Article
pubs.acs.org/est
Bioaccumulation, Biotransformation, and Toxicity of BDE-47, 6‑OHBDE-47, and 6‑MeO-BDE-47 in Early Life-Stages of Zebrafish (Danio
rerio)
Hongling Liu,*,†,# Song Tang,‡,# Xinmei Zheng,† Yuting Zhu,† Zhiyuan Ma,† Chunsheng Liu,†
Markus Hecker,‡,§ David M.V. Saunders,§ John P. Giesy,†,§,∥,⊥ Xiaowei Zhang,† and Hongxia Yu*,†
†
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, Jiangsu
210023, China
‡
School of Environment and Sustainability, University of Saskatchewan, Saskatoon, Saskatchewan S7N 5B3, Canada
§
Toxicology Centre, University of Saskatchewan, Saskatoon, Saskatchewan S7N 5B3, Canada
∥
Department of Veterinary Biomedical Sciences, University of Saskatchewan, Saskatoon, Saskatchewan S7N 5B3, Canada
⊥
Department of Biology and Chemistry, City University of Hong Kong, Kowloon, Hong Kong SAR, China
S Supporting Information
*
ABSTRACT: 2,2′,4,4′-Tetrabromodiphenyl ether (BDE-47),
6-hydroxy-tetrabromodiphenyl ether (6-OH-BDE-47), and 6methoxy-tetrabromodiphenyl ether (6-MeO-BDE-47) are the
most detected congeners of polybrominated diphenyl ethers
(PBDEs), OH-BDEs, and MeO-BDEs, respectively, in aquatic
organisms. Although it has been demonstrated that BDE-47
can interfere with certain endocrine functions that are
mediated through several nuclear hormone receptors (NRs),
most of these findings were from mammalian cell lines exposed
in vitro. In the present study, embryos and larvae of zebrafish
were exposed to BDE-47, 6-OH-BDE-47, and 6-MeO-BDE-47
to compare their accumulation, biotransformation, and
bioconcentration factors (BCF) from 4 to 120 hpf. In addition,
effects on expression of genes associated with eight different
pathways regulated by NRs were investigated at 120 hpf. 6-MeO-BDE-47 was most bioaccumulated and 6-OH-BDE-47, which
was the most potent BDE, was least bioaccumulated. Moreover, the amount of 6-MeO-BDE-47, but not BDE-47, transformed to
6-OH-BDE-47 increased in a time-dependent manner, approximately 0.01%, 0.04%, and 0.08% at 48, 96, and 120 hpf,
respectively. Expression of genes regulated by the aryl hydrocarbon receptor (AhR), estrogen receptor (ER), and
mineralocorticoid receptor (MR) was affected in larvae exposed to 6-OH-BDE-47, whereas genes regulated by AhR, ER, and
the glucocorticoid receptor (GR) were altered in larvae exposed to BDE-47. The greatest effect on expression of genes was
observed in larvae exposed to 6-MeO-BDE-47. Specifically, 6-MeO-BDE-47 affected the expression of genes regulated by AhR,
ER, AR, GR, and thyroid hormone receptor alpha (TRα). These pathways were mostly down-regulated at 2.5 μM. Taken
together, these results demonstrate the importance of usage of an internal dose to assess the toxic effects of PBDEs. BDE-47 and
its analogs elicited distinct effects on expression of genes of different hormone receptor-mediated pathways, which have expanded
the knowledge of different mechanisms of endocrine disrupting effects in aquatic vertebrates. Because some of these homologues
are natural products, assessments of risks of anthropogenic PBDE need to be made against the background of concentrations
from naturally occurring products. Even though PBDEs are being phased out as flame retardants, the natural products remain.
■
INTRODUCTION
environment, has raised concern about its potential adverse
effects to ecosystems and human health.3−5 In addition to the
synthetic BDE-47, its hydroxylated (OH−) or methoxylated
(MeO−) forms, 6-OH-BDE-47 and 6-MeO-BDE-47, have been
Polybrominated diphenyl ethers (PBDEs) have been extensively employed as flame retardants (FRs) in various consumer
and commercial products for decades.1,2 As a result of the
substantial production, long-term use, disposal, and recycling
processes, these chemicals are now frequently found in the
environment.3 The persistence, bioaccumulation potential, and
toxic potency (PBT criteria) of 2,2′,4,4′-tetrabromodiphenyl
ether (BDE-47), one of the primary PBDEs found in the
© 2015 American Chemical Society
Received:
Revised:
Accepted:
Published:
1823
August 6, 2014
December 24, 2014
January 7, 2015
January 7, 2015
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
suggested to be natural products of marine organisms6 and have
been detected in a wide variety of freshwater and marine
organisms including mollusks, mussels, shellfish, clam, fish, seal,
dolphin, and whale.4,7−13 Moreover, it has been conclusively
demonstrated that MeO-BDEs, and not PBDEs, are precursors
of OH-BDEs.6,17−19 In zebrafish (Danio rerio), 6-MeO-BDE-47
can be transformed into 6-OH-BDE-47; however, BDE-47
cannot be transformed into 6-OH-BDE-47.20 In addition,
interconversion between 6-MeO-BDE-47 and 6-OH-BDE-47
has been observed during dietary exposure of Japanese medaka
(Oryzias latipes).17
To date, increasing evidence has shown that exposure to
BDE-47 and its two natural analogs, 6-OH-BDE-47 and 6MeO-BDE-47, can elicit a number of adverse effects in aquatic
organisms including disruption of the endocrine system,21,22
disruption of molting,23 developmental defects,20,24−26 and
neurobehavioral toxicity.27−29 OH- and MeO-BDEs have been
shown to exhibit greater toxic potencies than PBDEs for certain
end-points such as estrogenicity and androgenicity.21,30
However, mechanisms of their toxicity are complex and have
not been fully resolved.31 PBDEs, as well as OH- and MeOBDEs, are structural analogs to thyroid hormones, T3 and T4, as
well as dioxin-like chemicals such as polychlorinated biphenyls
(PCBs), dioxins (TCDD), and furans (PCDF).31 This raises
the question of whether the adverse biological outcomes
resulting from exposure to PBDEs and OH- or MeO-BDEs are
due to their ability to simulate thyroid hormones or whether
they elicit effects similar to those of the above dioxin-like
chemicals.
Nuclear receptors (NRs) are a superfamily of ligandactivated, transcription factors that act globally to regulate a
broad range of biological processes, including development,
reproduction, and metabolism.32,33 NRs mediate signaling by
ligands such as endogenous hormones, lipids, and xenobiotics.34,35 Upon binding of a ligand to the ligand binding
domain of several kinds of NRs, a complex array of cellular
responses is initiated. Recently, several in vivo and in vitro
studies have investigated effects of BDE-47, 6-OH-BDE-47, or
6-MeO-BDE-47 on certain NR mediated physiological pathways, in particular the pathways involving the thyroid hormone
receptor (TR), estrogen receptor (ER), androgen receptor
(AR), and aryl hydrocarbon receptor (AhR). For example, in
adult fathead minnows (Pimephales promelas), dietary exposure
to BDE-47 induced transcription of TRα in the brain of
females, and decreased the transcription of TRβ in the brain of
fish of both sexes.41 In porcine ovarian follicles, both BDE-47
and 6-OH-BDE-47 did not alter expression of AR mRNA or
associated protein, but decreased expression of ERβ mRNA and
protein following exposure to BDE-47 and increase both ERα
and ERβ gene and protein expression following exposure to 6OH-BDE-47.42 In an AhR-responsive luciferase reporter assay,
6-OH-BDE-47 exhibited greater potency to induce AhR activity
than that of 6-MeO-PBDEs and BDE-47.43 However, so far,
most NR studies of BDE-47, 6-OH-BDE-47, and 6-MeO-BDE47 were completed by use of mammalian or cellular assays.
The zebrafish represents an excellent vertebrate model
organism in environmental toxicology studies,44−46 especially
in context with investigating effects of endocrine disrupting
chemicals (EDCs) on reproductive and developmental
systems.47−49 Moreover, developmental profiling of zebrafish
gene expression patterns has confirmed a high degree of
conservation in NR expression patterns between zebrafish and
other vertebrate models.50 Therefore, in the present study,
zebrafish embryos and larvae were used to determine the timecourse of accumulation, biotransformation, and bioconcentration factors (BCFs) of BDE-47, 6-OH-BDE-47, and 6-MeOBDE-47. Additionally, in order to gain a more comprehensive
understanding of the molecular mechanisms of the toxicity of
BDE-47 and related OH- and MeO-analogs on the endocrine
system, their effects on expression of genes associated with
eight nuclear hormone receptor pathways, particularly ER, AR,
AhR, TRα, peroxisome proliferator-activated receptor alpha
(PPARα), glucocorticoid receptor (GR), mineralocorticoid
receptor (MR), and pregnane x receptor (PxR), were
investigated and compared.
■
MATERIALS AND METHODS
Materials and Reagents. BDE-47 (98% purity) was
purchased from Chem Service (West Chester, PA, USA). 6MeO-BDE-47 and 6-OH-BDE-47 were synthesized at City
University of Hong Kong, and purities were more than 98% as
described previously.51 13C-PCB-178, 13C-2′-OH-BDE-99, and
13
C-BDE-139 were purchased from Cambridge Isotope
Laboratories (Andover, MA, USA). BDE-47, 6-OH-BDE-47,
and 6-MeO-BDE-47 were dissolved in dimethyl sulfoxide
(DMSO, Generay Biotech, Shanghai, China) to prepare stock
solutions and then diluted with embryonic rearing water (60
mg/L instant ocean salt in aerated distilled water) to the
desired test concentrations. Concentration of DMSO in final
test solutions did not exceed 0.1%. RNAlater, RNA
Stabilization Reagents, and RNeasy Mini Kit were purchased
from QIAGEN (Hilden, Germany). Maxima First Strand
cDNA Synthesis Kits were purchased from Fermentas (St
Leon-Rot, Germany). SYBR Real time PCR Master Mix Plus
Kits were purchased from Toyobo (Tokyo, Japan).
Animals and Exposure Experiment. Adult (7 months
old) AB strain zebrafish maintenance and culturing were
performed as previously described.20 The eggs were examined
under a stereomicroscope and only normally developed
embryos were used for exposure experiments. Briefly, 20
embryos were randomly distributed into a 25 mL glass beaker
containing 20 mL of exposure solution. Fish were exposed until
120 h post fertilization (hpf), by which time they had
developed into free-swimming larvae and most organs had
completed development.52 The control group received 0.1%
DMSO (v/v) only. 100% of the exposure solutions were
replaced by fresh exposure solution every 48 h. For 6-OH-BDE47, BDE-47, and 6-MeO-BDE-47 exposures, the experiments
included two parts: First, zebrafish embryos were exposed to 6OH-BDE-47 (0, 0.008, 0.02, 0.05, 0.1, 0.5 μM), 6-MeO-BDE47 (0, 0.02, 0.1, 0.5, 2.5 μM), and BDE-47 (0, 0.02, 0.1, 0.5, 2.5
μM) from 4 to 120 hpf to study the morphologic toxicity of
compounds as previously described.20 Second, on the basis of
the results of acute toxicity test, three comparable exposure
concentrations for each compound were chosen: 6-OH-BDE47 at 0.008, 0.02, and 0.05 μM, and at 0.1, 0.5, and 2.5 μM for
both 6-MeO-BDE-47 and BDE-47 from 4 to 120 hpf to study
the effects on expression of 63 genes involved in eight receptormediated pathways by q-RT-PCR. After exposure for 120 hpf,
larvae were anesthetized with ethyl 3-aminobenzoate methanesulfonate (MS-222, Suzhou Xin Yong Biological Medicine
Technology Co., Ltd., Jiangsu, China), and were preserved in
RNAlater RNA Stabilization Reagents until total RNA isolation.
Bioavailability Analysis and QA/QC. This experiment
was designed to analyze bioaccumulation of the three chemicals
in early life-stages of zebrafish. In each treatment, 600 zebrafish
1824
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
in R. A heatmap of gene expression results was implemented by
“pheatmap” package version 0.7.7 in R.
embryos were exposed to 6-OH-BDE-47, 6-MeO-BDE-47, or
BDE-47 at 300 μg/L (0.6, 0.58, and 0.62 μM, respectively)
from 4 to 120 hpf in glass beakers. Variation among
concentrations of the three compounds in both the exposure
medium and the embryos/larvae was determined. 80 embryos/
larvae and corresponding exposure solutions were collected at
12, 24, 48, 72, 96, and 120 hpf. Detailed protocols for
extraction, clean up, and quantification, and quality assurance
and quality control (QA/QC) are provided in previous
studies20,53,54 and the Supporting Information, Methods and
Table S1.
Quantitative RT-PCR. Total RNA was isolated from
zebrafish larvae using RNeasy Mini Kit. The concentration
and quality of total RNA were determined in accordance with
the procedures described in a previous study.45 First-strand
cDNA synthesis and quantitative RT-PCR were performed
using Maxima First Strand cDNA Synthesis and SYBR Realtime
PCR Master Mix Plus Kits.20 Quantitative RT-PCR was
performed by an Applied Biosystems Stepone Plus Real-time
PCR System (Foster City, California, USA). The primers were
either mined from previous literature55 or designed using
Primer 3 (http://bioinfo.ut.ee/primer3-0.4.0/). Primer sequences are listed in the Supporting Information, Table S2. The
housekeeping gene 18S small subunit rRNA (18S rRNA) was
used as an internal control.56 The thermal cycle was set at 95
°C for 2 min, followed by 40 cycles of 95 °C for 15 s and 60 °C
for 1 min. Melting curves were derived during RT-PCR to
validate that all cDNA samples amplified only a single product.
Levels of expression of genes were normalized to 18S rRNA
mRNA contents using the 2−ΔΔCt method. Each concentration
was measured in triplicate or quadruplicate in a composite
sample containing 20 larvae.
Nuclear Receptor Pathway Analysis. For genes relating
to AhR and ER pathways, the Agilent Literature Search
application was used to construct a biological interaction
network within the Cytoscape software v3.1.1 (Cytoscape
consortium, San Diego, CA, USA).57−59 The gene networks of
the other six NR pathways were retrieved by either WikiPathways (http://www.wikipathways.org)60 or SABioscience Gene
Network Central (http://www.sabiosciences.com/
genenetwork/genenetworkcentral.php), and integrated with
AhR and ER pathways as “associations” and visualized as one
network by Cytoscape. Only genes of interest were shown in
this pathway network. The resulting network genes (nodes)
were colored by the Enhanced Graphics application within
Cytoscape according to the significant fold changes of gene
expressions in the respective treatments.
Statistical Analysis. SPSS 12.0 (SPSS Inc., Chicago, IL,
USA) was used for statistical analysis. A Kolmogorov−Smirnov
test was used to verify the normality of the data, and the
homogeneity of variances was analyzed by Levene’s test as
previously described.20 If the data failed the Kolmogorov−
Smirnov test, logarithmic transformation was performed and
data was checked again for homogeneity of variances. A oneway analysis of variance (ANOVA) followed by LSD test was
used to evaluate differences between the control and exposure
groups. A value of p < 0.05 was considered statistically
significant. To capture the likely nonlinearity in concentrations
in exposure water or in zebrafish embryos−larvae across
different time-points, generalized additive models (GAMs)
were used by the “mgcv” package in R software version 3.10 (R
Core Team, Vienna, Austria). Hierarchical cluster analysis for
the gene expression was performed by the “complete” method
■
RESULTS
Morphologic Effects of 6-OH-BDE-47, 6-MeO-BDE-47
and BDE-47. Among the three compounds, 6-OH-BDE-47
was the most potent to zebrafish embryos/larvae (Supporting
Information, Figure S1B,C,D,F and Table S3). Exposure to all
concentrations caused delayed development of embryos for up
to 6−8 at 24 hpf. In embryos exposed to 0.5 μM 6-OH-BDE47, mortality significantly increased to 22.5 ± 4.1%, at 48 hpf.
At 48 hpf, in groups exposed to concentrations greater than 0.1
μM, the embryos developed hypopigmentation. At 72 hpf,
development was arrested in all embryos exposed to 0.5 μM 6OH-BDE-47 (Supporting Information, Figure S1B,C) whereas
development of embryos at 12−18 hpf was not altered. Larvae
exposed to 0.1 μM 6-OH-BDE-47 exhibited spinal curvatures
(Supporting Information, Figure S1D), decreased heartbeats,
and reduced body lengths (3516 ± 250 μm in 0.1 μM group
and 4040 ± 55 μm in control). The LC50 values of 6-OH-BDE47 for teratogenic effects were 0.28 μM (0.21−0.38) at 72 hpf,
0.13 μM (0.11−0.16) at 96 hpf, and 0.09 μM (0.04−0.10) at
120 hpf. The most sensitive toxicological end-point was spinal
curvature (Supporting Information, Figure S1F), which was
manifested in a concentration-dependent manner with maximal
effects resulting from exposure to 0.08 μM (0.07−0.09) 6-OHBDE-47 at 120 hpf.
There were no significant differences in developmental
alterations in individuals exposed to 6-MeO-BDE-47 and the
control group until 96 hpf. The most sensitive toxicological
end-points were concomitant spinal curvature and pericardial
edema at 120 hpf at 2.5 μM (Supporting Information, Figure
S1G). Although no statistically significant differences were
observed within 96 hpf following exposure to BDE-47 up to 2.5
μM, concomitant spinal curvature and pericardial edema
occurred in embryos at 120 hpf. The NOEC of spinal curvature
and pericardial edema was 0.5 μM of BDE-47, although there
were significant differences at 2.5 μM and the proportion of
affected larvae was 27.6 ± 18.3% (Supporting Information,
Figure S1H). In addition, at 120 hpf following exposure to 2.5
μM BDE-47, the body lengths of larvae (3751 ± 152 μm) were
significantly reduced compared to those in the control group
(4029 ± 201 μm).
Accumulation by Zebrafish. Concentrations of BDE-47,
6-OH-BDE-47, or 6-MeO-BDE-47 were below their analytical
method detection limits in the control group (Supporting
Information Table S4). At 120 hpf, 100% mortality occurred
following exposure to 300 μg/L 6-OH-BDE-47; therefore, no
data is available for this time-point. gas chromatography−mass
spectrometry (GC/MS) results indicated that the chemical
concentrations in exposure solutions decreased and the doses in
embryos and larvae increased in a time-dependent manner
(Figure 1). Moreover, the concentrations of BDE-47 and 6MeO-BDE-47 were approximately 10- to 100-fold greater than
concentrations of 6-OH-BDE-47 in zebrafish embryos and
larvae that were exposed to the same concentrations of the
three compounds (Figure 1). The three BDE congeners ranked
as follows regarding their in vivo accumulation (values from
greater to lesser potential): 6-MeO-BDE-47 > BDE-47 > 6OH-BDE-47.
Calculated BCF values for 6-OH-BDE-47 were 4.07, 9.88,
21.7, 26.9, and 23.3 at 12, 24, 48, 72, and 96 hpf, respectively
(Figure 2A). For 6-MeO-BDE-47, the BCF values were 17.2,
1825
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
concentration of 6-OH-BDE-47 (0.21 and 1.07 μg/g, wm vs
0.15 μg/g, wm) in larvae exposed to 0.05 μM of 6-OH-BDE-47
(Table 1).
Transcriptional Responses of NR Pathways to 6-OHBDE-47. A hierarchical cluster analysis of gene expression data
showed a dendrogram that highlighted five principal clusters
(Figure 3A). Exposures to 2.5 μM 6-MeO-BDE-47 resulted in a
unique clustering of gene expression data that revealed a
significantly different gene expression profile from the other
exposures (Figure 3A). Exposures to the same compound but
at different concentrations generally clustered in the same
group, especially for exposures to 0.008, 0.1, and 0.5 μM 6-OHBDE-47 (Figure 3A).
Zebrafish embryos exposed to 6-OH-BDE-47 had significant
alterations in the expression of genes associated with several
NR pathways. The most significant effects occurred along the
AhR pathway (Figures 3 and 4 and the Supporting Information,
Table S5) with exposure to 0.008 μM 6-OH-BDE-47 causing a
significant up-regulation in the expression of ahr1a, ahr1b, ahr2,
and arnt2 by 2.32-, 1.71-, 1.62-, and 1.62-fold, respectively (p <
0.05), and a significant down-regulation of ahrra and cyp19b
expression by 2.27- and 1.81-fold, respectively (p < 0.05).
Exposure to 0.02 μM 6-OH-BDE-47 significantly induced
expression of ahr1b and ahr2 by 1.63- and 1.67-fold,
respectively, and reduced the expression of ahrra and cyp1a1
by 1.96- and 1.75-fold (p < 0.05). Exposure to 0.05 μM 6-OHBDE-47 significantly reduced the expression of cyp1a1, cyp1b1,
arnt1la, ahrra, and cyp19b by 1.92-, 2.63-, 1.25-, 1.96-, and 2.23fold, respectively (p < 0.05). In addition to the ahr receptor,
following exposure to 0.008 and 0.05 μM 6-OH-BDE-47, the
expression of mr and er2b were significantly up-regulated by
1.51- and 1.83-fold, respectively (p < 0.05).
Transcriptional Responses of NR Pathways to 6-MeOBDE-47. Some NR-mediated pathways such as AhR, ER, AR,
TR, and GR were affected by exposure to 6-MeO-BDE-47, with
the greatest effects occurring at the greatest concentration
tested, 2.5 μM (Figures 3 and 4). Exposure to lesser
concentrations of 6-MeO-BDE-47, 0.1 μM, also induced the
expression of several genes in these pathways. Specifically,
arnt2, dut, ugtlal, ctnnb1, pa2g4a, dap3, and rela were
significantly induced by 2.08-, 1.67-, 2.23-, 1.56-, 1.78-, and
1.51-fold, respectively (p < 0.05). Exposure to 0.5 μM 6-MeOBDE-47 did not significantly alter the expressions of most
genes, except for the down-regulation cyp1a1 and er2a by 9.09and 1.56-fold, respectively (p < 0.05). However, exposure to a
greater concentration of 6-MeO-BDE-47, 2.5 μM, reduced the
expression of most altered genes. Along the AhR pathway, ahr2
was down-regulated by 1.92-fold and associated genes such as
cyp1a1, cyp1b1, cyp365a, and sp1 were also down-regulated by
50-, 100-, 3.03-, and 1.89-fold, respectively (p < 0.05).
However, both ahrra and ahrrb were significantly induced by
3.16- and 5.71-fold, respectively (p < 0.05). In the ER pathway,
er2a and ccnd1 were down-regulated by 1.43- and 1.85-fold,
respectively, whereas er2b was up-regulated by 1.44-fold
following exposure to 2.5 μM 6-MeO-BDE-47 (p < 0.05). In
the AR pathway, ar, ctnnb1, pa2g4b, and ncoa1 were downregulated by 1.85-, 1.89-, 1.47-, and 1.67-fold, respectively (p <
0.05). Exposure to 2.5 μM 6-MeO-BDE-47 also decreased the
expression of thra by 2.17-fold and TR associated genes such as
ncor and fus were significantly down-regulated by 2.08- and
1.85-fold (p < 0.05). Also, following exposure to 2.5 μM 6MeO-BDE-47, the expression of gr and tgfb1 were downregulated by 2.04- and 1.72-fold, respectively (p < 0.05).
Figure 1. Measured concentrations in exposure medium (μg/L) and
in zebrafish embryos−larvae (mg/g, wm) after exposure to 300 μg/L
of BDE-47 (0.62 μM), 6-MeO-BDE-47 (0.58 μM), or 6-OH-BDE-47
(0.6 μM) across time-points (hpf). Generalized additive model
(GAM) plots between concentrations (after log transformation) in
exposure water or in zebrafish embryos−larvae and time (hpf) are
given. Shaded areas are the 95% confidence intervals for each GAM.
The F-statistics, p-values, and adjusted R2 for the specific GAMs are
given in each plot, whereas D shows the deviance explained.
42.6, 89.7, 390, 1207, and 935 at 12, 24, 48, 72, 96, and 120 hpf,
respectively. For BDE-47, bioconcentration factor (BCF)
values were 25.7, 68.0, 489, 750, 2489, and 2430 at 12, 24,
48, 72, 96, and 120 hpf, respectively. These results indicated
that trends of BCF values after various durations of exposure
were similar for 6-MeO-BDE-47 and BDE-47, reaching a
maximum at 96 hpf, and showing a slight decrease at later timepoints.
Biotransformation in Zebrafish. Exposure of zebrafish to
300 μg/L 6-MeO-BDE-47 resulted in increasing tissue
concentrations of 6-OH-BDE-47 in their tissues were quantified
in a time-dependent manner, with 0.02, 0.11, 0.12, and 0.25 μg/
g, wm (wet mass) at 48, 72, 96, and 120 hpf, respectively
(Figure 2B). However, no transformation of BDE-47 into 6OH-BDE-47 or 6-MeO-BDE occurred. Also, 6-OH-BDE-47
was not transformed into either BDE-47 or 6-MeO- BDE-47.
Based on these transformation ratios, the amounts of
biotransformed 6-OH-BDE-47 were expected to be 0.04,
0.21, and 1.07 μg/g, wm after 120 hpf exposure to 0.1, 0.5,
or 2.5 μM 6-MeO-BDE-47, respectively (Table 1). Amounts of
biotransformed 6-OH-BDE-47 in larvae exposed to 0.5 and 2.5
μM of 6-MeO-BDE-47 were greater than the accumulated
1826
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
Figure 2. (A) Bioconcentration factors (BCF) calculated after exposure to 300 μg/L of BDE-47 (0.62 μM), 6-MeO-BDE-47 (0.58 μM), and 6-OHBDE-47 (0.6 μM) across time-points (hpf). BCF was calculated based on the ratio of measured concentrations in zebrafish embryos−larvae (μg/kg,
wm) and measured concentrations in exposure medium (μg/L) at a specific time (hpf). (B) Measured concentrations of biotransformed 6-OHBDE-47 in zebrafish embryos−larvae (μg/g, wm) after exposure to 300 μg 6-MeO-BDE-47/L (0.58 μM) across several time-points (hpf). A linear
regression between concentrations in zebrafish embryos−larvae and time (hpf) is given (R2 = 0.911; p < 0.05).
Table 1. Estimated Internal Doses of BDE-47, 6-OH-BDE-47, and 6-MeO-BDE-47 Exposures in Zebrafish Larvae at 120 hpfa
chemical
nominal concentration (μM)
BDE-47 (μg/g)
6-MeO-BDE-47 (μg/g)
6-OH-BDE-47 (μg/g)
biotransformed 6-OH-BDE-47 (μg/g)
BDE-47
0.1
9.05
0.5
45.27
6-MeO-BDE-47
2.5
226.33
6-OH-BDE-47
0.1
0.5
2.5
18.17
90.87
454.37
0.04
0.21
1.07
0.008
0.02
0.05
0.024
0.061
0.15
a
The calculation of internal dose for each compound was based on the ratio of the exposure concentration at 300 μg/L and the measured
concentration in larvae at 120 hpf.
Transcriptional Responses of NR Pathways to BDE-47.
Exposure to BDE-47 significantly affected the expression of
receptors in the AhR, ER, and GR pathways (Figures 3 and 4).
The number of altered genes increased in a concentration
dependent manner. The expression of gr was significantly
down-regulated by 1.28-fold following exposure to 0.5 μM
BDE-47 (p < 0.05). When exposed to 2.5 μM, expressions of
ahr1b, er2a, and er2b were significantly greater by 1.64-, 1.35-,
and 1.40-fold, respectively (p < 0.05) than those of the
respective genes controls. Expressions of genes along AhR and
ER pathways, such as cyp1a1, cyp19a, cyp3a65, and ccnd1, were
significantly less by factors of 5.88-, 2.88-, 2.04-, and 1.52-fold,
respectively, relative to that of the controls (p < 0.05).
then correlated with the concentrations of the compounds in
the ambient media. Because chemicals need to be accumulated
into organisms and distributed to target sites for the induction
of toxic effects, usage of target site effect concentrations are
postulated to best represent the hazards of a compound in
vivo.61,62 However, for small-bodied organisms such as the
zebrafish, effect concentrations at the target site are difficult to
determine, particularly at earlier stages of development. Average
body concentrations of contaminants in zebrafish embryos and
larvae are subject to competitive dynamic processes, which
include the ability of compounds to penetrate the chorion, in
vivo biotransformation, distribution, and excretion. Hence,
internal effect concentrations need to be determined. Moreover, the ratio of the internal concentration in a fish to the
surrounding concentration at a steady state represents the
compound’s BCF, which is an important metric for regulatory
assessment of chemicals.63
■
DISCUSSION
Hazard assessment of contaminants is typically based on the
exposure of aquatic organisms to chemical solutions for a
defined exposure time and the adverse outcomes observed are
1827
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
Figure 3. Dendrogram displaying similarities of chemicals and doses based on effects on genes in nuclear receptor pathways for BDE-47, 6-MeOBDE-47, and 6-OH-BDE-47 in zebrafish larvae at 120 hpf. (A) The dendrogram of hierarchical cluster analysis was calculated using the average gene
expression values (63 genes in total) of the three or four biological replicates per exposure. Samples names are composed by the name of exposure
compound followed by the exposure concentration (μM). Different colors in the dendrogram denoted five clustering groups. (B) The heatmap of
gene expression profiles was generated using the average gene expression values of the three or four biological replicates per exposure. The foldchanges of gene expression are given in the respective cells and genes involved in different receptor pathways are given different colors (see legend).
ences in lipophilicity were considered to be the most important
parameters for the different accumulation properties of the test
chemicals. For compounds such as 6-OH-BDE-47, the hydroxyl
group by making the compound more polar, result in greater
excretion, and might also be an important parameter for the
lesser in vivo concentration. Hence, our results confirmed that
in aquatic exposure tests, it is not sufficient to evaluate the
ecotoxicological risk of a compound based solely on the
exposure concentration. In addition to accumulation in the
body, the results of this study indicated that the amounts of 6MeO-BDE-47, but not BDE-47, that were transformed to 6OH-BDE-47 increased in a time-dependent manner, (approximately 0.01%, 0.04%, and 0.08% at 48, 96, and 120 hpf,
respectively), which is indicative of an increasing metabolic
capability of zebrafish embryos/larvae with increasing age.
Photomicrographs demonstrated that exposures to 6-OHBDE-47, 6-MeO-BDE-47, and BDE-47 resulted in developmental abnormalities in zebrafish embryos and larvae. The
embryo-toxic effects of BDE47, 6-OH-BDE47 and 6-MeOBDE47 have been investigated in a previous study with
In the present study, accumulation, biotransformation, varied
among BCF of BDE-47, 6-OH-BDE-47, and 6-MeO-BDE-47
during multiple developmental stages of zebrafish. The timepoints during which accumulation of BDE-47 and 6-MeOBDE-47 increased substantially coincide with the hatching
period of zebrafish embryos (48 hpf). In the late developmental
periods of the larvae (after 96 hpf), the bioaccumulation of
BDE-47 and 6-MeO-BDE-47 reached a plateau, which might be
due to an increase in metabolism and/or excretion activities as
well as the rapid growth of larvae at this time that might dilute
their body concentrations over this time period. In this study,
the BCF of 6-OH-BDE-47 at 96 hpf was 23.3, which is similar
to previously reported results,64 in which BCF values were
calculated in liver of zebrafish after 96 h exposure to 100 nM 6OH-BDE-47. At all six durations of exposure, 6-OH-BDE-47
was the least accumulated into the body, though it had the
greatest toxic potency. It is known that a compound with
greater log Kow generally has greater bioaccumulation potential.
Values of log Kow were 6.76, 7.17, and 6.59 for BDE-47, 6MeO-BDE-47, and 6-OH-BDE-47, respectively.65 Thus, differ1828
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
of this study (Supporting Information, Figure S1B,C,D).64
Furthermore, the lack of toxic effects of BDE47 or 6-MeOBDE47 at 2.5 μM until 120 hpf (Supporting Information,
Figure S1G,H) is consistent with previous findings that showed
that no toxicity was observed for both BDE47 and 6-MeOBDE47 in zebarfish at 72 hpf.64
Because disruptions of cellular molecules or processes are
thought to precede adverse outcomes, changes to normal
molecular processes might function as sensitive biomarkers to
predict adverse biological outcomes.66 Moreover, alteration of
NR mediated pathways has been shown to be associated with
adverse endocrine and developmental effects that were linked
with morphological deformities.20,55 For example, previous
studies have demonstrated that BDE-47 can alter thyroid status
and thyroid hormone-regulated gene transcription in the
pituitary and brain of adult fathead minnows,67 and both 6OH-BDE-47 and 6-MeO-BDE-47 were shown to affect
expression of TRα and TRβ genes in the TR pathway, which
can result in teratogenic effects such as pericardial edema,
developmental retardation, and curved spine in zebrafish
embryos.20 Also, all BDE-47, TBBPA and BPA have been
demonstrated to alter expression of genes along the
hypothalamus-pituitary-thyroid (HPT) axis of zebrafish larvae
as well as induce acute toxicity.68 Additionally, zebrafish has
been used to investigate effects of EDCs on the expression of
genes in six NR mediated pathways.55 In this study, two
additional receptor pathways-AR and PxR were added, and the
interactions of sixty-three genes involved in eight zebrafish
receptor pathways were integrated. This pathway network
might represent a novel tool for the examination of the
molecular function of each individual receptor as well as for the
study of their combinatorial regulatory network within NRs.
The structures of the three compounds tested are similar.
The cluster dendrogram showed that expression of genes in
individuals exposed to various concentrations of 6-OH-BDE-47
clustered together. Nevertheless, clustering is a function of
concentration for 6-MeO-BDE-47 and BDE-47. Patterns of
expression of genes following exposures to 2.5 μM BDE-47 and
0.5 μM 6-MeO-BDE-47 were grouped into the same cluster,
which indicates BDE-47 likely has fewer effects on zebrafish
NR-mediated pathways than 6-MeO-BDE-47 at similar waterborne exposure concentrations. Exposure to 2.5 μM of the
more bioaccumulative 6-MeO-BDE-47 resulted in a unique
gene expression profile compared to BDE-47 and 6-OH-BDE47. In addition, 1.07 μg/g, wm of biotransformed 6-OH-BDE47 were detected in larvae exposed to 2.5 μM of 6-MeO-BDE47, which was much greater than the detected amount of 6OH-BDE-47, 0.15 μg/g, which resulted from exposure to 0.05
μM 6-OH-BDE-47. The greater body burden of 6-OH-BDE-47
resulting from exposure to 2.5 μM 6-MeO-BDE-47 might
explain the significant and great effect on gene transcription of
NR pathways observed in this exposure group. The effects that
occurred in the greatest exposure group of 6-MeO-BDE-47,
therefore, were attributed to the combined effects of
biotransformed 6-OH-BDE-47 as well as 6-MeO-BDE-47.
The clustering of the three compounds correlated well with
their respective accumulation potency, indicating the great
importance of the usage of internal dose to assess the dose−
response relationship for studies of PBDEs, especially MeOPBDEs.
Further analyses of endocrine pathways indicated general
disruption of receptor pathways by all three BDEs congeners,
which correlated well with the observed teratogenic effects in
Figure 4. Interaction network of selected genes in nuclear steroid
receptor pathways of zebrafish. Nodes represent single genes, edges
either protein−protein or protein−DNA interactions. Statistically
significant changes (p < 0.05) in gene expression following different
concentrations of treatment of BDE-47, 6-MeO-BDE-47, and 6-OHBDE-47 at 120 hpf are given in the respective boxes (see legend).
zebrafish embryos exposed from 3 to 72 hpf.64 Those authors
showed that 6-OH-BDE-47 was the most toxic BDE-47
congener, inducing a range of developmental defects including
pericardial edema, yolk sac deformations, lesser pigmentation,
lessened heart rate, and delayed development at concentrations
of 25−50 nM, which is consistent with the morphology findings
1829
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
regulated by 6-OH-BDE-47 exposure, indicating 6-OH-BDE-47
might be an agonist of zebrafish PxR.
Reports on the effect of PBDEs on PPARα, MR, and GR are
limited. PPARα plays an important role in lipid homeostasis,
inflammation, adipogenesis, reproduction, and carcinogenesis.78
In this study, none of three compounds significantly affected
the expression of PPARα in zebrafish. However, treatment with
a PBDE mixture, BDE-71 and BDE-47 caused increases in
PPARγ transcript levels at day 8 in 3T3-L1 mouse embryo
fibroblast cells.79 GR and MR are essential for regulation of
multiple physiological functions, such as glucose metabolism,
mineral balance, and behavior.80 In this study, exposure to 2.5
μM 6-MeO-BDE-47 or 0.5 μM BDE-47 caused downregulation of GR, whereas 0.008 μM 6-OH-BDE-47 increased
MR expression, which suggests that GR or MR signaling
pathways might be involved in the endocrine disrupting effects
induced by PBDEs.
Altogether, the results of present study, which compared the
toxicities of BDE-47 with its OH- and MeO- analogs in
zebrafish via multiple quantitative approaches, ranging from in
vivo toxicity tests, bioaccumulation and biotransformation, to
the molecular analysis of response patterns of genes along NR
pathways, highlight the importance of the usage of internal dose
to evaluate the toxic effects for PBDEs, and the use of early lifestages of zebrafish as an efficient and reliable vertebrate model
to assess toxicological effects of endocrine disruptors. Our data
also elucidated several molecular aspects of BDE-47, 6-OHBDE-47, and 6-MeO-BDE-47 induced toxicities. Specifically,
the data provided valuable insights into the early interaction of
these compounds with steroid hormone receptor pathways,
which provided novel clues for their in vivo mechanisms of
subsequent endocrine disruption and developmental toxicities.
zebrafish. Specifically, 6-OH-BDE-47 altered the expression of
AhR, ER, and MR receptor-mediated pathways, whereas AhR,
ER, and GR were the primary pathways altered by BDE-47. Yet,
exposure to the more bioaccumulative 6-MeO-BDE-47 affected
AhR, ER, AR, GR, and TRα pathways. Molecular structures of
OH-BDEs closely resemble those of thyroid hormones (THs)
and 6-OH-BDE-47 can disrupt normal thyroid homeostasis and
functions as either an agonist or an antagonist.31,69 For
example, expression of genes along the hypothalamuspituitary-thyroid (HPT) axis that is responsible for regulation
of metabolism and early life-stage development was affected by
BDE-47 and its OH- or MeO- forms in zebrafish embryos−
larvae.20,68 However, in the present experiment, expression of
thra was not significantly altered following exposure to 6-OHBDE-47, which contradicted previous findings showing thra
was reduced in zebrafish exposed to 200 μg/L 6-OH-BDE-47.20
This difference might be due to the almost 10 times greater
concentrations used in the previous study. However, exposure
to 2.5 μM 6-MeO-BDE-47 significantly decreased the
expression of thra, nocr, and f us in TRα pathway. Because
significant quantities of biotransformed 6-OH-BDE-47 were
found in vivo, 6-MeO-BDE-47 may exert adverse effects
indirectly via transformation into 6-OH-BDE-47, which then
can bind directly to TH targeted genes by mimicking THs.
Apart from the TRα pathway, recent studies have also
focused on disruption of the AhR and ER pathways by PBDEs.
Cross talk between ER- and AhR-signaling pathways in fish has
been hypothesized previously.72,73 The pathway analyses
conducted in this study also suggest interactions between
these pathways as visualized in the constructed networks. All
three compounds altered AhR and ER pathways in zebrafish. 6OH-BDE-47 significantly increased expression of er2b in
zebrafish, which is consistent with previous in vitro findings
that both ERα and ERβ gene and protein expression were
induced by 6-OH-BDE-47.42 Exposure to 2.5 μM 6-MeO-BDE47 significantly reduced expression of er2a but induced er2b
indicating the compound might cause endocrine disrupting
effects through interfering with the ER signaling pathway.21 In
addition, 6-OH-BDE-47 increased the expression of several
AhR receptors including ahr1a, ahr1b, and ahr2 in vivo, while 6MeO-BDE-47 and BDE-47 only affected ahr2 and ahr1a
transcription, respectively. These results have confirmed that
OH-BDEs can induce greater dioxin-like activity than
corresponding MeO-BDEs and parent PBDEs in vitro.43,74
AR and PxR pathways have been previously shown to be
affected by PBDEs in vitro.21,75−77 In the MDA-kb2 human cell
line AR receptor binding assay, all three compounds exhibited
potent antiandrogenicity, with potencies ranking as follows: 6OH-BDE-47 (IC50 = 0.34 μM) > BDE47 (IC50 = 3.83 μM) >
6-MeO-BDE-47 (IC50 = 41.8 μM).21,76 However, in zebrafish,
both BDE-47 and 6-OH-BDE-47 did not significantly alter AR
expression, which is consistent with a study in porcine ovarian
follicular cells, showing BDE-47 and its OH- metabolites had
no effect on the expression of AR mRNA and protein
expression.42 The PxR, a steroid and xenobiotic nuclear
receptor (SXR), can be activated by BDE-47 in mice.75
Nevertheless, in zebrafish, BDE-47 significantly down-regulated
PxR associated genes of cyp3a65, hnf4a, and ugtlal, but
increased pou1f1. The incongruities between these results
could be due to differences between species and lesser
concentrations used in our experiments. Furthermore, PxR
associated genes cyp24a1 and hnf4a were significantly up-
■
ASSOCIATED CONTENT
* Supporting Information
S
Further details on the analytical methods and additional tables
and figures as noted in the text. This material is available free of
charge via the Internet at http://pubs.acs.org/.
■
AUTHOR INFORMATION
Corresponding Authors
*Dr. Hongling Liu. Tel: +86-25-89680356. Fax: +86-2589680356. E-mail: hlliu@nju.edu.cn.
*Dr. Hongxia Yu. Tel: +86-25-89680356. Fax: +86-2589680356. E-mail: yuhx@nju.edu.cn.
Author Contributions
#
These authors contributed equally to this work.
Notes
The authors declare no competing financial interest.
■
ACKNOWLEDGMENTS
We thank Dr. Richard A. Erickson (Upper Midwest Environmental Sciences Center, U.S. Geological Survey) for providing
the expertise in statistical analyses. This work was funded by
National Natural Science Foundation (No. 21377053 and
20977047) and Major National Science and Technology
Projects (No. 2012ZX07506-001 and 2012ZX07501-003-02)
of China. J.P.G. and M.H. were supported by the Canada
Research Chair Program. J.P.G. was supported by the Program
of 2012 “Great Level Foreign Experts” (#GDW20123200120)
funded by the State Administration of Foreign Experts Affairs,
China to Nanjing University, and the Einstein Professor
1830
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
(16) Malmvarn, A.; Zebuhr, Y.; Kautsky, L.; Bergman, K.; Asplund, L.
Hydroxylated and methoxylated polybrominated diphenyl ethers and
polybrominated dibenzo-p-dioxins in red alga and cyanobacteria living
in the Baltic Sea. Chemosphere 2008, 72 (6), 910−916.
(17) Wan, Y.; Liu, F.; Wiseman, S.; Zhang, X.; Chang, H.; Hecker,
M.; Jones, P. D.; Lam, M. H.; Giesy, J. P. Interconversion of
hydroxylated and methoxylated polybrominated diphenyl ethers in
Japanese medaka. Environ. Sci. Technol. 2010, 44 (22), 8729−8735.
(18) Wan, Y.; Wiseman, S.; Chang, H.; Zhang, X.; Jones, P. D.;
Hecker, M.; Kannan, K.; Tanabe, S.; Hu, J.; Lam, M. H.; Giesy, J. P.
Origin of hydroxylated brominated diphenyl ethers: Natural
compounds or man-made flame retardants? Environ. Sci. Technol.
2009, 43 (19), 7536−7542.
(19) Liu, F.; Wiseman, S.; Wan, Y.; Doering, J. A.; Hecker, M.; Lam,
M. H.; Giesy, J. P. Multi-species comparison of the mechanism of
biotransformation of MeO-BDEs to OH-BDEs in fish. Aquat. Toxicol.
2012, 114−115, 182−188.
(20) Zheng, X.; Zhu, Y.; Liu, C.; Liu, H.; Giesy, J. P.; Hecker, M.;
Lam, M. H.; Yu, H. Accumulation and biotransformation of BDE-47
by zebrafish larvae and teratogenicity and expression of genes along
the hypothalamus-pituitary-thyroid axis. Environ. Sci. Technol. 2012, 46
(23), 12943−12951.
(21) Hu, W.; Liu, H.; Sun, H.; Shen, O.; Wang, X.; Lam, M. H.;
Giesy, J. P.; Zhang, X.; Yu, H. Endocrine effects of methoxylated
brominated diphenyl ethers in three in vitro models. Mar. Pollut. Bull.
2011, 62 (11), 2356−2361.
(22) Chan, W. K.; Chan, K. M. Disruption of the hypothalamicpituitary-thyroid axis in zebrafish embryo-larvae following waterborne
exposure to BDE-47, TBBPA and BPA. Aquat. Toxicol. 2012, 108,
106−111.
(23) Davies, R.; Zou, E. Polybrominated diphenyl ethers disrupt
molting in neonatal Daphnia magna. Ecotoxicology 2012, 21 (5),
1371−1380.
(24) Dong, W.; Macaulay, L. J.; Kwok, K. W. H.; Hinton, D. E.;
Stapleton, H. M. Using whole mount in situ hybridization to examine
thyroid hormone deiodinase expression in embryonic and larval
zebrafish: A tool for examining OH-BDE toxicity to early life stages.
Aquat. Toxicol. 2013, 132, 190−199.
(25) Usenko, C. Y.; Hopkins, D. C.; Trumble, S. J.; Bruce, E. D.
Hydroxylated PBDEs induce developmental arrest in zebrafish.
Toxicol. Appl. Pharmacol. 2012, 262 (1), 43−51.
(26) Kallqvist, T.; Grung, M.; Tollefsen, K. E. Chronic toxicity of
2,4,2′,4′-tetrabromodiphenyl ether on the marine alga Skeletonema
costatum and the crustacean Daphnia magna. Environ. Toxicol. Chem.
2006, 25 (6), 1657−1662.
(27) Zhao, J.; Xu, T.; Yin, D. Q. Locomotor activity changes on
zebrafish larvae with different 2,2′,4,4′-tetrabromodiphenyl ether
(PBDE-47) embryonic exposure modes. Chemosphere 2014, 94, 53−
61.
(28) Chen, X.; Huang, C.; Wang, X.; Chen, J.; Bai, C.; Chen, Y.;
Chen, X.; Dong, Q.; Yang, D. BDE-47 disrupts axonal growth and
motor behavior in developing zebrafish. Aquat. Toxicol. 2012, 120−
121, 35−44.
(29) Chou, C. T.; Hsiao, Y. C.; Ko, F. C.; Cheng, J. O.; Cheng, Y. M.;
Chen, T. H. Chronic exposure of 2,2′,4,4′-tetrabromodiphenyl ether
(PBDE-47) alters locomotion behavior in juvenile zebrafish (Danio
rerio). Aquat. Toxicol. 2010, 98 (4), 388−395.
(30) Hamers, T.; Kamstra, J. H.; Sonneveld, E.; Murk, A. J.; Visser, T.
J.; Van Velzen, M. J.; Brouwer, A.; Bergman, A. Biotransformation of
brominated flame retardants into potentially endocrine-disrupting
metabolites, with special attention to 2,2′,4,4′-tetrabromodiphenyl
ether (BDE-47). Mol. Nutr. Food Res. 2008, 52 (2), 284−298.
(31) Ren, X. M.; Guo, L. H. Molecular toxicology of polybrominated
diphenyl ethers: Nuclear hormone receptor mediated pathways.
Environ. Sci. Process Impact. 2013, 15 (4), 702−708.
(32) Castrillo, A.; Tontonoz, P. Nuclear receptors in macrophage
biology: At the crossroads of lipid metabolism and inflammation.
Annu. Rev. Cell Dev. Biol. 2004, 20, 455−480.
Program of the Chinese Academy of Sciences. He was also
supported by a Visiting Distinguished Professorship in the
Department of Biology and Chemistry and State Key
Laboratory in Marine Pollution at City University of Hong
Kong.
■
REFERENCES
(1) Birnbaum, L. S.; Staskal, D. F. Brominated flame retardants:
Cause for concern? Environ. Health Perspect. 2004, 112 (1), 9−17.
(2) Qu, W. Y.; Bi, X. H.; Sheng, G. Y.; Lu, S. Y.; Fu, H.; Yuan, J.; Li,
L. P. Exposure to polybrominated diphenyl ethers among workers at
an electronic waste dismantling region in Guangdong, China. Environ.
Int. 2007, 33 (8), 1029−1034.
(3) Chen, D.; Letcher, R. J.; Burgess, N. M.; Champoux, L.; Elliott, J.
E.; Hebert, C. E.; Martin, P.; Wayland, M.; Chip Weseloh, D. V.;
Wilson, L. Flame retardants in eggs of four gull species (Laridae) from
breeding sites spanning Atlantic to Pacific Canada. Environ. Pollut.
2012, 168, 1−9.
(4) Valters, K.; Li, H.; Alaee, M.; D’Sa, I.; Marsh, G.; Bergman, A.;
Letcher, R. J. Polybrominated diphenyl ethers and hydroxylated and
methoxylated brominated and chlorinated analogues in the plasma of
fish from the Detroit River. Environ. Sci. Technol. 2005, 39 (15), 5612−
5619.
(5) Haraguchi, K.; Kotaki, Y.; Relox, J. R., Jr.; Romero, M. L.; Terada,
R. Monitoring of naturally produced brominated phenoxyphenols and
phenoxyanisoles in aquatic plants from the Philippines. J. Agric. Food
Chem. 2010, 58 (23), 12385−12391.
(6) Wiseman, S. B.; Wan, Y.; Chang, H.; Zhang, X. W.; Hecker, M.;
Jones, P. D.; Giesy, J. P. Polybrominated diphenyl ethers and their
hydroxylated/methoxylated analogs: Environmental sources, metabolic
relationships, and relative toxicities. Mar. Pollut. Bull. 2011, 63 (5−12),
179−188.
(7) Rotander, A.; van Bavel, B.; Riget, F.; Auethunsson, G. A.; Polder,
A.; Gabrielsen, G. W.; Vikingsson, G.; Mikkelsen, B.; Dam, M.
Methoxylated polybrominated diphenyl ethers (MeO-PBDEs) are
major contributors to the persistent organobromine load in sub-Arctic
and Arctic marine mammals, 1986−2009. Sci. Total Environ. 2012,
416, 482−489.
(8) Yogui, G. T.; Sericano, J. L. Polybrominated diphenyl ether flame
retardants in the U.S. marine environment: A review. Environ. Int.
2009, 35 (3), 655−666.
(9) Liu, X.; Jiao, Y.; Lin, C.; Sun, K.; Zhao, Y. PBDEs, hydroxylated
PBDEs and methoxylated PBDEs in bivalves from Beijing markets.
Chemosphere 2014, 110, 97−103.
(10) Sun, J.; Liu, J.; Liu, Y.; Jiang, G. Hydroxylated and methoxylated
polybrominated diphenyl ethers in mollusks from Chinese coastal
areas. Chemosphere 2013, 92 (3), 322−328.
(11) Yin, G.; Asplund, L.; Qiu, Y.; Zhou, Y.; Wang, H.; Yao, Z.; Jiang,
J.; Bergman, A. Chlorinated and brominated organic pollutants in
shellfish from the Yellow Sea and East China SeaEnviron. Sci. Pollut.
Res. Int. 2014, DOI: 10.1007/s11356-014-3198-8.
(12) Wang, H. S.; Du, J.; Ho, K. L.; Leung, H. M.; Lam, M. H. W.;
Giesy, J. P.; Wong, C. K. C.; Wong, M. H. Exposure of Hong Kong
residents to PBDEs and their structural analogues through market fish
consumption. J. Hazard. Mater. 2011, 192 (1), 374−380.
(13) McKinney, M. A.; De Guise, S.; Martineau, D.; Beland, P.;
Lebeuf, M.; Letcher, R. J. Organohalogen contaminants and
metabolites in beluga whale (Delphinapterus leucas) liver from two
Canadian populations. Environ. Toxicol. Chem. 2006, 25 (5), 1246−
1257.
(14) Malmvarn, A.; Marsh, G.; Kautsky, L.; Athanasiadou, M.;
Bergman, A.; Asplund, L. Hydroxylated and methoxylated brominated
diphenyl ethers in the red algae Ceramium tenuicorne and blue mussels
from the Baltic Sea. Environ. Sci. Technol. 2005, 39 (9), 2990−2997.
(15) Bowden, B. F.; Towerzey, L.; Junk, P. C. A new brominated
diphenyl ether from the marine sponge Dysidea herbacea. Aust. J. Chem.
2000, 53 (4), 299−301.
1831
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
and potential polybrominated diphenyl ether metabolites. Eur. J. Org.
Chem. 2003, 14, 2566−2576.
(52) Amsterdam, A.; Nissen, R. M.; Sun, Z.; Swindell, E. C.;
Farrington, S.; Hopkins, N. Identification of 315 genes essential for
early zebrafish development. Proc. Natl. Acad. Sci. U. S. A. 2004, 101
(35), 12792−12797.
(53) Wen, Q.; Liu, H. L.; Zhu, Y. T.; Zheng, X. M.; Su, G. Y.; Zhang,
X. W.; Yu, H. X.; Giesy, J. P.; Lam, M. H. Maternal transfer,
distribution, and metabolism of BDE-47 and its related hydroxylated,
methoxylated analogs in zebrafish (Danio rerio). Chemosphere 2014,
120C, 31−36.
(54) Wen, Q.; Liu, H. L.; Su, G. Y.; Wei, S.; Feng, J. F.; Yu, H. X.
Determination of polybrominated diphenyl ethers and their derivates
in zebrafish eggs. Chin. J. Anal. Chem. 2012, 40 (11), 1698−1702.
(55) Liu, C.; Wang, Q.; Liang, K.; Liu, J.; Zhou, B.; Zhang, X.; Liu,
H.; Giesy, J. P.; Yu, H. Effects of tris(1,3-dichloro-2-propyl) phosphate
and triphenyl phosphate on receptor-associated mRNA expression in
zebrafish embryos/larvae. Aquat. Toxicol. 2013, 128−129, 147−157.
(56) Tang, S.; Allagadda, V.; Chibli, H.; Nadeau, J. L.; Mayer, G. D.
Comparison of cytotoxicity and expression of metal regulatory genes
in zebrafish (Danio rerio) liver cells exposed to cadmium sulfate, zinc
sulfate and quantum dots. Metallomics 2013, 5 (10), 1411−1422.
(57) Cline, M. S.; Smoot, M.; Cerami, E.; Kuchinsky, A.; Landys, N.;
Workman, C.; Christmas, R.; Avila-Campilo, I.; Creech, M.; Gross, B.;
Hanspers, K.; Isserlin, R.; Kelley, R.; Killcoyne, S.; Lotia, S.; Maere, S.;
Morris, J.; Ono, K.; Pavlovic, V.; Pico, A. R.; Vailaya, A.; Wang, P.-L.;
Adler, A.; Conklin, B. R.; Hood, L.; Kuiper, M.; Sander, C.;
Schmulevich, I.; Schwikowski, B.; Warner, G. J.; Ideker, T.; Bader,
G. D. Integration of biological networks and gene expression data
using Cytoscape. Nat. Protoc. 2007, 2 (10), 2366−2382.
(58) Shannon, P.; Markiel, A.; Ozier, O.; Baliga, N. S.; Wang, J. T.;
Ramage, D.; Amin, N.; Schwikowski, B.; Ideker, T. Cytoscape: A
software environment for integrated models of biomolecular
interaction networks. Genome Res. 2003, 13 (11), 2498−2504.
(59) Saito, R.; Smoot, M. E.; Ono, K.; Ruscheinski, J.; Wang, P. L.;
Lotia, S.; Pico, A. R.; Bader, G. D.; Ideker, T. A travel guide to
Cytoscape plugins. Nat. Methods 2012, 9 (11), 1069−1076.
(60) Pico, A. R.; Kelder, T.; van Iersel, M. P.; Hanspers, K.; Conklin,
B. R.; Evelo, C. WikiPathways: pathway editing for the people. PLoS
Biol. 2008, 6 (7), e184.
(61) Escher, B. I.; Hermens, J. L. M. Modes of action in
ecotoxicology: Their role in body burdens, species sensitivity,
QSARs, and mixture effects. Environ. Sci. Technol. 2002, 36 (20),
4201−4217.
(62) Escher, B. I.; Hermens, J. L. M. Internal exposure: Linking
bioavailability to effects. Environ. Sci. Technol. 2004, 38 (23), 455A−
462A.
(63) Adolfsson-Erici, M.; Akerman, G.; McLachlan, M. S. Internal
benchmarking improves precision and reduces animal requirements
for determination of fish bioconcentration factors. Environ. Sci.
Technol. 2012, 46 (15), 8205−8211.
(64) van Boxtel, A. L.; Kamstra, J. H.; Cenijn, P. H.; Pieterse, B.;
Wagner, J. M.; Antink, M.; Krab, K.; van der Burg, B.; Marsh, G.;
Brouwer, A.; Legler, J. Microarray analysis reveals a mechanism of
phenolic polybrominated diphenylether toxicity in zebrafish. Environ.
Sci. Technol. 2008, 42 (5), 1773−1779.
(65) Yu, Y.; Yang, W.; Gao, Z.; Lam, M. H. W.; Liu, X.; Wang, L.; Yu,
H. RP-HPLC measurement and quantitative structure-property
relationship analysis of the noctanol-water partitioning coefficients of
selected metabolites of polybrominated dyphenyl ethers. Environ.
Chem. 2008, 5 (5), 332−339.
(66) Aardema, M. J.; MacGregor, J. T. Toxicology and genetic
toxicology in the new era of “toxicogenomics”: Impact of “-omics”
technologies. Mutat. Res. 2002, 499 (1), 13−25.
(67) Lema, S. C.; Dickey, J. T.; Schultz, I. R.; Swanson, P. Dietary
exposure to 2,2′,4,4′-tetrabromodiphenyl ether (PBDE-47) alters
thyroid status and thyroid hormone-regulated gene transcription in
the pituitary and brain. Environ. Health Perspect. 2008, 116 (12),
1694−1699.
(33) Chinenov, Y.; Gupte, R.; Rogatsky, I. Nuclear receptors in
inflammation control: Repression by GR and beyond. Mol. Cell.
Endocrinol. 2013, 380 (1−2), 55−64.
(34) Knoedler, J. R.; Denver, R. J. Kruppel-like factors are effectors of
nuclear receptor signaling. Gen. Comp. Endrocrinol. 2014, 203, 49−59.
(35) Omiecinski, C. J.; Vanden Heuvel, J. P.; Perdew, G. H.; Peters, J.
M. Xenobiotic metabolism, disposition, and regulation by receptors:
From biochemical phenomenon to predictors of major toxicities.
Toxicol. Sci. 2011, 120 (Suppl 1), S49−S75.
(36) Aranda, A.; Pascual, A. Nuclear hormone receptors and gene
expression. Physiol. Rev. 2001, 81 (3), 1269−1304.
(37) Jennings, P.; Limonciel, A.; Felice, L.; Leonard, M. O. An
overview of transcriptional regulation in response to toxicological
insult. Arch. Toxicol. 2013, 87 (1), 49−72.
(38) Jacobs, M. N.; Dickins, M.; Lewis, D. F. Homology modelling of
the nuclear receptors: Human oestrogen receptorbeta (hERbeta), the
human pregnane-X-receptor (PXR), the Ah receptor (AhR) and the
constitutive androstane receptor (CAR) ligand binding domains from
the human oestrogen receptor alpha (hERalpha) crystal structure, and
the human peroxisome proliferator activated receptor alpha (PPARalpha) ligand binding domain from the human PPARgamma crystal
structure. J. Steroid Biochem. Mol. Biol. 2003, 84 (2−3), 117−132.
(39) Prossnitz, E. R.; Arterburn, J. B.; Smith, H. O.; Oprea, T. I.;
Sklar, L. A.; Hathaway, H. J. Estrogen signaling through the
transmembrane G protein-coupled receptor GrPR30. Annu. Rev.
Physiol. 2008, 70, 165−190.
(40) Norman, A. W.; Mizwicki, M. T.; Norman, D. P. Steroidhormone rapid actions, membrane receptors and a conformational
ensemble model. Nat. Rev. Drug Discovery 2004, 3 (1), 27−41.
(41) Lema, S. C.; Dickey, J. T.; Schultz, I. R.; Swanson, P. Dietary
Exposure to 2,2′,4,4′-Tetrabromodiphenyl ether (PBDE-47) alters
thyroid status and thyroid hormone-regulated gene transcription in the
pituitary and brain. Environ. Health Perspect. 2008, 116 (12), 1694−
1699.
(42) Karpeta, A.; Ptak, A.; Gregoraszczuk, E. L. Different action of
2,2′,4,4′-tetrabromodiphenyl ether (BDE-47) and its hydroxylated
metabolites on ERalpha and ERbeta gene and protein expression.
Toxicol. Lett. 2014, 229 (1), 250−256.
(43) Su, G.; Xia, J.; Liu, H.; Lam, M. H.; Yu, H.; Giesy, J. P.; Zhang,
X. Dioxin-like potency of HO- and MeO- analogues of PBDEs’ the
potential risk through consumption of fish from Eastern China.
Environ. Sci. Technol. 2012, 46 (19), 10781−10788.
(44) Stegeman, J. J.; Goldstone, J. V.; Hahn, M. E. Perspectives on
zebrafish as a model in environmental toxicology. Fish Physiol. 2010,
29, 367−439.
(45) Tang, S.; Cai, Q.; Chibli, H.; Allagadda, V.; Nadeau, J. L.; Mayer,
G. D. Cadmium sulfate and CdTe-quantum dots alter DNA repair in
zebrafish (Danio rerio) liver cells. Toxicol. Appl. Pharmacol. 2013, 272
(2), 443−452.
(46) Dai, Y. J.; Jia, Y. F.; Chen, N.; Bian, W. P.; Li, Q. K.; Ma, Y. B.;
Chen, Y. L.; Pei, D. S. Zebrafish as a model system to study toxicology.
Environ. Toxicol. Chem. 2014, 33 (1), 11−17.
(47) Baker, M. E.; Hardiman, G. Transcriptional analysis of
endocrine disruption using zebrafish and massively parallel sequencing.
J. Mol. Endocrinol. 2014, 52 (3), R241−R256.
(48) Scholz, S.; Fischer, S.; Gundel, U.; Kuster, E.; Luckenbach, T.;
Voelker, D. The zebrafish embryo model in environmental risk
assessmentApplications beyond acute toxicity testing. Environ. Sci.
Pollut. Res. Int. 2008, 15 (5), 394−404.
(49) Scholz, S.; Mayer, I. Molecular biomarkers of endocrine
disruption in small model fish. Mol. Cell. Endocrinol. 2008, 293 (1−2),
57−70.
(50) Bertrand, S.; Thisse, B.; Tavares, R.; Sachs, L.; Chaumot, A.;
Bardet, P. L.; Escriva, H.; Duffraisse, M.; Marchand, O.; Safi, R.;
Thisse, C.; Laudet, V. Unexpected novel relational links uncovered by
extensive developmental profiling of nuclear receptor expression. PLoS
Genet. 2007, 3 (11), 2085−2100.
(51) Marsh, G.; Stenutz, R.; Bergman, A. Synthesis of hydroxylated
and methoxylated polybrominated diphenyl ethers - Natural products
1832
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Article
Environmental Science & Technology
(68) Chan, W. K.; Chan, K. M. Disruption of the hypothalamicpituitary-thyroid axis in zebrafish embryo-larvae following waterborne
exposure to BDE-47, TBBPA and BPA. Aquat. Toxicol. 2012, 108,
106−111.
(69) Ren, X. M.; Guo, L. H.; Gao, Y.; Zhang, B. T.; Wan, B.
Hydroxylated polybrominated diphenyl ethers exhibit different
activities on thyroid hormone receptors depending on their degree
of bromination. Toxicol. Appl. Pharmacol. 2013, 268 (3), 256−63.
(70) Zhou, H.; Wu, H.; Liao, C.; Diao, X.; Zhen, J.; Chen, L.; Xue, Q.
Toxicology mechanism of the persistent organic pollutants (POPs) in
fish through AhR pathway. Toxicol. Mech. Methods 2010, 20 (6), 279−
286.
(71) Shibata, H.; Spencer, T. E.; Onate, S. A.; Jenster, G.; Tsai, S. Y.;
Tsai, M. J.; O’Malley, B. W. Role of co-activators and co-repressors in
the mechanism of steroid/thyroid receptor action. Recent Prog. Horm.
Res. 1997, 52, 141−164 discussion 164-5.
(72) Gjernes, M. H.; Schlenk, D.; Arukwe, A. Estrogen receptorhijacking by dioxin-like 3,3′4,4′,5-pentachlorobiphenyl (PCB126) in
salmon hepatocytes involves both receptor activation and receptor
protein stability. Aquat. Toxicol. 2012, 124−125, 197−208.
(73) Grans, J.; Wassmur, B.; Celander, M. C. One-way inhibiting
cross-talk between arylhydrocarbon receptor (AhR) and estrogen
receptor (ER) signaling in primary cultures of rainbow trout
hepatocytes. Aquat. Toxicol. 2010, 100 (3), 263−270.
(74) Kojima, H.; Takeuchi, S.; Uramaru, N.; Sugihara, K.; Yoshida,
T.; Kitamura, S. Nuclear hormone receptor activity of polybrominated
diphenyl ethers and their hydroxylated and methoxylated metabolites
in transactivation assays using Chinese hamster ovary cells. Environ.
Health Perspect. 2009, 117 (8), 1210−1218.
(75) Pacyniak, E. K.; Cheng, X.; Cunningham, M. L.; Crofton, K.;
Klaassen, C. D.; Guo, G. L. The flame retardants, polybrominated
diphenyl ethers, are pregnane X receptor activators. Toxicol. Sci. 2007,
97 (1), 94−102.
(76) Liu, H.; Hu, W.; Sun, H.; Shen, O.; Wang, X.; Lam, M. H.;
Giesy, J. P.; Zhang, X.; Yu, H. In vitro profiling of endocrine disrupting
potency of 2,2′,4,4′-tetrabromodiphenyl ether (BDE47) and related
hydroxylated analogs (HO-PBDEs). Mar. Pollut. Bull. 2011, 63 (5−
12), 287−296.
(77) Fery, Y.; Buschauer, I.; Salzig, C.; Lang, P.; Schrenk, D.
Technical pentabromodiphenyl ether and hexabromocyclododecane as
activators of the pregnane-X-receptor (PXR). Toxicology 2009, 264
(1−2), 45−51.
(78) Abbott, B. D. Review of the expression of peroxisome
proliferator-activated receptors alpha (PPAR alpha), beta (PPAR
beta), and gamma (PPAR gamma) in rodent and human development.
Reprod. Toxicol. 2009, 27 (3−4), 246−257.
(79) Tung, E. W.; Boudreau, A.; Wade, M. G.; Atlas, E. Induction of
adipocyte differentiation by polybrominated diphenyl ethers (PBDEs)
in 3T3-L1 cells. PLoS One 2014, 9 (4), e94583.
(80) Greenwood, A. K.; Butler, P. C.; White, R. B.; DeMarco, U.;
Pearce, D.; Fernald, R. D. Multiple corticosteroid receptors in a teleost
fish: Distinct sequences, expression patterns, and transcriptional
activities. Endocrinology 2003, 144 (10), 4226−4236.
1833
DOI: 10.1021/es503833q
Environ. Sci. Technol. 2015, 49, 1823−1833
Supporting Information
Bioaccumulation, biotransformation and toxicity of BDE-47, 6-OH-BDE-47
and 6-MeO-BDE-47 in early life-stages of zebrafish (Danio rerio)
Hongling Liu1#*, Song Tang2#, Xinmei Zheng1, Yuting Zhu1, Zhiyuan Ma1, Chunsheng Liu1,
Markus Hecker2,3, David M.V. Saunders3, John P. Giesy1,3,4,5, Xiaowei Zhang1, Hongxia Yu1*
1
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment,
Nanjing University, Nanjing, Jiangsu 210023, China
2
School of Environment and Sustainability, University of Saskatchewan, Saskatoon, SK S7N
5B3, Canada
3
Toxicology Centre, University of Saskatchewan, Saskatoon, SK S7N 5B3, Canada
4
Department of Veterinary Biomedical Sciences, University of Saskatchewan, Saskatoon,
SK S7N 5B3, Canada
5
Department of Biology and Chemistry, City University of Hong Kong, Kowloon, Hong
Kong, SAR, China
#
These authors contributed equally to this work.
*
Correspondence to: Drs. Hongling Liu and Hongxia Yu, School of the Environment,
Nanjing University, Nanjing, Jiangsu 210023, China
Tel: +86-25-89680356; Fax: +86-25-89680356; Email: hlliu@nju.edu.cn (Dr. Hongling Liu)
and yuhx@nju.edu.cn (Dr. Hongxia Yu)
A summary of 13 pages including 1 figure and 5 tables
S1
Chemical Analysis Procedures
Bioavailability analysis
This experiment was designed to determine bioaccumulation of the three chemicals in
early life-stages of zebrafish. In each treatment, 600 zebrafish embryos were exposed to
6-OH-BDE-47, 6-MeO-BDE-47 or BDE-47 at 300 μg/L (0.6, 0.58, and 0.62 μM, respectively)
from 4 to 120 hpf in glass beakers (Diameter 150 mm). Concentration of each of the three
compounds was quantified in both exposure solution and embryos/larvae. 80 embryos/larvae
and corresponding exposure solutions were collected at 0, 12, 24, 48, 72, 96, and 120 hpf.
Detailed protocols for extraction, clean up, identification and quantification, and quality
assurance and quality control (QA/QC) are provided in previous publications.1-3 In brief,
samples of exposure solutions were collected and diluted to 5 mL with MilliQ water, and 1
mL hydrochloric acid (2 M HCl) were added. After vortex mixing, samples were extracted
with 6 mL n-hexane-methyl tert butyl ether mixture (MTBE) (1:1, v/v) three times and were
dried under nitrogen. Dried extracts were dissolved in 480 μL of derivatization solvent
(acetonitrile: methanol: water: pyridine = 5:2:2:1, v/v/v/v), and 40 μL methylchloroformate
(MCF) was added. Samples were then incubated at 25 oC with vortex for 1 h, after which 1.4
mL MilliQ water was added and extracted three times with 6 mL n-hexane. The extracts were
concentrated, and 100 μL of 13C-PCB-178 was added as the internal injection standard and
made up to 100 μL prior to GC/MS analysis. Embryos and larvae from each exposure were
composited, rinsed with MilliQ water, gently dried, and weighed. Internal dose and potential
biotransformation of individual OH-PBDEs, MeO-PBDEs and PBDEs were also determined.
Approximately 0.1 g sample was homogenized and transferred into amber serum bottles, and
spiked surrogate recovery standard (13C-2’-OH-BDE-99 and 13C-BDE-139), and 2 mL MilliQ
water, 50 µL HCl at 2 M, and 2 mL 3-propanol was added. Then they were extracted three
times with 10 mL of n-hexane/MTBE (1:1,v/v). Extracts were concentrated by rotary
evaporation and dried under nitrogen. After, the derivation of PBDEs, MeO-PBDEs and
OH-PBDEs and purification was followed by the previous method.1-3 Concentrations of 12
PBDEs, 12 OH-PBDEs and 12 MeO-PBDEs were determined by use of a TSQ Quantum
GC/MS (Thermo Scientific, USA) coupled with an Agilent DB-XLB column (15 m × 0.25
mm × 0.25 μm, J&W Scientific, USA) in 3 separate runs. Identification of specific PBDEs,
S2
OH-PBDEs and MeO-PBDEs was performed by comparing relative retention times versus
internal standard and product ions in SRM mode with the standards. Time-courses of
bioconcentration factors (BCF) for each chemical were also calculated based on measured
concentrations in zebrafish embryos and larvae divided by the determined concentration in
water.
QA/QC
QA/QC was conducted by the analysis of procedural blanks (RSD<22.6%, for 6
replicates) as previously described3 and the obtained recoveries of standard compounds were
ranged 93%-146%, 67%-113% and 103%-133% for PBDEs, OH-PBDEs and MeO-PBDEs,
respectively. Limit of detection (LOD) of each compound was defined as three times the SD
of the laboratory blanks. For congeners not detected in the blanks, we set the LOD at the
instrumental limit of quantification (LOQ). Method LODs ranged 9.2×10-2-2.9×101 ng/g wet
mass (wm) for individual PBDEs, OH-PBDEs and MeO-PBDEs congeners.
Concentrations
less than the method LOD were assumed to be not detected. PBDEs, OH-PBDEs and
MeO-PBDEs were all quantified in samples extracts relative to
13
C-PCB-178. Recoveries of
surrogate standards were 62.1%-104.7%, 82.9%-117.9%, and 44.9%-74.9% for BDE-47
(13C-BDE-139), 6-MeO-BDE-47 (13C-BDE-139), and 6-OH-BDE-47 (13C-2’-OH-BDE-99),
respectively.
Water Quality of Exposure Solutions
During the exposure, zebrafish embryos/larvae were maintained at 28 oC in egg water
(60 mg/L ocean salts) and the water quality parameters were measured every day including
pH (7.1-7.6), DO (95% saturation), hardness (15-18 mg/L as CaCO3) and ammonia (0-3
mg/L). Temperature, hardness, pH, oxygen concentration, and ammonia in all treatments
were similar from start and end of the exposures.
S3
Figure S1. Effects of exposures to 6-OH-BDE-47, 6-MeO-BDE-47 and BDE-47 on
morphologies of several development stages of zebrafish embryos and larvae. A) normal
zebrafish at 72 hpf; B) edema caused by 0.5 μM 6-OH-BDE-47 at 72 hpf; C) development
arrested by 0.5 μM 6-OH-BDE-47 at 72 hpf; D) spinal curvature caused by 0.1 μM
6-OH-BDE-47 at 72 hpf; E) normal zebrafish at 120 hpf; F) severe edema and spinal
curvature caused by 0.05 μM 6-OH-BDE-47 at 120 hpf, LC50=0.09 μM; G) spinal curvature
caused by 2.5 μM 6-MeO-BDE-47 at 120 hpf, NOEC=0.5 μM; H) spinal curvature caused by
2.5 μM BDE-47 at 120 hpf, NOEC=0.5 μM.
S4
Table S1. Results of repeatability, recovery and detection limit with spiked procedural
blanks.
Repeatability
Compounds
RSD
Detection Limit
Recovery
BDE-17
BDE-28
BDE-71
BDE-47
BDE-66
BDE-100
BDE-99
BDE-85
BDE-154
BDE-138
BDE-183
BDE-190
6'-MeO-BDE-17
2'-MeO-BDE-28
2'-MeO-BDE-68
6-MeO-BDE-47
2'-MeO-BDE-47
4'-MeO-BDE-49
4’-MeO-BDE-90
6-MeO-BDE-90
3-MeO-BDE-100
2-MeO-BDE-123
6-MeO-BDE-85
6-MeO-BDE-137
15.2%
14.6%
10.0%
9.5%
9.0%
10.1%
10.8%
11.9%
12.2%
12.9%
13.4%
12.4%
11.9%
11.7%
10.2%
9.1%
9.9%
10.2%
9.6%
11.4%
10.5%
11.4%
10.1%
12.1%
119.6%
144.0%
114.6%
119.1%
129.2%
101.9%
105.6%
116.0%
146.0%
93.0%
101.1%
109.5%
113.5%
109.7%
116.4%
119.6%
120.5%
127.3%
106.0%
112.6%
133.0%
116.5%
107.7%
103.0%
Instrument (ng/mL)
2.7
4.5×10-1
9.2×10-2
7.8×10-1
1.2
2.3
5.1×10-1
6.7×10-1
2.6
2.9
4.1
4.2
5.4×10-1
5.5×10-1
4.9
8.0
6.2
6.5
1.9×101
4.5
9.8
8.9
1.0×101
1.6×101
3'-OH-BDE-7
21.2%
21.5%
18.2%
22.6%
16.7%
11.6%
20.8%
20.9%
17.0%
15.4%
21.6%
17.3%
89%
91.9%
92.6%
81%
103.2%
88.9%
71.6%
67%
74.6%
112.8%
87%
91.4%
2.2×101
1.4×101
1.1×101
2.2×101
9.7
6.6
1.5×101
2.9×101
9.9
1.7×101
2.5×101
1.6×101
2'-OH-BDE-7
2'-OH-BDE-17
2'-OH-BDE-25
2'-OH-BDE-28
2-OH-BDE-47
2'-OH-BDE-68
2'-OH-BDE-66
2-OH-BDE-90
2-OH-BDE-85
4'-OH-BDE-49
2-OH-BDE-137
S5
Method (ng/g)
2.7
4.5×10-1
9.2×10-2
7.8×10-1
1.2
2.3
5.1×10-1
6.7×10-1
2.6
2.9
4.1
4.2
5.4×10-1
5.5×10-1
4.9
8.0
6.2
6.5
1.9×101
4.5
9.8
8.9
1.0×101
1.6×101
2.2×101
1.4×101
1.1×101
2.2×101
9.7
6.6
1.5×101
2.9×101
9.9
1.7×101
2.5×101
1.6×101
Table S2. Primer sequences of NR related genes for qRT-PCR.
Genes
11βhsd
18s
abcb311
adrb2a
adrb2b
ahr1b
ahr2
ahrra
ahrrb
aip
ar
arnt2
arntl1a
arntl1b
c1d
ccnd1
ctnnb1
cyp1a1
cyp1b1
cyp24a1
cyp3a65
dap3
dut
egfr
er1
er2a
er2b
flh
fus
gr
hdac3
hnf4a
hpse
hsp90aa1
il6
il8
lpl
mr
ncoa1
ncoa2
ncoa3
ncoa4
Forward Sequence (5’-3’)
tggtgaagtatgccatcgaa
ttgttggtgttgttgctggt
agcgtgtctcttttgggaga
gctgatctggtcatgggatt
cccgattacaagctgatggt
ggagagcacttgaggaaacg
atctccatgggcaaaacaag
gctgctgatgtttggactga
acctgggatttcatcagacg
ccatcacttgaagcctccat
acattctggaggccattgag
gaatggtctcggtccgtcta
tctcctgggggaaagaagat
ctcgctgaatgccatagaca
acggagagctgacagaccat
tgacttgccttgacttgtcg
atcctgtccaacctgacctg
cctgggcggttgtctatcta
gctcagctcggtaacactcc
aagacgtggaaggaccacac
ctgtgcatcatggaccaaac
tcgaccgttcatgtaaacca
tacagacgctggatgagacg
aacgcaaataatggcaggac
ggtccagtgtggtgtcctct
agcttgtgcacatgatcagc
ttgtgttctccagcatgagc
ctcttccaaatgccacgaat
ccaatatgcaggagcaggat
agaccttggtccccttcact
agccatgaaggtgtccattc
gccgacactacagagcatca
cggcagtctgaacagatgaa
ggatctggtgatcctgctgt
tcctggtgaacgacatcaaa
gtcgctgcattgaaacagaa
ctggccttctcaccaaacat
tttgagggaccagacaaacc
tgagagcctctgttggaggt
agagcctgtcagtcccaaga
aactcacctgcccacaaatc
gacaactgcggaaaagaagc
Reverse Sequence (5’-3’)
gcaaagctttttgagccatc
ggatgctcaacaggggttcat
atagcacagctagggccaga
atgtgatggcgatgtaacga
tatgagcaaccccactgtca
ggatccagatcgtcctttga
tccctcttgtgtcgataccc
gacgctgtgttcacgtcact
gctgtacagatgagccgtca
tgcatgtgctccaacttctc
acgtgcaagttacggaaacc
agctggtcacctgcagtctt
ccatcgctgcttcatcatta
cccgagacgactgtattggt
gccgacatcagatccagttt
gaaaaagcagggagcacttg
tctctgcatcctggtgtctg
tgaggaatggtgaagggaag
cgttagacacgaaccggaat
ctctgttgtggcagcgtaaa
ggtgaaggatggtgagagga
ctggatgctgagacacctga
aatgcagcaacacaaacagc
tctccagaaccacagtgcag
cacacgaccagactccgtaa
gctttcatccctgctgagac
ccacatatggggaaggaatg
gggcacgcagtagttatggt
cttccccgtctctctgtctg
cgcctttaatcatgggagaa
agaagctgcttgcaggactc
aggtgttcctggaccagatg
aacacgggacaaatccacat
tccagaacgggcatatcttc
tcatcacgctggagaagttg
cttaacccatggagcagagg
gcctttgaatcccaatgcta
cacactttggctgtcgaaga
ctctgaccctggtttggtgt
ggtcgtagccaccatcagtt
agaggcctgttgctggtcta
ctggggatttggcagagtta
S6
Gene ID
AY578180
NM_200713
NM_001006594
NM_001102652
NM_001089471
NM_001024816
NM_131264
NM_001035265
NM_001033920
NM_214712
NM_001083123
NM_131674
NM_131577
NM_178300
NM_001007059
NM_131025
NM_131059
AF210727
AY727864
NM_001089458
AY452279
NM_001098737
NM_001006005
AY332223
NM_152959
NM_180966
NM_174862
DQ118096
NM_201083
EF567112
NM_200990
NM_194368
NM_001045005
NM_131328
JN698962
XM_001342570
NM_131127
EF567113
XM_686652
NM_131777
XM_687846
NM_201129
ncor
ncor2
pa2g4a
pa2g4b
pgr
pou1f1
pparg
ppargc1a
pparα
pxr
rela
sp1
tgfb1
thra
ube2i
ugt1a1
vtg1
vtg2
vtg4
vtg5
agggtaaggagcagagcaca
ttgaaccagtttcaccacca
cgggaaaaggacatgaagaa
caaagacaccaccacgtttg
caacaggtggttgtggacag
cggagctttgtggagaagag
tgccgcatacacaagaagag
aatcaggattcggtgtggag
gattcaaatcttgccgtggt
ccagctaccagagccttgac
tataagccacacccacacga
tcctccattaatcggtcgag
aactactgcatggggtcctg
caatgtaccatttcgcgttg
tggaaagagggaagatgtgg
attggagaaggctcccaagt
ctgcgtgaagttgtcatgct
tactttgggcactgatgcaa
ctacaaggtggaggctctgc
agctaatgctctgcccgtta
gcaaaactggttcaggtggt
tgacaatggctgagttgctc
aagccgtcaacatgaactcc
gtgccaccattacgcttttt
atttggagatgtccgctttg
ttggtcatgaaggagctgtg
atgtggttcacgtcactgga
ttggatgcttcattgccata
tcgtcgctgagagactgaga
tggtcctccataaccagagc
gaatgggttgttttgcgtct
tgtgtgtgagcacaaaacga
ggacaattgctccaccttgt
gctcctgctctgtgttttcc
cgaatgaagtgaaggggtgt
ggaaaggatccgtgagcata
gaccagcattgcccataact
agacttcgtgaagcccaaga
ggaggacaaatcaccagcat
gttcagcctcaaacagcaca
S7
EF016488
NM_001007032
NM_001002170
NM_212641
NM_001166335
NM_212851
NM_131467
AY998087
NM_001161333
DQ069792
AY163839
NM_212662
AY178450
NM_131396
NM_131351
NM_001037428
NM_001044897
AY729644
NM_001045294
NM_001025189
Table S3. Toxicity of 6-OH-BDE-47, 6-MeO-BDE-47, and BDE-47 to early life-stages of zebrafish.
Chemicals
6-OH-BDE-47
6-MeO-BDE-47
BDE-47
Duration of
development (hpf)
48
72
72
72
96
120
120
120
120
120
Toxicity Endpoints
Hypopigmentation
Spinal curvatures, slow heartbeats, and reduced body lengths
Arrested development
LC50
LC50
LC50
NOEC
Spinal curvatures and pericardial edema
NOEC
Spinal curvatures, pericardial edema, and reduced body lengths
Concentration or Mean (95% CI)
(μM)
≥0.1
0.1
0.5
0.28 (0.21-0.38)
0.13 (0.11-0.16)
0.09 (0.04-0.10)
0.5
2.5
0.5
2.5
Table S4. Measured concentrations of BDE-47, 6-OH-BDE-47 and 6-MeO-BDE-47 in exposure solutions (ng/g) with a nominal concentration
of 300 μg/L at 0 hpf.
Chemical
Control
BDE-47
6-MeO-BDE-47
6-OH-BDE-47
BDE-17
BDE-28
BDE-71
BDE-47
BDE-66
BDE-100
BDE-99
BDE-85
BDE-154
BDE-138
n.d
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
22.39
27.02
n.d.
227.56
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
S8
BDE-183
BDE-190
6'-MeO-BDE-17
2'-MeO-BDE-28
2'-MeO-BDE-68
6-MeO-BDE-47
6-MeO-BDE-90
3-MeO-BDE-100
4-MeO-BDE-90
2-MeO-BDE-123
6-MeO-BDE-85
6-MeO-BDE-137
5-MeO-BDE-47
4'-MeO-BDE-49
3'-OH-BDE-7
2'-OH-BDE-17
2'-OH-BDE-25
2'-OH-BDE-28
2'-OH-BDE-7
6-OH-BDE-47
2'-OH-BDE-68
2'-OH-BDE-66
6-OH-BDE-90
6-OH-BDE-85
4'-OH-BDE-49
6-OH-BDE-137
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d..
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
S9
n.d.
n.d.
n.d.
n.d.
n.d.
493.19
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
282.73
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
S10
Table S5. Fold-change of gene expressions in NR pathways, * p<0.05 indicates significant difference between exposure groups and the control.
NRs
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AhR
AR
AR
AR
AR
AR
AR
AR
ER
ER
ER
Genes
cyp1a1
ahr1a
ahr1b
ahr2
ahrra
ahrrb
aip
arnt2
arntl1a
arntl1b
cyp1b1
ncor2
sp1
ar
ctnnb1
ncoa1
ncoa2
ncoa4
pa2g4a
pa2g4b
brca2
ccnd1
cyp19b
0
1.01
1.02
1.01
1.02
1.03
1.05
1.01
1.01
1.01
1.01
1.25
1.00
1.03
1.00
1.00
1.00
1.03
1.00
1.00
1.01
1.01
1.00
1.00
6-OH-BDE-47 (μM)
0.008
0.02
-1.16
-1.75*
2.32*
1.54
1.71*
1.63*
1.62*
1.67*
-2.27*
-1.96*
1.16
-1.35
1.13
-1.14
1.62*
1.40
-1.04
-1.18
1.15
1.11
-1.03
-1.52
1.03
1.53
-1.16
1.67
1.35
1.20
1.22*
1.24*
1.18
1.22
1.33
1.68
1.20*
1.39*
1.02
1.01
-1.15
-1.18
-1.25
-1.10
1.09
-1.03
-1.81*
-1.20
0.05
-1.92*
1.69
1.18
1.11
-1.64*
-1.96
-1.06
-1.01
-1.25*
-1.99
-2.63*
1.21
-1.39
-1.15
-1.11
-1.14
1.14
1.44*
1.91
1.40
1.00
-1.41*
-2.23*
0
1.01
1.02
1.01
1.00
1.00
1.00
1.04
1.01
1.01
1.00
1.24
1.02
1.03
1.00
1.00
1.00
1.01
1.00
1.00
1.01
1.01
1.00
1.00
6-MeO-BDE-47 (μM)
0.1
0.5
2.5
-1.54*
-9.09* -50.00*
1.49
-1.21
1.20
1.16
1.48
1.23
-1.02
-1.20
-1.92*
1.50
2.18
3.16*
1.50
2.15
5.71**
-1.20
-1.30
-1.00
2.08*
-1.18
-1.75
1.14
-1.19
-1.10
1.35
-1.47
-1.49
-1.05
-1.75
-100.0*
1.30
-1.36
-2.61
1.00
-1.41
-1.89*
1.08
-1.23
-1.85*
1.40*
1.01
-1.89*
1.11
1.04
-1.67*
1.31
1.07
-1.67
1.25
1.05
-1.25
1.56*
1.06
-1.33
1.14
-1.08
-1.47*
1.02
1.03
-1.09
1.15
-1.12
-1.85*
-1.10
2.43
1.07
S11
0
1.03
1.01
1.00
1.00
1.01
1.06
1.02
1.01
1.01
1.00
1.27
1.00
1.06
1.00
1.00
1.00
1.01
1.00
1.00
1.01
1.08
1.05
1.01
BDE-47 (μM)
0.1
0.5
-1.49
-1.72
1.10
1.03
1.29
1.53
-1.05
1.13
1.25
1.18
1.13
1.04
-1.24
1.06
1.22
1.51
1.49
1.91
1.15
1.07
1.17
1.44
1.07
-1.05
1.04
1.40
1.60
1.08
1.13
-1.02
-1.02
1.30
1.04
1.11
1.00
-1.03
1.04
-1.06
-1.03
-1.15
-1.77*
-1.54
-1.35
-1.28
1.39
-1.14
2.5
-5.88*
-1.75
1.64*
-1.09
1.38
-1.28
-1.28
-1.05
-1.28
-1.22
-6.25
-1.49
-2.50
-1.20
-1.45*
1.07
-1.12
1.02
-1.23*
-1.27*
1.22
-1.52*
-2.28*
ER
ER
ER
ER
ER
ER
ER
ER
ER
GR
GR
GR
GR
GR
MR
MR
MR
MR
MR
MR
MR
PPARα
PPARα
PPARα
PPARα
PPARα
er1
er2a
er2b
ncoa3
pgr
vtg1
vtg2
vtg4
vtg5
dap3
gr
hsp90aa1
rela
tgfb1
11βhsd3
adrb2a
adrb2b
egfr
hpse
mr
ube2i
dut
il6
il8
lpl
ppara
1.16
1.00
1.02
1.07
1.05
1.02
1.01
1.01
1.04
1.01
1.00
1.00
1.01
1.00
1.01
1.01
1.01
1.01
1.02
1.02
1.00
1.01
1.00
1.02
1.03
1.00
1.29
1.68
1.47
1.55
-1.36
-1.31
-1.22
-1.67
-1.39
1.42
1.12
-1.25
1.21
1.03
-1.35
1.00
-1.32
1.27
-1.28
1.51*
-1.06
-1.20
-1.20
-2.50*
1.39
-1.13
1.25
1.60
1.64
1.09
-1.12
-1.30
-1.21
-1.20
-1.34
1.28
-1.02
-1.06
1.19
1.04
-1.30
-1.01
-1.25
1.39
-1.06
1.42
-1.11
-1.08
1.56
-1.85*
-1.32
-1.02
1.59
1.23
1.83*
1.37
1.13
-1.02
1.10
-1.09
-1.13
2.18*
-1.30
1.49
1.34
3.27
-1.27
-1.12
-1.25
1.69
-1.16
-1.15
-1.22*
-1.27
2.29
-1.67
-1.51
-1.00
1.04
1.00
1.01
1.05
1.07
1.02
1.02
1.02
1.00
1.01
1.00
1.00
1.01
1.01
1.01
1.00
1.00
1.00
1.02
1.03
1.00
1.00
1.00
1.00
1.00
1.00
-1.24
-1.05
-1.16
-1.09
1.03
-1.35
-1.17
2.23
-3.85
1.78*
1.14
1.65
1.51*
1.09
1.40
-1.14
-1.14
1.47
-1.08
1.26
1.00
1.67*
-2.17
2.04
-1.10
1.15
S12
1.22
-1.56*
-1.05
1.08
1.02
-1.08
1.15
3.55
-5.26
-1.20
1.18
-1.04
1.03
-1.35
1.00
-1.01
1.02
1.07
-1.11
1.19
-1.11
1.02
1.01
2.47
1.28
1.26
1.88
-1.43*
1.44*
-1.75
-1.08
-1.14
1.19
1.06
-3.85
1.28
-2.04*
1.37
-1.96
-1.72*
-1.30
-1.05
1.22
-2.05
-1.45*
-1.89
-1.32*
1.30
2.30
2.21
1.47
-1.18
1.02
1.00
1.00
1.02
1.08
1.12
1.01
1.00
1.00
1.03
1.00
1.00
1.01
1.00
1.01
1.01
1.01
1.01
1.01
1.02
1.00
1.00
1.00
1.00
1.03
1.01
1.51
1.43*
1.27
1.23
-1.25
-1.33
1.44
-7.69*
1.07
1.14
-1.05
-1.07
1.03
1.09
-1.15
1.13
1.48*
-1.15
1.12
1.05
1.10
1.08
2.06
1.33
1.11
1.10
1.69
1.30
1.11
1.31
1.04
-1.36
1.46
-1.19
1.13
1.15
-1.28*
-1.17
-1.10
-1.12
-1.35*
1.05
1.31
-1.14
-1.00
1.09
1.07
1.07
1.72
1.49
1.02
1.40
1.38
1.35*
1.40*
-1.07
1.66
1.03
1.36
1.30
-1.29
1.15
-1.14
-1.01
-1.02
-1.08
-1.49*
1.17
1.36
-1.11
-1.09
-1.09
-1.25*
1.14
-1.56
1.13
-1.79
-1.08
PPARα
PPARα
PxR
PxR
PxR
PxR
PxR
PxR
PxR
TR
TR
TR
TR
TR
pparg
ppargcla
abcb3l1
cyp24a1
cyp3a65
hnf4a
pou1f1
pxr
ugtlal
c1d
fus
hdac3
ncor
thra
1.02
1.00
1.03
1.00
1.00
1.00
1.02
1.02
1.02
1.01
1.01
1.08
1.00
1.00
-1.17
1.24
-1.08
1.01
1.08
1.34*
-1.06
1.12
1.05
-1.49*
1.10
1.04
1.14
1.27
-1.22
1.18
1.12
1.07
1.05
1.38*
1.14
1.23
1.08
-1.23
1.22
1.52
-1.12
1.37
1.24
-1.05
1.04
2.05*
1.22
1.33*
1.16
-1.07
1.12
-1.08
-1.37
1.04
-1.45*
1.05
1.01
1.00
1.05
1.04
1.00
1.00
1.01
1.02
1.00
1.00
1.00
1.00
1.00
1.00
1.50
2.02
-1.19
-1.22
1.12
1.19
-1.02
1.58
2.23*
1.00
1.11
-1.37
1.24
1.42
-1.01
1.01
-1.15
-1.28
-1.23
1.15
1.17
-1.20
1.05
1.06
-1.16
-1.34
-1.19
-1.01
-2.26
-1.27
-1.02
-3.13*
-3.03*
-1.56
1.21
-1.23
-3.13
-1.09
-1.85*
-1.33
-2.08*
-2.17*
1.01
1.00
1.02
1.01
1.01
1.00
1.01
1.01
1.00
1.00
1.00
1.00
1.00
1.00
-1.06
1.07
1.53
1.69
1.18
1.08
1.31
-1.03
1.15
-1.04
1.33
1.05
1.17
-1.05
-1.23
-1.16
1.25
1.42
1.03
-1.06
1.19
-1.18
-1.25
-1.03
1.13
1.30
1.61
1.13
-1.20
-1.79*
1.04
-1.61
-2.04*
-1.56*
1.53*
-1.43
-2.70*
1.13
-1.11
1.69
1.06
-1.22
References
1. Wen, Q.; Liu, H. L.; Zhu, Y. T.; Zheng, X. M.; Su, G. Y.; Zhang, X. W.; Yu, H. X.; Giesy, J. P.; Lam, M. H., Maternal transfer, distribution,
and metabolism of BDE-47 and its related hydroxylated, methoxylated analogs in zebrafish (Danio rerio). Chemosphere 2014, 120C, 31-36.
2. Wen, Q.; Liu, H. L.; Su, G. Y.; Wei, S.; Feng, J. F.; Yu, H. X., Determination of Polybrominated Diphenyl Ethers and Their Derivates in
Zebrafish Eggs. Chine. J Anal. Chem. 2012, 40, (11), 1698-1702.
3. Zheng, X.; Zhu, Y.; Liu, C.; Liu, H.; Giesy, J. P.; Hecker, M.; Lam, M. H.; Yu, H., Accumulation and biotransformation of BDE-47 by
zebrafish larvae and teratogenicity and expression of genes along the hypothalamus-pituitary-thyroid axis. Environ. Sci. Technol. 2012, 46, (23),
12943-12951.
S13