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Aqueous photochemical degradation of hydroxylated PAHs: Kinetics, pathways, and multivariate effects of main water constituents
Science of the Total Environment 547 (2016) 166–172
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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv
Aqueous photochemical degradation of hydroxylated PAHs: Kinetics, pathways, and multivariate effects of main water constituents Linke Ge a, Guangshui Na a, Chang-Er Chen b, Jun Li a,c, Maowei Ju a, Ying Wang a, Kai Li a, Peng Zhang a,⁎, Ziwei Yao a a b c
Key Laboratory for Ecological Environment in Coastal Areas (SOA), National Marine Environmental Monitoring Center, Dalian 116023, China Lancaster Environment Centre, Lancaster University, Lancaster LA1 4YQ, United Kingdom College of Marine Science, Shanghai Ocean University, Shanghai 201306, China
H I G H L I G H T S
G R A P H I C A L
A B S T R A C T
• It is first reported on aqueous photochemical behavior of 4 hydroxylated PAHs. • Hydroxylated PAHs intrinsically photodegrade fast in sunlit surface waters. • Reaction types and transformation pathways of 9-Hydroxyfluorene were clarified. • Photolysis kinetics was affected by multivariate effects of main water constituents. • The photomodified toxicity of 9Hydroxyfluorene was examined using Vibrio fischeri.
Aqueous photochemical behavior of 4 hydroxylated PAHs is first reported on revealing the kinetics, mechanisms, toxicity, and multivariate effects of water constituents.
a r t i c l e
i n f o
Article history: Received 23 September 2015 Received in revised form 28 December 2015 Accepted 29 December 2015 Available online xxxx Editor: Kevin V. Thomas Keywords: Hydroxylated PAHs 9-Hydroxyfluorene Photodegradation kinetics Pathways Water multivariate effects Photomodified toxicity
⁎ Corresponding author. E-mail address: [email protected] (P. Zhang).
http://dx.doi.org/10.1016/j.scitotenv.2015.12.143 0048-9697/© 2015 Elsevier B.V. All rights reserved.
a b s t r a c t Hydroxylated polycyclic aromatic hydrocarbons (OH-PAHs) are contaminants of emerging concern in the aquatic environment, so it is of great significance to understand their environmental transformation and toxicity. This study investigated the aqueous photochemical behavior of four OH-PAHs, 9-Hydroxyfluorene (9-OHFL), 2Hydroxyfluorene, 9-Hydroxyphenanthrene and 1-Hydroxypyrene, under simulated sunlight irradiation (λ N 290 nm). It was observed that their photodegradation followed the pseudo-first-order kinetics. Based on the determined quantum yields, their calculated solar apparent photodegradation half-lives in surface waters at 45° N latitude ranged from 0.4 min for 9-Hydroxyphenanthrene to 7.5 × 103 min for 9-OHFL, indicating that the OH-PAHs would intrinsically photodegrade fast in sunlit surface waters. Furthermore, 9-OHFL as an example was found to undergo direct photolysis, and self-sensitized photooxidation via •OH rather than 1O2 in pure water. The potential photoreactions involved photoinduced hydroxylation, dehydrogenation and isomerization based on product identification by GC–MS/MS. 9-OHFL photodegraded slower in natural waters than in pure water, − which was attributed to the integrative effects of the most photoreactive species, such as Fe(III), NO− 3 , Cl and humic acid. The photomodified toxicity was further examined using Vibrio fischeri, and it was found that the toxicity of photolyzed 9-OHFL did not decrease significantly (p N 0.05) either in pure water or in seawater, implying
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the comparable or higher toxicity of some intermediates. These results are important for assessing the fate and risks of OH-PAHs in surface waters. © 2015 Elsevier B.V. All rights reserved.
1. Introduction Hydroxylated polycyclic aromatic hydrocarbons (OH-PAHs) are newly recognized contaminants derived from hydroxylation of PAHs. They are commonly used as biomarkers of human exposure to PAHs (Guo et al., 2013), and some of them have been proved to be of even more pronounced mutagenicity/toxicity than their original PAH precursors (Cochran et al., 2012). Various sources can contribute to the contaminants in the environment. Once absorbed in humans and animals, PAHs are metabolized to a complex mixture of products including OHPAHs (Rey-Salgueiro et al., 2009). In sediments/soils, some microorganisms may biotransform PAHs to OH-PAHs using active oxidative enzymatic systems (Johnson-Restrepo et al., 2008; Sepic et al., 2003). Importantly, PAHs undergo photochemical transformation or hydroxyl oxidation to form the corresponding hydroxy compounds in water, ice and atmospheric environment (Barrado et al., 2012; Beltran et al., 1995; Dolinová et al., 2006; Murayama and Dasgupta, 1996; Zhang et al., 2011). These transformation reactions lead to the ubiquity of OH-PAHs in the environment. The OH-PAHs have been found widely in various biota (Lutz et al., 2006). In the atmospheric and aquatic environment, they were detected at the concentrations of 6–35 pg μg−1 in the air of Nanjing, China (Shen and Wang, 2012), 3–200 pg m−3 in the aerosol of Madrid, Spain (Barrado et al., 2012), 15–68 ng L−1 in the treated effluent from sewage treatment plants of Venice, Italy (Pojana and Marcomini, 2007), and 0.49–5.8 ng L−1 in the seawater of Shizuoka, Japan (Itoh et al., 2006). Therefore, increasing attention has been paid to OH-PAHs in terms of their environmental fate and impact to be understood. Particularly, quantification of the pertinent processes is a prerequisite to assess the risk of OH-PAHs to the aquatic environment. In sunlit surface waters, photochemical degradation is a key process in determining the fate of organic micropollutants, including PAHs, hydroxylated polybrominated diphenyl ethers (OH-PBDEs), and antibiotics (Edhlund et al., 2006; Fasnacht and Blough, 2002; Ge et al., 2010; Guerard et al., 2009; Vione et al., 2014; Xie et al., 2013). Many studies reported that PAHs underwent either direct or sensitized photochemical reactions (Fasnacht and Blough, 2003). However, little is known about the aqueous photodegradation of OH-PAHs. It can be hypothesized that most OH-PAHs with two or more arene rings absorb surface solar radiation, allowing for the possibility of direct photodegradation. − In addition, it is possible for Fe(III), NO− 3 , Cl , and dissolved organic matter (DOM) to affect photochemical processes of OH-PAHs separately or collectively (Vione et al., 2014). These photoreactive water constituents had been identified as significant participants in the photochemical degradation of PAHs (Fasnacht and Blough, 2002; Sankoda et al., 2013; Xia et al., 2009). Though OH-PAHs and the corresponding PAHs are structurally similar, the hydroxylated substitution might make certain distinctions between their photochemical behavior, such as for the case of OH-PBDEs and PBDEs (Bastos et al., 2009; Xie et al., 2013). Accordingly, investigating the aqueous photodegradation on OH-PAHs would be of interest to test these hypotheses. More interestingly, photochemical processes can induce the photoenhanced toxicities of pollutants, such as some PAHs (Grote et al., 2005; Lampi et al., 2006; Petersen et al., 2008; Shemer and Linden, 2007), antibiotics (Ge et al., 2010; Jiao et al., 2008; Jung et al., 2008; Latch et al., 2005; Li et al., 2011) and anthraquinones (Wang et al., 2009a; Wang et al., 2009b). One kind of photoinduced toxicity can be identified as photomodification to more toxic photoproducts (Wang et al., 2009a). As the main derivatives of PAHs, OH-PAHs potentially show the photomodified toxicity to aquatic systems. Therefore, it
is of interest and necessary to examine the phototoxicity of OH-PAHs. To indicate the toxicity of chemicals, the bioluminescent bacterium Vibrio fischeri has been used effectively as one of the most sensitive aquatic organisms (El-Alawi et al., 2001; Ge et al., 2010; Jiao et al., 2008). Herein, for the first time, we investigated their photochemical behavior with four OH-PAHs — 9-Hydroxyfluorene (9-OHFL), 2Hydroxyfluorene (2-OHFL), 9-Hydroxyphenanthrene (9-OHPH) and 1-Hydroxypyrene (1-OHPY) as examples. The four OH-PAHs were shown to be most widely used biomarkers (Guo et al., 2013; Luan et al., 2006). Of them, 9-OHFL was an abundant OH-PAH that had high levels in the environment or biota (Johnson-Restrepo et al., 2008). Their photolyzed kinetics and quantum yields were determined. 9OHFL was selected to demonstrate the multivariate effects of water in− gredients [Fe(III), NO− 3 , DOM and Cl ] on the photolysis. The potential photolysis pathways and the toxicity evolvement were also discussed. 2. Materials and methods 2.1. Chemicals Four OH-PAHs, including 9-OHFL, 2-OHFL, 9-OHPH and 1-OHPY (purity ≥ 98%), were obtained from Sigma-Aldrich. Their structures are shown in Table S1. Humic acid sodium salt (HASS, CAS No. 68131-044) was also purchased from Sigma-Aldrich. All organic solvents were HPLC grade and other chemical reagents were of analytical grade. Pure water obtained from a Millipore MilliQ system was used. Local fresh water and seawater were collected, filtered and characterized as described in our previous study (Ge et al., 2010). 2.2. Photodegradation experiments and analysis A Pyrex-well cooled and filtered xenon lamp (500 W) was used to simulate sunlight (λ N 290 nm). Four individual OH-PAH solutions with the initial concentration of 0.5 μM in quartz tubes were put on a merry-go-round reactor to carry out the photochemical experiments. The incident light intensity (290–450 nm) over the reactive solutions was 7.04 mW cm−2. Photodegradation experiments and dark controls were duplicated at least in triplicate. To compare photoreaction efficiencies of the four OH-PAHs, their apparent quantum yields in pure water were measured using p-nitroanisole/pyridine as a chemical actinometer (Dulln and Mill, 1982; Edhlund et al., 2006). A Waters UPLC was employed to analyze the OH-PAH concentrations, detailed in the Supplementary material. To test whether the 9-OHFL underwent the self-sensitized photodegradation via reactive oxygen species (ROS, such as •OH and 1 O2), the ROS scavenging experiments and competition kinetic examination were carried out in pure water. The bimolecular rate constant for the reaction between 9-OHFL and •OH was determined according to Eq. 1 S
kOH ¼
ln ð½St =½S0 Þ R k ln ð½Rt =½R0 Þ OH
ð1Þ
where S is the substrate and R is the reference compound, acetophenone, with a known rate constant of k·OH = 5.9 × 109 M−1 s−1 (Buxton et al., 1988; Edhlund et al., 2006). Then photolytic intermediates of 9-OHFL in pure water were enriched by solid phase extraction and one half of extracted samples were derivatized with silylation reagents. The underivatized and derivatized samples were analyzed by GC–MS/MS. Detailed descriptions of sample
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pretreatment and instrumental analysis are provided in the Supplementary material. Furthermore, photodegradation kinetics of 9-OHFL in pure water, freshwater and seawater was investigated comparatively. To explain kinetic difference in the diverse waters, multivariate effects of main water constituents on the 9-OHFL photodegradation kinetics were explored by a four-factor central composite design, of which the layout was shown in Table 1. The design layout and statistical analysis were performed using Design-Expert (version 7.0.0, Stat-Ease Inc., Minneapolis, MN) according to Ge et al. (Ge et al., 2010). 2.3. Bioassay In order to examine the photomodified toxicity of 9-OHFL, the 15 min acute ecotoxicity of photolyzed samples in pure water was evaluated according to an international standard method (ISO 11348-3). Luminous intensities of luminescent marine bacteria V. fischeri were recorded by a Microtox 500 Analyzer. Then luminescence inhibition rates (I%) were calculated as toxicity indexes (ISO, 2007).
be from 0.4 min for 9-OHPH in midsummer to 7.5 × 103 min for 9OHFL in midwinter (Table 2). Furthermore, The t1/2,E values were verified in field experiments and found comparable to the determined values under sunlight irradiation (Table S2). Compared to some PAHs, such as fluorene, phenanthrene and pyrene (Pagni and Dabestani, 2005; Schwarzenbach et al., 2003), the OH-PAHs showed very short half-lives against solar irradiation. Thus, the apparent photolysis can be a key process in determining the fate of OH-PAHs in sunlit surface waters. Apart from the apparent photolysis, aqueous organic pollutants may suffer from indirect photodegradation. The indirect phototransformation kinetics of xenobiotics in surface waters can be predicted as a function of surface-water parameters, such as NO− 3 , DOM, and water depth. To do this, APEX, Aqueous Photochemistry of Environmentally-occurring Xenobiotics, was recently developed and proved to be a successful software tool (Bodrato and Vione, 2014; Vione et al., 2014). Further studies are needed to select specific water bodies and calculate the corresponding t1/2,E for OH-PAHs. 3.2. Photochemical reaction types and intermediates of 9-OHFL
3. Results and discussion 3.1. Photolysis kinetics of four OH-PAHs in pure water When exposed to the simulated solar irradiation, the four OH-PAHs were susceptible to photodegradation in pure water, but no loss was found in dark controls. The OH-PAH attenuation by photolysis was confirmed to follow pseudo first-order kinetics with good linear regression (r2 N 0.95) of ln(C/C0) vs time (t). The photodegradation rate constants (k) are listed in Table 2. It was found that 9-OHPH photodegraded the fastest, followed by 1-OHPY, 2-OHFL and 9-OHFL. Their k values differed over two orders of magnitude. Such differences in the photodegradation can be attributed to the different quantum yields (Table 2) and spectral overlaps of the absorption spectra with the source emission spectrum (Fig. S1). However, these factors finally depend on the chemical structures of these compounds, including the hydroxyl substituent and benzene rings (Table S1). In surface waters, most organic pollutants absorb sunlight directly, and undergo direct photolysis and potential self-sensitized photodegradation, which are collectively called apparent photolysis (Ge et al., 2010). The environmentally relevant half-lives (t1/2,E) for the apparent photolysis in sunlit surface waters can be calculated according to Eqs. 2 and 3 (Edhlund et al., 2006; Ge et al., 2010), kE ¼ 2:303Φ∑ðZ λ ελ Þ
ð2Þ
ln 2 kE
ð3Þ
t 1=2;E ¼
where ελ are the molar absorptivities of the substrates, and Zλ is tabular solar photon flux at noon of summer and winter, assuming continuous irradiation. Here, the apparent photolysis t1/2,E is dependent on seasons and latitudes. For 45° latitude, Zλ value at every specific wavelength was taken as the average of the 40° and 50° latitude values (Leifer, 1988; OECD, 1997). Then the corresponding t1/2,E values were calculated to Table 1 Factors, levels and initial concentrations in the four-factor central composite design. Factors (units)
Fe(III) (μM) NO− 3 (μM) Cl− (M) HASSa (mg C L−1) a
Factor levels and concentrations −2
−1
0
1
2
0.00 0.00 0.00 0.00
2.00 0.20 0.125 2.00
4.00 0.40 0.250 4.00
6.00 0.60 0.375 6.00
8.00 0.80 0.500 8.00
Humic acid sodium salt (HASS).
9-OHFL was investigated in order to clarify the photochemical reaction types and transformation pathways. From the ROS scavenging experiments, it was observed that the presence of NaN3 (scavenger of •OH and 1O2) and isopropanol (scavenger of •OH) induced a notable retardation of 9-OHFL degradation to a similar extent (Fig. 1a). Thus, it can be inferred that that 9-OHFL underwent the self-sensitized photooxidation via •OH, and 1O2 might not be involved in the degradation. Furthermore, the inference was examined by the competition kinetics with acetophenone and furfural, which showed that •OH induced the oxidative degradation of 9-OHFL, whereas 1O2 did not react with the substrate (Fig. 1b). Therefore, the 9-OHFL photoreactions in pure water involved the self-sensitized photooxidation via •OH rather than 1O2. Referring to Eq. 4 (Boreen et al., 2008), the •OH oxidation contribution fraction (R•OH) was estimated to be 14.5%. ROH ≈
kPW −kPWþisopropanol kPW
ð4Þ
where kPW and kPW + isopropanol are the pseudo-first-order photolytic rate constants of 9-OHFL in pure water and in NaN3-adding water, respectively. The hydroxyl radical reaction rate constant (k·OH) was further calculated to be (6.99 ± 0.32) × 109 M−1 s−1 based on the competition kinetics. The k·OH of 9-OHFL is slightly more than that of fluorene with k·OH = 2.77 × 109 M−1 s−1 (Pagni and Dabestani, 2005). In sunlit surface waters, •OH are ubiquitous with the concentrations ranging from 10−17 M to 10−15 M (Cooper et al., 1989). As aqueous dominant photooxidants, •OH can oxidize almost all classes of organic chemicals (Mill, 1999). The highest expected •OH level shows the shortest possible half-live (t1/2) for the oxidation process. Thus, the corresponding t1/2 is calculated to be 27.5 h for •OH oxidation toward 9-OHFL in sunlit surface waters. The oxidation is slower than that of some other compounds, such as anthracene (Mill, 1999). Total ion chromatograms (TIC) for underivatized and derivatized samples of photolyzed 9-OHFL in pure water were shown in Fig. S2, suggesting that there were 8 significant intermediates generated from the photodegradation of 9-OHFL. The product with molecular weight of 180 (P180) and the parent 9-OHFL showed same retention times (tR) before derivation. After derivation, the tR of P180 did not change, indicative of no inclusion of active hydrogens in its structure. The 9-OHFL and other products, such as isomers with molecular weights of 182 and 196, showed different tR before and after derivation, which indicated their structures contained active hydrogens, allowing their reaction with silylation reagents. One isomeride named as P182 displayed same tR and mass spectra with 2-OHFL.
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Table 2 Photolytic rate constants (k), quantum yields (Φ), and corresponding environmental half-lives (t1/2,E at 45° N latitude) based on apparent photolysis Including direct and self-sensitized photodegradation. Compounds
Abbr.
k (min−1)
t1/2,E (min)a
Φ
Summer 9-Hydroxyfluorene 2-Hydroxyfluorene 9-Hydroxyphenanthrene 1-Hydroxypyrene a
9-OHFL 2-OHFL 9-OHPH 1-OHPY
−3
−2
5.50 × 10 2.04 × 10−2 1.71 × 10−1 3.20 × 10−2
6.48 × 10 5.03 × 10−2 5.01 × 10−1 3.33 × 10−2
5.5 × 10 3.2 × 10 0.4 1.8
2
Winter 7.5 × 103 2.3 × 102 1.8 7.6
Values have been multiplied by 2 to correct for the increase in rate caused by the lens effect caused by the curved test tubes used in the experiments.
Based on the mass spectra (Fig. S3) and the NIST mass spectral library, molecular weights and structures were proposed for the products. According to the photoproducts, their photolysis pathways are proposed in Fig. 2, which involve photoinduced hydroxylation, dehydrogenation and isomerization. The hydroxylation formed multiple hydroxylated fluorenones, which verified the self-sensitized photooxidation of 9-OHFL via •OH. Hydroxylated or oxidized PAHs are often identified for photolysis of PAHs and phenolic compounds in in different waters (Calza et al., 2014; Sanches et al., 2011). Referring to PAHs (Kahan and Donaldson, 2007; Sanches et al., 2011), proposed mechanisms for the OH-PAH photolysis involve either direct ionization as the initial reaction step, or interaction with reactive oxygen species, such as •OH. 3.3. Multivariate effects of water constituents on 9-OHFL photolysis When irradiated by simulated sunlight (λ N 290 nm), the model compound underwent photodegradation in different waters, with rates in seawater and freshwater being lower than that in pure water (Fig. 3). The less photodegradable potential of 9-OHFL in natural waters could be attributed to the multivariate effects of different water constituents on the photodegradation (Ge et al., 2010; Lam et al., 2003). To verify the hypothesis, the multivariate effects of main natural water constituents (e.g., DOM, Cl−, Fe(III), NO− 3 ) on the photolysis were studied using the central composite experiments. The photodegradation rate constants (k) for the experiments were listed in Table S3. The k values and the levels (x1 − x4) of the four factors − (Fe(III), NO− 3 , Cl and HASS) were evaluated by fitting a full quadratic expression: k ¼ β0 þ β1 x1 þ β2 x2 þ β3 x3 þ β4 x4 þ β12 x1 x2 þ β13 x1 x3 þ β14 x1 x4 þ β23 x2 x3 þ β24 x2 x4 þ β34 x3 x4 þ β11 x1 2 þ β22 x2 2 þ β33 x3 2 þ β44 x4 2: ð5Þ The fitted regression parameters (βx) and the corresponding significance level (p) were shown in Table 3. If p b 0.05, the corresponding βx keys are significant contributing factors. Negative βx values of Fe − were observed, indicating that these three factors (III), NO− 3 and Cl suppressed the photodegradation of 9-OHFL. In contrast, HASS and the
interaction between Fe(III) and Cl− both enhanced the photodegradation. The terms β11, β33 and β44 are positive, which may be attributed to the nonlinear effects of Fe(III), NO− 3 and HASS on the 9-OHFL photodegradation. Thus, the less photodegradable potential of 9-OHFL in natural waters than in pure water was attributed to the integrative effects of these significant contributing factors. − The photoreactive species, Fe(III), NO− 3 , Cl and HASS, usually show multiple effects. They suppress the photodegradation of one pollutant due to their competitive photoabsorption and/or ROS scavenging, while they accelerate the photodegradation because of sensitizing effect. These effects sometimes coexist. Which one effect is stronger or more apparent depends on the pollutant properties, and the ambient conditions, such as light sources and aqueous pH values (Ge et al., 2009; Niu et al., 2013; Vione et al., 2014). Fe(III) and NO− 3 have strong overlapping absorption spectra with 9-OHFL near 290 nm (Mack and Bolton, 1999; Neamtu et al., 2009; Vione et al., 2014). Therefore, they competitively absorb actinic photons (λ = 290–315 nm). Although Fe (III) and NO− 3 might sensitize the 9-OHFL photodegradation, the experimental results indicated that the inhibiting effect caused by competitive photoabsorption was more pronounced than the sensitization effect. In contrast, overall, HASS enhanced the photodegradation by sensitization, though HASS also has overlapping absorption spectra with 9OHFL (Fig. S4). The inhibitive effect of Cl− might be due to its •OH scavenging and deceleration of the 9-OHFL self-sensitized photooxidation via •OH: Cl þ OH→ClHO − −
and Cl− might transform into Cl• radicals when interacting with Fe(III), which would facilitate the photodegradation (Chiron et al., 2006; Liu et al., 2009): 2þ
FeCl
þ hν→Fe2þ þ Cl :
3.4. Photomodified toxicity of 9-OHFL to V. fischeri More toxic products might be formed during the photochemical transformation of some compounds such as PAHs and antibiotics,
Fig. 1. Effects of NaN3 and isopropanol on photodegradation kinetics of 9-OHFL (C0 = 0.5 μM), and photooxidation reactivities of 9-OHFL with •OH and 1O2
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Fig. 2. Proposed transformation pathways for photodegradation of 9-OHFL in pure water and under simulated sunlight (λ N 290 nm). The photoproducts are labeled “Pmw”, with mw standing for the molecular weights of the intermediates.
which is defined as photomodified toxicities (Ge et al., 2010; Jiao et al., 2008; Li et al., 2011; Niu et al., 2013; Schwarzenbach et al., 2006; Shemer and Linden, 2007; Wang et al., 2009b). When 9-OHFL photodegraded, the parent compound decayed and intermediates formed, probably resulting in the changing toxicity of their solutions. The toxicity evolvement of the 9-OHFL in waters during the irradiation was assayed using V. fischeri (Fig. 4). As shown in Fig. 4, the 9-OHFL solutions have a similar toxicity evolvement trend. Either in pure water or in seawater, the luminescence inhibition rate of photolyzed 9-OHFL did not change significantly (p N 0.05), though the parent compound concentration decreased. This suggested that main intermediates still showed comparable or higher toxicity compared to parent compound. Combined with the product structures (Fig. 2), 9-fluorenone is more hydrophobic than 9-OHFL, which may result in easier penetration into the cells of luminescent bacteria and subsequent notable toxicity (Li et al., 2011; Mallakin et al., 1999; McConkey et al., 1997; Niu et al., 2013). Alternatively, fluorenone is hydroxylated, leading to the enhancement of toxicity.
4. Conclusion This study demonstrated the aqueous photochemical behavior of OH-PAHs, which followed pseudo-first-order photodegradation kinetics, with apparent quantum yields and solar photodegradation halflives varying according to their chemical structures. Compared with corresponding parent PAHs, the OH-PAHs would intrinsically photodegrade fast in sunlit surface waters. 9-OHFL for example, showed to undergo both direct photolysis and self-sensitized photooxidation via •OH rather than 1O2 in pure water. Photoinduced hydroxylation, dehydrogenation and isomerization were involved in these photoreactions. 9-OHFL photodegraded slower in natural waters than in pure water, attributing to the integrative effects of the most photoreactive species, − such as Fe(III), NO− 3 , Cl and humic acid. The toxicity of 9-OHFL solutions (in pure water or seawater) did not decrease significantly (p N 0.05) after irradiation potentially due to the comparable or higher toxicity of some intermediates. The important information about photochemical behavior of OH-PAHs would better understanding the fate of these chemicals in the aqueous environment and improve their risk assessment. Table 3 Parameter estimates and hypothesis tests for the coefficients of the quadratic model fitted to the photodegradation data of 9-OHFL.
Fig. 3. Degradation kinetics of 9-OHFL (C0 = 0.5 μM) in different waters under simulated solar irradiation (λ N 290 nm).
Parameter
βx key
βx × 105
Sum of squares × 109
pa
β0 β1 β2 β3 β4 β12 β13 β14 β23 β24 β34 β11 β22 β33 β44
Intercept Fe(III) NO− 3 Cl− HASS Fe(III)–NO− 3 Fe(III)–Cl− Fe(III)–HASS − NO− 3 Cl NO− 3 HASS Cl− HASS (Fe(III))2 2 (NO− 3 ) (Cl−)2 (HASS)2
382 −16.9 −15.7 −17.2 16.1 2.92 27.3 3.54 −5.21 −10.6 −2.08 28.6 −4.13 21.7 20.0
689 591 712 623 13.6 1192 20.1 43.4 181 6.94 2240 46.8 1292 1101
0.032 0.044 0.029 0.040 0.744 0.007 0.692 0.561 0.244 0.815 0.001 0.546 0.005 0.009
a
Tests as significant at the 95% confidence level.
L. Ge et al. / Science of the Total Environment 547 (2016) 166–172
Fig. 4. Bioluminescent inhibition rate (I) of photolyzed 9-Hydroxyfluorene (Concentration, C) to Vibrio fischeri in different waters.
Acknowledgments The study was supported by the National Natural Science Foundation of China (21577029, 21007013, 41476084, and 21207025), the Key Lab of Marine Bioactive Substance and Modern Analytical Technique (MBSMAT-2014-04), the Key Laboratory for Ecological Environment in Coastal Areas (201602), the Polar Science Strategic Research Foundation (20120320), and the State Oceanic Administration Specific Public Project (201105013) of China. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2015.12.143. References Barrado, A.I., Garcia, S., Barrado, E., Perez, R.M., 2012. PM2.5-bound PAHs and hydroxyPAHs in atmospheric aerosol samples: correlations with season and with physical and chemical factors. Atmos. Environ. 49, 224–232. Bastos, P.M., Eriksson, J., Bergman, Å., 2009. Photochemical decomposition of dissolved hydroxylated polybrominated diphenyl ethers under various aqueous conditions. Chemosphere 77, 791–797. Beltran, F.J., Ovejero, G., Garcia-Araya, J.F., Rivas, J., 1995. Oxidation of polynuclear aromatic hydrocarbons in water. 2. UV radiation and ozonation in the presence of UV radiation. Ind. Eng. Chem. Res. 34, 1607–1615. Bodrato, M., Vione, D., 2014. APEX (Aqueous Photochemistry of Environmentally occurring Xenobiotics): a free software tool to predict the kinetics of photochemical processes in surface waters. Environ. Sci.-Proc. Imp. 16, 732–740. Boreen, A.L., Edhlund, B.L., Cotner, J.B., McNeill, K., 2008. Indirect photodegradation of dissolved free amino acids: the contribution of singlet oxygen and the differential reactivity of DOM from various sources. Environ. Sci. Technol. 42, 5492–5498. Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., 1988. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (•OH/•O−) in aqueous solution. J. Phys. Chem. Ref. Data 17, 513–886. Calza, P., Vione, D., Minero, C., 2014. The role of humic and fulvic acids in the phototransformation of phenolic compounds in seawater. Sci. Total Environ. 493, 411–418. Chiron, S., Minero, C., Vione, D., 2006. Photodegradation processes of the antiepileptic drug carbamazepine, relevant to estuarine waters. Environ. Sci. Technol. 40, 5977–5983. Cochran, R.E., Dongari, N., Jeong, H., Beránek, J., Haddadi, S., Shipp, J., Kubátová, A., 2012. Determination of polycyclic aromatic hydrocarbons and their oxy-, nitro-, and hydroxy-oxidation products. Anal. Chim. Acta 740, 93–103. Cooper, W.J., Zika, R.G., Petasne, R.G., Fischer, A.M., 1989. In: Suffet, I.H., MacCarthy, P. (Eds.), Sunlight-induced photochemistry of humic substances in natural waters: major reactive species. American Chemical Society, Washington, DC. Dolinová, J., Ružička, R., Kurková, R., Klánová, J., Klán, P., 2006. Oxidation of aromatic and aliphatic hydrocarbons by OH radicals photochemically generated from H2O2 in ice. Environ. Sci. Technol. 40, 7668–7674. Dulln, D., Mill, T., 1982. Development and evaluation of sunlight actinometers. Environ. Sci. Technol. 16, 815–820. Edhlund, B.L., Arnold, W.A., McNeill, K., 2006. Aquatic photochemistry of nitrofuran antibiotics. Environ. Sci. Technol. 40, 5422–5427. El-Alawi, Y.S., Dixon, D.G., Greenberg, B.M., 2001. Effects of a pre-incubation period on the photoinduced toxicity of polycyclic aromatic hydrocarbons to the luminescent bacterium Vibrio fischeri. Environ. Toxicol. 16, 277–286.
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Fasnacht, M.P., Blough, N.V., 2002. Aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 36, 4364–4369. Fasnacht, M.P., Blough, N.V., 2003. Mechanisms of the aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 37, 5767–5772. Ge, L.K., Chen, J.W., Qiao, X.L., Lin, J., Cai, X.Y., 2009. Light-source-dependent effects of main water constituents on photodegradation of phenicol antibiotics: mechanism and kinetics. Environ. Sci. Technol. 43, 3101–3107. Ge, L.K., Chen, J.W., Wei, X.X., Zhang, S.Y., Qiao, X.L., Cai, X.Y., Xie, Q., 2010. Aquatic photochemistry of fluoroquinolone antibiotics: kinetics, pathways, and multivariate effects of main water constituents. Environ. Sci. Technol. 44, 2400–2405. Grote, M., Schüürmann, G., Altenburger, R., 2005. Modeling photoinduced algal toxicity of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 39, 4141–4149. Guerard, J.J., Chin, Y.P., Mash, H., Hadad, C.M., 2009. Photochemical fate of sulfadimethoxine in aquaculture waters. Environ. Sci. Technol. 43, 8587–8592. Guo, Y., Senthilkumar, K., Alomirah, H., Moon, H.B., Minh, T.B., Mohd, M.A., Nakata, H., Kannan, K., 2013. Concentrations and profiles of urinary polycyclic aromatic hydrocarbon metabolites (OH-PAHs) in several Asian countries. Environ. Sci. Technol. 47, 2932–2938. ISO, 2007. Water Quality—Determination of the Inhibitory Effect of Water Samples on the Light Emission of Vibrio fischeri (Luminescent Bacteria Test)—Part 3: Method Using Freeze-dried Bacteria (ISO 11348-3). The International Organization for Standardization. Itoh, N., Tao, H., Ibusuki, T., 2006. In-tube silylation in combination with thermal desorption gas chromatography–mass spectrometry for the determination of hydroxy polycyclic aromatic hydrocarbons in water. Anal. Chim. Acta 555, 201–209. Jiao, S.J., Zheng, S.R., Yin, D.Q., Wang, L.H., Chen, L.Y., 2008. Aqueous photolysis of tetracycline and toxicity of photolytic products to Luminescent bacteria. Chemosphere 73, 377–382. Johnson-Restrepo, B., Olivero-Verbel, J., Lu, S.J., Guette-Fernández, J., Baldiris-Avila, R., O'Byrne-Hoyos, I., Aldousa, K.M., Addink, R., Kannan, K., 2008. Polycyclic aromatic hydrocarbons and their hydroxylated metabolites in fish bile and sediments from coastal waters of Colombia. Environ. Pollut. 151, 452–459. Jung, J., Kim, Y., Kim, J., Jeong, D.H., Choi, K., 2008. Environmental levels of ultraviolet light potentiate the toxicity of sulfonamide antibiotics in Daphnia magna. Ecotoxicology 17, 37–45. Kahan, T.F., Donaldson, D.J., 2007. Photolysis of polycyclic aromatic hydrocarbons on water and ice surfaces. J. Phys. Chem. A 111, 1277–1285. Lam, M.W., Tantuco, K., Mabury, S.A., 2003. PhotoFate: a new approach in accounting for the contribution of indirect photolysis of pesticides and pharmaceuticals in surface waters. Environ. Sci. Technol. 37, 899–907. Lampi, M.A., Gurska, J., McDonald, K.I., Xie, F., Huang, X.D., Dixon, D.G., Greenberg, B.M., 2006. Photoinduced toxicity of polycyclic aromatic hydrocarbons to Daphnia magna: ultraviolet-mediated effects and the toxicity of polycyclic aromatic hydrocarbon photoproducts. Environ. Toxicol. Chem. 25, 1079–1087. Latch, D.E., Packer, J.L., Stender, B.L., VanOverbeke, J., Arnold, W.A., McNeill, K., 2005. Aqueous photochemistry of triclosan: formation of 2,4-dichlorophenol, 2,8-dichlorodibenzo-pdioxin, and oligomerization products. Environ. Toxicol. Chem. 24, 517–525. Leifer, A., 1988. The Kinetics of Environmental Aquatic Photochemistry: Theory and Practice. American Chemical Society, Washington, DC. Li, Y., Niu, J.F., Wang, W.L., 2011. Photolysis of enrofloxacin in aqueous systems under simulated sunlight irradiation: kinetics, mechanism and toxicity of photolysis products. Chemosphere 85, 892–897. Liu, H., Zhao, H., Quan, X., Zhang, Y., Chen, S., 2009. Formation of chlorinated intermediate from bisphenol A in surface saline water under simulated solar light irradiation. Environ. Sci. Technol. 43, 7712–7717. Luan, T.G., Yu, K.S.H., Zhong, Y., Zhou, H.W., Lan, C.Y., Tam, N.F.Y., 2006. Study of metabolites from the degradation of polycyclic aromatic hydrocarbons (PAHs) by bacterial consortium enriched from mangrove sediments. Chemosphere 65, 2289–2296. Lutz, S., Feidt, C., Monteau, F., Rychen, G., Le Bizec, B., Jurjanz, S., 2006. Effect of exposure to soil-bound polycyclic aromatic hydrocarbons on milk contaminations of parent compounds and their monohydroxylated metabolites. J. Agric. Food Chem. 54, 263–268. Mack, J., Bolton, J.R., 1999. Photochemistry of nitrite and nitrate in aqueous solution: a review. J. Photochem. Photobiol. A Chem. 128, 1–13. Mallakin, A., McConkey, B.J., Miao, G., McKibben, B., Snieckus, V., Dixon, D.G., Greenberg, B.M., 1999. Impacts of structural photomodification on the toxicity of environmental contaminants: anthracene photooxidation products. Ecotoxicol. Environ. Saf. 43, 204–212. McConkey, B.J., Duxbury, C.L., Dixon, D.G., Greenberg, B.M., 1997. Toxicity of a PAH photooxidation product to the bacteria Photobacterium phosphoreum and the duckweed Lemna gibba: effects of phenanthrene and its primary photoproduct, phenanthrenequinone. Environ. Toxicol. Chem. 16, 892–899. Mill, T., 1999. Predicting photoreaction rates in surface waters. Chemosphere 38, 1379–1390. Murayama, M., Dasgupta, P.K., 1996. Liquid chromatographic determination of nitrosubstituted oolynuclear aromatic hydrocarbons by sequential electrochemical and fluorescence detection. Anal. Chem. 68, 1226–1232. Neamtu, M., Popa, D.M., Frimmel, F.H., 2009. Simulated solar UV-irradiation of endocrine disrupting chemical octylphenol. J. Hazard. Mater. 164, 1561–1567. Niu, J.F., Li, Y., Wang, W.L., 2013. Light-source-dependent role of nitrate and humic acid in tetracycline photolysis: kinetics and mechanism. Chemosphere 92, 1423–1429. OECD, 1997. Guidance Document on Direct Phototransformation of Chemicals in Water. OECD Environmental Health and Safety Publication. Series on Testing and Assessment No.7, Paris, France. Pagni, R.M., Dabestani, R., 2005. Recent developments in the environmental photochemistry of PAHs and PCBs in water and on solids. Environmental Photochemistry Part II. Springer Berlin Heidelberg, pp. 193–219.
172
L. Ge et al. / Science of the Total Environment 547 (2016) 166–172
Petersen, D.G., Reichenberg, F., Dahllöf, I., 2008. Phototoxicity of pyrene affects benthic algae and bacteria from the Arctic. Environ. Sci. Technol. 42, 1371–1376. Pojana, G., Marcomini, A., 2007. Determination of monohydroxylated metabolites of polycyclic aromatic hydrocarbons (OH-PAHs) from wastewater-treatment plants. Int. J. Environ. An. Ch. 87, 627–636. Rey-Salgueiro, L., Martínez-Carballo, E., García-Falcón, M.S., González-Barreiro, C., SimalGándara, J., 2009. Occurrence of polycyclic aromatic hydrocarbons and their hydroxylated metabolites in infant foods. Food Chem. 115, 814–819. Sanches, S., Leitão, C., Penetra, A., Cardoso, V.V., Ferreira, E., Benoliel, M.J., Barreto Crespo, M.T., Pereira, V.J., 2011. Direct photolysis of polycyclic aromatic hydrocarbons in drinking water sources. J. Hazard. Mater. 192, 1458–1465. Sankoda, K., Kuribayashi, T., Nomiyama, K., Shinohara, R., 2013. Occurrence and source of chlorinated polycyclic aromatic hydrocarbons (Cl-PAHs) in tidal flats of the Ariake Bay, Japan. Environ. Sci. Technol. 47, 7037–7044. Schwarzenbach, R.P., Gschwend, P.M., Imboden, D.M., 2003. Environmental Organic Chemistry. second ed. John Wiley & Sons, Inc., New Jersey. Schwarzenbach, R.P., Escher, B.I., Fenner, K., Hofstetter, T.B., Johnson, C.A., von Gunten, U., Wehrli, B., 2006. The challenge of micropollutants in aquatic systems. Science 313, 1072–1077. Sepic, E., Bricelj, M., Leskovsek, H., 2003. Toxicity of fluoranthene and its biodegradation metabolites to aquatic organisms. Chemosphere 52, 1125–1133. Shemer, H., Linden, K.G., 2007. Aqueous photodegradation and toxicity of the polycyclic aromatic hydrocarbons fluorene, dibenzofuran, and dibenzothiophene. Water Res. 41, 853–861.
Shen, G.F., Wang, G.H., 2012. Molecular composition and seasonal variation of hydroxylated polycyclic aromatic hydrocarbons in atmospheric PM2.5 of Nanjing, China. J. Earth Environ. 3, 1066–1069. Vione, D., Minella, M., Maurino, V., Minero, C., 2014. Indirect photochemistry in sunlit surface waters: photoinduced production of reactive transient species. Chem. Eur. J. 20, 10590–10606. Wang, Y., Chen, J.W., Ge, L.K., Wang, D.G., Cai, X.Y., Huang, L.P., Hao, C., 2009a. Experimental and theoretical studies on the photoinduced acute toxicity of a series of anthraquinone derivatives towards the water flea (Daphnia magna). Dyes Pigments 83, 276–280. Wang, Y., Chen, J.W., Lin, J., Wang, Z., Bian, H.T., Cai, X.Y., Hao, C., 2009b. Combined experimental and theoretical study on photoinduced toxicity of an anthraquinone dye intermediate to Daphnia magna. Environ. Toxicol. Chem. 28, 846–852. Xia, X., Li, G., Yang, Z., Chen, Y., Huang, G.H., 2009. Effects of fulvic acid concentration and origin on photodegradation of polycyclic aromatic hydrocarbons in aqueous solution: importance of active oxygen. Environ. Pollut. 157, 1352–1359. Xie, Q., Chen, J.W., Zhao, H.X., Qiao, X.L., Cai, X.Y., Li, X.H., 2013. Different photolysis kinetics and photooxidation reactivities of neutral and anionic hydroxylated polybrominated diphenyl ethers. Chemosphere 90, 188–194. Zhang, Y., Yang, B., Gan, J., Liu, C.G., Shu, X., Shu, J.N., 2011. Nitration of particle-associated PAHs and their derivatives (nitro-, oxy-, and hydroxy-PAHs) with NO3 radicals. Atmos. Environ. 45, 2515–2521.
Contents lists available at ScienceDirect
Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv
Aqueous photochemical degradation of hydroxylated PAHs: Kinetics, pathways, and multivariate effects of main water constituents Linke Ge a, Guangshui Na a, Chang-Er Chen b, Jun Li a,c, Maowei Ju a, Ying Wang a, Kai Li a, Peng Zhang a,⁎, Ziwei Yao a a b c
Key Laboratory for Ecological Environment in Coastal Areas (SOA), National Marine Environmental Monitoring Center, Dalian 116023, China Lancaster Environment Centre, Lancaster University, Lancaster LA1 4YQ, United Kingdom College of Marine Science, Shanghai Ocean University, Shanghai 201306, China
H I G H L I G H T S
G R A P H I C A L
A B S T R A C T
• It is first reported on aqueous photochemical behavior of 4 hydroxylated PAHs. • Hydroxylated PAHs intrinsically photodegrade fast in sunlit surface waters. • Reaction types and transformation pathways of 9-Hydroxyfluorene were clarified. • Photolysis kinetics was affected by multivariate effects of main water constituents. • The photomodified toxicity of 9Hydroxyfluorene was examined using Vibrio fischeri.
Aqueous photochemical behavior of 4 hydroxylated PAHs is first reported on revealing the kinetics, mechanisms, toxicity, and multivariate effects of water constituents.
a r t i c l e
i n f o
Article history: Received 23 September 2015 Received in revised form 28 December 2015 Accepted 29 December 2015 Available online xxxx Editor: Kevin V. Thomas Keywords: Hydroxylated PAHs 9-Hydroxyfluorene Photodegradation kinetics Pathways Water multivariate effects Photomodified toxicity
⁎ Corresponding author. E-mail address: [email protected] (P. Zhang).
http://dx.doi.org/10.1016/j.scitotenv.2015.12.143 0048-9697/© 2015 Elsevier B.V. All rights reserved.
a b s t r a c t Hydroxylated polycyclic aromatic hydrocarbons (OH-PAHs) are contaminants of emerging concern in the aquatic environment, so it is of great significance to understand their environmental transformation and toxicity. This study investigated the aqueous photochemical behavior of four OH-PAHs, 9-Hydroxyfluorene (9-OHFL), 2Hydroxyfluorene, 9-Hydroxyphenanthrene and 1-Hydroxypyrene, under simulated sunlight irradiation (λ N 290 nm). It was observed that their photodegradation followed the pseudo-first-order kinetics. Based on the determined quantum yields, their calculated solar apparent photodegradation half-lives in surface waters at 45° N latitude ranged from 0.4 min for 9-Hydroxyphenanthrene to 7.5 × 103 min for 9-OHFL, indicating that the OH-PAHs would intrinsically photodegrade fast in sunlit surface waters. Furthermore, 9-OHFL as an example was found to undergo direct photolysis, and self-sensitized photooxidation via •OH rather than 1O2 in pure water. The potential photoreactions involved photoinduced hydroxylation, dehydrogenation and isomerization based on product identification by GC–MS/MS. 9-OHFL photodegraded slower in natural waters than in pure water, − which was attributed to the integrative effects of the most photoreactive species, such as Fe(III), NO− 3 , Cl and humic acid. The photomodified toxicity was further examined using Vibrio fischeri, and it was found that the toxicity of photolyzed 9-OHFL did not decrease significantly (p N 0.05) either in pure water or in seawater, implying
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the comparable or higher toxicity of some intermediates. These results are important for assessing the fate and risks of OH-PAHs in surface waters. © 2015 Elsevier B.V. All rights reserved.
1. Introduction Hydroxylated polycyclic aromatic hydrocarbons (OH-PAHs) are newly recognized contaminants derived from hydroxylation of PAHs. They are commonly used as biomarkers of human exposure to PAHs (Guo et al., 2013), and some of them have been proved to be of even more pronounced mutagenicity/toxicity than their original PAH precursors (Cochran et al., 2012). Various sources can contribute to the contaminants in the environment. Once absorbed in humans and animals, PAHs are metabolized to a complex mixture of products including OHPAHs (Rey-Salgueiro et al., 2009). In sediments/soils, some microorganisms may biotransform PAHs to OH-PAHs using active oxidative enzymatic systems (Johnson-Restrepo et al., 2008; Sepic et al., 2003). Importantly, PAHs undergo photochemical transformation or hydroxyl oxidation to form the corresponding hydroxy compounds in water, ice and atmospheric environment (Barrado et al., 2012; Beltran et al., 1995; Dolinová et al., 2006; Murayama and Dasgupta, 1996; Zhang et al., 2011). These transformation reactions lead to the ubiquity of OH-PAHs in the environment. The OH-PAHs have been found widely in various biota (Lutz et al., 2006). In the atmospheric and aquatic environment, they were detected at the concentrations of 6–35 pg μg−1 in the air of Nanjing, China (Shen and Wang, 2012), 3–200 pg m−3 in the aerosol of Madrid, Spain (Barrado et al., 2012), 15–68 ng L−1 in the treated effluent from sewage treatment plants of Venice, Italy (Pojana and Marcomini, 2007), and 0.49–5.8 ng L−1 in the seawater of Shizuoka, Japan (Itoh et al., 2006). Therefore, increasing attention has been paid to OH-PAHs in terms of their environmental fate and impact to be understood. Particularly, quantification of the pertinent processes is a prerequisite to assess the risk of OH-PAHs to the aquatic environment. In sunlit surface waters, photochemical degradation is a key process in determining the fate of organic micropollutants, including PAHs, hydroxylated polybrominated diphenyl ethers (OH-PBDEs), and antibiotics (Edhlund et al., 2006; Fasnacht and Blough, 2002; Ge et al., 2010; Guerard et al., 2009; Vione et al., 2014; Xie et al., 2013). Many studies reported that PAHs underwent either direct or sensitized photochemical reactions (Fasnacht and Blough, 2003). However, little is known about the aqueous photodegradation of OH-PAHs. It can be hypothesized that most OH-PAHs with two or more arene rings absorb surface solar radiation, allowing for the possibility of direct photodegradation. − In addition, it is possible for Fe(III), NO− 3 , Cl , and dissolved organic matter (DOM) to affect photochemical processes of OH-PAHs separately or collectively (Vione et al., 2014). These photoreactive water constituents had been identified as significant participants in the photochemical degradation of PAHs (Fasnacht and Blough, 2002; Sankoda et al., 2013; Xia et al., 2009). Though OH-PAHs and the corresponding PAHs are structurally similar, the hydroxylated substitution might make certain distinctions between their photochemical behavior, such as for the case of OH-PBDEs and PBDEs (Bastos et al., 2009; Xie et al., 2013). Accordingly, investigating the aqueous photodegradation on OH-PAHs would be of interest to test these hypotheses. More interestingly, photochemical processes can induce the photoenhanced toxicities of pollutants, such as some PAHs (Grote et al., 2005; Lampi et al., 2006; Petersen et al., 2008; Shemer and Linden, 2007), antibiotics (Ge et al., 2010; Jiao et al., 2008; Jung et al., 2008; Latch et al., 2005; Li et al., 2011) and anthraquinones (Wang et al., 2009a; Wang et al., 2009b). One kind of photoinduced toxicity can be identified as photomodification to more toxic photoproducts (Wang et al., 2009a). As the main derivatives of PAHs, OH-PAHs potentially show the photomodified toxicity to aquatic systems. Therefore, it
is of interest and necessary to examine the phototoxicity of OH-PAHs. To indicate the toxicity of chemicals, the bioluminescent bacterium Vibrio fischeri has been used effectively as one of the most sensitive aquatic organisms (El-Alawi et al., 2001; Ge et al., 2010; Jiao et al., 2008). Herein, for the first time, we investigated their photochemical behavior with four OH-PAHs — 9-Hydroxyfluorene (9-OHFL), 2Hydroxyfluorene (2-OHFL), 9-Hydroxyphenanthrene (9-OHPH) and 1-Hydroxypyrene (1-OHPY) as examples. The four OH-PAHs were shown to be most widely used biomarkers (Guo et al., 2013; Luan et al., 2006). Of them, 9-OHFL was an abundant OH-PAH that had high levels in the environment or biota (Johnson-Restrepo et al., 2008). Their photolyzed kinetics and quantum yields were determined. 9OHFL was selected to demonstrate the multivariate effects of water in− gredients [Fe(III), NO− 3 , DOM and Cl ] on the photolysis. The potential photolysis pathways and the toxicity evolvement were also discussed. 2. Materials and methods 2.1. Chemicals Four OH-PAHs, including 9-OHFL, 2-OHFL, 9-OHPH and 1-OHPY (purity ≥ 98%), were obtained from Sigma-Aldrich. Their structures are shown in Table S1. Humic acid sodium salt (HASS, CAS No. 68131-044) was also purchased from Sigma-Aldrich. All organic solvents were HPLC grade and other chemical reagents were of analytical grade. Pure water obtained from a Millipore MilliQ system was used. Local fresh water and seawater were collected, filtered and characterized as described in our previous study (Ge et al., 2010). 2.2. Photodegradation experiments and analysis A Pyrex-well cooled and filtered xenon lamp (500 W) was used to simulate sunlight (λ N 290 nm). Four individual OH-PAH solutions with the initial concentration of 0.5 μM in quartz tubes were put on a merry-go-round reactor to carry out the photochemical experiments. The incident light intensity (290–450 nm) over the reactive solutions was 7.04 mW cm−2. Photodegradation experiments and dark controls were duplicated at least in triplicate. To compare photoreaction efficiencies of the four OH-PAHs, their apparent quantum yields in pure water were measured using p-nitroanisole/pyridine as a chemical actinometer (Dulln and Mill, 1982; Edhlund et al., 2006). A Waters UPLC was employed to analyze the OH-PAH concentrations, detailed in the Supplementary material. To test whether the 9-OHFL underwent the self-sensitized photodegradation via reactive oxygen species (ROS, such as •OH and 1 O2), the ROS scavenging experiments and competition kinetic examination were carried out in pure water. The bimolecular rate constant for the reaction between 9-OHFL and •OH was determined according to Eq. 1 S
kOH ¼
ln ð½St =½S0 Þ R k ln ð½Rt =½R0 Þ OH
ð1Þ
where S is the substrate and R is the reference compound, acetophenone, with a known rate constant of k·OH = 5.9 × 109 M−1 s−1 (Buxton et al., 1988; Edhlund et al., 2006). Then photolytic intermediates of 9-OHFL in pure water were enriched by solid phase extraction and one half of extracted samples were derivatized with silylation reagents. The underivatized and derivatized samples were analyzed by GC–MS/MS. Detailed descriptions of sample
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pretreatment and instrumental analysis are provided in the Supplementary material. Furthermore, photodegradation kinetics of 9-OHFL in pure water, freshwater and seawater was investigated comparatively. To explain kinetic difference in the diverse waters, multivariate effects of main water constituents on the 9-OHFL photodegradation kinetics were explored by a four-factor central composite design, of which the layout was shown in Table 1. The design layout and statistical analysis were performed using Design-Expert (version 7.0.0, Stat-Ease Inc., Minneapolis, MN) according to Ge et al. (Ge et al., 2010). 2.3. Bioassay In order to examine the photomodified toxicity of 9-OHFL, the 15 min acute ecotoxicity of photolyzed samples in pure water was evaluated according to an international standard method (ISO 11348-3). Luminous intensities of luminescent marine bacteria V. fischeri were recorded by a Microtox 500 Analyzer. Then luminescence inhibition rates (I%) were calculated as toxicity indexes (ISO, 2007).
be from 0.4 min for 9-OHPH in midsummer to 7.5 × 103 min for 9OHFL in midwinter (Table 2). Furthermore, The t1/2,E values were verified in field experiments and found comparable to the determined values under sunlight irradiation (Table S2). Compared to some PAHs, such as fluorene, phenanthrene and pyrene (Pagni and Dabestani, 2005; Schwarzenbach et al., 2003), the OH-PAHs showed very short half-lives against solar irradiation. Thus, the apparent photolysis can be a key process in determining the fate of OH-PAHs in sunlit surface waters. Apart from the apparent photolysis, aqueous organic pollutants may suffer from indirect photodegradation. The indirect phototransformation kinetics of xenobiotics in surface waters can be predicted as a function of surface-water parameters, such as NO− 3 , DOM, and water depth. To do this, APEX, Aqueous Photochemistry of Environmentally-occurring Xenobiotics, was recently developed and proved to be a successful software tool (Bodrato and Vione, 2014; Vione et al., 2014). Further studies are needed to select specific water bodies and calculate the corresponding t1/2,E for OH-PAHs. 3.2. Photochemical reaction types and intermediates of 9-OHFL
3. Results and discussion 3.1. Photolysis kinetics of four OH-PAHs in pure water When exposed to the simulated solar irradiation, the four OH-PAHs were susceptible to photodegradation in pure water, but no loss was found in dark controls. The OH-PAH attenuation by photolysis was confirmed to follow pseudo first-order kinetics with good linear regression (r2 N 0.95) of ln(C/C0) vs time (t). The photodegradation rate constants (k) are listed in Table 2. It was found that 9-OHPH photodegraded the fastest, followed by 1-OHPY, 2-OHFL and 9-OHFL. Their k values differed over two orders of magnitude. Such differences in the photodegradation can be attributed to the different quantum yields (Table 2) and spectral overlaps of the absorption spectra with the source emission spectrum (Fig. S1). However, these factors finally depend on the chemical structures of these compounds, including the hydroxyl substituent and benzene rings (Table S1). In surface waters, most organic pollutants absorb sunlight directly, and undergo direct photolysis and potential self-sensitized photodegradation, which are collectively called apparent photolysis (Ge et al., 2010). The environmentally relevant half-lives (t1/2,E) for the apparent photolysis in sunlit surface waters can be calculated according to Eqs. 2 and 3 (Edhlund et al., 2006; Ge et al., 2010), kE ¼ 2:303Φ∑ðZ λ ελ Þ
ð2Þ
ln 2 kE
ð3Þ
t 1=2;E ¼
where ελ are the molar absorptivities of the substrates, and Zλ is tabular solar photon flux at noon of summer and winter, assuming continuous irradiation. Here, the apparent photolysis t1/2,E is dependent on seasons and latitudes. For 45° latitude, Zλ value at every specific wavelength was taken as the average of the 40° and 50° latitude values (Leifer, 1988; OECD, 1997). Then the corresponding t1/2,E values were calculated to Table 1 Factors, levels and initial concentrations in the four-factor central composite design. Factors (units)
Fe(III) (μM) NO− 3 (μM) Cl− (M) HASSa (mg C L−1) a
Factor levels and concentrations −2
−1
0
1
2
0.00 0.00 0.00 0.00
2.00 0.20 0.125 2.00
4.00 0.40 0.250 4.00
6.00 0.60 0.375 6.00
8.00 0.80 0.500 8.00
Humic acid sodium salt (HASS).
9-OHFL was investigated in order to clarify the photochemical reaction types and transformation pathways. From the ROS scavenging experiments, it was observed that the presence of NaN3 (scavenger of •OH and 1O2) and isopropanol (scavenger of •OH) induced a notable retardation of 9-OHFL degradation to a similar extent (Fig. 1a). Thus, it can be inferred that that 9-OHFL underwent the self-sensitized photooxidation via •OH, and 1O2 might not be involved in the degradation. Furthermore, the inference was examined by the competition kinetics with acetophenone and furfural, which showed that •OH induced the oxidative degradation of 9-OHFL, whereas 1O2 did not react with the substrate (Fig. 1b). Therefore, the 9-OHFL photoreactions in pure water involved the self-sensitized photooxidation via •OH rather than 1O2. Referring to Eq. 4 (Boreen et al., 2008), the •OH oxidation contribution fraction (R•OH) was estimated to be 14.5%. ROH ≈
kPW −kPWþisopropanol kPW
ð4Þ
where kPW and kPW + isopropanol are the pseudo-first-order photolytic rate constants of 9-OHFL in pure water and in NaN3-adding water, respectively. The hydroxyl radical reaction rate constant (k·OH) was further calculated to be (6.99 ± 0.32) × 109 M−1 s−1 based on the competition kinetics. The k·OH of 9-OHFL is slightly more than that of fluorene with k·OH = 2.77 × 109 M−1 s−1 (Pagni and Dabestani, 2005). In sunlit surface waters, •OH are ubiquitous with the concentrations ranging from 10−17 M to 10−15 M (Cooper et al., 1989). As aqueous dominant photooxidants, •OH can oxidize almost all classes of organic chemicals (Mill, 1999). The highest expected •OH level shows the shortest possible half-live (t1/2) for the oxidation process. Thus, the corresponding t1/2 is calculated to be 27.5 h for •OH oxidation toward 9-OHFL in sunlit surface waters. The oxidation is slower than that of some other compounds, such as anthracene (Mill, 1999). Total ion chromatograms (TIC) for underivatized and derivatized samples of photolyzed 9-OHFL in pure water were shown in Fig. S2, suggesting that there were 8 significant intermediates generated from the photodegradation of 9-OHFL. The product with molecular weight of 180 (P180) and the parent 9-OHFL showed same retention times (tR) before derivation. After derivation, the tR of P180 did not change, indicative of no inclusion of active hydrogens in its structure. The 9-OHFL and other products, such as isomers with molecular weights of 182 and 196, showed different tR before and after derivation, which indicated their structures contained active hydrogens, allowing their reaction with silylation reagents. One isomeride named as P182 displayed same tR and mass spectra with 2-OHFL.
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Table 2 Photolytic rate constants (k), quantum yields (Φ), and corresponding environmental half-lives (t1/2,E at 45° N latitude) based on apparent photolysis Including direct and self-sensitized photodegradation. Compounds
Abbr.
k (min−1)
t1/2,E (min)a
Φ
Summer 9-Hydroxyfluorene 2-Hydroxyfluorene 9-Hydroxyphenanthrene 1-Hydroxypyrene a
9-OHFL 2-OHFL 9-OHPH 1-OHPY
−3
−2
5.50 × 10 2.04 × 10−2 1.71 × 10−1 3.20 × 10−2
6.48 × 10 5.03 × 10−2 5.01 × 10−1 3.33 × 10−2
5.5 × 10 3.2 × 10 0.4 1.8
2
Winter 7.5 × 103 2.3 × 102 1.8 7.6
Values have been multiplied by 2 to correct for the increase in rate caused by the lens effect caused by the curved test tubes used in the experiments.
Based on the mass spectra (Fig. S3) and the NIST mass spectral library, molecular weights and structures were proposed for the products. According to the photoproducts, their photolysis pathways are proposed in Fig. 2, which involve photoinduced hydroxylation, dehydrogenation and isomerization. The hydroxylation formed multiple hydroxylated fluorenones, which verified the self-sensitized photooxidation of 9-OHFL via •OH. Hydroxylated or oxidized PAHs are often identified for photolysis of PAHs and phenolic compounds in in different waters (Calza et al., 2014; Sanches et al., 2011). Referring to PAHs (Kahan and Donaldson, 2007; Sanches et al., 2011), proposed mechanisms for the OH-PAH photolysis involve either direct ionization as the initial reaction step, or interaction with reactive oxygen species, such as •OH. 3.3. Multivariate effects of water constituents on 9-OHFL photolysis When irradiated by simulated sunlight (λ N 290 nm), the model compound underwent photodegradation in different waters, with rates in seawater and freshwater being lower than that in pure water (Fig. 3). The less photodegradable potential of 9-OHFL in natural waters could be attributed to the multivariate effects of different water constituents on the photodegradation (Ge et al., 2010; Lam et al., 2003). To verify the hypothesis, the multivariate effects of main natural water constituents (e.g., DOM, Cl−, Fe(III), NO− 3 ) on the photolysis were studied using the central composite experiments. The photodegradation rate constants (k) for the experiments were listed in Table S3. The k values and the levels (x1 − x4) of the four factors − (Fe(III), NO− 3 , Cl and HASS) were evaluated by fitting a full quadratic expression: k ¼ β0 þ β1 x1 þ β2 x2 þ β3 x3 þ β4 x4 þ β12 x1 x2 þ β13 x1 x3 þ β14 x1 x4 þ β23 x2 x3 þ β24 x2 x4 þ β34 x3 x4 þ β11 x1 2 þ β22 x2 2 þ β33 x3 2 þ β44 x4 2: ð5Þ The fitted regression parameters (βx) and the corresponding significance level (p) were shown in Table 3. If p b 0.05, the corresponding βx keys are significant contributing factors. Negative βx values of Fe − were observed, indicating that these three factors (III), NO− 3 and Cl suppressed the photodegradation of 9-OHFL. In contrast, HASS and the
interaction between Fe(III) and Cl− both enhanced the photodegradation. The terms β11, β33 and β44 are positive, which may be attributed to the nonlinear effects of Fe(III), NO− 3 and HASS on the 9-OHFL photodegradation. Thus, the less photodegradable potential of 9-OHFL in natural waters than in pure water was attributed to the integrative effects of these significant contributing factors. − The photoreactive species, Fe(III), NO− 3 , Cl and HASS, usually show multiple effects. They suppress the photodegradation of one pollutant due to their competitive photoabsorption and/or ROS scavenging, while they accelerate the photodegradation because of sensitizing effect. These effects sometimes coexist. Which one effect is stronger or more apparent depends on the pollutant properties, and the ambient conditions, such as light sources and aqueous pH values (Ge et al., 2009; Niu et al., 2013; Vione et al., 2014). Fe(III) and NO− 3 have strong overlapping absorption spectra with 9-OHFL near 290 nm (Mack and Bolton, 1999; Neamtu et al., 2009; Vione et al., 2014). Therefore, they competitively absorb actinic photons (λ = 290–315 nm). Although Fe (III) and NO− 3 might sensitize the 9-OHFL photodegradation, the experimental results indicated that the inhibiting effect caused by competitive photoabsorption was more pronounced than the sensitization effect. In contrast, overall, HASS enhanced the photodegradation by sensitization, though HASS also has overlapping absorption spectra with 9OHFL (Fig. S4). The inhibitive effect of Cl− might be due to its •OH scavenging and deceleration of the 9-OHFL self-sensitized photooxidation via •OH: Cl þ OH→ClHO − −
and Cl− might transform into Cl• radicals when interacting with Fe(III), which would facilitate the photodegradation (Chiron et al., 2006; Liu et al., 2009): 2þ
FeCl
þ hν→Fe2þ þ Cl :
3.4. Photomodified toxicity of 9-OHFL to V. fischeri More toxic products might be formed during the photochemical transformation of some compounds such as PAHs and antibiotics,
Fig. 1. Effects of NaN3 and isopropanol on photodegradation kinetics of 9-OHFL (C0 = 0.5 μM), and photooxidation reactivities of 9-OHFL with •OH and 1O2
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Fig. 2. Proposed transformation pathways for photodegradation of 9-OHFL in pure water and under simulated sunlight (λ N 290 nm). The photoproducts are labeled “Pmw”, with mw standing for the molecular weights of the intermediates.
which is defined as photomodified toxicities (Ge et al., 2010; Jiao et al., 2008; Li et al., 2011; Niu et al., 2013; Schwarzenbach et al., 2006; Shemer and Linden, 2007; Wang et al., 2009b). When 9-OHFL photodegraded, the parent compound decayed and intermediates formed, probably resulting in the changing toxicity of their solutions. The toxicity evolvement of the 9-OHFL in waters during the irradiation was assayed using V. fischeri (Fig. 4). As shown in Fig. 4, the 9-OHFL solutions have a similar toxicity evolvement trend. Either in pure water or in seawater, the luminescence inhibition rate of photolyzed 9-OHFL did not change significantly (p N 0.05), though the parent compound concentration decreased. This suggested that main intermediates still showed comparable or higher toxicity compared to parent compound. Combined with the product structures (Fig. 2), 9-fluorenone is more hydrophobic than 9-OHFL, which may result in easier penetration into the cells of luminescent bacteria and subsequent notable toxicity (Li et al., 2011; Mallakin et al., 1999; McConkey et al., 1997; Niu et al., 2013). Alternatively, fluorenone is hydroxylated, leading to the enhancement of toxicity.
4. Conclusion This study demonstrated the aqueous photochemical behavior of OH-PAHs, which followed pseudo-first-order photodegradation kinetics, with apparent quantum yields and solar photodegradation halflives varying according to their chemical structures. Compared with corresponding parent PAHs, the OH-PAHs would intrinsically photodegrade fast in sunlit surface waters. 9-OHFL for example, showed to undergo both direct photolysis and self-sensitized photooxidation via •OH rather than 1O2 in pure water. Photoinduced hydroxylation, dehydrogenation and isomerization were involved in these photoreactions. 9-OHFL photodegraded slower in natural waters than in pure water, attributing to the integrative effects of the most photoreactive species, − such as Fe(III), NO− 3 , Cl and humic acid. The toxicity of 9-OHFL solutions (in pure water or seawater) did not decrease significantly (p N 0.05) after irradiation potentially due to the comparable or higher toxicity of some intermediates. The important information about photochemical behavior of OH-PAHs would better understanding the fate of these chemicals in the aqueous environment and improve their risk assessment. Table 3 Parameter estimates and hypothesis tests for the coefficients of the quadratic model fitted to the photodegradation data of 9-OHFL.
Fig. 3. Degradation kinetics of 9-OHFL (C0 = 0.5 μM) in different waters under simulated solar irradiation (λ N 290 nm).
Parameter
βx key
βx × 105
Sum of squares × 109
pa
β0 β1 β2 β3 β4 β12 β13 β14 β23 β24 β34 β11 β22 β33 β44
Intercept Fe(III) NO− 3 Cl− HASS Fe(III)–NO− 3 Fe(III)–Cl− Fe(III)–HASS − NO− 3 Cl NO− 3 HASS Cl− HASS (Fe(III))2 2 (NO− 3 ) (Cl−)2 (HASS)2
382 −16.9 −15.7 −17.2 16.1 2.92 27.3 3.54 −5.21 −10.6 −2.08 28.6 −4.13 21.7 20.0
689 591 712 623 13.6 1192 20.1 43.4 181 6.94 2240 46.8 1292 1101
0.032 0.044 0.029 0.040 0.744 0.007 0.692 0.561 0.244 0.815 0.001 0.546 0.005 0.009
a
Tests as significant at the 95% confidence level.
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Fig. 4. Bioluminescent inhibition rate (I) of photolyzed 9-Hydroxyfluorene (Concentration, C) to Vibrio fischeri in different waters.
Acknowledgments The study was supported by the National Natural Science Foundation of China (21577029, 21007013, 41476084, and 21207025), the Key Lab of Marine Bioactive Substance and Modern Analytical Technique (MBSMAT-2014-04), the Key Laboratory for Ecological Environment in Coastal Areas (201602), the Polar Science Strategic Research Foundation (20120320), and the State Oceanic Administration Specific Public Project (201105013) of China. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2015.12.143. References Barrado, A.I., Garcia, S., Barrado, E., Perez, R.M., 2012. PM2.5-bound PAHs and hydroxyPAHs in atmospheric aerosol samples: correlations with season and with physical and chemical factors. Atmos. Environ. 49, 224–232. Bastos, P.M., Eriksson, J., Bergman, Å., 2009. Photochemical decomposition of dissolved hydroxylated polybrominated diphenyl ethers under various aqueous conditions. Chemosphere 77, 791–797. Beltran, F.J., Ovejero, G., Garcia-Araya, J.F., Rivas, J., 1995. Oxidation of polynuclear aromatic hydrocarbons in water. 2. UV radiation and ozonation in the presence of UV radiation. Ind. Eng. Chem. Res. 34, 1607–1615. Bodrato, M., Vione, D., 2014. APEX (Aqueous Photochemistry of Environmentally occurring Xenobiotics): a free software tool to predict the kinetics of photochemical processes in surface waters. Environ. Sci.-Proc. Imp. 16, 732–740. Boreen, A.L., Edhlund, B.L., Cotner, J.B., McNeill, K., 2008. Indirect photodegradation of dissolved free amino acids: the contribution of singlet oxygen and the differential reactivity of DOM from various sources. Environ. Sci. Technol. 42, 5492–5498. Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., 1988. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (•OH/•O−) in aqueous solution. J. Phys. Chem. Ref. Data 17, 513–886. Calza, P., Vione, D., Minero, C., 2014. The role of humic and fulvic acids in the phototransformation of phenolic compounds in seawater. Sci. Total Environ. 493, 411–418. Chiron, S., Minero, C., Vione, D., 2006. Photodegradation processes of the antiepileptic drug carbamazepine, relevant to estuarine waters. Environ. Sci. Technol. 40, 5977–5983. Cochran, R.E., Dongari, N., Jeong, H., Beránek, J., Haddadi, S., Shipp, J., Kubátová, A., 2012. Determination of polycyclic aromatic hydrocarbons and their oxy-, nitro-, and hydroxy-oxidation products. Anal. Chim. Acta 740, 93–103. Cooper, W.J., Zika, R.G., Petasne, R.G., Fischer, A.M., 1989. In: Suffet, I.H., MacCarthy, P. (Eds.), Sunlight-induced photochemistry of humic substances in natural waters: major reactive species. American Chemical Society, Washington, DC. Dolinová, J., Ružička, R., Kurková, R., Klánová, J., Klán, P., 2006. Oxidation of aromatic and aliphatic hydrocarbons by OH radicals photochemically generated from H2O2 in ice. Environ. Sci. Technol. 40, 7668–7674. Dulln, D., Mill, T., 1982. Development and evaluation of sunlight actinometers. Environ. Sci. Technol. 16, 815–820. Edhlund, B.L., Arnold, W.A., McNeill, K., 2006. Aquatic photochemistry of nitrofuran antibiotics. Environ. Sci. Technol. 40, 5422–5427. El-Alawi, Y.S., Dixon, D.G., Greenberg, B.M., 2001. Effects of a pre-incubation period on the photoinduced toxicity of polycyclic aromatic hydrocarbons to the luminescent bacterium Vibrio fischeri. Environ. Toxicol. 16, 277–286.
171
Fasnacht, M.P., Blough, N.V., 2002. Aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 36, 4364–4369. Fasnacht, M.P., Blough, N.V., 2003. Mechanisms of the aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 37, 5767–5772. Ge, L.K., Chen, J.W., Qiao, X.L., Lin, J., Cai, X.Y., 2009. Light-source-dependent effects of main water constituents on photodegradation of phenicol antibiotics: mechanism and kinetics. Environ. Sci. Technol. 43, 3101–3107. Ge, L.K., Chen, J.W., Wei, X.X., Zhang, S.Y., Qiao, X.L., Cai, X.Y., Xie, Q., 2010. Aquatic photochemistry of fluoroquinolone antibiotics: kinetics, pathways, and multivariate effects of main water constituents. Environ. Sci. Technol. 44, 2400–2405. Grote, M., Schüürmann, G., Altenburger, R., 2005. Modeling photoinduced algal toxicity of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 39, 4141–4149. Guerard, J.J., Chin, Y.P., Mash, H., Hadad, C.M., 2009. Photochemical fate of sulfadimethoxine in aquaculture waters. Environ. Sci. Technol. 43, 8587–8592. Guo, Y., Senthilkumar, K., Alomirah, H., Moon, H.B., Minh, T.B., Mohd, M.A., Nakata, H., Kannan, K., 2013. Concentrations and profiles of urinary polycyclic aromatic hydrocarbon metabolites (OH-PAHs) in several Asian countries. Environ. Sci. Technol. 47, 2932–2938. ISO, 2007. Water Quality—Determination of the Inhibitory Effect of Water Samples on the Light Emission of Vibrio fischeri (Luminescent Bacteria Test)—Part 3: Method Using Freeze-dried Bacteria (ISO 11348-3). The International Organization for Standardization. Itoh, N., Tao, H., Ibusuki, T., 2006. In-tube silylation in combination with thermal desorption gas chromatography–mass spectrometry for the determination of hydroxy polycyclic aromatic hydrocarbons in water. Anal. Chim. Acta 555, 201–209. Jiao, S.J., Zheng, S.R., Yin, D.Q., Wang, L.H., Chen, L.Y., 2008. Aqueous photolysis of tetracycline and toxicity of photolytic products to Luminescent bacteria. Chemosphere 73, 377–382. Johnson-Restrepo, B., Olivero-Verbel, J., Lu, S.J., Guette-Fernández, J., Baldiris-Avila, R., O'Byrne-Hoyos, I., Aldousa, K.M., Addink, R., Kannan, K., 2008. Polycyclic aromatic hydrocarbons and their hydroxylated metabolites in fish bile and sediments from coastal waters of Colombia. Environ. Pollut. 151, 452–459. Jung, J., Kim, Y., Kim, J., Jeong, D.H., Choi, K., 2008. Environmental levels of ultraviolet light potentiate the toxicity of sulfonamide antibiotics in Daphnia magna. Ecotoxicology 17, 37–45. Kahan, T.F., Donaldson, D.J., 2007. Photolysis of polycyclic aromatic hydrocarbons on water and ice surfaces. J. Phys. Chem. A 111, 1277–1285. Lam, M.W., Tantuco, K., Mabury, S.A., 2003. PhotoFate: a new approach in accounting for the contribution of indirect photolysis of pesticides and pharmaceuticals in surface waters. Environ. Sci. Technol. 37, 899–907. Lampi, M.A., Gurska, J., McDonald, K.I., Xie, F., Huang, X.D., Dixon, D.G., Greenberg, B.M., 2006. Photoinduced toxicity of polycyclic aromatic hydrocarbons to Daphnia magna: ultraviolet-mediated effects and the toxicity of polycyclic aromatic hydrocarbon photoproducts. Environ. Toxicol. Chem. 25, 1079–1087. Latch, D.E., Packer, J.L., Stender, B.L., VanOverbeke, J., Arnold, W.A., McNeill, K., 2005. Aqueous photochemistry of triclosan: formation of 2,4-dichlorophenol, 2,8-dichlorodibenzo-pdioxin, and oligomerization products. Environ. Toxicol. Chem. 24, 517–525. Leifer, A., 1988. The Kinetics of Environmental Aquatic Photochemistry: Theory and Practice. American Chemical Society, Washington, DC. Li, Y., Niu, J.F., Wang, W.L., 2011. Photolysis of enrofloxacin in aqueous systems under simulated sunlight irradiation: kinetics, mechanism and toxicity of photolysis products. Chemosphere 85, 892–897. Liu, H., Zhao, H., Quan, X., Zhang, Y., Chen, S., 2009. Formation of chlorinated intermediate from bisphenol A in surface saline water under simulated solar light irradiation. Environ. Sci. Technol. 43, 7712–7717. Luan, T.G., Yu, K.S.H., Zhong, Y., Zhou, H.W., Lan, C.Y., Tam, N.F.Y., 2006. Study of metabolites from the degradation of polycyclic aromatic hydrocarbons (PAHs) by bacterial consortium enriched from mangrove sediments. Chemosphere 65, 2289–2296. Lutz, S., Feidt, C., Monteau, F., Rychen, G., Le Bizec, B., Jurjanz, S., 2006. Effect of exposure to soil-bound polycyclic aromatic hydrocarbons on milk contaminations of parent compounds and their monohydroxylated metabolites. J. Agric. Food Chem. 54, 263–268. Mack, J., Bolton, J.R., 1999. Photochemistry of nitrite and nitrate in aqueous solution: a review. J. Photochem. Photobiol. A Chem. 128, 1–13. Mallakin, A., McConkey, B.J., Miao, G., McKibben, B., Snieckus, V., Dixon, D.G., Greenberg, B.M., 1999. Impacts of structural photomodification on the toxicity of environmental contaminants: anthracene photooxidation products. Ecotoxicol. Environ. Saf. 43, 204–212. McConkey, B.J., Duxbury, C.L., Dixon, D.G., Greenberg, B.M., 1997. Toxicity of a PAH photooxidation product to the bacteria Photobacterium phosphoreum and the duckweed Lemna gibba: effects of phenanthrene and its primary photoproduct, phenanthrenequinone. Environ. Toxicol. Chem. 16, 892–899. Mill, T., 1999. Predicting photoreaction rates in surface waters. Chemosphere 38, 1379–1390. Murayama, M., Dasgupta, P.K., 1996. Liquid chromatographic determination of nitrosubstituted oolynuclear aromatic hydrocarbons by sequential electrochemical and fluorescence detection. Anal. Chem. 68, 1226–1232. Neamtu, M., Popa, D.M., Frimmel, F.H., 2009. Simulated solar UV-irradiation of endocrine disrupting chemical octylphenol. J. Hazard. Mater. 164, 1561–1567. Niu, J.F., Li, Y., Wang, W.L., 2013. Light-source-dependent role of nitrate and humic acid in tetracycline photolysis: kinetics and mechanism. Chemosphere 92, 1423–1429. OECD, 1997. Guidance Document on Direct Phototransformation of Chemicals in Water. OECD Environmental Health and Safety Publication. Series on Testing and Assessment No.7, Paris, France. Pagni, R.M., Dabestani, R., 2005. Recent developments in the environmental photochemistry of PAHs and PCBs in water and on solids. Environmental Photochemistry Part II. Springer Berlin Heidelberg, pp. 193–219.
172
L. Ge et al. / Science of the Total Environment 547 (2016) 166–172
Petersen, D.G., Reichenberg, F., Dahllöf, I., 2008. Phototoxicity of pyrene affects benthic algae and bacteria from the Arctic. Environ. Sci. Technol. 42, 1371–1376. Pojana, G., Marcomini, A., 2007. Determination of monohydroxylated metabolites of polycyclic aromatic hydrocarbons (OH-PAHs) from wastewater-treatment plants. Int. J. Environ. An. Ch. 87, 627–636. Rey-Salgueiro, L., Martínez-Carballo, E., García-Falcón, M.S., González-Barreiro, C., SimalGándara, J., 2009. Occurrence of polycyclic aromatic hydrocarbons and their hydroxylated metabolites in infant foods. Food Chem. 115, 814–819. Sanches, S., Leitão, C., Penetra, A., Cardoso, V.V., Ferreira, E., Benoliel, M.J., Barreto Crespo, M.T., Pereira, V.J., 2011. Direct photolysis of polycyclic aromatic hydrocarbons in drinking water sources. J. Hazard. Mater. 192, 1458–1465. Sankoda, K., Kuribayashi, T., Nomiyama, K., Shinohara, R., 2013. Occurrence and source of chlorinated polycyclic aromatic hydrocarbons (Cl-PAHs) in tidal flats of the Ariake Bay, Japan. Environ. Sci. Technol. 47, 7037–7044. Schwarzenbach, R.P., Gschwend, P.M., Imboden, D.M., 2003. Environmental Organic Chemistry. second ed. John Wiley & Sons, Inc., New Jersey. Schwarzenbach, R.P., Escher, B.I., Fenner, K., Hofstetter, T.B., Johnson, C.A., von Gunten, U., Wehrli, B., 2006. The challenge of micropollutants in aquatic systems. Science 313, 1072–1077. Sepic, E., Bricelj, M., Leskovsek, H., 2003. Toxicity of fluoranthene and its biodegradation metabolites to aquatic organisms. Chemosphere 52, 1125–1133. Shemer, H., Linden, K.G., 2007. Aqueous photodegradation and toxicity of the polycyclic aromatic hydrocarbons fluorene, dibenzofuran, and dibenzothiophene. Water Res. 41, 853–861.
Shen, G.F., Wang, G.H., 2012. Molecular composition and seasonal variation of hydroxylated polycyclic aromatic hydrocarbons in atmospheric PM2.5 of Nanjing, China. J. Earth Environ. 3, 1066–1069. Vione, D., Minella, M., Maurino, V., Minero, C., 2014. Indirect photochemistry in sunlit surface waters: photoinduced production of reactive transient species. Chem. Eur. J. 20, 10590–10606. Wang, Y., Chen, J.W., Ge, L.K., Wang, D.G., Cai, X.Y., Huang, L.P., Hao, C., 2009a. Experimental and theoretical studies on the photoinduced acute toxicity of a series of anthraquinone derivatives towards the water flea (Daphnia magna). Dyes Pigments 83, 276–280. Wang, Y., Chen, J.W., Lin, J., Wang, Z., Bian, H.T., Cai, X.Y., Hao, C., 2009b. Combined experimental and theoretical study on photoinduced toxicity of an anthraquinone dye intermediate to Daphnia magna. Environ. Toxicol. Chem. 28, 846–852. Xia, X., Li, G., Yang, Z., Chen, Y., Huang, G.H., 2009. Effects of fulvic acid concentration and origin on photodegradation of polycyclic aromatic hydrocarbons in aqueous solution: importance of active oxygen. Environ. Pollut. 157, 1352–1359. Xie, Q., Chen, J.W., Zhao, H.X., Qiao, X.L., Cai, X.Y., Li, X.H., 2013. Different photolysis kinetics and photooxidation reactivities of neutral and anionic hydroxylated polybrominated diphenyl ethers. Chemosphere 90, 188–194. Zhang, Y., Yang, B., Gan, J., Liu, C.G., Shu, X., Shu, J.N., 2011. Nitration of particle-associated PAHs and their derivatives (nitro-, oxy-, and hydroxy-PAHs) with NO3 radicals. Atmos. Environ. 45, 2515–2521.