Polybrominated diphenyl ethers (PBDEs) effects on Chironomus sancticaroli larvae after short-term exposure

Ecotoxicology and Environmental Safety 139 (2017) 308–315

Contents lists available at ScienceDirect

Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv

Polybrominated diphenyl ethers (PBDEs) effects on Chironomus sancticaroli larvae after short-term exposure

MARK

⁎

Angela Maria Palacio-Cortésa, , Isadora de Lourdes Signorini-Souzaa, Emerson Luis Yoshio Harab, Rodrigo Geonildo Disnerc, Debora Rebechia, Marco Tadeu Grassib, Marta Margarete Cestaric, Mario Antônio Navarro-Silvaa a b c

Zoology Department, Federal University of Paraná, Brazil Chemistry Department, Federal University of Paraná, Brazil Genetic Department, Federal University of Paraná, Brazil

A R T I C L E I N F O

A B S T R A C T

Keywords: Midge Acetyl cholinesterase DNA

In-vivo effects of polybrominated diphenyl ethers (PBDEs) containing 3, 4 and 5 bromine atoms were tested on fourth-instar larvae of Chironomus sancticaroli (Diptera: Chironomidae) after 48 h of exposure, by measuring the activity of the acetyl cholinesterase, alpha and beta esterases and glutathione S-transferase. The PBDE congeners 2,2′,4-triBDE (BDE-17), 2,2′,4,4′-tetraBDE (BDE-47) and 2,2′,4,4′,5-pentaBDE (BDE-99) were evaluated at 0.5, 1.0, 2.0 and 3.0 ng mL−1. Acetyl cholinesterase activity decreased significantly (p≤0.05) at all evaluated concentrations of the three PBDE congeners, except for larvae exposed to BDE-17 at 1.0 and 2.0 ng mL−1. The significant inhibition of acetyl cholinesterase activity ranged from 18% (BDE-47 at 0.5 ng mL−1) to 72% (BDE47 at 2.0 ng mL−1). The enzymes alpha and beta esterase were also affected by the three congeners, reducing their activity from 14% (BDE-99 at 1.0 ng mL−1) to 52% (BDE-47 at 2.0 ng mL−1) and from 7% (BDE-99 at 2.0 ng mL−1) to 34% (BDE-47 at 3.0 ng mL−1) respectively. Substantial increments in glutathione S-transferase activity were similarly observed, varying from 138% (BDE-99 2.0 at ng mL−1) to 346% (BDE-17 at 1.0 ng mL−1). DNA strand breaks were detected exclusively in larvae exposed to BDE-99 at 2.0 and 3.0 ng mL−1 (H=11.7, p=0.019). These results showed that C. sancticaroli larvae were sensitive to the PBDEs treatments under the experimental conditions.

1. Introduction

the immune system of organisms (Dufault et al., 2005; Fernie et al., 2005; Man et al., 2011; Metcalfe et al., 2013; Ward et al., 2014). Even though some of the formulations of PBDEs were withdrawn from the market and banned by the Stockholm Convention, they are still present in environmental compartments (European Union, 2004; Great Lakes Chemical Corporation 2003; Stockholm Convention, 2009). PBDEs can even be detected at concentrations as low as picograms per litre the water column, sediments and soils (Lohmann et al., 2013; Richman et al., 2013; Shaw and Kannan, 2009). The most abundant congeners identified in environmental samples are 2,2′,4,4′-tetrabromodiphenyl ether (BDE-47) and the 2,2′,4,4′,5-pentabromodiphenyl ether (BDE-99), which are components of the penta-BDE commercial mixture (Dornbos et al., 2015; La Guardia et al., 2006). However, 2,2′,4-Tribromodiphenyl ether (BDE-17) has also been reported in environment compartments (Olukunle et al., 2015; Wong et al., 2016). Previous studies on Brazil displayed the presence of BDE-47 and BDE-99 in mussels, fish and the adipose breast tissue of humans;

Polybrominated diphenyl ethers (PBDEs) are a class of brominated flame retardants that are added in household and commercial polymerbased products, including textiles, electronic devices and furniture, to increase the fire resistance of these products (Alaee et al., 2003; WHO, 1997). There are 209 congeners, which are commercialized according to their degree of bromination as penta-BDE, octa-BDE and deca-BDE (de Wit, 2002;). These compounds are released into the environment during the manufacturing process of the aforementioned products, or leach from them into wastewaters as well as the atmosphere. In addition, they undergo global transport, are persistent and can bioaccumulate in aquatic and terrestrial organisms (de Wit et al., 2010; Hale et al., 2003, 2002; Law et al., 2006; North, 2004). The concerns about the effects of some congeners on humans and the environment are related to their effects on the reproductive, developmental systems, neurotoxic, carcinogenic and endocrine disruption, and their impact on

⁎

Correspondence to: Laboratory of Medical Entomology and Ecotoxicology, Zoology Department, Federal University of Paraná, CEP: 81531-980, Brazil. E-mail address: [email protected] (A.M. Palacio-Cortés).

http://dx.doi.org/10.1016/j.ecoenv.2017.01.052 Received 27 September 2016; Received in revised form 20 December 2016; Accepted 31 January 2017 0147-6513/ © 2017 Published by Elsevier Inc.

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

genotoxic effects of PBDE congeners on aquatic insects endemic to South America. Even though some Chironomidae species have been used as a model to assess the bioaccumulation of PBDEs (Bartrons et al., 2012, 2007), there are no data on the biochemical or molecular responses of this organism to brominated flame retardants. Taking all of the aforementioned aspects into account, the aim of the present study was to investigate the effects of BDE-17, BDE-47 and BDE-99 on the activity of the enzymes acetyl cholinesterase, alpha and beta esterase and glutathione S-transferase. In addition, DNA damage was also assessed using Chironomus sancticaroli larvae as a model organism under in vivo exposure conditions.

the BDE-17 congener was also reported in those samples (da Silva et al., 2013; Kalantzi et al., 2009). Considering the aquatic systems as destination of diverse environmental pollutants, habitat organisms can be affected. The non-biting midge Chironomus sancticaroli Strixino & Strixino, 1981 is a freshwater benthic insect, endemic from Latin America whose larvae are an important recycler of organic matter (Armitage et al., 1995; TrivinhoStrixino, 2011) and are an essential food source for predaceous organisms. Chironomid food chain position, short lifecycles, their easy identification and maintenance in the laboratory, and their physiological tolerance to different environmental conditions during the larval stages are features that contribute to their use as models to monitor the risk assessment of freshwater and sediment toxicity (Callisto et al., 2002; EPA, 1996; Lee and Choi, 2009; OECD/OCDE, 2010; Osmulski and Leyko, 1986; Qi et al., 2015a). The potentially harmful effects caused by environmental contaminants can be measured by biochemical and molecular markers that give early signals of the presence of those compounds in an ecosystem (McCarthy and Shugart, 1990; Walker, 2014). Cholinesterases and nonspecific esterases can change their activity when the organism is exposed to environmental contaminants. For instance, the AChE is inhibited by a wide range of compounds such as phosphates, carbamates, metal species, and surfactants, among other mixtures (de Lima et al., 2013; Domingues et al., 2010; Guilhermino et al., 2000). Its inhibition produces neurotoxicity in the organisms due to the accumulation of acetylcholine in the synapse area, leading to the over stimulation of the nervous system, which leads to the mortality (Walker et al., 2001). Chronic and acute tests on several Chironomus species have documented the AChE inhibition after exposure to organophosphorated and organochlorinated compounds (Printes et al., 2011; Rebechi et al., 2014), pyrethroids and polycyclic aromatic hydrocarbons (Ibrahim et al., 1998; Qi et al., 2015a, b). The non-specific esterases, α-esterase (EST-α) and β-esterase (ESTβ) are associated with the detoxification process of organisms and are also involved in the metabolic resistance process of insects (Hemingway and Ranson, 2000). EST-α and EST-β inhibition have been observed in larvae of C. sancticaroli and C. tentans Fabricius, 1805, after exposure to organophosphorated compounds (Rakotondravelo et al., 2006; Rebechi et al., 2014). Likewise, the antioxidant defence enzyme glutathione-Stransferase (GST) helps to protect the organism from contaminants and endogenous compounds throughout their conjugation with the tripeptide glutathione (GSH), which makes them more hydrophilic, promoting their excretion from the organism. Moreover, the GST forms covalent bonds with the electrophilic compounds produced by phase I enzymes, protecting the DNA and other cellular macromolecules of the activated species (Boelsterli, 2007; Enayati et al., 2005). It has also been suggested that the GST is involved in the formation of BDEglutathione metabolites in rodents (Hakk and Letcher, 2003) and avian species (Fernie et al., 2005), being also responsible for the metabolism of PBDEs via debromination in fish species (Noyes et al., 2010; Roberts et al., 2011). This enzyme has been used to identify physical and chemical stress in Chironomus riparius (Choi et al., 2000; Nair and Choi, 2011; Nair et al., 2013) and Chironomus tepperi Skuse, 1889 (Jeppe et al., 2014). The DNA damage, in terms of strand breaks, caused by contaminants can be evaluated using the comet assay, that detects genotoxicity in cells even after a few hours of exposure to toxicants at low concentrations (Frenzilli et al., 2009; Labieniec et al., 2007). Some studies have documented that the DNA of Chironomidae larvae is vulnerable to the effect of xenobiotics like metal species, bisphenol A, nonylphenol, pentachlorophenol, phenanthrene, tributyltin and triclosan, suggesting the use of Chironomus larvae as a model in genotoxic studies (Al-Shami et al., 2013; Martínez-Paz et al., 2013; Morais et al., 2014; Park and Choi, 2009). Despite the frequent detection of the PBDEs and their persistence in aquatic environments, there are no data on the biochemical and

2. Materials and methods 2.1. Exposure of chironomids Chironomus sancticaroli larvae were obtained from the colony of the Laboratory of Medical Entomology and Ecotoxicology maintained at the Federal University of Paraná. The colony was kept in aerated aquaria following the protocol of Maier et al. (1990), under 25 ± 2 °C, 80% relative humidity and photophase: scotophase (12:12). Larvae were fed three times a week with Dog Chow®. Voucher specimens of this colony are in the Entomology Museum Padre Jesus Santiago Moure of the Zoology Department at the Federal University of Parana (DZUP) with accession numbers from 249269 to 249276. Freshly laid egg masses from the colony were transferred to trays containing reconstituted water with 1.2 mg L−1 hydrated CaSO4, 0.08 mg L−1 KCl, 2.44 mg L−1 MgSO4·7H2O, and 1.92 mg L−1 Na2CO3, conductivity of 160 μS cm−1, pH 7.2 and hardness 16 mg L−1. Larvae were fed TetraMin® fish three times per week and constant aeration was maintained until they reached the fourth-instar, when they were exposed to PDBE. Bioassays were carried out in glass vessels containing ten larvae, 50 mL of reconstituted water and 13 g of sand 50–70 mesh Sigma®. The effects of individual PBDE congeners were tested at 0.5, 1.0, 2.0 and 3.0 ng mL−1 after 48 h of exposure. The control groups were exposed to acetone and water. Temperature, conductivity, pH and dissolved oxygen concentrations were determined at 0 h and 48 h of the exposure. Bioassays were conducted in a BOD chamber under 25 ± 2 °C, 80% relative humidity and photophase: scotophase (12:12). 2.2. Experimental solutions Analytical grade standards of BDE-17 (CAS No. 147217-75), BDE47 (CAS No 5436-43-1) and BDE-99 (CAS No 60348-60-9) in isooctane (50 μ mL−1) were purchased from Accustandard®. Stock solutions at 1000 μg L−1 for dosing were prepared in acetone and stored in amber glass vials at −20 °C until the bioassays started. 2.3. Enzyme activities The biochemical effects of PBDE were studied in 13 pools, each composed of five larvae. This totalled 65 individuals per each of the seven replicates of each concentration. A total of 910 larvae were then necessary to carry out this portion of the study. Exposed larvae were kept in eppendorf tubes and stored at −80 °C before enzyme activity measurements. Each pool was homogenized in 620 μL of Milli-Q type water and centrifuged at 12.000g for 1 min at 4 °C. 96 well microplates (Greiner bio-one) and a BioTek® microplate reader (BioTek Instruments, Inc.) were used to carry out the analyses. Total protein concentration was determined according Bradford's method (1976), using bovine serum albumin as the standard. A volume of 10 μL of supernatant and 250 μL of Bradford reagent (Sigma®) was placed in a microplate and the absorbance measured at 595 nm. All enzyme activities were run in triplicate wells. Cholinesterase (ChE) activity tests were performed in microplates 309

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

loaded with 25 μL of larvae homogenate, 200 μL of 0.75 mmol L−1 5´,5´- dithiobis (2-nitrobenzoic acid) (DTNB) in 100 mmol L−1 potassium phosphate buffer, pH 7.5 and 50 μL of 10 mmol L−1 acetylthiocholine iodine in Milli-Q water (Ellman et al., 1961). The mixture was incubated during 30 min at room temperature, and then the activity was measured every 30 s during five minutes at 405 nm and expressed as μmol min−1 mg−1 of protein. Alpha and beta esterase activities were measured using 200 μL of 0.3 mmol L−1 alpha naphthyl acetate and beta naphthyl acetate in potassium phosphate buffer added to 10 μL of the larvae homogenate (Valle et al., 2006). The mixture was incubated at room temperature during 15 min and then 50 μL of 0.3% Fast Blue B salt in 70% SDS and 30% Milli-Q water was added. After 5 min of incubation the activity was measured at 570 nm. Activity was expressed as nmol alpha naphthol min−1 mg−1 of protein and nmol beta naphthol min−1 mg−1 of protein respectively. To measure the glutathione S-transferase activity, 195 μL of 21 mmol L−1 2,4-dinitroclorobenzene (CDNB) and 100 mmol L−1 reduced glutathione (GSH) in 100 mmol L−1 potassium phosphate buffer, pH 6.5 was added to 15 μL of the larvae homogenate (Keen et al., 1976). The absorbance increase was measured during 20 min at intervals of 1 min at 340 nm and the activity was expressed as nmol min−1 mg−1 of protein.

were performed at 37 kHz and 154 W for 5 min. The formed emulsion was disrupted during 5 min of centrifugation at 5000 rpm and the organic phase was decanted at the bottom of the conical tube (Fontana et al., 2009). Decanted chloroform was removed by using a 250 μL chromatographic syringe, transferred to a 250 μL glass insert which was placed into a 2 mL GC glass vial and then the extracts were immediately subjected to GC-MS analysis. GC-MS analyses were performed using a Shimadzu 2010-Plus gas chromatograph coupled to a triple quadrupole mass detector TQ-8040 (Kyoto, Japan). Separation was carried out on a Rtx-5MS capillary column (30 m×0.25 mm i.d., 0.25 µm film thickness), also from Shimadzu. Helium (purity 99.99%) was employed as a carrier gas at a constant column flow of 1.0 mL min−1. The GC oven temperature was programmed from 60 °C during 3–200 °C at 20 °C min−1, to 260 °C at 10 °C min−1 and then until 300 °C during 4 min at 5 °C min−1. Total analysis time was 28 min. Injections (1 μL) were done in the splitless mode, using an injector temperature of 280 °C. The triple quadrupole mass spectrometer was operated in the electron impact ionization (EI) positive mode (+70 eV). The mass range was scanned in the full scan mode from 200 to 600 m/z. The ion source and transfer line were kept at 250 °C and 280 °C, respectively.

2.4. Genotoxicity assay

Enzyme activity results were expressed as mean ± S.D. and subjected to by one-way analysis of variance (ANOVA). Tukey test was used to identify the significant differences between treatments. The comet assay data were analysed using the non-parametric KruskalWallis test followed by the Student-Newman-Keuls when necessary. All differences were considered significant at p≤0,05 through BioEstat version 5.0 software (Ayres et al., 2003).

2.6. Data analyses

DNA damage was evaluated following the alkaline comet assay methodology (Lee and Choi, 2006) with minor modifications. Fifteen replicates were carried out, each containing 10 larvae for every PBDE concentration, as well as water and solvent control, totaling 2.100 larvae. After 48 h of PBDE exposure, larvae were transferred to 250 μL of fetal bovine serum and mechanically grinded. The homogenate was centrifuged during five min at 1.000g and the pellet was re-suspended in 120 μL of 0.5% low-melting point agarose. 60 μL of this suspension was casted on a microscope slide coated with normal agarose. Slides were incubated during 24 h in a lysis solution of 10 mmol L−1 Tris, 100 mmol L−1 dimethyl ethylene acetic acid (EDTA), 2.5 mol L−1 sodium chloride and 10% dimethyl sulfoxide. To unwind the DNA, slides were kept for 30 min in 300 mmol L−1 sodium hydroxide and 1 mmol L−1 dimethyl ethylene acetic acid. After that, electrophoresis was performed for 25 min at 1 Vcm−1. Subsequently the slides were neutralized during 15 min in 0.4 mol L−1 Tris, fixed with absolute ethanol for five minutes and finally stained with 0.20 mg mL−1 ethidium bromide. Comet formation was observed using a LEICA® epifluorescence microscopy 400× magnification. To detect DNA strand breaks 50 cells per slides were analysed. A visual classification was made based on fragment cells migration (Kobayashi et al., 1995).

3. Results Nominal and measured concentrations of the BDE-17, BDE-47 and BDE-99 studied with the respective uncertainties considering a 95% limit of confidence are presented in Table 1. None of the control larvae exhibited alteration in enzyme activity or DNA damage. Once the water and the solvent controls did not influence the larvae response, larvae belonging to the solvent control were compared with those of the treatments. The AChE activity of C. sancticaroli larvae was inhibited at almost all evaluated concentrations of the three studied PBDE congeners, except the BDE-17 at 1.0 and 2.0 ng mL−1. The AChE activity inhibition of larvae exposed to 0.5 and 3.0 ng mL−1 of BDE-17 showed 30% and 34% inhibition, which represents 149 ± 14 and 141 ± 15 μmol min−1 mg−1 of protein (Fig. 1A). The AChE activity of larvae exposed to BDE-47 was inhibited between 72% (BDE-47 at 2.0 ng mL−1) and 18% (BDE-47 at 0.5 ng mL−1), which corresponds to 58 ± 12 and 174 ± 30 μmol min−1 mg−1 of protein, respectively. Furthermore, the strong effect of this congener was notorious at 2.0 and 3.0 ng mL−1 when compared with 0.5 and 1.0 ng mL−1 (Fig. 1B). Inhibition of the AChE activity induced by BDE-99 was similar for all evaluated concentrations, ranging from 35% to 26%, which corresponds to 137 ± 15 to 158 ± 22 μmol min−1 mg−1 of protein (Fig. 1C). The EST-α activity in larvae, expressed in nmol of -α-naphtol min−1 mg−1 of protein, was affected by the three PBDE congeners (Fig. 2). The activity of EST-α of larvae exposed to BDE-17 ranged from 2.7 ± 0.3 to

2.5. Chemical analyses For the chemical analyses, water and sand were transferred to glass containers, and kept at −8.0 °C until PBDE determination. Sample volumes of 5 mL were placed into 15 mL conical bottom glass centrifuge tubes. Prior to the extraction, 100 ng mL−1 of PCB 209 were added to each sample as internal standard, 100 μL of chloroform were added as extraction solvent and 500 μL of NaCl 6.15 mol L−1 were added to favor the salting-out effect. The tube was immersed into an ultrasonic water bath Unique USC-1800 (São Paulo, Brazil). Extractions Table 1. Nominal and measured concentrations of BDE congeners in water exposure (ng mL−1). BDE-17 Nominal Measured

0.5 < 0.78

BDE-47 1.0 1.12 ± 0.14

2.0 1.95 ± 0.31

3.0 3.14 ± 0.35

0.5 < 0.81

BDE-99 1.0 1.14 ± 0.13

310

2.0 2.02 ± 0.25

3.0 3.10 ± 0.30

0.5 < 0.79

1.0 1.10 ± 0.12

2.0 2.04 ± 0.22

3.0 3.16 ± 0.32

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

Fig. 1. Average activity of the acetylcholinesterase enzyme expressed in μmol min−1 mg−1 of protein in Chironomus sancticaroli larvae after 48 h of exposure at 0.5, 1.0, 2.0 and 3.0 ng mL−1 of different PBDEs. A: BDE-17, B: BDE-47 and C: BDE-99. Bars with different letters indicate significantly different values (n=13, mean ± SD, Tukey's test, p < 0.05).

naphtol min−1 mg−1 of protein ranged from 2.7 ± 0.3 (31%) to 3.0 ± 0.2 (23%) for the BDE-17 (Fig. 3A); larvae exposed to BDE-47 presented enzymatic activity values ranging from 2.6 ± 0.2 (34%) to 3.2 ± 0.2 (19%) nmol of β-naphtol min−1 mg−1 of protein (Fig. 3B), and the range activity of the EST-β of larvae exposed to BDE-99 showed values from 3.1 ± 0.2 (21%) to 3.6 ± 0.3 (8%) (Fig. 3C). The GST activity of larvae was notoriously influenced by all PBDE congeners. The enzyme values in μmol min−1 mg−1 of protein ranged from 1079 ± 65 to 1252 ± 72 in larvae exposed to the BDE-17 which represents activity increases of 384% and 446% respectively (Fig. 4A).

3.2 ± 0.5, which indicates inhibitions of 39% and 28%, respectively (Fig. 2A); moreover, the BDE-47 showed activities fluctuating between 2.1 ± 0.3 to 3.5 ± 0.2, which corresponds to 52% and 21% inhibition (Fig. 2B). Furthermore, the BDE-99 exhibited activities from 3.4 ± 0.2 to 3.8 ± 0.5, which signify 23% and 14% inhibition, respectively (Fig. 2C). All congeners at all concentrations had influenced the activity of this enzyme. Among them BDE-47 at 2.0 ng mL−1 was the one responsible for the highest inhibition. The three PBDE congeners also influenced the EST-β activity of C. sancticaroli larvae (Fig. 3). The EST-β activity expressed in nmol of β-

Fig. 2. Average activity of the esterase alpha enzyme expressed in nmol alpha naphthol min−1 mg−1 of protein in Chironomus sancticaroli larvae after 48 h of exposure at 0.5, 1.0, 2.0 and 3.0 ng mL−1 of different PBDEs. A: BDE-17, B: BDE-47 and C: BDE-99. Bars with different letters indicate significantly different values (n=13, mean ± SD, Tukey's test, p < 0.05).

311

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

Fig. 3. Average activity of the esterase beta enzyme expressed in nmol beta naphthol min−1 mg−1 of protein in Chironomus sancticaroli larvae after 48 h of exposure at 0.5, 1.0, 2.0 and 3.0 ng mL−1 of different PBDEs. A: BDE-17, B: BDE-47 and C: BDE-99. Bars with different letters indicate significantly different values (n=13, mean ± SD, Tukey's test, p < 0.05).

4. Discussion

Larvae exposed to BDE-47 showed activities from 676 ± 95 (240%) to 910 ± 49 (324%) μmol min−1 mg−1 of protein (Fig. 4B), larvae exposed to the BDE-99 exhibited activities from 669 ± 49 (238%) to 707 ± 58 (252%) (Fig. 4C). DNA damages caused by the three PBDE studied were evaluated by alkali single cell gel electrophoresis. Larvae exposed to any concentration of the BDE-17 and BDE-47 did not show significant DNA strand break (H=2.9, p=0.57; H=5.6, p=0.22). However, BDE-99 showed a moderated genotoxic effect (H=11.7, p=0.019) at 2.0 and 3.0 ng mL−1 (Table 2).

Non-biting aquatic midges have been used to monitor the aquatic and sediment health conditions of different kinds of contaminants and C. sancticaroli larvae have shown biochemical alterations after exposure to PBDEs. In our study the AChE activity reduction in fourth-instar larvae caused by BDE-47 and BDE-99 congeners is consistent with others published experiments using different organisms. The mussel Mytilus galloprovincialis Lamark, 1819 showed reduction of the AChE activity induced by PBDEs after 48 h exposure at concentrations from 2 to 15 ng mL−1 (Vidal-Liñán et al., 2015). The presence of PBDE in tissues of Balanus amphitrite Darwin, 1854 also demonstrated a direct

Fig. 4. Average activity of the glutathione S-transferase enzyme expressed in nmol min−1 mg−1 of protein in Chironomus sancticaroli larvae after 48 h of exposure at 0.5, 1.0, 2.0 and 3.0 ng mL−1 of different PBDEs. A: BDE-17, B: BDE-47 and C: BDE-99. Bars with different letters indicate significantly different values (n=13, mean ± SD, Tukey's test, p < 0.05).

312

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

Table 2. DNA strand damage in Chironomus sancticaroli larvae after 48 h of exposure at 0.5, 1.0, 2.0 and 3.0 ng mL−1 of BDE-17, BDE-47 and BDE-99. Comet assay results are expressed as medians (lower to upper quartiles) of the scores, H-values based on Kruskal-Wallis (p < 0.05) and p-values based on Student-Newman-Keuls comparisons with the control condition. Each concentration included 50 cells per slide. PBDE

BDE-17 BDE-47 BDE-99

Concentration (ng mL−1)

H-values

Control

0.5

1.0

2.0

3.0

21 (15–49) 24 (22–34) 15.5 (7–23.2)

25 (19–40) 28 (19.7–41) 21.5 (13.2–33.2)

37.5 (21.5–50.5) 34 (24.5–46) 13.5 (9–25.5)

31 (22.5–52) 22 (13–30.0) 34 (21.5–48.5)*

30 (13.5–42.7) 22 (14–34) 33 (17–41.5)*

2.9 5.6 11.71*

* p < 0.05

obtained in cultured primary rat hippocampal neurons at concentrations between 1.0 to 20 ng mL−1, where authors noticed an increase in DNA damage related to an increase in the concentration of BDE-47 (P. He et al., 2008a, b). Correspondingly, effects of the BDE-47 at similar and even higher concentrations, which were used in this study (1.0, 2.0, 4.0 and 8 μg mL−1) were observed in human neuroblastoma cells after a 24 h exposure in vitro conditions (W. He et al., 2008a, b). The differences might be associated with the model organisms and experimental conditions. In addition, the lack of a positive genotoxic effect, mainly from the BDE-47, could be associated with the ability of chironomids to adapt to environmental stressors that encompass the efficiency of their defence mechanisms.

relationship with AChE activity. The authors of that study used a mixture of PBDEs where BDE-47 and BDE-99 were in higher concentrations (Chen et al., 2015). Danio rerio (Hamilton 1822) larvae also exhibited a decrease in AChE activity when their parents were exposed for 150 days to a commercial mixture that included the congeners 47 and 99 at concentrations ranging from 0.1 to 4 ng mL−1 (Chen et al., 2012). Despite the fact that all aforementioned studies were conducted using not only different organisms, but also distinct experimental conditions such as time of exposure, PBDE concentrations, the presence or absence of mixtures, among others, all of them share one result: the positive relationship between AChE inhibition and PBDE. Although the BDE-17 has already been described as a less toxic congener (De Wit, 2002), we were able to observe that AChE activity was significantly reduced at 0.5 and 3.0 ng mL−1. This results could indicate a hormetic response, that depicts a biphasic dose response with a low dose stimulation and a high-dose inhibitory effect (Calabrese and Baldwin, 2002; Calabrese, 2008). BDE-47 induced a hormetic effect in human hepatoma HepG2 cells after exposure to increasing concentrations (Wang et al., 2012). Furthermore, BDE-47 at 1 and 10 ng mL−1 produced a simillar effect in the osmotic regulation and metabolic energy alteration of the mussel Mytilus galloprovincialis (Ji et al., 2014). EST-α and EST-β activities showed similar sensitivity to inhibition by BDE-17, BDE-47 and BDE-99 at all evaluated concentrations, although in general no significant differences were observed when PBDE concentrations were increased from 0.5 up to 3.0 ng mL−1. This is the first time that non-specific esterases using the synthetic substrates alpha and beta naphthyl acetate are used to monitor the effect of PBDEs in Chironomids. It is well known that these enzymes can be inhibited by carbamate and organophosphate compounds, as described for the insect Chironomus sancticaroli and C. tentans (Rakotondravelo et al., 2006; Rebechi et al., 2014), gastropods like Biomphalaria glabrata and oligochaetes like Lumbricus terrestris and Lumbricus variegatus (Kristoff et al., 2010; Sanchez-Hernandez et al., 2009) that showed the esterase activity reduced. Induction of the GST activity by the three PBDE congeners in C. sancticaroli larvae indicate the response of the organisms to the presence of these compounds and may correspond to the initial detoxification process that protects cell integrity. Studies have demonstrated that PBDEs triggered an oxidative stress in cultured cells (He et al., 2008a, b; Jin et al., 2010; Shao et al., 2008) and in whole organisms (Fernie et al., 2005; Ghosh et al., 2013; Giordano et al., 2008; Zhao et al., 2011) when individual PBDE congeners or a mixture of them where evaluated, including the BDE-47 and BDE-99. Although we observed an increase of the GST activity, further indicators of ROS are necessary to conclude that oxidative stress effects happened. Regarding DNA integrity, BDE-99 was the only compound that produced DNA strand breaks in larvae. Similar results were also observed in the freshwater mussel Dreissena polymorpha (Pallas 1771) when another environmentally relevant pentabrominated compound, BDE-100 at 0.1 and 1 ng mL−1 (48 h exposure), was studied (Parolini and Binelli, 2012). On the other hand, the absence of a genotoxic effect on C. sancticaroli larvae caused by the BDE-47 contrasts with the results

5. Conclusions To our knowledge, this is the first study on the effects of PBDE on the physiology of C. sancticaroli. Our results show that BDE-17, BDE-47 and BDE-99 induced changes in enzymes associated with neurotoxic effects. Although cholinesterases and non-specific esterases have been used to identify exposure to organophosphate and carbamate pesticides, we have collected evidence of its sensitivity to the presence of the polybrominated diphenyl ethers studied. In order to advance our knowledge about PBDEs toxicity on C. sancticaroli, studies on population growth, reproductive success and kinetic behaviour should be performed. Conflict of interest The authors declare that they have no conflict of interest. Acknowledgements The authors would like to thank the financial support of the cooperation program between Newton Fund, Conselho Nacional das Fundações Estaduais de Amparo à Pesquisa - Fundação Araucária and Research Councils UK (RCUK) (Grant number 45191.460.30251.29012015); the National Counsel of Technological and Scientific Development (Grant numbers 305470/2012-4 and 309576/2014-8). A.M. Palacio-Cortés is grateful to CAPES-PNPD (Grand Number 20132185) for the fellowship. References Alaee, M., Arias, P., Sjödin, A., Bergman, Å., 2003. An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 29, 683–689. http:// dx.doi.org/10.1016/S0160-4120(03)00121-1. Al-Shami, S.A., Che Salmah, M.R., Mohd Nor, S.A., Ahmad, A.H., Mohd Ishadi, N.A., Dieng, H., 2013. Genotoxicity in Chironomus kiiensis (Chironomidae: diptera) after exposure to polluted sediments from rivers of north peninsular malaysia: implication for ecotoxicological monitoring. River Res. Appl. 29, 1195–1204. Armitage, P.D., Pinder, L.C., Cranston, P., 1995. The Chironomidae – The Biology and Ecology of Non-bitting Midges. Chapman & Hall, London. Ayres M., Ayres Jr M., Ayres D., Santos A., 2003. BioEstat 3.0 Aplicações estatísticas nas áreas das ciênciasbiológicas e médicas. Sociedade Civil Mamirauá, Belém. Bartrons, M., Grimalt, J.O., Catalan, J., 2007. Concentration changes of organochlorine compounds and polybromodiphenyl ethers during metamorphosis of aquatic insects.

313

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

retardant-treated polyurethane foam as a source of brominated diphenyl ethers to the US environment. Chemosphere 46, 729–735. Hale, R.C., Alaee, M., Manchester-Neesvig, J.B., Stapleton, H.M., Ikonomou, M.G., 2003. Polybrominated diphenyl ether flame retardants in the North American environment. Environ. Int. 29, 771–779. http://dx.doi.org/10.1016/S0160-4120(03)00113-2. He, P., He, W., Wang, A., Xia, T., Xu, B., Zhang, M., Chen, X., 2008a. PBDE-47-induced oxidative stress, DNA damage and apoptosis in primary cultured rat hippocampal neurons. Neurotoxicology 29, 124–129. http://dx.doi.org/10.1016/ j.neuro.2007.10.002. He, W., He, P., Wang, A., Xia, T., Xu, B., Chen, X., 2008b. Effects of PBDE-47 on cytotoxicity and genotoxicity in human neuroblastoma cells in vitro. Mutat. Res. 649, 62–70. http://dx.doi.org/10.1016/j.mrgentox.2007.08.001. Hemingway, J., Ranson, H., 2000. Insecticide resistance in insect vectors of human disease. Annu Rev. Entomol. 45, 371–391. http://dx.doi.org/10.1146/ annurev.ento.45.1.371. Ibrahim, H., Kheir, R., Helmi, S., Lewis, J., Crane, M., 1998. Effects of organophosphorus, carbamate, pyrethroid and organochlorine pesticides, and a heavy metal on survival and cholinesterase activity of Chironomus riparius Meigen. Bull. Environ. Contam. Toxicol. 60, 448–455. http://dx.doi.org/10.1007/s001289900646. Jeppe, K.J., Carew, M.E., Long, S.M., Lee, S.F., Pettigrove, V., Hoffmann, A.A., 2014. Genes involved in cysteine metabolism of Chironomus tepperi are regulated differently by copper and by cadmium. Comp. Biochem. Physiol. C. Toxicol. Pharm. 162, 1–6. http://dx.doi.org/10.1016/j.cbpc.2014.02.006. Ji, C., Zhao, J., Wu, H., 2014. Gender-specific metabolic responses in gonad of mussel Mytilus galloprovincialis to 2,2′,4,4′-tetrabromodiphenyl ether. Environ. Toxicol. Pharm. 37, 1116–1122. http://dx.doi.org/10.1016/j.etap.2014.04.007. Jin, S., Yang, F., Hui, Y., Xu, Y., Lu, Y., Liu, J., 2010. Cytotoxicity and apoptosis induction on RTG-2 cells of 2,2′,4,4′-tetrabromodiphenyl ether (BDE-47) and decabrominated diphenyl ether (BDE-209). Toxicol. Vitr. 24, 1190–1196. http://dx.doi.org/10.1016/ j.tiv.2010.02.012. Kalantzi, O.I., Brown, F.R., Caleffi, M., Goth-Goldstein, R., Petreas, M., 2009. Polybrominated diphenyl ethers and polychlorinated biphenyls in human breast adipose samples from Brazil. Environ. Int. 35, 113–117. http://dx.doi.org/10.1016/ j.envint.2008.09.008. Keen, J.H., Habig, W.H., Jakoby, W.B., 1976. Mechanism for the several activities of the glutathione S-transferases. J. Biol. Chem. 251, 6183–6188. Kobayashi, C., Sugiyama, H., Morikawa, Y., Hayashi, M., Sofuni, T., 1995. A comparison between manual microscopic analysis and computerized image analysis in the single cell gel electrophoresis. MMS Commun. 3, 103–115. Kristoff, G., Guerrero, N.R.V., Cochón, A.C., 2010. Inhibition of cholinesterases and carboxylesterases of two invertebrate species, Biomphalaria glabrata and Lumbriculus variegatus, by the carbamate pesticide carbaryl. Aquat. Toxicol. 96, 115–123. http:// dx.doi.org/10.1016/j.aquatox.2009.10.001. La Guardia, M.J., Hale, R.C., Harvey, E., 2006. Detailed polybrominated diphenyl ether (PBDE) congener composition of the widely used penta-, octa-, and deca-PBDE technical flame-retardant mixtures. Environ. Sci. Technol. 40, 6247–6254. Labieniec, M., Biernat, M., Gabryelak, T., 2007. Response of digestive gland cells of freshwater mussel Unio tumidus to phenolic compound exposure in vivo. Cell Biol. Int. 31, 683–690. http://dx.doi.org/10.1016/j.cellbi.2006.12.005. Law, K., Halldorson, T., Danell, R., Stern, G., Gewurtz, S., Alaee, M., Marvin, C., Whittle, M., Tomy, G., 2006. Bioaccumulation and trophic transfer of some brominated flame retardants in a Lake Winnipeg (Canada) food web. Environ. Toxicol. Chem. 25, 2177–2186. Lee, S.B., Choi, J., 2006. Multilevel evaluation of nonylphenol toxicity in fourth-instar larvae of Chironomus riparius (Diptera, Chironomidae). Environ. Toxicol. Chem. 25, 3006–3014. http://dx.doi.org/10.1897/05-601r1.1. Lee, S.W., Choi, J., 2009. Multi-level ecotoxicity assay on the aquatic midge, Chironomus tentans (Diptera, Chironomidae) exposed to octachlorostyrene. Environ. Toxicol. Pharm. 28, 269–274. http://dx.doi.org/10.1016/j.etap.2009.05.004. de Lima, D., Roque, G.M., de Almeida, E.A., 2013. In vitro and in vivo inhibition of acetylcholinesterase and carboxylesterase by metals in zebrafish (Danio rerio). Mar. Environ. Res. 91, 45–51. http://dx.doi.org/10.1016/j.marenvres.2012.11.005. Lohmann, R., Klanova, J., Kukucka, P., Yonis, S., Bollinger, K., 2013. Concentrations, fluxes, and residence time of PBDEs across the tropical Atlantic Ocean. Environ. Sci. Technol. 47, 13967–13975. http://dx.doi.org/10.1021/es403494b. Maier, K.J., Kosalwat, P., Knight, A.W., 1990. Culture of Chironomus decorus (Diptera: Chironomidae) and the effect of temperature on its life history. Environ. Entomol. 19, 1681–1688. Man, Y.B., Lopez, B.N., Wang, H.S., Leung, A.O.W., Chow, K.L., Wong, M.H., 2011. Cancer risk assessment of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in former agricultural soils of Hong Kong. J. Hazard Mater. 195, 92–99. http://dx.doi.org/10.1016/j.jhazmat.2011.08.010. Martínez-Paz, P., Morales, M., Martínez-Guitarte, J.L., Morcillo, G., 2013. Genotoxic effects of environmental endocrine disruptors on the aquatic insect Chironomus riparius evaluated using the comet assay. Mutat. Res. Toxicol. Environ. Mutagen 758, 41–47. http://dx.doi.org/10.1016/j.mrgentox.2013.09.005. McCarthy, J., Shugart, L., 1990. Biological markers of environmental contamination. In: McCarthy, J., Shugart, L. (Eds.), Biomarkers of Environmental Contamination. Lewis Publishers, Boca Raton, pp. 3–14. Metcalfe, C.D., Kleywegt, S., Letcher, R.J., Topp, E., Wagh, P., Trudeau, V.L., Moon, T.W., 2013. A multi-assay screening approach for assessment of endocrine-active contaminants in wastewater effluent samples. Sci. Total Environ. 454–455, 132–140. http://dx.doi.org/10.1016/j.scitotenv.2013.02.074. Morais, S., Pesenti, E.C., Cestari, M.M., Navarro-silva, M.A., 2014. Genotoxic effect of phenanthrene on Chironomus sancticaroli (Diptera: Chironomidae). Zoologia 31, 323–328.

Environ. Sci. Technol. 41, 6137–6141. http://dx.doi.org/10.1021/es0703271. Bartrons, M., Grimalt, J., Mendoza, G., Catalan, J., 2012. Pollutant dehalogenation capability may depend on the trophic evolutionary history of the organism: PBDEs in freshwater food webs. PLoS One 7, e1829. http://dx.doi.org/10.1371/ journal.pone.0041829. Boelsterli, U.A., 2007. Mechanistic Toxicology: the Molecular Basis of How Chemicals Disrupt Biological Targets. CRC Press, Boca Raton. Bradford, M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248–254. http://dx.doi.org/10.1016/0003-2697(76)90527-3. Calabrese, E., 2008. Why it is important to toxicology and toxicologists. Environ. Toxico Chem. 27, 1451–1474. http://dx.doi.org/10.1897/07-541.1. Calabrese, E., Baldwin, L., 2002. Defining hormesis. Hum. Exp. Toxicol. 21, 91–97. http://dx.doi.org/10.1191/0960327102ht217oa. Callisto, M., Moreno, P., Gonçalves Júnior, J.F., Leal, J.J.F., Esteves, F.A., 2002. Diversity and biomass of chironomidae (Diptera) larvae in an impacted coastal lagoon in Rio De Janeiro, Brazil. Braz. J. Biol. 62, 77–84. Chen, L., Yu, K., Huang, C., Yu, L., Zhu, B., Lam, P.K.S., Lam, J.C.W., Zhou, B., 2012. Prenatal transfer of polybrominated diphenyl ethers (PBDEs) results in developmental neurotoxicity in zebrafish larvae. Environ. Sci. Technol. 46, 9727–9734. http://dx.doi.org/10.1021/es302119g. Chen, L., Lam, J.C.W., Zhang, X., Pan, K., Guo, C., Lam, P.K.S., Wang, W., Liu, H., Qian, P.-Y., 2015. Relationship between metal and polybrominated diphenyl ether (PBDE) body burden and health risks in the barnacle Balanus amphitrite. Mar. Pollut. Bull. 100, 383–392. http://dx.doi.org/10.1016/j.marpolbul.2015.08.020. Choi, J., Roche, H., Caquet, T., 2000. Effects of physical (hypoxia, hyperoxia) and chemical (potassium dichromate, fenitrothion) stress on antioxidant enzyme activities in Chironomus riparius mg. (diptera, chironomidae) larvae: potential biomarkers. Environ. Toxicol. Chem. 19, 495–500. http://dx.doi.org/10.1002/ etc.5620190231. Domingues, I., Agra, A.R., Monaghan, K., AMVM, Soares, AJA, Nogueira, 2010. Cholinesterase and glutathione-S-transferase activities in freshwater invertebrates as biomarkers to assess pesticide contamination. Environ. Toxicol. Chem. 29, 5–18. http://dx.doi.org/10.1002/etc.23. Dornbos, P., Chernyak, S., Rutkiewicz, J., Cooley, T., Strom, S., Batterman, S., Basu, N., 2015. Hepatic polybrominated diphenyl ether (PBDE) levels in Wisconsin river otters (Lontra canadensis) and Michigan bald eagles (Haliaeetus leucocephalus). J. Gt. Lakes Res. 41, 222–227. http://dx.doi.org/10.1016/j.jglr.2014.12.023. Dufault, C., Poles, G., Driscoll, L.L., 2005. Brief postnatal PBDE exposure alters learning and the cholinergic modulation of attention in rats. Toxicol. Sci. 88, 172–180. http:// dx.doi.org/10.1093/toxsci/kfi285. Ellman, G.L., Courtney, K.D., Andres, V., Featherstone, R.M., 1961. A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharm. 7, 88–95. http://dx.doi.org/10.1016/0006-2952(61)90145-9. Enayati, A.A., Ranson, H., Hemingway, J., 2005. Insect glutathione transferases and insecticide resistance. Insect Mol. Biol. 14, 3–8. http://dx.doi.org/10.1111/j.13652583.2004.00529.x. EPA, 1996. Chironomid sediment toxicity test, ecological effects tests guidelines. EPA 712-C-96-313. Washington, DC. European Union, 2004. Directive 2003/11/EC of the European Parliament and of the council of 6 February 2003 amending for the 24th time Council Directive 76/769/ EEC relating to restrictions on the marketing and use of certain dangerous substances and preparations. Available: 〈http://eur-lex.europa.eu/legal-content/EN/TXT/PDF/ ?Uri=CELEX:32003L0011 & from=EN〉.( accessed 20.05.15). Fernie, K.J., Shutt, J.L., Mayne, G., Hoffman, D., Letcher, R.J., Drouillard, K.G., Ritchie, I.J., 2005. Exposure to polybrominated diphenyl ethers (PBDEs): changes in thyroid, vitamin A, glutathione homeostasis, and oxidative stress in American kestrels (Falco sparverius). Toxicol. Sci. 88, 375–383. http://dx.doi.org/10.1093/toxsci/kfi295. Fontana, A.R., Wuilloud, R.G., Martínez, L.D., Altamirano, J.C., 2009. Simple approach based on ultrasound-assisted emulsification-microextraction for determination of polibrominated flame retardants in water samples by gas chromatography-mass spectrometry. J. Chromat. A 1216 (1), 147–153. http://dx.doi.org/10.1016/ j.chroma.2008.11.034. Frenzilli, G., Nigro, M., Lyons, B.P., 2009. The comet assay for the evaluation of genotoxic impact in aquatic environments. Mutat. Res. 681, 80–92. http://dx.doi.org/10.1016/ j.mrrev.2008.03.001. Ghosh, R., Lokman, P.M., Lamare, M.D., Metcalf, V.J., Burritt, D.J., Davison, W., Hageman, K.J., 2013. Changes in physiological responses of an Antarctic fish, the emerald rock cod (Trematomus bernacchii), following exposure to polybrominated diphenyl ethers (PBDEs). Aquat. Toxicol. 128–129, 91–100. http://dx.doi.org/ 10.1016/j.aquatox.2012.11.019. Giordano, G., Kavanagh, T.J., Costa, L.G., 2008. Neurotoxicity of a polybrominated diphenyl ether mixture (DE-71) in mouse neurons and astrocytes is modulated by intracellular glutathione levels. Toxicol. Appl. Pharm. 232, 161–168. http:// dx.doi.org/10.1016/j.taap.2008.06.018. Great Lakes Chemical Corporation News Release, 2003. Great Lakes will cease production of penta-PBDE flame retardant by end of 2004. Gt. Lakes Chem Corp online 〈www. greatlakes.com〉 issued Nov. 3:2003 Guilhermino, L., Lacerda, M., Nogueira, A.J., Soares, A.M.V., 2000. In vitro and in vivo inhibition of Daphnia magna acetylcholinesterase by surfactant agents: possible implications for contamination biomonitoring. Sci. Total Environ. 247, 137–141. http://dx.doi.org/10.1016/S0048-9697(99)00485-4. Hakk, H., Letcher, R.J., 2003. Metabolism in the toxicokinetics and fate of brominated flame retardants–a review. Environ. Int. 29, 801–828. http://dx.doi.org/10.1016/ S0160-4120(03)00109-0. Hale, R.C., La Guardia, M.J., Harvey, E., Mainor, T.M., 2002. Potential role of fire

314

Ecotoxicology and Environmental Safety 139 (2017) 308–315

A.M. Palacio-Cortés et al.

congeners in three fish species. Environ. Sci. Technol. 45, 1999–2005. http:// dx.doi.org/10.1021/es103934x. Sanchez-Hernandez, J.C., Mazzia, C., Capowiez, Y., Rault, M., 2009. Carboxylesterase activity in earthworm gut contents: potential (eco)toxicological implications. Comp. Biochem. Physiol. C. Toxicol. Pharm. 150, 503–511. http://dx.doi.org/10.1016/ j.cbpc.2009.07.009. Shao, J., Eckert, M.L., Lee, L.E.J., Gallagher, E.P., 2008. Comparative oxygen radical formation and toxicity of BDE 47 in rainbow trout cell lines. Mar. Environ. Res. 66, 7–8. http://dx.doi.org/10.1016/j.marenvres.2008.02.007. Shaw, S.D., Kannan, K., 2009. Polybrominated diphenyl ethers in marine ecosystems of the American continents: foresight from current knowledge. Rev. Environ. Health 24, 157–229. da Silva, S.F.G., IMRA, Brüning, Montone, Taniguchi, R.C., Cascaes, S., Dias, M.J., Lavandier, P.S., Hauser-Davis, R.C., Moreira I, R.A., 2013. Polybrominated diphenyl ethers (PBDES) and polychlorinated biphenyls (PCBS) in mussels and two fish species from the estuary of the Guanabara Bay, southeastern Brazil. Bull. Environ. Contam Toxicol. 91, 261–266. http://dx.doi.org/10.1007/s00128-013-1026-3. Stockholm Convention, 2009. Access. 20 March 2016〈http://chm.pops.int/ TheConvention/ThePOPs/TheNewPOPs/tabid/2511/Default.aspx〉. Trivinho-Strixino, S., 2011. Chironomidae (Insecta, Diptera, Nematocera) do Estado de São Paulo, Sudeste do Brasil. Biota Neotrop. 11, 675–684. Valle D., Montella I., Ribeiro R., Medeiros P., Martins-Jr A., Lima J., 2006. Quantification methodology for enzyme activity related to insecticide resistance in Aedes aegypti. Fundação Oswaldo Cruz and Secretaria de Vigilância em Saúde, Rio de Janeiro Vidal-Liñán, L., Bellas, J., Fumega, J., Beiras, R., 2015. Bioaccumulation of BDE-47 and effects on molecular biomarkers acetylcholinesterase, glutathione-S-transferase and glutathione peroxidase in Mytilus galloprovincialis mussels. Ecotoxicology 24, 292–300. http://dx.doi.org/10.1007/s10646-014-1377-5. Walker, C., 2014. Ecotoxicology: effects of pollutants on the natural environment. CRC Press, London. Walker, C.H., Hopkin, S.P., Sibly, R.M., Peakall, D.B., 2001. Principles of ecotoxicology. Taylor & Francis, London. Wang, L., Zou, W., Zhong, Y., An, J., Zhang, X., Wu, M., Yu, Z., 2012. The hormesis effect of BDE-47 in HepG2 cells and the potential molecular mechanism. Toxicol. Lett. 209, 193–201. http://dx.doi.org/10.1016/j.toxlet.2011.12.014. Ward, M., Colt, J., Deziel, N., Whitehead, T., Reynolds, P., Gunier, R., Nishioka, M., Dahl, G., Rappaport, S., Buffer, P., Metayer, C., 2014. Residential levels of polybrominated diphenyl ethers and risk of childhood acute lymphoblastic leukemia in California. Env. Heal Perspect. 122, 1110–1116. http://dx.doi.org/10.1289/ehp.1307602. WHO (1997) Flame retardans: a general introduction. World Health Organization. Environmental Health Criteria 192. de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583–624. http://dx.doi.org/10.1016/S0045-6535(01)00225-9. de Wit, C.A., Herzke, D., Vorkamp, K., 2010. Brominated flame retardants in the Arctic environment - trends and new candidates. Sci. Total Environ. 408, 2885–2918. http://dx.doi.org/10.1016/j.scitotenv.2009.08.037. Wong F., Suzuki G., Michinaka C., Yuan B., Takigami H., de Wit C., 2016. Dioxin-like activities, halogenated flame retardants, organophosphate esters and chlorinated paraffins in dust from Australia, the United Kingdom, Canada, Sweden and China. Chemosphere. In Press. doi.org/http://dx.doi.org/10.1016/j.chemosphere.2016.10. 074. Zhao, A., Liu, H., Zhang, A., Wang, X., Zhang, H., Wang, H., 2011. Effect of BDE-209 on glutathione system in Carassius auratus. Environ. Toxicol. Pharm. 32, 35–39. http:// dx.doi.org/10.1016/j.etap.2011.03.004.

Nair, P.M.G., Choi, J., 2011. Identification, characterization and expression profiles of Chironomus riparius glutathione S-transferase (GST) genes in response to cadmium and silver nanoparticles exposure. Aquat. Toxicol. 101, 550–560. http://dx.doi.org/ 10.1016/j.aquatox.2010.12.006. Nair, P.M.G., Park, S.Y., Choi, J., 2013. Evaluation of the effect of silver nanoparticles and silver ions using stress responsive gene expression in Chironomus riparius. Chemosphere 92, 592–599. http://dx.doi.org/10.1016/j.chemosphere.2013.03.060. North, K.D., 2004. Tracking polybrominated diphenyl ether releases in a wastewater treatment plant effluent, Palo Alto, California. Environ. Sci. Technol. 38, 4484–4488. Noyes, P.D., Kelly, S.M., Mitchelmore, C.L., Stapleton, H.M., 2010. Characterizing the in vitro hepatic biotransformation of the flame retardant BDE 99 by common carp. Aquat. Toxicol. 97, 142–150. http://dx.doi.org/10.1016/j.aquatox.2009.12.013. OECD/OCDE (2010) Sediment-water Chironomid life-cycle toxicity test using spiked water or spiked sediment. OECD Guidelines for the testing of chemicals. 233: pp. 1–29. Olukunle, O.I., Sibiya, I.V., Okonkwo, O.J., Odusanya, A.O., 2015. Influence of physicochemical and chemical parameters on polybrominated diphenyl ethers in selected landfill leachates, sediments and river sediments from Gauteng, South Africa. Environ. Sci. Pollut. Res. Int. 22, 2145–2154. http://dx.doi.org/10.1007/ s11356-014-3443-1. Osmulski, P.A., Leyko, W., 1986. Structure, function and physiological role of Chironomus haemoglobin. Comp. Biochem. Physiol. B 85, 701–722. Park, S.Y., Choi, J., 2009. Genotoxic effects of nonylphenol and bisphenol A exposure in aquatic biomonitoring species: freshwater crustacean, Daphnia magna, and aquatic midge, Chironomus riparius. Bull. Environ. Contam. Toxicol. 83, 463–468. http:// dx.doi.org/10.1007/s00128-009-9745-1. Parolini, M., Binelli, A., 2012. Cyto-genotoxic effects induced by three brominated diphenyl ether congeners on the freshwater mussel Dreissena polymorpha. Ecotoxicol. Environ. Saf. 79, 247–255. http://dx.doi.org/10.1016/j.ecoenv.2012.01.008. Printes, L.B., Fernandes, M.N., Espíndola, E.L.G., 2011. Laboratory measurements of biomarkers and individual performances in Chironomus xanthus to evaluate pesticide contamination of sediments in a river of southeastern Brazil. Ecotoxicol. Environ. Saf. 74, 424–430. http://dx.doi.org/10.1016/j.ecoenv.2010.10.033. Qi, H.X., Li, H.Z., Ma, P., You, J., 2015a. Integrated sediment quality assessment through biomarker responses and bioavailability measurements: application in Tai Lake, China. Ecotoxicol. Environ. Saf. 119, 148–154. http://dx.doi.org/10.1016/ j.ecoenv.2015.05.007. Qi, H.X., Ma, P., Li, H., You, J., 2015b. Assessment of sediment risk in the North End of Tai lake, China: integrating chemical analysis and chronic toxicity testing with Chironomus dilutus. Arch. Environ. Contam. Toxicol. 69, 461–469. http://dx.doi.org/ 10.1007/s00244-015-0162-7. Rakotondravelo, M.L., Anderson, T.D., Charlton, R.E., Zhu, K.Y., 2006. Sublethal effects of three pesticides on activities of selected target and detoxification enzymes in the aquatic midge, Chironomus tentans (diptera: chironomidae). Arch. Environ. Contam. Toxicol. 51, 360–366. http://dx.doi.org/10.1007/s00244-005-0227-0. Rebechi, D., Richardi, V.S., Vicentini, M., Guiloski, I.C., Assis, H.C.S., Navarro-Silva, M.A., 2014. Low malathion concentrations influence metabolism in Chironomus sancticaroli (Diptera, Chironomidae) in acute and chronic toxicity tests. Rev. Bras. Entomol. 58, 296–301. http://dx.doi.org/10.1590/S0085-56262014000300012. Richman, L.A., Kolic, T., MacPherson, K., Fayez, L., Reiner, E., 2013. Polybrominated diphenyl ethers in sediment and caged mussels (Elliptio complanata) deployed in the Niagara River. Chemosphere 92, 778–786. http://dx.doi.org/10.1016/ j.chemosphere.2013.04.011. Roberts, S.C., Noyes, P.D., Gallagher, E.P., Stapleton, H.M., 2011. Species-specific differences and structure-activity relationships in the debromination of PBDE

315