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Article

Mechanistic Insight into Degradation of Cetirizine under UV/Chlorine Treatment: Experimental and Quantum Chemical Studies

by
Boyi Zhu
†,
Fangyuan Cheng
†,
Wenjing Zhong
,
Jiao Qu
,
Ya-nan Zhang
* and
Hongbin Yu
*
State Environmental Protection Key Laboratory of Wetland Ecology and Vegetation Restoration, School of Environment, Northeast Normal University, Changchun 130117, China
*
Authors to whom correspondence should be addressed.
†
These authors contributed equally to this work.
Water 2022, 14(9), 1323; https://doi.org/10.3390/w14091323
Submission received: 26 March 2022 / Revised: 12 April 2022 / Accepted: 18 April 2022 / Published: 19 April 2022
(This article belongs to the Topic Advanced Oxidation Process: Applications and Prospects)
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Abstract

:
UV/chlorine treatment is an efficient technology for removing organic pollutants in wastewater. Nevertheless, degradation of antihistamines in the UV/chlorine system, especially the underlying reaction mechanism, is not yet clear. In this study, the degradation of cetirizine (CTZ), a representative antihistamine, under UV/chlorine treatment was investigated. The results showed that CTZ could undergo fast degradation in the UV/chlorine system with an observed reaction rate constant (kobs) of (0.19 ± 0.01) min−1, which showed a first-increase and then-decrease trend with its initial concentration increased. The degradation of CTZ during the UV/chlorine treatment was attributed to direct UV irradiation (38.7%), HO• (35.3%), Cl• (7.3%), and ClO• (17.1%). The kobs of CTZ decreased with the increase in pH and the increase in concentrations of a representative dissolved organic matter, Suwannee River natural organic matter (SRNOM), due to their negative effects on the concentrations of reactive species generated in the UV/chlorine system. The detailed reaction pathways of HO•, ClO•, and Cl• with CTZ were revealed using quantum chemical calculation. This study provided significant insights into the efficient degradation and the underlying mechanism for the removal of CTZ in the UV/chlorine system.

Graphical Abstract

1. Introduction

Pharmaceuticals and personal care products (PPCPs) have attracted widespread attention because of their ubiquity in the aqueous environment and potentially high risks to the ecosystem [1,2,3]. Antihistamines, as one kind of PPCPs, are mainly used to alleviate human allergies [4]. In recent years, antihistamines have been frequently detected in the environment due to the extensive production and widespread use [5]. The concentrations of antihistamines in surface water have been found to be mainly at ng L−1 level, and high concentrations up to µg L−1 level were also reported [6,7]. The antihistamines in surface water are mainly discharged from wastewater treatment plants (WWTPs), of which effluent contains antihistamines with concentrations as high as mg L−1 level [7]. Although antihistamines have been produced and used as specific histamine H1-receptor antagonists, the toxic effects, e.g., cardiotoxicity and sublethal effects, of frequently used antihistamines to aquatic organisms have been observed [6,8,9]. Thus, it is of great significance to remove antihistamines in wastewater before their discharging into natural waters.
Nowadays, the mainly used treatment technologies in WWTPs cannot effectively remove organic pollutants [10,11], especially PPCPs [12,13]. Antihistamines have been also ineffectively or inefficiently degraded by the traditionally used technologies due to their high concentration in the effluent from WWTPs [7,14]. Advanced oxidation processes (AOPs), including ozonation, ultraviolet (UV)/H2O2, etc., have been proven to be effective technologies for degrading refractory PPCPs, including antihistamines [5,15,16]. The antihistamines, e.g., cimetidine (CMD), cetirizine (CTZ), and diphenhydramine (DPD), have been shown to undergo degradation in TiO2-initiated photocatalysis [17], ozonation and peroxymonosulfate oxidation [5,18], and UV/H2O2 treatment [19], respectively.
Among various AOPs, UV/chlorine technology is attractive due to its efficient degradation of various pollutants [20,21,22]. In a UV/chlorine system, reactive species with high reactivity such as hydroxyl radicals (HO•) and chlorine free radicals (including Cl•, ClO•, and Cl2•−) were generated, which dominated the degradation of organic contaminants [23]. For example, Lei et al. [24] found that two representative antihistamines, cimetidine and famotidine, could undergo fast reaction with HO•, Cl•, and Cl2•−, and the reaction rate constants were 6.50 × 109 M−1 s−1, 1.46 × 1010 M−1 s−1, and 0.43 × 1010 M−1 s−1 and 1.72 × 1010 M−1 s−1, 2.78 × 109 M−1 s−1, and 1.65 × 109 M−1 s−1, respectively. Therefore, it can be reasonably speculated that antihistamines can be efficiently degraded in the UV/chlorine process. However, little attention has been paid on the removal of antihistamines. Furthermore, the contribution of these reactive species to the degradation of antihistamines in the UV/chlorine system is also unclear.
The degradation of PPCPs in a UV/chlorine system was reported to be influenced by aqueous environmental factors [25]. For example, pH can affect free chlorine’s dissociation and subsequently influence the generation of reactive species [26], and also change the reactivity of ionizable active pharmaceuticals with the generated reactive species during the UV/chlorine process [27]. During the UV/chlorine treatment, dissolved organic matter (DOM), which is a ubiquitous component in wastewater, was shown to inhibit the degradation of two PPCPs (naproxen and gemifibrozil) via inner filter effect and quenching of the generated reactive species [25]. The water matrix could not only influence the degradation efficiency but also affect the degradation pathways of PPCPs by changing the generation and presence of reactive species during the UV/chlorine treatment. The reactivity of the generated reactive species toward PPCPs was shown to be structure-dependent [28].
The reaction of the reactive species in a UV/chlorine system with PPCPs could lead to the formation of toxic by-products [29], and the formation of halogenated intermediates in the reactions of reactive chlorine radicals with PPCPs is a great concern in the use of the UV/chlorine technology [28]. Therefore, it is essential to investigate the detailed reaction pathways of the reactive species such as HO•, Cl•, Cl2•−, and ClO• with PPCPs. The reactions of these reactive species with organic pollutants mainly occur through single electron transfer, abstraction of hydrogen (H-abstraction), and addition pathways [30,31,32]. Quantum chemical calculation has been previously applied to investigate the reactions of reactive species with organic contaminants successfully [32,33], and the results demonstrated that H-abstraction and addition are the major reaction pathways [21]. Nevertheless, the reaction pathways of antihistamines with the generated reactive species in the UV/chlorine system are still unclear and worthy of urgent research.
Thus, the degradation of antihistamines in a UV/chlorine system was investigated with CTZ as a representative, which is among the most detected antihistamines with the highest concentrations in the effluent and surface water [7]. The effect of pH, initial concentration, and SRNOM (a representative of DOM) on the degradation of CTZ was revealed. Furthermore, the contribution of the reactive species on the degradation of CTZ was investigated, and the detailed reaction pathways of these reactive species with CTZ were calculated using quantum chemical calculation. The results of this study were helpful for providing alternative technology to remove antihistamines in wastewater as well as deep insight into the degradation mechanisms of emerging PPCPs in the UV/chlorine system.

2. Materials and Methods

2.1. Chemicals

CTZ (98%), sodium hypochlorite solution (available chlorine 5%), nitrobenzene (NB, 98%), and tert-butanol (TBA, chromatographical purity) were purchased from J&K Scientific Ltd. (Beijing, China). SRNOM (2R101N) was obtained from the International Humic Substances Society. Na2HPO4 and NaH2PO4 were purchased from Tianjin Damao Chemical (Tianjin, China). Methanol (chromatographicfal purity) was purchased from TEDIA (Fairfield, CT, USA). Ultrapure water was produced using an instrument purchased from Chengdu Ultrapure Technology Co., Ltd. (Chengdu, China).

2.2. UV/Chlorine Degradation Experiments

The experiments were performed in an OCRS-PX32T rotatable photochemical reactor that was obtained from Kaifeng HXsei Science Instrument Factory (Kaifeng, China) (Figure S1). The light source used in this study was a 500 W mercury lamp with main wavelengths of 254 nm and 297 nm, of which light intensity was determined to be (1.45 ± 0.07) mW cm−2 and (0.57 ± 0.03) mW cm−2 using a UV-B dual-channel ultraviolet radiation meter (Photoelectric Instrument Factory of Beijing Normal University, Beijing, China). During the treatment, the temperature was controlled at (25 ± 1) °C using a circulating cooling water system. All experiments were conducted in triplicate.
The initial concentration of CTZ and free chlorine was 10 µM and 100 μM, respectively (Phosphate buffer solution (PBS), pH 7.0). The UV-Vis absorption spectra of NaClO and CTZ are shown in Figure S2. UV/chlorine treatment was performed in the photo-reactor for 10 min. During the irradiation, dark control experiments were conducted with quartz tubes (30 mL) covered using aluminum foil. Samples were obtained and placed in dark conditions at room temperature during the UV/chlorine treatment. The concentration of residual chlorine was determined with the method that we used in our previous study [34].
The effects of the initial concentration of CTZ (2, 5, 10, 25, and 50 µM) and free chlorine (100, 200, 500, and 700 µM) on the degradation of CTZ was investigated by changing the concentrations. SRNOM was added with concentrations of 2.0, 5.0, 10.0, and 15.0 mg L−1 to investigate its effect on the degradation of CTZ. NB (2 µM) was used as the quencher of HO• [35], sodium bicarbonate (NaHCO3, 100 mM) was used to scavenge the HO•, Cl•, and Cl2•− [28], and TBA (100 mM) was employed as the quencher of HO•, Cl•, and ClO• [36]. During the degradation of CTZ, Cl− with high concentration of 50 mM was added to investigate the roles of Cl• and Cl2•− as it can convert Cl• to Cl2•− [27].

2.3. Analytical Methods

An Agilent 1260 II HPLC (Agilent Technologies Inc., Santa Clara, CA, USA) with a diode array detector was used to quantify CTZ. During the quantification, a Welch Ultimate™ AQ-C18 (250 mm × 4.6 mm, 5 m) analysis column (Welch Materials Inc., Maryland, USA) was used. The mobile phase for the detection of CTZ was methanol and 0.1% phosphoric acid solution aqueous (pH 2.3) with a ratio of 60:40, and the wavelength selected for the detection of CTZ was 200 nm.

2.4. Quantum Chemical Calculation Methods

All quantum chemical calculation was performed with Gaussian 16 software [37]. The optimization and single-point energy calculation were performed using density functional theory (DFT) with the function of B3LYP. The optimization was performed at the level of 6-31+G(d, p), and the single-point energy calculation was performed at the level of 6-311++G(3df, 2p). The integral equation form of the polarization continuum model (IEFPCM) was used to consider the solvent effect of water. The transition state (TS) was obtained and characterized by the virtual vibration frequency (only one), and was verified with intrinsic reaction coordinate (IRC) analysis. The thermodynamic energies, i.e., Gibbs free energy and enthalpy, were obtained and zero-point correction was performed while calculating these energies.

3. Results and Discussion

3.1. Degradation of CTZ and Influencing Factors during UV/Chlorine Treatment

During direct UV irradiation, obvious degradation of CTZ was observed and no obvious degradation in the dark controls was observed (Figure 1). These results indicate that CTZ was photodegradable under direct UV irradiation, and the degradation of CTZ followed pseudo first-order kinetics. The observed degradation rate constant (kobs) of CTZ under direct UV irradiation was calculated to be (0.08 ± 0.01) min−1. In a previous study, the photodegradation of CTZ under UV irradiation was also observed [38]. In the UV/chlorine system, the degradation rate of CTZ was faster than that under direct UV irradiation (p < 0.05, Figure 1). The kobs of CTZ increased to (0.19 ± 0.01), (0.29 ± 0.02), (0.42 ± 0.03), and (0.50 ± 0.02) min−1 with concentrations of free chlorine increased to 100, 200, 500, and 700 µM, respectively. These results demonstrated the important role of free chlorine in the removal of CTZ. The degradation efficiency reached up to 84.9% within 10 min of treatment and exceeded 90% within 15 min of treatment at the free chlorine concentration of 100 µM in the UV/chlorine system. Thus, the free chlorine concentration of 100 µM was selected in the following experiments.
In the UV/chlorine system, the initial concentration of organic pollutants was shown to affect their degradation kinetics and efficiencies [39]. Therefore, the effect of initial concentrations of CTZ on its degradation was studied, and the results are exhibited in Figure 2 and Figure S3 in the Supplementary Materials. As shown in Figure 2, the kobs of CTZ showed a first-increase and then-decrease trend under UV irradiation as well as the UV/chlorine treatment. The fastest degradation rate of CTZ was observed with its initial concentration of 10 µM. With the increase in initial concentration of CTZ from 10 µM to 50 µM, the kobs of CTZ under direct UV irradiation and in the UV/chlorine system decreased 36.4% and 64.4%, respectively.
The effect of initial concentration on the degradation of CTZ under direct UV irradiation was attributed to the competitive light absorption effect at the CTZ with high concentrations. The differences in kobs values in the UV/chlorine system compared with those under direct UV irradiation also showed a first-increase and then-decrease trend (Figure 2), indicating that the concentration-dependence in the degradation of CTZ was not only attributed to the UV-induced degradation but also the competing reactions to the generated reactive species.
The influence of pH on the degradation of CTZ was studied and the results showed that the kobs of CTZ decreased with the increase in pH in the UV/chlorine system (Figure 3 and Figure S4 in the Supplementary Materials). There were two possible mechanisms for the pH-dependence of CTZ degradation: (1) the different reactivities of the generated reactive species toward CTZ of different forms, i.e., neutral form and anionic form; (2) the dissociation of HOCl/OCl− with pKa of 7.5 [40] led to different concentrations of the generated reactive species [41]. In this case, the former was not suitable as more than 99.7% CTZ (pKa = 3.5) existed with anionic form in the solutions with pH above 6.0. Thus, in the UV/chlorine system, the decrease in kobs of CTZ with the increase in pH was due to the dissociation of HOCl to OCl− as the generation quantum yield of reactive species was higher and radical scavenging effect was weaker for HOCl compared with those of OCl− [42].
The DOM in the wastewater, which was shown to significantly inhibit the removal of primidone and caffeine in the UV/chlorine system, is considered to be a great defect during the practical application of UV/chlorine technology [39]. As exhibited in Figure 3 and Figure S5 in the Supplementary Materials, in the UV/chlorine system, the presence of SRNOM obviously decreased the kobs of CTZ (p < 0.05), and the inhibitory effect of SRNOM increased with the increase in its concentration. The kobs of CTZ decreased from (0.19 ± 0.01) min−1 without SRNOM to (0.18 ± 0.01), (0.16 ± 0.01), (0.14 ± 0.01), and (0.12 ± 0.01) min−1 in the presence of SRNOM with concentrations of 2.0, 5.0, 10.0, and 15.0 mg L−1, respectively, of which decreased rates were 7.3%, 17.8%, 28.4%, and 38.9%, respectively. The inhibition on the degradation of CTZ induced by SRNOM was attributed to: (1) the light screening effect of SRNOM as it can absorb light with wavelengths from 200 to 400 nm (Figure S1); (2) the competition with reactive radicals as reported in a previous study [39]. Thus, it could be concluded that the DOM in wastewater is an unfavorable factor during the degradation of antihistamines in wastewater using UV/chlorine technology.

3.2. Roles of Reactive Species in Degradation of CTZ

In the UV/chlorine system, the underlying degradation mechanisms of CTZ were revealed by performing quenching experiments. As can be seen in Figure 4, the presence of Cl− significantly decreased the kobs of CTZ in the UV/chlorine system (p < 0.05), indicating the involvement of Cl• in the degradation of CTZ as Cl− can quench Cl• to generate Cl2•− rapidly [27]. In the presence of NB, the kobs of CTZ decreased to (0.12 ± 0.01) min−1 from (0.19 ± 0.01) min−1 in the PBS (p < 0.05), which indicated the involvement of HO• during the degradation of CTZ. The presence of HCO3−, which could quench HO•, Cl•, and Cl2•− [28], decreased the kobs of CTZ to (0.11 ± 0.01) min−1 (p < 0.05). The contribution of Cl• and Cl2•− (kobs(NB)−kobs(HCO3−)) was comparable with that of Cl• (kobs(PBS)−kobs(Cl−)) in the degradation of CTZ, which demonstrated the negligible effect of Cl2•− on the degradation of CTZ. The addition of TBA, which quench HO•, Cl•, and ClO• [36], further decreased the kobs of CTZ compared with that of HCO3−, which indicated the involvement of ClO• in the degradation of CTZ.
In the UV/chlorine system, the presence of SRNOM could inhibit the degradation of CTZ (Figure 3). Thus, the involvement of the reactive species in the degradation of CTZ in SRNOM solutions was also investigated. The addition the quenchers decreased the kobs of CTZ in the SRNOM solutions (Figure 5), which was similar with those in the PBS. Thus, in the presence of SRNOM, HO•, Cl•, and ClO• were also involved in the degradation of CTZ.
Contribution ratios of the reactive species (HO•, Cl•, and ClO•) and direct UV irradiation to the degradation of CTZ were calculated, and the values are shown in Table S1 in the SI. The contribution ratio of UV-induced degradation of CTZ was 38.7% in the PBS. For the reactive species, the contribution ratio of HO•, Cl•, and ClO• was 35.3%, 7.3%, and 17.1%, respectively. The remaining 1.6% degradation of CTZ was attributed to other reactive species. Among these reactive species, HO• played a crucial role during the degradation of CTZ. In SRNOM solutions, the contribution ratio of direct UV-induced degradation of CTZ increased to 42.2% compared with that in the PBS. The contribution ratio of HO• and Cl• decreased to 29.6% and 4.8%, respectively, compared with the values in the PBS. However, the contribution ratio of ClO• increased to 21.2% from 17.1%. These results demonstrated the different role of SRNOM on the quenching of these reactive species.

3.3. Degradation Pathways of CTZ in UV/Chlorine System

Considering the complexity of the reactions for CTZ in the UV/chlorine system, the DFT method was used to investigate the degradation pathways of CTZ in the UV/chlorine system. In wastewater, CTZ mainly exists in anionic form as it is with low pKa of 3.5. Therefore, the reaction pathways of anionic CTZ with the important reactive species (HO•, Cl•, and ClO•) were calculated with DFT methods. As shown in a previous study, major reactions of these reactive species with small organic chemicals are through H-abstraction and addition pathways [21]. For anionic CTZ, there are 24 H-abstraction and 12 addition pathways (Figure 6). All these possible reaction pathways of HO•, Cl•, and ClO• with CTZ were calculated, and the calculated values of Gibbs free energy change (ΔG), activation free energy (ΔG‡), and enthalpy change (ΔH) for these reaction pathways are listed in Tables S2 and S3. The reaction complexes, transition states, and reaction intermediates of these reactions are shown in Figures S6–S10 in the Supplementary Materials.
For the HO•-initiated H-abstraction and addition reactions, the values of ΔG and ΔH were all < 0 (Figure 6 and Table S2), implying that H-abstraction from CTZ and HO• addition to the unsaturated bonds of the phenyl group of CTZ were all thermodynamically spontaneous. The values of ΔG‡ for the addition reactions were from 0.87 to 5.59 kcal mol−1, indicating that these reactions were also dynamically possible. The HO• additions on the C16 and C20 sites were with low ΔG‡ values (0.87 and 0.89 kcal mol−1, respectively), which indicated that the addition reactions at the two sites were more favorable. For the H-abstraction reactions, the ΔG‡ values for the pathways from the phenyl group (H41-49) and α-carbon (α-C) are from 1.72 to 10.52 kcal mol−1, and the pathways from the hydrocarbon chain (except that from the α-C) were all barrierless, which indicated that the H-abstraction reaction from the hydrocarbon chain of CTZ was more favorable. These results demonstrated the high reactivity of HO• with CTZ, which is in accordance with the extremely high reaction rate constant of HO• with CTZ (8.94 × 109 M−1 s−1) [43]. Meanwhile, these results are also in agreement with the experimental results that HO• plays an important role during the degradation of CTZ in the UV/chlorine system.
For the ClO• addition pathways, only the ΔG and ΔH for the reaction at the C24 site were negative (Figure 7 and Table S3), indicating that ClO• addition to the unsaturated bonds at the C24 site was thermodynamically spontaneous and exothermic. Thus, although the ΔG‡ values of the addition reaction at the C24 site (17.21 kcal mol−1) were relatively higher compared with that at other sites, ClO• addition reaction could only occur spontaneously at this site. This was due to the generation of a thermodynamically more stable intermediate with an oxygen atom on C24 instead of Cl. The H-abstraction pathways on the phenyl group of CTZ were all thermodynamically nonspontaneous and endothermic as they were all with positive ΔG and ΔH (Table S3). The abstraction reactions of H39, H40, H50, and H51 at the hydrocarbon chain were thermodynamically spontaneous and exothermic due to their negative ΔG and ΔH (Figure 7 and Table S3). Among these reactions, the abstraction of H50 and H51 were more favorable due to the lower ΔG‡ values (2.22 and 1.83 kcal mol−1, respectively).
For the reaction of CTZ with Cl•, only H-abstraction of H36 was with negative ΔG (−3.08 kcal mol−1), and of which ΔG‡ was 9.54 kcal mol−1, implying that H-abstraction of H36 was thermodynamically spontaneous and dynamically possible. The reaction complex, transition state, and intermediate of this reaction are shown in Figure S11. The results implied that the Cl• showed low reactivity toward CTZ, which is consistent with the experimental results that Cl• played a minor role (7.3%) in the degradation of CTZ in the UV/chlorine system.
For the reactions of CTZ with HO•, Cl•, and ClO•, H-abstraction is the main reaction pathway, which leads to the generation of C-centered radicals. C-centered radicals undergo subsequent reactions with dissolved O2 or H2O to form hydroxylated products [30,32]. Addition reaction of HO• and ClO• with CTZ could also lead to the generation of hydroxylated products and the dechlorinated product. Meanwhile, no chlorinated products were generated during the reaction of these reactive species with CTZ. Thus, the degradation of CTZ during the UV/chlorine treatment could decrease the toxicity induced by CTZ.

4. Conclusions

The degradation of an emerging organic pollutant CTZ in the UV/chlorine system was investigated and the underlying reaction mechanism was revealed by combining the methods of experiment and quantum chemical calculation. The results demonstrated that CTZ could be removed efficiently in the UV/chlorine system, of which degradation was initial-concentration dependent. In the UV/chlorine system, besides direct UV-induced degradation, the generated HO•, Cl•, and ClO• contributed greatly to the degradation of CTZ, among which HO• played a crucial role in the degradation of CTZ. The HO• and ClO• initiated the degradation of CTZ mainly through addition and H-abstraction reactions, and the Cl• could only initiate H-abstraction reaction.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w14091323/s1, Figure S1: Photochemical reactor system diagram; Figure S2: UV-Vis absorption spectra of NaClO, CTZ, and SRNOM; Figure S3: Degradation kinetics of CTZ with different concentrations in UV/chlorine system; Figure S4: Degradation kinetics of CTZ in UV/chlorine system with different pH values; Figure S5: Degradation kinetics of CTZ in UV/chlorine system in the presence of SRNOM with different concentration; Figure S6: Reaction complexes, transition states and intermediates of the addition reaction pathways of CTZ with HO•; Figure S7: Reaction intermediates of barrierless H-abstraction reaction pathways of CTZ with HO•; Figure S8: Reaction complexes, transition states and intermediates of the hydrogen abstraction pathways of CTZ with HO•; Figure S9: Reaction complexes, transition states and intermediates of the addition reaction pathways of CTZ with ClO•; Figure S10: Reaction complexes, transition states and intermediates of the hydrogen abstraction reaction pathways of CTZ with ClO•; Figure S11: Reaction complexes, transition states and intermediates of the hydrogen abstraction reaction of H36 of CTZ with Cl•; Table S1: Contribution ratio of different reactive species and UV degradation to the degradation of CTZ in UV/chlorine system; Table S2: Calculated Gibbs free energy change (ΔG), enthalpy change (ΔH) and activation free energy (ΔG‡) values for possible reaction pathways of HO• with CTZ; Table S3: Calculated Gibbs free energy change (ΔG), enthalpy change (ΔH) and activation free energy (ΔG‡) values for possible reaction pathways of ClO• with CTZ (in kcal mol−1).

Author Contributions

Conceptualization, Y.-n.Z. and H.Y.; methodology, B.Z. and F.C.; software, B.Z. and F.C.; validation, B.Z., F.C. and W.Z.; formal analysis, B.Z.; investigation, B.Z., F.C. and W.Z.; resources, J.Q., Y.-n.Z. and H.Y.; data curation, B.Z. and F.C.; writing—original draft B.Z.; writing—review and editing, Y.-n.Z. and H.Y.; supervision, Y.-n.Z. and H.Y.; funding acquisition, J.Q., Y.-n.Z. and H.Y. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the National Natural Science Foundation of China, grant number 22176030, 42130705, and 21976027; the Fundamental Research Funds for the Central Universities, grant number 2412019FZ019 and 2412020FZ015, and the Jilin Province Science and Technology Development Projects, grant number 20210101110JC.

Acknowledgments

This work was supported by the National Natural Science Foundation of China (22176030, 42130705, and 21976027), the Fundamental Research Funds for the Central Universities (2412019FZ019 and 2412020FZ015), and the Jilin Province Science and Technology Development Projects (20210101110JC).

Conflicts of Interest

The authors declare no conflict of interest.

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Water 14 01323 g001 550
Figure 1. Degradation kinetics of CTZ in UV/chlorine system with different concentration of free chlorine (pH = 7.0, [CTZ]0 = 10 µM. The error bars represent the 95% confidence interval (n = 3)).
Figure 1. Degradation kinetics of CTZ in UV/chlorine system with different concentration of free chlorine (pH = 7.0, [CTZ]0 = 10 µM. The error bars represent the 95% confidence interval (n = 3)).
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Figure 2. Observed degradation rate constant (kobs) of CTZ with different initial concentration in UV (circle) and UV/chlorine (square) system (pH = 7.0, (Free chlorine)0 = 100 µM. The insert figure represents the differences in kobs values in UV/chlorine system compared with those under UV treatment alone. The error bars represent the 95% confidence interval (n = 3)).
Figure 2. Observed degradation rate constant (kobs) of CTZ with different initial concentration in UV (circle) and UV/chlorine (square) system (pH = 7.0, (Free chlorine)0 = 100 µM. The insert figure represents the differences in kobs values in UV/chlorine system compared with those under UV treatment alone. The error bars represent the 95% confidence interval (n = 3)).
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Figure 3. Observed degradation rate constant (kobs) of CTZ in UV/chlorine system in different pH or in the presence of SRNOM with different initial concentrations ([CTZ]0 = 10 µM, (Free chlorine)0 = 100 µM. The error bars represent the 95% confidence interval (n = 3)).
Figure 3. Observed degradation rate constant (kobs) of CTZ in UV/chlorine system in different pH or in the presence of SRNOM with different initial concentrations ([CTZ]0 = 10 µM, (Free chlorine)0 = 100 µM. The error bars represent the 95% confidence interval (n = 3)).
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Figure 4. Observed degradation rate constant (kobs) of CTZ (10 µM) in UV/chlorine system ((Free chlorine)0 = 100 µM) under different conditions (PBS, pH = 7.0; (Cl−) = 50 mM; (NB (nitrobenzene)) = 2 µM; (HCO3−) = 100 mM; (TBA (tert-butanol)) = 100 mM) and in UV system as a control (the error bars represent the 95% confidence interval (n = 3)).
Figure 4. Observed degradation rate constant (kobs) of CTZ (10 µM) in UV/chlorine system ((Free chlorine)0 = 100 µM) under different conditions (PBS, pH = 7.0; (Cl−) = 50 mM; (NB (nitrobenzene)) = 2 µM; (HCO3−) = 100 mM; (TBA (tert-butanol)) = 100 mM) and in UV system as a control (the error bars represent the 95% confidence interval (n = 3)).
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Figure 5. Observed degradation rate constant (kobs) of CTZ (10 µM) in the presence of SRNOM (5 mg L−) in UV/chlorine system ((Free chlorine)0 = 100 µM) under different conditions (PBS, pH = 7.0; (Cl−) = 50 mM; (NB (nitrobenzene)) = 2 µM; (HCO3−) = 100 mM; (TBA (tert-butanol)) = 100 mM) and in UV system as a control (the error bars represent the 95% confidence interval (n = 3)).
Figure 5. Observed degradation rate constant (kobs) of CTZ (10 µM) in the presence of SRNOM (5 mg L−) in UV/chlorine system ((Free chlorine)0 = 100 µM) under different conditions (PBS, pH = 7.0; (Cl−) = 50 mM; (NB (nitrobenzene)) = 2 µM; (HCO3−) = 100 mM; (TBA (tert-butanol)) = 100 mM) and in UV system as a control (the error bars represent the 95% confidence interval (n = 3)).
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Figure 6. Addition and H-abstraction reaction pathways for reaction of CTZ with HO• (ΔG: Gibbs free energy change; ΔG ‡: activation free energy; unit in kcal mol−1).
Figure 6. Addition and H-abstraction reaction pathways for reaction of CTZ with HO• (ΔG: Gibbs free energy change; ΔG ‡: activation free energy; unit in kcal mol−1).
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Figure 7. Detailed addition (C24 site) and H-abstraction (H39, H40, H50, and H51) pathways for the reaction of CTZ + ClO• (the values in the brackets are length of bonds marked with the dashed lines, and the unit is Å; TS and IM denote transition state and intermediate, respectively).
Figure 7. Detailed addition (C24 site) and H-abstraction (H39, H40, H50, and H51) pathways for the reaction of CTZ + ClO• (the values in the brackets are length of bonds marked with the dashed lines, and the unit is Å; TS and IM denote transition state and intermediate, respectively).
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Zhu, B.; Cheng, F.; Zhong, W.; Qu, J.; Zhang, Y.-n.; Yu, H. Mechanistic Insight into Degradation of Cetirizine under UV/Chlorine Treatment: Experimental and Quantum Chemical Studies. Water 2022, 14, 1323. https://doi.org/10.3390/w14091323

AMA Style

Zhu B, Cheng F, Zhong W, Qu J, Zhang Y-n, Yu H. Mechanistic Insight into Degradation of Cetirizine under UV/Chlorine Treatment: Experimental and Quantum Chemical Studies. Water. 2022; 14(9):1323. https://doi.org/10.3390/w14091323

Chicago/Turabian Style

Zhu, Boyi, Fangyuan Cheng, Wenjing Zhong, Jiao Qu, Ya-nan Zhang, and Hongbin Yu. 2022. "Mechanistic Insight into Degradation of Cetirizine under UV/Chlorine Treatment: Experimental and Quantum Chemical Studies" Water 14, no. 9: 1323. https://doi.org/10.3390/w14091323

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