Effects of the brominated flame retardant, TBCO, on development of zebrafish (Danio rerio) embryos
Introduction
Brominated flame retardants (BFRs) are added to industrial and consumer products, including electronics, plastics, insulations, and textiles to inhibit ignition or slow their combustion (Birnbaum and Staskal, 2004). Brominated flame retardants are classified as monomers, reactive, or additive, depending on how they are incorporated into products (Alaee et al., 2003). Monomers are directly incorporated as the products are polymerized, reactive compounds are chemically bound to the products, and additive compounds are blended with the polymer composition of products (Alaee et al., 2003; Harju et al., 2009). Of these types, additive BFRs are most heavily used which is problematic as they are most likely to leach out of products (Harju et al., 2009). Detection of BFRs in municipal wastewater effluents, areas surrounding landfills, domestic workplaces, and in human breast milk has led to concerns about effects on the health of humans and wildlife (Alaee et al., 2003; Harju et al., 2009). Multiple BFRs are present and persist in freshwater and marine ecosystems, where they bioaccumulate in aquatic organisms, including fish (Sühring et al., 2016).
Concern about adverse effects of BFRs on humans and wildlife, coupled with the need to remain amenable with code for product standards, has established a cycle of banning and implementing usage of novel BFRs. Polybrominated diphenyl ethers (PBDEs) were the most commonly used additive BFRs for approximately 30 years until their discovery in aquatic biota, human tissue, and human breast milk resulted in their ban in the European Union (EU) as well as in North America (Alaee et al., 2003; De Wit 2002; Ryan et al., 2002; Sutton et al., 2014). Following the ban of PBDEs, hexabromocyclododecane (HBCD) became the most frequently used additive BFR in the EU and North America, especially for industrial and domestic insulation purposes (De Wit 2002; Law et al., 2014). Early studies demonstrated that HBCD bioaccumulated in aquatic organisms (Sellström et al., 1998). Exposure of early life-stage zebrafish (Danio rerio) to HBCD results in various developmental malformations as well as induction of oxidative stress (Deng et al., 2009; Usenko et al., 2016). Several PBDEs and HBCD are listed in Annex A of the Stockholm Convention.
The novel brominated flame retardant 1,2,5,6-tetrabromocyclooctane (TBCO) is a potential replacement for HBCD. TBCO exists as two diastereoisomers - α and β -TBCO (Riddell et al., 2009). Usage of TBCO is primarily in textiles, paints, and plastics (Danish EPA Report, 1999). Limited information is available regarding environmental concentrations, behaviour, and any adverse effects of TBCO. Despite being detected in eggs from herring gulls ((Larus argentatus) in the North American Great Lakes, concentrations were not quantifiable (Covaci et al., 2011). Concentrations of TBCO were as great as 12 ng/g ww in dab (Limanda limanda) and 1.2 ng/g dw in sediments from freshwater and marine sampling stations in the German Bight and the Elbe river (Sühring et al., 2016). Concentrations of β-TBCO in atmospheric samples from the West Antarctic were as great as 0.21 pg/m3 (Zhao et al., 2020). Exposure of embryos of Japanese medaka (Oryzias latipes) to TBCO at 2.7 ± 0.4, 29.6 ± 7.5, and 151.3 ± 81.9 μg/L resulted in bioaccumulation, and caused decreased hatching rate and success (Sun et al., 2016). Results of a 21-day reproduction assay with Japanese medaka and in vitro studies of sex steroid synthesis and sex steroid receptor signaling demonstrated that TBCO has endocrine-disrupting effects, leading to impairment of reproduction (Mankidy et al., 2014; Saunders et al., 2013, 2015). Using a toxicogenomics approach, it was demonstrated that exposure of early life-stage of Japanese medaka to TBCO at 29.6 ± 7.5 μg/L disrupts transcript and protein abundances in pathways related to cardiac function and vision, and these effects were confirmed by bioassays which found a small reduction in heart rate and impaired visual acuity (Sun et al., 2016).
As usage of TBCO increases, concentrations in aquatic ecosystems are likely to increase. Thus, it is important to understand potential adverse effects of TBCO on aquatic organisms. To date, only one study has investigated embryotoxicity of TBCO (Sun et al., 2016). Given that species of fishes can differ significantly in their sensitivity to toxicants, studies with additional species are warranted. Using zebrafish as a model species, objectives of this study were to determine the acute toxicity of TBCO to early life-stages of zebrafish and to determine mechanisms of toxicity.
Section snippets
Preparation and quantification of TBCO
1,2,5,6-tetrabromocyclooctane (TBCO) was obtained from SynQuest Laboratories (SynQuest Laboratories, Inc. Alachua, FL, USA) and dissolved in 100% acetone (Thermo Scientific, Mississauga, ON, Canada), in a water bath at 60 °C. The exposure solution of greatest nominal concentration (200 μg/L) was prepared by 1:1000 dilution of the stock and diluted to prepare solutions with lesser concentrations. Exposure solutions were prepared in dechlorinated City of Lethbridge water (average water quality:
Embryotoxicity
Exposure to TBCO caused a concentration dependant increase in the cumulative mortality of embryos (Fig. 1). Percentages of mortality of embryos exposed to the freshwater control, acetone control, low TBCO, medium TBCO, and high TBCO was 18.5 ± 4.0, 12.3 ± 4.7, 31.1 ± 9.4, 58.3 ± 16.6, and 71.0 ± 13.4%, respectively (Fig. 1). Of these, the mortality of embryos exposed to the medium and high concentration of TBCO was significantly greater than mortality of embryos exposed to the freshwater
Credit author statement
Darren van Essen: Conceptualization, Investigation, Methodology, Data curation, Writing – original draft. Chloe Devoy: Investigation, Methodology, Data Curation: Justin Miller: Investigation, Methodology, Data Curation. Paul Jones: Investigation, Methodology, Data Curation. Steve Wiseman: Conceptualization, Project administration, Funding Acquisition, Writing – review and editing.
Declaration of competing interest
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.
Acknowledgements
Steve Wiseman was supported by a Tier II Canada Research Chair in Aquatic and Mechanistic Toxicology, infrastructure grants from the Canadian Foundation for Innovation (Project #35224) and Government of Alberta Research Capacity Program, and a Discovery Grant from the Natural Science and Engineering Research Council (NSERC) of Canada (Project #704454). Darren Van Essen was supported by an Undergraduate Research Summer Assistant scholarship from NSERC (NSERC-URSA). The authors would like to
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