Degradation of 17β-estradiol by UV/persulfate in different water samples

Endocrine disrupting chemicals (EDCs) are a diverse group of natural and synthetic organic substances which cause risk to human beings and animals via influencing the processes and functions of the endocrine system (Geyer et al. 2000; Mai et al. 2008; Mazellier et al. 2008; Zhao et al. 2008; Zhang et al. 2013; Ahmad et al. 2015; Bajard et al. 2021). Among the EDCs, the main synthetic chemicals include pesticides, plastics, and other industry-related materials (such as bisphenol A) produced by humans, natural compounds such as estrone (E1), 17β-estradiol (E2) released by humans and animals (Irmak et al. 2005; Yoon et al. 2020). Over the past few years, EDCs have been frequently detected at a level from picograms per liter to nanograms per liter in surface waters and sewage effluent (Zhao et al. 2008; Brienza et al. 2014; Park et al. 2018). Furthermore, recent studies have demonstrated that EDCs have been identified as having the highest endocrine disrupting activity, even at very low concentrations (ng/L) in the aquatic environment and are not commonly regulated (Zhang et al. 2013). For example, a Canadian field study showed that long-term exposure to EDCs at ambient concentrations (5–6 ng/L) can lead to feminization of wild fish in lake waters and eventually cause fish population decline (Kidd et al. 2007).

As a representative EDC, E2, an extremely potent estrogen receptor modulator, is widely used to treat advanced prostate cancer and estrogen deficiency (Lee & Peart 1998; Servos et al. 2005; Weber et al. 2005; Mispagel et al. 2009). As well, E2 has also been used as a feed supplement to aquaculture, with great success (Pechsiri & Yakupitiyage 2005; Flynn & Benfey 2007). Due to its extensive use, many studies have reported that E2 with heavy concentrations has been detected in marine and freshwater systems or drinking water in various countries (Mai et al. 2008; Sun et al. 2017; Jiang et al. 2018; Ma & Yates 2018; Mauricio et al. 2018). For example, Duong's research shows that an average concentration of 9.1 ng/L for E2 has been detected in the surface water of China (Duong et al. 2010).

Furthermore, the potentially detrimental effects of E2 on human health have caused increasing concern, with some studies suggesting that the toxicity of E2 may transfer from the parental generation to subsequent generations (Pedram et al. 2006; Bradley et al. 2009; Cajthaml 2015). Due to its impact both on public and ecological health concerns, research on water treatment methods regarding the removal of E2 is a critically important issue (Bila et al. 2007; Maniero et al. 2008; Jiang et al. 2012; Ting et al. 2017). However, E2 can resist the degradation of conventional sewage treatment operation, with typical chemical and biological treatment methods not adequately achieving removal of its biological activity (Bradley et al. 2009; Jiang et al. 2010; Mashtare et al. 2013). As reported, the concentration of E2 in effluent of wastewater treatment plants (WWTP) is 0.1–88 ng/L which is much higher than its typical detection (Li et al. 2016).

In the past few years, advanced oxidation processes (AOPs) have been investigated to attenuate EDCs from various types of water because these techniques are considered to be an efficient method to destroy the biological activity of EDCs (Wu et al. 2014; Ahmad et al. 2015; Lian et al. 2017; Joao et al. 2021). EDCs' degradation by AOPs, such as ozonation, photo-Fenton reaction, and UV/hydrogen peroxide (UV/H₂O₂) has been well investigated (Gibson et al. 2005; Zhao et al. 2008; Zhang et al. 2010; Nakrst et al. 2011). Among the AOPs, persulfate (PS), which generates the highly reactive sulfate free radical (•SO₄⁻) that can rapidly degrade the wide variety of compounds to nontoxic CO₂ and H₂O, has gained considerable attention (Ahmad et al. 2015; Lin & Tsai 2019; Xiong & Pei 2021). PS can completely decompose EDCs due to its strong redox potential (2.5–3.1 V) which is stronger than that of •OH (1.9–2.7 V) (Lian et al. 2017). Moreover, PS can be activated by a series of methods, including heat, UV irradiation or transition metal to catalysis (Fu et al. 2019). As typical UV-based AOPs, UV/Na₂S₂O₈ (UV/PS) processes not only exhibit great potential removal of refractory organics with less energy consumption, but also do not cause secondary pollution (Zhang et al. 2018). For instance, Ye et al. (2016) found that UV/PS treatment was an effective method to degrade ciprofloxacin (CIP). The time of UV/PS complete degradation of CIP was no longer than 20 minutes and the apparent rate constants 0.2413 min⁻¹.

Degradation of other man-made organic compounds by UV/PS have been studied previously and the results show that the reaction process is capable of oxidizing the vast majority of organic contaminants (Mai et al. 2008). Nonetheless, to the best of our knowledge, few studies have been done with the focus on the degradation of E2 by UV/PS, and using the UV/PS process in marine aquaculture water treatment has yet to be performed. Accordingly, the aim of this investigation is to elucidate the reaction kinetics, mechanisms, and degradation pathways of the UV/PS process in degrading E2 in marine aquaculture water. The fate of E2 in marine aquaculture water was analyzed in the presence of potential interfering factors such as anions (Br⁻, Cl⁻, HCO₃⁻, SO₄²⁻), cation (Na⁺, Mg²⁺, K⁺), and pH value. Furthermore, the degradation pathways of E2 were explored based on the main intermediates identified by LC/MS/MS. As well, the degradation pathways of E2 were investigated based on the main intermediates identified by LC/MS/MS. The inorganic disinfection by-product, namely, bromate (BrO₃⁻) of E2 treated by UV/PS was also explored.

E2 (>98% purity) was purchased from Sigma-Aldrich (St. Louis, MO, USA) and used without further purification in the experiments. Both sodium persulfate (Na₂S₂O₈) and sodium sulfite (Na₂SO₃) were analytical grade and obtained from Tianjin Komeo Chemical Reagent Company (Tianjin, China). High performance liquid chromatography (HPLC) grade methanol was supplied by Anhui Fulltime Specialized Solvents & Reagents Co., Ltd (China). Other chemicals were analytical grade without further purification and purchased from China Pharmaceutical Group Chemical Reagents Co., Ltd (China). Deionized water was obtained from a Millipore Milli-Q Ul-trapure Gradient A10 purification system. Water samples used in the experiment containing 0.1 mmol/L E2 were prepared by spiking 100 mL E2 stock solution (1 mmol/L in methanol) into 1 L deionized water. According to the components of actual water samples, the synthetic sea water was prepared by adding essential anions into ultrapure water including Cl⁻ (6.6 g/L), Br⁻ (22 mg/L), HCO₃⁻ (0.047 g/L), and SO₄²⁻ (0.903 g/L). In the actual marine aquaculture, the water was a mixture of freshwater and seawater at a ratio of 2:1. Therefore, the salinity and various ion concentrations of marine aquaculture water were one-third of seawater. The pH of all water samples was adjusted to 7 with phosphate buffer.

E2 degradation experiments were carried out in a photoreactor (BL-GHX-V, Shanghai Bilang Instruments Co., Ltd, China). The UV light source was a high-pressure mercury lamp (254 nm) equipped with a quartz cover which was vertically fixed in the center of the reactor. Eight quartz tubes with a height of 18.5 cm and an internal diameter of 2.0 cm were evenly placed around the high-pressure mercury lamp. An agitator was placed inside each quartz tube ensuring complete mixing of the reaction solution. A steady flow of cooling system was applied to maintain the temperature of reaction solutions at about 25 °C during the irradiation process.

The TOC was measured using a Liquid TOC analyzer (TOC-VSH, Shimazu), and the determination of bromate (BrO₃⁻) was performed using an ion chromatography system (Dionex ICS-5000, Thermo USA).

Under acidic conditions, the generation of •S₂O₈⁻ in water samples may result in decrease of •SO₄⁻ contents in accordance with Equation (2). However, the higher •S₂O₈⁻ generation rate may cause radical with radical scavenger reactions, which could result in inhibition of the reactivity of •SO₄⁻. A similar inhibition mechanism was observed at an alkaline solution as shown in Equation (3). As reported, •SO₄⁻ has a longer oxidation time than •OH; •SO₄⁻ has a longer oxidation time on pollution and can improve the degradation effect. Under the same reaction conditions, •SO₄⁻ has a higher degradation degree on compounds than •OH. With the reaction in alkaline solution, •OH gradually increases in the reaction system, and the content of •SO₄⁻ decreases gradually. Compared with •OH, it was easier for •SO₄⁻ to cause electron transfer on E2 and it selectively oxidized E2. The above oxidation mechanism leads to •SO₄⁻ degrading E2 faster than •OH. On the other hand, when the system was under alkaline conditions, it resulted in an increase in the production of •OH which affects the content of •SO₄⁻. Therefore, the degradation of E2 is slowed down by the mass production of •OH under strong alkali conditions.

It can be seen from Figure 2(b) that within half an hour in the UV/PS system, the degradation of E2 was less than 90% in the absence of Br⁻; however, the degradation of E2 reached over 99% when the concentration of Br⁻ increased to 66 mg/L. This suggests that the presence of Br⁻ can accelerate the rate of UV/PS degradation of E2. It can also be seen from Figure 2(b) that when a high concentration of Br⁻ is present in the reaction system, the reaction rate is greater than that of the reaction system containing no Br⁻, and the reaction rate increases as Br⁻ increases. The acceleration mechanism for this phenomenon is similar to that of Cl⁻. When the UV/PS reaction system contains a higher concentration of Br⁻, HBrO is formed in the system, thereby accelerating the degradation of E2.

In addition to the Br⁻ and Cl⁻, there are other matters such as HCO₃⁻ with high concentration in seawater which may also affect the rate at which the UV/PS system degrades E2. As shown in Figure 2(c), when 142 mg/L of HCO₃⁻ is present in the reaction system, the degradation rate of E2 has reached about 55% at 15 min in the experiment. At the same time, when in the reaction system 47 mg/L of HCO₃⁻ is present, the degradation rate of E2 is about 78%. Correspondingly, in the system without HCO₃⁻, the degradation rate of E2 is about 80% when the reaction is carried out for 15 min. Figure 2(c) also shows that as the corresponding concentration of HCO₃⁻ in the reaction system increases, the degradation rate of E2 in the UV/PS system decreases. It is possible that the pH of the system changed due to the addition of HCO₃⁻. As HCO₃⁻ in the reaction system increased, the pH gradually increased. As shown in Equation (3), the conversion of •SO₄⁻ to •OH led to lower reaction rate in the reaction system when in the alkaline condition. The presence of higher levels of SO₄²⁻ in seawater may influence the reactions. As shown in Figure 3(a), SO₄²⁻ did not exhibit an obvious effect on the degradation of E2 in the UV/PS system. On the other hand, the effect of cations present in seawater on the degradation of E2 in UV/PS systems is also discussed in this paper, and similar to the effect of SO₄²⁻, cations such as Na⁺, Ca²⁺, Mg²⁺, and K⁺ have almost no effect on the reaction rate (Figure 3(b)).

The process of oxidative degradation of E2 by UV/PS system also is a mineralization process. In order to determine the degree of mineralization of E2, the ability of UV/PS to degrade E2 was explored. In this study, TOC removal rate of UV/PS degradation of E2 was studied in brackish water. As shown in Figure 4, when the reaction time increases, the TOC removal rate increases gradually. The mineralization rate of E2 reaches 22.16% at 1 h. Then, the reaction proceeds for 2 h, 4 h, 5 h, and 8 h, and the mineralization rates were 45.47%, 65.31%, 72.76%, and 90.97%, respectively.

Since the most important oxidants in the UV/PS system are •SO₄⁻ and •OH, which have strong oxidizing properties, the results show that although the reaction rate of UV/PS degrading E2 is affected by the anions, the reaction pathways were substantially the same. On the other hand, computational chemistry based on quantum theory has become an important analytical method in this paper. According to the frontier orbital theory, the highest energy level (HOMO) of the E2 molecule and the lowest energy level (LUMO) of the non-occupied electron are the keys to determining the chemical reaction (Aihara 1999). The position of the material having a large charge density is most susceptible to attack by electrophiles, causing the chemical bonds associated with it to break (Zhan et al. 2003).

The HOMO, LUMO, and net charge distribution of E2 are shown in Figure 5. The HOMO of the E2 molecule and the electron cloud on the LUMO are concentrated in the benzene ring. It can be inferred that the benzene ring structure is both the electron donor of the E2 molecule and the main acceptor part of the electron (Hu et al. 2002). The HOMO of a molecule is more susceptible to attack by electrophiles to generate an electrophilic reaction, and the nucleophile is more likely to attack the atom with a higher charge density on the LUMO (Serna-Galvis et al. 2017). This paper calculates the atomic net charge distribution in the E2 molecular structure, and the main reactive sites on E2 are inferred. From the microscopic level, the mode of action of •SO₄⁻ and E2 is combined with the high-resolution LC/MS/MS first-order mass spectrum display. The intermediate product has an accurate molecular mass that is inferred from the possible pathways of the E2 degradation process.

The UV/PS degradation of E2 in fresh and brackish water was carried out for 10 min, 30 min, and 2 h, and the mixture was analyzed by full sweep in positive ion mode via LC/MS/MS. Three more obvious product peaks were found in the two solutions for 2 h, named P1, P2, and P3. In addition, a total of six distinct peaks were found to be named Pm1–Pm6 in the solution in which the reaction was carried out for 10 and 30 min, respectively. Information about these intermediate products of the mass spectrum is shown in Supplementary Material, Figure S1(a)–(j).

The intermediate by-products of E2 were estimated by the molecule mass difference and isotope patterns in the mass spectra. The mass spectral characterized data for these organic products were confirmed by LC/MS which is shown in Supplementary Material, Figure S1. Based on the organic intermediates, two pathways (A and B) are extrapolated as shown in Figure 6. In pathway A, the hydroxyl group of the E2 five-membered ring is oxidized to form E1 (m/z = 270), and in the process of oxidation of E1, the two paths of A1 and A2 are, respectively, formed to produce two products of P1 (m/z = 252) and P2 (m/z = 196). In the pathway A1, the C6 position on E1 is oxidized to a hydroxyl group to form Pm3 (m/z = 286), and Pm3 continues to be oxidized by oxygenation to dehydrogenate to P1. It is presumed that the chemical bond between C10 and C5 of the benzene ring of E2 is broken by oxidation. The chemical bond between C2 and C4 also ruptures, and the chemical bond between C5 and C6 on the six-membered ring also breaks, resulting in the formation of a five-membered ring by C6 and C10, and the oxygenation reaction at the C1 position to forming P1. In the pathway A2, on the basis of the structural characteristics of E1, the chemical bond between C1 and C10, C4 and C5 is broken to remove the benzene ring, thereby forming Pm4 (m/z = 206), and the alkyl group on Pm4 undergoes oxygenation and dehydrogenation to form Pm6 (m/z = 222). Pm6 continues to oxidize and dehydrogenate to oxidize the alkyl group on the original five-membered ring to a carboxyl group to form Pm7 (m/z = 236). The decarboxylation reaction that occurred at the C13 position on the five-membered ring and the five-membered ring was cleaved at the C17 and C14 positions, and the hydroxyl group linked to C17 was oxidized to the carboxyl group, resulting in the production of P2.

In pathway B, P3 (m/z = 334) with a retention time of 5.20 min, the spectrum is shown in Supplementary Material, Figure S1(c). First, the C2 site on the E2 benzene ring is electrophilically attacked to form a hydroxyl group to form Pm2 (m/z = 288). Then, the site connecting the hydroxyl group at the benzene ring site of Pm2 is oxidized to form a carboxyl group and simultaneously causes a chemical bond between C2 and C3. Moreover, E2 electron cloud distribution characteristics are also known to be susceptible to electro-electrical attack at the C6 site (Hu et al. 2003). A hydroxyl group formed at the C6 site during the subsequent oxidative degradation process which led to forming Pm5. The hydroxyl group in Pm5 undergoes a dehydrogenation reaction to become a carbonyl group to form P3. The analysis of the reaction sites of E2 in the degradation process, combined with mass frontier software led to speculating the possible degradation pathways. The study found that the above intermediate products were gradually broken during the degradation reaction, and the fragments with smaller mass fractions (Supplementary Material, Figure S1(k)–(q)) were broken. The mass spectrum of each product's fragment is shown in Figure 5, and E2 eventually mineralized into H₂O and CO₂.

As reported, when Br⁻ is present in the solution, UV/PS will produce BrO₃⁻ while degrading the contaminants (Li et al. 2014). This paper preliminarily explores the general rule of BrO₃⁻ generation during the degradation of E2 by UV/PS technology in a brackish water system. This study found that BrO₃⁻ was not detected after 2 h of degradation of E2 by UV/PS in a brackish water system. It has been reported that the formation of BrO₃⁻ during UV/PS oxidation is inhibited in the presence of humic acid. Thus, this paper explores the formation of BrO₃⁻ in a brackish water system without E2. As shown in Figure 7(a), the experiment times were set at 5 min, 10 min, 30 min, and 2 h, respectively. With the increase of time, the BrO₃⁻ formed in the brackish water system showed an increasing trend, while the Br⁻ content showed a downward trend. When the reaction was set as 10 min, BrO₃⁻ with a content of 1.08 mg/L was detected in the system. After 2 h, the BrO₃⁻ production amount reached 14.88 mg/L. At the same time, Br⁻ in the reaction system was almost completely consumed. The experimental results show that the •SO₄⁻ formed in the system can oxidize Br⁻ to BrO₃⁻ when UV/PS is used to treat the water containing Br⁻.

The above experiment proves that in the absence of E2 in the reaction system, BrO₃⁻ is formed. This paper investigated the effect of E2 on the formation of BrO₃⁻ in the UV/PS system under the background value of brackish water. During the experiments, the reaction time was fixed and the concentration of E2 increased from 0 to 0.1 mmol/L. As shown in Figure 7(b), when the experiment was carried out for 30 min, the content of BrO₃⁻ decreased with the increase of E2 concentration in the brackish water. In the absence of E2, 7.38 mg/L of BrO₃⁻ was formed in the reaction system. When the concentration of E2 was set as 0.01 mmol/L, the amount of BrO₃⁻ in the reaction system was 5.934 mg/L. When the concentration of E2 was further increased to 0.02 mmol/L, the concentration of BrO₃⁻ was reduced to 4.029 mg/L. The reason for this phenomenon may be that the oxidized BrO₃⁻ formed in the reaction system can react with E2 in the system. The presence of E2 can inhibit the formation of BrO₃⁻ in the reaction system.

The aim of our work was to investigate the fate of E2 in the UV/PS degradation process. The results show that the pH of the water will affect the rate of degradation of E2 by UV/PS. Under neutral conditions, the degradation rate is the fastest. Over-alkali or per-acid will inhibit the reaction. Anions such as Br⁻ and Cl⁻ in water will promote UV/PS degradation of E2. In the reaction process, HCO₃⁻ has an inhibitory effect on the reaction, while SO₄²⁻ and cation have no effect on the degradation rate. UV/PS degradation of E2 was also a mineralization process, with the mineralization rate reaching 90.97% at 8 h. In the UV/PS system, the same intermediate products were detected in both brackish water and fresh water. The reaction of hydroxyl, deoxygenation, and hydrogenation are the primary degradation reactions. The E2 reaction site is mainly located on the benzene ring. There is mainly carbonylation on the five-membered ring. The bond cleavage between C10 and C5 leads to the removal of the benzene ring and the carboxylation of the C2 and C3 sites. Otherwise, in the presence of halogen ions, halogenated DBPs are not produced during UV/PS degradation of E2. On the other hand, the presence of E2 in the UV/PS system inhibits the formation of bromate.

This study was supported by the National Science Foundation China (No. 42030502, No. 42090041), the Natural Science Foundation of Guangxi Province (2020GXNSFAA159135). This study was also supported by the State Key Laboratory for Conservation and Utilization of Subtropical Agro-Bioresources.

All relevant data are included in the paper or its Supplementary Information.