reveals the reaction rate constant ratio between PRO and BA with SO•− 4 /HO· at pH 7.0, which was achieved by competition kinetics methods. The relative rate of PRO to BA with SO•− 4 was k1 = 16.164, and the second-order rate constant for the reaction of PRO with SO•− 4 was calculated to be 1.94 × 1010 M−1 s−1. In a similar way, the second-order rate constant for the reaction of PRO with HO· was calculated to be 6.77 × 109 M−1 s−1 at the base of the relative rate of PRO to BA with HO· (k2 = 1.148). It is noteworthy that the second-order rate constant of PRO with SO•− 4 was higher than that with HO·, which might explain the phenomenon that the UV/PS oxidation process was more effective than UV/H2O2 process for PRO degradation. These data implied further that SO•− 4 was more effective than other oxidation species for the removal of PRO, which makes SO•− 4 -induced AOPs a promising technology for the elimination of PRO and structurally related contaminants. 3.4. Effect of initial pH

0

100 200 300 400 500 Scavenger dosage (mM)

600

Fig. 2. Effect of different dosages of EtOH and TBA on kobs of PRO in UV/PS process. Experimental conditions: [PRO]0 = 20 μM, [PS]0 = 0.4 mM, [EtOH]0 = 0–600 mM, [TBA]0 = 0–600 mM, pH = 7.0.

The initial pH of solution plays a complex role in the decomposition of organic compounds in SO•− 4 -based AOPs, as it may affect the formation of reactive species and the ionization state of organic pollutants (Kolthoff and Miller, 1951; Dogliotti and Hayon, 1967). Under acidic pH, PS may undergo acid-catalyzed decomposition, depleting PS dosage

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(a)

4 k=16.164

0.14

UV/H2O2

0.13

UV/PS

0.12

Kobs(min-1)

ln([PRO]0/[PRO]t)

5

0.11

3

0.10

2

0.09

1

0.08

k=1.148

0.07

0

4

1.5

Fig. 3. Determination of the second-order rate constants for the reaction of PRO with SO•− 4 and HO·. Experimental conditions: for k(SO•− 4 /PRO), [PRO]0 = [BA]0 = 20 μM, [PS]0 = 0.4 mM, [TBA]0 = 100 mM, and pH = 7.0; for k(HO·/PRO), [PRO]0 = [BA]0 = 20 μM, [H2O2]0 = 0.4 mM, and pH = 7.0.

through non-radical pathways without SO•− 4 generation (Kolthoff and − Miller, 1951). Under basic pH, SO•− 4 can react with H2O or OH , and then was transformed to HO· (Eqs. (3)–(4)) (Zhao et al., 2013; Fang et al., 2013). 2− þ SO•− 4 þ H2 O→HO  þSO4 þ H ; − 2− SO•− 4 þ OH →HO  þSO4 ;

5

6

2.0

k½H2 Ob3  103 M−1 s−1

ð3Þ

k ¼ ð6:5  1:0Þ  107 M−1 s−1

ð4Þ

We conducted experiments to analyze changes in pH before and after the experiment. The result is shown in Fig. S4. After 30 min of experimentation, the pH of the solution was reduced. This phenomenon is consistent with the experimental analysis. The effect of initial pH of solution on kobs of PRO is presented in Fig. 4a. As clearly seen, the kobs value increased from 0.0731 to 0.1354 min−1 with increasing initial solution pH from 4.0 to 10.0. The pH-dependent speciation of PRO in the solution is shown in Fig. 4b. PRO has two forms of existence at different pH ranges due to the pKa value of 9.53(Wang et al., 2017). Thus PRO mainly exists in the protonated form when the solution is below its pKa, and when the solution pH is higher than the pKa, the proportion of the deprotonated form of PRO increases. It may been seen that the deprotonated form of PRO gradually increased with increasing pH. Above phenomenon suggested that the deprotonated form of PRO had higher activity to react with SO•− 4 and/or HO·, in contrast to protonated or neutral forms of PRO (Anquandah et al., 2013). It has been demonstrated that the deprotonated form of PRO facilitated its oxidization because of the weaker electron-withdrawing effect of the amine groups (Yang et al., 2010). The low kobs value under acidic condition could also be attributed to lower formation efficiencies of SO•− 4 (Tan et al., 2013). Furthermore, with the decreasing solution pH, the percentage of less reactive protonated PRO gradually increased, which might also be responsible for the observed low kobs value of PRO. 3.5. Effect of initial PS dosage The initial dosage of PS plays a crucial role on the removal of emerging pollutants in UV/PS process, as it may directly affect the equilibrium concentration of SO•− 4 (Liu et al., 2016; Zhang et al., 2016a). The effect of initial PS dosage on kobs value of PRO is shown in Fig. 5. As shown, the kobs increased markedly from 0.0197 to 0.1433 min−1 with increasing initial PS dosages from 0.05 to 0.8 mM. Furthermore, a linear

(b)

Fraction of PRO (%)

0.5 1.0 ln([BA]0/[BA]t)

7 8 Solution pH

9

10

1.0 0.8 0.6

PRO HPRO+

0.4 0.2 0.0 0

2

4

6

pH

8

10

12

14

Fig. 4. (a) Effect of initial pH on kobs of PRO in UV/PS process; (b) distribution of dissociation species of PRO. Experimental conditions: [PRO]0 = 20 μM, [PS]0 = 0.4 mM, pH = 4.0–10.0.

0.15 y=0.163x+0.01405 R2=0.998

0.12

kobs(min-1)

0.0

0.09 0.06 0.03 0.00 0.0

0.2

0.4 0.6 PS dosage (mM)

0.8

Fig. 5. Effect of initial PS dosage on kobs of PRO in UV/PS process. Experimental conditions: [PRO]0 = 20 μM, [PS]0 = 0.05–0.8 mM, pH = 7.0.

T. Chen et al. / Science of the Total Environment 690 (2019) 878–890

0.18

be formed with increased initial PS dosages during the UV-activated PS oxidation process. Several previous studies have reported that the non-linear increase of kobs can be observed at high PS concentrations because of the scav•− enging of SO•− 4 by excess PS and the recombination of SO4 (Eqs. (5) and (6)) (Neta et al., 1988; Liu et al., 2016; Zhang et al., 2016a). However, in this study, the inhibition impact of PS was not discovered, likely because the highest PS dosage (0.8 mM) did not attain the critical level, which began to slow down PRO degradation.

0.15 0.12 0.09 0.06 0.03

2− 2− •− 6 −1 −1 SO•− s 4 þ S2 O8 →SO4 þ S2 O8 ; k ¼ 1:2  10 M

ð5Þ

•− 2− 8 −1 −1 s SO•− 4 þ SO4 →S2 O8 ; k ¼ 4:0‐8:1  10 M

ð6Þ

3.6. Effect of initial PRO concentration

0.00 10

20 30 40 PRO concentration (μM)

50

Fig. 6. Effect of initial PRO concentration on kobs of PRO in UV/PS process. Experimental conditions: [PRO]0 = 10–50 μM, [PS]0 = 0.4 mM, pH = 7.0.

relationship between kobs and PS dosage (kobs = 0.163 × PS + 0.01405, R2 = 0.998) was observed. Similar findings have been reported for the degradation of sulfamethazine and methyl paraben at various PS dosages in UV/PS process (Gao et al., 2012; Dhaka et al., 2017). It is generally believed that higher oxidizing species (SO•− 4 and HO·) levels will

1.0

UV/PS UV/PS+1mM NO-3

0.8

UV/PS+1mM HCO-3

[PRO]/[PRO]0

(a)

UV/PS+1mM Cl-

0.6

The effect of initial concentration of PRO on kobs value was evaluated. As depicted in Fig. 6, the observed kobs decreased from 0.1581 to 0.0285 min−1 when the initial concentration of PRO ranged from 10 to 50 μM. This is because the increased larger number of PRO molecules and their transformation products could compete the reactive radicals (Shah et al., 2013). Furthermore, with the increase of the initial PRO concentration, most of the UV photons was absorbed by PRO molecules and their transformation products, which thereby inhibited the absorption of UV irradiation by the PS, and reduced the formation of SO•− 4 for PRO degradation. Similar phenomenons have been observed for the degradation of antipyrine and oxytetracycline in UV/PS process (Tan et al., 2013; Zhang et al., 2016b).

(b) 0.16

0.12

Kobs(min-1)

Kobs(min-1)

883

0.08

0.4 0.2

0.04

0.0

0.00 0

10 15 20 Time (min)

25

30

UV/PS

UV/PS +HCO-3

UV/PS +Cl-

UV/PS +NO-3

(d) 1.0

0.08

0.8

0.06

Kobs(min-1)

[PRO]/[PRO]0

(c)

5

0.6

0.04

0.4 UV/PS UV/PS+1mg/L HA UV/PS+5mg/L HA UV/PS+10mg/L HA

0.2

0.02

0.00

0.0 0

5

10 15 20 Time (min)

25

30

0

1 5 HA (mg/L)

10

− − Fig. 7. Effect of inorganic ions (a-b) and NOM (c–d) on PRO degradation in UV/PS process. Experimental conditions: [PRO]0 = 20 μM, [PS]0 = 0.4 mM, [HCO− 3 ]0 = [Cl ]0 = [NO3 ]0 = 1 mM, [HA]0 = 0–10 mg C·L−1.

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(a)

(b)

0.0

1.0 0.8

-1.0

0.6

C/C0

ln ([PRO]/[PRO]0)

-0.5

-1.5 UW kobs=0.082 min-1

-2.0

PRO change TOC change PS change

0.4 0.2

RW kobs=0.040 min-1 WW kobs=0.010 min-1

-2.5 0

5

10

15

0.0

20

25

30

0

40

Time (min)

80 120 Time (min)

160

Fig. 8. (a) Effects of water matrices on the degradation of PRO (UW: ultrapure water, RW: river water, WW: secondary wastewater effulent); (b) degradation and mineralization of PRO as well as PS consumption in UV/PS process. Experimental conditions: [PRO]0 = 20 μM (3.574 mg/L TOC), [PS]0 = 0.4 mM, pH = 7.0.

3.7. Effect of natural water constituents − Inorganic anions (e.g., Cl−, NO− 3 , and HCO3 ) and natural organic matter (NOM) are ubiquitous natural waters constituents, and have been proved to exhibit different impacts on the removal of organic contaminants by SO•− 4 -induced AOPs (Ghauch and Tuqan, 2012; Fan et al., 2015). Thus, it is necessary to investigate the influence of these natural water body constituents on the degradation of PRO for potential practical application of UV-activated PS oxidation in the treatment of PRO in natural waters.

− • SO•− 4 by NO3 to produce less reactive species NO3(Eq. (12)) (Exner et al., 1992). Furthermore, the light screening effect of NO− 3 could weaken the UV light intensity (Sörensen and Frimmel, 1997), and − thus reduce the formation of SO•− 4 . Similar influence of NO3 on destruction of florfenicol in UV/PS process have also been observed (Gao et al., 2015).

3.7.1. Effects of inorganic anions – Fig. 7a exhibits the effect of Cl−, NO− 3 , and HCO3 at 1 mM level on the destruction of PRO. The results revealed that the presence of Cl− and HCO–3 showed significant promotion effect on the degradation of PRO, – whereas NO− 3 significantly suppressed the reaction. HCO3 usually act as an radical scavenger in AOPs, and can quench SO•− and HO· to gen4 erate less reactive carbonate radicals (HCO•3 and CO•− 3 ) (Eqs. (7)–(9)) (Liang et al., 2006). Although the redox potentials of HCO•3 and CO•− 3 is lower than SO•− 4 and HO·, they can react with electron-rich compounds, for instance, phenols and anilines, at a fairly high rate (Larson and Zepp, 1988; Canonica et al., 2005). In this study, the enhancement in PRO degradation rate might be attributed to the reactions of HCO•3/CO•− 3 with PRO that contains an electron-rich naphthalene ring.

3.7.2. Effects of NOM Fig. 7b exhibits the effect of humic acid (HA), a typical constituent of NOM, on PRO degradation. As shown, the degradation of PRO was drastically inhibited in the presence of different amounts of HA. The removal rate of PRO decreased consistently from 91.8% to 26.1% with increasing the dosages of HA from 0 to 10 mg C·L−1. The results could be explained by that NOM could act as the scavengers of SO•− 4 and HO· due to the electron-rich moieties within NOM molecules structure, which were prone to react with these electrophilic radicals (Westerhoff et al., 2007; Gara et al., 2009). Besides, as the light filter, HA could compete with the target pollutant and PS for UV light, which might affect the direct UV photodegradation of PRO and significantly reduce the generation of SO•− 4 (Khan et al., 2013; Xie et al., 2015). As a result, the degradation of PRO drastically inhibited by the introduction of HA. The inhibition effect of HA in SO•− 4 -induced oxidation process has been observed in the degradation methyl paraben using UV-activated PS method (Dhaka et al., 2017).

•− 6 −1 −1 HO  þHCO− s 3 →CO3 þ H2 O; k ¼ 8:5  10 M

ð7Þ

3.8. Effects of water matrices on the degradation of PRO

− 2− • 6 −1 −1 SO•− s ð pH ¼ 8:4Þ 4 þ HCO3 →SO4 þ HCO3 ; k ¼ 1:6  10 M

ð8Þ

HCO•3 ↔Hþ þ CO•− 3 ; pK a ¼ 9:5

ð9Þ

In order to investigate the feasibility of the UV/PS process on the elimination of PRO under environmentally relevant conditions, the degradation of PRO was performed in different natural water matrices (ultrapure water, river water and secondary wastewater effluent). As shown in Fig. 8a, the degradation process of PRO was moderately

−

•

2− 8 −1 −1 SO•− s 4 þ Cl →SO4 þ Cl ; k ¼ 3:1  10 M •

−

•−

ð10Þ

Cl þ Cl →Cl2 ; k ¼ 2:1  1010 M−1 s−1

ð11Þ

•− 2− • 4 −1 −1 NO− s 3 þ SO4 →SO4 þ NO3 ; k ¼ 5:0  10 M

ð12Þ

− •− In SO•− 4 -induced AOPs, Cl can react with SO4 to generate reactive • •− chlorine species such as Cl and Cl2 (Eqs. (10)–(11)) (Neta et al., 1988). These chlorine species were prone to react with electron-rich compounds, such as phenols, producing chlorinated compounds (Anipsitakis et al., 2006). Therefore, the promoting effect of Cl− on PRO degradation might be explained through the reactions of chlorine species with the naphthalene ring of PRO. The inhibiting effect of NO− 3 on PRO degradation could be explained through the scavenging of

Table 1 Economic comparison of UV, UV/H2O2 and UV/PS processes for PRO removal in DI water (pH 7.0; UV intensity: 0.285 mW cm−2). Molar fraction

k (min−1)

UV UV/H2O2 UV/PS

– 20 2.5 5 10 20 40

EEO (kWhm−3 order−1)

(oxidant/PRO) 0.002 0.052 0.020 0.031 0.048 0.082 0.143

3840.00 147.69 384.00 247.74 160.00 93.66 53.70

T. Chen et al. / Science of the Total Environment 690 (2019) 878–890

885

Table 2 Mass spectrometry information and proposed structure for the transformation products of PRO in UV/PS process via UPLC/Q-TOF-MS. TP

Retention time (min)

Molecular formula

ESI(+) MS m/za

ESI(+) MS2 m/z

PRO

7.72

C16H21NO2

260

218, 183, 157, 116, 98, 74, 56

TP 308

4.24

C16H21NO5

308

290, 249, 230, 175, 131, 116, 98

TP 292-1

2.74

C16H21NO4

292

274, 248, 230, 161, 133, 116, 72

TP 292-2

6.72

C16H21NO4

292

274, 250, 232, 177, 159, 131, 116, 98, 72

TP 290

2.76

C16H19NO4

290

248, 230, 174, 159, 131

TP 282

2.77

C14H19NO5

282

264, 222, 149, 134, 116, 98, 72

TP 276-1

6.32

C16H21NO3

276

276, 116

TP 276-2

6.32

C16H21NO3

276

276, 116

TP 266

3.22

C14H19NO4

266

184, 159, 149, 116

Proposed structure

(continued on next page)

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T. Chen et al. / Science of the Total Environment 690 (2019) 878–890

Table 2 (continued) TP

Retention time (min)

Molecular formula

ESI(+) MS m/za

ESI(+) MS2 m/z

TP 264

2.38

C14H17NO4

264

222, 160, 148, 116

Proposed structure

TP: transformation product. a m/z values shown are for protonated molecular ions [M + H]+.

decreased in RW compared with in WW for the UV/PS process. The degradation process was significantly inhibited in WW. The degradation rate constants decreased from 0.082 min−1 to 0.010 min−1 for ultrapure water and secondary wastewater effluent, respectively. Probably because large amounts of NOM would scavenge the reactive radical species and compete for the UV radiation, remarkably inhibiting the generation of radical species. This phenomenon is consistent with the research of Frontistis (Frontistis, 2019) who studied the degradation of the nonsteroidal anti-inflammatory drug piroxicam using UVactivated persulfate. The author found that in the secondary effluent, the apparent kinetic constant decreased almost 14 folds. These results suggests that a pretreatment was necessary for the application of UV based AOPs in some real wastewater waters. 3.9. Mineralization of PRO and PS consumption TOC is usually used as a measure of mineralization of target pollutants. Fig. 8b shows the TOC removal and PS consumption corresponding to the degradation of PRO during the oxidation process over a period of 180 min. Following 30 min of reaction time, 91.8% of the PRO was removed against 6.9% of TOC removal. Following 60 min of reaction time, the PRO was almost completely removed, while only a ~10.3% removal of TOC was achieved, which suggested that the intermediates of PRO were recalcitrant against the UV-activated PS oxidation process. A similar finding was reported by Liu et al. (Liu et al., 2016), who studied the mineralization of oxytetracycline in UV/PS process, and found only a 15.7% of TOC was removed subsequent to 10 h UV irradiation. When increasing the irradiation time from 60 to 180 min, the removal of TOC attained 45.3%, which indicated that the intermediates were gradually mineralized, albeit very slowly. It is noteworthy that the PS consumption gradually increased during the mineralization of PRO. Under 180 min of UV irradiation, PS consumption reached 59.7%. Above phenomenon suggested the mineralization of PRO and PS consumption would be enhanced with prolonging the reaction time, and confirmed the potential practical applicability of this treatment technology in the removal of contaminants from natural ambient water. 3.10. Economic comparison of UV/PS and UV/H2O2 processes Due to the US/UV/H2O2 process is electric-energy intensive and electric energy can represent a major fraction of the operating costs, simple figures-of-merit based on electric energy consumption can be very informative. Hence, Bolton et al. defined the figures-of-merit electric energy per order (EEO) to use in the first-order kinetic regime of AOPs.

On the basic of that the energy electric energy consumed by UV lamp is expressed in terms of electrical energy per order (EE/OUV), which is defined as the electric energy in kWh required to degrade the target pollutant by one order of magnitude in 1 m3 of water. The economic comparison of UV, UV/PS and UV/H2O2 processes for PRO degradation were calculated and summarized. The EE/OUV can be calculated based on the Eq. (13). EE=OUV ≡

 P  t  1000 38:4  P  −3 −1  ¼ kWhm order C0 V k V  log Ct

ð13Þ

where P is the UV lamp power (kW) to the AOPs system; t is the irradiation time (h); V is the volume of solution in reactor (L); k is the firstorder rate constant (min−1); C0 is the initial pollutant concentrations; Ct is the final pollutant concentrations (Bolton et al., 2001). The results were presented in Table 1. The higher electrical energy consumption means lower process efficiency. The pure UV process consumed 3840.00 kWhm−3 order−1, whereas UV/H2O2 process consumed 147.69 kWhm−3 order−1. In contrast, UV/PS process showed the less electrical energy consumption (93.696 kWhm−3 order−1). The EE/OUV

Table 3 FEDs on atoms of PRO calculated using the Gaussian 09 program at the b3lyp/6-311 + g (d, p). (number) Atom

2FED2HOMO

2FED2LUMO

FED2HOMO + FED2LUMO

Point charge

1C 2C 3C 4C 5C 6C 9C 10 C 13 C 14 C 18 O 19 C 20 C 24 C 27 O 29 N 31 C 32 C 33 C

0.018759 0.022399 0.002535 0.000002 0.032123 0.010547 0.013544 0.015686 0.004713 0.010480 0.027078 0.000148 0.000107 0.000235 0.000024 0.003394 0.000211 0.000079 0.000363

0.007970 0.027880 0.000460 0.000316 0.022160 0.014100 0.034041 0.034196 0.013438 0.014522 0.006279 0.000009 0.000028 0.000017 0.000009 0.000037 0.000006 0.000006 0.000003

0.013365 0.025140 0.001498 0.000159 0.027141 0.012323 0.023792 0.024941 0.009076 0.012501 0.016679 0.000078 0.000067 0.000126 0.000016 0.001716 0.000109 0.000043 0.000183

−0.137078 0.202275 −0.129349 −0.065541 −0.075926 −0.113547 −0.035099 −0.086280 −0.100504 −0.113865 −0.371957 −0.003379 0.036025 −0.082476 −0.444297 −0.410060 −0.088639 −0.275392 −0.279436

See Fig. S3 for atom numbering.

T. Chen et al. / Science of the Total Environment 690 (2019) 878–890

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3.11. Identification of transformation products

Fig. 9. The molecular orbital of the PRO: (a) HOMO; (b) LUMO.

As shown above, the effective destruction of PRO was achieved in UV/ PS process. However, evaluating the formation of transformation products (TPs) prior is necessary in order to examine the actual applicability of this technology. In this study, the identification of transformation products was performed via UPLC/Q-TOF-MS technique, operating in positive mode (ESI+), as well as the PRO molecular orbital calculations. Following 20 min of UV-activated PS oxidation, new peaks appeared in the total ion chromatogram (TIC) (as shown in Fig. S5). The mass spectra data and proposed structure of transformation products for PRO degradation are summarized in Table 2. We identified transformation products in total, with m/z of 264, 266, 276, 282, 292, and 308, respectively. It may be clearly seen that several transformation products exhibit the same precursor ion, which suggested that they might be constitutional isomers. For this study, the initially identified entities were TP 276–1 and TP 276–2 isomers, which represented the addition of 16 mass units to the PRO molecule, imputable to monohydroxylated transformation products. TP 292 was identified as naphthalene ring-opening product, which resulted from the addition of two oxygen atoms to PRO molecule. The MS2 spectra of TP 292–1 exhibited different patterns with TP 292–2. The fragment ions m/z 131 and 159 of TP 292–2 confirmed the presence of two aldehyde moieties, whereas, the fragment ion m/z 133 and 161 of TP 292–1 supported the presence of aldehyde and hydroxyl moieties. TP 308 was identified as the oxidation product of the TP 276–2. TP 290 was identified as the dehydrated product of TP 308 of the branched ether chain. TP 266 was identified as the aldehyde group cleavage product of TP 292–2, as well as TP 282, which also was identified as an aldehyde group cleavage product of TP 308. TP 264 was identified as an aldehyde group cleavage product of TP 290. 3.12. Transformation mechanisms and pathways

values revealed that increase of rate constant lowers the consumption of electrical energy. The Table 1 displayed the EE/OUV increases with decreasing the PS concentration.

The identification tests of radical species demonstrated that the UVactivated PS oxidation process predominantly generated SO•− 4 . FEDs on

Fig. 10. Proposed transformation pathways for PRO degradation in UV/PS process. Experimental conditions: [PRO]0 = 20 μM, [PS]0 = 0.4 mM, pH = 7.0. [(1) Hydroxylation; (2) Naphthalene ring-opening; (3) Aldehyde group cleavage; (4) Dehydration.]

888

T. Chen et al. / Science of the Total Environment 690 (2019) 878–890

Scheme 1. Hydroxylation at the aromatic ring.

atoms of PRO calculated using the Gaussian 09 program at the b3lyp/ 6–311 + g (d, p) in Table 3. As revealed in the Fig. 9, the HOMO and LUMO of the PRO means that the electrons flow from the position of HOMO to LUMO, thus these sites are vulnerable to radical species attack. As shown in Fig. 10, based on the predominant identified radical species and the elucidated transformation product structures, detailed transformation mechanism and three pathways of PRO in UV/PS process were preliminarily proposed. Initially, it was suggested that hydroxylation that occurred in naphthalene ring (attack at C5) through action of SO•− 4 formed monohydroxylated TP 276–1 isomers. According to the 2FED2LUMO, C10 was attacked by a nucleophilic reaction to generate TP 276–2. Hydroxylation is a familiar transformation pathway in oxidation reaction of SO•− 4 . As pKa of PRO is 9.53(Kibbey et al., 2007), the reactivity of the secondary amine moiety was very weak under the experimental conditions studied in terms of the chemical structure of PRO. Thus, naphthalene group might be the predominant hydroxylation reaction site, which is consistent with theoretical calculations. Where the oxidation step on the secondary amine moiety might be considered as a secondary reaction site. According to the common reaction mechanisms of SO•− 4 with aromatic ring, SO•− 4 firstly reacts with the aromatic ring and then can result 2− in the generation of a short-lived SO•− 4 adduct, which eliminates a SO4 toward the generation of a radical cation. As shown in Scheme 1, electron transfer occurs to form a C-centered radical cation while SO•− 4 participates in the reaction, which can react with H2O via hydrogenation due to the ionization potential of aromatic ring (He et al., 2014; Norman et al., 1970; Li et al., 2018). Hydroxycyclohexadienyl radical is thus generated, which may be attacked by dissolved oxygen quickly. Related kinetic and thermochemical studies demonstrate the occurrence of this reaction (Grebenkin and Krasnoperov, 2004). Eventually, hydrogen superoxide radical and a hydroxylated transformation product were formed (Atkinson, 2000). On the basis of the above mechanism and the results of molecular orbital calculations, the monohydroxylation TP 276–1 isomers could be formed. Besides, the formation of monohydroxylated TP 276–2 (attack at C10) was possibly derived from the direct attack on the carbon-centered radical via HO·, − which was produced through the reactions of SO•− 4 with H2O or OH (Zhao et al., 2013; Fang et al., 2013). The continuous attack of SO•− 4 on each side of a double bond of an activated naphthalene group system led to the ring being opened via a 1,3-dipolar cycloaddition, which resulted in the TP 292–2, possessing two aldehyde groups. Moreover, the attack of SO•− 4 on the naphthalene group also produced the intermediate m/z 294 that has been reported in the degradation of PRO by K2FeVIO4 (Wilde et al., 2013), but not in this work, probably because of its rapid transformation rate in the UVactivated persulfate process. The simultaneous formation of TP 308 indicated that the ring-opening step was followed by the attack of SO•− 4 to produce the secondary oxidized product. The naphthalene group and secondary amine moiety are potential target sites for attack of SO4•−. The hydroxylation of the secondary amine moiety has been previously suggested (Santiago-Morales et al., 2013; Romero et al., 2011; Benner and Ternes, 2009). The dehydration referred to an elimination of H2O, and have been reported to be a significant reaction pathways in SO•− 4 based oxidation processes (Liu et al., 2016). In terms of the PRO

structure, the dehydration at hydroxyl group of side chain led to the formation of a stable double bond, making it a more likely reaction site. In this study, TP 290 was formed by the dehydration of TP 308. Further, TP 292–1 was formed by the hydroxylation of the secondary amine group and the dehydration of intermediate m/z 294. It has been reported that the double bond in the α-position of the aldehyde groups was easily attacked by ozone and Fe(VI), and thus causing the cleavage of the group (Benner and Ternes, 2009; Anquandah et al., 2013). Both TP 266, TP 282, and TP 264 were the transformation products of aldehyde group cleavage, which was attacked by SO•− 4 . In addition, the hydroxylation of the secondary amine moiety of TP 266 also led to the formation of TP 282.

4. Conclusions In summary, this work report the degradation of PRO in UV/PS oxidation process was systematically investigated. It was demonstrated that effective degradation of PRO was achieved at experimental conditions, and the degradation obeyed pseudo-first-order reaction kinetics. The degradation rate constant (kobs) was observed to increase with higher initial PS dosages or decreased initial PRO concentrations. Increasing initial pH could also enhance degradation efficacy. SO•− 4 was identified as the main radical species responsible for PRO removal. The second-order rate constants for the reaction of PRO with SO•− 4 and HO· were anticipated to be 1.94 × 1010 M−1 s−1 and 6.77 × 109 M−1 s−1, respectively at neutral pH. The presence of Cl− and HCO− 3 showed promotion effect on PRO degradation. In contrast, natural water NOM and NO− 3 showed inhibitory effect on the degradation of PRO. The mineralization of PRO and PS consumption were both enhanced through prolonging the reaction time. Transformation products were identified by UPLC/Q-TOF-MS technique. Moreover, the molecular orbital calculation of propranolol and HOMO and LUMO were obtained by B3LYP/6-311G + (d, p) level and Multiwfn, respectively. Hence, the degradation mechanism of PRO is revealed in more depth and the transformation pathways including hydroxylation, dehydration, naphthalene ring opening, and the cleavage of aldehyde groups were illustrated in UV/PS process. This study provides significant information on the practical application of this treatment technology for PRO removal from natural ambient water, also helps to better understand and evaluate the characteristics of propranolol degradation in advanced oxidation process based on sulfate radicals and has practical significance to the treatment of other organic pollutants.

Acknowledgments This work was financially supported by the National Natural Science Foundation of China (No. 21677040). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.07.034.

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