Chemosphere 93 (2013) 2507–2513

Contents lists available at ScienceDirect

Chemosphere journal homepage: www.elsevier.com/locate/chemosphere

Effects of four commonly used UV filters on the growth, cell viability and oxidative stress responses of the Tetrahymena thermophila Li Gao a,b, Tao Yuan a,⇑, Chuanqi Zhou a, Peng Cheng c, Qifeng Bai a, Junjie Ao a, Wenhua Wang a, Haimou Zhang c a b c

School of Environmental Science and Engineering, Shanghai Jiao Tong University, Shanghai 200240, China School of Resource and Environment, Ningxia University, Yinchuan 750021, China School of Life Science, Hubei University, Wuhan 430062, China

h i g h l i g h t s  Two selected UV filters can constrain the growth of T. thermophila.  Increasing of exposure time will exacerbate the impairment of cell viability.  1.0 lg L

1

of BP-3 and 4-MBC can induce oxidative injuries of T. thermophila.

a r t i c l e

i n f o

Article history: Received 15 April 2013 Received in revised form 6 September 2013 Accepted 10 September 2013 Available online 13 October 2013 Keywords: Benzophenone-3 4-Methyl-benzylidene camphor Growth inhibition Cell viability Oxidative stress Environmentally relevant concentration

a b s t r a c t UV filters are increasingly used in sunscreens and other personal care products. Although their residues have been widely identified in aquatic environment, little is known about the influences of UV filters to protozoan. The growth inhibition effects, cell viability and oxidative stress responses of four commonly used UV filters, 2-ethylhexyl 4-methoxycinnamate (EHMC), benzophenone-3 (BP-3), 4-methyl-benzylidene camphor (4-MBC) and octocrylene (OC), to protozoan Tetrahymena thermophila were investigated in this study. The 24-h EC50 values with 95% confidence intervals for BP-3 and 4-MBC were 7.544 (6.561–8.675) mg L1 and 5.125 (4.874–5.388) mg L1, respectively. EHMC and OC did not inhibit the growth of T. thermophila after 24 h exposure at the tested concentrations. The results of cell viability assays with propidium iodide (PI) staining were consistent with that of the growth inhibition tests. As for BP-3 and 4-MBC, the relatively higher concentrations, i.e. of 10.0 and 15.0 mg L1, could lead to the cell membranes impairment after 4 h exposure. With the increase of the exposure time to 6 h, their adverse effects on cell viability of T. thermophila were observed at the relatively lower concentration groups (1.0 mg L1 and 5.0 mg L1). In addition, it is noticeable that at environmentally relevant concentration (1.0 lg L1), BP-3 and 4-MBC could lead to the significant increase of catalase (CAT) activities of the T. thermophila cells. Especially for the BP-3, the oxidative injuries were further confirmed by the reduction of glutathione (GSH) content. It is imperative to further investigate the additive action of UV filters and seek other sensitive endpoint, especially at environmentally relevant concentration. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Ultraviolet (UV) light in sunlight is widely accepted as the underlying cause of overexposure to sun radiation for adverse health effects on skin, eyes and immune system. Sunscreen cosmetic products have been used for nearly 75 years and comprise substances (commonly referred as ultraviolet (UV) filters) able to absorb UV radiation and protect human skin from direct exposure to the deleterious wavelengths of sunlight (Giokas et al., 2007). UV filters are being increasingly used in personal care products and ⇑ Corresponding author. Tel.: +86 21 54742823. E-mail address: [email protected] (T. Yuan). 0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.09.041

other materials in order to provide protection against UV-A (320–400 nm) and UV-B (280–320 nm) radiation. More specifically, UV filters absorb photons and rapidly return to ground state by thermally emitting the energy through a series of vibrational transitions (vibrational relaxation). To date, numerous commercial formulations with varying compositions are marketed and afford protection against both types of sunlight radiation (UV-A and UV-B). In the European Union regions, 26 organic compounds are permitted for use as sunscreen agents (Diaz-Cruz et al., 2008). Chemical organic filters absorb mostly UV-B radiation (e.g., PABA derivatives, salicylates, cinnamates and camphor derivatives). There are some UV filters can partly absorb the range of UV-A (e.g., benzophenones, anthranilates and dibenzoylmethanes).

2508

L. Gao et al. / Chemosphere 93 (2013) 2507–2513

These compounds have been released to the aquatic environment either directly via wash-off from skin and cloth, or indirectly via wastewater. It is estimated that 10 000 tons of UV filters are produced annually for the global market (Danovaro et al., 2008). Unfortunately, they could not be removed completely in conventional waste water treatment plants (WWTPs) (Liu et al., 2012). Residues of UV filters have been found in wastewater (Li et al., 2007; Rodil et al., 2012), sludge (Gago-Ferrero et al., 2011; Zhang et al., 2011), swimming-pool water (Giokas et al., 2004; Cuderman and Heath, 2007), surface waters (Poiger et al., 2004; Tarazona et al., 2010), sediments (Schlenk et al., 2005; Kameda et al., 2011; Amine et al., 2012), fish (Buser et al., 2006; Mottaleb et al., 2009; Fent et al., 2010b), soil (Sanchez-Brunete et al., 2011), and even drinking water (Liu et al., 2011; Diaz-Cruz et al., 2012). The occurrences of four important organic UV filter compounds (4-methylbenzylidene camphor, 4-MBC; ethylhexyl methoxy cinnamate, EHMC; octocrylene, OC; benzophenone-3, BP-3) in wastewater in Switzerland had been investigated (Balmer et al., 2005). In sunscreens and suncare products, 4-MBC and EHMC are widely used for UV-B protection. OC and BP-3 play the role of both UV-B and UV-A protection. The four commonly used UV filters were identified in WWTP influents and effluents, with the maximum concentrations of 19 lg L1 for EHMC and 2.7 lg L1 for 4-MBC, respectively. It was reported that the maximum concentration of BP-3 was 3.3 lg L1 in the surface water samples, which were collected from the beaches located in the Mediterranean coast of Spain during the summer season (Tarazona et al., 2010). The maximum concentrations of UV filters in lake were reported by Moeder et al. (2010) and Rodil et al. (2009). For example, OC, 4MBC and EHMC were determined up to 4.3 lg L1, 2.6 lg L1 and 3.0 lg L1 in the Cospuden lake, respectively. UV filters were identified in river water, with the concentration range of several ng L1 to lg L1 (Negreira et al., 2010). Because of their high lipophilicity (mostly with Log KOW 4–8), UV filters could be accumulated in sediments and biota. These substances, namely BP-3, OC, EHMC and 4-MBC had all been found in sediments samples, with maximum concentrations of 642 lg kg1 (OC) in lake sediments in Frankfurt, Germany (Kaiser et al., 2012b). In addition, Bachelot et al. (2012) investigated UV filters accumulation in marine mussels from French coastal regions. The maximum concentrations of EHMC and OC in mussel tissues were 256 and 7 112 lg kg1, respectively. Furthermore, UV filters could have been bioaccumulated in the food chain. Fent et al. (2010b) investigated a number of UV filters in aquatic ecosystems in Switzerland. The maximum concentrations of lipidweighted EHMC were 337 lg kg1 in 48 macroinvertebrate and fish samples collected from six rivers, and 701 lg kg1 in 5 cormorants. The results suggested the potential food-chain accumulation of UV filters. Since the in vivo and in vitro endocrine disruption effects of UV filters were reported, they had been considered as a new class of estrogenic (Schlumpf et al., 2001) or antiandrogenic (Ma et al., 2003) chemicals. As for aquatic ecotoxicity studies, the published researches mainly focus on the endocrine disruption effects of UV filters to fish and invertebrates. For example, it was reported that benzophenone-1 (BP-1), benzophenone-2 (BP-2), 3-benzylidene camphor (3-BC) and ethyl-4-aminobenzoate (Et-PABA) could lead to the induction of vitellogenin (Holbech et al., 2002; Weisbrod et al., 2007). Similarily, vitellogenin plasma concentration in male fathead minnows (Pimephales promelas) was significantly increased at 244.5 lg L1 of EHMC exposure. In addition, EHMC induced significant histological changes in testes and ovaries at 394 lg L1 (Christen et al., 2011). On the other hand, Zucchi et al. (2011) investigated the interfering effects of BP-4 on the sex hormone system of fish. When exposed up to 3 d after hatching for 14 d, BP-4 at

3000 lg L1 induced a low estrogenic activity and interference with early thyroid development in eleuthero-embryos. In adult males BP-4 displayed multiple effects on gene expression in different tissues. Furthermore, the expression of genes involved in hormonal pathways in fathead minnows showed that EHMC displayed low but multiple hormonal activities including estrogenic (down-regulation of 3b-HSD), antiestrogenic (down-regulation of esr1), and antiandrogenic activity (down-regulation of ar in the liver of females) (Christen et al., 2011). On the other hand, Schmitt et al. (2008) investigated the ecotoxicity effects of UV filters on two benthic invertebrates, Potamopyrgus antipodarum and Lumbriculus variegatus. They observed an increased unshelled embryo production when P. antipodarum was exposed to either 3-BC or 4-MBC in the low concentrations (0.28 and 6.33 mg kg1 sediment dry weight for 3-BC, 1.71 and 7.65 mg kg1 sediment dw for 4-MBC) for 56 d. However, the mortality was increased significantly when P. antipodarum was exposed at high concentrations (31.6 and 32.9 mg kg1 sediment dw for 3-BC and 4-MBC, respectively). In addition, it was noted that the reproduction of L. variegatus could be inhibited by 3-BC and 4MBC after 28 d exposure. At the concentration of 6.47 and 29.1 mg kg1 sediment dw for 3-BC and 6.18 mg kg1 sediment dw for 4-MBC, the reproduction of L. variegatus was significantly lower compared to the solvent control. Moreover, EHMC was reported to cause a toxic effect on reproduction in snails, i.e. P. antipodarum and Melanoides tuberculata, with the lowest observed effect concentrations (LOEC) of 0.4 mg kg1 and 10 mg kg1, respectively (Kaiser et al., 2012a). Besides the potential endocrine disruption effects, a few other aquatic ecotoxicology studies were reported in the recent years. For example, Kim et al. (2011) tested the acute toxicity of the selected benzotriazole UV filters on a freshwater crustacean Daphnia pulex. The 24-h and 48-h LC50 values of UV-571 for D. pulex were estimated to be 6.35 and 2.59 mg L1, respectively. Danovaro et al. (2008) found that UV filters EHMC, BP-3 and 4-MBC were able to induce the lytic viral cycle in symbiotic zooxanthellae with latent infections and caused complete bleaching of Acropora spp. even at very low concentration. Protozoa are the simplest unicellular eukaryotic organisms. They present in both aquatic and terrestrial ecosystems that constitute a next between primary producers and pluricellular consumers (Gallego et al., 2007). Tetrahymena is a ubiquitous freshwater ciliated protozoan that has been extensively investigated and is increasingly being used in toxicology studies (Mortimer et al., 2010; Stefanidou et al., 2011; Lang and Kohidai, 2012). For example, Zhang et al. (2012) selected Tetrahymena as the model organism and applied microcalorimetry to determine the acute toxicity of chlorobenzenes. In addition, it also represents an ecotoxicological tool for water quality assessment with an aquatoxic potency correlating with that observed for fish (Sinks and Schultz, 2001; Seward et al., 2002). Unfortunately, to the best of our knowledge, little is known about the influences of UV filters on the freshwater protozoan, e.g. Tetrahymena thermophila, although the occurrences of UV filters in freshwater have been frequently reported. Moreover, the reports are relatively scarce about the endpoint investigations of the aquatic organisms exposed to the UV filters at environmental relevant concentration, i.e. ng L1 to lg L1 level. In this study, we investigated the growth inhibition effects of four common used UV filters on the T. thermophila. In addition, a DNA intercalating dye (propidium iodide, PI) was applied to detect the cell viability of T. thermophila exposed to the environmental relevant concentration of UV filters. Meanwhile, the antioxidant enzymes activities were tested to investigated the oxidative stress. This study attempts to improve the understandings of the ecological risk of UV filters in aquatic environment.

2509

L. Gao et al. / Chemosphere 93 (2013) 2507–2513

2. Materials and methods

2.4. Cell viability assays

2.1. Chemicals

The cell viability of T. thermophila was measured as described in Mortimer et al. (2010) with small modification for the use of propidium iodide (PI, Sigma–Aldrich). Briefly, after 2 h, 4 h or 6 h exposure to the UV filters, 180 lL per well of cell suspension was transferred to 96-well black microplate. 20 lL of PI solution was added, with the final concentration of 10 lg mL1. Cells were incubated at 30 °C in the dark for 15 min. Fluorescence was measured with a microplate reader (Varioskan Flash, Thermo Fisher Scientific, USA) at excitation and emission wavelengths of 530 and 620 nm, respectively.

UV filters were obtained as follows: EHMC (96%) was purchased from TCI, Japan. BP-3 (98%) and OC (97%) were products of Sigma– Aldrich, USA. 4MBC (>99%) was obtained from Alfa, USA. Their structures and relevant physic-chemical properties are given in Table 1. The stock and work solutions of UV filters were prepared in DMSO and stored in the dark at 4 °C. 2.2. Species and culture medium T. thermophila (SB 210) was provided by the Institute of Hydrobiology, Chinese Academy of Sciences, Wuhan, PR China. The cells were cultured at 30 °C in a liquid medium containing 1% (w/v) tryptone (Oxoid), 0.1% yeast extract (Oxoid) and 0.2% glucose. Culture medium and consumables were sterilized in high-pressure steam at 120 °C for 20 min before use. 2.3. Growth inhibition test To prepare the cultures for toxicity testing, 5 mL of the 24 h precultures were transferred to 45 mL of sterile medium in a 250 mL Erlenmeyer flask. The initial population density was higher than 5  103 cells mL1. The individual UV filters were added to cell suspension with final concentrations ranging from 0 (control) to 15 mg L1, respectively. Each concentration was conducted in triplicate. The final DMSO-concentration was restricted lower than 0.1%. Then T. thermophila cells were exposed for 24 h (30 °C), respectively. The density of the cell population was measured spectrophotometrically at 492 nm using a microplate reader (Varioskan Flash, Thermo Fisher Scientific, USA). The optical density (OD) value is directly proportional to the number of T. thermophila cells. Percent of cell growth inhibition was calculated as:

Percent of growth inhibition ð%Þ ¼ ðODC  ODT Þ=ðODC  ODB Þ  100 wherein, ODC and ODT are the measured values of the control groups and the tested groups, respectively. ODB is the value of blank medium without cells.

2.5. Cellular antioxidant enzymes and glutathione analyses Protozoan culture was exposed to the UV filters at two concentrations: 1 lg L1 and 1 mg L1, which are in the same order of magnitude with the maximum environmentally relevant concentration and 24-h EC50 values of T. thermophila for BP-3 and 4MBC, respectively. The T. thermophila cells (about 1  106 /mL) were exposed to UV filters for 24 h. Afterwards, cells were harvested by centrifugation (300 g, 5 min) and washed twice with cold phosphate buffer solution (PBS, pH 7.4) at room temperature. The cells were resuspended in PBS and homogenized using freeze–thaw process three times. The homogenate was centrifuged at 10 400 rpm for 20 min. The supernatant was used as the cell extract. The activities of cellular enzymes, including superoxide dismutase (SOD), catalase (CAT) and glutathione peroxidase (GPx), total protein concentrations and the contents of glutathione (GSH) were measured respectively by the assay kits provided by Nanjing Jiancheng Biotechnique Institute (Nanjing, China). The crude cell extracts were standardized per milligram of total protein. 2.6. Statistical analysis Experiments were carried out at least in triplicate. All values were expressed as the mean ± standard deviation (SD). For the purpose of EC50 values calculation the data were fitted to the classical sigmoidal four parameter dose–response model (log(inhibitor) vs. response – Variable slope):

y ¼ b þ ða  bÞ=ð1 þ 10^ ðLogEC 50  xÞhÞ

Table 1 Structures and some physic-chemical properties of the selected UV filters. Molecular weight

CAS number

Log Kow*

Solubility (mg L1)

2-Ethylhexyl 4-methoxycinnamate (EHMC)

290.41

5466-77-3

5.80

0.1548

Benzophenone-3 (BP-3)

228.25

131-57-7

3.79

68.56

4-Methyl-benzylidene camphor (4-MBC)

254.38

36861-47-9

5.92

0.1966

Octocrylene (OC)

361.49

6197-30-4

6.88

0.003808

Compounds

*

Chemical structure

Log Kow, the logarithmised octanol – water partition coefficient, calculated from EPISUITE 4.1.

2510

L. Gao et al. / Chemosphere 93 (2013) 2507–2513

wherein y is the response, b represents minimum of the response, a represents maximum of the response, h is shape parameter, x is logarithm of inhibitor concentration. EC50 values were calculated with their 95% confidence interval from the data obtained using the software package GraphPad PrismÒ (version 5.0 Demo, GraphPad Software, USA). Data from cell viability assay and cellular enzyme activity test were analyzed using SPSS statistical package version 17.0. Differences in cell viabilities and enzyme activities were made by analysis of variance (ANOVA). In addition, post hoc comparison (LSD test) was carried out to check for differences between the exposed and control groups. A value of p < 0.05 was considered statistically significant. 3. Results and discussion 3.1. Effects of UV filters on the growth inhibition of T. thermophila T. thermophila cells were exposed to UV filters with concentrations of 0 (control), 0.001 mg L1, 0.01 mg L1, 0.1 mg L1, 1.0 mg L1, 5.0 mg L1, 10.0 mg L1 and 15.0 mg L1, respectively. Despite the use of DMSO (0.1%), some UV filters would precipitate in the culture medium when the final concentration was more than 15.0 mg L1 in the preliminary experiments. Therefore, in this test, the maximum final concentration of UV filters was designed as 15.0 mg L1. In the tested nominal concentration range, it was not observed the significant growth inhibition effects of EHMC and OC on T. thermophila after 24 h exposure. Dose–response relationships of T. thermophila to BP-3 and 4-MBC were shown in Fig. 1. The maximum of percent of growth inhibition for BP-3 and 4-MBC was 96.7% and 98.6%, respectively. The 24-h EC50 values (concentrations causing a 50% decrease in viability) with 95% confidence intervals for BP-3 and 4-MBC in this study were 7.544 (6.561– 8.675) mg L1 and 5.125 (4.874–5.388) mg L1, respectively (Table 2). 120

percent of growth inhibition (%)

(a)

100 80 60 40 20 0 -20 1E-3

0.01

0.1

1

10

concentration of BP-3 (mg/L) 120

percent of growth inhibition (%)

(b)

100 80 60

EHMC, BP-3, 4-MBC and OC are the most commonly used UV filters in cosmetics. They are usually co-applied to increase UV protection effects. As mentioned above, it is quite few of the information on growth inhibition effects of UV filters on aquatic organisms. Recently, Sieratowicz et al. (2011) investigated the growth inhibition effects of BP-3 and 4-MBC on alga (Desmodesmus subspicatus) after 72 h exposure. It was found that the 72-h EC50 of BP-3 and 4-MBC on alga are 0.96 and 7.66 mg L1, respectively. In this study, the 24-h EC50 of BP-3 and 4-MBC are 7.5 and 5.1 mg L1, respectively (Table 2). The results suggested T. thermophila to be more sensitive bioindicator for 4-MBC, whereas the alga is for BP-3. In addition, Fent et al. (2010a) performed acute immobilization tests of EHMC, 4-MBC and BP-3 with Daphnia magna and obtained 48-h EC50 values of 0.29, 0.56 and 1.9 mg L1, respectively. Sieratowicz et al. (2011) also got the similar data (0.57, 0.80 and 1.67 mg L1 for EHMC, 4-MBC and BP-3, respectively). Furthermore, it was implied that the immobilization effects of EHMC, 4MBC, BP-3 and BP-4 to D. magna increased with Log KOW of the compound (Fent et al., 2010a). However, of the four UV filters in this study, 4-MBC appeared to be most acutely toxic to T. thermophila, even if OC (Log KOW 6.88) is more lipophilic than 4-MBC (Log KOW 5.92).

3.2. Cell viabilities Propidium iodide (PI) is a DNA intercalating dye, which is generally excluded by viable cells. Thus the fluorescence power indicates an impairment of plasma membrane. In this study, PI staining was used to evaluate the toxic effects of four UV filters on the cell viabilities of T. thermophila at three exposure time (2 h, 4 h and 6 h). The results of the PI fluorescence assay were expressed as relative fluorescence unit (RFU) values, i.e. a percentage compared to non-exposed controls. Therefore, with the increase of RFU values, the cell viabilities decreased. Compared with those in the control groups, fluorescence values measured in all tested groups were not changed significantly after 2 h exposure. However, with the increase of exposure time to 4 h, the RFU values increased significantly in the higher concentration groups of BP-3 and 4-MBC (i.e. 10.0 and 15.0 mg L1) (Fig. 2a and b). Similarly, Mortimer et al. (2010) applied PI assay for the evaluation of cell viability when T. thermophila exposed to nanoparticles of ZnO and CuO. It was found that both nanoparticale and bulk ZnO could impair plasma membrane of T. thermophila after 4 h exposure. They suggested that 4 h PI assay of T. thermophila could be used for the toxicity screening of nanoparticles (Mortimer et al., 2010). Our results support that it could be also applied for the high throughput testing of UV filters or other organic pollutants. In addition, when exposure time was increased to 6 h, the cytotoxicity effects of BP-3 and 4-MBC were observed in the lower concentration groups of 1.0 mg L1 (4-MBC, p < 0.01) and 5.0 mg L1 (BP-3, p < 0.05; 4-MBC, p < 0.01) (Fig. 2a and b). The results suggested that the cell viability of T. thermophila might be affected by long term exposure of the BP-3 and 4-MBC at the environmental occurrence concentration level. In this study, it was not found that

40 Table 2 24-h EC50 of the tested UV filters to T. thermophila.

20 0 -20 1E-3

0.01

0.1

1

10

concentration of 4-MBC (mg/L) Fig. 1. Dose–response curve of T. thermophila exposed to BP-3 (a) and 4-MBC (b) for 24 h. Data are expressed as mean ± SD (n = 3).

Compounds

24-h EC50 (mg L1)

95% Confidence interval (mg L1)

EHMC BP-3 4-MBC OC

>15 7.544 5.125 >15

n.c. 6.561–8.675 4.874–5.388 n.c.

n.c. = Not calculable.

L. Gao et al. / Chemosphere 93 (2013) 2507–2513

2511

as an antioxidant to prevent damage to important cellular components caused by ROS such as free radicals and peroxides (Couto et al., 2013). The results of antioxidant enzyme activities and contents of GSH were shown in Fig. 3. It was found that the CAT activities in the T. thermophila exposed to 1.0 lg L1 of BP-3 and 4-MBC, respectively, were significantly increased (p < 0.05) (Fig. 3a). However, no significant changes of the enzyme activities of SOD and GPx were observed in other experimental treatments (Fig. 3b and c). As to the content of GSH, a significant decrease was found only in the 1.0 lg L1 of BP-3 exposure group (p < 0.05) (Fig. 3d).

Fig. 2. Cell viability of T. thermophila exposed to BP-3 (a) and 4-MBC (b) after 2 h, 4 h and 6 h, respectively. Results were expressed as RFU values of PI. The data points represent the means of eight culture wells with standard deviations. Significant difference from the control are indicated as ⁄p < 0.05, ⁄⁄p < 0.01 or ⁄⁄⁄p < 0.001.

cell viabilities of T. thermophila were significantly affected by EHMC and OC at the tested concentrations and exposure time. In general, the results of cell viability assays were consistent with that of the growth inhibition tests. It was noted that EHMC and OC at the tested concentrations did not show any obvious cytotoxicity to T. thermophila cells, and accordingly did not affect their growth. As for BP-3 and 4-MBC, the relatively higher concentrations, i.e. of 10.0 and 15.0 mg L1, could lead to the cell membranes impairment after 4 h exposure. As mentioned above, with the increase of the exposure time to 6 h, their adverse effects on cell viability of T. thermophila were observed at the relatively lower concentration groups (1.0 mg L1 and 5.0 mg L1). The cytotoxicity results of BP-3 and 4-MBC were helpful to explain their effects on growth inhibitions on T. thermophila (Fig. 1). 3.3. Activities of antioxidant enzymes and contents of GSH Reactive oxygen species (ROS) can be generated in living organisms exposed to environmental contaminants. Changes in antioxidant enzyme activities can reflect the oxidative stress caused by ROS. The antioxidant enzymes are the first line of defense against ROS. SOD converts superoxide radical (O 2 ) to peroxide (H2O2), and of CAT and GPx to further detoxify H2O2 and organic hydroperoxides. In addition, there are some other low molecular weight scavengers such as GSH. GSH is a tripeptide with a gamma peptide linkage between the amine group of cysteineand the carboxyl group of the glutamate side-chain. GSH exists in both reduced (GSH) and oxidized (glutathione disulfide, GSSG) states. In the reduced state, the thiol group of cysteine is able to donate a reducing equivalent to other unstable molecules, such as ROS. In donating an electron, glutathione itself becomes reactive, but readily reacts with another reactive glutathione to form GSSG. GSH can be regenerated from GSSG by the enzyme glutathione reductase. In such a process, GSH can act

Fig. 3. Anti-oxidative stress of T. thermophila after exposure to UV filters at the concentration of 1.0 lg L1 and 1.0 mg L1 respectively. (a) CAT activities, (b) SOD activities, (c) GPx activities and (d) contents of GSH. Results are expressed as mean ± SD (n = 3). Significant difference from the control are indicated as ⁄p < 0.05.

2512

L. Gao et al. / Chemosphere 93 (2013) 2507–2513

The antioxidant enzymes could be induced by a slight oxidative stress due to compensatory response; however, a severe oxidative stress suppresses the activities of these enzymes due to oxidative damage and a loss in compensatory mechanisms. CAT is mainly located in the peroxisomes. It is responsible for the reduction of hydrogen peroxide produced from the metabolism of long chain fatty acids in peroxisomes (Yi et al., 2007). The increase of CAT activity implied that the tested chemicals might stimulate the production of H2O2 in organisms (Barata et al., 2005). In this study, the results suggested that 1.0 lg L1 of BP-3 and 4-MBC respectively could exacerbate the oxidative stress of T. thermophila in terms of H2O2 production. Especially for the BP-3, the oxidative injuries were further confirmed by the reduction of GSH content in T. thermophila (Fig. 3d). The findings are even more worrisome since it was reported that the maximum concentrations of BP-3, 4-MBC and other organic UV filters in water environment are more than 1.0 lg L1 (Moeder et al., 2010; Tarazona et al., 2010). To date, the study of UV filters effects on aquatic organisms at environmentally relevant concentration is rare. Kunz et al. (2004) reported that the exposure of Xenopus laevis to 4-MBC and 3-BC respectively at environmentally relevant concentrations did not negatively affect the thyroid system and sex ratio of frogs. However, the results in this study indicated that the environmentally relevant concentration of UV filters can induce some adverse effects on T. thermophila, at least in terms of oxidative stress situation. 4. Conclusions There is currently a lack of information on the growth inhibition effects of UV filters to aquatic organisms. This study showed that in the tested nominal concentration range, EHMC and OC did not render significant growth inhibition on T. thermophila with 24 h of exposure. However, BP-3 and 4-MBC could inhibit the growth of T. thermophila. The 24-h EC50 values with 95% confidence intervals were 7.544 (6.561–8.675) mg L1 and 5.125 (4.874–5.388) mg L1, respectively. Of the four UV filters in this study, 4-MBC appeared to be most acutely toxic to T. thermophila, even if OC (Log KOW 6.88) is more lipophilic than 4-MBC (Log KOW 5.92). The results of cell viability assays with PI staining were consistent with that of the growth inhibition tests. EHMC and OC at the tested concentrations did not show any obvious cytotoxicity to T. thermophila cells, and accordingly did not affect their growth. BP-3 and 4-MBC could lead to the cell membranes impairment after 4 h exposure at the relatively higher concentrations, i.e. of 10.0 and 15.0 mg L1. Furthermore, with the increase of the exposure time to 6 h, their adverse effects on cell viability of T. thermophila were observed at the relatively lower concentration groups (1.0 mg L1 and 5.0 mg L1). The results of CAT activity and GSH content suggested that BP-3 and 4-MBC could exacerbate the oxidative stress of T.thermophila at environmentally relevant concentration. Although it was not observed the growth inhibitions and cell viability impairments of T. thermophila exposed to the selected UV filters at the environmentally relevant concentration, the adverse effects of UV filters must be taken into accounts as the aquatic organisms are exposed to them for the entire lifetime. Furthermore, the co-presence of other UV filters and compounds may contribute to the additive action or other sensitive endpoint changes (under submission). Considering the widespread use of UV filters, further investigations of their ecotoxicities and the mechanisms to aquatic organisms are imperative. Acknowledgments The research was supported by the National Science Foundation of China (No. 21277092), the Key Program of Shanghai Committee

of Science and Technology, China (No. 10JC1407800), the Key Innovation Research Project of Shanghai Education Committee (No. 12ZZ027) and the Key Discipline Construction Project of Shanghai Municipal Public Health (No. 12GWZX0401). Appreciation is extended to Professor Wei Miao in Institute of Hydrobiology, Chinese Academy of Sciences, for the complimentary of T. thermophila. References

Amine, H., Gomez, E., Halwani, J., Casellas, C., Fenet, H., 2012. UV filters, ethylhexyl methoxycinnamate, octocrylene and ethylhexyl dimethyl PABA from untreated wastewater in sediment from eastern Mediterranean river transition and coastal zones. Marine Pollution Bulletin 64 (11), 2435–2442. Bachelot, M., Li, Z., Munaron, D., Le Gall, P., Casellas, C., Fenet, H., Gomez, E., 2012. Organic UV filter concentrations in marine mussels from French coastal regions. Science of the Total Environment 420, 273–279. Balmer, M.E., Buser, H.R., Muller, M.D., Poiger, T., 2005. Occurrence of some organic UV filters in wastewater, in surface waters, and in fish from Swiss lakes. Environmental Science and Technology 39 (4), 953–962. Barata, C., Varo, I., Navarro, J.C., Arun, S., Porte, C., 2005. Antioxidant enzyme activities and lipid peroxidation in the freshwater cladoceran Daphnia magna exposed to redox cycling compounds. Comparative Biochemistry and Physiology Part C: Toxicology and Pharmacology 140 (2), 175–186. Buser, H.R., Balmer, M.E., Schmid, P., Kohler, M., 2006. Occurrence of UV filters 4methylbenzylidene camphor and octocrylene in fish from various Swiss rivers with inputs from wastewater treatment plants. Environmental Science and Technology 40 (5), 1427–1431. Christen, V., Zucchi, S., Fent, K., 2011. Effects of the UV-filter 2-ethyl-hexyl-4trimethoxycinnamate (EHMC) on expression of genes involved in hormonal pathways in fathead minnows (Pimephales promelas) and link to vitellogenin induction and histology. Aquatic Toxicology 102 (3–4), 167–176. Couto, N., Malys, N., Gaskell, S., Barber, J., 2013. Partition and turnover of glutathione reductase from saccharomyces cerevisiae: a proteomic approach. Journal of Proteome Research 12 (6), 2885–2894. Cuderman, P., Heath, E., 2007. Determination of UV filters and antimicrobial agents in environmental water samples. Analytical and Bioanalytical Chemistry 387 (4), 1343–1350. Danovaro, R., Bongiorni, L., Corinaldesi, C., Giovannelli, D., Damiani, E., Astolfi, P., Greci, L., Pusceddu, A., 2008. Sunscreens cause coral bleaching by promoting viral infections. Environmental Health Perspectives 116 (4), 441–447. Diaz-Cruz, M.S., Llorca, M., Barcelo, D., 2008. Organic UV filters and their photodegradates, metabolites and disinfection by-products in the aquatic environment. Trac-Trends in Analytical Chemistry 27 (10), 873–887. Diaz-Cruz, M.S., Gago-Ferrero, P., Llorca, M., Barcelo, D., 2012. Analysis of UV filters in tap water and other clean waters in Spain. Analytical and Bioanalytical Chemistry 402 (7), 2325–2333. Fent, K., Kunz, P.Y., Zenker, A., Rapp, M., 2010a. A tentative environmental risk assessment of the UV-filters 3-(4-methylbenzylidene-camphor), 2-ethyl-hexyl4-trimethoxycinnamate, benzophenone-3, benzophenone-4 and 3-benzylidene camphor. Marine Environmental Research 69 (Suppl), S4–S6. Fent, K., Zenker, A., Rapp, M., 2010b. Widespread occurrence of estrogenic UV-filters in aquatic ecosystems in Switzerland. Environmental Pollution 158 (5), 1817– 1824. Gago-Ferrero, P., Diaz-Cruz, M.S., Barcelo, D., 2011. Occurrence of multiclass UV filters in treated sewage sludge from wastewater treatment plants. Chemosphere 84 (8), 1158–1165. Gallego, A., Martin-Gonzalez, A., Ortega, R., Gutierrez, J.C., 2007. Flow cytometry assessment of cytotoxicity and reactive oxygen species generation by single and binary mixtures of cadmium, zinc and copper on populations of the ciliated protozoan Tetrahymena thermophila. Chemosphere 68 (4), 647–661. Giokas, D.L., Sakkas, V.A., Albanis, T.A., 2004. Determination of residues of UV filters in natural waters by solid-phase extraction coupled to liquid chromatographyphotodiode array detection and gas chromatography-mass spectrometry. Journal of Chromatography A 1026 (1–2), 289–293. Giokas, D.L., Salvador, A., Chisvert, A., 2007. UV filters: from sunscreens to human body and the environment. TrAC Trends in Analytical Chemistry 26 (5), 360– 374. Holbech, H., Norum, U., Korsgaard, B., Poul, B., 2002. The chemical UV-filter 3benzylidene camphor causes an oestrogenic effect in an in vivo fish assay. Pharmacology and Toxicology 91 (4), 204–208. Kaiser, D., Sieratowicz, A., Zielke, H., Oetken, M., Hollert, H., Oehlmann, J., 2012a. Ecotoxicological effect characterisation of widely used organic UV filters. Environmental Pollution 163, 84–90. Kaiser, D., Wappelhorst, O., Oetken, M., Oehlmann, J., 2012b. Occurrence of widely used organic UV filters in lake and river sediments. Environmental Chemistry 9 (2), 139–147. Kameda, Y., Kimura, K., Miyazaki, M., 2011. Occurrence and profiles of organic sunblocking agents in surface waters and sediments in Japanese rivers and lakes. Environmental Pollution 159 (6), 1570–1576. Kim, J.W., Chang, K.H., Isobe, T., Tanabe, S., 2011. Acute toxicity of benzotriazole ultraviolet stabilizers on freshwater crustacean (Daphnia pulex). Journal of Toxicological Sciences 36 (2), 247–251.

L. Gao et al. / Chemosphere 93 (2013) 2507–2513 Kunz, P.Y., Galicia, H.F., Fent, K., 2004. Assessment of hormonal activity of UV filters in tadpoles of frog Xenopus laevis at environmental concentrations. Marine Environmental Research 58 (2–5), 431–435. Lang, J., Kohidai, L., 2012. Effects of the aquatic contaminant human pharmaceuticals and their mixtures on the proliferation and migratory responses of the bioindicator freshwater ciliate Tetrahymena. Chemosphere 89 (5), 592–601. Li, W., Ma, Y., Guo, C., Hu, W., Liu, K., Wang, Y., Zhu, T., 2007. Occurrence and behavior of four of the most used sunscreen UV filters in a wastewater reclamation plant. Water Research 41 (15), 3506–3512. Liu, Y.S., Ying, G.G., Shareef, A., Kookana, R.S., 2011. Simultaneous determination of benzotriazoles and ultraviolet filters in ground water, effluent and biosolid samples using gas chromatography-tandem mass spectrometry. Journal of Chromatography A 1218 (31), 5328–5335. Liu, Y.S., Ying, G.G., Shareef, A., Kookana, R.S., 2012. Occurrence and removal of benzotriazoles and ultraviolet filters in a municipal wastewater treatment plant. Environmental Pollution 165, 225–232. Ma, R., Cotton, B., Lichtensteiger, W., Schlumpf, M., 2003. UV filters with antagonistic action at androgen receptors in the MDA-kb2 cell transcriptional-activation assay. Toxicological Sciences 74 (1), 43–50. Moeder, M., Schrader, S., Winkler, U., Rodil, R., 2010. At-line microextraction by packed sorbent-gas chromatography-mass spectrometry for the determination of UV filter and polycyclic musk compounds in water samples. Journal of Chromatography A 1217 (17), 2925–2932. Mortimer, M., Kasemets, K., Kahru, A., 2010. Toxicity of ZnO and CuO nanoparticles to ciliated protozoa Tetrahymena thermophila. Toxicology 269 (2–3), 182–189. Mottaleb, M.A., Usenko, S., O’Donnell, J.G., Ramirez, A.J., Brooks, B.W., Chambliss, C.K., 2009. Gas chromatography-mass spectrometry screening methods for select UV filters, synthetic musks, alkylphenols, an antimicrobial agent, and an insect repellent in fish. Journal of Chromatography A 1216 (5), 815–823. Negreira, N., Rodriguez, I., Rubi, E., Cela, R., 2010. Dispersive liquid-liquid microextraction followed by gas chromatography-mass spectrometry for the rapid and sensitive determination of UV filters in environmental water samples. Analytical and Bioanalytical Chemistry 398 (2), 995–1004. Poiger, T., Buser, H.R., Balmer, M.E., Bergqvist, P.A., Muller, M.D., 2004. Occurrence of UV filter compounds from sunscreens in surface waters: regional mass balance in two Swiss lakes. Chemosphere 55 (7), 951–963. Rodil, R., Schrader, S., Moeder, M., 2009. Non-porous membrane-assisted liquidliquid extraction of UV filter compounds from water samples. Journal of Chromatography A 1216 (24), 4887–4894. Rodil, R., Quintana, J.B., Concha-Grana, E., Lopez-Mahia, P., Muniategui-Lorenzo, S., Prada-Rodriguez, D., 2012. Emerging pollutants in sewage, surface and drinking water in Galicia (NW Spain). Chemosphere 86 (10), 1040–1049. Sanchez-Brunete, C., Miguel, E., Albero, B., Tadeo, J.L., 2011. Analysis of salicylate and benzophenone-type UV filters in soils and sediments by simultaneous extraction cleanup and gas chromatography-mass spectrometry. Journal of Chromatography A 1218 (28), 4291–4298.

2513

Schlenk, D., Sapozhnikova, Y., Irwin, M.A., Xie, L., Hwang, W., Reddy, S., Brownawell, B.J., Armstrong, J., Kelly, M., Montagne, D.E., Kolodziej, E.P., Sedlak, D., Snyder, S., 2005. In vivo bioassay-guided fractionation of marine sediment extracts from the Southern California Bight, USA, for estrogenic activity. Environmental Toxicology and Chemistry 24 (11), 2820–2826. Schlumpf, M., Cotton, B., Conscience, M., Haller, V., Steinmann, B., Lichtensteiger, W., 2001. In vitro and in vivo estrogenicity of UV screens. Environ Health Perspect 109 (3), 239–244. Schmitt, C., Oetken, M., Dittberner, O., Wagner, M., Oehlmann, J., 2008. Endocrine modulation and toxic effects of two commonly used UV screens on the aquatic invertebrates Potamopyrgus antipodarum and Lumbriculus variegatus. Environmental Pollution 152 (2), 322–329. Seward, J.R., Hamblen, E.L., Schultz, T.W., 2002. Regression comparisons of Tetrahymena pyriformis and Poecilia reticulata toxicity. Chemosphere 47 (1), 93–101. Sieratowicz, A., Kaiser, D., Behr, M., Oetken, M., Oehlmann, J., 2011. Acute and chronic toxicity of four frequently used UV filter substances for Desmodesmus subspicatus and Daphnia magna. Journal of Environmental Science and Health Part A 46 (12), 1311–1319. Sinks, G.D., Schultz, T.W., 2001. Correlation of Tetrahymena and Pimephales toxicity: evaluation of 100 additional compounds. Environmental Toxicology and Chemistry 20 (4), 917–921. Stefanidou, M.E., Hatzi, V.I., Terzoudi, G.I., Loutsidou, A.C., Maravelias, C.P., 2011. Effect of cocaine and crack on the ploidy status of Tetrahymena pyriformis: a study using DNA image analysis. Cytotechnology 63 (1), 35–40. Tarazona, I., Chisvert, A., Leon, Z., Salvador, A., 2010. Determination of hydroxylated benzophenone UV filters in sea water samples by dispersive liquid-liquid microextraction followed by gas chromatography–mass spectrometry. Journal of Chromatography A 1217 (29), 4771–4778. Weisbrod, C.J., Kunz, P.Y., Zenker, A.K., Fent, K., 2007. Effects of the UV filter benzophenone-2 on reproduction in fish. Toxicology and Applied Pharmacology 225 (3), 255–266. Yi, X., Ding, H., Lu, Y., Liu, H., Zhang, M., Jiang, W., 2007. Effects of long-term alachlor exposure on hepatic antioxidant defense and detoxifying enzyme activities in crucian carp (Carassius auratus). Chemosphere 68 (8), 1576–1581. Zhang, Z., Ren, N., Li, Y.F., Kunisue, T., Gao, D., Kannan, K., 2011. Determination of benzotriazole and benzophenone UV filters in sediment and sewage sludge. Environmental Science and Technology 45 (9), 3909–3916. Zhang, T., Li, X., Min, X., Fang, T., Zhang, Z., Yang, L., Liu, P., 2012. Acute toxicity of chlorobenzenes in Tetrahymena: estimated by microcalorimetry and mechanism. Environmental Toxicology and Pharmacology 33 (3), 377–385. Zucchi, S., Bluthgen, N., Ieronimo, A., Fent, K., 2011. The UV-absorber benzophenone-4 alters transcripts of genes involved in hormonal pathways in zebrafish (Danio rerio) eleuthero-embryos and adult males. Toxicology and Applied Pharmacology 250 (2), 137–146.