PBDEs in the Environment Time trends, bioaccumulation and the identification

PBDEs in the Environment
Time trends, bioaccumulation and the identification
of their successor, decabromodiphenyl ethane
Amelie Kierkegaard
Doctoral thesis
Department of Applied Environmental Science
Stockholm University
Stockholm 2007
© Amelie Kierkegaard, Stockholm 2007
ISBN 91-7155-410-6
Typesetting: Intellecta Docusys
Printed in Sweden by Intellecta Docusys, Västra Frölunda 2007
Distributor: Stockholm University Library
ii
Abstract
Polybrominated diphenyl ethers (PBDEs) are important chemical flame retardants, but also environmental pollutants. Their behaviour in the environment is a function of their inherent molecular properties, largely governed by
the number and character of the bromine atoms substituted, and the microenvironment where they reside. In this thesis different aspects of importance
for the understanding of the behaviour of lower brominated and higher brominated PBDEs in the environment are addressed.
The contamination of a Swedish freshwater system with lower brominated BDEs was assessed by a retrospective study of pike from Lake Bolmen covering the time period 1967 to 2000. The concentrations of tetra- to
hexaBDEs increased exponentially up to the mid-1980s and then leveled
off/decreased slowly, possibly reflecting the voluntary reduction in production and usage of lower brominated BDEs in Europe. Methoxylated PBDEs
were found to be present in similar concentrations to the PBDEs. However,
there was no correlation between the levels of the two substance groups, and
it was therefore concluded that they originated from different sources.
To understand the low abundance of higher brominated BDEs in wildlife
despite their extensive use and high levels in e.g. sediment, the dietary uptake of the fully brominated BDE, BDE209, was studied in fish. Although it
was not expected to be taken up due to its large size and hydrophobicity, it
was absorbed to a small extent via the diet. Once absorbed, BDE209 was
reductively debrominated to nona- to hexa-brominated BDE congeners.
Reductive debromination in vivo was also demonstrated in dairy cows exposed to higher brominated BDEs in their natural diet. The transfer of
BDE209 to milk was low (< 0.2 %). In contrast to PCBs and lower brominated BDEs, there was no equilibrium between adipose tissues and milk fat,
and for congeners with a log Kow > 7 a progressively smaller fraction of the
ingested PBDEs was transferred to the milk. The results indicate that while
lower brominated BDEs are excreted in the milk of dairy cows exposed to
PBDEs, the higher brominated BDEs are accumulated in the meat.
At the same time that PBDEs were receiving increasing regulatory attention, the next generation of brominated flame retardants was introduced. In
this thesis decabromodiphenyl ethane, a replacement for the technical
BDE209 formulation, was identified for the first time in the environment.
This thesis identified differences in uptake, metabolism and excretion for
brominated compounds compared to the previously thoroughly characterized
organochlorines. This knowledge will be useful for future risk assessments
given the ongoing use of these brominated aromatic compounds.
iii
“To dare is to lose one's footing momentarily.
Not to dare is to lose oneself.”
Sören Kierkegaard
iv
List of papers
This thesis is based upon the following papers which are referred to in the
text by their Roman numerals.
I
Polybrominated diphenyl ethers (PBDEs) and their methoxylated derivatives in fish from Swedish waters with emphasis on
temporal trends, 1967-2000.
Amelie Kierkegaard, Anders Bignert, Ulla Sellström, Mats Olsson,
Lillemor Asplund, Bo Jansson and Cynthia A. de Wit
Environ. Pollut., 2004, 130, 187-198.
II
Dietary uptake and biological effects of decabromodiphenyl
ether in the rainbow trout (Oncorhynchus mykiss).
Amelie Kierkegaard, Lennart Balk, Ulla Tjärnlund, Cynthia de Wit
and Bo Jansson
Environ. Sci. Technol., 1999, 33, 1613-1617.
III
Fate of higher brominated diphenyl ethers in lactating cows.
Amelie Kierkegaard, Lillemor Asplund, Cynthia A. de Wit,
Michael S. McLachlan, Gareth O. Thomas, Andrew J. Sweetman,
and Kevin C. Jones
Environ. Sci. Technol., 2007, 41, 417-423
IV
Identification of the flame retardant decabromodiphenyl ethane
in the environment.
Amelie Kierkegaard, Jonas Björklund and Ulrika Fridén.
Environ. Sci. Technol., 2004, 38, 3247-3253.
The papers are reprinted with the kind permission of the publishers, paper I
by Elsevier and papers II-IV by the American Chemical Society.
I, Amelie Kierkegaard, made the following contributions: in Papers I and IV, I was
responsible for the planning, chemical analysis, data evaluation, and writing the
manuscript. In papers II and III, I took part in the planning of the project, was responsible for the chemical analysis, data evaluation and manuscript writing. The
exposure, sampling and biological part of paper II was performed by Lennart Balk
and his co-workers. The mass balance study that provided the samples in paper III
was planned and conducted by Gareth Thomas and his co-workers. The other coauthors each made valuable contributions to the planning, data evaluation, and/or
writing of the manuscripts.
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Contents
Abstract .......................................................................................................... iii
List of papers................................................................................................... v
Contents ........................................................................................................ vii
Abbreviations ............................................................................................... viii
Introduction .....................................................................................................1
Objectives .......................................................................................................2
Brominated flame retardants...........................................................................4
PBDEs ........................................................................................................5
Properties and usage................................................................................................. 5
Environmental occurrence ......................................................................................... 8
Toxicity .......................................................................................................................8
Next generation BFRs ................................................................................9
Decabromodiphenyl ethane ....................................................................................... 9
BFR Look alikes .......................................................................................10
Methoxy-PBDEs....................................................................................................... 10
Analytical procedure......................................................................................12
Sample matrices.......................................................................................12
Extraction..................................................................................................13
Clean-up ...................................................................................................15
Instrumental analysis................................................................................16
Gas chromatography ............................................................................................... 16
Mass spectrometry................................................................................................... 18
Identification & quantification....................................................................20
Quality of the analysis ..............................................................................23
Results and Discussion.................................................................................27
Environmental levels - Temporal trends...................................................27
PBDEs...................................................................................................................... 27
Methoxy-BDEs ......................................................................................................... 30
Bioaccumulation .......................................................................................33
Dietary absorption.................................................................................................... 33
Biotransformation..................................................................................................... 36
Aquatic versus terrestrial environment .................................................................... 43
Risk assessment / Implications for exposure characterisation ................44
Decabromodiphenyl ethane – a next generation BFR .............................47
Conclusions...................................................................................................50
Acknowledgements .......................................................................................52
References....................................................................................................53
vii
Abbreviations
BAF
BCF
BFR
BMF
CV
DDT
DeBDethane
ECD
ECNI
ECS
EI
GC
GPC
HRMS
Kow
LOD
LOQ
LRM
LRMS
MeO-PBDE, MeO-BDE
MS
PBDE, BDE
PCB
PCDD/F
PLE
POP
SLE
SRM
STP
viii
bioaccumulation factor
bioconcentration factor
brominated flame retardant
biomagnification factor
coefficient of variation
2,2-bis(4-chlorophenyl)-1,1,1-trichloroethane
decabromodiphenyl ethane
electron capture detector
electron capture negative ionization
effective cross section
electron ionization
gas chromatography
gel permeation chromatography
high resolution mass spectrometry
octanol/water partition coefficient
limit of detection
limit of quantification
laboratory reference material
low resolution mass spectrometry
methoxylated polybrominated diphenyl ether,
methoxylated brominated diphenyl ether congener
mass spectrometry
polybrominated diphenyl ether, brominated diphenyl ether congener
polychlorinated biphenyl
polychlorinated dibenzo-p-dioxin/furan
pressurized liquid extraction
persistent organic pollutant
solid/liquid extraction
standard reference material
sewage treatment plant
Introduction
Generally, halogenation makes an aromatic hydrocarbon less volatile and
less water soluble, and also increases its stability. The chemical inertness of
halogenated hydrocarbons also makes them attractive in many industrial
processes. Unfortunately, resistance to degradation and lipophilicity are also
characteristics of persistent organic pollutants, POPs.
The halogen substituent has a strong influence on the physical-chemical
properties of halogenated hydrocarbons. While the small fluorine atom is
strongly bound to the carbon, the bond between the large iodine and the carbon is comparably weak. The resulting differences in the chemical properties
are reflected in the uses of halogenated hydrocarbons in technical applications. For instance, the stability of fluorinated hydrocarbons is a useful property in high temperature applications, whereas the thermal decomposition of
brominated hydrocarbons is essential for their function as flame retardants.
According to the Stockholm convention, POPs are defined as “chemicals
that remain intact in the environment for long periods, become widely distributed geographically, accumulate in the fatty tissue of living organisms
and are toxic to humans and wildlife” (1). The history of POPs includes a
number of chlorinated organic compounds, many of which are still present in
the environment worldwide, despite regulatory actions having been taken.
Examples are pesticides that were deliberately produced to be toxic, such as
DDT, industrial chemicals unintentionally released to the environment, such
as PCBs, and chemicals formed as by-products in manufacturing and combustion processes, such as chlorinated dioxins. The detrimental effects of
these chemicals were observed after decades of usage, when the environment
was already extensively contaminated. One of the goals of environmental
chemistry is to prevent this situation from arising again by providing a
knowledge base that allows the environmental behaviour of chemicals to be
foreseen or recognized at an early stage.
Although the use of many chlorinated hydrocarbons has decreased, the
development and use of hydrocarbons substituted with other halogens continues, with consequences for the environment that are largely unknown.
Examples of such compounds are the brominated flame retardants, of which
the polybrominated diphenyl ethers (PBDEs) are a prominent representative.
Of vital importance for the risk assessment of PBDEs is the manner in which
the bromine-carbon bond and changes in the degree of bromine substitution
influence their behaviour in the environment.
1
Objectives
The overall objective of this thesis was to improve the understanding of the
behaviour of brominated flame retardants in the environment. The results
obtained serve to identify potential environmental risks of using brominated
flame retardants (BFRs) and to assist environmental authorities in making
risk assessments. In this rapidly moving field, the research in this thesis was
conducted in the context of the state of the art at the time the work was done.
Paper I
A screening study of lower brominated diphenyl ethers, PBDEs, (up to
pentabrominated, denoted pentaBDE) in Swedish biota set the starting point
of this thesis (2). The results showed that these compounds were widespread,
and the highest concentrations were detected in aquatic ecosystems. In Paper I the following questions were addressed:
• Are PBDEs an increasing environmental problem in Sweden? Are the
levels of PBDEs in the environment increasing as a result of increasing
usage?
Two studies of retrospective temporal trends were initiated, one in pike
from Lake Bolmen and the other in guillemot eggs from Stora Karlsö in the
Baltic Proper (3). Both trends were reconstructed using archived material.
By the end of the 1990s another family of brominated compounds, methoxylated PBDEs (MeO-PBDEs), were identified in fish and seal from the Baltic
Sea (4). Their presence was also established in the pike samples from Lake
Bolmen and other Swedish freshwater systems. The concentrations of the
MeO-PBDEs in these samples were in the same range or higher than the
PBDEs. The structural similarities between the two groups were obvious, but
the origin of the MeO-PBDEs was unknown. By including the two most
abundant MeO-PBDEs in the study, it was possible to investigate if the temporal trends of the two groups of compounds were correlated to one another.
The question raised was therefore:
• Are PBDEs the source for MeO-PBDEs present in Swedish freshwater
systems?
2
Paper II
Other studies went on to show that tetra- and pentaBDEs are present in
biota all over the world. However, there were hardly any reports on the most
frequently used PBDE product, the Deca-mix formulation, containing
mainly the fully brominated congener, decabromodiphenylether or BDE209.
A follow-up study of sediment and fish was therefore performed (5) in the
same region where PBDEs were first detected in biota (6), the river Viskan,
which received wastewater from textile industries known to have used the
Deca-mix product. The sediment had high concentrations of BDE209, but
only traces were found in fish. These findings led to the questions:
• Why is the most frequently used BDE congener, BDE209, not found in
biota? Is it at all absorbed in fish? Could BDE209 metabolism/debromination be a source of the more abundant lower brominated
BDEs found in fish?
Paper III
The results showed that BDE209 was in fact absorbed from the gut in
rainbow trout, although to a small extent. However, further studies of bioaccumulation in eggs from peregrine falcons revealed a dominance of higher
brominated congeners (7), a pattern so far not seen in any organism from the
aquatic environment. This led to the hypothesis that the bioavailability of
higher brominated BDEs is greater in terrestrial organisms. To investigate
the fate of higher brominated BDEs, a mass balance was performed in a terrestrial organism of large importance for human exposure to many persistent
organic pollutants, namely the cow. Since occupational exposure studies had
shown that BDE209 was present in humans without known sources of exposure, an alternative exposure pathway was suggested to be via the food. The
questions addressed were:
• How do higher brominated BDEs behave in cows, exposed via their natural diet, regarding uptake, possible debromination and excretion to the
milk? Could dairy products and/or meat be an exposure route for higher
brominated BDEs to humans?
Paper IV
While an extensive research effort resulted in close regulatory scrutiny of
both the lower brominated and higher brominated BDEs, the search for other
persistent brominated chemicals in the environment continued. In a survey of
BFRs in Swedish sewage sludge (8) a brominated substance eluting after
BDE209 was detected, this led to the question:
• Is the next generation of heavy BFR already present in the environment?
3
Brominated flame retardants
The increasing use of flammable polymeric materials in a range of commercial and consumer products has increased the demand for fire protection.
Flame retardants are chemicals that, when added to materials, reduce their
flammability. Different types of chemicals act in different ways to reduce
flammability e.g. by cooling via the release of water, or by creating an uncombustible barrier on the surface of the material. Halogenated flame retardants inhibit fire mainly by scavenging oxidizing free radicals that are produced during combustion (9). The lack of oxidizing radicals prevents the
propagation of the fire. Although all halogens have this function, the brominated compounds are the most effective. They have a high scavenging efficiency, a low cost, decompose at lower temperatures compared to chlorinated flame retardants (the other major class of halogenated flame retardants),
and are compatible with a wide range of polymers (10,11). BFRs are particularly suitable for petroleum-based plastic and synthetic materials, and consequently they are common in home and office equipment such as casings for
electrical appliances, copiers, TVs, computers, and mobile phones. In fact,
90% of all electronic appliances produced contain BFRs (11). Other important areas of use for BFRs are in furniture upholstered with flame retarded
foams and textiles as well as insulating foams and other building materials.
BFRs can either be reactive, i.e. covalently bound to the polymer, or additive, meaning that they are blended in the polymer system. Another way to
introduce BFRs into the plastics is via the use of brominated monomers that
are mixed with the original monomers before the polymerisation process,
resulting in a polymer with built-in flame retardant properties. The most
common BFRs are tetrabromobisphenol A (TBBPA), PBDEs and hexabromocyclododecane (HBCD). The former is mainly used as a reactive BFR
while the latter two are additive products and therefore more prone to leak
from the material (12). This thesis focuses on the additive BFR with the
largest usage, namely the PBDEs.
4
PBDEs
Properties and usage
PBDEs are aromatic compounds substituted with up to 10 bromine atoms
(Figure 1). Like PCBs, the theoretical number of possible congeners is 209
and the same numbering system as suggested by Ballschmiter et al. (13) is
applied. The substitution pattern of the BDE congeners included in this
summary is given in Table 1.
Figure 1. Chemical structures of the major substances studied in this thesis, i.e.
PBDEs, MeO-BDEs and decabromodiphenyl ethane.
Commercial PBDEs are manufactured as three mixtures named after one of
their major components: Penta-mix, Octa-mix and Deca-mix formulations.
The industrial production is based on the bromination of diphenyl ether, a
process that is terminated at different degrees of bromination (14). Due to
5
the chemical properties of the oxygen directing the bromine to para and
ortho positions, and the steric hindrance of the substituted bromines, only a
limited number of congeners are present in the PBDE products (10,15). The
major components in the Penta-mix are tetra- and pentaBDEs, BDE47 and
BDE99 (Figure 1, Table 1) as major congeners, in Octa-mix they are heptaand octaBDEs, such as BDE183 (Table 1), and in the Deca-mix formulation
the main component is decabromodiphenyl ether or BDE209 (Figure 1,
Table 1) with small amounts of nonaBDEs.
Table 1. Substitution pattern of BDE congeners included in this summary.
BDE #
28
37
47
49
77
99
100
104
126
140
153
154
155
171
173
182
Bromine substitution
2,4,4’-triBDE
3,4,4'-triBDE
2,2’,4,4’-tetraBDE
2,2',4,5'-tetraBDE
3,3',4,4'-tetraBDE
2,2’,4,4’,5-pentaBDE
2,2’,4,4’,6-pentaBDE
2,2',4,6,6'-pentaBDE
3,3',4,4',5-pentaBDE
2,2',3,4,4',6'-hexaBDE
2,2’,4,4’,5,5’-hexaBDE
2,2',4,4',5,6'-hexaBDE
2,2',4,4',6,6'-hexaBDE
2,2',3,3',4,4',6-heptaBDE
2,2',3,3',4,5,6-heptaBDE
2,2',3,4,4',5,6'-heptaBDE
BDE #
183
188
190
194
196
197
198
201
202
203
204
206
207
208
209
Bromine substitution
2,2’,3,4,4’,5’,6-heptaBDE
2,2',3,4',5,6,6'-heptaBDE
2,3,3',4,4',5,6-heptaBDE
2,2',3,3',4,4',5,5'-octaBDE
2,2',3,3',4,4',5,6'-octaBDE
2,2',3,3',4,4',6,6'-octaBDE
2,2',3,3',4,5,5',6-octaBDE
2,2',3,3',4,5',6,6'-octaBDE
2,2',3,3',5,5',6,6'-octaBDE
2,2',3,4,4',5,5',6-octaBDE
2,2',3,4,4',5,6,6'-octaBDE
2,2',3,3',4,4',5,5',6-nonaBDE
2,2',3,3',4,4',5,6,6'-nonaBDE
2,2',3,3',4,5,5',6,6'-nonaBDE
2,2',3,3',4,4’,5,5',6,6'-decaBDE
PBDEs are applied in a wide range of products, such as thermoplastics used
in home or office furnishings, casings for electrical appliances, polyurethane
foam, and synthetic fabrics. The typical addition varies between 5-30% by
weight of the material to be flame retarded (14). The worldwide consumption of PBDE in 2001 (and 1999) was 67 000 tonnes, whereby Deca-mix
formulation accounted for 83% (16,17). The usage of the PBDE products
differs greatly between the continents. Compared to North America, both
Europe and Asia have no or a low consumption of the Penta-mix product and
in Europe also of the Octa-mix product (16,17). This is mainly due to differences in regulations and/or voluntary initiatives. The European Union has
banned the use of Penta-mix and Octa-mix formulations from 2004 (18) and,
at least in Japan, the industry has voluntarily phased out the lower brominated products (17). The Swedish import of technical Deca-mix is decreasing (see Figure 2) and a national ban on usage in for example textiles has
6
recently been adopted (2007) (19). However, most of the Deca-mix is imported in treated products and semi-finished goods (20).
Figure 2. Annual import of technical Deca-mix formulation to Sweden (tonnes)
(21). Imported Deca-mix incorporated in goods and products is not included.
The physical-chemical properties of the PBDEs are in many ways similar to
well known environmental contaminants, such as PCBs. The PBDEs are
highly hydrophobic compounds with logarithms of the octanol/water partition coefficients (log Kow) ranging from 5.08 to 8.70 for monoBDE to decaBDE (22). While the log Kow increases with the number of bromine substituents, the water solubility and vapour pressure both decrease. Hence, with
increasing degree of bromination there is a diminishing tendency for PBDEs
to be found dissolved in water or in the gas phase, they are rather sorbed to
particles in the air or in the sediment or soil. The ability for long range transport is high for lower brominated BDEs, whereas for higher brominated
BDEs the ability is linked to the distance that particles in the air are transported (22,23).
Although the bromine-carbon bond generally is weaker than the corresponding chlorine-carbon bond, PBDEs are regarded as persistent in the
environment. The physical-chemical properties of the BDE congeners differ
greatly, which leads to differences in their environmental behaviour. For
example, the reactivity or susceptibility to transformation measured as the
rate of the hydrolysis reaction with sodium methoxide has been shown to
increase with the degree of bromination (24). Furthermore, higher brominated BDEs are known to be photolytically degradable in the presence of
UV light, resulting in mainly lower brominated BDEs and brominated
dibenzofurans (25,26). So far, photolysis reactions have only been confirmed
in laboratory experiments, while for instance BDE209 in sewage sludge
7
showed no degradation 20 years after application of the contaminated sludge
to agricultural soil (27).
Environmental occurrence
The first finding of PBDEs in biota was reported in 1981 for pike from the
Swedish river Viskan (6). In the years that followed, the presence of PBDEs
was confirmed in biota from contaminated sites and from remote regions
(e.g. 2,5,28-34), indicating that they are global pollutants. In 1998 it was
reported that the level of PBDEs in human milk from Swedish mothers had
increased exponentially over the previous two decades (35). In the same
year, PBDEs were detected in sperm whales, indicating that PBDEs had
reached deep Ocean sea waters (36). After that, the number of investigations
on environmental levels of PBDEs worldwide increased exponentially. Several comprehensive reviews have been published that summarize the presence of PBDEs in biotic as well as abiotic environmental matrices (37-41).
The experience gained from previous studies on other POPs set the initial
focus on the aquatic environment. Generally, the dominant PBDEs reported
in biota were the tetra- to hexabrominated congeners. The most abundant
congeners are BDE47, BDE99 and BDE100, which also are the major congeners in the technical Penta-mix formulation (42,43), albeit with differing
composition. In sediment, dust and sewage sludge the major congener is
generally BDE209 followed by the congeners present in the Penta-mix product. The pattern is similar to that of the commercial mixtures, and thus reflects the usage and composition of the PBDE formulations (39).
Characterizing the levels and profiles of the PBDEs in the environment
was one of the objectives of this thesis, and is further discussed in the Results and Discussion.
Toxicity
The knowledge of PBDE toxicity is still insufficient to predict their potential
health risks. The interpretation of early studies is furthermore obscured because commercial mixtures were used and the effects seen were at least
partly due to more potent halogenated impurities in the products. The current
knowledge has been summarized in several reviews such as (14,44-47) and
will only briefly be presented below.
Generally, the acute toxicity of PBDEs is low and lower brominated
BDEs cause effects at comparably lower doses than higher brominated
BDEs (44). The most critical endpoint of PBDEs is the developmental neurotoxicity in mice that has been reported for a range of congeners (48-51)
including BDE209. Neonatal exposure during a sensitive period of brain
development in the fetus affected the spontaneous behaviour of the offspring
and caused impaired learning and memory functions after they reached ma8
turity. PBDEs also seem to have effects on thyroid hormone transport and
metabolism, suggesting them as potential endocrine disrupters (45). Studies
in rats and mice have shown reduced levels of thyroid hormone in the serum
after exposure to PBDEs (44,45,47). At high dosages, BDE209 has further
been shown to induce tumors in rats and mice (44).
Ecotoxicological studies in birds also show that environmentally relevant
doses of lower brominated BDEs cause immunosuppression, oxidative
stress, reduced thyroid function, and decreased vitamins A and E (52-54).
BDE47 and BDE99 were further found to have negative effects on the reproduction and development of the invertebrate, Nitocra spinipes (55).
Next generation BFRs
The growing concern about the global presence of PBDEs in the environment has led to a search for alternatives. One approach has been to decrease
the emissions of BFRs into the environment, beginning with the production
and following through to the disposal of the products. Closing the system at
the production sites, recycling of BFR treated products, and the use of polymers in which the BFR is covalently bound to the monomer are examples of
this. Another approach is to find environmental friendly replacements. Examples of BFRs that have been suggested to replace PBDEs are bis(2,4,6tribromophenoxy)ethane (BTBPE) for the Octa-mix formulation and decabromodiphenyl ethane (DeBDethane) for the Deca-mix. The latter is included
in this study and is described below.
Decabromodiphenyl ethane
Decabromodiphenyl ethane, or 1,2-bis(pentabromodiphenyl)ethane (see
Figure 1), was introduced to the market already in the mid-1980s (56) but
became commercially important as an alternative to the Deca-mix formulation in the early 1990s (57). This additive BFR was also designed to meet the
strict European regulations on maximum amounts of brominated dioxins/furans allowed in the product (57,58). The structural difference to
BDE209 with carbon linking the aromatic rings also reduces the potential for
producing dioxins or furans under pyrolysis conditions (59).
DeBDethane is marketed under the trade names SAYTEX® 8010 (Albemarle Corp.) and Firemaster® 2100 (Chemtura Corp.) and has the same
applications as the Deca-mix PBDE-product, i.e. as an additive to different
polymeric materials like high-impact polystyrene (HIPS), and in textiles.
Typical applications are in consumer electronics, such as TV cabinets, in
building and cable insulation, and in adhesives. The recommended additions
are the same as for the Deca-mix formulation. There are currently no figures
on the global consumption of DeBDethane, but with increasing concern
9
about the PBDEs, the use of DeBDethane is predicted to increase in the future. Europe has no production of its own, but the import in 2001 was estimated to be a few thousand tonnes, primarily to Germany (57). Furthermore,
in Japan there has been a clear shift in consumption away from Deca-mix to
DeBDethane (Figure 3) (17).
Figure 3. Annual consumption (tonnes) of Deca-mix and DeBDethane in Japan
from Watanabe et al. (17).
Based on the structural resemblance of DeBDethane to BDE209, its physical-chemical properties are assumed to be similar, i.e. a low volatility and
low water solubility. The inclusion of the ethane bridge between the aromatic rings makes it slightly more hydrophobic compared to BDE209, but
also gives the molecule conformational flexibility.
The knowledge on the toxicity of DeBDethane is scarce. The oral toxicity
in rats was low, possibly due to poor absorption efficiency of the compound
(60).
BFR Look alikes
Methoxy-PBDEs
A new group of brominated compounds structurally related to PBDEs was
initially reported in Baltic herring, salmon and grey seals (4). The substances, which were present at similar concentrations to the PBDEs, were
identified as methoxylated tetra- and pentaBDEs (4). The presence of MeOPBDEs was later confirmed in white-tailed sea eagle (61), in salmon (62)
and in red algae (63) from the Baltic region. Additionally, early findings of
10
MeO-BDEs in biota from other regions were reported in pilot and beluga
whales from Svalbard (64) and in dolphins, whales and dugongs from Australia (65,66). The common feature of the different MeO-PBDE congeners in
Swedish biota was the position of the methoxy group, i.e. ortho to the ether
bridge. The major congeners were identified as 6-methoxy-2,2’,4,4’tetrabromodiphenyl ether (6-MeO-BDE47) and 2’-methoxy-2,3’4,5’tetrabromodiphenyl ether (2’-MeO-BDE68) (see Figure 1) (62).
There is currently no known anthropogenic source of these compounds.
Unlike the PBDEs, both hydroxylated and methoxylated PBDEs are among
the wide range of naturally produced organobromine compounds (for a review see Gribble, 67). They have been identified in, for example, marine
sponges and green algae (68-70). An alternative hypothesis to natural synthesis is that they may be metabolites of PBDEs, formed either by metabolizing enzymes via methylation after an initial hydroxylation in vivo or via methylating microorganisms after excretion (4). Investigating the relationship
between the MeO-PBDEs and the PBDEs was one of the objectives of this
study, and this subject is further discussed in the section on environmental
levels – temporal trends.
11
Analytical procedure
The analytical methods used in this study were developed from a goaloriented perspective. As the experience in analysing a class of pollutants
increases, the demands placed on the analytical methods typically rise. As
the maturity of the analytical methods evolves, it becomes possible to make
more detailed and reliable statements about the environmental occurrence
and behaviour of the pollutants. An example of this development is the quantification in papers I, II and IV, which was accomplished by using the
commercial mixtures as external standards, while a large number of single
congeners was commercially available in paper III. Nevertheless, pioneering work is a prerequisite for stimulating interest in a new pollutant class and
mobilising the resources required for further analytical method development.
Generally, the chemical trace analysis of non-volatile organic environmental contaminants involves a number of steps of equal importance. These
are sampling, extraction, clean-up, chromatographic separation and detection, and finally identification and quantification. Even though all of the
steps are relevant, those parts that distinguish BFRs from other organic pollutants will be emphasized in the following discussion.
Sample matrices
The type of samples and the way the samples are collected are of importance
for the analytical results. The choice of sample matrix depends on the type of
question posed and should consider whether the samples are representative
of the biological population or the geographical area that is investigated. In
environmental monitoring these considerations are especially important,
since it is necessary to ensure comparability between samples from different
geographical locations and over time. A well-defined procedure for sampling/subsampling reduces the variation between parallel samples, thus facilitating the detection of differences due to the variables targeted in the design of the experiment/sampling program.
Most of the work in this thesis was concerned with biota. Sampling biota
requires knowledge about the ecology of the organism, e.g. whether it is
stationary or migratory, as well as many biological variables, such as age,
gender, sexual maturity and nutritional status. The influence of these variables on contaminant levels in biota and the consequences for biota sampling
12
have been thoroughly discussed by Bignert and colleagues at the Department
of Contaminant Research, Swedish Museum of Natural History (71,72). The
temporal trend in paper I was based on pike, a stationary predatory fish used
in the monitoring of PCBs and DDTs in freshwater systems (73,74). In the
feeding study (paper II), juvenile rainbow trout were used due to the welldocumented background knowledge from their use as experimental fish in
biochemical studies (75). Paper III involved samples of grass (silage) and
adipose tissues, organs, milk and feces from cows, which together allowed
the assessment of the total chemical input and output in the cows over the
sampling period. Whereas the tissues were possibly not perfectly representative due to a potential inhomogeneous PBDE distribution in the cows, the
milk fat and feces samples were generated from pooled samples that were
integrated over a defined time period.
For the purpose of screening “new pollutants” originating from the technosphere, sewage sludge may be a particularly sensitive sample matrix in the
sense that it integrates the emissions of chemicals that are used in society,
both from domestic sources, traffic and small industries (39). The screening
for DeBDethane in paper IV was accomplished with sewage sludge samples
from sewage treatment plants (STP) distributed all over Sweden (8,76).
Sediment is believed to be the major sink for highly brominated BDEs released into effluents (17,39). The sediment analysed in paper IV originated
from an area heavily polluted by Deca-mix (77) and therefore suspected to
represent a potential sink for emissions of DeBDethane.
Extraction
The physical-chemical properties of the PBDEs (and MeO-PBDEs) make
them extractable into organic solvents by methods typically used for traditional lipophilic POPs. Extraction, clean-up and separation/detection methods for PBDEs have been reviewed by Covaci et al. (78). Methods frequently used to extract PBDEs from solid matrices are liquid/solid extractions (LSE), either by sequential extractions after vibrating, rotating or sonicating the sample (5,79,80), or by a continuous solvent extraction as in
column (4,81) or Soxhlet extraction (79,82), whereby the latter uses hot solvents. The extraction efficiency is dependent on the power of the solvent, the
accessibility of the matrix to the solvent, and extraction time. A high
temperature will increase the efficiency of the solvent by decreasing its
viscosity (thereby increasing solute mobility) and increasing the solubility of
the analytes. However, the use of high temperatures and exposure to UVlight may confound the extraction of octa- to decaBDE due to their
propensity for degradation.
Other extraction methods used for PBDEs are pressurized liquid extraction (PLE) (83-85), supercritical fluid extraction (SFE) (86-88) microwave
13
assisted extraction (MAE) (89,90) and more recently, solid-phase microextraction (SPME), with or without prior solvent extraction (91,92). Unfortunately, only a few of the “new” extraction methods reported in the literature
include higher brominated BDEs. PLE, which lately has become a common
method for PBDE extraction, applies temperature and pressure to increase
the diffusive power of the solvents. Apart from being automated, it has the
advantage of the possibility to use on-line cleanup (84,93).
Generally, lipophilic compounds such as PBDEs (and MeO-PBDEs), are
associated with the lipids in the tissues and therefore extraction methods that
effectively extract lipids will be efficient in extracting the PBDEs as well.
The lipid content is usually determined gravimetrically, either as part of the
extraction clean-up (this study) or by applying a total-lipid method such as
Bligh & Dyer to an aliquot of the sample. The biota samples in the present
thesis (papers I-III) were extracted according to the cold LSE method originally described by Jensen et al. (94). The method comprises repetitive extractions by n-hexane/acetone and n-hexane/diethyl ether. The extraction
efficiency of the lipids, mainly triglycerids, in fatty fish such as herring or
rainbow trout (paper II), was comparable to the well known method of
Bligh & Dyer (95). Later, however, it was found to be less accurate (lipid
extraction efficiency ~ 75%) for lean fish such as pike. This was suggested
to be due to the increased proportion of phospholipids in the lean fish, and
the method was therefore modified to increase the extraction efficiency in
matrices with a low fat content (< 1%) (96). Nevertheless, the pike samples
in the time trend study (paper I) were all extracted by the original biased
method. Moreover, the fat content decreased significantly over the studied
period, from the late 1960s to 2000, implying that the lipid content was more
biased at the end of the period compared to the beginning. To increase the
comparability of the samples, the trend was therefore presented on a fresh
weight basis.
A modified version of the LSE original method by Jensen et al., where
acetone is replaced by isopropanol was used for the extraction of cow feces
in paper III (96). For the extraction of the low-level (with respect to PBDE)
samples, i.e. adipose tissue, organs and milk fat, all with a fat content > 1%,
the original solvents were used to reduce the contamination from isopropanol (97). The silage samples were Soxhlet extracted with dichloromethane.
Sediment, sewage sludge and mineral supplement samples in paper III
and IV were cold-solvent extracted with acetone/n-hexane according to
Nylund et al. (98), whereas for sediment acetone/toluene was used. The air
sample and the insulation tube samples were sonicated with dichloromethane
and n-hexane, respectively (paper IV).
14
Clean-up
The crude extracts of environmental samples normally contain nonpolar and
semipolar compounds that are coextracted in large quantities with lipophilic
target analytes, such as the PBDEs. The coextractants may coelute with the
analytes (precluding or falsifying identification), distort the chromatography
of the analytes, cause unwanted matrix-effects in the ion source, or even
change the performance of the GC column and injection system. The general
strategy of sample clean-up was to keep the number of steps as few and efficient as possible, both to reduce the loss of analyte and to minimize the procedural blanks, e.g. from the solvents used.
The bulk of coextractants present in the sample extract are removed by
destructive or non-destructive methods. Liquid partitioning with concentrated sulfuric acid or column “filtering” by impregnating the acid on silica
are destructive methods frequently used for lipid removal in biological samples. In case more labile compounds are to be analysed, a non-destructive
method, like gel permeation chromatography (separation mainly based on
molecular size) is preferred (78). Other non-destructive methods for lipid
reduction are the use of adsorbents, like aluminium oxide, florisil and to
some extent silica, applied singly or in combination. The capacities of the
latter two are low compared to alumina or treatment with sulfuric acid (78).
All substances analysed in this study are resistant to strong acid. Sulfuric
acid was used for bulk removal of lipids and other components in the majority of the samples in this study (paper I-IV). The lipid content in the samples varied from
< 1% in pike up to 100 % in the adipose tissue. In lowlevel samples where a small final extract volume was necessary, an additional sulfuric acid impregnated silica column was used to further reduce the
remaining lipids (paper I, III, IV). For the analysis of milk fat (paper III),
the degree of lipid removal needed for an accurate determination of BDE209
could not be achieved with reasonable amounts of sulphuric acid. Therefore,
saponification of the lipids with potassium hydroxide (1 M KOH in ethanol)
for one hour at 60 °C was used. This method is simple and powerful with
respect to lipid removal and keeps the solvent consumption low compared to
e.g. GPC, but it can also be hard on labile analytes. Degradation of octachlorodibenzofuran/dioxin and o,p’-DDT and p,p’-DDT have previously
been reported (99,100,100a). This is the likely explanation for the comparably low recovery (53%, coeffient of variation: 12%) recorded for the surrogate standard, 13C-BDE209 in milk fat (paper III).
Further clean-up usually involves fractionation of different classes of
compounds or additional removal of interfering compounds. In papers III
and IV an extra column was needed to remove acid-resistant nonpolar saturated hydrocarbons, present in varying amounts in the samples. These coextractants were removed by fractionation on a silica column, 1 g activated at
450 ºC over night (paper IV). In order to get a more reproducible deactiva15
tion of the silica, the activated silica was deactivated with water (2%) in
glass ampoules that were immediately sealed until use (paper III).
The presence of elemental sulfur in sediment and sewage sludge, often in
high concentrations (mg/g dry weight in typical sewage sludge), distorts the
chromatography and interferes in electron capture detection (101). Common
methods for sulfur removal are fractionation with GPC or the use of a metal,
usually elemental copper, to form metal sulfide (for an overview see 102).
The copper powder/granulates can be added to the extract or mounted on the
top of a silica column (82). The sulphur present in sewage sludge and sediment (paper IV) was removed by the use of tetrabutylammonium sulphite to
oxidize the sulfur to the water soluble thiosulfate (103). This method was,
however, not optimal for DeBDethane, since it caused degradation via debromination in standard solutions. This problem was probably less pronounced in the sewage sludge and sediment samples due to the presence of a
protecting matrix. For the objective of this study it was considered tolerable.
Instrumental analysis
Gas chromatography
The final separation/detection of most BFRs and look alikes such as MeOPBDEs is generally accomplished by gas chromatography/mass spectrometry (GC/MS) analysis and, in some cases, also by GC equipped with an electron capture detector (ECD). The instrumental aspects of the analysis of
PBDEs were recently reviewed by Stapleton (104). Due to the limited number of congeners present in the commercial products, a 30 m non-polar or
semi-polar column is normally sufficient for the separation of lower brominated BDEs in environmental samples. A comparative study of 126 BDE
congeners run on 7 capillary columns (~ 30 m) suggested DB-XLB to be the
best choice based on coelutions with other BDE congeners or other brominated substances known to occur in environmental samples (105). However,
the presence of unidentified congeners produced via abiotic or biotic debromination may necessitate further separation. Likewise, other brominated
substances of anthropogenic or natural origin such as MeO-PBDEs may
require more careful assessment of the chromatographic separation (see later
discussion on coelutions). The analysis of higher brominated BDEs, in particular BDE209, requires different instrumental conditions. For example, the
DB-XLB column suited for lower brominated BDEs was highly discriminating against BDE209 (106). The difficulty with the analysis of BDE209 is
that the temperatures needed for its vaporization in the injector and in the
column (if it is to elute within a reasonable time) overlap with the temperature range in which it is thermally degraded. Degradation in the column is
seen as a raised baseline before the peak, whereas degradation in the injector
16
results in the elution of lower brominated BDEs (mostly octa- to nonaBDEs)
as distinct peaks. To reduce the thermal degradation of BDE209, the residence time in the column as well as in the injector should be minimized
(106). A short column of 15 m or less with a thin stationary phase (0.1 µm)
is a good choice. The different GC conditions needed for the lower and
higher brominated BDEs also imply that a complete PBDE analysis is optimally performed on two separate columns.
The injectors usually applied for PBDEs are hot vaporizing injectors such
as splitless, pulsed splitless, programmable temperature vaporization (PTV)
or cold on-column. The degradation of higher brominated BDEs encountered
in splitless injectors is caused partly by the temperature itself, but also by
interactions with active surfaces in the injector. The transfer efficiency can
be low for splitless injectors, a fact that is beneficial for dirty samples in
trapping non-volatile components in the sample that may be detrimental to
the column. However, it also discriminates against high boiling analytes.
There is no discrimination in on-column injectors, but they require “clean”
samples.
Figure 4. Fraction of BDE209 (in %) degraded to nonaBDEs quantified using
HRMS with on-column injection or LRMS with splitless injection (3 min. at 280 ºC,
paper III). The results are the mean of 41 samples and 11 procedural blanks, respectively. Vertical bars represent the 95 % confidence interval for the mean of the
sum of the nonaBDEs.
In paper III the cumulative degradation of BDE209 over the whole analytical procedure was quantified using two different injection systems. Quantified on a molar basis, an average of 4 % of the total BDE209 content degraded to nonaBDEs in samples analysed with a cold on-column injector,
probably reflecting the degradation during extraction/cleanup (Figure 4).
17
When a splitless injector was used, the corresponding degradation was 7%,
whereby BDE207 and BDE208 accounted for the major difference in the
products formed. BDE206 was detected at similar yields in both systems,
which suggests that this degradation happened earlier in the analysis and was
not governed by thermal reactions in the injector. The formation of BDE207
and BDE208 seemed to be accentuated in the procedural blanks, possibly
because other components in the matrix shielded the analytes from exposure
to surfaces that could trigger catalytic reactions.
Although degradation of DeBDethane was encountered during extraction/clean-up, it appeared to be less sensitive to thermal degradation than
BDE209. Thus the conditions selected for BDE209 were applied to DeBDethane as well (paper IV).
Mass spectrometry
In a few investigations (e.g. 34) the electron capture detector (ECD), an inexpensive halogen selective detector, has been used in PBDE analysis. The
risk for coelutions is, however, large (107). Thus, today, a mass spectrometer
is almost always used.
The most frequently applied MS techniques for PBDEs are low resolution MS (LRMS), operated in either electron ionization (EI) mode or in the
electron capture negative ionization (ECNI) mode, or high resolution MS
(HRMS) operated in EI mode. Similarly to the organochlorines the brominated compounds, with the natural bromine isotope ratio of 50.5% of 79Br
and 49.5% 81Br, give rise to fragment ions with isotope distributions characteristic for the number of bromine substituents on the fragment ion. ECNI
involves the ionization of a reagent gas, like ammonia or methane that via
collision with the high energy electrons emitted from the filament generates
less energetic thermal electrons that subsequently ionize mainly electronegative compounds. The thermal electrons are captured by the analyte either in a
dissociative (1) or a non-dissociative (2) process.
AB + e-therm
A- + B
(1)
AB + e-therm
AB-
(2)
Brominated compounds generally form ions from a dissociative electron
capturing process, often dominated by the production of bromide ions.
Monitoring the bromide ions is one of the most sensitive detection methods
for PBDEs (28,108,109) with instrumental limits of detection (LOD, explained later in this section) of 7-400 fg for di- to heptaBDEs (110,111). The
detection is selective for brominated substances but gives no structural information and the identification relies solely on differences in the retention
time. The fragmentation patterns of brominated substances depend on their
18
chemical structure, but also on the temperature of the ion source, the electron
energy, the system pressure, the choice of reagent gas, and to some extent
also on the design of the ion source. The ion source parameters have been
optimized in several investigations for lower brominated BDEs (110,111)
and for BDE209 (112). The instrumental conditions used in paper I aimed
to achieve a robust method. Therefore the ion source was kept at a high temperature to favour a complete fragmentation of the analytes and to reduce the
condensation of heavy contaminants (in particular BDE209, 84) in the
source. The sensitivity drop encountered with methane as the reagent gas
(also reported for isobutene, 110) due to the deposition of carbonized material in the source was avoided by using ammonia, which instead has a cleansing function. The detection limits achieved with ammonia are similar to
those of methane (111). The instrumental conditions used for the analysis of
higher brominated BDEs and DeBDethane in paper II-IV were similar although different instruments were used.
A common alternative to ECNI is electron ionisation (EI) either in LRMS
or HRMS. The ionization is accomplished by the emission of high energy
electrons generated from the filament that collide with the analytes, producing characteristic molecular and/or fragment ions. Generally EI gives more
structural information about the compounds ionized. The analysis of PBDEs
in EI is more selective than ECNI since the ions monitored are [M-2Br]+ or
[M]+, which allows distinction between BDE congeners of different homologue groups as well as other classes of brominated compounds. The instrumental LOD for PBDEs analysed in LRMS-EI is however at least one order
of magnitude higher than in ECNI, with decreasing sensitivity from monoto heptaBDEs (110,111), although improvements have been reported for
LRMS-EI combined with large volume injection (113) as well as the use of
ion storage MS (114). Still, with HRMS-EI both the selectivity and the sensitivity are better than with LRMS-EI, and the instrumental LOD for lower
brominated BDEs are similar to those in ECNI-LRMS (115,116).
The major advantage of EI compared to other ionization methods is the
possibility to apply isotope dilution by the use of 13C-labeled internal standards, which improves the precision of the analysis. The use of isotope dilution is also possible in ECNI for BDE209 due to the abundancy of the
phenoxide ion, [C6Br5O]− (112). The formation of phenoxide ions in ECNI
was previously reported by Buser (109). Another advantage of phenoxide
ions over bromide ion detection is an enhanced signal-to-noise ratio due to
reduced chemical noise at higher masses (112). The occurrence of higher
mass fragments in ECNI decreases with decreasing number of bromine substituents, but by optimizing the source parameters to enhance the production
of phenoxide ions the instrumental LOD has been shown to approach that of
bromide detection for 14 of 39 tri- to heptaBDEs investigated (110). The
phenoxide fragments can also be used to differentiate between congeners
within the same homologue group, such as octaBDE with 3 and 5 bromines
19
in the two aromatic rings versus a congener with 4 bromines in each ring,
since different phenoxide ions are produced (43). Some of the unidentified
brominated substances detected in paper II were characterized by the isotope distribution of the phenoxide ions as discussed in the following section.
In paper III, octa- and nonaBDEs were quantified using larger fragments
in ECNI-LRMS (phenoxide ions) as well as in EI-HRMS ([M-2Br]+). A
prerequisite for the cow study was the achievement of low detection limits to
enable the quantification of a range of congeners in all samples collected.
The production of the larger fragments in ECNI varies depending on the
pattern and number of bromine substituents on the BDE congener. For example, the nona-BDE206 and the octa-BDE196 have a lower response compared to the nona-BDE207 and the octa-BDE197, respectively, possibly
indicating that the number of ortho-substituted bromines favours the response. Furthermore, La Guardia et al. (43) reported that hepta-BDE171,
which is substituted with 3 and 4 bromines in the two rings, produced detectable phenoxide ions from both rings, whereas no high mass ions were
detected from its homologues BDE190 and BDE173, both of which have 2
and 5 bromines substituted in the rings. A symmetrical octaBDE like
BDE197 (and hypothetically BDE194 and BDE202) or BDE209 gives a
greater response since only one phenoxide ion is formed, independent of
what half of the molecule is ionized. However, even though some octa- and
nona-brominated congeners were detected with satisfying sensitivity in
ECNI in paper III, others had a lower response. This, together with the less
disparate response factors within homologue groups in EI, favoured the
choice of HRMS-EI. For BDE209 however, ECNI was preferred due to a
higher sensitivity.
Identification & quantification
A prerequisite for an accurate quantification is the unambiguous identification of the peak by minimizing coelutions and the use of good internal and
external standards. As mentioned above, bromide ion detection in ECNI
makes no distinction between brominated substances. Thus, there is a risk
for erroneous results due to coelutions. Coelutions for congeners reported in
environmental samples that are present in the technical mixtures, are summarized in Table 2.
Among the most frequently occurring PBDEs in environmental samples,
some congeners are more subject to coelutions. For example, BDE28 and
BDE49 may coelute with other BDE congeners, and BDE99, BDE153 and
BDE154 with other BFRs or natural brominated compounds. However, for
BDE47, the most abundant congener in biota, no major coelutions have been
reported. Moreover, coelutions can be region or matrix specific. For in-
20
stance, the presence of the naturally produced MeO-PBDEs in samples from
the Baltic Sea should be considered in the quantification of BDE99 and
Table 2. Coelutions reported for congeners present in the technical mixtures (and
reported in environmental samples) detected with ECNI after separation on various
nonpolar to semipolar 30 m columns (i.d. 0.25 mm). Coeluting compounds detected
in environmental samples are marked in bold, numbers represent BDE congeners.
Data is summarized from a (105), b (43), c (117), d (3) e (78), f Paper I, g (62).
Coelutes with
BDE#
28
49
DB-1a
DB1-HT b
DB-5 a,c,d,e,f CP-Sil-8 g
DB5-HT b
HT-5 a
DB-XLB a
J&W Sci.
J&W Sci.
J&W Sci.,CAgilent
f
Chrompack
J&W Sci.
SGE Int.
J&W Sci.
33
16,33
33
16, 33, 38
71, 48
68
degr.prod of HBCD*
71, 48
68
16, 33
68,80
47
66
75
85
degr.prod of HBCD
42
51
6-MeO-BDE99
100
97
120
126
118
119
155
138
153
173
183
185
197
203
42
51
5-Cl-6-MeO-BDE47
6'-Cl-2'-MeO-BDE68
2',6-diMeO-BDE68**
degr.prod of HBCD
99
154
46, 48, 68,
71
4'-MeO-BDE49
unknown br.‡
97
116
109
97
120
126
TBBPA
HBCD
181,190
190
BB153
Me-TBBPA
unknown br.
BB153†
Me-TBBPA††
190
BB169
166
HBCD
190, 171
190
BB169
204
198
175
192
204
198
182
* Degradation products from hexabromocyclododecane, ** named BC-11 by the author (117),
† Bromobiphenyl 153, †† Dimethyl tetrabromobisphenol-A. ‡ unknown brominated substance
possibly BDE100 (62). Furthermore, pentabromoethylbenzene may erroneously be assigned as BDE37 on a DB5 column (L. Asplund, personal
comm.). A minor coelution was encountered for BDE154 (paper I), which
is why height instead of area was used in the quantification. The possible
coelutions among the higher brominated BDEs, mainly hexa-, hepta- and to
some extent octaBDEs, are less known since the number of commercially
available congeners still are limited compared to lower brominated BDEs.
21
Coelution problems in EI are restricted to congeners within the same
homologue groups and to other compounds brominated or non-brominated
with interfering mass fragments/molecular ions present in the samples. For
the latter type of coelutions HRMS is highly selective. Still, isobaric ions of
for instance tetraBDEs [M-2Br]+, pentaPCB [M]+ and heptaPCB [M-2Cl]+
requiring a resolution of 82 000 and 23 000 respectively to separate, have
been reported (107,118). For instance, BDE47 coelutes with CB180
(2,2',3,4,4',5,5'-heptachlorobiphenyl) on a 30 m HP-5 column, and on a corresponding nonpolar column with CB191 (107). Other potential coelutions
are tetra- to hexabromodibenzofurans with the corresponding PBDEs. For
example [M+4]+ from tetraBDF and [M+6]+ from penta- to hexaBDFs respectively, require a resolution of about 30 000 to differentiate from the corresponding [M+2]+ and [M+4]+ of the tetra- to hexaBDEs (119). However,
since the isomer distribution is two amu higher for the BDE congeners, the
ratio between the two most abundant ions should reveal major contributions
from the corresponding brominated furan. Regarding coelutions with other
BDEs however, only one of those presented in Table 2, namely BDE155
and BDE126, represent interference between homologues whereas the remaining are within homologues and therefore valid also for HRMS (105).
Due to the limited number of congeners available, potential coeluting isomers of the hepta- and octaBDEs quantified both in EI and ECNI in paper
III could therefore not be excluded. BDE203 in the samples may thus potentially be assigned to BDE198, BDE197 to BDE204 and BDE173 to BDE190
on a 15 m short column (see Table 2).
A good internal standard (surrogate standard) should have physicalchemical properties similar to the analyte. An ideal choice is therefore 13Clabeled standards because they perfectly mimic losses during the analytical
procedure. However, the use of 13C-labeled standards is as previously mentioned restricted to MS-EI or for BDE209 in ECNI. Another approach is to
establish the differences in behaviour between analyte and the surrogate
standard of choice in the analytical procedure, by recovery experiments.
Examples of surrogate standards used in ECNI are less abundant BDEs, such
as BDE77 (115), BDE104, BDE140 (120) or for example decabromobiphenyl for BDE209 (79)(for additional examples see 78). Monofluorinated
derivatives of PBDEs comprise another alternative having similar properties
but eluting before the corresponding BDE congener (ortho- and metasubstituted) (121). The internal standard in paper I, 2,2',5,6'tetrachlorobiphenyl (CB53), was the original internal standard used in the
monitoring of PCBs and DDTs (122). CB53 was far from ideal for PBDE
analysis, but it was a prerequisite for using the archived extracts. The extraction/clean-up method was the same over the studied period and the relative
recoveries between the surrogate standard and the BDE congeners were established. At the time of the BDE209 feeding study (paper II), no 13Clabeled BDE209 was commercially available. The choice of dechlorane was
22
based on its long term use in previous PBDE investigations (2,5).
Dechlorane was also used in the preliminary quantification of DeBDethane
(paper IV), which enabled direct comparison to previous results of BDE209
from a larger survey of sewage sludge samples (8). In paper III, the labeled
congeners 13C-BDE183 and 13C-BDE209 were used.
For a long time, PBDEs in the environment were quantified against the
technical products due to the lack of single congeners. The Penta-mix,
Bromkal 70-5DE, the Deca-mix formulation, Dow FR-300BA, and the technical DeBDethane product Saytex 8010®, were used in paper I, II and IV,
respectively. To increase the accuracy of quantification in retrospective
analysis (paper I), the Bromkal 70-5DE product was characterized and the
major components quantified (42). Further characterization of the commercial mixtures has recently been published for the purpose of identifying the
source of congeners present in the environment (43,105). The two MeOBDEs present in pike (paper I) were initially quantified using the response
factors of BDE47, which was the closest eluting BDE congener with the
same degree of bromination. A conversion factor was later calculated for the
retrospective correction of the original quantifications.
PBDEs are believed to accumulate in the body lipids like many classic
organic pollutants. In this case the concentrations normalized to lipid weight
will enable the levels in tissues and organs with different fat content or even
between different species to be compared. In some cases it may still be more
relevant to normalize to fresh weight, e.g. for food analysis aimed at human
exposure. In the present thesis, the intention was to normalize to fat content,
but due to several factors, the concentrations in paper I and II were expressed on a fresh weight basis. In paper I the fat content decreased from
about 0.8% to 0.4% from the late 1960s to 2000. Apart from the problems
encountered with the low extraction efficiency as previously discussed, there
was no correlation for pike sampled in a given year at a given location, between lipid content and fresh weight concentration of the PBDEs. This fact
by itself is an argument against lipid weight normalization (123). In paper
II, the lipid content of the muscle tissue decreased during the experiment
whereas the lipid content in liver was about the same. To remove decreasing
lipid content as a confoundry factor in the uptake curve of BDE209 a conservative approach was taken and concentrations were expressed on a wet
weight basis.
Quality of the analysis
Establishing a good quality control and practice for the analytical procedure
is necessary for the comparison of analytical data of different origins. A
number of criteria and rules for the analysis of PBDEs in biota have recently
been suggested (124) and the results from four interlaboratory studies show
23
satisfactory results for lower brominated BDEs, while the higher brominated
BDEs (BDE183, BDE209) still need attention (125,126). The accuracy of
the method can be determined by analysing standard reference material
(SRM). There are currently seven SRMs with certified and reference values
for PBDE congeners (127) and another two candidate SRMs (128). The
SRM matrices are cod liver oil, fish tissue, whale blubber, human serum, and
flounder (candidate SRM), all of which contain tri- to heptaBDEs. Housedust (SRM) (129) and sediment (candidate SRM), also contain high levels of
BDE209. In the following, the special measures taken for the analysis of
BDE209 are emphasized.
A Recovery experiment based on spiked samples, e.g. 13C-labeled recovery standards, quantified against an injection standard is a measure of how
much of the analyte is lost in the analytical procedures (absolute recovery).
When internal standards other than 13C-labeled are used, it is vital to investigate the recovery of the analytes compared to the recovery of the internal
standard (relative recovery) and if there is a large discrepancy to correct for
the difference. Relative and absolute recoveries of the PBDEs versus CB53
were determined in spiked samples at two concentration levels (paper I).
The relative recoveries were considered acceptable (on average 97-106%),
i.e. within the range of the expected variation from parallel samples and
were therefore not corrected for. The relative recoveries of dechlorane used
in paper II and IV were 110-113 % for tetra- to hexaBDEs in fish (paper
II, unpublished) and 114 % for BDE209 (5) and 89% for BDE209 in sediment (unpublished). Despite the somewhat different yield for the BDE congeners compared to dechlorane, corrections were not made. Dechlorane has
on the other hand later been applied in several interlaboratory studies and
has shown good results (126,130).
The extraction efficiency in solid matrices can not be evaluated using
spiked samples. Adding a surrogate standard to an environmental matrix can
as a rule not simulate the form in which the analyte is associated with the
matrix. A common way to examine the extraction efficiency is to successively extract the same sample using either the same method or different
methods, so called exhaustive extractions. The extraction efficiency for biota
and sediment with the methods used in this study was satisfactory for all
BDE congeners, with less than 1% remaining for all BDE congeners studied.
The use of a laboratory reference material (LRM) provides a measure of
the precision of the method or the within-laboratory reproducibility over
time. It also serves as a warning/alarm system for both occasional discrepancies and long term deviations. LRMs of herring and pike were included in
the temporal trend study from 1996. The coefficients of variation for the low
level pike were 10-26 % and for herring it was 6-19 %. These are within the
expected range suggested by Horwitz et al. (131). LRM was also prepared
and analysed for the different matrices in paper III. However, with the two
24
extraction occasions per matrix performed, the data set was too small for any
statistical analysis.
One or several procedural blanks covering the whole or parts of the analytical procedure were analysed in parallel with the samples. In low level
samples the concentration in the blanks often dictates if the samples can be
quantified. Thus, a protocol for checking solvents and use of separate glassware is recommended. In paper III, two blanks for each extraction occasion
were used (containing on average 38 pg BDE209/blank), and for the milk fat
analysis three additional blanks were employed covering parts of the procedure. In all results presented (paper I-IV) the amount in the blanks was subtracted from the samples. The addition of a matrix, such as corn oil to the
blanks, may present an alternative to avoid losses due to the lack of a
“keeper” compared to the samples. This may also prevent a more extensive
degradation of BDE209 during extraction/cleanup as well as in the injector/column as previously discussed (Figure 4).
PBDEs can be expected in a laboratory environment equipped with computers and other electronic devices. Lower brominated BDEs have been reported in laboratory air (132), in offices (133), and high concentrations of
BDE209 have been detected in dust (134,135). Subsequently, all handling of
extracts exposed to the air should be minimized and the extracts should always be protected from particle deposition. The contamination of the solvents (with both higher and lower brominated BDEs) varies between batches
and needs to be tested before use. In general, solvent usage should be kept to
a minimum. All glassware was heated to 450 ºC overnight and solvent rinsed
before use. Other sources of BDE209 contamination are plastic materials
such as polyethylene bottles, insulating foams and lubricant oils present in
for instance blenders. For the same reason, unnecessary electric appliances
and upholstered furniture/chairs were avoided as well as unpackaging of
goods in the laboratory where extraction and clean-up took place.
The degradation of BDE209 warrants attention, not only the thermal degradation occurring in the GC/MS system but also the degradation induced by
UV light or in contact with active surfaces/materials in the analytical process. The laboratory was equipped with UV-filters on windows and the fluorescents lamps. The extracts were kept in brown glassware or covered in
aluminium foil. Despite these precautions, BDE209 degraded throughout the
whole procedure. The use of isotope dilution compensates for the loss of
BDE209, ensuring an accurate quantification, but a problem arises if
nonaBDEs and octaBDEs are quantified in the same samples. The major
products from the degradation of BDE209 are lower brominated BDEs
formed by a successive debromination. In paper III, an approach to trace
the degradation is presented. The surrogate standards 13C-BDE183 and 13CBDE209 were used for quantification of all higher brominated BDEs. By
tracing 13C-nona-, 13C-octa- and 13C-heptaBDEs (apart from 13C-BDE183)
formed by the degradation of 13C-BDE209 in the samples and subtracting a
25
proportional fraction from the native octa- and nonaBDEs, it was possible to
reconstruct and quantify the amount of octa- and nonaBDEs that was present
from the start. The correction was made assuming a first order degradation
process. A procedural blank containing only 13C-BDE209 was used to trace
the potential debromination to 13C-BDE183. Although the input of nonaBDEs from BDE209 was traced by this method, the concurrent degradation of
nonaBDEs to octaBDEs was not accounted for. Additionally, separate
calibration standard series for BDE209 and the lower brominated BDEs
were used. In general, the calibration curves used in this thesis were
prepared from standard solutions at 6-9 concentration levels, analysed 2-3
times in mixed order with the samples.
The lowest level that can be quantified is ultimately defined by the limit
of detection (LOD) achieved for the sample matrix with the applied method.
The International Union of Pure and Applied Chemistry (IUPAC) has defined the limit of detection as “the concentration or quantity derived from the
smallest measure that can be detected with reasonable certainty for a given
analytical procedure” (136). The smallest measure, XL, is given by the equation:
k is a factor chosen according to the level of confidence desired. A k factor
of 3 thus represents the amount where the signal is just distinguishable from
the background (with a confidence level of 99.7 % separated from the noise)
and is often used to define the LOD (137). The limit of quantification
(LOQ), representing the lowest level that can be reliably measured in a sample, requires a higher precision and a factor (k) of 10 is therefore recommended (137). A common simplification of the above method is to use a
signal to noise ratio (S/N) measured around the analyte retention in the
chromatogram. The LOQ concentration is estimated from the signal that
corresponds to 2 to 5 times the S/N ratio.
In the analysis of PBDEs, the LOQ for the major congeners in the technical products (BDE47, BDE99, BDE209) are often defined by their presence
in the procedural blanks. The LOQ in this thesis was determined for every
sample series based on the S/N or the blank levels. In papers I, III and IV it
was defined as 5 times the S/N, (corresponding to k > 10 in the IUPAC definition) or, if present in the procedural blanks, as 5 times that amount. In paper II the LOQ for BDE209 was defined as S/N > 3 due to a larger peak
width compared to later analyses, which was caused by instrumental differences. Despite the reduced height/area ratio, the peak was clearly distinguished from the noise. Measured values above the LOD but below the LOQ
were used in paper I and III in the statistical evaluations and in the figures,
respectively, as they represented the best available estimates.
26
Results and Discussion
This thesis made several contributions to the understanding that has been
garnered over the last 15 years about the behaviour of BFRs, in particular
PBDEs, in the environment. One focus was on the identification and assessment of BFRs as an environmental problem. For the lower brominated
BDEs, this included determining whether environmental levels were increasing and therefore a growing risk (paper I). For DeBDethane, a second generation BFR, the contribution was establishing its presence in the environment (paper IV). A second focus was in the field of bioaccumulation, in
particular the absorption, biotransformation, and excretion of higher brominated BDEs (papers II, III). Despite the differences between the systems
studied, aquatic versus terrestrial organisms, fish versus mammals, and
spiked food versus naturally contaminated food, certain common characteristics of bioaccumulation behaviour were identified. These will be discussed in
the following section, addressing:
• Levels and temporal trends of BDEs and MeO-PBDEs in the environment
• Dietary absorption, biotransformation and excretion of higher brominated BDEs in organisms from the aquatic and the terrestrial environment, and the resulting implications for environmental and human risk assessments
• Occurrence and potential impact of DeBDethane, a second generation BFR
Environmental levels - Temporal trends
PBDEs
The first papers reporting the presence of PBDEs in environmental samples
(in 1979) were from abiotic matrices such as sewage sludge, soil and sediment sampled close to manufacturing facilities in the USA (138,139). In
1981, PBDEs (tetra-hexaBDEs) were for the first time identified in biota (6).
The highly sensitive detection method of bromide ions in ECNI (108), enabled the detection of PBDEs in wildlife (seal, sea eagle and guillemot) from
remote regions suggesting that these compounds were global pollutants (28).
27
Other early reports of PBDEs in the environment included avian tissues and
eggs, bottlenose dolphins and human adipose tissue from the USA
(29,140,141) fish, shellfish, and sediment from Japan (30,142) and fish and
seals from Europe (31,32). A more comprehensive screening of Swedish
biota from aquatic as well as terrestrial ecosystems was performed in the
beginning of the 1990s (2). It revealed that PBDEs (BDE47, BDE99,
BDE100), like PCBs, were present in terrestrial organisms but more abundant in aquatic species.
Analysis of dated segments of a laminated sediment core indicated that
PBDE levels were increasing in Baltic sediment (98). Against this background, two retrospective temporal trend studies were initiated, one in guillemot eggs from the Baltic and the other in pike from Lake Bolmen, an area
without known sources of PBDEs in the south of Sweden (paper I). Both
trends were first presented in 1993 (143), showing increasing levels (sum of
BDE47, BDE99 and BDE100) from 1967 to 1991 in pike, whereas the guillemot trend indicated a decrease at the end of the period (1989-1992). This
was in contradiction to the previously reported PBDE trend in cod liver from
the North Sea (31) and in eel from the Rhine and Meuse rivers, but in concurrence with the increasing time trend of eel from the Roer River (144).
Since then, numerous papers have been published documenting the widespread distribution of PBDEs in remote regions as well as in “hot spot” areas
close to manufacturing plants/user sites. High PBDE concentrations, up to
60 ppm on lipid weight, have for example been reported in fish and earth
worms from Sweden (5,27), and in carp, rainbow trout, and Forster’s tern
eggs from Virginia, Washington, and California, respectively (USA) (145147), all collected from contaminated sites. Concentrations of similar magnitude, but with a larger fraction of higher brominated BDEs, were detected in
eggs from Nordic populations of wild peregrine falcons (7,148,149). High
levels have also been reported in a number of marine mammals, such as porpoise (150), bottlenose and white-beaked dolphins (151),(150) and pilot
whales (152). The presence of PBDEs in deep marine food webs (36,150) as
well as in biota from the Arctic (41) and Antarctic (153) regions confirms
their global scale distribution and long range transport. Extensive reviews of
environmental levels have recently been published (37-41,154).
The initial temporal trend measured in pike from Lake Bolmen, was later
extended to include the hexaBDEs, BDE153 and BDE154, and two methoxylated tetraBDEs (paper I). All BDE congeners showed increasing concentrations up to the mid-1980s, which then leveled off or decreased to the
end of the period, in 2000. The additional data from 2005 (Figure 5, unpublished data, K. Nylund) confirm the decreasing trend.
28
Figure 5. Arithmetic mean concentration (pg/g wet weight) of A) BDE47 and B) 6MeO-BDE47 in pike from Lake Bolmen, 1967-2005. One outlier with extremely
high concentrations in the sample series of 2005 was excluded (n=7).
The variation, both within and between the sampling years, encountered in
the pike study was extensive. Between years, the variation may be explained
by external factors like temperature or food supply. The within-year variation is to a small extent explained by the analytical uncertainty but is more
likely caused by variation in concentrations from food items and from individual physiological differences, e.g. in the absorption efficiency and in the
rate at which the compounds are metabolized and excreted. A way to reduce
the within-year and between-year variation is to normalize the data to the
concentration of a similar compound that is influenced by climatic or physiological differences in a similar way. For example, normalizing the levels of
BDE154 to CB153 markedly reduced the variation (Figure 6). CB153,
2,2',4,4',5,5'-hexachlorobiphenyl, is a legacy contaminant that has had stable
concentrations in this lake between 1988 and 2000 (data missing for 20012004). Even though the decrease of BDE154 was not statistically significant
by itself, the BDE154/CB153 ratio was. Thus, relative to CB153, BDE154
was decreasing. This reinforces the interpretation that BDE154, like the
other PBDEs, decreased between 1991 and 2005 in pike from Lake Bolmen.
The initial increase and peak of PBDE levels in pike from Lake Bolmen
(paper I) was similar to the trend observed in guillemot eggs from the Baltic
(3) and in roach from another Swedish fresh water system, Lake Krankesjön
(155). The more dramatic decrease in the guillemot trend may be explained
by the location of Lake Bolmen, it being closer to local sources of emission.
A recently published trend of PBDE levels in mussels from the English
Channel (1981-2003) was similar to the Bolmen-trend, with a peak at the
beginning of the 1990s followed by a leveling off or slow decrease (156).
29
Figure 6. Concentration ratios of BDE154 and CB153 in pike from Lake Bolmen,
1991-2005 (A). The ratios are based on the years for which both BDE154 and
CB153 were analysed in the same specimens (outlier of 2005 included), and show a
significant decreasing trend (log-linear regression curve (P = 0.0012). The corresponding concentrations (geometric means, 95% confidence intervals) versus time
plots for B) BDE154 and C) CB153 are also presented.
While the decrease likely reflects the voluntary reduction in production and
usage of PBDEs in Europe, temporal trends from other parts of the world
differ. For example, the levels of lower brominated BDEs in a range of
North American wildlife have increased exponentially (157-161) up to year
2000. In 1999 and 2001, the American continent accounted for 95% or more
of the global demand of the Penta-mix (16,17) and has up to now only voluntary agreements on reduction in its production and usage. These trends
have lately slowed down, while the levels in Arctic biota are expected to
continue to increase, partly due to the ongoing processes of long range transport to the Arctic regions (38). Thus, temporal trends for seals and beluga
whales from the Canadian Arctic showed no signs of leveling off (summarized by de Wit et al., 41). Likewise, the PBDE levels (including BDE209)
in peregrine falcon eggs from South Greenland increased continuously from
1986 to 2003 (149) although the exposure is partly from their migration to
Central and North America during the winter.
Methoxy-BDEs
The two tetrabrominated MeO-BDEs, 6-MeO-BDE47 (6-methoxy-2,2’,4,4’tetrabromodiphenyl ether) and 2’-MeO-BDE68 (2-methoxy-2’,3,4’,5tetrabromodiphenyl ether) (Figure 1), that were present in the pike samples
(paper I), had previously been found in fish, seals and predatory birds from
30
the Baltic region (4,61,162). The levels reported were all in the same range
as the PBDE levels, but in contrast to the PBDEs they had no commercial
application. The lack of any known direct anthropogenic source, and their
structural resemblance to the PBDEs led to several hypotheses, one of which
was that these compounds originated from the metabolism of PBDEs (4).
Salmon plasma was shown to contain hydroxylated PBDEs (162), which
could be interpreted as supporting this hypothesis. Furthermore, the corresponding MeO-BDEs were present in the neutral fraction of the plasma,
which indicated that they had a common source (162,163). Suggested biotransformation pathways are enzymatic oxidation followed by methylation,
enzymatic or by microorganisms in the intestine, or direct methoxylation by
microorganisms in e.g. the sediment (4). Hydroxylated metabolites formed
from PBDEs had previously been reported in fish and rodents (164-166).
Furthermore, methoxy-hydroxy-BDEs were detected in rats exposed to 14CBDE209 via the diet (166). However, although 6-OH-BDE47 was tentatively
identified in fish dosed with 14C-BDE47, its methylated form was absent
(poster of 164). Thus, even if 6-MeO-BDE47 is a plausible metabolite of
BDE47 via hydroxylation and subsequent methylation, there is no apparent
precursor for 2’-MeO-BDE68 (62,63), suggesting that there are other
sources of this compound.
Unlike PBDEs, both hydroxylated and methoxylated PBDEs belong to
the wide collection of naturally produced brominated organic compounds
(for a review see Gribble, 67). Both MeO-BDEs prevalent in the Baltic biota
and in Bolmen pike had previously been found in large quantities in tropical
marine sponges (66,69,167,168) and 2’-MeO-BDE68 has also been observed
in green algae from Japan (70), further supporting their natural origin. In
addition, high concentrations have been recorded in marine mammals from
the Southern Hemisphere in which only traces of PBDEs were detected (65).
Although MeO-PBDEs are naturally produced in sponges and algae from
the marine environment, it was surprising to also find them commonly present in fish from freshwater systems (paper I). The purpose of including the
MeO-PBDEs in the retrospective temporal trend study of PBDEs was to
investigate if there was any covariation between concentrations of the
PBDEs and the MeO-PBDEs that could support the theory that the MeOPBDEs were products of PBDEs. The results were clear, none of the MeOBDEs showed any correlation to the PBDEs, and the potential metabolite of
BDE47, 6-MeO-BDE47, had a decreasing trend over the time period, in
contrast to PDE47, which initially showed an exponential increase (Figure
5). A temporal trend in cod liver samples from the Norwegian Arctic coast
collected during 1987-1998 showed increased levels (a factor of 10) of
MeO-tetraBDEs in 1992 and 1993 compared to the years preceding and following (169). The high concentrations found in Bolmen pike in the late
1960s compared to later years suggest a similar variability in concentrations.
In conclusion, the results support the hypothesis that the two MeO-BDEs
31
were naturally produced. The natural origin of the same MeO-BDEs was
recently confirmed in True’s beaked whale from the North Atlantic by natural abundance radiocarbon measurements, which were consistent with a
natural formation as opposed to an industrial synthesis based on petrochemical feedstock (170).
The majority of the naturally produced OH-/MeO-PBDEs reported in the
literature so far have the hydroxy/methoxy group at the ortho position (62).
While the metabolism of PBDEs at the ortho position could not be excluded,
the presence of hydroxylated/methoxylated BDEs at the meta or para positions is more indicative of metabolism of PBDEs (62,171). For example, the
4’-OH-BDE49 identified in Baltic salmon (62), in various fish species from
the Detroit river (172), and in glaucus gull from the Norwegian Arctic (173)
was suggested to be a metabolite of BDE47 (by 1,2-shift of bromine and
hydrogen) and/or BDE49 (62).
The natural source of the MeO-BDEs is still not fully understood. A tentative spatial trend of herring collected in 1987 showed increased concentrations from the Kattegatt/North Sea to the Bothnian Bay (paper I) and no
agreement with the corresponding BDE47 concentrations. This was also
observed in 2001 and 2002 (174). A range of OH-/MeO-PBDE pairs have
been identified in salmon blood, blue mussels and in the red macroalgae sp.
Ceramium from the Baltic region (62,63). Their presence in red algae may
represent one of the sources in the Baltic and can possibly explain the lower
levels found in herring from the Kattegatt.
Table 3. Characteristics of the sampling location and concentration of 6-MeOBDE47 and 2’-MeO-BDE68 in fish samples collected in Swedish lakes.
Species, Lake,
sampling year
n
Pike, Bolmen,
141+
17*
1967-2000
Pike, Storvindeln,
1*
LRM, 1993
Pike, Storvindeln,
7
1978-2000
Roach, Krankesjön,
64
1980-1996**
Pike, Roxen,
9
1972
Perch, Stensjön
Hjärtsjön, Bysjön,
24
2001***
Sampling
season
Region
Degree of
productivity
MeO-BDE68
(ng/g lw)
MeO-BDE47
(ng/g lw)
Spring
Woodland
Mesotrophic
25-240
50-500
Winter
Remote
Oligotrophic
21
7
Spring
Remote
Oligotrophic
4-20
2-4
Autumn
Agricult.
Eutrophied
n.d.
n.d.
Spring
Agricult.
Eutrophied
0.3
0.4
Autumn
Woodland
Oligotrophic/
mesotrophic
~0.3-0.6
~0.07-0.15
* pooled samples, ** Results from (155), *** Unpublished results (L. Asplund), n.d. not
detected.
32
The natural origin of the MeO-PBDEs in freshwater systems is intriguing. In
order to investigate if the source of the MeO-PBDEs was produced by a
primary producer and thus related to eutrophication, fish samples from a
number of Swedish lakes distributed all over Sweden, from oligotrophic to
eutrophic and from populated areas to remote regions, were analysed (Table
3) (paper I). No correlation between MeO-PBDE levels and any of the variables was found. Although the data set in this study is limited, it suggests
that the source is neither favoured by eutrophication nor comes from emissions connected to human activity.
Bioaccumulation
According to the Stockholm Convention, bioaccumulation is one of the key
characteristics in the classification of POPs (1). While bioconcentration is
the result of the diffusive equilibration of the chemical between the organism
and the media in which the organism lives, i.e. water for aqueous and air for
terrestrial organisms, bioaccumulation represents the combined uptake from
the bioconcentration and from the diet. Some bioaccumulating compounds
biomagnify in the food chain. For hydrophobic compounds, biomagnification can be defined as bioaccumulation leading to higher lipid weight normalized concentrations in an organism compared with its food. Here, the
first focus is on one of the key processes determining a chemical’s ability to
biomagnify, namely the dietary absorption efficiency. This process depends
on the physical-chemical properties of the substance, the type of food, and
the strength of the chemical’s sorption to the food. The second focus concerns other factors that affect the extent of bioaccumulation, namely biotransformation after absorption as well as excretion, these will be addressed
later in this chapter.
Dietary absorption
Despite early findings of BDE209 in human adipose tissue in 1991 (141), it
was, based on its high molecular weight and extreme hydrophobicity,
claimed not to be bioavailable and therefore unable to bioaccumulate in
wildlife (175). The initial screening studies performed pointed in the same
direction, in contrast to the lower brominated BDEs, no BDE209 was detected in fish despite high concentrations in sediment (5,176). The uptake of
BDE209 was therefore investigated in rainbow trout fed dried cod chips
fortified with the technical Deca-mix (paper II). Both liver and muscle levels of BDE209 increased with time of exposure, thus demonstrating that it
was absorbed from the gut in fish. The absorption efficiency was difficult to
calculate due to the extensive debromination encountered (see later discussion), hence only a rough estimate was presented. This comprised the molar
33
sum of the debromination products in the muscle tissue based on estimated
response factors interpolated from those of the hexaBDEs and BDE209,
which were the only congeners available as pure standards at that time.
Comparing the mean molar sum of the lower BDEs and BDE209 with the
mean dietary dose of BDE209, the uptake was extremely low, less than 0.13
% after 120 days. This was in contrast to the efficient uptake observed in
pike subject to dietary exposure to tetra- to hexaBDE (177). The efficiency
in that study was partly attributed to a lipid-mediated facilitation achieved by
using living rainbow trout as the matrix for the congener cocktail. Recently,
an uptake study of BDE209 in rainbow trout similar to that in paper II was
performed (178). The dietary absorption was now determined to 3.5 %. The
dose vehicle used was similar to the one in paper II, except that cod oil was
used instead of corn oil as the solvent for BDE209 and food pellets were
used instead of cod homogenate. The daily dose, however, was 3 orders of
magnitude lower (~10 µg/kg body weight). In addition to the differences in
the lipid content of the food and the lower daily dose, the lower uptake in
paper II may have been due to the fact that the technical product was not
fully dissolved in the corn oil before blending it with the cod homogenate/gelatine mixture. Hence the BDE209 may have been present in the fish
gut as aggregates, which could hinder absorption. The use of a suitable vehicle was suggested to at least partly explain the high dietary absorption of
BDE209 observed in rats (> 26 %) (166,179) compared to previous studies
in the same species (< 0.5 %) (180,181).
Decreasing absorption efficiency with increasing hydrophobicity, (log
Kow > 6.5), molecular size or molecular weight has previously been observed
for halogenated organic compounds in fish (182,183) and in cows (184,185).
These characteristics all coincide with increasing degree of halogenation of
the substance. There are currently two major theories on the mechanism
responsible for the reduced dietary absorption of large superhydrophobic
substances. One is based on the size of the molecule physically restricting its
passage through the membrane. A molecular effective cross section (ECS) of
9.5 Å has been proposed as the limit for uptake via the gills and similar limits were also suggested for the gut (186,187). According to this theory, not
even the pentaBDEs, which have an ECS of 9.6 Å (177), should be taken up.
The other theory is based on the combined effect of the compound’s lipid
and water solubilities, usually expressed as the log Kow. The first restriction
encountered during dietary absorption is here defined as the ability to dissolve in the lipid/bile salt micelles formed in the gut that act as the transport
medium to the intestinal epithelium. The second restriction is connected to
the passage through a stagnant boundary layer of water through which the
chemicals must diffuse following dissociation of the micelles (188). For
superhydrophobic substances such as BDE209 the rate limiting restriction is
believed to be the passage through the water layer.
34
The dietary absorption efficiencies of PBDEs have been investigated in
pike, zebra fish, lake trout, common carp and in farmed Atlantic salmon
(177,189-193). Although there were differences between the species, all
except one study (190) showed equal or higher absorption of the lower
BDEs in comparison to PCBs (177,191-193). Surprisingly, none of the studies found a clear correlation between the bioaccumulation from food of the
BDE congeners and their log Kow (BDE209 was included in only one study
(190). There were, however, large differences in the apparent absorption
efficiencies between BDE congeners. Because of the inherent difficulties in
collecting feces from fish, the uptake was based on the tissue concentrations
of the studied substances and thus did not include any biotransformation.
Since a range of BDE congeners were added to the food in all studies, the
rapid biotransformation of some of the BDEs was suggested to account for a
low apparent absorption efficiency of, for example BDE99 and BDE153 in
carp (191). Similarly, the unexpectedly high apparent absorption efficiencies
may have been confounded by the bioformation of lower brominated BDEs
(189-191,193). The extremely high absorption efficiency, close to 100 %, for
BDE47 and its very slow approach to steady state were attributed to bioformation via debromination of higher brominated congeners (191-193).
In cows (paper III), the dietary absorption could unfortunately not be
quantified due to the variability of the PBDE levels in the silage. However,
decreasing absorption with increasing log Kow, bromination degree or molecular size was indirectly indicated by a comparison of the ratios of the
PBDE concentrations between adipose tissue and silage, focusing on the
lowest ratio in each homologue group (Figure 7). The stepwise decreasing
ratio from hepta to deca homologues may reflect the influence of bromination degree, molecular size or log Kow on the dietary absorption.
Figure 7. Ratio of adipose tissue and silage concentrations for hepta- to decaBDE in
the slaughtered cow (paper III). The lowest ratio within each homologue group is
highlighted. * silage levels < LOQ (measured values, not LOQ, are plotted). The
congeners are presented in the order of elution from the GC-column.
35
High absorption efficiencies have been reported for BDE209 in rats (>26 %)
(166,179) and in grey seals (89 %) (194). This is in accordance to previous
findings for other POPs (such as PCBs and PCDD/Fs) showing higher assimilation efficiencies for birds and mammals compared to fish and other
poikilothermic organisms (188). Furthermore, in contrast to birds and humans the absorption of substances in both fish and cows has been shown to
be dramatically reduced for substances with log Kow > 6.5 (182,184).
Biotransformation
The biotransformation of PBDEs is not fully understood and the number of
investigations are few (for a review see Hakk & Letcher, 46). Similar to the
metabolism of other aromatic xenobiotics, a suggested pathway for PBDEs
is oxidative hydroxylation and/or oxidative debromination via an epoxide,
mediated by the cytochrome P450 enzyme system. In general, laboratory in
vivo-studies identifying polar metabolites of PBDEs have, with the exception
of one fish study (164), been limited to rodents. These studies have all identified the presence of hydroxylated metabolites of the BDE congeners studied (164,165,171,195). In pike that were exposed to food spiked with 14CBDE47, six hydroxy-PBDEs were detected and no lower brominated PBDEs
were produced (164). Similar results have been reported in rodents exposed
to tetra- to heptaBDEs (e.g. 71,196).
The metabolism of BDE209, however, seems to be more complex. Due to
the lack of unsubstituted carbon atoms an oxidation via an arene oxide is
most likely preceded by a debromination (46). However, the outcome from
the rat studies varied from no hydroxylated or debrominated metabolites
detected in the blood (171), to several debrominated-hydroxylated and hydroxylated-methoxylated metabolites identified (166,179), and recently to
lower brominated BDEs indicating debromination of BDE209 (197).
Debromination as metabolic pathway has previously been shown in fish
and rats exposed to other brominated substances. Already in 1977, lower
brominated biphenyls were detected in Atlantic salmon exposed to a technical octabromobiphenyl mixture administered via the water and via the food
(198). Later, reductive debromination of hexabromobenzene was demonstrated in rats (199). It was concluded that intestinal microorganisms were
not involved. Similarly, in rainbow trout that were exposed via the diet to the
commercial Deca-mix product, a range of hexa- to nonabrominated BDEs
were detected (paper II). The levels of all of these congeners increased with
the length of exposure. However, assessment of metabolic debromination
was confounded by the presence of lower brominated BDEs in the amended
food. The technical Deca-mix formulation used in the experiment, DOW
FR-300BA, contained more nonaBDEs than later formulations, possibly
reflecting changes in manufacturing practices (43). Furthermore, despite
precautions, additional degradation of BDE209 to for instance BDE206 may
36
have occurred during the initial separation of the PBDEs from planar impurities and in the preparation/storage of the amended cod chips (see Figure 2 in
paper II). Nevertheless, whereas the congener profile in the feed resembled
that observed after photolytic degradation of BDE209 (25,26), the profile in
fish was different (Figure 8). In fish the first eluting congeners in each
homologue group from the hepta- to nonaBDEs were prominent, while the
late eluting congeners were more abundant after photodegradation. Despite
the apparent accumulation, it was not possible to exclude uptake of minor
ingredients in the diet. By estimating the total amount of the different congeners in the fish compared to the potential dose via the feed, not even nondetected impurities could be ruled out. Another way to exclude selective
uptake was to assess the formation of lower brominated congeners during the
depuration period. Although both BDE154 and the first eluting heptaBDE
showed an increase during this period in contrast to their potential precursors, it was not statistically significant. However, estimated bioaccumulation
factors (muscle:feed) after 120 days of exposure differed up to a factor of
400 between isomers within the homologue groups. Since the mechanism of
dietary absorption for hydrophobic compounds in general is believed to be a
non-selective diffusive process, bioformation was a more likely explanation
for the disparate bioaccumulation behavior of the isomers.
Since then, reductive debromination of PBDEs has been confirmed in fish
(178,190,192,193,200,201), in starlings (202), and by anaerobic microorganisms by single bacteria cultures as well as by inoculum from a sewage sludge
digester (203,204). Laboratory experiments with common carp, Atlantic
salmon and rainbow trout (178,190,192,193,200) have shown debromination
to occur with both lower and higher brominated BDEs as precursors. Furthermore, mosquitofish and pumpkinseed captured in the receiving waters of
a wastewater treatment plant contained a number of unidentified hexa- to
octaBDEs not present in the sediment or effluent waters, which indicated
that debromination had occurred also outside the laboratory (201).
Carp was shown to debrominate (possibly BDE153 to) BDE99 to BDE47
and BDE183 to BDE154 (192) as well as BDE209 to octa- down to pentaBDEs (200). In contrast to rainbow trout (paper II), the parent BDE congener
was never detected in the tissues. Carp are stomachless fish with the liver
tissue situated along the intestine. This may result in a more efficient absorption and a greater metabolic capacity. A comparative in vitro study of the
degradation of BDE209 by liver microsomes from rainbow trout and carp
revealed a more efficient debromination (65 % of the total mass of BDE209)
by carp than rainbow trout (22 %) over 24 hours (178). The differences in
the rate of metabolism were also confirmed in vivo by exposing rainbow
trout via the diet. An almost identical congener profile was obtained in this
study compared to the previous study (paper II).
The enzymes involved in the debromination have not yet been identified.
It has been suggested that the deiodinase enzymes responsible for the re37
moval of an iodine atom from thyroxine (T4) to triiodothyronine (T3) are
involved (178). These enzymes, which are essential for the regulation of
thyroid hormones, are selective for meta-substituted iodine, which is one
position reported to be debrominated in fish (192). Another possibility is
reductive dehalogenation via the cytochrome P450 enzyme system. Then
again, no increase in plausible debromination products was observed in Atlantic tomcod intraperitoneally injected with BDE209 and the cytochrome
P450 1A (CYP1A) inducer CB126 (205). Whether other cytochrome P450
enzymes are involved is not known, however, both cytochrome P450 and
deiodinase enzyme systems are present in the microsomal fraction used in
the in vitro studies described above.
Figure 8. Mass chromatogram (bromide ions in ECNI) for A) rainbow trout muscle
after 120 days of dietary exposure to Deca-mix, B) Deca-mix in toluene photodegraded for 120 min (data from U. Sellström) and C) Penta-, Octa-, and Deca-mix
technical PBDE products. Coelutions (e.g. BDE197/204 and BDE203/198) cannot
be excluded.
38
Figure 9. Combined mass chromatograms ([M]+ for hexaBDEs and [M-2Br]+ for
hepta- to decaBDE in HRMS, EI) for A) cow’s milk, B) adipose tissue, C) cow’s
feces and D) grass silage. Coelutions cannot be excluded.
Typical signs of debromination were also encountered in the cow (paper
III). The congener profile in the cow’s adipose tissues, organs and in the
milk differed from that in the major feed, the silage, and the feces (Figure
9). Three major congeners, BDE197, BDE196 and BDE207, seemed to accumulate in the cow. In accordance with the uptake study in fish (paper II),
39
the differences between isomers within the homologue groups pointed towards debromination after absorption as the most likely explanation.
The comparison of congener profiles between feces and feed on the one
hand and organs/tissues and milk on the other is difficult since other processes apart from debromination influence the outcome. The profile in the
tissues also reflects the differences in the absorption efficiency and clearance
rate, and, for milk, the transfer efficiency from the mammary gland to the
milk. These differences are substantial between homologue groups. By normalizing the concentrations to the congener with the lowest level within each
homologue group, the similarities and dissimilarities between the silage, the
feces, the adipose tissue and the milk are revealed (Figure 10). It seems
clear that compared to isomers within the homologue groups there is a preferential accumulation after absorption of BDE207 among the nonaBDEs, of
BDE197 and BDE196 among the octaBDEs, and possibly by BDE182
among the heptaBDEs. BDE182 was below the quantification limit in the
silage and feces but clearly detectable in the adipose tissue and thus only
indicative.
Figure 10. Concentrations normalized to the congener with the lowest concentration
within each homologue group. The heptaBDE was normalized to BDE183, the octaBDEs to BDE203, and the nonaBDEs to BDE208. * < LOQ (measured values, not
LOQ, are plotted). The congeners are presented in the order of elution from the GCcolumn.
Another species from the terrestrial food chain that has been shown to be
capable of reductive debromination is the European starling. 76 days after
the insertion of BDE209 amended implants under the skin, a range of lower
40
brominated BDEs were detected in the muscle and in the liver (202). The
BDE congeners that accumulated most were BDE207, BDE208, BDE197,
and BDE206.
Humans exposed to the Deca-mix also had elevated concentrations of
octa- and nonaBDEs in the blood (97,206). From the chromatograms the
most abundant congeners seem to be BDE207 and an octaBDE (by comparing elution order possibly BDE197). Recently, the debromination of
BDE209 in rats identified BDE207, BDE197 and BDE201 as major products
after 21 days of dietary exposure (197). Thus, at least two of the major accumulating congeners, BDE207 and BDE197, are common to cows, starlings rats and possibly also to humans.
The major BDE congeners accumulated in the cow were not the same as
those with the highest rate of formation in the fish study. The debromination
products from BDE209 in rainbow trout were recently identified in a feeding
study that was similar to that in paper II (178). The major heptaBDE
formed was identified as BDE188, and the two major octaBDEs as BDE202
and BDE201. The debromination pathways are difficult to assess since the
accumulation of a congener may represent both a bioformation from higher
brominated congeners and/or a resistance to further degradation. Depending
on the rate of formation versus the rate of degradation, some congeners accumulate to a larger extent than others. Nevertheless, it is clear that fish are
capable of removing meta-substituted bromine atoms, for example in transforming BDE99 to BDE47 as reported in carp (192). In addition, fish can
also remove bromine atoms in the para-position, as the accumulation of
BDE202 implicates. BDE congeners with bromine substituents in the orthoposition appear to be more recalcitrant in fish. The debromination pathway
in cows (paper III) seems to be more selective. The major products are restricted to the removal of bromines in the meta-position such as BDE207
debrominating to BDE197 and BDE196 debrominating to BDE182, and
possibly to bromines being removed from the ortho-position. Thus, BDE196
may originate from an ortho-debromination of BDE207 or a metadebromination of BDE206, both of which were present in the silage. In contrast to the debromination pathway demonstrated in fish, no apparent removal of bromines in the para-position was observed in cows. Tentative
schemes of the debromination pathways of BDE209 encountered in fish
(paper II, (178,190,192)) and in cows (paper III) are presented in Figure
11. It is tempting to attribute differences in bioaccumulation profiles to differences in the capacity of biotransformation between species. Could a less
efficient metabolism in terrestrial biota explain the larger proportion of
higher brominated BDEs in some of these organisms?
41
Figure 11. Tentative pathways for the metabolic debromination of BDE209. The
congeners shown represent a summary of identified lower brominated BDEs detected in fish (paper II, (178,190,192), and in cows from paper III. The suggested
pathways for the degradation in papers II and III are marked as shadows in the
background (no quantitative measures). Solid arrows represent debromination at the
meta- and para-positions. The dotted arrows represent a tentative debromination at
the ortho-position.
42
Aquatic versus terrestrial environment
Similar to most classical organochlorines, the research on PBDEs was for a
long time focused on the aquatic environment. Generally, in fish, fish-eating
birds and marine mammals, the lower brominated BDEs dominate the BDE
pattern typically with BDE47 as the most abundant congener
(38,40,77,148,207,208). The few early investigations including samples
from terrestrial biota reported low levels of (lower brominated) PBDEs, for
instance in rabbits, moose, reindeer and starlings (2). However, the discovery of high levels of higher brominated BDEs in the eggs of peregrine falcons (7) raised the question of whether organisms in the terrestrial food
chain may be more exposed to higher brominated BDE congeners than
aquatic organisms. This was further supported by the bioaccumulation of
octa- to decaBDE reported in earthworms collected from sewage sludgeamended soil (27). Earthworms represent the base of many terrestrial food
chains and thus may constitute a pathway for the bioaccumulation of higher
brominated BDEs at higher trophic levels. BDE209 was also the dominant
BDE congener in the soil samples from the reference sites, indicating that
atmospheric deposition may be an important source of the PBDE contamination in terrestrial ecosystems. A similar profile was also found in
air/deposition samples from the Baltic Proper, an area remote from any direct emissions (209).
A congener profile indicating a preferential accumulation of higher brominated BDEs was also observed in other terrestrial birds of prey such as
merlins and golden eagles (148). Furthermore, a survey of birds from the UK
showed a clear, positive correlation between the presence of BDE209 and
birds feeding from the terrestrial food web (210). Recently, the bioaccumulation of higher brominated BDEs was reported in eight terrestrial bird species
from the Beijing area in Northern China (211). BDE209 concentrations of up
to 12 ppm on a lipid weight basis were found in the liver of the common
kestrel, which is the highest BDE209 level reported in wildlife. The kestrels
were year-round residents of the urban areas of Beijing and thereby exposed
to PBDE sources such as domestic waste. Foxes and grizzly bears are other
terrestrial species reported to have a congener profile dominated by BDE209
(212,213). The fox is an omnivore that, when possible, supplements its diet
with domestic waste. The difference between terrestrial and aquatic food
chains was further reinforced by the PBDE profile in grizzly bears from British Columbia, although all specimens contained BDE209, only in those feeding entirely on terrestrial food was BDE209 the dominant congener, whereas
BDE47 was the major congener in those that ate salmon part of the year
(213).
In paper III, BDE209 was the dominant congener in the cow tissues examined, which is consistent with the observations for the other terrestrial
feeders listed above. However, the concentrations were low, on average 3.7
43
ng/g lipid weight in adipose tissues, which is at the lower end of the range of
levels detected in foxes and grizzly bears. This may reflect the extent and
composition of the dietary exposure, as in the example of the grizzly bears,
but may also be attributable to physiological differences between species.
There are other variables besides the contamination of the diet that vary
between species and that could also play a role in the differences in bioaccumulation of BDE209 between aquatic and terrestrial organisms as well as
among terrestrial organisms. One is the dietary absorption efficiency. Investigations with other chemicals have shown that the dietary absorption of the
higher chlorinated PCDD/Fs (log Kow comparable to that of BDE209) is
much lower in cows than in terrestrial predators (e.g. humans or birds) (188).
This could explain the low levels of BDE209 in cows compared to other
terrestrial organisms. The other explanation is related to differences in the
biotransformation efficiency. Thus, despite a low absorption of BDE209, the
dominance of BDE209 in cows may be explained by a low rate of metabolism and excretion. Biotransformation may also be responsible for the different BDE congener pattern observed in German peregrine falcons compared
to sparrowhawks, despite their similar feeding habits (38). In humans, the
apparent half life of BDE209 in serum is short (15 days) (214) even though
the dietary absorption is high for highly chlorinated POPs, which indicates a
rapid metabolism of BDE209. The comparably low levels of BDE209 in
aquatic food webs may be attributable to a lower uptake (paper II, 178,188)
and/or a more efficient metabolism of this compound by aquatic organisms.
Which of these two factors, dietary absorption or metabolism, is more important for the bioaccumulation of BDE209 is, however, far from clear, and
may vary from species to species.
Risk assessment / Implications for exposure characterisation
An important pathway for human exposure to PCBs and PCDD/Fs is via
atmospheric deposition to soil and vegetation, transfer to grazing cows, and
sequestration in dairy products and meat (215). Since BDE209 is present not
just in occupationally exposed workers but also in the general population
(206,216), there must be important sources of exposure outside of the workplace. The short half-life of BDE209 in human serum implies that exposure
must be semi-continuous in order to explain the blood concentrations observed (214). One proposed source of human exposure to higher brominated
BDEs is via inhalation/ingestion of household dust. Dust has been reported
to contain considerable amounts of BDE209 (134,217,218). Another pathway may be via food.
44
Meat and dairy products from cattle account for a large part of the human
exposure to classical chlorinated POPs. Insight into these potential vectors of
human exposure to BDE209 was gained in paper III. A clear discrimination
against the transfer of higher brominated BDEs to the milk was observed.
Whereas the adipose tissues from different lipid compartments in the cow
had similar concentrations on a lipid weight base, indicating that the tissues
were close to equilibrium, the concentrations in milk fat were 1-2 orders of
magnitude lower. This was in contrast to the PCBs and the lower brominated
BDEs in the same cows, for which the concentrations in milk fat and adipose
tissues were similar (219,220). A comparable behaviour has been reported
for the lactational transfer of higher chlorinated PCBs and for tetra- to hexaBDEs in marine mammals with a high throughput of milk fat (88,221,222).
The low transfer encountered for higher brominated BDEs in cows was
therefore suggested to be caused by a slow rate of transfer of superhydrophobic compounds into the milk in combination with the high daily yield of
milk fat (about 1 kg/day). Recently, the discrimination in the transfer of
higher brominated BDEs was confirmed in human milk versus serum from
Japanese mothers (223). BDE209 was the dominant congener in the serum
but not in milk, which had 10 times lower concentrations. Compared to humans the transfer of BDE209 in cows was even smaller, a factor of 100
compared to adipose tissue, possibly related to the larger elimination rate of
fat per day. The consequence of this discrimination in dairy cows is that,
instead of being excreted via the milk like most lipophilic contaminants in
lactating cows, the highly brominated BDEs are accumulated in the fat and
meat. From the perspective of human exposure, this can be expected to influence the dominant pathways of dietary intake.
Several food basket analyses in Europe have identified fish as the major
source of lower brominated BDEs (224-226). However, North American
studies have recognized meat as the primary contributor, probably due to the
fact that less fish is eaten there (227,228). Only a few studies have included
higher brominated BDEs (227,229,230), and in these the BDE209 levels
varied widely between food types and items. For example, BDE209 was the
dominant congener in cheese but not in butter or milk in one investigation
(230), while it was dominant in butter but not in cheese and milk in another
study (229). More work is needed to clarify which foods contribute most to
dietary exposure to BDE209. The results of paper III suggest that particular
attention should be payed to beef.
Debromination of higher brominated BDEs can also have consequences
for PBDE risk assessment, since the toxicity of PBDEs seems to be inversely
related to the degree of bromination (44). Due to the analytical difficulties
associated with the analysis of BDE209, there are fewer investigations including this congener compared to lower brominated BDEs. However, the
number of papers reporting the presence of BDE209 in biota is increasing. In
order to assess debromination in environmental samples, there is a need to
45
identify and quantify the debromination products. The limited number of
commercially available reference standards for hepta- to octaBDEs is a constraining factor. Furthermore, the difficulties connected to the quantification
of octa- and nonaBDEs due to the degradation of BDE209 during the procedures of sample preparation and instrumental analysis pose challenges that
were addressed in paper III.
A different area that indirectly is influenced by debromination is the determination of bioconcentration factors (BCF), bioaccumulation factors
(BAF), or half-lives of BDE congeners from environmental samples or in
laboratory experiments where BDE congener mixtures are used. As previously discussed for the absorption efficiencies in fish, the successive debromination from higher to lower brominated congeners will give rise to overestimated absorption efficiencies, BCFs, BAFs or half-lives for the debromination products and underestimated factors for the debromination precursors.
Such processes may, for example, explain the unexpectedly high biota-soil
BAFs that were reported for BDE197 and BDE207 in earthworms in sewage
sludge amended soil (27). Likewise, the differences in apparent half-life
between isomers of higher brominated BDEs in human blood may be related
to bioformation/debromination (214).
PBDE bioaccumulation has been shown to be highly species and congener specific. Some of these differences may be at least partly caused by differences in metabolic capacity (146,231,232). If the major metabolic pathway of the species is via debromination, it will influence the BDE congener
accumulation in its predator. Biomagnification factors (BMFs) are therefore
species dependent and highly specific for the food web they describe. For
example, if the debromination capacity increases at the top of the food chain,
the BMFs of the debromination products will be elevated and dependent on
the levels of precursors present in the prey. As with the BAFs, the consequences are an overestimation of the biomagnification potential for the debromination products and an underestimation for the debromination precursors. Such mechanisms may for instance explain the unexpectedly high biomagnification potential reported for a heptaBDE in marine food webs from
the Baltic Sea and Atlantic Ocean (207).
As of August 2004, the European Union banned the marketing and use of
the Penta-mix and Octa-mix formulations. The United States has currently
no national legislation concerning PBDEs, but the sole manufacturer of the
Penta-mix and Octa-mix voluntarily stopped production of these PBDE
products at the end of 2004 (233). Canada has recently (December 2006)
proposed strategies to prohibit the use of the Penta-mix and the Octa-mix
and to regulate the use of the Deca-mix. In Europe, the Deca-mix is exempt
from the prohibition of PBDEs. However, Sweden has from the first of
January 2007 introduced a national ban on the use of Deca-mix in other applications than electronic products and vehicles (19). Thus, although the
major applications are in electronic goods, it is not allowed in for example
46
textiles. Still, from a global perspective the Deca-mix product is the most
frequently used and the least regulated PBDE product. Furthermore, the
global presence of BDE209 in the environment may indirectly, via biotic or
abiotic debromination, be a long term source of future exposure of biota to
lower brominated BDEs.
Decabromodiphenyl ethane – a next generation BFR
DeBDethane, a BFR used to replace the Deca-mix formulation, was identified for the first time in the environment (paper IV). DeBDethane was present in sediment from an area heavily polluted by the Deca-mix, in sewage
sludge from 25 (of 50) Swedish STPs and in an air sample from an electronic dismantling facility. In all samples more BDE209 than DeBDethane
was present, but the ratio differed (see Figure 12). Its presence in sewage
sludge did not show any correlation to STP size or latitude but seemed to be
more frequent in highly populated areas. The occurrence was similar to that
of BDE209, indicating a diffuse input from households and urban runoff
(8,76). DeBDethane has since then been reported in sewage sludge from
Spain (234) and Canada (235). The concentrations were in the lower range
(up to 15 and 32 ng/g dry weight) of those reported in paper IV.
Figure 12. Concentrations (ng/g dry weight) of DeBDethane and BDE209 in sewage sludge from 50 Swedish STPs. The DeBDethane concentrations are estimates
and the BDE209 concentrations are from Nylund et al. (8,76). Zero-levels represent
“non-detects”.
47
The presence of DeBDethane was also confirmed in both air and dust samples (tentatively identified) from a Swedish electronic recycling facility
(236,237). Further evidence of its applications in domestic electronic products is that detectable DeBDethane levels are reached in normal household
dust. For example, DeBDethane was recently reported in 4 of the 5 household dust samples, and in fact exceeded the BDE209 levels in one sample
(238). The concentrations did not correlate to the BDE209 concentrations,
and DeBDethane was not detected in human plasma from any of the inhabitants. Additionally, DeBDethane was detected along with BDEs in tree bark
collected in Arkansas, U.S., and their occurrence was traced to emissions
from two major BFR manufacturing plants in southern Arkansas (239).
Information on the bioavailability/absorption of DeBDethane is so far
scarce. However, recently DeBDethane was included in a study investigating
the bioaccumulation of a number of BFRs in a food web from Lake Winnipeg, Canada (240). Although DeBDethane was not detected in zooplankton,
mussels or whitefish, it was present in some of the other fish species, such as
walleye and burbot. A significant relationship between trophic level and
lipid-based concentrations led the authors to suggest that DeBDethane biomagnified in the food web investigated.
The behaviour of DeBDethane is in many ways similar to that of
BDE209, both being large, superhydrophobic and thermolabile compounds.
Both compounds also degrade to some extent to lower brominated congeners
during sample preparation/analysis. Two nonaBDethane congeners were also
present in the technical product, Saytex 8010® (paper IV). BDE209 is
known to be photolytically labile undergoing degradation to lower brominated congeners (e.g. 25). Whether DeBDethane is susceptible to degradation via UV light is not known. However, preliminary results from exposing
a DeBDethane solution to a daylight fluorescent lamp (Metal Halide Arc) in
the laboratory showed that DeBDethane is degraded, producing two nonaBDethane congeners as well as a number of peaks tentatively identified as
octa-brominated products (Figure 13).
A study of the thermal decomposition (pyrolysis) of flame retarded highimpact polystyrene (HIPS) showed that DeBDethane was degraded mainly
to brominated toluenes, while BDE209 produced lower brominated BDEs
and brominated furans (241). The difference was suggested to be caused by
the weaker C-C-bond in DeBDethane compared to the ether bond in the
BDE209 structure. Whether a degradation to pentabromotoluene (PBT) is
feasible in the environment is not known. Interestingly, pentabromotoluene
has recently been detected in air, river sediment and in blubber from beluga
whales (242-244). Although pentabromotoluene is/has been used as a BFR
itself (9) and may also be released from the production of BFR oligomers
(245), further investigation of its possible formation from DeBDethane is
warranted. Future work could investigate the covariation of the concentra-
48
tions as an indication of a common source of DeBDethane and pentabromotoluene.
Figure 13. Mass chromatogram (bromide ions in ECNI) of technical DeBDethane in
n-hexane, A) photodegraded and B) original solution (unpublished data). The raised
baseline before the DeBDethane peak is due to thermal degradation in the column.
The present study illustrates the importance of carefully assessing the environmental behaviour of new chemicals, particularly when they are replacements of known problematic substances. It is likely that the use of DeBDethane will increase in the future, particularly after the recent regulatory initiatives to reduce the use of the Deca-mix in, for instance, Sweden and Canada. On the basis of predicted future widespread use in the UK, an
environmental risk evaluation of DeBDethane is currently being performed
by the British EPA (draft for peer review 2006).
49
Conclusions
The work presented in this thesis has highlighted some aspects of the behaviour of BFRs in the environment. The results contribute to the exposure
characterization for the environmental risk assessment of the PBDEs in particular, but also a representative of the next generation BFRs, decabromodiphenyl ethane. From the results the following conclusions were drawn:
• The concentrations of lower brominated BDEs in pike from a Swedish
lake reflect the past usage of PBDEs in Sweden/Europe. The levels increased up to the mid 1980s and then slowly decreased from the late
1980s up to now.
• The structurally related methoxy-PBDEs, which were present in
amounts similar to lower brominated BDEs in fresh water pike, do not
originate from the anthropogenic PBDEs.
• Despite its molecular size and hydrophobicity, the fully brominated
PBDE, BDE209, can to a small extent be absorbed in fish and cows,
when ingested via the diet.
• In vivo reductive debromination of BDE209 to lower brominated BDEs
occurs in fish as well as in cows exposed via the diet.
• The transfer of PBDEs to the milk in cows decrease dramatically with
increasing molecular size, bromination degree or log Kow of the BDE
congener.
• A representative for the next generation of BFRs, decabromodiphenyl
ethane that is marketed as a replacement for the DecaBDE product is
already present in the environment.
For PBDEs, the stability and toxicity decrease with increasing degree of
bromination. This behaviour is in many ways different from that of the classical chlorinated pollutants. As shown in this thesis, the higher brominated
BDEs are subject to degradation, producing lower brominated BDEs both in
organisms as well as in the laboratory during analysis. Others have shown
their ability to debrominate in the presence of UV-light. While the usage of
50
lower brominated BDEs is subject to regulations, the highly brominated
BDEs continue to be used in large volumes. A key question is therefore
whether the continued use of the higher brominated BDEs will result in increased contamination of biota with the more environmentally detrimental
lower brominated BDEs. The lack of persistence of the higher brominated
BDEs, which is generally a useful property of an environmental contaminant, may in this case be detrimental, as BDE209 becomes a source of more
persistent lower brominated BDEs in the environment. The differences in the
isomer patterns of BDE homologues between the technical products and
what is observed in biota suggests that this is happening.
51
Acknowledgements
My supervisors Lillemor Asplund, Cynthia de Wit and Michael
McLachlan, are gratefully acknowledged for their support and guidance
at different stages of this process.
Others of importance for the outcome are Bo Jansson, a pioneer in the
analysis of brominated substances, Ulla Sellström, my closest colleague
with hands-on experience of PBDE-analysis, my coauthors for their
exclusive contributions, especially Jonas Björklund, Gareth Thomas,
Anders Bignert, Mats Olsson, Lennart Balk and Ulrika Fridén.
Yngve Zebühr and Michael Strandell are acknowledged for their skilful
technical assistance with the HRMS analysis.
Sören Jensen for advice in the analytical method development, Eva
Jakobsson and Göran Marsh for gifts in the form of synthesized
standards.
Finally, I am grateful for the help from my colleagues at ITM, who made
small and large contributions along the way.
Thank you all!
52
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