Abstract

In this study, photochemical degradation of two emerging pharmaceutical chemicals, mefenamic acid (MF) and triclosan (TCS), was investigated to clarify the role of treated wastewater effluent matrices on their environmental photolysis. Target compounds were individually exposed to simulated sunlight in different media: ultrapure buffered water and synthetic field water with treated municipal wastewater effluent. The results in ultrapure buffered water showed that the direct photolysis processes in aquatic environments are not relevant to the elimination of MF. However, in samples containing treated wastewater effluent, photochemical degradation of MF was clearly enhanced. Our results indicate that MF undergoes indirect photolysis by reactive intermediates produced in an effluent matrix. Further quenching experiments suggested that photochemically produced hydroxyl radicals and excited triplet state dissolved organic matter drive the degradation of MF. In contrast to MF, TCS photochemical degradation proceeds through rapid direct photolysis. TCS was quickly degraded in ultrapure buffered water but it is considerably hampered in samples containing wastewater effluent. The declined degradation of TCS in the synthetic field water was discussed in terms of underlying optical filter effects by coexisting chromophoric substances. Results emphasize the importance of taking local water chemistry into consideration when predicting natural attenuation of pharmaceutical chemicals in receiving areas.

INTRODUCTION

In the past decades, aquatic pollution by residual pharmaceuticals and personal care products (PPCPs) has been a major concern. Discharges from wastewater treatment plants (WWTPs) have been recognized as one of the most important sources of PPCPs (Tanoue et al. 2015). Previous reports have detected PPCPs in aquatic environments such as wastewater effluent (Daouk et al. 2016), rivers and fish samples (Tanoue et al. 2015). Although concentrations of many PPCPs in natural aquatic environments are generally in the ranges from parts per trillion (ng L−1) to parts per billion (μg L−1), several compounds have been shown to remain in environments at levels that can pose potential ecological risks. Among PPCPs, mefenamic acid (MF) is widely used as a nonsteroidal anti-inflammatory drug and has been frequently detected in aquatic environments due to the low removal efficiency of WWTP processes (Dai et al. 2014; Tanoue et al. 2015). On the other hand, triclosan (TCS) has been extensively added to numerous consumer products as an antibacterial agent and also widely found remaining in aquatic environments. Studies on toxicity showed that exposure to MF and TCS can potentially induce adverse effects on aquatic organisms such as luminescence inhibition, alteration of swimming performance, and death (Chen et al. 2015; Ishibashi et al. 2004; Nassef et al. 2010). The importance of studies regarding the environmental fate of MF and TCS is emphasized by the potential concern that their aquatic levels can exceed predicted no-effect concentrations (Daouk et al. 2016; Guo & Iwata 2017).

Among environmental processes, photodegradation is expected to be an alternative to control the aforementioned concerns. In sunlit aquatic environments, not only direct photolysis but also indirect photolysis is important for transformation and elimination of micropollutants. For example, nitrate and dissolved organic matter (DOM) can induce indirect degradation processes by photochemically produced reactive intermediates (PPRIs) such as hydroxyl radicals (•OH), excited triplet state DOM (*DOM) and singlet oxygen (1O2), though DOM can also inhibit photodegradation, acting as radical quencher and optical filter (Ryan et al. 2011; Chiwa et al. 2015; Shang et al. 2015). For these reasons, photodegradation of PPCPs in pure water and synthetic field water systems that consist of natural organic matter (NOM) have been well explained (Werner et al. 2005; Yamamoto et al. 2009), but to date only a few studies have investigated photochemical reactions of PPCPs in WWTP effluent matrices under environmentally relevant conditions (Ryan et al. 2011; Wang et al. 2017). It would also be necessary to clarify the photochemical degradation processes in a WWTP effluent system because of the unique origin and chemical compositions of effluent organic matter that differ from NOM (Shon et al. 2006). In addition, municipal WWTP effluents that have received aerobic biological treatment processes are generally rich in nitrate, an important source of •OH (Chiwa et al. 2015). Recently, Wang et al. (2017) studied photochemical degradation of PPCPs under wastewater stabilizing pond systems to show the important effects of effluent matrices. This indicates that effluent matrices also play crucial roles in environmental fates of PPCPs in effluent receiving streams. However, to the best of our knowledge, photochemical reactions of MF and TCS in effluent matrices have not been simultaneously and systematically investigated within the same study. Owing to the different chemical structure and physicochemical properties, WWTP effluent matrices would exhibit different extents of impact on the aqueous photodegradation of MF and TCS.

The main purpose of this laboratory-scale study was to reveal the role of effluent matrices on photochemical degradation processes of MF and TCS in order to characterize the elimination of PPCPs in effluent-impacted streams. Samples of PPCPs dissolved in ultrapure water and in collected WWTP effluent were individually exposed to simulated sunlight. Quenching experiments were performed in order to identify important PPRIs involved. The present results are expected to contribute in better understanding pollutant fates, ultimately assisting in environmental risk assessments of aquatic systems.

MATERIALS AND METHODS

MF (98%), TCS (96%), cimetidine (98%), isopropanol (IPA), and isoprene were all purchased from Wako Pure Chemical Industries (Osaka, Japan). Ultrapure water was provided by a Millipore ultrapure water system. A 100 W Xe solar simulator lamp (LCS-100A, Oriental Instrument, USA), equipped with a 1.5 G air mass filter, was used as the light source in photolysis experiments. The light intensity of the Xe lamp was set at 100 mW/cm2 in all experiments. PPCPs analysis was performed by a high performance liquid chromatography (HPLC) system equipped with a UV-visible (UV-Vis) detector (SPD 10A, Shimadzu, Japan). Suwanee river humic acid was obtained from International Humic Substances Society.

The treated wastewater samples were taken from a WWTP located in Saitama Prefecture, Japan. The WWTP employs an oxidation ditch system as a second stage treatment process. The final effluent samples from the WWTP were collected in a 500 mL glass vessel and transported to a laboratory immediately after sampling, under dark conditions. The sample was then passed through a 0.45 μm membrane fibre filter (A045A, Advantec) in order to remove particles and microbes. Dissolved organic carbon (DOC) and nitrate present in the effluent were measured using a total organic carbon analyzer (TOC-VCSN, Shimadzu, Japan) and an ion chromatograph (IC-2010, Tosoh, Japan), respectively. Optical characteristics of the WWTP effluent samples were also measured by a UV-Vis analyzer (UVmini-1240, Shimadzu, Japan) and a spectrofluorometer (FP-8300, Jasco, Japan) to determine specific UV absorption at 254 nm (SUVA254) and humification index (HIX) (Coble et al. 2014). The SUVA254 is an index for relative contents of aromatic hydrocarbons and determined by dividing the UV absorption at 254 nm with DOC. The HIX provides information on degree of humification of DOM and determined based on the ratio of average emission intensity from 435–480 nm to those of 300–345 nm with the excitation wavelength of 255 nm. The key parameters of the used effluent sample are summarized in Table 1. Although the SUVA and HIX values of the effluent sample were relatively low, the water chemistry shows that the used sample certainly contains humic substances as important components of DOM. On the other hand, Du et al. (2018) reported SUVA and HIX (determined by 435–480 nm/300–480 nm) of Suwannee River humic acid as 5.77 and 17.74, respectively.

Table 1

Relevant parameters of the used effluent samples

pH Nitrate (mg L−1DOC (mg L−1SUVA254 (L m−1 mg−1HIX 
8.2 6.6 9.8 0.5 1.9 
pH Nitrate (mg L−1DOC (mg L−1SUVA254 (L m−1 mg−1HIX 
8.2 6.6 9.8 0.5 1.9 

To examine effects of direct and indirect photolysis, photolysis experiments were performed under ultrapure water conditions and in treated sewage sample matrices using a Xe lamp sunlight simulator. As discharges from WWTPs comprise tens of percent of river flow in effluent receiving streams (Mano & Okamoto 2016), synthetic field water was prepared by diluting the WWTP effluent sample by a factor of 10 (1:9. v/v) with 1.0 mM phosphate buffer solution (ultrapure, pH 7). To highlight the role of effluent matrices, in this study synthetic field water was prepared, diluting with pure water. For the photolysis experiments, standard stock solutions of individual PPCPs were prepared in methanol and added to 10 mL volumetric flasks. Since methanol can impair indirect photochemical reactions by acting as a quencher of •OH, the solvent was removed by a gentle nitrogen stream and then replaced with either the WWTP effluent samples or ultrapure water buffered by the phosphate buffer (1.0 mM). MF and TCS with concentrations of 1 mg L−1 were individually irradiated with the simulated sunlight in horizontally placed capped-quartz cells (12.5 mm × 12.5 mm × 58 mm). In addition, quenching experiments were conducted to investigate degradation pathways of PPCPs via indirect photolysis under WWTP effluent matrices. IPA and isoprene are frequently used as quenchers of •OH and excited triplet state, respectively (Werner et al. 2005; Ryan et al. 2011). These were added to effluent samples at the final concentrations of 1%. Moreover, Suwanee River humic acid was added to ultrapure water to experimentally evaluate inhibition effects of DOM on TCS degradation. The humic acid was prepared at two levels (1 mg L−1 and 2 mg L−1) and irradiated with the lamp in the presence of TCS. The photolysis experiments were basically performed in triplicate (n = 3) for each condition. At each sampling time-point (TCS: 0, 3, 6, 9, and 12 min; MF: 0, 30, 60, 90, and 120 min), an amount was taken and injected (20 μL) into the HPLC system. Samples were chromatographically separated by an octadecyl-silica column (Shim-pack VP-ODS; length: 150 mm; internal diameter: 4.6 mm; Shimadzu, Japan) at a column temperature of 30 °C, using a mixture of phosphate buffer solution (1.0 mM, pH 2.6) and methanol (1:4, v/v) at a flow rate of 1 mL min−1, in isocratic mode. TCS and MF were detected at 220 nm and 225 nm, respectively. Quantitative limits of MF and TCS in samples, determined based on three times the standard deviations obtained from repeated measurements (n= 7) of the lowest level of calibration curves, were 26.4 μg L−1 and 36 μg L−1, respectively. Pseudo-first order rate constants for the degradation of PPCPs were calculated by linear regression of logarithmic response of PPCPs as a function of time.

RESULTS AND DISCUSSION

MF

MF photodegradation profiles in different media are shown in Figure 1. Assuming the degradation process to be of pseudo-first order, rate constants were calculated and are summarized in Table 2. Results in the buffered ultrapure water (UPW) system represent effects of direct photolysis. MF in the ultrapure water system was strongly tolerant to degradation as almost all MF persisted in the samples, even after 120 min of exposure to simulated sunlight. The persistent behavior of MF in the ultrapure water system is consistent with a previous report by Yamamoto et al. (2009). Werner et al. (2005) also reported that MF undergoes slow direct photolysis, with an estimated half-life time of 33 h. These results were rationalized by the lower quantum yields and relaxing effects of the excited triplet state of MF upon irradiation when dissolved in polar media, including water (Werner et al. 2005; Yamamoto et al. 2009). In contrast, degradation of MF in the treated effluent sample was clearly enhanced. The response of MF after 120 min of irradiation under WWTP effluent conditions decreased to 20.4 ± 1.3% (mean ± SD) compared to that of the initial time and the degradation rate constant drastically increased in the effluent sample (Table 2). These results show that degradation behavior of MF in effluent are different from those in ultrapure water and the enhanced degradation of MF in effluent matrices could be attributed to effects of indirect photolysis. Recently, Wang et al. (2017) also reported that photodegradation of several PPCPs such as caffeine and carbamazepine can be accelerated in treated wastewater effluent conditions. Taken together, our results suggest that MF exhibits only negligible levels of degradation through direct photolysis, meanwhile the indirect photolysis processes involving PPRIs formed upon irradiation significantly drive its degradation.

Table 2

Mean degradation rate constants of PPCPs under different conditions within ±95% confidence intervals

Compounds Photolysis conditions (n = 3) Photodegradation rate constants (min−1
MF UPW N.A. 
Synthetic field water 1.3 × 10−2 ± 8.9 × 10−4 
Synthetic filed water +1% IPA 4.5 × 10−3 ± 1.3 × 10−3 
Synthetic field water +1% IPA +1% isoprene 5.9 × 10−4 ± 6.4 × 10−4 
Cimetidine Synthetic field water 4.7 × 10−4 ± 4.8 × 10−4 
TCS UPW 0.16 ± 1.8 × 10−2 
Synthetic field water 5.5 × 10−2 ± 1.6 × 10−2 
Synthetic field water +1% IPA 5.1 × 10−2 ± 1.2 × 10−2 
UPW +1 mg L−1 humic acid 4.6 × 10−2 ± 3.8 × 10−3 
UPW +2 mg L−1 humic acid 5.8 × 10−2 ± 7.1 × 10−3 
Compounds Photolysis conditions (n = 3) Photodegradation rate constants (min−1
MF UPW N.A. 
Synthetic field water 1.3 × 10−2 ± 8.9 × 10−4 
Synthetic filed water +1% IPA 4.5 × 10−3 ± 1.3 × 10−3 
Synthetic field water +1% IPA +1% isoprene 5.9 × 10−4 ± 6.4 × 10−4 
Cimetidine Synthetic field water 4.7 × 10−4 ± 4.8 × 10−4 
TCS UPW 0.16 ± 1.8 × 10−2 
Synthetic field water 5.5 × 10−2 ± 1.6 × 10−2 
Synthetic field water +1% IPA 5.1 × 10−2 ± 1.2 × 10−2 
UPW +1 mg L−1 humic acid 4.6 × 10−2 ± 3.8 × 10−3 
UPW +2 mg L−1 humic acid 5.8 × 10−2 ± 7.1 × 10−3 

Photolysis rate constant of MF in ultrapure water (UPW) was not calculated due to the low degradability and is represented as N.A.

Figure 1

Degradation of MF under buffered ultrapure water (UPW) and treated wastewater effluent conditions with simulated sunlight exposure. Symbols and error bars represent averages values and standard deviations, respectively (n = 3).

Figure 1

Degradation of MF under buffered ultrapure water (UPW) and treated wastewater effluent conditions with simulated sunlight exposure. Symbols and error bars represent averages values and standard deviations, respectively (n = 3).

To investigate PPRIs that contribute to the degradation of MF under WWTP effluent conditions, quenching experiments were performed using IPA (•OH quencher) and isoprene (excited triplet state quencher). The degradation of MF in the effluent sample was greatly inhibited with the addition of quenchers (Table 2). Spiking IPA (1%) to the effluent sample prior to photolysis resulted in the decrease of the degradation rate constant of MF by 32.7%. Moreover, with the combination of IPA and isoprene, the rate constant was reduced by 4.5% compared to that obtained in the effluent sample without quenchers. The reduced degradation rate constant of MF in the presence of IPA and isoprene was 5.9 × 10−4 min−1 with an estimated half-life time of 1,171 min (19.5 h), showing similar behavior to those in ultrapure water conditions. These results strongly imply that •OH and *DOM are important PPRIs in the indirect photolysis of MF under the studied system. Another possible PPRI is 1O2 that can be produced by the interaction of *DOM and O2. In order to assess the 1O2 contribution for the indirect photodegradation processes, we performed photolysis experiments using cimetidine solutions (1 mg L−1) since it has been reported that almost all photochemical degradation using cimetidine solutions occurs via reactions with 1O2 (Ryan et al. 2011). Thus if 1O2 is produced during photolysis, a significant decrease of cimetidine should be observed. However, results demonstrated that degradation of cimetidine in the effluent sample was markedly slower than those of MF and TCS (Table 2). This means that the role of 1O2 is insufficient to contribute to the degradation of MF and the sample is composed of DOM that has limited capability to yield 1O2 under air equilibrated conditions, again highlighting the importance of *DOM and •OH. The indirect photolysis rate of MF by these PRRIs would be different for different water matrices. Werner et al. (2005) have reported enhanced degradation of MF by indirect photolysis in the presence of fluvic acid, which could be mostly attributed to reaction of MF with *DOM. In addition, MF undergoes efficient degradation in the presence of •OH (Chen et al. 2015; Davis et al. 2017). For sources of •OH in the effluent samples, there are two possible main pathways: nitrate and DOM. Nitrate is frequently found in effluent and it generates •OH in sunlit surface water by the following reaction (Liu et al. 2010). 
formula
(1)

However, Werner et al. (2005) suggested insignificant contribution of •OH to the indirect photolysis of MF. The differences might be due to differences in the water matrix such as levels of nitrate. In addition, formation rate of •OH from photolysis of DOM depends on their types. Generally, effluent organic matter is composed of various classes of compounds not only NOM but also synthetic organic compounds, and soluble microbial products produced during biological treatment processes (Shon et al. 2006). According to a previous study conducted by Dong & Rosario-Ortiz (2012), such organic matter in WWTP effluents can undergo photochemical reactions to form higher amounts of •OH compared to those from NOM. As shown in Table 1, moderate levels of nitrate and DOC were detected in the WWTP effluent sample, which can evince the formation of PPRIs during photolysis. We exclude contributions of photo-Fenton reactions to •OH formation, because such processes take place preferentially under acidic condition, which is not the present case (Jacobs et al. 2012).

The results indicate that MF can undergo efficient photodegradation in effluent-impacted streams during river transport. However, it should be noted that photolysis does not necessarily mean reduction of total risks related to MF. Although we did not examine any intermediates of MF in this study, Chen et al. (2015) detected a series of products including nitrosylated MF derivatives upon UV radiation in nitrate solution, some of which can exhibit higher toxicity than the parent compound.

TCS

TCS photodegradation profiles under different conditions are shown in Figure 2. Compared to MF, TCS showed trends of relatively rapid degradation under both buffered ultrapure water and effluent sample conditions upon exposure to simulated sunlight. Control experiments showed that concentrations of TCS hardly decreased in dark conditions for both matrices, which means that effects of biodegradation and sorption are negligible, and that temporal decrease trends of TCS in the irradiated samples occurred by photochemical degradation. Although reports on TCS degradation rate constants vary considerably according to experimental conditions such as type of light sources and solution pH, the relatively rapid degradation of TCS in this study is generally consistent with previous reports (Latch et al. 2005; Sanchez-Prado et al. 2006; Buth et al. 2009). In the present study, TCS was generally much more susceptible to degradation and likely to degrade faster under ultrapure water conditions than in effluent samples. Remaining TCS after 12 min was 12.6 ± 2.3% and 52 ± 11.1% of the total initial amount, in the buffered ultrapure water and WWTP effluent samples, respectively. As shown in Table 2, the calculated degradation rate constant of TCS in the treated wastewater effluent was reduced to 34.3% of that in the ultrapure water. This result is consistent with a previous report by Sanchez-Prado et al. (2006). Moreover, the degradation rate constant of TCS in the effluent sample with 1% IPA was within the same order of magnitude compared to that in the absence of IPA (Table 2). According to the results, it is implied that (1) degradation of aqueous TCS predominantly occurs via direct photolysis while indirect photolysis is insignificant, and (2) effluent matrices can inhibit the apparent photodegradation rates. Under the tested condition in this study, TCS (pKa: 8.1) would dominantly present as neutral form. Although neutral form TCS is more photochemically stable than phenolate form TCS, it would also efficiently absorb sunlight radiation (Latch et al. 2003). Even if PRRIs were generated during irradiation, acceleration effects would be limited when target compounds receive direct photolysis at higher rates. On the other hand, the inhibited degradation of TCS in the effluent condition can be considered to be due to inhibition of photolysis by competitive light absorption and/or •OH scavenging caused by coexisting DOM (Shang et al. 2015). However, effects of scavenging of •OH would not be important for TCS since quenching experiments using IPA showed limited effects of indirect photolysis. Tixier et al. (2002) have suggested that DOM is one of the important factors to inhibit degradation of TCS by light absorption effects. As shown in Table 1, the used effluent sample is rich in DOC which would possibly hinder the direct light absorption processes of TCS. To experimentally examine the optical filter effects of organic matter on degradation of TCS, additional photolysis experiments were performed using Suwanee River humic acid as a model of DOM (1 mg L−1 and 2 mg L−1). The addition of DOM to ultrapure water solution resulted in dramatically decreased degradation of TCS (Figure 3), which could be attributed to light absorption by chromophoric moieties of DOM (Shang et al. 2015). These results would support that TCS dominantly degrades via direct photolysis and further give a rational explanation to results obtained in the synthetic field water experiments. We also note that the DOM compositions of Suwannee River humic acid are different from those of the used effluent organic matter as shown by different SUVA and HIX. However, because degradation rates of TCS in the presence of humic acid are similar to that of synthetic field water (Table 2), it might be likely that coexisting DOM inhibits the direct photolysis processes of TCS regardless of its detailed characteristics. Based on these results, photolysis of TCS can be inhibited in effluent matrices by optical filter effects of DOM. Therefore environmental TCS in effluent receiving streams would be more persistent than those expected from photolysis experiments conducted in ultrapure water systems.

Figure 2

Degradation of TCS under buffered ultrapure water (UPW) and treated wastewater effluent with and without photolysis. Controls in UPW are results of duplicate experiments (n = 2).

Figure 2

Degradation of TCS under buffered ultrapure water (UPW) and treated wastewater effluent with and without photolysis. Controls in UPW are results of duplicate experiments (n = 2).

Figure 3

Effects of humic acid (HA) on degradation of TCS. Kinetics data in ultrapure water (UPW) samples are the same as shown in Figure 2.

Figure 3

Effects of humic acid (HA) on degradation of TCS. Kinetics data in ultrapure water (UPW) samples are the same as shown in Figure 2.

In this study, we experimentally confirmed that treated WWTP effluent matrices substantially affect the apparent photodegradation rate constants of MF and TCS present in aquatic environments. WWTP effluent matrices had either acceleration or inhibition impacts on apparent half-life times of PPCPs according to their direct photolysis rates. Recent studies have made special efforts to characterize dynamics of PPCPs in streams. However, it has been shown that natural attenuation behaviors of PPCPs are different according to the water systems (Li et al. 2016; Hanamoto et al. 2018). The results in the present study emphasize the importance of taking local water chemistry into consideration when predicting and rationalizing natural attenuation of PPCPs in receiving areas.

ACKNOWLEDGEMENTS

The authors thank Saitama Prefectural Public Sewage Works Bureau for their kind assistance in effluent sampling. We thank the anonymous reviewers for their comments that improved our work. We are also grateful to Hibiki Moroi and Saeka Ishikawa (Graduate School of Science and Engineering, Saitama University) for their valuable assistance with sampling and photolysis experiments.

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(
2
),
351
362
.