This study is the first to comprehensively investigate the photochemical fate of pentoxifylline in natural water systems. Among all attenuation processes, indirect sunlight photolysis is very likely the primary process for pentoxifylline degradation. The combination of dissolved organic matter (represented by fulvic acid), NO3− and HCO3− closely simulated the photolysis rate in real surface water (Jingmei River, Taipei) (t1/2 = 34.7 ± 2.2 h), indicating that these three parameters are the most important determinants of the photolysis fate of pentoxifylline. The results also showed that HCO3− contributes to photodegradation of pentoxifylline in the NO3− system due to the participation of ·CO3−. Although pentoxifylline was degraded, total organic carbon remained constant during the photolytic process, indicating that byproducts were generated in the solution. Five byproducts that have not been previously reported are proposed for the first time in this study. Among all of the byproducts, two byproducts (P4 (M.W. = 250) and P5 (M.W. = 227)) were only detected in the presence of NO3− and HCO3−, implying that ·CO3− possibly caused different photolytic pathways of pentoxifylline.
Pentoxifylline, which belongs to a group of vasodilators called methylxanthines, is used for the treatment of peripheral arterial disease and helps with blood circulation in human bodies. Pentoxifylline increases the blood flow by decreasing the viscosity of blood and decreases the potential for platelet aggregation and thrombus formation.
According to National Health Insurance medicine usage data from 2015 in Taiwan, the usage of pentoxifylline was 21.5 tons. Due to its wide usage, pentoxifylline residues have been detected in various water matrices of different countries. For example, in Taiwan, pentoxifylline has been observed in drug production facilities and hospitals, with maximum concentrations of 1,370 and 302 ng/L, respectively (Lin et al. 2008). It has also been detected in the influents (98–191 ng/L) and effluents (56–147 ng/L) of wastewater treatment plants (WWTPs) and even in the groundwater (0.4–2.4 ng/L) in Taiwan (Lin et al. 2009, 2015). In other countries, such as Romania and Canada, pentoxifylline has been detected at maximum concentrations of 299 and 45 ng/L in WWTP influents and effluents, respectively (Hua et al. 2006; Moldovan et al. 2009). Pentoxifylline has also been found in the Rhine River in Germany at a maximum concentration of 570 ng/L (Sacher et al. 2008). Several researchers have investigated the acute toxicity of pentoxifylline (Hansen 1994; Beshay et al. 2001; Ruddock & Hirst 2005); however, there is still a lack of information regarding its chronic toxicity, as well as potential risks to aquatic organisms due to the presence of pentoxifylline residues in aqueous environments.
Currently, little information is available regarding the environmental fate of pentoxifylline. Although pentoxifylline can be strongly hydrolyzed (t1/2 < 1 day at pH = 13 at high temperatures (80 °C)), it is stable in typical environmental waters (pH ∼ 7, T ∼ 20 °C) with very slow hydrolysis rates (no change in 4 days) (Mone & Chandrasekhar 2010); moreover, it cannot be eliminated through volatilization because its Henry's law constant is low (9.8 × 10−14 atm m3/mole). Pentoxifylline has a low octanol-water partitioning coefficient (logKow = 0.56) (Udrescu et al. 2008), implying that it is unlikely to be attenuated through sorption or biotransformation in aqueous environments. Therefore, sunlight photolysis may be the predominant process determining the natural fate of pentoxifylline. Researchers have suggested that sunlight photolysis may significantly reduce many types of recalcitrant pharmaceuticals (e.g. cephalosporins, sulfonamides, fluoroquinolones, β-agonists, controlled drugs and chemotherapeutic drugs) in aqueous environments (Boreen et al. 2004; Wang & Lin 2012; Yang et al. 2013; Lin et al. 2014b; Sturini et al. 2015). However, to date, no information exists regarding the sunlight-mediated photochemical behavior of pentoxifylline.
Sunlight photochemical reactions in aqueous environments include direct and indirect photolysis (Schwarzenbach et al. 2005). The former occurs via light absorption by compounds, resulting in bond cleavage; the latter is triggered via light absorption by photosensitizers, such as dissolved organic matter (DOM), and NO2−, which produce reactive species that subsequently react with the target chemicals. has also been thought to affect the photolysis behavior; however, the overall effect of seems to be compound-dependent. Researchers have recently reported that the presence of enhances the photolysis rate of 5-fluouracil, 4-halogenophenol, catechol and cephalosporins due to the generation of carbonate radicals (·CO3−) in the presence of (Vione et al. 2009; Wang & Lin 2012; Lin et al. 2013). Many other studies have also reported that inhibits the photolysis rate when is present in the solution (Lam et al. 2003; Chowdhury et al. 2011). is present at high concentrations in surface water (1–3 mM) (Hem & Geological Survey (US) 1985; Brezonik & Fulkerson-Brekken 1998), and the estimated concentration of ·CO3− (10−13–10−14 M) is higher than that of hydroxyl radicals (·OH) (10−16–10−18 M) (Mill et al. 1980; Russi et al. 1982; Huang & Mabury 2000b), with a difference of more than two orders of magnitude under sunlight irradiation. Therefore, it is interesting to investigate the influence of on the phototransformation behavior of pentoxifylline.
The purpose of this work was to investigate the sunlight photodegradation behavior of pentoxifylline in natural surface water environments. Water parameters, including DOM, and , as well as their competing effects, were studied. Phototransformation byproducts and pathways during the solar photodegradation of pentoxifylline were also investigated. To the best of our knowledge, this is the first investigation of the sunlight photolysis of pentoxifylline in surface waters.
MATERIALS AND METHODS
Pentoxifylline (99%), formic acid, ammonium acetate, sodium nitrate (NaNO3) and sodium bicarbonate (NaHCO3) were purchased from Sigma-Aldrich (St Louis, MO, USA). HPLC-grade methanol was purchased from Avantor Performance Materials (Phillipsburg, NJ, USA). The Suwannee River Fulvic Acid (FA) standard (1S101F) was obtained from the International Humic Substance Society (St Paul, MN, USA). FA, NaNO3 and NaHCO3 were used as the source of DOM, and in this study, respectively. Individual stock standard solutions (pentoxifylline, FA, NaNO3 and NaHCO3) were prepared on a weight basis (1,000 mg/L) using Milli-Q water and were stored in amber glass bottles at 4 °C for a maximum of 30 days. The physicochemical properties of pentoxifylline are shown in Table S1 (available with the online version of this paper).
Surface water sampling
Grab samples (2 L) were collected in amber glass bottles from the Jingmei River, which is located in southern Taipei, in July 2012. The Jingmei River, one of the main rivers located in the southwest of the Taipei Basin, has a length of 28 km and a drainage area of 114 km2. The average flow rate of the Jingmei River is 1,952,460 m3/day. The stream drainage range of the upstream Jingmei River includes wastewater from approximately 150 hospitals and clinics and 10 animal husbandries, and our past works have shown that various pharmaceuticals were found constantly and at significant concentrations in this drainage area (Lin et al. 2008, 2010a, 2010b, 2014a). All water samples (except for the ones used for the incubation experiments from the Jingmei River) were vacuum-filtered through 0.22-μm cellulose acetate membrane filters (Advantec, Toyo Roshi Kaisha, Japan) and stored in a 4 °C refrigerator until analysis. The water parameters, including the pH, dissolved organic carbon (DOC), and alkalinity of the grab samples, were analyzed.
Photolysis experiment setup
Photochemical experiments were conducted in a sunlight simulator (Suntest CPS; Atlas, Chicago, IL, USA) equipped with a 1.5-kW xenon arc lamp; detailed operating conditions have been previously reported (Lin et al. 2014a, 2014b). A Suprax filter was fitted to allow a total passing wavelength range of 290 − 800 nm, and the irradiation intensity and the illumination time was set to 765 W/m2 and 40 h for all photolysis experiments, respectively. For the photolysis experiments, the target compound pentoxifylline and other reagents (FA, and ) were spiked into Milli-Q water in the reactor to achieve the desired concentrations. Individual standards of pentoxifylline in Milli-Q water (20 μg/L) were placed in capped quartz glass reaction tubes (1.6-cm i.d. × 13.5-cm depth, volume of 27 mL) and exposed to radiation from a sunlight simulator maintained at 20 ± 1 °C with a thermostat. Dark control experiments of the same concentrations were scrupulously maintained in darkness. The solution pH was adjusted to 7.0 (using 1.0 N sulfuric acid and 1.0 N sodium hydroxide) for all synthetic waters and was maintained at 7.0 throughout the reaction, except for those meant to simulate the Jingmei River waters (which were adjusted to a pH of 7.7 to match). The synthetic waters were prepared with the Suwannee River FA standard, NaNO3 and NaHCO3. The overall experimental framework of this study is provided in Figure S1 (available with the online version of this paper).
An incubation test was conducted in the dark at 20 °C for pentoxifylline (20 μg/L) in the Jingmei River water (unfiltered) for 46 h.
The chromatographic separation of pentoxifylline and byproducts was performed using an Agilent 1200 module (Agilent, Palo Alto, CA, USA) equipped with a ZORBAX Eclipse XDB-C18 column (150 × 4.6 mm, 5 μm). Mass spectrometric measurements were performed on a Sciex API 4000 (Applied Biosystems, Foster City, CA, USA) equipped with an electrospray ionization (ESI) interface. For the byproduct study, pentoxifylline was investigated at a high initial concentration (20 mg/L). The full scan mode was used to detect byproducts in the degradation mixture and to obtain the mass spectra of those byproducts (Figure S2, available with the online version of this paper). The signal areas of the byproducts were quantified with liquid chromatography-tandem mass spectrometry (LC-MS/MS). Detailed information regarding the LC–MS/MS operations and the byproduct investigation is presented in the Supplementary text.
An ACD MS fragmenter (Advanced Chemical Development, Toronto, ON, Canada) was used to generate a tree fragmentation for the structure of pentoxifylline based on mass spectrometry fragmentation rules; additionally, the selected ESI ionization mode and the number of fragmentation steps were used to identify the byproducts of photolysis. The tree fragmentations of the parent compound pentoxifylline were used to identify the product ions from the mass spectra of their byproducts, and the product ions were then combined to predict the structures of the byproducts (Lin et al. 2014a; Lai et al. 2015).
A total organic carbon (TOC) analysis was conducted using an OI Analytical model 1030 (OI Analytical, College Station, TX, USA) with an autosampler (OI Analytical model 1088). Prior to the analysis, the samples were acidified and sparged with nitrogen to remove inorganic carbon. The standard solutions for calibration were prepared using potassium hydrogen phthalate.
RESULTS AND DISCUSSION
Indirect photolysis in natural surface waters
Influence of DOM
Influence of NO3−
Influence of HCO3− in the NO3− system
Compared to ·OH, ·CO3− is a highly selective oxidant and is less scavenged by DOM, which leads to a higher steady-state concentration in natural surface water systems (Huang & Mabury 2000b; Canonica et al. 2005). In addition, the existence of has been reported to induce a higher generation rate of reactive species during the photolysis of (Vione et al. 2009). Previous studies have also demonstrated that ·CO3− has an affinity for reacting with compounds that contain electron-rich moieties (e.g. aromatic anilines, amino acids, phenols, nitrogen-containing and sulfur-containing organics) through the reactions of electron transfer and hydrogen abstraction (Neta et al. 1988; Huang & Mabury 2000a; Wang & Lin 2012; Liu et al. 2016).
Photolysis byproducts (NO3− system versus NO3− + HCO3− system)
In the photolysis of , ·OH was generated. Because pentoxifylline is unable to undergo direct photolysis (Figure 1(b)), byproducts P1–P3 were produced through a ·OH attack on pentoxifylline. In addition, in the presence of both and , ·OH and ·CO3− were produced and further reacted with the target compound pentoxifylline. Therefore, ·CO3− may be involved in the generation of byproducts P4 and P5. P4 is the demethylation product of pentoxifylline. The role of ·CO3− in the degradation of oxytetracycline by a UV-based advanced oxidation process has been investigated, and it has been reported that ·CO3− proceeds the demethylation reaction (Liu et al. 2016). A similar phenomenon was also noted in another study (Mazellier et al. 2007), which demonstrated the demethylation byproduct of fenuron by UV photolysis of Co(NH3)5CO3+ in the presence of ·CO3−. In summary, different types of byproducts were formed in the two systems ( system and system), which implies that different photodegradation pathways occurred.
This study provides the first comprehensive investigation of the photochemical fate of pentoxifylline in surface waters. The results indicated that pentoxifylline is fairly persistent in the aquatic environment once it is released to the environment, and sunlight photolysis is very likely the predominant attenuation process. In waters such as the Jingmei river (t1/2 = 34.7 ± 2.2 h), it would take approximately 5–7 days for the concentration of pentoxifylline to decrease by half (assuming 5–7 h of bright sunlight a day). However, the presence of DOM, and greatly affects the photolysis rate of pentoxifylline. During the photolysis of , the presence of contributes to photodegradation of pentoxifylline due to the participation of ·CO3−. Despite the fact that pentoxifylline was photodegraded through indirect sunlight photolysis, TOC data (no mineralization) indicated that pentoxifylline was only photo-transformed into degradation byproducts. Five photodegradation byproducts (P1–P5) of pentoxifylline, which have not been reported previously, are proposed for the first time in this work; among the degradation byproducts, byproducts P4 and P5 were only observed in the and system. This finding implies that compared to ·OH, ·CO3− likely leads to a different degradation pathway. Given that is abundant in natural waters (approximately 1–3 mM), the involvement of in the photochemical fate should not be neglected. An additional toxicity assessment of the photodegradation byproducts is also essential to understand their risk toward ecosystems and humans.
We are grateful for the support from the Ministry of Science and Technology (MOST) (MOST 103-2221-E-002-240-MY5), Taiwan, ROC.