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Author's personal copy
Marine Pollution Bulletin 63 (2011) 179–188
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Polybrominated diphenyl ethers and their hydroxylated/methoxylated analogs:
Environmental sources, metabolic relationships, and relative toxicities
Steve B. Wiseman a, Yi Wan a, Hong Chang a, Xiaowei Zhang a, Markus Hecker a,b, Paul D. Jones a,d,
John P. Giesy a,c,e,f,g,h,i,⇑
a
Toxicology Centre, University of Saskatchewan, Saskatoon, Saskatchewan, Canada S7N 5B3
Cardno-ENTRIX, Inc., Saskatoon, Saskatchewan, Canada S7N 5B3
Dept. of Biology and Chemistry, State Key Laboratory in Marine Pollution, City University of Hong Kong, Kowloon, Hong Kong, SAR, China
d
School of Environment and Sustainability, University of Saskatchewan, Saskatoon, Saskatchewan, Canada S7N 5C8
e
Dept. of Biomedical Veterinary Sciences, University of Saskatchewan, Saskatoon, Saskatchewan, Canada S7N 5B3
f
Dept. of Zoology, Center for Integrative Toxicology, Michigan State University, East Lansing, MI, USA
g
State Key Laboratory of Marine Environmental Science, College of Oceanography and Environmental Science, Xiamen University, Xiamen, PR China
h
School of Biological Sciences, University of Hong Kong, Hong Kong, SAR, China
i
Zoology Department, College of Science, King Saud University, P.O. Box 2455, Riyadh 11451, Saudi Arabia
b
c
a r t i c l e
Keywords:
Natural products
Review
Flame retardants
OH-BDE
MeO-BDE
Biotransformation
i n f o
a b s t r a c t
Brominated compounds are ubiquitous in the aquatic environment. The polybrominated diphenyl ether
(PBDE) flame retardants are anthropogenic compounds of concern. Studies suggest that PBDEs can be biotransformed to hydroxylated brominated diphenyl ethers (OH-BDE). However, the rate of OH-BDE formation observed has been extremely small. OH-BDEs have also been identified as natural compounds
produced by some marine invertebrates. Another class of compounds, the methoxylated BDEs (MeOBDEs), has also been identified as natural compounds in the marine environment. Both the OH-BDEs
and MeO-BDEs bioaccumulate in higher marine organisms. Recent studies have demonstrated that
MeO-BDEs can be biotransformed to OH-BDEs and this generates greater amounts of OH-BDEs than could
be generated from PBDEs. Consequently, MeO-BDEs likely represent the primary source of metabolically
derived OH-BDEs. Given that for some endpoints OH-BDEs often exhibit greater toxicity compared to
PBDEs, it is prudent to consider OH-BDEs as chemicals of concern, despite their seemingly ‘‘natural’’
origins.
Ó 2011 Elsevier Ltd. All rights reserved.
1. Introduction
Polybrominated diphenyl ethers (PBDEs or BDEs) are manufactured as brominated flame retardants (BFRs) that are used as additives in a variety of household and industrial products. These
organohalogen compounds are intended to reduce flame propagation by gas phase radical quenching, thus reducing the spread of
fire in particular in electronics, electrical products, and textiles.
The general structure of brominated diphenyl ethers (BDE) has
two phenyl rings, each with varying degrees of hydrogen or bromine (Br) substitutions, which are joined by an ether bond
(Fig. 1). As with polyhalogenated biphenyls there are 209 possible
BDE congeners depending on the number and position of the Br
atom(s). Production of PBDEs began in the 1960s and continues
⇑ Corresponding author. Address: Toxicology Centre, University of Saskatchewan,
44 Campus Drive, Saskatoon, SK, Canada S7N 5B3. Tel.: +1 306 966 2096; fax: +1
306 966 4796.
E-mail address: john.giesy@usask.ca (J.P. Giesy).
0025-326X/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2011.02.008
to this day. The PBDEs were produced as three commercial formulations; pentabromodiphenyl ether (penta-BDE), octabromodiphenyl ether (octa-BDE) and decabromodiphenyl ether (deca-BDE)
(Costa et al., 2008). Based on the 2001 production of 67,400 tons
it is estimated that total PBDE production over the last 30 years
has exceeded 2 million tons (Shaw and Kannan, 2009).
Concern over the potential effects of PBDEs on environmental
and human health led to a ban on production and usage of the
penta-BDE and octa-BDE formulations in Europe in 2004 and a
voluntary ban on the production of these formulations in certain
American states followed shortly thereafter (Betts, 2008).
Widespread distributions of penta-BDE and octa-BDE in the environment led the Stockholm Convention to include these formulations on the list of persistent organic pollutants (POPs) that are
persistent, bioaccumulative, toxic to exposed organisms, and
undergo long-range transport in the environment (Stockholm
Convention, 2009). Nonetheless, large-scale production and usage
of deca-BDE continues and releases into the environment are
increasing (Hale et al., 2006; Shaw and Kannan, 2009).
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2
2.1. Hydroxylation of PBDEs – in vitro and in vivo exposure studies
2’
3
1
1’
3’
Br
Br
4
6
5
6’
4’
5’
Fig. 1. The basic structure of a polybrominated diphenyl ether (PBDE). The sum of
x + y = 1–10. Relative to the ether bond the ortho positions are 2,20 ,6,60 ; the meta
positions are 3,30 ,5,50 ; and the para positions are 4,40 .
Some flame retardants, such as tetrabromobisphenol A (TBBPA),
are chemically bonded to polymers, making it difficult for them to
leach from the parent materials. Conversely, PBDEs are ‘‘additive’’
flame retardants that are blended with polymers and as a result
are more likely to leach out of products and into the environment.
Large-scale production of PBDEs and their ability to leach from
products has led to their ubiquitous environmental distribution.
PBDEs enter the environment as emissions from primary production facilities, from manufacturing plants that incorporate PBDEs
into their products, and from the products in which they are used.
The disposal of PBDE-containing compounds through recycling
processes or in landfill sites also contributes significantly to the
environmental load of PBDEs. PBDEs have been found in municipal
waste disposal and recycling sites worldwide, including Southeast
Asia (Leung et al., 2010; Muenhor et al., 2010), the United States
(Schecter et al., 2009), and Central America (Athanasiadou et al.,
2008; Gullett et al., 2010). Despite a ban on production of some
PBDEs formulations, disposal of products containing penta-BDE
and octa-BDE formulations continues to increase their concentrations in the environment.
Lesser brominated PBDEs (hexabrominated and less) are expected to undergo long-range atmospheric transport (Moon et al.,
2007; Noël et al., 2009; Wang et al., 2005) and become associated
with various environmental matrices worldwide. Once in the
aquatic environment, PBDEs sorb to sediment and waterborne suspended organic matter. Organisms inhabiting both local discharge
sites and remote locations accumulate PBDEs. Some of the greatest
concentrations of PBDEs have been measured in freshwater and
marine organisms. A thorough review of concentrations of PBDEs
in North American marine ecosystems, including concentrations
in fish, birds and marine mammals has been provided by Shaw
and Kannan (2009).
2. PBDE analogs – hydroxylated and methoxylated BDEs
Structural analogs of PBDEs, including hydroxylated- (OH-BDE)
and methoxylated- (MeO-BDE) BDEs are frequently detected in the
environment. Unlike PBDEs, neither OH-BDEs nor MeO-BDEs are
artificially synthesized or used in industrial processes. However,
both OH-BDEs and MeO-BDEs are known to be produced naturally
in the marine environment (Malmvärn et al., 2005,, 2008). It has
also been reported that OH-BDEs are transformation products of
PBDE flame retardants (Chen et al., 2006; Cheng et al., 2008;
Erratico et al., 2010; Hakk et al., 2002, 2006, 2009; Kierkegaard
et al., 2001; Malmberg et al., 2005; Marsh et al., 2006; McKinney
et al., 2006; Qiu et al., 2007; Stapleton et al., 2009; Staskal et al.,
2006). Given the interest in the toxicity of PBDEs and the structural
similarities between each of these classes of compounds, there is
interest in determining the sources and relationship(s) among
PBDEs, OH-BDEs and MeO-BDEs.
Transformation, either by physicochemical process in the environment or by biotransformation is an important determinant of
bioaccumulation, fate, toxicokinetics, and toxicity of organic molecules. Consequently, it is important to understand the biotransformation of PBDEs. Several in vivo and in vitro studies have reported
formation of OH-BDEs from various PBDE congeners. Both monoand di-OH-BDE were detected in tissues of rats (Chen et al.,
2006; Hakk et al., 2002) and mice exposed to BDE-99 (Staskal
et al., 2006). The biotransformation of BDE-100 to mono-OH-PDEs
and di-OH-BDEs has been reported in rats (Hakk et al., 2006) and
mice (Staskal et al., 2006). Also, biotransformation of BDE-154 to
mono-OH-BDEs, di-OH-BDEs, and trace concentrations of a triOH-BDE was observed in rats (Hakk et al., 2009). Several monoOH-BDEs were measured in rats after they had been administered
BDE-47 (Marsh et al., 2006; Staskal et al., 2006). BDE-153 was reported to be biotransformed into six mono-OH-BDEs in female
mice (Staskal et al., 2006). When biotransformation of the commercial PBDE mixture DE-71 (a mixture of BDE-47, -85, -99,
-100, -153, -154) was studied in mice, six mono-OH-BDE metabolites were reported to be formed (Qiu et al., 2007). Sixteen monoOH-BDEs and two di-OH-BDEs were identified in rats exposed
intraperitoneally to a PBDE mixture (Malmberg et al., 2005).
Hydroxylation of parent PBDEs in non-mammalian vertebrates is
less studied. Only one study was found that reported the presence
of six OH-BDEs that were detected in several tissues, including
blood, after exposure of northern pike (Esox lucius) to dietary 14C
BDE-47 (Kierkegaard et al., 2001).
Biotransformation of PBDEs has also been studied in vitro. In human hepatocytes exposed to BDE-99, two mono-OH-BDEs were
identified (Stapleton et al., 2009). Three mono-OH-BDEs were detected in rat microsomes incubated with BDE-99 (Erratico et al.,
2010). Using microsomes isolated from control, phenobarbitolor b-napthoflavone treated rats both mono- and di-OH-BDEs
were identified in microsomes incubated with BDE-15 and -28.
Mono-OH-BDEs transformation products were detected when
these microsomes were incubated with BDE-47. No biotransformation products were detected in microsomes exposed to BDE-100
(Cheng et al., 2008). Formation of a single OH-BDE was reported
to occur in beluga whale (Delphinapterus leucas) microsomes
exposed to BDE-15 (McKinney et al., 2006).
Together, these results suggest that PBDEs might undergo biotransformation to OH-BDEs. However, where data is available the
amount of the OH-BDEs produced is small. Based on the data presented by Erratico et al. (2010), the concentration ratios between
metabolites (4-OH-BDE-90, 50 -OH-BDE-99 and 60 -OH-BDE-99)
and exposed BDE-99 were approximately 0.056%, 0.043% and
0.01%, respectively. Likewise, in rats exposed to a single dose of a
PBDE mixture containing 3 lmol/kg body weight of BDE-47, -99,
-100, -153, -154, -183 and -209 the mean plasma concentrations
of 4-OH-BDE-42, 3-OH-BDE-47 and 40 -OH-BDE-49 were 0.009%,
0.025% and 0.021% relative to the parent compounds, respectively
(Malmberg et al., 2005). When BDE-47 was incubated with microsomes isolated from phenobarbitol-treated rats, 20 -OH-BDE-66,
3-OH-BDE-47, 4-OH-BDE-42, 40 -OH-BDE-49, 5-OH-BDE-47, and
6-OH-BDE-47 accounted for 0.30%, 0.84%, 0.10%, 0.54%, 0.074%,
and 0.022% of the parent BDE-47, respectively (Hamers et al.,
2008). Two OH-BDE metabolites, 50 -OH-BDE-99 and an unidentified
tetra-OH-BDE, accounted for approximately 0.1–3% of the initial
exposure concentration in human hepatocytes exposed to BDE99 (Stapleton et al., 2009). The proportions of OH-BDE reported
to have been formed during biotransformation of PBDE are so small
that it brings into question whether there was actually conversion
or whether the results were an artifact due to contaminants in the
chemicals used or in the study animals. While it is likely that the
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S.B. Wiseman et al. / Marine Pollution Bulletin 63 (2011) 179–188
authors checked the PBDEs standards used for the exposure studies
for the presence of OH-BDE, it is not known whether the presence
of MeO-BDE impurities was determined. Given recent findings in
our laboratory (Wan et al., 2009, 2010a) that MeO-BDEs are metabolized to OH-BDE, the presence of any MeO-BDE impurities in the
PBDE standards may have contributed to the low levels of OH-BDEs
reported. In a recent study investigating the biotransformation of
BDE-47, 6-OH-BDE-47 and 6-MeO-BDE-47 in Japanese Medaka
(Oryzias latipes), low levels of MeO-BDEs were detected in the commercial fish food to which the exposure chemicals had been spiked.
This illustrates the importance of carefully analyzing all experimental agents that could potentially contribute to PBDEs,
MeO-BDEs, and OH-BDEs detected in animal tissues in controlled
exposure studies.
Not all exposure studies have found OH-BDEs as biotransformation products of PBDEs. No OH-BDE biotransformation products
were detected in S9 fractions from Sprague–Dawley rats, or microsomes from White-leghorn Chicken (Gallus gallus) or rainbow trout
(Oncorhynchus mykiss) exposed to either BDE-99 or a PBDE mixture
(Wan et al., 2009). Similarly, OH-BDEs were not detected in
Japanese Medaka exposed to BDE-47 via their diet (Wan et al.,
2010a). No OH-BDE metabolites were detected in either Rainbow
trout (Stapleton et al., 2006), common carp (Cyprinus carpio)
(Stapleton et al., 2006; Benedict et al., 2007) or Chinook salmon
(Oncorhynchus tshawytscha) (Browne et al., 2009) exposed to
BDE-99. OH-BDEs were not produced by beluga whale microsomes
exposed to either BDE-28 or -47 (McKinney et al., 2006). Also, in
microsomes that were prepared from the Chinese sturgeon
(Acipenser sinensis) and incubated with BDE-47, -99, -154 or -183
measurable amounts of OH-BDEs were not detected (Zhang et al.,
2010). The fact that OH-BDEs have been reported to occur in some
studies but not others, combined with the small rates of biotransformation reported for those studies in which OH-BDEs were
reported to occur, brings into question the veracity of the
biotransformation of PBDEs to OH-BDEs.
2.2. Methoxylation of PBDEs
The occurrence of MeO-BDEs in wildlife tissues has prompted
researchers to hypothesize that PBDEs may undergo methoxylation. Based on bacterial methylation of phenolic compounds it
has been proposed that MeO-BDEs could be formed via methylation of either PBDEs or OH-BDEs by intestinal microflora or microorganisms in sediments (Haglund et al., 1997). However,
MeO-BDEs have not been observed to be formed in controlled
exposure studies with PBDEs performed either in vitro or in vivo.
No MeO-BDEs were detected in BDE-15, -28 or -47 exposed hepatic
microsomes from either Beluga whale or rats (McKinney et al.,
2006). MeO-BDEs were not detected when S9 fractions from livers
of Sprague–Dawley rats, or microsomes from livers of Whiteleghorn Chicken or rainbow trout were exposed to either BDE-99
or a mixture of PBDE congeners (Wan et al., 2009). Similarly, MeOBDEs were not detected in Japanese Medaka exposed to BDE-47
via the diet (Wan et al., 2010a) or in several PBDE exposed hepatic
microsomes isolated from Chinese sturgeon (Zhang et al., 2010).
2.3. Mechanism(s) of PBDE metabolism
Several lines of evidence suggest that biotransformation of
PBDEs to OH-BDEs is mediated by cytochrome P450 (Cyp450 or
Cyp) enzymes. Formation of OH-BDE metabolite in beluga whale
microsomes was dependent upon NADPH, a co-factor required
for Cyp-catalyzed reactions (McKinney et al., 2006). Structure(s)
of OH-BDEs detected in controlled laboratory studies suggest that
both oxidation and oxidative debromination are involved in biotransformation of PBDE (Hakk et al., 2002, 2006, 2009; Orn and
181
Klasson-Wehler, 1998; Qiu et al., 2007; Staskal et al., 2006). It
has been proposed that the first step in generation of OH-BDEs is
the Cyp450-catalyzed formation of epoxide intermediates (Hakk
et al., 2009). This intermediate could then undergo ring opening
without loss of bromine or with the loss of one or two bromine
atoms which would result in formation of mono-OH-BDEs. The
Cyp450-generated intermediate(s) could then be acted upon by
the enzymes epoxide hydrolase and dihydrodiol dehydrogenase,
with or without debromination, which could form di-OH-BDEs. It
has been estimated that 43 different OH-BDEs could be formed
from PBDEs (BDE-47, -49, -100, -153, -154, -183) via direct
hydroxylation followed by the 1,2 bromine shift (Malmberg
et al., 2005).
Given the similarity in structure between PBDEs and dioxins,
such as 2,3,7,8-tetrachlrodibenzo-p-dioxin (2,3,7,8-TCDD) several
studies have investigated induction of Cyp4501A in response to
PBDE exposure. Although several authors have reported increased
Cyp1A expression in livers of animals exposed to DE-71, it has been
found that contamination with polybrominated dibenzo-p-furan
(PBDF) was responsible for this effect (Hanari et al., 2006; Kuiper
et al., 2006). Other researchers have been unable to detect alteration in Cyp1A1 or Cyp1A2 mRNA expression in response to
BDE-47, -99 or -209 (Pacyniak et al., 2007). Neither Cyp1A1 enzyme activity (ethoxyresosrufin O-deethylase; EROD) nor mRNA
expression were detected in MCF-7, H4IIE or HepG2 cells exposed
to BDE-47, -77, -99, -100, -153, -154, -183, or -209 (Peters et al.,
2004). No increase in EROD activity was detected when primary
hepatocytes of cynomolgus monkeys (Macaca fascicularis) were exposed to BDE-47, -77, -99, -100, -153, -154, or -183 (Peters et al.,
2006a). Since induction of Cyp1A is dependent on activation of
aryl-hydrocarbon receptor (AhR), these results suggests that PBDEs
are not AhR agonists.
The most potent dioxins, and dioxin-like-compounds, which
are well known inducers of Cyp1A expression have a rigid planar
configuration allowing them to bind to and activate the AhR.
Some flexibly planar molecules such as some PCB congeners are
also able to attain a planar configuration and so have the ability
to bind and activate AhR although at considerably decreased
potency compared to dioxins. However, PBDEs are both noncoplanar and have Br substituted for Cl and are therefore unlikely
to act as AhR agonists. Indeed, studies have shown that PBDEs
bind to the AhR but do not activate signaling (Peters et al.,
2006a,b). The occurrence of polybrominated dibenzo-p-dioxins
(PBDD), PBDFs, and polybrominated biphenyls (PBBs) in commercial PBDE mixtures has been reported (Hanari et al., 2006; Kuiper
et al., 2006; Wahl et al., 2008), and it is likely that trace amounts
of these compounds could have been responsible for inductions
observed in some studies. These results are similar to those that
have been reported perviously for the analogous polychlorinated
biphenyl ethers (PCDE) where it has been reported that the PCDEs
up-regulated expression of Cyp genes and activities of their associated enzymes (Howie et al., 1990), but it was later found that all
of the activity attributed to the PCDEs was in fact due to trace
contamination of the research chemicals with polychlorinated
dibenzo-p-dioxins (PCDD) and polychlorinated dibenzo-furans
(PCDFs) (Koistinen et al., 1996).
Several Cyp genes are inducible by PBDEs. Increased hepatic
expression of Cyp2B and Cyp3A enzymes has been demonstrated
in rats exposed to PBDEs (Pacyniak et al., 2007; Sanders et al.,
2005; van der Ven et al., 2008; Zhou et al., 2001). Recently,
Pacyniak et al. (2007) reported that BDE-47, -99, and -209 activate
the pregnane-X-receptor (PXR) in mice and humans. The PXR is a
ligand-activated transcription factor belonging to the superfamily
of nuclear hormone receptors that is crucial to the regulation of
xenobiotic biotransforming enzymes, including Cyp3A and Cyp2B
(Kliewer et al., 2002).
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been isolated from at least 40 species of the red alga Laurencia.
Approximately 600 brominated compounds, including sesquiterpenoids, diterpenoids, triterpentoids, and C15 acetogenins have been
isolated from Laurencia (Suzuki and Vairappan, 2005). Marine
sponges are a source of organohalogen compounds including
brominated pyrroles, indoles, phenols and tyrosines. In some
cases, organohalogens isolated from sponges are actually produced by associated bacteria or microalgae (Faulkner, 2002). A
survey of five Antarctic sponge species (Kirkpatrickia variolosa,
Artemisina apollinis, Phorbas glaberrima, Halichondria sp. and
Leucetta antarctica) identified 146 brominated compounds, with
each sponge species containing at least 35 brominated compounds
of natural origin, and 14 brominated compounds were identified in
each of the 5 species (Vetter and Janussen, 2005). The marine
sponge Dysidea contains approximately 25 BDEs, including the
natural BDE 4,6-dibromo-2-(20 ,40 -dibromo) phenoxyanisole, which
has been found to occur at concentrations as great as 3.8 mg/kg
ww in blubber from pygmy sperm whale (Kogia breviceps) (Vetter
et al., 2002). Concentrations of this compound found in marine
mammals exceed concentrations of any anthropogenic contaminant, including PBDEs, PCBs, DDT, or other chlorinated pesticides
or their biotransformation products (Vetter et al., 2002).
3. OH-BDEs and MeO-BDEs in wildlife
3.1. Natural organohalogens
Organohalogen compounds have been detected in nearly all
environments worldwide. A variety of these organohalogen compounds are synthetic, having resulted from direct manufacturing
processes or as byproducts of industrial processes. The quantity
and diversity of anthropogenic organohalogen compounds, however, pales in comparison to that of organohalogens produced by
natural biogenic or abiogenic processes with more than 4000 natural organohalogen compounds having been identified (Gribble,
2003). Of these compounds approximately 2200 are organochlorines, 1950 are organobromines, 95 are organoiodines, 100 are
organoflourines, and a few hundred are organochlorobromines.
The marine environment is the richest source of biogenic organohalogens. Nearly all of the natural organobromines discovered
are produced by marine organisms, including bacteria, plants and
animals (Gribble, 2003). Since life most probably originated in
the oceans, and there are relatively large concentrations of iodine,
Br and chlorine (Cl) in seawater, it is no surprise that marine organisms have synthesized a wide range of organohalogen compounds
in response to their environment. Although Cl is more abundant
than Br in marine systems, marine organisms seem to produce
more brominated than chlorinated compounds. This may be due
to the greater concentrations of bromoperoxidase (BPO) relative
to chloroperoxidase (CPO) in marine organisms (Gribble, 1998,
2000) and it has also been hypothesized to be related to the lesser
energy of the Br-C than the Cl-C chemical bond (Wan et al., 2010b).
A variety of marine organisms produce organohalogens
(Gribble, 1998, 2000, 2003). Marine plants produce up to 25% of all
known natural organohalogens. Organohalogen compounds have
Br
3.2. Ortho-substituted OH-BDEs and MeO-BDEs
Naturally occurring MeO-BDEs and OH-BDEs share a common
structural feature in that the OH or MeO group is located in the
ortho position relative to the diphenyl ether bond (Athanasiadou
et al., 2008). 20 -OH-BDE-28, 20 -OH-BDE-68, 2-OH-BDE-123, 60 -OHBDE-17, 6-OH-BDE-47, 60 -OH-BDE-49, 6-OH-BDE-85, 6-OH-BDE90, 6-OH-BDE-99, 6-OH-BDE-137 have been structurally identified
(Fig. 2) and confirmed as natural compounds in either the marine
Br
OH
OH
OH
Br
Br
Br
2’-OH-BDE-68
Br
Br
Br
6’-OH-BDE-17
Br
Br
Br
Br
Br
OH
Br
OH
Br
Br
Br
Br
OH
Br
Br
Br
OH
Br
6-OH-BDE-90
Br
Br
OH
Br
6-OH-BDE-137
Br
Br
OH
Br
Br
Br
Br
OH
6-OH-BDE-85
2-OH-BDE-123
6’-OH-BDE-49 Br
Br
Br
Br
6-OH-BDE-47
Br
Br
Br
Br
Br
2’-OH-BDE-28
Br
Br
Br
Br
6-OH-BDE-99
Fig. 2. Structures of OH-BDEs identified as natural compounds. In each case the –OH group is located in the ortho position relative to the diphenyl ether bond.
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sponge (Dysidea herbacea) or its associated filamentous cyanobacterium (Oscillatoria spongeliae), red alga (Ceramium tenuicorne), or
green alga (Cladophora fascicularis) (Bowden et al., 2000; Fu et al.,
1995; Handayani et al., 1997; Malmvärn et al., 2005, 2008). These
compounds were also identified in wildlife at higher trophic levels
(Marsh et al., 2004). Profiles of OH-BDEs in marine organisms were
generally dominated by 6-OH-BDE-47 and to a lesser extent 20 -OHBDE-68 (Kelly et al., 2008; McKinney et al., 2006; Verreault et al.,
2005; Wan et al., 2009; Zhang et al., 2010).
Several ortho-substituted MeO-BDE congeners are also natural
products (Fig. 3). 6-MeO-BDE-47, 20 -MeO-BDE-68, 20 -MeO-BDE28, 6-MeO-BDE-85, and 6-MeO-BDE-137 have been isolated from
red algae, and marine sponges or their associated cyanobacteria
(Malmvärn et al., 2005, 2008). Using 14C stable isotope analysis
6-MeO-BDE-47 and 20 -MeO-BDE-68 isolated from North Atlantic
True’s beaked whales (Mesoplodon mirus) were determined to be
natural compounds (Teuten et al., 2005). These compounds were
also identified in archived whale oil collected in 1921, before production of synthetic PBDEs (Teuten and Reddy, 2007). Temporal
trends of 6-MeO-BDE-47 and 20 -MeO-BDE-68 in pike from Swedish
waters from 1967 to 2000 were different than those of PBDEs. The
greatest concentrations of the MeO-BDEs were detected in fish collected before 1970 while concentrations of PBDEs showed increasing trends up to the mid-1980s (Kierkegaard et al., 2004).
3.3. meta/para-Substituted OH-BDEs and MeO-BDEs
Numerous OH- and MeO-BDEs that have been identified in
marine wildlife have not been classified as natural products. The
following OH-BDEs have been identified in marine wildlife:
2-OH-BDE-75, 2-OH-BDE-123, 2-OH-BDE-153, 3-OH-BDE-47,
40 -OH-BDE-17, 4-OH-BDE-42, 40 -OH-BDE-49, 5-OH-BDE-47,
6-OH-BDE-47, 6-OH-BDE-49, 60 -OH-BDE-85, (Jörundsdóttir et al.,
2009; Kelly et al., 2008; Liu et al., 2010; Marsh et al., 2004;
McKinney et al., 2006; Routti et al., 2009; Verreault et al., 2005;
Wan et al., 2009; Zhang et al., 2010). The following MeO-BDEs have
been identified in marine wildlife: 2-MeO-BDE-28, 2-MeO-BDE123, 3-MeO-BDE-47, 4-MeO-BDE-17, 40 -MeO-BDE-17, 4-MeOBDE-42,
4-MeO-BDE-49,
40 -MeO-BDE-49,
5-MeO-BDE-47,
0
0
5 -MeO-BDE-100, 6 -MeO-BDE-17, 60 -MeO-BDE-49, 60 -MeO-BDE66, 6-MeO-BDE-68, 6-MeO-BDE-90, 6-MeO-BDE-99 (Haglund
et al., 2010; Jörundsdóttir et al., 2009; Liu et al., 2010; Kelly
et al., 2008; Marsh et al., 2004; Verreault et al., 2005, 2007; Wan
et al., 2009).
Several of these compounds listed above are ortho-substituted
and therefore may be natural products. The meta- and parasubstituted OH-BDEs have been reported to be biotransformation
OCH3
Br
products produced during PBDE exposure (Chen et al., 2006; Cheng
et al., 2008; Erratico et al., 2010; Hakk et al., 2002, 2006, 2009;
Kierkegaard et al., 2001; Malmberg et al., 2005; Marsh et al.,
2006; McKinney et al., 2006; Qiu et al., 2007; Stapleton et al.,
2009; Staskal et al., 2006). However, as MeO-BDEs have never been
identified in PBDE exposure studies the true sources of the metaand para-substituted MeO-PBDEs remain unclear. Given the
abundance of naturally brominated compounds the possibility that
these meta- and para-substituted MeO-PBDEs are natural products
must be considered. Recently, Wan et al. (2009) demonstrated that
6-OH-BDE-49 was formed via debromination of OH-PentaBDEs,
which can be produced via demethoxylation of MeO-PentaBDE
congeners. Therefore, it is possible that MeO-PentaBDE congeners
could also be an important contributor of para-substituted
OH-PBDEs.
4. Relationships among PBDEs, OH-BDEs, and MeO-BDEs
4.1. Relationships in wildlife tissues
Concentrations of PBDEs, MeO-BDEs, or OH-BDEs, or some combinations of each have been reported for several marine organisms.
A correlation between any two of these compounds has been suggested as a non-definitive indicator of a biotransformation relationship between compounds. In a survey of 15 MeO-BDEs in
blood plasma samples of Glaucous gulls (Larus hyperboreus) from
the Norwegian arctic strong correlations between 6-MeO-BDE-47
and BDE-47 and between 6-OH-BDE-47 and BDE-47 were reported
(Verreault et al., 2005). There were also strong correlations with
other BDE congeners. Those authors suggested that 6-OH-BDE-47
and 6-MeO-BDE-47 in glaucous gull plasma resulted from uptake
from natural sources (Verreault et al., 2005). A positive relationship
was observed between concentrations of 6-MeO-BDE-47 and BDE47 in beluga whale, ringed seal (Pusa hispida) and several seaducks
(common eider; Somateria mollissima sedentaria and white-winged
scoter; Melanitta fusca) from the Canadian Arctic (Kelly et al.,
2008). Statistically significant positive correlations were observed
between 6-MeO-BDE-99 and BDE-99, and between 60 -MeO-BDE66, 6-MeO-BDE-49 and 4-MeO-BDE-42 and BDE-47 in beluga
whale (Kelly et al., 2008). Concentrations of 6-MeO-BDE-47 and
20 -MeO-BDE-68 were strongly correlated with several PBDEs in
Japanese common squid (Todarodes pacificus) (Kim and Stapleton,
2010). Because no study has demonstrated MeO-PBDE formation
during exposures to PBDEs these positive correlations may instead
reflect similar bioaccumulation of PBDEs and MeO-BDEs in the species analyzed in these studies. In fact, in other studies correlations
between concentrations of BDE-47 and 6-MeO-BDE-47 in tissues of
OCH3
Br
Br
Br
Br
Br
Br
Br
Br
2’-MeO-BDE-68 Br
2’-MeO-BDE-28
Br
Br
Br
6-MeO-BDE-47
Br
Br
Br
OCH3
Br
Br
OCH3
6-MeO-BDE-85
Br
Br
OCH3
Br
Br
6-MeO-BDE-137
Fig. 3. Structures of MeO-BDEs identified as natural compounds. In each case the –OCH3 group is located in the ortho position relative to the diphenyl ether bond.
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marine animals collected in different global locations were not significant (Verreault et al., 2005; Haglund et al., 1997; McKinney
et al., 2006; Stapleton et al., 2006). This is consistent with the
hypothesis that PBDEs are not precursors of MeO-PBDEs.
Correlations between OH-BDEs and PBDEs or MeO-BDEs have
also been investigated. Concentrations of OH-BDEs were not correlated with concentrations of PBDEs in beluga whale (Kelly et al.,
2008). Concentrations of PBDEs, MeO-BDEs and OH-BDE congeners
were measured in livers of tuna, five albatross species, and polar
bear from remote marine locations (Wan et al., 2009). No significant correlation between total PBDEs and total OH-BDEs was observed. However, concentrations of 6-OH-BDE-47 were correlated
with 6-MeO-BDE-47 and total concentrations of MeO-BDEs were
correlated with total concentrations of OH-BDEs in albatrosses
and polar bear. In tuna, there was no correlation between concentrations of OH-BDEs and MeO-BDEs. Concentrations of OH-BDEs
were also positively correlated with concentrations of bromophenols (BRPs), another class of natural organobromine compounds.
When samples from all species were considered together, a strong
positive correlation between MeO-BDEs and OH-PBDEs was evident and was made even stronger when BRP concentrations were
grouped with OH-BDEs in the analysis. Recently, a significant correlation between 6-OH-BDE-47 and 6-MeO-BDE-47 in anadromous
Chinese sturgeon from the Yangtze River was determined, while no
correlations were found between BDE-47 and 6-OH-BDE-47 or 6MeO-BDE-47 (Zhang et al., 2010).
The co-occurrence of OH-/MeO-BDE pairs in tissues might be
further support for the hypothesis that metabolic relationships
exist between OH-BDEs and MeO-BDEs. Such co-occurrences
have been reported in several studies. The following pairs, 6-OHBDE-47/6-MeO-BDE-47, 60 -OH-BDE-49/60 -MeO-BDE-49, 20 -OHBDE-68/20 -MeO-BDE-68, 6-OH-BDE-90/6-MeO-BDE-90 were identified in blood plasma of Salmon (Salmo salar) from the Baltic sea.
6-OH-BDE-47/6-MeO-BDE-47 and 20 -OH-BDE-68/20 -MeO-BDE-68
were also observed in livers of tuna, polar bear and albatrosses
(Wan et al., 2009). 3-OH-BDE-47/3-MeO-BDE-47, 40 -OH-BDE-49/
40 -MeO-BDE-49, and 6-OH-BDE-47/6-MeO-BDE-47 were detected
in blood plasma from glacous gulls and 40 -OH-BDE-49/40 -MeOBDE-49 in polar bear plasma collected from the Norwegian Arctic
(Verreault et al., 2005). 20 -MeO-BDE-68/2-OH-BDE-68, 6-MeOBDE-47/6-OH-BDE-47, 5-MeO-BDE-47/5-OH-BDE-47 were found
in Chinese sturgeon from the Yangtze River (Zhang et al., 2010).
Studies of red algae further support a relationship between naturally produced OH-BDE and MeO-BDEs (Malmvärn et al., 2005,
2008). In red alga PBDEs occur at approximately 5% of the concentrations of MeO-BDEs (BDE-47 = 0.1 ng/g dry weight versus 6MeO-BDE-47 = 2.4 ng/g dry weight). The observation that 20 -OHBDE-68 is present at a concentration that is 10-fold greater than
that of 20 -MeO-BDE-68 and total concentrations of OH-BDEs are
30-fold greater than total concentrations of MeO-BDE is consistent
with MeO-PBDEs being formed from OH-BDEs, not from PBDEs
(Malmvärn et al., 2008). Biotransformation of OH-BDE to MeOBDEs would be dependent upon O-methylation reactions. Such
reactions have been demonstrated in bacteria (Allard et al., 1987)
and fungi (Hundt et al., 2000). In addition, some marine macroalgae are capable of methylating mercury and lead (Pongratz and
Heumann, 1998). As further evidence for a natural production of
these compounds, several OH-BDEs and MeO-BDEs (20 -OH/MeOBDE-68 or 6-OH/MeO-BDE-90) have no known or likely PBDE precursors (Malmvärn et al., 2008).
4.2. Metabolic relationships between PBDEs, OH-BDEs and MeO-BDEs
The occurrence of OH-BDEs has been reported in several controlled PBDE exposure studies. However, concentrations of PBDE
used in laboratory studies have generally been in the lg/g ww
range but the OH-BDEs are detected at only trace concentrations,
often less than 1% of the concentration of the ‘parent’ PBDEs
(Hamers et al., 2008; Malmberg et al., 2005; Erratico et al., 2010;
Stapleton et al., 2009). In contrast, environmental concentrations
of OH-BDEs in marine organisms are generally in the pg/g to
ng/g range (Verreault et al., 2005; Kelly et al., 2008; Routti et al.,
2009). These concentrations are much greater than would be expected based on the measured concentrations of PBDEs in wildlife
tissues and the conversion rates of PBDEs to OH-BDEs observed in
laboratory studies. This is consistent with OH-BDEs found in marine tissue samples more likely originating from non-PBDE sources
than from PBDE precursors. Although OH-BDEs have been detected
in studies where animals were exposed to PBDEs, there also appears to be a metabolic relationship between OH-BDEs and MeOBDEs. Recent evidence strongly supports the hypothesis that
MeO-BDEs and OH-BDEs detected in wildlife are related. When
S9 fractions from Sprague–Dawley rats and microsomes from
White-leghorn Chicken and rainbow trout were exposed to either
BDE-99 or a BDE mixture no OH-BDEs were detected. However,
when the S9 fraction or microsomes were exposed to 6-MeOBDE-47 approximately 10% of the MeO-BDE was converted to
6-OH-BDE-47. When the same reactions were performed with a
mixture of MeO-BDEs a mixture of OH-BDEs was formed (Wan
et al., 2009). To further elucidate the relationships among PBDEs,
OH-BDEs, and MeO-BDEs, Japanese Medaka were exposed to each
compound via their diet (Wan et al., 2010a). Neither OH-BDEs nor
MeO-BDEs were detected in fish exposed to BDE-47. However, fish
fed 6-MeO-BDE-47 accumulated significant concentrations of 6OH-BDE-47 (approximately 6% of MeO-BDE-47 concentrations) in
liver and eggs collected over the duration of the exposure period.
Formation of 6-OH-BDE-47 from 6-MeO-BDE-47 but not from
BDE-47 was confirmed in hepatic microsomes from Japanese Medaka. The biotransformation ratio for the conversion of MeO-PBDEs
to OH-PBDEs (in vitro: 10%; in vivo: 6%) was 60- to 1000-fold greater than those between PBDEs and OH-PBDEs reported elsewhere
(<0.01–1%). Thus, production of OH-PBDEs from naturally
occurring MeO-PBDEs could be an important contributor for the
occurrence of OH-PBDEs found in wildlife.
The study by Wan et al. (2010a) also yielded interesting and
important information regarding the biotransformation of MeOBDEs. In Japanese Medaka exposed to 6-OH-BDE-47 significant concentrations of 6-MeO-BDE-47 were detected in liver, whole carcass
minus the liver, and in eggs collected over the course of the experiment. Biotransformation of 6-OH-BDE-47 to 6-MeO-BDe-47 did
not take place in isolated microsomes suggesting that this reaction
takes place in the extra-hepatic tissue. Conversion of OH-PBDEs to
MeO-PBDEs has been suggested previously (Haglund et al., 1997).
However, this is the first study to demonstrate generation of a
MeO-BDE from an OH-BDE (Wan et al., 2010a). The observation that
either 6-MeO-BDE-47 or 6-OH-BDE-47 can act as a precursor for the
other compound suggests that inter-conversion of OH-BDEs and
MeO-BDEs can take place in exposed organisms. Therefore, bioaccumulation of natural OH-PBDEs can contribute to the load of
MeO-BDEs and bioaccumulation of MeO-BDEs can contribute to
the load of OH-BDEs in marine organisms.
5. Toxicology of PBDEs and their analogs
The toxicology of PBDEs has been reviewed elsewhere
(Birnbaum and Staskal, 2004; Costa et al., 2008; Gill et al., 2004;
Legler and Brouwer, 2003) and is beyond the scope of this review.
However, given the structural similarities between PBDEs, MeOBDEs, and OH-BDEs it is interesting to compare their toxicities.
To this end an overview of some of the endocrine disrupting effects
of these compounds is presented here.
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5.1. Estrogenicity and androgenicity of PBDEs and OH-BDEs
The issue of environmental toxicants exhibiting endocrine disrupting properties has received a great deal of attention during
the past two decades. In particular, the issue of sex hormone disruption is a major concern. Some chemicals have structural similarities to the sex hormones 17b-estradiol (E2) and testosterone
(T) and are able to bind to the estrogen (ER) or androgen (AR)
receptors thereby altering cell signaling pathways. Some chemicals
act as endocrine disruptors by altering sex hormone steroidogenesis leading to altered circulating levels of these hormones.
Endocrine disrupting effects of PBDEs and OH-BDEs have been
reported. In particular, the hormone receptor agonist and antagonist properties of PBDEs and OH-BDEs have been investigated in
several in vitro studies. In vitro estrogenicity of PBDEs and OH-BDEs
using the T47D breast cancer cell line that is stably transfected
with an estrogen-responsive luciferase reporter gene construct
was investigated (Meerts et al., 2001). Of the eleven PBDEs that
exhibited dose-dependent estrogen receptor (ER) agonist activity
the most potent congeners (BDE-100 > -75 > -51 > -30 > -119)
were 250,000–390,000 times less potent than the natural ligand,
E2. However, 2 of the 3 OH-BDEs congeners had estrogenic activity
greater than any of the PBDE congeners used. Three of the PBDE
congeners, BDE-153, -166, and -190, each displayed antiestrogenic activity. Neither of the OH-BDEs assayed exhibited
anti-estrogenic activity. The results demonstrate that the presence
of an OH group greatly increases the endocrine disrupting potential
of PBDEs. The estrogenicity of several PBDEs and 6-OH-BDE-47 has
also been profiled in vitro (Hamers et al., 2006). In agreement with
the results presented by Meerts et al. (2001), less brominated
PBDEs exhibited estrogenic activity while greater brominated
PBDEs displayed anti-estrogenic activity. However, the most
potent anti-estrogenic compound assayed was 6-OH-BDE-47.
Several studies have demonstrated the potential of PBDEs, OHBDEs and MeO-BDEs to impact estradiol and testosterone synthesis. Concentrations of testosterone and estradiol were measured
in blood of rat pups exposed to BDE-99 during gestational days
10–18 (Lilienthal et al., 2006). The two BDE-99 doses (1 and
10 mg/g bw/day) resulted in dose-dependent decreases in estradiol
concentrations in male offspring at 3 weeks after birth. Even
160 days after birth testosterone concentrations in the 10 mg/kg
BDE-99-exposed male offspring were significantly less than in control animals. Using the H295R cell line, He et al. (2008) determined
the effects of several MeO-PBDEs on testosterone and estradiol steroidogenesis. Few changes in either testosterone or estradiol production were observed. Exposure to 6-MeO-BDE-85 slightly
increased testosterone production but had a greater effect on increased estradiol production. 6-MeO-BDE-137 increased testosterone production. Less production of estradiol was observed in cells
exposed to 20 -MeO-BDE-23, but no change in testosterone production was observed. Relative to OH-BDEs, the MeO-BDEs had a
greater effect on mRNA abundance of steroidogenic enzymes in
the H295R cell line (He et al., 2008). In particular, greater mRNA
abundance of cytochrome P450 11B2, which is responsible for biosynthesis of aldosterone, was measured in cells exposed to MeOBDEs compared to cells exposed to OH-BDEs.
are two main subtypes known as TRs: TRa mediates actions of
T3 in the brain and TRb mediates actions of T3 in the liver and other
tissues (Forrest and Vennstrom, 2000).
Because OH-BDEs are structurally similar to the thyroid hormones (Fig. 4) they have the ability to interact with the thyroid
hormone receptors. 4-OH-BDE-90 and 3-OH-BDE-47 are competitive inhibitors of T3 binding to TRa (Kitamura et al., 2008) and 4OH-BDE-90 significantly inhibits TRa and TRb controlled gene
transcription (Kojima et al., 2009). A recent study demonstrated
that OH-BDEs, but not PBDEs, significantly activate TRb reporter
gene expression. Of the OH-BDEs tested the naturally occurring
6-OH-BDE-47 was one of several congeners that were strong activators of gene expression (Li et al., 2010).
A regulated supply of T4 is required for a variety of physiological processes such as tissue growth, development, and differentiation, including brain development (Yen, 2001). The Metabolites of
the polyhalogenated aromatic hydrocarbons (PHAHs), including
hydroxylated polychlorinated biphenyls (OH-PCBs) hydroxylated
polychlorinated dibenzo-p-dioxins (OH-PCDDs) bind to the transport protein transthyretin (TTR) (Brouwer et al., 1990, 1998;
McKinney and Waller, 1994). By doing so the natural ligand thyroxine (T4) can be displaced, which leads to a decrease in plasma
T4 levels (Darnerud et al., 1996).
Due to their structural similarity to (T4) and triiodothyronine
(T3) the effects of PBDEs and OH-BDEs on the TH system have received considerable attention. Based on previous studies suggesting
that OH-BDEs are generated by microsomal biotransformation of
PBDE, Meerts et al. (2000) investigated interactions of PBDEs and
biotransformed PBDEs with human TTR, in vitro. Only after PBDEs
were incubated with microsomes isolated from rats exposed to
either phenobarbitol (inducer of Cyp2B), b-napthoflavone (inducer
of Cyp1A), or clofibrate (inducer of Cyp4A) were interactions with
the TTR observed. Two synthetic OH-BDEs, 2,6 dibromo-4-(2,4,6tribromophenoxyphenol) and 2-bromo-4-(2,4,6-tribromophenoxyphenol, designed to resemble T3 and T4, respectively, had greater
affinity for TTR than either T3 or T4. The naturally occurring 6-OHBDE-47 has a greater affinity than BDE-47 for TTR (Meerts et al.,
2001). A much greater binding affinity (several orders of magnitude) of both 40 -OH-BDE-49 and 6-OH-BDE-47 than either BDE-47
or 6-MeO-BDE-47 for TTR has also been demonstrated in herring
gulls (Larus argentatus) and glaucous gulls (Ucán-Marín et al.,
I
I
OH
CH2
O
CH
C
NH2
O
OH
I
I
3,5,3’,5’ - Tetraiodothyronine (thyroxine, T 4)
I
I
5.2. Effects on the thyroid hormone system
OH
The thyroid hormone system is a well known target of environmental toxicants (Brouwer et al., 1998; Brucker-Davis, 1998). The
thyroid hormone thyroxine (T4) is produced in the thyroid gland
and is secreted into blood where it binds to transport proteins
and is delivered to target tissues where it is converted to its biologically active form T3. The actions of T3 are largely mediated by activation of the thyroid hormone receptors (TRs). In mammals there
CH2
O
CH
C
NH2
O
I
3,5,3’ - Triodothyronine (T 3)
Fig. 4. Structures of triiodothyronine (T3) and thyroxine (T4).
OH
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2009; 2010). In these gull species the para substituted 40 -OH-BDE49 had a greater affinity than the ortho-substituted 6-OH-BDE-47
(Ucán-Marín et al., 2009; 2010). Using a surface plasmon resonance
(SPR) biosensor-based screening method it was demonstrated that
several OH-BDEs (3-OH-BDE-47, 6-OH-BDE-47, 5-OH-BDE-47, 40 OH-BDE-49, 60 -OH-BDE-49, 6-OH-BDE-99) have high affinity for
both TTR and thyroid binding globulin (TBG), the primary T4 transport protein in humans, accounting for transport of 75% of T4. For
both transport proteins the OH-BDEs had a greater affinity than
the PBDEs (BDE-47, BDE-49, BDE-68, and BDE-99) (Marchesini
et al., 2008). These same studies also demonstrated that 6-MeOBDE-47 had slight affinity for TTR and TBG while 20 -MeO-BDE-68
had a slight affinity for TBG. In each case the affinity of the MeOBDEs was comparable to the affinity of the PBDEs.
Given the importance of a regulated supply of T4 any disruption
could be detrimental to organism health. This is particularly true in
the case of the developing human fetus. OH-BDEs have been detected in people from the United States (Qiu et al., 2009), Nicaragua
(Athanasiadou et al., 2008), The Netherlands (Meijer et al., 2008),
China (Yu et al., 2010), Japan (Kawashiro et al., 2008), Spain
(Lacorte and Ikonomou, 2009), and South Korea (Wan et al.,
2010c). In several of these cases the OH-BDEs were detected in
pregnant women and their fetuses (Kawashiro et al., 2008; Qiu
et al., 2009; Wan et al., 2010c). OH-BDEs, as well as MeO-BDEs,
have also been detected in milk from nursing mothers (Lacorte
and Ikonomou, 2009). The naturally occurring 6-OH-BDE-47 has
been detected in South Korean and American women and in each
of their matching fetuses. Interestingly, the concentrations were
greater in the fetal serum samples than in the maternal samples
(Qiu et al., 2009; Wan et al., 2010c). This is a particularly interesting finding because the developing fetus might be more sensitive
to the effects of chemicals than the adult (Uchida et al., 2002).
The mechanism(s) whereby OH-BDEs accumulate in the developing fetus are unknown but evidence suggests that TTR might be
involved. TTR is able to cross the placental membrane thereby
facilitating delivery of T4 to the developing fetus (Brouwer et al.,
1998). As mentioned above, 6-OH-BDE-47 has a stronger binding
affinity for mammalian TTR than does T4 (Marchesini et al., 2008;
Meerts et al., 2000). In addition, placental cells synthesize TTR
(Mckinnon et al., 2005), and this TTR is secreted and binds T4
resulting in increased internalization of the TTR–T4 complex
(Landers et al., 2009). As such, placental TTR may also be responsible for the accumulation of 6-OH-BDE-47 by the developing fetus.
This of particular interest because synthesis and secretion of fetal
T4 by the hypothalamus-pituitary-thyroid (HPT) axis does not
initiate until the end of the first trimester of development (Calvo
et al., 2002). Consequently, disruption of the supply of maternal
T4 during the first 16 wks of gestation could have a number of
consequences including disruption of neurological development
(Morreale de Escobar et al., 2000, 2004). Fortunately, the concentrations of 6-OH-BDE in fetal blood reported to date were 4-fold
less than the concentrations required to affect T4 interactions with
TTR (Meerts et al., 2000; Marchesini et al., 2008).
6. Conclusion and future directions
The issue of PBDE toxicity is set to remain with us well into the
future. The stability of these compounds coupled with ongoing
usage of the deca-BDE formulations and the number of PBDE containing products that will be recycled or discarded in the near future assures that environmental levels will continue to rise. The
aquatic environment and the oceans in particular are the ultimate
sinks for these chemicals. As such, it is important that we continue
to explore all facets of PBDE toxicology, from environmental fate to
adverse health effects in exposed organisms. An important compo-
nent of this research is a clear understanding of PBDE biotransformation. Importantly, there is a need to clearly establish whether
the toxicologically more relevant OH-BDEs are metabolites of
PBDEs or of natural origin. Currently, the combination of low levels
of OH-BDEs claimed to be metabolites of PBDEs, and the failure to
detect these metabolites in many exposure studies, cast considerable doubt as to the origin of OH-BDEs detected in controlled laboratory settings.
Structural analogs of PBDEs, namely OH-BDEs and MeO-BDEs,
have been identified as natural products produced by a myriad of
marine invertebrates. Relatively great levels of these compounds
have been identified in numerous species throughout the marine
food web. An increasing number of studies aimed at identifying,
quantifying, and characterizing the toxicological effects of these
compounds have appeared in the literature over the past decade
and more studies will be required to understand their role in the
marine food web. Recent work suggests that the OH-BDEs found
in marine organisms, and in humans, may not originate solely from
biotransformed PBDEs. Rather, greater proportions of OH-BDEs
might be contributed from biotransfomation of the naturally
occurring MeO-BDEs.
The toxicity of OH-BDEs has been investigated and for some
endpoints they are more potent than PBDEs. We have briefly highlighted the endocrine disrupting activity of OH-BDEs compared to
PBDEs. MeO-BDEs and OH-BDEs are found in everyday dietary
items, including marine fish, and mammals, and even in fish oils
consumed as nutritional supplements. Given the propensity for
MeO-BDEs to be converted to toxic OH-BDEs, determining whether
or not PBDEs are biotransformed to OH-BDEs will be an important
step in understanding the toxicity of PBDEs.
Acknowledgements
Preparation of this manuscript was supported by a Discovery
Grant from the National Science and Engineering Research Council
of Canada (Project 326415-07) and a grant from Western Economic
Diversification Canada (Projects 6578 and 6807). J.P.G. was supported by the Canada Research Chair program, an at large Chair
Professorship at the Department of Biology and Chemistry and
State Key Laboratory in Marine Pollution, City University of Hong
Kong, the Einstein Professor Program of the Chinese Academy of
Sciences and the Distinguished Visiting Professor Program of King
Saud University.
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