Toxicity and bioaccumulation of three hexabromocyclododecane diastereoisomers in the marine copepod Tigriopus japonicas

Accepted Manuscript Title: Toxicity and Bioaccumulation of Three Hexabromocyclododecane Diastereoisomers in the Marine Copepod Tigriopus japonicas Authors: Haizheng Hong, Dongmei Lv, Wanxin Liu, Lingming Huang, Leyun Chen, Rong Shen, Dalin Shi PII: DOI: Reference:

S0166-445X(17)30112-1 http://dx.doi.org/doi:10.1016/j.aquatox.2017.04.010 AQTOX 4642

To appear in:

Aquatic Toxicology

Received date: Revised date: Accepted date:

19-1-2017 12-4-2017 14-4-2017

Please cite this article as: Hong, Haizheng, Lv, Dongmei, Liu, Wanxin, Huang, Lingming, Chen, Leyun, Shen, Rong, Shi, Dalin, Toxicity and Bioaccumulation of Three Hexabromocyclododecane Diastereoisomers in the Marine Copepod Tigriopus japonicas.Aquatic Toxicology http://dx.doi.org/10.1016/j.aquatox.2017.04.010 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Toxicity and Bioaccumulation of Three Hexabromocyclododecane Diastereoisomers in the Marine Copepod Tigriopus japonicas Haizheng Hong a,b,c,*, Dongmei Lv a, Wanxin Liu a, Lingming Huang a, Leyun Chen a, Rong Shena, Dalin Shi a,b

a

State Key Laboratory of Marine Environmental Science, Xiamen University, Xiamen 361102,

China b

Key Laboratory of the Ministry of Education for Coastal and Wetland Ecosystems, Xiamen

University, Xiamen 361102, China c

Center for Marine Environmental Chemistry and Toxicology, Xiamen University, Xiamen

361102, China

*ADDRESS CORRESPONDENCE TO: Haizheng Hong, Ph. D. Email: [email protected] Phone: +86-592-2189879

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Highlights 

α-, β- and γ-HBCDs caused developmental delay in T. japonicus.



α-, β- and γ-HBCDs had similar developmental toxicity in T. japonicus.

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α- and β-HBCDs had higher lethal toxicity in T. japonicus than γ-HBCD.

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The BCF of HBCDs was in the order of α-HBCD>γ-HBCD>β-HBCD in T. japonicus.



α- and β-HBCDs had higher potency to induced oxidative stress than γ-HBCD.

Abstract The three major hexabromocyclododecane (HBCD) diastereoisomers, i.e. α-, β- and γ-HBCD, have distinct physical and chemical properties that may potentially result in different levels of bioaccumulation and toxicity in aquatic organisms. To assess the impact of diastereomeric variation in HBCDs, the marine copepod Tigriopus japonicus was exposed to α-, β- and γ-HBCD in isolation. Results showed that all the three diastereoisomers had a similar potency to cause growth delay in T. japonicas. Variation was observed in the overall survival rate with exposure to α- and β-HBCD, and this resulted in significantly higher lethal toxicity in T. japonicas than that with exposure to γ-HBCD. Exposure to α-, β- and γ-HBCD led to the generation of ROS in T. japonicas, a possibly toxic mechanism. Both α- and β-HBCD showed a higher potential to induce oxidative stress, which may be a factor in the higher lethal toxicity observed with α- and β-HBCD exposure. It is of note that T. japonicus was found to be more sensitive to all three diastereoisomers in the F1 generation than in the F0 generation. The bioconcentration potential of HBCD diastereoisomers can be ranked in the order α-HBCD>γ-HBCD>β-HBCD and was found to be higher in T. japonicus than has been reported for fish species.

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Introduction Hexabromocyclododecanes (HBCDs) are primarily used in polystyrene insulation foam boards, textile and electronic products, being the third most widely used brominated flame retardants world-wide (Yang et al., 2012). HBCDs are now ubiquitously occurring environmental contaminants, having been listed for consideration for global elimination by the 2013 Stockholm Convention on Persistent Organic Pollutants (www.pops.int). Dissolved HBCD concentrations in the range of ng L-1 have been detected in water bodies (He et al., 2013; Wu et al., 2010). For example, levels as high as 2100 ng L-1 total HBCDs have been detected in the Kuzuryu river (Japan), which is heavily contaminated with effluents from local textile industries (Oh et al., 2014). HBCDs accumulate in sediments, due to their high octanol-water partition coefficient (Kow, 5.625), with sediment in the Kuzuryu river containing up to 7800 μg kg-1 d.w. (dry weight) HBCDs (Oh et al., 2014). Critically, the high Kow value also results in significant HBCD accumulation in aquatic organisms, resulting in trophic transfer and contamination throughout the ecosystem. In field survey, the measurement of HBCDs in lower trophic level aquatic organisms, such as crustaceans or other invertebrates, is very limited (Son et al., 2015; Tomy et al., 2004; Tomy et al., 2008), although high concentrations of HBCDs have been frequently detected in several freshwater and marine fish species collected from highly industrialized areas of Europe and China (Allchin and Morris, 2003; Koppen et al., 2010; Xian et al., 2008). Nevertheless, laboratory studies have shown that bioconcentration factors (BCFs) of HBCDs were higher in the copepod Tigriopus japonicus than in the fish Cyprinus carpio specularis (mirror carp) (Shi et al., 2017; Zhang et al., 2014). Therefore, it is likely that HBCDs 3

may be accumulated through water exposure and cause adverse impacts on aquatic invertebrates. Commercially available HBCD, also known as technical HBCD (tHBCD), is a mixture of mainly three diastereoisomers (i.e., α-, β- and γ-HBCD) with varying compositions of tHBCD ranging from 1-12% α-HBCD, 10-13% β-HBCD, and 75-89% γ-HBCD (Koeppen et al., 2007).Variations in the stereo-structural characteristics of these HBCDs result in significant differences in their physical and chemical properties (e.g., polarity, solubility and Kow), which may lead to distinctive environmental fates (MacGregor and Nixon, 2004). Laboratory analysis has shown that the BCF and biomagnification factor (BMF) of α-HBCD in fish species (zebrafish Danio rerio or mirror carp Cyprinus carpio specularis) are much higher than those of β- and γ-HBCD (Du et al., 2012a; Zhang et al.,2014). In addition, it was found that β- and γ-HBCD can be metabolized to form α-HBCD in adult fish (Du et al., 2012a; Law et al., 2006a; Zhang et al., 2014). Therefore, α-HBCD is a predominant HBCD diastereoisomer found in both freshwater and marine fish samples, usually accounting for ~80-90% of the total HBCD load (He et al., 2013; Law et al., 2006b; Li et al., 2012; Son et al., 2015; Xia et al., 2011). Conversely, in crustaceans collected from both freshwater and marine environments, the concentrations of α-HBCD were often close to γ-HBCD concentrations (Son et al., 2015; Tomy et al., 2004; Tomy et al., 2008). It is of note that the difference in bioaccumulation potential for the three aforementioned diastereoisomers in species other than fish is still unknown. Toxicological studies on tHBCD show that exposure induces the generation of reactive oxygen species (ROS) in aquatic organisms, causing higher malformation rates during embryonic development (Deng et al., 2009; Du et al., 2012b; Hong et al., 2014; Shi et al., 2017). However, the toxicity of individual HBCD diastereoisomers, an essential aspect in the risk assessment of 4

HBCD exposure, remains largely unknown. Du et al. (2012b) found that in freshwater Danio rerio the developmental toxicity of the three HBCD diastereoisomers could be ranked in the order: γ-HBCD>β-HBCD>α-HBCD. Based on these results, the metabolism of β- or γ-HBCD into α-HBCD in fish species could be considered a detoxification process, which would suggest that performing a risk assessment based on toxicity data from tHBCD exposure may overestimate the risk of HBCDs. Alternatively, in the marine fish Oryzias melastigma, α-, β-and γ-HBCD were found to have a similar ability to induce oxidative stress, accordingly resulting in similar developmental toxicity (Hong et al., 2015). At present, little comparison has been made of the toxicity of individual HBCD diastereoisomers to non-fish aquatic species. Considering the notable inter-species variation, in order to evaluate the risk of HBCDs to aquatic, particularly marine ecosystems, it is necessary to evaluate the bioaccumulation potential and toxicity of individual HBCD diastereoisomers on non-fish species, to reduce uncertainty and better understand full ecosystem effects of HBCD exposure. Marine copepods are important crustaceans that contribute significantly to the flow of vital nutrients within marine ecosystems. In the present study, we selected the marine copepod T. japonicus as a model species for the comparison of toxicity of three HBCD diastereoisomers, as it is a widely distributed marine copepod species, has a reasonable life-cycle duration and exhibits distinct sexual dimorphism characteristics with both nauplius and copepodid phases. In addition, T. japonicus can adapt to a wide range of temperature and salinity environments, therefore making it a useful and established model for estuarine and marine ecotoxicological studies (Guo et al., 2012; Lee et al., 2008a; Raisuddin et al., 2007). Here we assessed the bioaccumulation and toxicity of α-, β-, and γ-HBCD on T. japonicus through water-borne exposure, with the transcription of genes 5

related to oxidative stress and apoptosis examined, to compare the toxic mechanisms of exposure to HBCD diastereoisomers in marine copepods.

Materials and Method Chemicals and reagents. α-, β- and γ-HBCD used for exposure were purified from tHBCD as previously described by Hong et al. (2015), with a purity of >98%. The analytical standards for α-HBCD, β-HBCD and γ-HBCD were obtained from AccuStandard, Inc. (New Haven, Connecticut, USA). Stable isotope labeled [13C]-α-HBCD, [13C]-β-HBCD and [13C]-γ-HBCD were obtained from Cambridge Isotope Laboratories, Inc. (Tewksbury, MA, USA). The organic solvents acetonitrile, acetone, hexane and methanol used for sample extraction and liquid chromatography-tandem mass spectrometry (LC-MS/MS) analysis, were all of analytical grade (purity >99.5%) and were purchased from Tedia Inc. (Fairfield, OH, USA). All other reagents were of analytical grade and were obtained from Sigma-Aldrich Chemical Co. (Saint Louis, MO, USA). T. japonicus maintenance. Copepod T. japonicas populations were maintained at 20 ± 1 °C in a 12:12 h light :dark cycle in 0.22 µm-filtered artificial seawater at a salinity of 30‰. An equal mixture of four algal species, i.e., Thalassiosira pseudonana, Isochrysis galbana, Tetraselmis suecica and Nannochloropsis oceanica, were used as daily feed. All other copepod exposure conditions were the same as described for maintenance conditions, except where indicated. Two generation full life-cycle tests. The exposure concentrations utilized for two-generation full life-cycle toxicity test for all three diastereoisomers tested were 0, 8, 30, 80, 300 and 800 µg L-1, with five replicates included in each exposure group. The control group was exposed to solvent 6

only conditions, which comprised of DMSO at the same concentration as exposed to experimental groups (0.01%, v/v). Exposure protocols have been detailed fully previously (Shi et al., 2017). In brief, the first-generation life-cycle test (F0) was performed with newly hatched nauplii (<24 h post-hatch) in a freshly prepared exposure media, with over 80% of the total volume renewed daily and nauplii cultured until adult females developed egg sacs. The green alga T. suecica was used as a daily feed at a density of 1.3 × 105 cells mL-1, added approximately 2 h before media renewal. Copepod mortality and progression through developmental stages were assessed daily using a stereomicroscope, with mortality rate (%) calculated along with the time for newly hatched nauplii to develop to the C1 copepodid stage (N-C) and to adults bearing egg sacs (N-A) (Lee et al., 2008a; Raisuddin et al., 2007). Five egg-sac bearing female copepods were selected from each exposure group, and were individually transferred to glass beakers and cultured for 10 days to assess both the number of clutches and the number of nauplii per clutch as a measure of fecundity. For the second-generation life-cycle test (F1 test), the experimental exposure conditions were the same as those described for the F0 generation, with the exception that developmental and fecundity tests and subsequent toxicity evaluation were not performed on the F1 generation for exposure concentrations causing over 50% mortality in the test population in the F0 generation. The lowest observable effect concentrations (LOECs) were determined as the lowest nominal exposure concentrations that caused significant effects (e.g. mortality and developmental delay) on T. japonicas, while the no observable effect concentrations (NOECs) were determined as the maximum nominal concentrations in the exposure that did not cause a significant effect. Exposure for bioaccumulation and depuration study. The experiment was conducted in 2-L 7

glass tanks, with three replicates included for each exposure group. The uptake phase lasted for 96 h followed by a 48-h depuration period. In the uptake phase, T. japonicus was exposed to a nominal concentration of 2 µg L-1 of either α-, β- or γ-HBCD in seawater, with a solvent only control group maintained without any addition of HBCDs. The selected exposure concentration was below the water solubility of the three individual HBCD diastereoisomers, i.e., 48.8, 14.7 and 2.1 µg L-1 for α-, β- and γ-HBCD, respectively (MacGregor and Nixon, 2004). To maintain the concentration of HBCDs at 2 µg L-1, the walls of the glass tanks were pre-equilibrated to the same concentration of exposure solution, and the exposure media were freshly prepared and renewed every 12 h until depuration which was performed in clean artificial seawater. During the whole exposure period, T. suecica was centrifuged to remove culture media and was added at a density of 5 × 105 cells mL-1 daily, 1 h before media renewal. Sampling was performed at 0, 6, 12, 24, 48, 72 and 96 h during the uptake phase and at 102, 108, 120 and 144 h during the depuration phase (Fig. 3). At each sampling time-point, 100 copepods were collected from each test group and frozen immediately in liquid nitrogen and stored at -80ºC until further analysis. In addition, 20 mL of exposure solution was collected for the measurement of HBCD concentrations in the media. Sample extraction and LC-MS/MS quantification. To determine the actual concentrations of HBCD diastereoisomers present in the exposure media, 10 mL of exposure media was sampled for solid phase extraction using a CNW® HC-C18 SPE column (CNW Technologies, Shanghai, China), which had been preconditioned with 20 mL of methanol followed by 20 mL of water. Columns were washed with 10 mL H2O for desalting, and HBCDs were eluted with 9 mL methanol and the eluate was concentrated to 0.5 mL under a steady N2 flow. To determine the concentration of HBCD diastereoisomers accumulated by T. japonicus, 8

copepod samples from different time-points within each exposure group were homogenized and extracted using a solution of hexane: acetone (1:1, v/v). HBCDs concentrations in extracts were analyzed by LC-MS/MS following the methods described previously by Shi et al. (2017). In brief, C-labeled α-HBCD, β-HBCD and γ-HBCD were spiked into samples as internal standards, with

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samples homogenized and extracted twice in 5 mL hexane: acetone (1:1, v/v) solution using ultrasonication for 30 min. The resulting supernatants were combined and concentrated to 0.5 mL under a steady N2 flow. LC-MS/MS analysis was conducted using an Agilent 1290-6490 UPLC-triple quadruple mass spectrometry system (Agilent Technologies, Palo Alto, California, USA), equipped with an electro-spray ionization probe operated in negative ion-mode. The separation of α-, β- and γ-HBCD isomers was achieved using an Agilent XDB C18 HPLC column (2.1 mm × 150 mm, particle size 3.5 µm) at a flow rate of 0.25 mL min-1. The mobile phase consisted of two solvents: mobile phase A (30% methanol in water, v/v) and mobile phase B (30% methanol in acetonitrile, v/v), with gradient elution used for HPLC separation as follows: linear gradient from 70 to 100% B (0-10 min); isocratic flow at 100% B (10-15 min); and re-equilibrate at 70% B (10 min). α-, βand γ-HBCD isomers and their corresponding 13C-labeled internal standards were monitored in multiple reaction monitoring mode with transition events at m/z 640.8 > 80.8 and m/z 652.8 > 80.8, respectively. A 5-point calibration curve was generated (20, 50, 100, 200, 500 ng mL-1) to confirm the linearity of response of the mass spectrometer (R2> 0.99). None of the HBCD stereoisomers were detectable in procedural blanks, which were included in each experimental batch to confirm that no contamination was introduced via sample bottles or organic solvents. The recoveries of α-, β9

and γ-HBCD were found to be 86%, 83%, and 95%, respectively, in the spiked blanks of seawater media and 88%, 91% and 93%, respectively, in the spiked blanks of copepod samples. Exposure induced gene transcription analysis. To explore the mechanisms of HBCD toxicity, we conducted an isolated exposure experiment where adult T. japonicus were exposed to 0, 30 and 100 µg L-1 of either α-, β- or γ-HBCD. Each exposure group was cultured in a glass tank containing 1 L of exposure media, and three replicates were included for each group. The solvent control group received the same concentration of DMSO as the treated groups (0.01%, v/v). The exposure media were freshly prepared and renewed every 24 h, and samples were collected following 5-day exposure with 120 copepods per sample, frozen immediately in liquid nitrogen and stored at -80ºC until further analysis. Real-time quantitative PCR (qPCR). Total RNA was extracted, with an equal amount of RNA reverse-transcribed using mixtures of oligo (dT) primer (5’-TTTTTTTTTTTTTTTTTTTT-3′) and random primers using M-MLV reverse transcriptase (BGI, Shenzhen, China) to generate cDNA. qPCR was carried out using a CFX96TM Real-Time System (Bio-Rad Laboratories), with the sequence of qPCR primers for the genes superoxide dismutase (SOD), P53, 8-oxoguanine DNA glycosylase (OGG1), corticotrophin-releasing hormone binding protein (CRH-BP), ecdysone receptor (EcR) and vitellogenin (VTG) and Actin, based on previously published papers (Hwang et al., 2010a; Kim et al., 2011; Kim et al., 2012; Lee et al., 2008b). Among them, SOD and OGG1 are considered as oxidative stress biomarker genes, P53 plays an important role in both apoptosis and cell cycle arrest, and CRH-BP, EcR and VTG are endocrine related genes (Ravanat et al., 2002; Harris and Levine, 2005; Morgan and Liu, 2011; Kim et al., 2011; Kim et al., 2012). The thermal cycle program consisted of an initial denaturation step at 95°C for 3 min, followed by 50 10

cycles of 95°C for 10 s and 65°C for 35 s. Dissociation curve analysis was performed to confirm that only targeted PCR products were amplified and detected, with two qPCR analysis utilized for each sample as technical replicates. The transcription levels of the tested genes were analyzed using the 2−ΔΔCt method and normalized according to concentrations of Actin mRNA. Data analysis. To determine the bioaccumulation and depuration parameters, as described in Shi et al. (2017), results were fitted to the following equations (Eq (1) -(6)): Ct = C0e−k2t

(1)

Data obtained during the depuration phase were fitted into Eq (1), where Ct is the concentration of HBCD diastereoisomers in copepod at time t; C0 is the concentration of HBCD diastereoisomers in copepods at the start of the depuration phase; and k2 represents the depuration rate constant. A linear regression of ln(concentration) versus time was performed, where k2 is the regression line slope. Cf(t) = Cw(t)

k1 k2

(1 − e−k2t ) (2)

Data obtained during the uptake phase were fitted into Eq (2), where k2 is the depuration rate constant obtained by Eq (1); k1 represents the uptake rate constant; and Cf(t) and Cw(t) are the HBCD concentrations in copepods and exposure media, respectively, at time t. Concentrations vs. time during the uptake phase were regressed using non-linear curve fitting to obtain k1. The BCFk values (based on a kinetic model) and BCFss values (based on steady-state concentrations) were calculated using Eq (3) and Eq (4), respectively: BCFk = BCFss =

k1 k2 Css Cw

(3) (4)

where Css is the HBCD concentration in copepods after reaching a steady state; and Cw is the 11

HBCD concentrations in the exposure media. The lipid normalized kinetic BCF (BCFkL) was calculated using Eq (5), which was expressed on a 5% lipid content basis (OECD 305): BCFkL =

0.05 Ln

BCFk (5)

where Ln is the mean lipid fraction (wet weight), which has previously been determined to be 5.6% in T. japonicus (Shi et al., 2017). The depuration half-life (t1/2) was calculated using Eq (6): t1/2 =

ln2 k2

(6)

Statistical analysis was conducted using Sigma Plot v.12.5 (Systat Software, Inc.) or SPSS Statistics v.19.0 (IBM software, Inc.). Data were first tested for normality using the Shapiro-Wilk test and then for homogeneity-of-variance by Levene test. After passing the normality test and homogeneity-of-variance test, significance differences were analyzed by one-way analysis of variance (ANOVA) followed by an LSD test.

Results Mortality caused by HBCD diastereoisomers. Exposure of T. japonicus to γ-HBCD at concentrations of up to 800 µg L-1 did not result in significant mortality in either the F0 or F1 generations (Table 1), while α- and β- HBCD were found to induce significant lethal toxicity. Exposure to 80, 300 and 800 µg L-1 α-HBCD gave rise to a cumulative mortality of 23%, 76% and 100%, respectively, by the end of the F0 toxicity test period (Fig. 1A), with a faster onset of mortality observed at higher exposure doses (Fig. 1A). Due to the high mortality rate, the 300 and 12

800 µg L-1 α-HBCD exposure groups were discontinued for all subsequent F1 testing. F1 generation exposed to 80 µg L-1 α-HBCD resulted in 43% cumulative mortality rate, a notable increase from the F0 generation mortality rate at the same concentration (Fig. 1B). The lethal toxicity of β-HBCD was found to be higher than observed with α-HBCD exposure. At concentrations as low as 30 µg L-1, β-HBCD exposure resulted in 41% cumulative mortality in the F0 generation and the mortality increased to 62% in the F1 generation at the same exposure concentration (Fig. 1C-D). As the survival rates observed following exposure to 80, 300 and 800 µg L-1 β-HBCDs were lower than 50% in the F0 generation, particularly the two highest doses (300 and 800 µg L-1) causing 100% mortality within 3 days of exposure, the exposure was discontinued for the F1 generation test (Fig. 1C). The actual concentrations of α-, β-, and γ-HBCD HBCD in the daily freshly prepared exposure media were sampled and measured three times during the whole exposure period, and the results showed that they were very close (within ±15%) to the nominal concentrations (Table S1). Developmental and reproductive toxicity of HBCD diastereoisomers. All the three HBCD diastereoisomers resulted in a dose dependent increase in the duration of the developmental period in T. japonicas from nauplii to both copepodid (N-C) and adult (N-A) phase. Such an effect was observed in both F0 and F1 generations (Fig. 2). The LOECs observed for α- and γ-HBCD induced developmental delay in both the N-C and N-A phases were 30 µg L-1 in F0 generation and 8 µg L-1 in F1 generation (Fig. 2 and Table 1), showing notably increased sensitivity to both α- and γ-HBCD among the F1 generation. β-HBCD exposure resulted in high levels of mortality, and its effects on developmental duration were thus evaluated only at the doses of 0, 8 and 30 µg L-1. The 13

LOEC established for β-HBCD induced developmental delay during N-A phase was 30 µg L-1 in the F0 and F1generations. Therefore, all three diastereoisomers presented a similar ability to delay the development of T. japonicus. No significant difference was observed in the reproductive characteristics of T. japonicus, including sex ratio, the number of nauplii per clutch and the total nauplii brooded over ten days (fecundity) after exposure to α-, β- or γ-HBCD (Table S2). Bioaccumulation and depuration of HBCD diastereoisomers. At the nominal concentration of 2 µg L-1, the average measured concentrations of α-, β-and γ-HBCD in the exposure media during the uptake phase was 1.34±0.14, 1.26±0.13 and 1.45±0.27 µg L-1, respectively. The measured media concentrations deviated within 20% of the mean values (Fig. 3), suggesting that the exposure concentrations were relatively stable. In the control groups, the concentrations of HBCDs in T. japonicus and exposure media were less than 2% of the values in the exposure groups. During the 96-h uptake phase, the concentrations of each of the three HBCD diastereoisomers in T. japonicus increased over time, approaching saturation by 96 h (Fig. 3). The three uptake and depuration curves established, all followed the models described by Eq (2) for uptake and by Eq(1) for depuration (R12= 0.99, R22> 0.95, respectively; Table 2). α- and γ-HBCD presented with similar uptake rate constants (k1), which were significantly higher than that of β-HBCD (Table 2). The depuration rate constants (k2) did not vary considerably and can be ranked in the order of γ-HBCD>β-HBCD>α-HBCD (Table 2). As a result, the BCFk, calculated as k1/ k2, was in the order of α-HBCD>γ-HBCD>β-HBCD. In addition, the values of BCFss also followed the same order of ranking as BCFk, confirming that overall α-HBCD had the highest 14

bioconcentration potential, followed by γ-HBCD and then β-HBCD. Particularly, the values of BCFss were significantly different among the three diastereoisomers (Table 2). Gene transcription induced by HBCD diastereoisomers. Following exposure to 30 or 100 µg L-1 of either α- or β-HBCD for 5 days, the transcription of the oxidative stress biomarker gene SOD in T. japonicus was significantly up-regulated in a dose dependent manner, while no significant increase in the induction of SOD by γ-HBCD was observed (Fig. 4A). The transcription of OGG1, a gene responsible for the repair of DNA oxidative damage, along with the P53 gene, which plays an important role in DNA repair and apoptosis, were both up-regulated upon exposure to all three HBCD diastereoisomers (Fig. 4B-C). In particular, at the higher dose of 100 µg L-1, the induction of OGG1 and P53 transcription was higher with α- and β-HBCD exposure than with γ-HBCD exposure. α-, β- and γ-HBCD also induced the transcription of CRH-BP, the binding protein for corticotrophin-releasing hormone (CRH), which is important in the modulation of CRH activity (Fig. 4D). Similarly, the fold-change observed in CRH-BP transcription was more significant in the α- and β-HBCD exposure groups than in the γ-HBCD exposure group. None of the three diastereoisomers significantly induced the transcription of EcR and VTG at the dose of 30 or 100 µg L-1 (Fig. 4E-F). Discussion In previous studies tHBCD exposure resulted in developmental delay, without causing an increase in mortality in T. japonicus (Shi et al.,2017). γ-HBCD is the predominant diastereoisomer in the tHBCD mixture and there was no distinct bioconversion observed among HBCD diastereoisomer variants in T. japonicus (Shi et al., 2017), suggesting that tHBCD toxicity can be largely attributed to the γ-HBCD content. Consistently, in the present study we found that like 15

tHBCD γ-HBCD did not induced significant mortality. In addition, the concentrations, at which γ-HBCD induced growth delay in T. japonicus during both N-C and N-A phases, were in the same range as those of tHBCD (Shi et al., 2017). When comparing the toxicity of the three diastereoisomers tested, α-, β- and γ-HBCD showed a similar potency to induce the generation of ROS and increase malformation rate during embryonic development in the marine fish Oryzias melastigma (Hong et al., 2015). In contrast, in the freshwater zebrafish Danio rerio, variations in the developmental toxicity of diastereoisomers were observed, with the ability to induce ROS and developmental malformation ranked as: γ-HBCD>β-HBCD>α-HBCD (Du et al., 2012b). In the present study, the ranking of potency for the induction of developmental delay by α-, β- and γ-HBCD in the marine copepod T. japonicus was the same as that established in the marine fish O. melastigma. In addition, the LOECs for α-, β- and γ-HBCD induced developmental delay was 8 or 30 µg L-1, with the F1 generation being more sensitive to HBCDs than the F0 generation (Table 1). These LOECs were in the range of environmentally realistic concentrations (e.g., up to 2.1 µg L-1 HBCDs being detected in the Kuzuryu river)(Oh et al., 2014). Therefore, environmentally realistic concentrations may pose a significant threat to the marine copepod T. japonicus. Our study demonstrates that α- and β-HBCD caused considerably higher rates of lethal toxicity than γ-HBCD in T. japonicus, with the order of ranking being β-HBCD>α-HBCD»γ-HBCD, which is different from what has been shown for zebra fish and marine medaka in embryo-juvenile stages (Hong et al., 2015; Du et al., 2012b). Therefore, variations in the order of toxicity of HBCD diastereoisomers, depending on different test species (e.g., fish vs. copepod) and toxic endpoints monitored, show varying mechanisms of action 16

relative to the toxicity model, which should be taken into consideration when evaluating the effects of HBCDs on aquatic ecosystems. The three HBCD diastereoisomers tested showed higher BCFs in T. japonicus, based on both steady state and kinetic methods, than in the mirror carp fish (Zhang et al., 2014), which may be attributed to the higher uptake rate constant (k1) observed in T. japonicus. As marine copepods are suspension feeders, uptake of HBCDs from the surrounding seawater (i.e., bioconcentration) is likely an important exposure pathway for them, as seen with species such as Mysis and blue mussels (Haukåset al., 2010; Tomy et al., 2004). It is thus essential to determine the bioconcentration potential (i.e., BCFs) in suspension feeding marine invertebrates. Although laboratory studies have shown that mirror carp fish has lower BCFs values than T. japonicus (Zhang et al., 2014), fish usually contain higher concentration of HBCDs than invertebrates in marine ecosystems (Haukåset al., 2010; Tomy et al., 2004; Son et al., 2015), which is likely due to biomagnification of HBCDs via ingestion as shown in a variety of fish species (Son et al., 2015; Tomy et al., 2004). Nevertheless, our study demonstrates that bioconcentration of HBCDs into invertebrates through water exposure is an important pathway for HBCDs being bioaccumulated in aquatic ecosystems. The BCFs, particularly the BCFss, of the three diastereoisomers in T. japonicas were found to be statistically different (p<0.05, Table 2) and can be ranked in the order of α-HBCD>γ-HBCD>β-HBCD (Table 2). The higher BCF for α-HBCD than that for γ-HBCD may be responsible for the similar fractions of α-HBCD and γ-HBCD in the total HBCD content and for the much higher ratios of α-HBCD/γ-HBCD than that in tHBCD mixture, as observed in marine crustacea (Tomy et al., 2004; Haukås et al., 2010; Son et al., 2015), despite γ-HBCD 17

constituting the majority of tHBCD (Koeppen et al., 2007). Additionally, as β-HBCD only exists as a minor fraction in tHBCD (Koeppen et al., 2007) and its BCF value was much lower than those of α- and γ-HBCD (Table 2), β-HBCD was very often not detected in marine invertebrates (Haukås et al., 2010; Sonet al., 2015). Considering the higher bioaccumulation potential and much higher lethal toxicity of α-HBCD than γ-HBCD in T. japonicus, it is very important to evaluate the ecotoxicological risk of individual HBCD diastereoisomers separately. The highest (i.e., α-HBCD ) and the lowest (i.e., β-HBCD ) BCF values of the three diastereoisomers only varied in by 2 folds in T .japonicus. However, the BCF of α-HBCD in the mirror carp was reported to be 10-fold higher than the BCF values for β-HBCD or γ-HBCD (Zhang et al., 2014). It has been well established that β-HBCD and γ-HBCD can be bio-isomerized into α-HBCD in the livers of adult zebrafish, mirror carp and juvenile rainbow trout after long term exposure (>20 days) (Du et al., 2012a; Law et al., 2006a; Zhang et al., 2014). Moreover, α-HBCD is known to be more resistant to metabolic degradation in vitro by adult rainbow trout liver enzymes, when an NADPH regenerating system is also present (Abdallah et al., 2014). Therefore, biotransformation may contribute significantly to the much higher BCF value found for α-HBCD than for β-HBCD or γ-HBCD in fish species. In contrast, the present study showed that in each of the α-, β- or γ-HBCD exposure group, only the specific HBCD diastereoisomer utilized for exposure was detected in T. japonicus (data not shown), suggesting no apparent bioconversion among the three diastereoisomers, which may be the main reason that the BCF of α-HBCD was only slightly greater than that of β-HBCD or γ-HBCD in T. japonicus. Moreover, no HBCD metabolites (e.g., mono-hydroxy-HBCD and di-hydroxy-HBCD) were observed in marine crustacea, whereas they have been observed in fish species (Son et al., 2015). 18

Therefore, species-specific metabolism and bioconversion activity could result in significantly different bioaccumulation potential among HBCD diastereoisomers. Previous studies have reported that tHBCD induces the transcription of oxidative stress biomarker genes SOD, glutathione S-transferase(GST) and catalase (CAT) (Shi et al., 2017). In addition, tHBCD induces the OGG1 gene, which encodes 8-oxoguanine DNA glycosylase, a DNA base-excision repair protein which repairs DNA oxidative damages caused by ROS (Ravanat et al., 2002; Shi et al., 2017). Furthermore, increased levels of DNA oxidative damage results in the up-regulation of the P53 gene, promoting cell cycle arrest and mitochondria-mediated apoptotic pathways (Harris and Levine, 2005; Honget al., 2014). In the present study, we found that α-and β-HBCD had a higher potency to induce oxidative stress in T. japonicus than γ-HBCD, as indicated by the up-regulation of SOD, OGG1 and P53 transcription (Fig. 4), which could be a contributing factor to the higher lethality observed with α-and β-HBCD exposure than with γ-HBCD exposure. In addition, CRH-BP, which binds to CHR to modulate the hormones activity, was significantly up-regulated by all HBCD diastereoisomers tested, with notably higher expression observed following α-and β-HBCD exposure than with γ-HBCD exposure (Fig. 4D). CRH can further up-regulate the transcription of nuclear factor kappa B (NF-κB), whose signaling pathway could interfere with ROS generation and cell death regulation via various mechanisms (Smith et al., 2006; Morgan and Liu, 2011). Therefore, exposure to HBCDs can also reduce bioactive CRH concentrations by increasing the expression of CRH-BP, thereby causing developmental toxicity. Other than the genes presented and discussed above, the three HBCD diastereoisomers did not affect the transcription of other endocrine related genes monitored, e.g., ecdysone receptor (EcR) and vitellogenin (Fig. 4E-F). EcR plays an important role in copepod 19

development and molting (Hwanget al.,2010b), while vitellogenin is an important indicator of the activation of estrogen receptors. It is also possible that α-and β-HBCD exposure may result in high mortality in T. japonicus due to other toxic mechanisms in combination with those discussed here. To investigate this, a proteomic or transcriptomic approach should be applied, allowing better understanding of the molecular mechanisms of the toxicity caused by individual HBCD diastereoisomers. Conclusions. The three major HBCD diastereoisomers, α-, β- and γ-HBCD presented a similar potential to induce developmental toxicity and were found to cause significant developmental delay in T. japonicus. Importantly, exposure to α- and β-HBCD stereoisomers resulted in considerably higher lethality than γ-HBCD exposure. In addition, the BCFs of the three HBCD diastereoisomers were found to be ranked in the order α-HBCD>γ-HBCD>β-HBCD in T. japonicas, notably higher than the BCFs reported in mirror carp fish. Exposure to α-, β- and γ-HBCD all induced ROS generation in T. japonicas, which may be a causative factor in the induction of toxicity. In particular, α- and β-HBCD stereoisomers showed a greater potential to induce oxidative stress, which may partially explain the higher lethality observed with exposure to them, although further investigation is required to investigate other toxic mechanisms involved. Acknowledgements This study was supported by the National Natural Science Foundation of China (No. 41206090), the Natural Science Foundation of Fujian Province, China (No. 2015J01174), and the Recruitment Program of Global Youth Experts.

20

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Xia, C.H., Lam, J.C.W., Wu, X.G., Sun, L.G., Xie, Z.Q., Lam, P.K.S., 2011. Hexabromocyclododecanes (HBCDs) in marine fishes along the Chinese coastline. Chemosphere 82, 1662-1668. Xian, Q., Ramu, K., Isobe, T., Sudaryanto, A., Liu, X., Gao, Z., Takahashi, S., Yu, H., Tanabe, S., 2008. Levels and body distribution of polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecanes (HBCDs) in freshwater fishes from the Yangtze River, China. Chemosphere 71, 268-276. Yang, R., Wei, H., Guo, J., Li, A., 2012. Emerging brominated flame retardants in the sediment of the Great Lakes. Environ. Sci. Technol. 46, 3119-3126. Zhang, Y., Sun, H., Ruan, Y., 2014. Enantiomer-specific accumulation, depuration, metabolization and isomerization of hexabromocyclododecane (HBCD) diastereomers in mirror carp from water. J. Hazard. Mater. 264, 8-15.

24

Figure captions Fig 1. Cumulative mortality rate in T. japonicas exposed to α-HBCD (A and B) and β-HBCD (C and D) in the F0 generation (A and C) and F1 generation (B and D) life-cycle toxicity tests, performed at exposure concentrations of 0, 8, 30, 80, 300 and 800 µg L-1. Exposure concentrations that induced over 50% cumulative mortality in the F0 generation (i.e., 300 and 800 µg L-1 α-HBCD, and 80, 300 and 800 µg L-1 β-HBCD) were discontinued for the F1 generation test. Values are presented as mean ± SD (n = 5). Fig 2. Effects of α-HBCD (A and B), β-HBCD (C and D) and γ-HBCD (E and F) exposure on the developmental duration of N-C phase (from newly hatched nauplius to C1 copepodid stage) and N-A phase (the whole maturation period, from newly hatched nauplius to adult bearing egg sacs) in T. japonicas (Raisuddin et al., 2007), in the F0 (A, C, E) and F1 generations (B, D, F). α-, βand γ-HBCD concentrations of 0, 8, 30, 80, 300 and 800 µg L-1 were used for exposure. In the exposure groups with concentrations that caused cumulative mortality higher than 50% (i.e., 300 and 800 µg L-1 α-HBCD, and 80, 300 and 800 µg L-1 β-HBCD), developmental durations were not calculated and exposure was discontinued for the F1generation. Values are presented as the mean ± SD (n = 5), with significant differences from the controls indicated by asterisks (one-way ANOVA, followed by the LSD post hoc test: * p< 0.05, ** p< 0.01, *** p< 0.001). Fig 3. Concentrations of HBCDs detected in T. japonicus (left y-axis) and exposure media (right y-axis), when copepods were exposed to 2 µg L-1 α-HBCD (A), β-HBCD(B) or γ-HBCD (C) for 96 h followed by 48 h depuration. Values are presented as the mean ± SD (n = 3). Fig. 4. Induction of gene transcription for SOD (A), OGG1 (B), P53 (C), CRH-BP (D), EcR (E) and VTG (F) in T. japonicus following exposure to 0, 30 or 100 µg L-1 α-, β- or γ-HBCD over a 5-day period. Values are presented as the mean ± SD (n = 3) and significant differences from the controls are indicated by asterisks (one-way ANOVA, followed by the LSD post hoc test: * p< 0.05, ** p< 0.01, *** p< 0.001).

25

Fig 1.

F1

F0

120

120

A

B

100 Cum mulative Mortality (%)

Cum mulative Mortality (%)

100 80 60 40

control 8 μg/L 30 μg/L 80 μg/L 300 μg/L 800 μg/L

80 60 40 20

20

0

0 0

5

10

15

0

20

5

15

20

15

20

Time (days)

Time (days) 120

120

C

D

100 Cumulative Mortality (%))

100 Cumulative Mortality (%)

10

80 60 40

80 60 40 20

20

0

0 0

5

10 Time (days)

15

20

0

5

10 Time (days)

Fig 2. 20

N-C N-A

F0

F1 20

A

**

*

***

15

Days

Days

15

10

10

***

***

*** 5

***

***

5

Survival rate  < 50%

0 8

30

80

300

800

0

20

8

30

80

300

800

20

C

D

*

15

*

Days s

15

10

10

* Survival rate  < 50%

5

5

0

No sufficient  offspring from  F0 for F1  generation test

0 0

8

20

E

30

80

300

***

***

***

800

0

***

20

**

15

8

30

***

F

80

**

300

800

***

***

15

10

***

***

***

***

Days

Days

No sufficient  offspring from  F0 for F1  generation test

0 0

Days s

***

B

***

10

**

5

***

***

***

***

5

0

0 0

8

30

80

300

Concentration (μg L-1)

800

0

8

30

80

300

Concentration (μg L-1)

800

Fig 3. Uptake by T. japonicus Depuration by T. japonicus Concentration in media 4

120

3

100 80

2 60 40

1

20 0 0

20

40

60

100

120

140

0 160 4

B 80

3

60 2 40 1

20

0 0

20

40

60

80

100

120

140

Time (h)

120

β-HBCD in media (μg L-1)

1 ww) β-HBCD in T. japonicus (ng mg-1

80 Time (h)

100

γ-HBCD in T. japonicus (ng mg-1 ww)

α-HB BCD in media (μg L-1)

A

0 160 4

C 100 3 80 2

60 40

1 20 0 0

20

40

60

80 Time (h)

100

120

140

0 160

γ-HBCD in media (μg g L-1)

α-HBCD in n T. japonicus (ng mg-1 ww)

140

Fig 4.

4

6 5

**

3

Relative change R

Relative change

A  SOD **

2

*

1

*

B  OGG1

control 30 μg/L 100 μg/L

**

4 3

*

**

2 1

0

4

C  P53 *

3

*

**

2

1

0

3

* *

2

*** *

1

0

3

3

E  EcR

F  VTG Relattive change

Rela ative change

D  CHR‐BP

** Relative change e

Relative change e

4

0

2

1

0

2

1

0

control

α‐HBCD β‐HBCD γ‐HBCD

control

α‐HBCD β‐HBCD γ‐HBCD

Table 1. The no observable effect concentrations (NOECs) and lowest observable effect concentrations (LOECs) of mortality and developmental delay in N-C and N-A phases in T. japonicus with exposure to α-, β- and γ-HBCD in the two generation (F0 and F1) toxicity assays. Developmental delay Mortality N-C

N-A

NOEC (µg L-1)

LOEC (µg L-1)

NOEC (µg L-1)

LOEC (µg L-1)

NOEC (µg L-1)

LOEC (µg L-1)

F0

30

80

8

30

8

30

F1

30

80

<8

8

<8

8

F0

8

30

>30

>30

8

30

F1

8

30

8

30

<8

30

F0

>800

>800

8

30

8

30

F1

>800

>800

<8

8

8

30

α-HBCD

β-HBCD

γ-HBCD

26

Table 2. Bioaccumulation and depuration parameters for α-, β- and γ-HBCD in T. japonicas through low-dose water-borne exposure.

R12

k1 -1

R22

-1

(L kg d )

k2

BCFk -1

BCFss

-1

(d )

(L kg )

-1

(L kg )

BCFkL

t1/2

-1

(d)

(L kg )

α-HBCD

0.99

(3.2±0.3)×104

a

0.96

0.19±0.09 a

(1.7±0.2)×105

a

(8.7±0.6)×104

a

(1.5±0.2)×105

a

3.6

β-HBCD

0.99

(2.3±0.2)×104

b

0.97

0.23±0.05 a

(9.8±0.8)×104

b

(5.3±0.3)×104

b

(8.8±0.7)×104

b

3.0

γ-HBCD

0.99

(3.1±0.2)×104

a

0.96

0.26±0.07 a

(1.2±0.1)×105

a,b

(6.5±0.4)×104

c

(1.1±0.1)×105

a,b

2.7

k1, uptake rate constant; k2, depuration rate constant; BCFk, kinetic bioconcentration factor; BCFss, steady-state bioconcentration factor; BCFkL, lipid-normalized BCFk; t1/2, depuration half-time. Values of k1, k1, BCFk, BCFss and BCFkL are presented as the mean ± SD (n = 3). Different letters in superscription indicate significant difference (p<0.05) among the three diastereoisomers (one way ANOVA followed by LSD test).

27