E n v i r o n m e n t a l l y r e l e v a n t c h e m i c a l d i s r u p t o r s o f
o x i d a t i v e p h o s p h o r y l a t i o n i n B a l t i c S e a b i o t a
- E x p o s u r e a n d t o x i c p o t e n t i a l s
Anna-Karin Dahlberg
Environmentally relevant chemical
disruptors of oxidative phosphorylation
in Baltic Sea biota
Exposure and toxic potentials
Anna-Karin Dahlberg
©Anna-Karin Dahlberg, Stockholm University 2015
ISBN 978-91-7649-127-0
Printed in Sweden by Publit, Stockholm 2015
Distributor: Department of Environmental Science and Analytical Chemistry, Stockholm University
Front cover: Bladderwrack (F. vesiculosus) from the Baltic Sea, photo by Jukka www.flicr.com
Blue mussels (M. edulis), photo by Travis www.flicr.com
Long-tailed duck (C. hyemalis), photo by Dominic Sherony www.flicr.com
Herring (C. harengus), photo by Thomas Quine www.flicr.com
To my beloved parents
viii
List of papers
This thesis is based on the following articles and manuscripts, referred to in
the text by their roman numerals (I-IV). The published articles are repro-
duced here with kind permission of the publisher. Co-shared first authorship
is indicated with an asterisk.
I Disruption of oxidative phosphorylation (OXPHOS) by
hydroxylated polybrominated diphenyl ethers (OH-PBDEs)
present in the marine environment.
Jessica Legradi*, Anna-Karin Dahlberg*, Peter Cenijn, Göran Marsh,
Lillemor Asplund, Åke Bergman and Juliette Legler.
Environmental Science & Technology. 2014, 48, 14703-147011.
II Hydroxylated and methoxylated polybrominated diphenyl ethers
in Long-tailed duck (Clangula hyemalis) and their main food,
Baltic blue mussels (Mytilus trossulus x Mytilus edulis)
Anna-Karin Dahlberg, Vivian Lindberg Chen, Kjell Larsson, Åke Bergman
and Lillemor Asplund.
Manuscript
III Anthropogenic and naturally produced brominated substances in
Baltic herring (Clupea harengus) from two sites in the Baltic Sea.
Anna-Karin Dahlberg, Anders Bignert, Jessica Legradi, Juliette Legler
and Lillemor Asplund.
Manuscript
IV Recovery discrepancies of OH-PBDEs polybromophenols in human
plasma and cat serum versus herring and long-tailed duck plasma.
Anna-Karin Dahlberg*, Jessica Norrgran*, Lotta Hovander, Åke Bergman
and Lillemor Asplund.
Chemosphere. 2014, 94, 97-103.
ix
Table of contents
1 Introduction ......................................................................................... 1 1.1 Aim of thesis .................................................................................................... 2
2 Background .......................................................................................... 3 2.1 The Baltic Sea .................................................................................................. 3 2.2 Adverse effects in Baltic marine wildlife ..................................................... 5 2.3 Natural production of brominated substances ........................................... 7 2.4 Brominated substances discussed in this thesis ....................................... 8
2.4.1 Polybrominated diphenyl ethers ........................................................ 9 2.4.2 Hydroxylated polybrominated diphenyl ethers ............................. 12 2.4.3 Methoxylated polybrominated diphenyl ethers ............................. 15 2.4.4 Polybrominated phenols and anisoles ............................................ 16
3 Oxidative phosphorylation .............................................................. 18 3.1 Disruption of oxidative phosphorylation ................................................... 19 3.2 Methods to study OXPHOS in vitro ............................................................ 21
3.2.1 The TPP assay ..................................................................................... 21 3.2.2 The TMRM assay ................................................................................. 22 3.2.3 Cytotoxicity ......................................................................................... 22
3.3 Mixture toxicity ............................................................................................. 22
4 Chemical analysis ............................................................................. 24 4.1 Description of species studied .................................................................... 24
4.1.1 Blue mussel (Mytilus trossulus x Mytilus edulis) .......................... 24 4.1.2 Long-tailed duck (Clangula hyemalis) ............................................ 24 4.1.3 Herring (Clupea harengus) ............................................................... 25
4.2 Samples and sampling ................................................................................. 26 4.2.1 Blue mussels ....................................................................................... 27 4.2.2 Long-tailed duck ................................................................................. 27 4.2.3 Herring ................................................................................................. 27
4.3 Analytical methods ....................................................................................... 28 4.3.1 Extraction ............................................................................................ 29 4.3.2 Separation of neutral and phenolic compounds ........................... 29 4.3.3 Derivatization ...................................................................................... 30 4.3.4 Lipid removal ...................................................................................... 30 4.3.5 Instrumental analysis ........................................................................ 31 4.3.6 Quality assurance/Quality control ................................................... 32
x
4.4 An analytical challenge ................................................................................ 33 4.4.1 Further method evaluation (additional results) ............................ 33
5 Results and Discussion .................................................................... 36 5.1 OH-PBDEs potency to disrupt OXPHOS in vitro....................................... 36
5.1.1 TPP data on OH-PBDEs ..................................................................... 36 5.1.2 TMRM data on OH-PBDEs ................................................................. 37 5.1.3 Cytotoxicity ......................................................................................... 38 5.1.4 Synergistic effects .............................................................................. 38
5.2 Blue mussels .................................................................................................. 40 5.3 Long-tailed duck ........................................................................................... 42 5.4 Baltic herring ................................................................................................. 44
6 Concluding remarks and future perspective .............................. 47
7 Sammanfattning (summary in Swedish) .................................... 49
8 Acknowledgement ............................................................................ 52
9 Appendix A......................................................................................... 54
10 References ......................................................................................... 55
xi
Abbreviations
ADP
AhR
APPI
ATP
BAs
BPs
CA
Adenosine diphosphate
Aryl hydrocarbon receptor
Atmospheric pressure photo ionization
Adenosine triphosphate
Brominated anisoles
Brominated phenols
Concentration addition
4,4’-DDE
4,4’-DDT
DNP
EC
ECNI
EI
ER
ESI
ETC
FCCP
GC
GM
GPC
GR
HOCs
HPCs
IA
LLE
LOD
LOEC
LOQ
MeO-PBDEs
HRMS
LRMS
NOEC
OH-PBDEs
OXPHOS
PBDEs
PCBs
1,1-dichloro-2,2-bis(4-chlorophenyl) ethene
1,1,1-trichloro-2,2-bis(4-chlorophenyl) ethane
2,4-dinitrophenol
Effect concentration
Electron capture negative ionization
Electron ionization
Estrogen receptor
Electrospray
Electron transport chain
4-(trifluoromethoxy)phenylhydrazone
Gas chromatography
Geometric mean
Gel permeation chromatography
Glucocorticoid receptor
Halogenated organic compounds
Halogenated phenolic compounds
Independent action
Liquid-liquid extraction
Limit of detection
Lowest observed effect concentration
Limit of quantification
Methoxylated polybrominated diphenyl ethers
High resolution mass spectrometry
Low resolution mass spectrometry
No observed effect concentration
Hydroxylated polybrominated diphenyl ethers
Oxidative phosphorylation
Polybrominated diphenyl ethers
Polychlorinated biphenyls
xii
PCP
PFBCl
POPs
ppm
SF6847
SIM
T3
T4
TBG
TMRM
TPP+
TR
TTR
TU
Pentachlorophenol
Pentafluorobenzolyl chloride
Persistent organic pollutants
Parts per million
3,5-di(tert-butyl)-4-
hydroxybenzylidenemalononitrile
Single ion monitoring
3,3’,5-triiodo-L-thyronine
3,3’,5,5’-tetraiodo-L-thyronine
Thyroid binding globulin
Tetramethylrhodamine methyl ester perchlorate
Tetraphenylphosphonium ion
Thyroid hormone receptor
Transthyretin
Toxic unit
1
1 Introduction
During the 20th century the chemical industry has developed new products
and materials that have considerably improved the life of mankind. How-
ever, the awareness of the risks anthropogenic chemicals can pose has also
grown over the recent half century. Today several halogenated organic com-
pounds (HOCs) of anthropogenic origin are known to be harmful for humans
and wildlife due to their persistent, toxic and bioaccumulative properties,
and their production and use is nowadays banned or restricted[1]. At the
same time, the knowledge regarding natural/biogenic production of HOCs
has increased tremendously[2].
With seawater being rich in halides (i.e. chloride, bromide and iodide) it is
not surprisingly that a large number of HOCs are naturally produced by ma-
rine biota. The chemical structures produced ranges from simple haloalkanes
to complex halogenated aromatic structures[2]. In the Baltic Sea a large
number of brominated phenolic compounds, including hydroxylated
polybrominated diphenyl ethers (OH-PBDEs) have been identified. The
Baltic Sea is a unique brackish water body with a sensitive ecosystem.
During the last half century the Baltic Sea has suffered from chemical pollu-
tion and eutrophication due to human activities, which have severely affect-
ed the ecosystem[3,4].
In recent years have the toxicology sciences done a lot of research on HOCs
and an extensive knowledge base has been created. Many studies have con-
firmed that phenolic metabolites of HOCs in fact can be more biological
active than their neutral parent compounds[5]. OH-PBDEs of both natural
and anthropogenic (metabolic) origin have been linked to several toxicologi-
cal effects, including effects on the endocrine- and nervous systems[6,7].
OH-PBDEs have also been reported to be acutely toxic and induce develop-
mental arrest in zebrafish[8,9]. The underlying mechanism for this toxicity
was suggested by van Boxtel and co-workers to be due to disruption of mito-
chondrial oxidative phosphorylation (OXPHOS)[9]. OXPHOS is the meta-
bolic pathway used by cells to produce energy (i.e. adenosine triphosphate,
ATP) under aerobic conditions. Exposure to compounds which disturb this
mechanism may thus lead to severe biological consequences as shown for
several “classical” uncouplers of OXPHOS[10].
2
1.1 Aim of thesis
This thesis focuses on brominated organic chemicals of natural and/or
anthropogenic origin, with special emphasis on OH-PBDEs present in Baltic
biota. The work presented herein comprises both toxicological and analytical
studies with the overall aim of assessing OH-PBDEs potency for OXPHOS
disruption and to determine the exposure of these compounds in Baltic wild-
life (i.e. in mussel, fish and waterfowl). The specific objectives of each paper
are presented as follows:
Paper I: OH-PBDEs share many structural characteristics with potent dis-
ruptors of OXPHOS (i.e. being weak lipophilic acids). The objective of this
study was to determine the OXPHOS disrupting potential of several OH-
PBDEs present in Baltic biota, using two in vitro bioassays. Single OH-
PBDE congeners as well as mixtures were tested to study potential mixture
effects.
Paper II: The objective was to determine sea ducks exposure to brominated
organic substances by analysing livers from long-tailed duck (Clangula
hyemalis) wintering in the Baltic Sea. Further to assess their dietary intake
prior to migration and reproduction, the concentration of brominated sub-
stances in Baltic blue mussels collected from sites used by sea ducks for
foraging were analysed.
Paper III: The condition and fat content in Baltic herring has shown de-
creasing trends since the 1980’s. Meanwhile has the concentrations of OH-
PBDEs increased in herring livers. Given the toxicity of the OH-PBDEs to
zebrafish, the aim of this study was to further assess the environmental con-
centrations of these compounds in Baltic herrings by analyse OH-PBDEs
and related brominated organic compounds in fish sampled from the north-
ern Baltic Proper and the southern Bothnian Sea.
Paper IV: The aim of this paper was to evaluate the applicability of a liquid-
liquid extraction method used for analysing HPCs in human plasma, for OH-
PBDE analysis in blood matrices from mammal, bird and fish.
3
2 Background
2.1 The Baltic Sea
The Baltic Sea is a semi-enclosed brackish water body located in the north-
ern hemisphere (Figure 1.). The sea is connected to the North Sea via the
three narrow and shallow Danish straits and consists of several deep basins
separated by narrow sills. Large input of freshwater from the catchment area
surrounding the Baltic Sea together with limited water exchange with the
North Sea, results in a salinity gradient ranging from 30‰ down to 3‰ with
increasing distance from the North Sea. Further are the salt- and freshwater
masses in the Baltic Sea vertically stratified and separated by a halocline.
The stratification is caused by colder saltwater having higher density than
freshwater. The stratification prevents vertical mixing of the water and oxy-
genation of deep bottoms. Seasonally a thermocline is also present in the
Baltic Sea which further prevents vertical mixing of the water during sum-
mer[3].
Figure 1. The Baltic Sea with the salinity gradients depicted.
Baltic Proper
Bothnian Sea
BothnianBay
Gulf of Riga
400 km
N
S
30‰
20‰
8‰
7‰
6‰
5‰4‰
3‰
5‰
4‰
3‰
4
Being neither a true marine nor a freshwater sea, the number of species liv-
ing in the Baltic Sea is limited. The few marine and freshwater species
which have adapted to the brackish water environment is under constant
physiological stress due to the osmotic pressure caused by the low salinity of
the sea water. The low biodiversity makes the ecosystem sensitive to exter-
nal pressures such as eutrophication, extensive fishing and hazardous sub-
stances[11].
Today nearly the entire Baltic Sea is eutrophicated. Agricultural land use and
other human activities in the surrounding catchment area results in riverine
and atmospheric discharges of nutrients (e.g. inorganic nitrogen and phos-
phorus) which together with the limited water exchange in the Baltic Sea
have resulted in elevated nutrient levels. Despite actions to reduce inflow of
nutrients into the Baltic Sea, the concentrations of nitrogen and phosphorous
are still too high. High nutrient concentrations favor growth of phytoplank-
ton which results in increased sedimentation of organic matter. During deg-
radation of organic matter large quantities of oxygen is consumed. Together
with the vertical stratification of the water column which prevents oxygena-
tion at greater depths, this has resulted in prevalent anoxia in the deep sea
bottoms of the Baltic Proper and occasional hypoxia in more shallow coastal
areas. Anoxic sediments contribute to increased phosphorous levels in the
Baltic Sea by release of inorganic phosphate. In the Baltic Sea growth of
nitrogen fixating bacteria is limited by the availability of phosphorous. As a
result of high phosphorous levels, large summer blooms of nitrogen fixating
bacteria (e.g. Aphanizomenon flos-aquae and Nodularia spumigena) are now
a reoccurring phenomenon in the Baltic Sea. Increased production of phyto-
plankton and nitrogen fixating bacteria also reduces the light penetration
depth, which affects the distribution and specie composition of the macro
algae community in favor of fast growing filamentous macro algae[3,12].
In addition to eutrophication the Baltic Sea is heavily polluted by anthropo-
genic substances, including persistent organic pollutants (POPs) of which
some have caused severe adverse effects on the Baltic marine wild-
life[13,14]. POPs are chemically stable organic compounds, toxic for hu-
mans and wildlife. Due to their chemical stability these chemicals are only
transformed very slowly in the environment. Further can POPs undergo
long-range transport and thus become wildly distributed. POPs are highly
lipophilic and able to bioaccumulate and biomagnify in the food web, with
the risk of reaching high concentrations in top predators[15]. Given the
threat these chemicals pose on humans and wildlife, 179 parties have signed
a global treaty to eliminate or reduce release of POPs listed in the Stockholm
Convention[16]. Today this list consists of 23 halogenated compounds, and
additionally four halogenated substances are under consideration for inclu-
sion[15].
5
In the Baltic Sea, the concentrations of several POPs listed in the Stockholm
Convention are continuously measured in marine biota by the National Swe-
dish Contaminant Monitoring Programme. POPs included in the monitoring
programme are for example 1,1,1-trichloro-2,2-bis(4-chlorophenyl) ethane
(4,4’-DDT) and polychlorinated biphenyls (PCBs)[17], former used as insec-
ticide and heat exchange fluids in electrical appliances, respectively[18,19].
The concentrations of polybrominated diphenyl ethers (PBDEs), used as
brominated flame retardants, are also continuously measured. PBDEs are
further discussed in section 2.4.1.
In addition to halogenated substances of anthropogenic origin, a large num-
ber of organohalogens are also known to be naturally biosynthesized in the
marine environment[2]. Natural production of brominated substances is
more thoroughly discussed in section 2.3.
2.2 Adverse effects in Baltic marine wildlife
That the Baltic ecosystem is sensitive to external pressures such as hazard-
ous substances become evident in the 1960-70’s when the Swedish popula-
tion of white-tailed sea eagle (Haliaeetus albicilla) faced extinction, due to
reproductive failure[20]. This adverse effect was later found to be primarily
caused by 1,1-dichloro-2,2-bis(4-chlorophenyl) ethene (4,4’-DDE), the main
metabolite and degradation product of technical DDT. The concentration of
4,4’-DDE was strongly correlated with eggshell thinning, which was found
to be the underlying cause for the population decline[13]. At the same time,
other adverse effects such as uterine lesions, sterility and skeletal defor-
mations were observed among Baltic Sea populations of grey seal (Halicho-
erus grypus), harbor seal (Phoca vitulina) and ringed seal (Phoca hispida),
effects likely caused by high concentrations of chlorinated compounds[21].
An extensive conservatory program to save the Swedish population of white-
tailed sea eagle and declining concentrations of DDE and PCB has resulted
in recovery of population[22]. However, white-tailed sea eagles living in the
southern Bothnian Sea still show signs of impaired reproduction, for which
the cause is yet unknown[22]. The grey seal and harbor seal populations
have increased after the population decline in the 1960’s and 1970’s. Where-
as ringed seal and the Kalmarsund subpopulation of harbour seal is still clas-
sified as vulnerable species[23]. Although the reproductive health has im-
proved, pathological changes such as colonic ulcers and thickening of the
adrenal cortices (Adrenocortical hyperplasia) are still observed in Baltic
grey seals[24].
6
Today several adverse effects of unknown cause/s are being observed Baltic
Sea wildlife of which some will be discussed in the following sections.
In grey seals, decreasing trends in blubber thickness has been observed since
year 2000[25]. The fat content in Baltic herring (Clupea harengus) from the
Baltic Sea has also decreased in the Baltic Proper since the 1980’s[25].
For several sea duck species, the Baltic Sea is an important wintering and
breeding area. In 2009 the populations of five of the most common sea duck
species were observed to have declined dramatically (by 4.2 million birds or
60%) since the early 1990’s when the first large survey on waterfowls in the
Baltic Sea was performed. For common eider (Somateria mollissima) and
long-tailed duck (Clangula hyemalis) the population showed decreases of
51% and 65%, respectively[26]. There are various pressures such as preda-
tion, oil pollution, by-catches, hazardous substances, diseases and larger
ecosystem changes, potentially affecting the food quality of sea ducks,
which may have contributed to the decline[26].
Further has common eider and herring gull (Larus argentatus), living in the
Baltic Sea been reported to be suffering from a mortal paralytic syndrome
with clinical symptoms such as paralysis of wings and legs, diarrhea, star-
gazing and ataxia[27], symptoms which were found to be remedied by injec-
tions of thiamine (vitamin B1)[27].Thiamine deficiency has also been report-
ed in Baltic salmon (Salmo salar) and is commonly referred to as the M74
syndrome and affects the development of the salmon fry at the yolk-sac
stage[28]. The cause of M74 is still unresolved but it has been shown to be
associated with thiamine deficiency[29]. In a study determining the concen-
tration of organohalogens in muscle, blood and roe from healthy Baltic
salmon and salmon producing offspring with M74, no association was found
for e.g. DDT, PCBs or PBDEs. However presence of a large number (>100)
of halogenated phenolic compounds (brominated and/or chlorinated), includ-
ing OH-PBDEs were found in the salmon blood[30]. Further work resulted
in structural identification of several OH-PBDEs and methoxylated
polybrominated diphenyl ethers (MeO-PBDEs) of which eight compounds
had not previously been reported in the environment[31].
High mortality in herring roe has also been observed in the Baltic Sea. The
mortality was found to change during the spawning season and follow the
life cycle of filamentous brown algae (i.e. Pilayella littoralis)[32]. Results
from further field and laboratory experiments in which herring roe was ex-
posed to algae exudates, suggests that the increased mortality was due to
toxic substances released by the filamentous brown algae[33].
7
2.3 Natural production of brominated substances
Today more than 4000 halogenated organic compounds are known to be
naturally produced in terrestrial and marine environments by living organ-
isms or formed in natural abiotic processes (e.g. in volcanoes and forest
fires). The chemical structures ranges from simple haloalkanes to complex
halogenated aromatic structures. The majority of these compounds contain
chlorine and/or bromine substituents, but a few organoiodine and organoflu-
orine compounds have also been identified as natural products[2]. For a
thorough review of the diversity of naturally produced organohalogens the
book “Naturally occurring organohalogen compounds – a comprehensive
update” by Gordon Gribble[2] is recommended for further reading.
The marine environment is likely the single largest source for biogenic orga-
nohalogens and there is a plethora of species such as bacteria, algae, spong-
es, corals and sea squirts producing them[2]. Sea water is rich in chloride
ions ( ̴0.5M), whereas bromide ( ̴1mM) and iodide ( ̴1µM) is present in only
smaller amounts[34]. The high halogen content makes sea water a natural
substrate reservoir for haloperoxidases involved in the biosynthesis of orga-
nohalogens in biota. Haloperoxidases are enzymes that catalyze oxidation of
halides (i.e. chloride, bromide and iodide) using hydrogen peroxide, result-
ing in subsequent halogenation of organic compounds[34]. As bromine is
less electronegative than chlorine, bromine is more easily oxidized, which
likely explains the abundance of organobromine compounds in the marine
environment despite bromide ions being present at lower concentration than
chloride ions in sea water[34].
Bromoperoxidases (which can catalyze oxidation of bromide and iodide)
have been found in all classes of marine algae and in other marine organ-
isms[34]. It has been suggested that formation of brominated compounds can
be a pathway for marine algae to scavenge hydrogen peroxide formed during
photosynthesis, which at high concentrations can be harmful to the cell[35].
It has also been suggested that the halogenated organic compounds play a
role as chemical defense in algae against grazers[36] and epiphytic settle-
ment[37].
8
2.4 Brominated substances discussed in this thesis
Among the various chemical classes of brominated substances naturally
produced in the marine environment this thesis focus on a few groups there-
of i.e. simple brominated phenols (BPs), simple brominated anisoles (BAs),
OH-PBDEs and MeO-PBDEs, with special focus on OH-PBDEs. PBDEs of
anthropogenic origin are also discussed as an additional source for
OH-PBDEs via e.g. metabolic transformation. General chemical structures
of the compounds discussed in the thesis are given in Figure 2.
Figure 2 General chemical structures and abbreviations of the compounds discussed.
9
2.4.1 Polybrominated diphenyl ethers
PBDEs have been industrially produced since the early 1970’s and used as
additive flame retardants in various polymers (e.g. polyurethanes, polyam-
ides, polystyrene and styrene copolymers) used in textile, furniture and elec-
tronic equipment[38]. PBDEs and other bromine containing flame retard-
ants, prevent fire to propagate by scavenge the free radicals formed during
combustion[38]. However, since PBDEs are not chemically bound to the
material, they are prone to migrate out of the material and subsequently leak
into the environment.
Commercially PBDEs have been produced in three technical mixtures
known as PentaBDE, OctaBDE and DecaBDE, classified according to their
average bromine content[39]. The PBDE congener composition of some
commercial PentaBDE, OctaBDE and DecaBDE products are presented
elsewhere[40]. Since 2009, Penta- and OctaBDE are included in the Stock-
holm Convention on Persistent Organic Pollutants (POPs) under Annex
A[1], and the use of DecaBDE is restricted within the European Union by
the RoHS directive[41]. In the United States companies committed to phase
out DecaBDE by the end of 2013[42]. In 2001 the estimated global use of
PentaBDE, OctaBDE and DecaBDE were 7,500, 3,800 and 56,000 metric
tonnes, respectively[39]
2.4.1.1 Physicochemical properties
PBDEs are lipophilic compounds with low vapor pressures. PBDEs physico-
chemical properties are influenced by their degree of bromination. For ex-
ample does their lipophilicity (log Kow) increase with increasing number of
bromine substituents, i.e. from 5.9-6.2 for tetraBDEs, 6.5-7.0 for pentaB-
DEs, 8.4-8.9 for octaBDEs and 10 for decaBDE[43]. On the opposite, their
vapor pressure decrease with increasing number of bromine substituents.
Conformational studies of PBDEs have shown that they exist in twist to
skewed conformations where increasing number of bromine substituents
ortho to the diphenyl ether linkage force the molecule to adopt a more
skewed conformation[44,45]
2.4.1.2 Environmental occurrence – with focus on the Baltic Sea
Today PBDEs have become ubiquitous environmental contaminants found
also in remote areas like the Arctic. For levels and trends of PBDEs in the
environment several reviews are recommended[46-48]. PBDEs have shown
to bioaccumulate in lipid rich tissue and biomagnify in marine food
webs[46].
10
In the Baltic Sea, PBDEs have been reported in marine top predators such as
ringed seal (liver)[49] and white tailed sea eagle (egg and serum)[50,51].
Increasing concentrations of PBDEs (i.e. BDE-47, BDE-100, BDE-99) were
observed in Guillemot (Uria aalge) eggs from the 1960’s until early 1990’s,
when the concentrations started to decrease[17]. Similar trends with increas-
ing concentrations from the 1970’s to the mid 1980’s, were also observed in
cod (Gadus morhua) liver in the Baltic Sea[17]. In herring (Clupea ha-
rengus) muscle, decreasing concentrations are seen at most sampling sites
included in the Swedish Monitoring Program since the late 1990’s, but the
concentrations are generally higher in the Baltic Proper compared to the
Swedish west coast[17].
2.4.1.3 Metabolic transformation of PBDEs
There are several studies on metabolic transformation of PBDEs which have
been reviewed elsewhere[39,52]. For highly brominated PBDEs reductive
debromination has been suggested to be the first metabolic step, followed by
hydroxylation (inserted either directly or formed via an arene oxide interme-
diate) by cytochrome P450 enzymes. Formation of OH-PBDEs metabolites
have been confirmed in several studies[39]. In one study, sixteen OH-
PBDEs and two di-OH-PBDEs were identified in rat plasma five days after
the rats were given an intraperitoneal dose of equimolar concentrations of
seven PBDEs (i.e. BDE-47, BDE-99, BDE-100, BDE-153, BDE-154, BDE-
183 and BDE-209)[53]. The metabolites formed had the hydroxyl group
inserted in either meta or para position to the diphenyl ether bond[53]. For-
mation of meta and para substituted OH-PBDEs metabolites have also been
reported by others and seem to be a structural characteristic for OH-PBDEs
originating from metabolic transformation[52]. However, it should be stated
that a few ortho substituted OH-PBDEs have also been reported as metabo-
lites[31,54]. OH-PBDE metabolites have also been confirmed in fish i.e.
northern pike (Esox lucius) and common sole (Solea solea), after PBDE
dietary exposure[55,56]. However, in common sole OH-PBDE formation
was found to be a minor transformation route compared to metabolic trans-
formation via debromination[56].
2.4.1.4 Toxicological effects
Several experimental animal studies have shown that prenatal or neonatal
exposure to PBDEs has effects on thyroid homeostasis, neurobehavioral and
reproductive systems. Although the thyroid hormone system seem to be the
main target by which PBDE elicit its toxicity, reproductive effects associated
with estrogenic and androgenic processes have also been reported. For a
more comprehensive understanding of the toxicological effects of PBDEs,
the following articles are recommended[6,39,57]
11
Disruption of thyroid hormone system has been studied in rodents and found
to be associated with reduction in thyroxine (T4) concentrations in the
blood[39]. 3,3’,5,5’-tetraiodo-L-thyronine (T4) is a prohormone produced by
the thyroid gland and a precursor for formation of the active thyroid hor-
mone, 3,3’,5-triiodo-L-thyronine (T3). The thyroid hormone plays an essen-
tial role for normal fetal development and for regulation of growth and basal
metabolism throughout life. The T3 hormone acts by binding to the nuclear
thyroid hormone receptor (TR) which initiates transcription of thyroid hor-
mone responsive genes[58]. Metabolites of PBDEs (i.e. OH-PBDEs) have
shown strong affinity to both the thyroid hormone receptor[59] and thyroid
hormone transporting proteins; transthyretin (TTR) and thyroid binding
globulin (TBG)[60,61]. Interestingly, PBDEs show no binding affinity to
TTR and no or only weak affinity to TBG[60,61], which indicate that meta-
bolic transformation is important for PBDEs to elicit this effect in vivo.
Thyroid hormones play an important role during early brain develop-
ment[58], and accordingly behavioral effects of PBDE exposure have been
studied. In one of the first studies published, Eriksson et al (2001) exposed
10 day old mice to a single oral dose of BDE-47 and BDE-99 and reported
dose-dependent permanent aberrations in spontaneous behavior (i.e. locomo-
tion, rearing and total activity) in 2- and 4- month old mice[62]. Effects also
observed after exposure of other PBDE congeners using similar experi-
mental design, as reviewed elsewhere[6]. Together these studies strongly
indicate that PBDE causes adverse neurobehavioral effects. Further, PBDEs
have been shown to affect neuronal cell viability, cell differentiation and
migration and neuronal signaling in vitro, as reviewed elsewhere[6]. For
some neurotoxic endpoints such as altered calcium homeostasis and induced
vesicular neurotransmitter release (exocytosis), 6-OH-BDE47 showed a
higher potency than BDE-47[63].
PBDEs have also shown to affect other endocrine pathways than the thy-
roidal. For example were anti-androgenic, estrogenic and anti-estrogenic
activities observed in vitro[64]. A recent publication also reported altered
expression of genes regulated by the aryl hydrocarbon receptor (AhR), es-
trogen receptor (ER) and glucocorticoid receptor (GR) in zebrafish larvae
exposed to BDE-47[65].
12
2.4.2 Hydroxylated polybrominated diphenyl ethers
OH-PBDEs may originate from metabolic transformation of PBDEs as
discussed in chapter 2.4.1.3 or from natural production. Naturally produced
OH-PBDEs are widespread in the marine environment and have been report-
ed in various marine species such as sponge[66,67], sea squirt[68], macro
algae[69]and cyanobacteria[69,70]. Biogenic demethylation of naturally
produced MeO-PBDEs has also been confirmed as a pathway for formation
of OH-PBDEs in fish[71,72].
2.4.2.1 Physicochemical properties
OH-PBDEs are weak lipophilic acids and as for PBDEs their physicochemi-
cal properties are dependent on the degree of bromination. For the OH-
PBDEs studied in paper I, the log Kow and pKa ranged between 5-8 and 5-9,
respectively. The acidity of OH-PBDEs makes them depending on pKa fully
or partly deprotonated in sea water which result in higher water solubility of
the compounds.
2.4.2.2 Biosynthesis in marine bacteria and algae
OH-PBDEs are believed to be biosynthesized via formation of BPs and free
radical dimerization[73](Figure 3). 4-Hydroxybenzoic acid has been pro-
posed as precursor for formation of 2,4-dibromophenol (2,4-diBP) and
2,4,6-tribromophenol (2,4,6-triBP) in both algae[74] and marine bacte-
ria[75]. Bromination at the 2 and 4 position on the phenolic ring is likely
favored due to the ortho and para directing property of the hydroxyl group.
The precursor, 4-hydroxybenzoic acid can be formed from chorismate, an
intermediate in the shikimate acid pathway[73] which is the common ana-
bolic pathway in bacteria and plants for formation of the aromatic amino
acids phenylalanine, tyrosine and tryptophan[76]. Hydroxybenzoic acid has
also been proposed to be formed from degradation of tyrosine in marine
algae[74].
13
Figure 3 Proposed biosynthetic pathway for formation of hydroxylated polybrominated diphenyl ethers.
Modified figure adopted from [73,76].3-deoxy-D-arabino-heptulosonate (DAHP), 3-dehydroquinate
(DHQ), 5-enolpyruvylshikimate-3-phosphate (EPSP)
Dimerization of BPs to OH-PBDEs has been shown to be catalyzed by
bromoperoxidase in algae[77] and by cytochrome P450 in marine
bacteria[75]. Theoretically, dimerization via radical formation can result in
OH-PBDEs having the hydroxyl group in either ortho and para position to
the diphenyl ether bond as shown in Figure 3. However, naturally produced
OH-PBDEs identified in the marine environment are almost exclusively
ortho substituted. However, three para-substituted OH-PBDEs were recently
reported to be formed in vitro, by enzymes from marine bacteria[75].
2.4.2.3 Environmental occurrence – with focus on the Baltic Sea
OH-PBDEs are naturally produced in the Baltic Sea, by cyanobacteria
(Aphanizomenon flos-aquae, Nodularia spumigena)[70,78], red algae
(Ceramium tenuicorne, Furcellaria lumbricalis)[69,70], brown algae
(Pilayella littoralis, Dictyosiphon foenicolaceus)[69,78] and green algae
(Cladophora glomerata, Enteromorpha intestinalis)[69]. The concentration
of OH-PBDEs in Ceramium tenuicorne and Cladophora glomerata have
been shown to vary seasonally, with highest concentrations in the sum-
mer[69]. Several ortho substituted OH-PBDEs have also been found in blue
mussels (Mytilus edulis) in high concentrations (50ng/g f.w.)[79]. Further-
more have the concentration of OH-PBDE in blue mussels been observed to
mimic the seasonal variation in algae[69,79].
14
Several OH-PBDEs have also been confirmed or tentatively identified in
Baltic salmon (Salmo salar) plasma[31]. The structures of the major com-
pounds found suggest that they originate from natural sources rather than
PBDE metabolism[31]. OH-PBDEs have also been found in ringed seal
plasma[49], common guillemot eggs[80] and in serum from white-tailed sea
eagle nestlings[51]. A recent study has also shown that the concentrations of
2’-OH-BDE68 and 6-OH-BDE47 in Baltic herring has increased from 1980
to 2010[25,81].
2.4.2.4 Toxicological effects
The toxicity of OH-PBDEs is not as well studied as for PBDEs. However,
the scientific interest has grown over the last decade as more studies indicate
that OH-PBDEs are highly biological active, having both endocrine and
neurotoxic effects[6,7].
OH-PBDEs have potential to disrupt the thyroid hormone system and action.
OH-PBDEs have shown strong binding affinity to both the thyroid hormone
receptor[59] and thyroid transporting proteins (TTR and TBG)[60,61]. The
high affinity are likely due to OH-PBDEs structural resemblance to the thy-
roid hormones T3 and T4, as exemplified with 4’-OH-BDE121 in Figure 4.
Effects on other endocrine pathways than the thyroidal have also been re-
ported for OH-PBDEs. Low brominated OH-PBDEs have for example been
reported to show estrogenic activity whereas higher brominated OH-PBDEs
showed anti-estrogenic activity in vitro[82]. The difference in estrogenic
activity between high and low brominated OH-PBDEs was found to corre-
late with the observation of two different binding modes to the estrogen re-
ceptor which were dependent on the molecular size of the OH-PBDE[82].
Further, 6-OH-BDE47 has shown anti-androgenic activity in vitro, having
similar potency as the anti-androgenic drug flutamide[64].
Figure 4. Chemical structures of 4’-OH-BDE121 and the thyroid hormones T3 and T4.
15
6-OH-BDE47 is acutely toxic to adult and developing zebrafish when ex-
posed to concentrations in the nanomolar range[9]. Developmental defects
such as pericardial edema, yolk sac malformations, reduced pigmentation
and lowered heart rate and developmental arrest were observed in the
zebrafish embryo[9]. Microarray analysis of gene expression in the zebrafish
embryonic fibroblast cells (PAC2) led to identification of two functional
gene groups involved in carbohydrate metabolism and proton transport. Fur-
ther work measuring respiration and membrane potential in isolated
zebrafish mitochondria showed that 6-OH-BDE47 inhibits complex II in the
mitochondrial electron transport chain[9].
Developmental arrest in zebrafish has also been observed for 3-OH-BDE47,
5-OH-BDE47[8]. In this study genes involved in thyroid hormone regula-
tion, neurodevelopment and stress response were found to be upregulated
after exposure to 6-OH-BDE47[8]. Others have reported altered expression
of genes controlled by ER, mineralocorticoid receptor and AhR in zebrafish
larvae exposed to 6-OH-BDE47[65]. 6-OH-BDE47 have also been reported
to cause cytotoxic effects (i.e. increased apoptosis, cell cycle block and
cause DNA damage) and oxidative stress in human hepatoma cells[83]
2.4.3 Methoxylated polybrominated diphenyl ethers
MeO-PBDEs are naturally produced in the marine environment and to the
best of the author’s knowledge there are no known anthropogenic sources for
MeO-PBDEs.
2.4.3.1 Physicochemical properties
MeO-PBDEs are neutral lipophilic compounds and their physicochemical
properties are influenced by their degree of bromination. The methoxylated
analogues to the OH-PBDEs studied in paper I have log Kow values rang-
ing from 6 to 8, based on computer modeling calculations using ACD/Labs
software, version 11.02.
2.4.3.2 Environmental occurrence – with focus on the Baltic Sea
MeO-PBDEs have been found in cyanobacteria (Aphanizomenon flos-aquae,
Nodularia spumigena)[70,78], red algae (Ceramium tenuicorne, Furcellaria
lumbricalis)[69,70], brown algae (Pilayella littoralis, Dictyosiphon foenico-
laceus)[69,78] and green algae (Cladophora glomerata, Enteromorpha intes-
tinalis)[69] from the Baltic Sea. The concentration of MeO-PBDEs has
shown to vary seasonally in blue mussel, but with increase in summer being
less pronounced as for OH-PBDEs[79]. MeO-PBDEs have also been report-
ed in Baltic herring[84], common guillemot eggs [80], white-tailed sea eagle
eggs[50] and in serum from white-tailed sea eagle nestlings[51].
16
2.4.3.3 Toxicological effects
Information regarding MeO-PBDEs toxicity is limited. However one study
reports teratogenic and morphological effects after exposing zebrafish em-
bryo to 5-MeO-BDE47[85]. Another study showed no significant develop-
mental alterations in zebrafish development with 6-MeO-BDE47 during
early development. However, pericardial edema and curved spine was ob-
served in exposed larvae after 120 hours post fertilization[65]. Biotransfor-
mation of 6-MeO-BDE47 in the zebrafish embryo/larvae was also studied
and interestingly the 6-OH-BDE47 concentration increased in a time de-
pendent manner[65]. This may be an indication that the effect observed is
caused by the more potent 6-OH-BDE47 rather than 6-MeO-BDE47.
MeO-PBDEs potential to disturb thyroid hormone homeostasis is likely lim-
ited since MeO-PBDEs (i.e. 6-MeO-BDE47 and 2’-MeO-BDE68) have
shown only slight or no binding affinity to TTR and TBG[60].
2.4.4 Polybrominated phenols and anisoles
BPs found in the environment can originate from both natural and anthropo-
genic sources[2,86]. 2,4-diBP, 2,4,6-triBP and pentabromophenol (PentaBP)
are industrially produced chemicals and used as reactive intermediates in
production of brominated resins and polymers[86]. 2,4,6-TriBP is the most
widely used simple brominated phenol[87] and has also been used as a wood
preservative[86]. Simple brominated phenols can also be formed from deg-
radation of brominated benzenes and PBDEs in the environment[86]. To the
best of the author’s knowledge, no anthropogenic source for BAs is known.
2.4.4.1 Physicochemical properties
BPs are lipophilic acids with low vapor pressures. Their log Kow range from
3 (2,4-diBP) to 5 (pentaBP) and their pKa range from 8 (2,4-diBP) to 4 (pen-
taBP)[86]. Corresponding anisoles lipophilicity range from log Kow 4 (2,4-
diBA) to 6 (pentaBP) based on computer modeling calculations using
ACD/Labs software, version 11.02.
2.4.4.2 Environmental occurrence – with focus on the Baltic Sea
The concentration of 2,4,6-triBP in herring (liver) have increased with 6%
during 1980-2010 in southern Baltic Proper[25,81]. Whereas no such trend
is visible in the northern Baltic Proper and likewise, not in the southern
Bothnian Sea either[25,81].
2.4.4.3 Toxicological effects
The toxicity BPs have been reviewed by others [86,88], while data on toxici-
ty of BAs seem to be lacking. Dietary exposure to 2,4,6-triBP have been
17
reported to affect reproduction (i.e. reduced fertilization success and dis-
turbed gonad morphology) in zebrafish[89]. Altered sex ratio towards males
has also been observed in zebrafish after chronic exposure to environmental
levels of 2,4,6-triBP during embryo and larvae development[90]. Further,
were offspring to exposed fish found to be affected by malformations, re-
tarded growth and reduced survival[90]. 2,4,6-triBP have also been reported
to bind to TTR with 10 times higher binding affinity to TTR than T4[64].
18
3 Oxidative phosphorylation
Mitochondria are intracellular organelles that play a pivotal role in energy
production in eukaryotic cells as the majority of all adenosine triphosphate
(ATP) is produced there. A mitochondrion is composed of two lipid mem-
branes. The outer mitochondrial membrane separates the organelle from the
cytosol whereas the inner mitochondrial membrane encloses the mitochon-
drial matrix. The mitochondrial matrix contains enzymes involved in key
catabolic processes such as the tricarboxylic acid cycle (TCA cycle) and
fatty acid (β) oxidation. The inner mitochondrial membrane is impermeable
to ions and contains the membrane bound proteins involved in oxidative
phosphorylation (OXPHOS)[10,91].
OXPHOS is the main metabolic pathway used by cells to produce ATP un-
der aerobic conditions. The process of OXPHOS involves the electron
transport chain (ETC) and the ATP synthase (Figure 5). The ETC consist of
four membrane bound complexes (I-IV) which transfer electrons from
NADH and Succinate (derived from catabolism of nutrients) via series of
redox-reactions to the final electron acceptor, molecular oxygen. The energy
released during these redox-reactions is used by complex I, III and IV to
create a proton gradient (ΔpH+) and a separation of charge (ΔΨ) across the
inner mitochondrial membrane. The ATP synthase use the energy stored in
this electrochemical proton gradient (i.e. the proton-motive force) to fuel
phosphorylation of adenosine diphosphate (ADP) to ATP[10]. Under normal
conditions the proton motive force is approximately 180 mV[92]. The rate of
substrate oxidation and electron transfer in the ETC is tightly coupled to the
electrochemical proton gradient, a process known as respiratory control.
When the cellular energy (ATP) demand is high, the electrochemical gradi-
ent decrease which stimulates substrate oxidation by the ETC[10].
19
Figure 5. Simple representation of the complexes involved in the electron transport chain (ETC) and the
ATP synthase located in the inner mitochondrial membrane. The electron flow and proton transfer is
depicted with red and orange arrows, respectively.
NADH dehydrogenase (Complex I), succinate dehydrogenase (Complex II), Ubiquinone (Q),
ubiquinone:cytochrome c oxidoreductase (Complex III), cytochrome c (Cyt c) and
cytochrome c oxidase (Complex IV)
3.1 Disruption of oxidative phosphorylation
The tight control between ETC substrate oxidation, the electrochemical pro-
ton gradient and ATP production, depends greatly on the structural integrity
of the inner mitochondrial membrane and its impermeability to ions. An
uncontrolled transfer of ions across the inner mitochondrial membrane
against the proton gradient created by the ETC will affect the electrochemi-
cal proton gradient negatively and result in increased substrate oxidation and
oxygen consumption. In such case, the mitochondrion is utilizing energy
(metabolic substrates) without it being coupled to ATP formation. Instead
the energy from substrate oxidation is released as heat. Exposure to agents
dissipating the electrochemical potential in this manner will result in in-
creased substrate oxidation but reduced ATP production. Rapid depletion of
ATP can be acutely toxic, whereas low dose exposure over a prolonged time
can lead to less efficient energy metabolism (i.e. body weight loss) and in-
creased thermogenesis[10,93].
20
Some xenobiotics cause such uncontrolled proton leak by damaging the
structural integrity of the inner mitochondrial membrane. Other xenobiotics
act as proton carriers across the membrane, translocating protons from the
intermembrane space to the mitochondrial matrix[10]. Generally, so called
protonophoric uncouplers share common structural characteristics such as
being lipophilic weak acids (pKa 5-7) with an acid dissociable group, a
strong electron withdrawing moiety and bulky lipophilic groups[94]. Substi-
tuted phenols are a well-studied class of protonophoric uncouplers and both
mono-molecular and bi- molecular models have been proposed to describe
their proton translocation in the mitochondrial membrane[10] (Figure 6).
Examples of potent protonophoric uncouplers are e.g. 2,4-dinitrophenol
(DNP) and pentachlorophenol (PCP)[94].
Figure 6 Molecular models of proton translocation across the inner mitochondrial membrane by substi-
tuted phenols. a) Monomolecular model b) bimolecular model. The protonated (HA) and
deprotonated (A-) forms of the substituted phenol are depicted.
Other mitochondrial toxins exert their toxicity by inhibiting the protein com-
plexes involved in the ETC. Dependent on which complex is inhibited this
can hamper or stop the electron and proton transfer by the ETC, resulting in
suppressed substrate oxidation, oxygen consumption and ATP produc-
tion[10]. Examples of potent inhibitors are rotenone (which inhibits Com-
plex I), Antimycin A (inhibits complex III), potassium cyanide (inhibits
complex IV)[95]. There are also compounds which inhibit the ATP synthase
(e.g. Oligomycin[95]), resulting in decreased ATP formation[10].
Other compounds (e.g. Valinomycin[96]) depolarize the membrane potential
by facilitate transfer of small ions across (e.g. K+) the inner mitochondrial
membrane. These so called ionophores can either form ion channels or lipid
soluble ion complexes which mediate ion transport across the mem-
brane[10].
21
3.2 Methods to study OXPHOS in vitro
OXPHOS can be studied in vitro in both isolated mitochondria and in cells.
Usually measurement of both mitochondrial respiration and membrane po-
tential provides the best mechanistic information. The methods using isolat-
ed mitochondria are generally well established and the experimental settings
can be modified (e.g. by using inhibitors) to provide mechanistic insight how
xenobiotic disturb OXPHOS. However, the drawback of using isolated mito-
chondria is the lack of the cellular context. Hence, cell assays provides a
more physiological relevant test condition as all cellular functions are pre-
served in the cell. However, the impermeability of the cell membrane to
some commonly used substrates and inhibitors used to study OXPHOS can
be problematic. Nevertheless, new techniques to study bioenergetics in intact
cells are increasing as one advantage is the possibility for high throughput
screening[92].
In paper I the OXPHOS disrupting potential of 18 OH-PBDE congeners of
both natural and anthropogenic (metabolic) origin, were determined using a
classic liver mitochondrial assay (TPP assay) and a mitochondrial potential
cell assay (TMRM assay).
3.2.1 The TPP assay
The TPP assay is a well-established assay in which respiration (O2) and
membrane potential (ΔΨ) is measured simultaneously in isolated mitochon-
dria[95,97]. The experiment was performed in a closed vessel and the oxy-
gen level was measured using a Clarke-type electrode[95]. The membrane
potential was assessed by help of a tetraphenylphosphonium ion (TPP+) elec-
trode. TPP+ is a lipid soluble cation which is transferred electrophoretically
into the mitochondria where it accumulates in proportion to the membrane
potential[97]. Hence, the amount of free TPP+ in the reaction medium is a
relative measurement of ∆Ψ. The test compound was added during titration
and by recording changes in respiration and membrane potential simultane-
ously, effects on OXPHOS could be studied. 6-OH-BDE47 has previously
been reported to be a potent inhibitor of complex II of the ETC[9]. By add-
ing rotenone to inhibit complex I, substrate oxidation could only be mediated
by complex II. This allowed us to study inhibitory effects as well as pro-
tonophoric uncoupling in the same experiment. A decrease in membrane
potential and respiration after addition of test compound is indicative for
inhibition of ETC, whereas a decreased membrane potential and increased
respiration is symptomatic for protonophoric uncoupling.
22
3.2.2 The TMRM assay
Tetramethylrhodamine methyl ester (TMRM) is a fluorescent lipophilic rho-
damine dye which accumulates in the mitochondrial matrix in proportion to
the mitochondrial membrane potential (ΔΨ)[98]. In paper I, TMRM was
used to study changes in mitochondrial membrane potential in zebrafish
embryonic fibroblast (PAC2) cells after addition of test compound. The
zebrafish cell line was chosen as it provides an outlook for future toxicity
studies using zebrafish and zebrafish embryo. By measure the TMRM fluo-
rescence in the cell spectrophotometrically a relative measurement of the
mitochondrial membrane potential was obtained. The TMRM method is
described in detail in paper I. The bioassay responds to compounds which
alter the mitochondrial membrane potential but provides no detailed mecha-
nistic information.
3.2.3 Cytotoxicity
Two bioassays (LDH- and MTT assay) were used in paper I to detect poten-
tial cytotoxic effects to confirm that the altered mitochondrial membrane
potential observed in the TMRM assay was not due to nonspecific cytotoxi-
city caused by other mechanism than OXPHOS disruption, mitochondrial
toxicity or disturbed mitochondrial membrane integrity.
3.3 Mixture toxicity
OH-PBDEs are found as mixtures in environment. The toxicity of chemicals
in combination can differ from what predicted from the toxicity of single
compounds as the compounds can interact which may either increase or de-
crease the toxic effect of the mixture. There are in principle two concepts
which are used to assess combined effects of chemicals (i.e. concentration
addition and independent action). Both models describe the quantitative rela-
tionship between the toxicity of simple compounds and their mixture, but are
based on very different theories on how to assess mixture toxicity[99].
Concentration addition (CA) is built on the assumption that the compounds
included in the mixture share a common mode of action and behave as they
are simple dilutions of one another, showing no interactions. In such case,
the effect exerted by one compound could be replaced by another compound
contributing to an equal toxic effect without changing the overall mixture
toxicity. The mixture toxicity can thus be described by summarizing each
compounds effect contribution to the mixture, also called toxic unit (TU)
summation. The effect contribution is based on the single compounds indi-
vidual potencies (e.g. EC50 or EC10) and their concentrations in the mixture.
23
An important feature with the CA model is that it takes into account all
compounds in the mixture and their effect contribution to the complete mix-
ture toxicity, including compounds which are below their No observed effect
concentration (NOEC) when tested as single compounds[99].
The model of independent action (IA) on the other hand is a probability
based method using the statistical concept for independent random events to
calculate the combined effects of chemicals. In the model of IA the effect
contribution of the single compounds making up the mixture are assessed
separately, assuming no interactions. Compounds which are below NOEC
will not contribute to the calculated mixture effect using IA. The model of
independent action is used to assess mixtures which contain compounds
which exert their toxicity by different modes of action[99].
As both the CA and IA model assume no interactions amongst the chemicals
in the mixture. Observed deviations from the predicted effect value by either
method can thus be used to describe and quantify interactions (i.e. synergism
or antagonism) within a mixture[99]. In paper I mixture effects of OH-
PBDEs potency to disrupt OXPHOS were studied in the TMRM assay. An
artificial mixture (composed of four OH-PBDEs) was tested and the com-
pounds were combined so that each compound contributed equally to the
mixture’s toxicity at each dose level (for details see paper I). The choice of
using equitoxic concentrations of OH-PBDEs was derived from the TU con-
cept in the CA model, using altered mitochondrial membrane potential as
shared toxic endpoint. If the OH-PBDEs act as dilutions of each other (i.e.
show additive effects but no interactions) the mixtures effect concentration
(EC50 or EC10) should be one TU.
In addition to artificial mixtures, a mixture composed of seven OH-PBDEs
reported in Baltic blue mussels were assessed in paper I. The seven com-
pounds (i.e. 6-OH-BDE47, 2’-OH-BDE68, 6-OH-BDE85, 6-OH-BDE90, 6-
OH-BDE99, 2-OH-BDE123 and 6-OH-BDE137) were combined at fixed
ratios to mirror the internal exposure in blue mussel. The mixture was tested
at dose levels 1 - 1000 times above the reported environmental concentra-
tion. Details regarding concentrations and compound ratios tested are given
in paper I.
24
4 Chemical analysis
4.1 Description of species studied
4.1.1 Blue mussel (Mytilus trossulus x Mytilus edulis)
The Baltic blue mussel is a marine suspension feeding bivalve that success-
fully has adapted to the brackish water environment in the Baltic Sea. The
Baltic blue mussel differ genetically[100], physiologically[101] and morpho-
logically (i.e. being smaller in size and more elongated) from marine living
blue mussels[102]. Since predation pressure and competition with other spe-
cies are negligible in the Baltic Proper, blue mussels have become a biomass
dominant that colonize hard bottoms down to approximately 30 m
depths[103]. The distribution of blue mussels in the Baltic Sea is restricted
by the salinity gradient, and their abundance decrease rapidly north of the
Åland Sea, due to lowered salinity[103].
4.1.2 Long-tailed duck (Clangula hyemalis)
The long-tailed duck is a small mussel feeding sea duck (Figure 7) found in
the northern hemisphere. The long-tailed duck breeds in fresh water habitats
in northern Europe and western Siberia but migrate to ice-free brackish and
marine waters in the winter to forage[26]. The Baltic Sea is the most im-
portant wintering area for the European population of long-tailed ducks. The
majority of these sea ducks breeds in western Siberia but migrate to the
Baltic Proper during October-December where they stay until mid-May.
During winter and spring, the birds are found foraging on mussel banks
located offshore and along the coasts of the Baltic Proper [26]. Long-tailed
ducks dive and forage down to 35 m depths and in the Baltic Sea their diet
consists mainly of bivalves, supplemented with other benthic organisms,
crustaceans and small fish and fish egg[104-106].
25
In 2009, the population of long-tailed duck was surveyed and found to con-
sist of 1,482,000 birds in the Baltic Sea. This represents a population decline
by 65% (or 2.8 million birds) since the early 1990’s[26] and long-tailed duck
is now classified as a vulnerable specie on the IUCN global red list of threat-
ened species[107]. Dramatic population declines have also been observed for
other mussel feeding sea duck species e.g. common eider (Somateria mollis-
sima) and velvet scoter (Melanitta fusca), in the Baltic Sea[26].
Figure 7 Long-tailed ducks, one female and four males in flight. Photo by: Kjell Larsson
4.1.3 Herring (Clupea harengus)
Baltic herring and sprat (Sprattus sprattus) are the two dominant pelagic
zooplanktivorous species in the Baltic Sea and prey for e.g. cod (Gadus
morhua), salmon (Salmo salar) and seals (e.g. Halichoerus grypus )[108].
From 1980’s to 2010, decreasing trends in condition (weight versus length)
and fat content in Baltic herring has been observed in the Bothnian Bay
(Harufjärden), the northern Baltic Proper (Landsort) and in the southern
Baltic Proper (Utlängan)[17]. During the last three decades, dramatic chang-
es in the fish community has been observed with a large decrease in popula-
tion of cod and an increased abundance of sprat[109]. As sprat and herring to
some extent share the same feeding preferences[110], an increased competi-
tion with sprat over food resources can have affected the condition of the
herring negatively.
26
High mortality of herring roe has been observed in the Baltic Sea as men-
tioned in chapter 2.2. Baltic herring spawn in coastal areas and depending on
region, the spawning season varies from March to June[111]. Herring place
their roe on the sea bottom in shallow coastal waters and a study of the
spawning bed selection of herring has shown that herring prefer hard bot-
toms with rich vegetation[112]. Algae species used by herring as substrate
for their roe are for example Pilayella sp. and Cladophora sp[112]. Both
these species from the Baltic Sea have shown to contain OH-PBDEs, and
especially Pilayella can contain high concentrations in June[69].
4.2 Samples and sampling
The sampling sites of blue mussel, long-tailed duck and herring analysed in
paper II-IV are presented in Figure 8
Figure 8. Sampling sites of blue mussel, long-tailed duck and herring.
Gotland
400 km
N
S
Blue mussel
Long-tailed duck
Herring
27
4.2.1 Blue mussels
The blue mussels analysed in paper II were collected during March, May
and June (2011/12) from areas used by sea ducks for foraging. The sampling
sites included both coastal areas as well as mussel banks in open waters
(Figure 8). The sampling depth ranged from 6-19 m, depending on site. The
mussels were collected using a bottom scraper dragged from a boat. The
mussels (1-2 cm of length) were dissected and the mussel soft body was
removed from the shell by tweezers. Mussels collected from the same sam-
pling site and sampling time, were pooled and the tissue homogenized to
reduce the effects between individuals.
4.2.2 Long-tailed duck
The long-tailed duck (n=10) analysed in paper II were collected from two
locations in the Baltic Sea in February 2000 and May 2009 (Figure 8) The
ducks collected in February (n=8) were from an offshore mussel bank (Ho-
burgs bank) located south of Gotland. The birds had drowned in fishnets and
were collected by local fishermen. Two birds from Åland archipelago were
shot by authorised hunters according to permit. The liver was dissected and
the tissue homogenised, each bird was analysed individually.
Plasma from long-tailed duck was used in paper IV. The blood was collect-
ed from birds post mortem and stored in heparin treated tubes. The blood
was separated into plasma and red blood cells by centrifugation.
4.2.3 Herring
Herring analysed in paper III were collected from two sites in the Baltic Sea
(Figure 8). The fish (n=12) from Askö (May 2012) were collected using
gillnets according to ethical permit. The fish from Ängskärsklubb (June
2012) (n=12) were sampled within the Swedish Environmental Monitoring
Program and kindly provided by the Swedish Museum of Natural History.
The fish were sexed post mortem and found to be sampled prior to spawning.
Plasma from herring was used in paper IV. The blood was sampled from the
caudal vein with a syringe treated with heparin. The fish were euthanised
directly after blood sampling according to ethical procedures and permit.
The blood was separated into plasma and red blood cells by centrifugation.
28
4.3 Analytical methods
A simple schematic presentation of extraction and sample clean-up applied
in paper II, III and IV are presented in Figure 9.
Figure 9. Scheme of the extraction and sample clean-up used in paper II, III and IV.
Denaturation/Extraction2-propanol, iso-hx:DEE (3:1 v/v)
Gravimetric lipid determination
Paper II and IIIBlue mussel, herring and long-tailed duck (liver)
SeparationKOH partitioning
PhenolsNeutrals
Instrumental analysisGC-LRMS(ECNI)
Lipid removalH2SO4 treatment
andH2SO4:SiO2 column
Lipid removalGel permeation chromatography
(long-tailed duck only)
Lipid removalH2SO4 treatment
andH2SO4:SiO2 column
Paper IVPlasma/serum from
human, cat, herring and long-tailed duck
DerivatizationDiazomethane
Denaturation/Extraction2-propanol, HCl (6M), c-hx:MTBE (1:1 v/v)
Gravimetric lipid determination
SeparationKOH partitioning
PhenolsNeutrals
Instrumental analysisGC-LRMS(ECNI)
Lipid removalH2SO4 treatment
andH2SO4:SiO2 column
Lipid removalH2SO4 treatment
andH2SO4:SiO2 column
DerivatizationDiazomethane
29
4.3.1 Extraction
Over the years several extraction methods have been developed for extrac-
tion lipids and lipophilic compounds in biological tissues[113-117]. The
biological samples analysed in paper II (i.e. blue mussel and liver from
long-tailed duck) and paper III (i.e. herring) were extracted using the liquid-
liquid extraction (LLE) method described by Jensen et al.[118], with the
minor modification that normal-hexane (n-hx) was replaced with iso-hexane
(iso-hx) as solvent. This method is based on the former method by Jensen et
al[115] which were modified to make the extraction faster and simpler by
removing the filtrating step. The solvent ratio between n-hx and diethyl ether
(DEE) was also changed from 9:1 (v/v) to 3:1 (v/v) in the new method. The
method by Jensen et al[118] has been shown to extract lipids more efficient-
ly than other methods (i.e. Bligh and Dyer[113] and Smedes[116]) and to
give satisfactory recovery of phenols (e.g. simple chlorinated/brominated
phenols, OH-PCBs and OH-PBDEs) in herring and blue mussel[118].
The extraction procedure is described in detail paper II and paper III. In
short, the sample is denaturated with 2-propanol and extracted after addition
of iso-hx:DEE (3:1, v/v) under natural pH. The extraction is repeated once
more and the combined extracts is thereafter washed with a weak aqueous
solution of hydrochloric acid (0.2 M) to remove 2-propanol and any un-
wanted water soluble co-extracts from the organic phase.
I paper IV the applicability of a LLE method developed by Hovander et al.
for analysis of HPCs (i.e. OH-PCBs) in human plasma[119], was evaluated
for OH-PBDE analysis in blood matrices (serum/plasma) from mammal
(human, cat), bird (long-tailed duck) and fish (herring). The method applied
in paper IV is the same as described by Hovander et al. [119], with the mi-
nor modification that n-hx was replaced with cyclo-hexane (c-hx) as solvent.
Briefly described, the sample (serum/plasma) is denaturated with hydrochlo-
ric acid (6M) and 2-propanol and extracted after addition of c-hx:MTBE
(1:1, v/v) under acidic pH. The extraction is repeated once more with
c-hx:MTBE (1:1, v/v) and the combined extract is then washed with an
aqueous solution to remove 2-propanol and water soluble co-extracts.
4.3.2 Separation of neutral and phenolic compounds
Neutral and phenolic compounds in paper II-IV were separated into two
fractions by potassium hydroxide partitioning. The phenolic compounds are
deprotonated by the potassium hydroxide and partition into the alkaline
phase whereas the neutral lipophilic compounds remain in the organic phase.
The phenolic compounds can then be retrieved from the alkaline phase by
acidify the water phase and re-extract it with organic solvent.
30
Sample matrix (e.g. free fatty acids) can interfere with the potassium hydrox-
ide partitioning and result in low recovery of OH-PBDEs as shown in shark
(liver)[120] and herring (unpublished data). To avoid matrix effects gel
permeation chromatography (GPC), a non-destructive clean-up method, was
used in paper II to remove most of the lipids in the long-tailed duck liver
samples prior to the potassium partitioning step. This clean-up step was in-
troduced due to the high lipid content (mean: 5.8%) in the liver samples.
In paper IV, the herring samples were divided prior to the potassium parti-
tioning step to reduce the lipid content in each fraction to avoid matrix
effects.
4.3.3 Derivatization
The phenolic compounds analysed in paper II-IV were methylated prior to
instrumental analysis to enable analysis using gas chromatography- mass
spectrometry (GC-MS). The derivatization was performed using diazome-
thane, synthesized in house from N-methyl-N-nitroso-p-toluene-sulfonamide
according to the method described by Vogel[121]. The methoxy-derivatives
formed after derivatization with diazomethane are stable i.e. can be stored
and tolerate destructive clean-up methods such as sulfuric acid treatment.
However, diazomethane is carcinogenic and explosive and must thus be
handled with great care. The laboratory work with diazomethane was ap-
proved by the Swedish work environment authority.
Acetylating and silylating agents can also be used as derivatizating agents.
However, their main disadvantage is that their products generally are less
stable than methoxy-derivatives[122]. However, pentafluorobenzolyl chlo-
ride (PFBCl) has successfully been used to acetylate halogenated phenolic
compounds (HPCs) in biological samples with help of tetrabutylammonium
as a phase transfer catalyst, yielding products (PFB acetyl derivatives) able
to withstand concentrated sulfuric acid treatment[118].
4.3.4 Lipid removal
Both non-destructive and destructive clean-up methods were used to clean
the samples from lipids which would otherwise obstruct the instrumental
analysis.
Partitioning with concentrated sulfuric acid is an efficient but destructive
method that was used in paper II, III and IV. However care should be taken
when applying this method since some compounds are degraded by concen-
trated sulfuric acid. For the phenolic compounds to withstand the sulfuric
acid treatment, the phenolic compounds were methylated.
31
The mussel and fish samples in paper II, III and IV were cleaned up on
silica gel columns impregnated with sulfuric acid to remove the last remain-
ing lipids in the samples.
Gel permeation chromatography (GPC), a non-destructive clean-up method,
was used in paper II to remove most of the lipids in the long-tailed duck
liver samples prior to the potassium partitioning step. The stationary phase
used consisted of small neutral porous beads of styrene divenylbenzene
(with 3% cross linkage). Large compounds with a molecular size larger than
the exclusion limit of the stationary phase (i.e. ≥2000 daltons) will pass the
beads unhindered, whereas small compounds diffuse into the pores of the
beads where they are retained. Hence, small compounds will take longer
time to pass through the GPC column than larger compounds. Further, the
choice of mobile phase influence the retention time of the analytes as the
solvent will affect swelling of the beads and the adsorption of the analytes to
the stationary phase[123]. In paper II, a mixture composed of iso-
hx:dichloromethane (1:1 v/v) with 0.5% formic acid (v/v) was used as mo-
bile phase. The formic acid was used to facilitate elution of OH-PBDEs, by
increasing the polarity of the mobile phase and more importantly ensure that
the OH-PBDEs remained protonated during clean-up.
4.3.5 Instrumental analysis
The brominated substances discussed in this thesis were analyzed using gas
chromatography-low resolution mass spectrometry (GC-LRMS) operated in
electron capture negative ionization (ECNI) mode. The analyses were per-
formed using single ion monitoring (SIM), scanning the bromide ions,
m/z 79 and 81. This method gives a limit of quantification (LOQ) in the
same order of magnitude as gas chromatography-high resolution mass spec-
trometry (GC-HRMS) operated in electron ionization (EI) mode for the
brominated compounds analysed, but provides no structural information.
Hence identification is based on the retention times of the authentic refer-
ence standards employed. The authentic reference standards used were either
purchased or synthesized in house.
Lack of available reference standards are often a limiting factor for identifi-
cation of new compounds, as true identification requires e.g. matching reten-
tion times and full scan spectra of the unknown compound and the reference
compound. The OH-PBDEs and MeO-PBDEs quantified in this thesis have
previously been identified in Baltic biota (i.e. red algae, blue mussel and
salmon) by comparing relative retention times (on both polar and nonpolar
columns) and ECNI and EI full scan spectra, against synthesized reference
compounds[31,124].
32
Liquid chromatography-mass spectrometric techniques (LC-MS) using elec-
trospray (ESI) and atmospheric pressure photo ionization (APPI) have been
used by several others for OH-PBDEs analysis[125-129]. However, a slight-
ly better chromatographic separation and probably higher LOQ using GC
compared to LC have resulted in preference of the former technique[130].
4.3.6 Quality assurance/Quality control
Measures were taken to limit sample contamination during extraction and
clean-up. Hence, all solvent, acids and salt used in the sample work up in
paper II-IV were of highest purity available. All glass equipment were also
washed and heated at 300 ᵒC prior to use to avoid contamination. To meas-
ure any potential background contamination during the analysis, procedural
solvent blanks were prepared in parallel with each batch of samples. Blank
contamination was mainly a problem with the long-tailed duck liver samples
(paper II) were small amounts of PBDEs, MeO-PBDEs and OH-PBDEs
were found in the procedural solvent blanks. Likely these residues originate
from sample cross contamination in the GPC injection vault.
The limit of detection (LOD) was set to three times the background noise
(signal to noise ratio = 3). The limit of quantification (LOQ) was set as three
times the LOD or five times the mean concentration in procedural solvent
blanks. All samples were blank subtracted when blank contamination was
present. The efficiency of the extraction and clean-up method was assessed
by the use of surrogate standards. The surrogate standard should share struc-
tural and chemical properties as the analytes studied (to mimic their behavior
in the sample work up) and the compound of choice must not be present as
an environmental contaminant in the sample.
Very low recoveries of the surrogate standard (4’-OH-BDE121) were unex-
pectedly observed when OH-PBDEs were analysed in herring plasma (un-
published data). The method applied for the analyses has been developed
by Hovander et al.[119], and is well established. The low recovery of 4’-
OH-BDE121 pointed out the need to evaluate this method for analysis of
OH-PBDEs in fish plasma. The method was evaluated, as described in pa-
per IV using nine OH-PBDEs, three simple bromophenols and three MeO-
PBDEs as test compounds. The recoveries of the analytes were tested in
plasma from herring, long-tailed duck, humans as well as in cat serum. The
results are further discussed in chapter 4.4
33
4.4 An analytical challenge
The results of the recovery study and the set up are presented in paper IV
and indicated satisfactory recoveries of MeO-PBDEs across all species. High
and reproducible recoveries were also obtained for BPs and OH-PBDEs in
human plasma and cat serum. On the contrary, poor recoveries and large
variations were obtained for plasma samples from herring and long-tailed
duck. This initiated an extensive and tedious work to trace the reason(s) for
this significant obstacle in the analysis of such samples
4.4.1 Further method evaluation (additional results)
The observation of poor OH-PBDEs recoveries, called for some in-depth
method evaluation using herring plasma. A structured evaluation scheme
was set up using nine OH-PBDEs (5 ng/compound) which were added to
herring plasma (0.5 g) at different steps in the sample work up as shown in
Figure 10. As control OH-PBDEs were spiked to solvent which were ex-
tracted and clean-up along with the serum samples. The tests were per-
formed in duplicate and the recoveries determined in relation to three ref-
erences samples (solvent spiked with the same amount of OH-PBDEs as
each sample). The references phenols were methylated with diazomethane
and partitioned with sulfuric acid to remove impurities from the derivatiza-
tion agent, but were otherwise handled as little as possible. A volumetric
standard (BDE-138, 5 ng) was added to references and samples prior to in-
strumental analysis by GC/MS (ECNI). The recoveries of the references
were set to 100%. The recovery of each OH-PBDE congener added to serum
and solvent at different steps of the work up is presented in Figure 10. The
samples spiked prior to denaturation (B) showed the poorest recoveries but
also samples spiked prior to derivatization (D) gave very low recoveries for
some OH-PBDEs (i.e. 2’-OH-BDE68, 4’-OH-BDE49 and 4’-OH-BDE121).
While other OH-PBDEs were recovered in fairly good yields. The low re-
coveries of 2’-OH-BDE68, 4’-OH-BDE49 and 4’-OH-BDE121 are indicated
to be matrix dependent since the spiked, extracted and cleaned-up solvent
samples (A) had satisfactory recoveries.
34
Figure 10 Schematic presentation of the recovery experiment. The recovery of each OH-PBDE at the end
of the sample work up after addition at different stages of the sample work up is presented.
Differences in recovery between sample B and C (Figure 10) were tested for
pH dependency i.e. three extractions of herring plasma were performed un-
der acidic, neutral and alkaline conditions. Briefly, three herring plasma
samples (0.5 ng/sample) were spiked with 4’-OH-BDE121 (5ng) and 2-
propanol (2 mL) and c-hx:DEE (3:1, v/v) was added. One sample was acidi-
fied by adding hydrochloric acid (6M, 0.3mL) whereas trimethylamine
(99%, 0.1 mL) was added to another sample to increase the pH. The three
samples were vortexed and after centrifugation the liquid phase was trans-
ferred to new test tubes. The remaining pellet was re-extracted with 2-
propanol (2 mL) and c-hx:DEE (3:1, v/v, 3 mL). The pooled liquid phase
which contained triethylamine was acidified by adding a few droplets of
hydrochloric acid (6M) before all samples were partitioned with hydrochlo-
ric acid (0.2 M, 5 ml) and further treated with potassium hydroxide partition-
ing, derivatization (diazomethane) and sulfuric acid partitioning. A vol-
umetric standard (BDE-138) was added prior to instrumental analysis and
the recovery was determined against a reference, prepared as previously
described. The recovery of 4’-OH-BDE121 was found to be poorest when
extracted under acidic conditions (3%), whereas neutral (25%) and alkaline
extraction (45%) gave slightly better yields.
Extractionc-hx:MTBE (1:1 v/v)
Gravimetric lipid determination
SeparationKOH partitioning
PhenolsNeutrals
Instrumental analysisGC-LRMS(ECNI)
Lipid removalH2SO4 treatment
andH2SO4:SiO2 column
DerivatizationDiazomethane
Denaturation2-propanol, HCl (6M)
x 2 x 2A B C D E
OH-PBDEs added to herring plasma
OH-PBDEs added to solvent
A
B
C
D
E
Recovery (%)
2’-OH-BDE28
2’-OH-BDE68
6-OH-BDE85
2-OH-BDE123
6-OH-BDE137
3-OH-BDE47
5-OH-BDE47
4’-OH-BDE49
4’-OH-BDE121
76 71 83 83 74 83 80 84 81
7 0 47 47 0 12 16 0 0
51 7 72 72 16 81 79 24 0
40 9 78 78 22 63 68 30 0
66 7 95 95 72 68 81 23 1
35
The poor recovery of OH-PBDEs in sample E (Figure 10) was hypothesized
to be due to co-extractives which obstructed the derivatization of some of the
OH-PBDEs.
Since both the herring and the long-tailed duck plasma samples were red
stained by haemolysis, potential effects of hemin/hematin on the derivatiza-
tion efficiency was further investigated by a colleague (Dennis Lindqvist,
Personal communication). In this experiment, hemin (0.2 mg) was added to
aqueous sodium chloride solution (2.5 mL), spiked with 6-OH-BDE47 and
4’-OH-BDE121 and the sample extracted according to Hovander et al.[119].
After potassium hydroxide partitioning to isolate the phenolic fraction, the
sample was divided in two subsamples and the phenols were methylated
with either diazomethane or pentafluorobenzolyl chloride (PFBCl). The test
was performed in duplicates and the recoveries of 6-OH-BDE47 and 4’-OH-
BDE121 after derivatization with diazomethane were 56±2% and 19±1%,
respectively, whereas PFBCl yielded good recoveries of both 6-OH-BDE47
(95±1%) and 4’-OH-BDE121 (96±4%). The result indicate that co-extracted
hemin/hematin in blood samples can affect diazomethane methylation effi-
ciency of OH-PBDEs. However, the use of PFBCl as derivatisation agent in
biological samples is problematic since the efficiency of this derivatisation
agent is known to be sensitive to other potential co-extractables (i.e. free
fatty acids), as discussed elsewhere[118].
These findings emphasises the importance of proper method evaluations,
especially when working with complex biological matrices. Most important-
ly, care needs to be taken when analysing OH-PBDEs in plasma or serum.
Samples with haemolysis should be avoided as it can have profound effects
on the result as shown here and in Paper IV. Since blood is a favorable ma-
trix to work with for analysing OH-PBDEs, due to their specific retention to
blood proteins (e.g. transthyretin). The need to develop a reliable extraction-,
derivatization- and clean-up method for OH-PBDEs which are insensitive to
hemolysis and preferable also could be applied to whole blood is most need-
ed.
36
5 Results and Discussion
This chapter is aimed to highlight and discuss major findings of paper I-III.
5.1 OH-PBDEs potency to disrupt OXPHOS in vitro
In paper I, the OXPHOS disrupting potential of 18 OH-PBDEs were deter-
mined in vitro. The compound studied included both OH-PBDEs of natural
and anthropogenic (metabolic) origin, and comprised compounds previously
reported in biota from the Baltic Sea (table S1, paper I).
5.1.1 TPP data on OH-PBDEs
In the TPP assay, all OH-PBDEs tested were found to disrupt OXPHOS
either by protonophoric uncoupling and/or via inhibition of the ETC at micro
Molar (µM) concentrations in isolated rat liver mitochondria (Table 1). The
lowest observed effect concentration (LOEC) was found to differ with a
factor of 100 within the group. Three compounds (i.e. 3-OH-BDE47, 6-OH-
BDE47, 6-OH-BDE85) were found to be as potent (LOEC=0.01 µM) as the
model uncouplers; SF6847 (LOEC= 0.02µM), FCCP (LOEC=0.05µM) and
PCP (LOEC=0.05µM), used as positive controls. Nearly all compounds
tested were found to act via both uncoupling and inhibition within the con-
centration range tested (including the positive controls; FCCP, SF6847, PCP
and DNP). Only 6-OH-BDE47 and 3’-OH-BDE154 did solely act as inhibi-
tors whereas 5-OH-BDE47 and 6-OH-BDE90 did only act via protonophoric
uncoupling.
37
Table 1. OH-PBDEs potency to uncouple and/or inhibit oxidative phosphorylation in the TPP- and
TMRM assay. The compounds occurrence in Baltic biota is indicated. Table adopted from paper I.
TPP assay TMRM assay
uncoupling inhibition altered membrane
potential
compound LOEC (µM) LOEC (µM) LOEC (µM) EC50 (µM) Baltic biota
2’-OH-BDE28 3.8 16 n.d. >40
2’-OH-BDE66 1.0 10 50 39
2’-OH-BDE68 0.5 5.0 n.d. >100 A, C, B, H, S, G, R
2’-OH-6’-Cl-BDE68 1.0 3.8 12 19 B, S
2-OH-BDE123 0.7 3.8 12 15 A, C, B, H
3-OH-BDE47 0.01 2.0 n.d. >100 R
3-OH-BDE153 2.5 8.8 50 56
3’-OH-BDE154 n.e. 1.3 20 28
3-OH-BDE155 3.2 32 n.d. >100
4’-OH-BDE17 4.4 34 75 60
5-OH-BDE47 0.1 n.e. n.d. 100
6-OH-BDE47 n.e. 0.02 6.0 6.0 A, C, B, H, S, L, G, R
6-OH-5-Cl-BDE47 0.4 3.8 16 21 S
6’-OH-BDE49 0.8 4.0 50 41 S
6-OH-BDE85 0.01 1.8 8.0 5.6 A, C, B
6-OH-BDE90 1.3 n.e. 20 17 A, C, B, H, S, R
6-OH-BDE99 0.5 0.7 14 27 A, C, B, H, S, L
6-OH-BDE137 1.0 2.5 10 21 A, C, B, H
FCCP 0.05 1.0 0.75 1.0
SF6847 0.02 0.38 2.5 2.1
PCP 0.05 2.0 30 33
DNP 1.0 80 n.d. >90
BDE-47 n.e. n.e. n.d. >100
no effect (n.e.), not detected (n.d.), cyanobacteria (C), algae (A), blue mussel (B), herring (H), salmon (S),
long-tailed duck (L), common guillemot (G), ringed seal (R). For references please see table S1, paper I
and paper II and III
5.1.2 TMRM data on OH-PBDEs
In the TMRM assay, 13 out of 18 OH-PBDEs were found to disrupt
OXPHOS in zebrafish embryonic fibroblast cells (Table 1). 6-OH-BDE47
(LOEC=6.0µM) and 6-OH-BDE85 (LOEC=8.0µM) were the two most
potent OH-PBDEs tested with effect concentrations in the same magnitude
as the model uncouplers, FCCP (LOEC=0.75µM) and SF6847 (LOEC=2.5
µM). Five compounds (i.e. DNP, 2’-OH-BDE28, 2’-OH-BDE68, 3-OH-
BDE47 3-OH-BDE155 and 5-OH-BDE47) which showed effect in the TPP
assay did not disrupt OXPHOS in the TMRM assay. The negative response
might be due to insufficient uptake by the cell/mitochondria, interaction with
38
the cell media or possible that these compounds are unable to disturb
OXPHOS in zebrafish embryonic fibroblast cells, at least within the concen-
tration range tested.
For the other compounds, a positive correlation (Pearson’s product moment
correlation, r=0.68, P=0.004) between the TPP and TMRM LOEC values
was obtained which indicate that the TMRM assay well predicts the effects
of OXPHOS disruption. Since we were not able to pin point why some com-
pounds showed a negative response in the TMRM assay compared to the
TPP assay, the use of both assays is recommended also in the future when
assessing chemicals potential for OXPHOS disruption.
5.1.3 Cytotoxicity
None of the compounds tested showed sign of cytotoxicity in the LDH as-
say. In the MTT assay, a decrease in formazan production was observed for
some compounds at 50µM whereas other showed no effect or an increased
formazan production compared to control (indicative of increased mitochon-
drial metabolism). However, altogether the results show that the observed
disruption in mitochondrial membrane potential in the TMRM assay is likely
not caused by confounding cytotoxicity.
5.1.4 Synergistic effects
When the OH-PBDEs were assessed as mixtures in the TMRM assay, syner-
gistic effects were observed. In toxicology and ecotoxicology, synergistic
effects are rarely observed and if they occur they are usually very specific
for the chemical mixture tested and end-point studied[99]. For the four com-
pound mixture (composed of 6-OH-BDE47, 6-OH-BDE99, 6-OH-5-Cl-
BDE47, 2’-OH-6’-Cl-BDE68) the obtained TUs were 0.55 TU (EC50) and
0.34 TU (EC10), hence showing strong synergism (Figure 11a). However,
when the potent inhibitor 6-OH-BDE47 was removed from the mixture, the
three compound mixture (combined at equitoxic concentrations) showed
additive effects (0.96 TU) when derived from EC50 values, and synergism
(0.22 TU) was only observed at lower concentrations (EC10 values) as shown
in Figure 11b. The four compound mixture was also assessed using the IA
model, assuming that OH-PBDEs act independently, because OH-PBDEs
were found to act both as protonophoric uncouplers and inhibitors of the
ETC as shown in the TPP assay. The IA model predicted a lower effect than
observed, again indicating synergism.
39
Figure 11 Concentration response curves in the TMRM assay for a four compound mixture (a) and a
three compound mixture (b). The concentrations in graph a and b are expressed as toxic units (TU)
calculated based on EC50 and EC10 concentrations derived from the individual OH-OPBDEs dose-
response curves depicted in graph c. Figures adopted from paper I.
Valmas et al. (2008) have previously reported strong synergism when com-
bining protonophoric uncouplers (FCCP and PCP) and inhibitors of the ETC
(phosphine, azide)[131]. The synergistic effect was suggested to be caused
by simultaneous inhibition of the ETC and depletion of the proton gradient
by uncoupling, causing a rapid collapse of the mitochondrial membrane po-
tential[131]. As the OH-PBDEs tested in paper I were found to act via pro-
tonophoric uncoupling and/or inhibition, this combined effect on the mito-
chondrial membrane potential, could likely explain the synergism observed.
This hypothesis is also strengthened by the observation in paper I that the
mixture toxicity decreased when the potent inhibitor 6-OH-BDE47 was re-
moved from the mixture.
0.1 1 100
50
100
150
EC50= 0.55 (0.45-0.65) TU
EC50
EC10
EC10= 0.34 (0.25-0.43) TU
a)
TU
Eff
ect (%
)
0.1 1 100
50
100
150
EC50= 0.96 (0.57-1.35) TU
EC50
EC10
EC10= 0.22 (0.10-0.53) TU
b)
TU
Eff
ect (%
)
1 10 1000
20
40
60
80
100
6-OH-BDE47
6-OH-BDE99
6-OH-5-Cl-BDE47
2'-OH-6'-Cl-BDE68
c)
Eff
ect (%
)
Concentration (µM)
40
Strong synergism (0.22 TU, EC50) was also observed when the seven com-
pound mixture, mimicking the internal exposure in Baltic blue mussels, was
studied. The synergism was even stronger than for the four compound mix-
ture (0.55 TU, EC50), indicating that the synergistic effects increase with
increasing number of compounds. Further, the mixture were found to disrupt
the mitochondrial membrane potential at dose levels only 100 to 1000 times
higher than reported in Blue mussels from the Baltic Sea (Σ7OH-PBDEs: 50
ng/g f.w.) (Figure 12).
Figure 12. Dose response curve of a seven compound mixture added to mimic the exposure of OH-PBDEs
in Baltic blue mussel. The concentration factor of one represent the environmental concentration in blue
mussels and a factor of 100 a concentration 100 times more concentrated. Figure adopted from paper I.
5.2 Blue mussels
The blue mussel is a key species in the Baltic Sea ecosystem[103] and the
main food source for long-tailed ducks while wintering in the Baltic
Sea[106]. The concentrations of brominated substances (i.e. BPs, BAs, OH-
PBDEs, MeO-PBDEs and PBDEs) were assessed in blue mussels, collected
in March and May from nine sites in the Baltic Proper used by long-tailed
ducks for foraging (Paper II). The geometric mean (GM) concentrations of
ΣOH-PBDEs, ΣMeO-PBDEs and ΣPBDEs are presented in Figure 13.
On average the Blue mussels collected in spring were found to contain ana-
lytes in the following decreasing order, with geometric means (GM) present-
ed on fresh weight basis in parenthesis: ΣMeO-PBDEs (GM 1.3 ng/g f.w),
ΣBPs (GM 0.87 ng/g f.w.), ΣOH-PBDEs (GM 0.82 ng/g f.w.), ΣBAs (GM
0.20 ng/g f.w.) and ΣPBDEs (GM 0.13 ng/g f.w.). 2,4,6-triBP, 6-OH-BDE47
and 6-OH-BDE85 were the three dominating compounds in the phenolic
fraction, where 6-OH-BDE47 and 6-OH-BDE85 comprised as much as 50%
and 21% of the sum OH-PBDEs, respectively.
0.01 0.1 1 10 100 1000 100000
50
100
150
Concentration factor
Eff
ect (%
)
41
The mean concentration of ΣOH-PBDEs (GM 0.82 ng/g f.w.) in blue mussel
collected during March-May is approximately 60 times lower than concen-
tration measured in the Blue mussels sampled in June (ΣOH-PBDEs: 53 ng/g
f.w.) in paper II. The higher exposure seen in summer is expected since
concentrations of OH-PBDEs in the Baltic Sea vary seasonally with an in-
crease during summer in both algae[69] as well as in blue mussels[79]. Fur-
ther, blue mussels collected in June contained similar OH-PBDE concentra-
tions as previously reported from the same site in June (ΣOH-PBDEs (50
ng/g f.w.)[79]. The results confirm that the high exposure to these com-
pounds in the Baltic blue mussel most likely is an annually reoccurring
summer phenomenon.
It is not yet understood how blue mussels are affected in vivo by long time
exposure to environmental levels of OH-PBDEs, a fact that requires further
research. Exposure studies with the potent protonophoric uncoupler penta-
chlorophenol (PCP) have shown to affect the metabolic rate in blue mus-
sel[132] and fresh water clam (Pisidium amnicum)[133]. In the study with
blue mussels, the lowest water concentration tested (50 µg/L PCP) were
found to enhance the metabolic rate (measured as oxygen uptake and heat
output) by 35% after two days exposure under normal aerobic conditions.
The effect concentration corresponded to a tissue concentration of
2.9 µg/g f.w. in blue mussel[132]. The measured effect concentration is ap-
proximately 50 times higher than the concentration of ΣOH-PBDEs meas-
ured in blue mussels collected in June (Paper II).
Others have also reported PCP to affect the anaerobic metabolism in blue
mussels[134]. During prolonged period of anaerobiosis blue mussels mainly
derive their energy from malate, which originate from phosphoenolpyruvate
derived in the glycolysis. In the mitochondria, part of the malate is oxidized
to pyruvate and acetate which yield NADH whereas another portion of mal-
ate is converted to fumarate and reduced to succinate by fumarate reductase.
The electrons to fuel the reduction of fumarate is derived from the NADH
formed from the oxidation of malate and puryvate which are funnelled via
complex I of ETC and rhodoquinone to the fumarate reductase[135]. The
ratio of fumarate:succinate were found to increase with increasing PCP ex-
posure as reported in a study with blue mussels exposed to PCP under anoxic
conditions[134]. The authors of the study suggest a metabolic block in the
electron transfer system as potentially responsible for the observed ef-
fect[134].
PCP was also found to inhibit ETC at higher concentration in the TPP assay
in paper I. However, inhibition of complex I could not be studied in the
experimental setup used in the TPP assay in paper I, as rotenone was used
to block this complex to enable simultaneous studies of protonophoric un-
42
coupling and inhibition of complex II, III and IV. Hence, further studies are
needed to determine if OH-PBDEs also have potential to affect the
anaerobic metabolism in blue mussels.
As OH-PBDEs showed to disrupt OXPHOS in paper I the effect of OH-
PBDEs on aerobic metabolism in blue mussels can potentially affect the
condition of the blue mussel, and their nutritional value for mussel feeders
such as sea ducks.
5.3 Long-tailed duck
Long-tailed ducks need to digest large quantities of mussels each day to
meet their daily energy requirement. Their estimated daily intake of blue
mussel meat was assessed in paper II, based on the assumption that winter-
ing long-tailed ducks have a daily energy expenditure of 1500kJ/day[105].
Long-tailed ducks average daily consumption was estimated to be approxi-
mately 450 g mussel meat (calculations given in paper II). Based on the
mean fresh weigh concentrations of brominated substances in blue mussels
collected in March and May, this correspond to a daily intake of 390 ng
ΣBPs, 90ng ΣBAs, 370ng ΣOH-PBDEs, 590ng ΣMeO-PBDEs and 59ng
ΣPBDEs for the long-tailed ducks. Although this is an estimate, it provides a
first assessment of long-tailed duck dietary intake of these substances while
foraging in the Baltic Sea.
Dietary exposure to protonophoric uncouplers (i.e. PCP) has been reported
to significantly decrease the body weight and reduce the lipid content in
mallard ducklings (Anas platyrhynchos)[136]. The lowest observed adverse
effect level (LOAEL) after 11 day exposure was found for birds feed a diet
with a concentration of 961 µg/g PCP (which corresponded to a liver tissue
concentration of 31µg/g).
Concentration of brominated substances in livers of long-tailed duck winter-
ing in the Baltic Sea is presented in paper II. The geometric mean (GM)
concentrations of ΣOH-PBDEs, ΣMeO-PBDEs and ΣPBDEs are shown in
Figure 13. The liver concentration of ΣOH-PBDEs (GM 0.34 ng/g f.w.) is
approximately 5 order of magnitude below LOAEL for PCP (31µg/g f.w.)
reported to cause a 50% reduction in lipid content in Mallard ducklings after
11 day exposure[136]. How adult long-tailed ducks are affected by long term
exposure to low doses of OH-PBDEs while foraging in the Baltic Sea (from
October to mid-May) is not known. As mussel feeding sea ducks have poten-
tial to be affected by compounds disturbing OXPHOS directly and/or indi-
rectly via reduced food quality of their feed, further studies on the in vivo
effect of OH-PBDE exposure in sea ducks and mussels are needed.
43
OH-PBDEs
Blue mussel Long-tailed duck0
50
100
150
ng
/g lip
id w
eig
ht
MeO-PBDEs
Blue mussel Long-tailed duck0
50
100
150
ng
/g lip
id w
eig
ht
PBDE
Blue mussel Long-tailed duck0
5
10
15
20
ng
/g lip
id w
eig
ht
OH-PBDEs
Blue mussel Long-tailed duck0.0
0.5
1.0
1.5
ng
/g f
resh
we
igh
t
MeO-PBDEs
Blue mussel Long-tailed duck0
1
2
3
4
5
ng
/g f
resh
we
igh
t
PBDEs
Blue mussel Long-tailed duck0.0
0.2
0.4
0.6
0.8
ng
/g f
resh
we
igh
t
a)
b)
Figure 13. Geometric mean concentrations and range (min-max) of ΣOH-PBDEs, ΣMeO-PBDEs and
ΣPBDEs in Baltic blue mussels collected in March and May (n=11) and in livers of long-tailed duck
(n=10) presented on a) lipid weight basis and b) fresh weight basis. Figure adopted from Paper II.
The significantly lower concentrations of ΣBPs, ΣBAs, ΣOH-PBDEs and
ΣMeO-PBDEs in long-tailed duck liver tissue compared to blue mussel
(Figure 13) despite a constant daily intake, suggest that these compounds are
rapidly excreted in long-tailed duck liver tissue whereas PBDEs seem to be
retained.
The PBDE congener profile in long-tailed duck were found to be similar as
previously reported in liver tissue from common eider (Somateria mollissi-
ma) and white-winged scoter (Melanitta deglandi) from the Canadian arc-
tic[137] (Figure 14). The result indicate that the PBDE congener profile in
benthic feeding sea ducks differ from what is seen in sea birds[138], fish and
fish feeding mammals[46], where BDE-47 usually dominate.
44
Figure 14. PBDE congener profile in blue mussels, sea ducks and herring compared to a technical
PentaBDE mixture (Bromkal 70-5DE).
5.4 Baltic herring
Baltic herring is one of the dominant pelagic zooplanktivorous species in the
Baltic Sea. Concentrations of brominated substances (i.e. BPs, BAs, OH-
PBDEs, MeO-PBDEs and PBDEs) were assessed in Baltic herring as de-
scribed in paper III. The herring, collected in May and June, were from the
northern Baltic Proper (Askö) and the southern Baltic Sea (Ängskärsklubb).
The geometric mean (GM) concentrations of ΣOH-PBDEs, ΣMeO-PBDEs
and ΣPBDEs are presented in Figure 15 showing no significant differences
in ΣOH-PBDEs levels between Askö (GM 9.4 ng/g l.w.) and Ängskärsklubb
(GM 10 ng/g l.w.). 6-OH-BDE47 dominated the OH-PBDE congener profile
in herring from both locations, where it comprised 87% (Askö) and 91%
(Ängskärsklubb) of the sum OH-PBDEs.
6-OH-BDE47 was one of the toxicologically most potent OH-PBDEs tested
in paper II. This compound has also been found to be acutely toxic to adult
and developing zebrafish[8,9]. In adult zebrafish lethality was observed after
12 min when exposed to a water concentration of 1µM 6-OH-BDE47 (corre-
sponding to a liver concentration of 3.5 µg/g f.w.)[9]. The mean (min-max)
concentration of 6-OH-BDE47 in herring liver from the northern Baltic
Proper (Landsort) sampled in autumn during 1980-2010 were 2.0 (0.89-3.8)
ng/g f.w.[81].
0
10
20
30
40
50
60
70
80
90
100
Bromkal (70-5DE) Arctic blue mussel(M. edulis)
Baltic blue mussel(M. trossulus x
edulis)
White-winged scoter(M. deglandi)
Common eider(S. Mollissima)
Long-tailed duck(C. hyemalis)
Baltic herring(C. harengus)
% c
on
gen
er c
on
trib
uti
on
BDE-183
BDE-153
BDE-154
BDE-99
BDE-100
BDE-47
BDE-28
La Guardia et al. (2006) Kelly et al. (2008) Kelly et al. (2008) Kelly et al. (2008)Paper II Paper II Paper III
45
OH-PBDEs have shown to be taken up in fish after both water and dietary
exposure[72,139]. 6-OH-BDE47 has also been reported to be formed as a
metabolite of BDE-47 in Northern pike (Esox lucius)[55] but little or no
transformation has been seen in zebrafish (Danio rerio)[9,72] or japanese
medaka (Oryzias latipes)[71]. As the metabolic capacity of herring is not
fully understood, the metabolic contribution to the here reported 6-OH-
BDE47 concentration cannot be excluded, although natural production is
most likely the major source in the Baltic Sea. However, metabolic inter-
conversion of 6-MeO-BDE47 and 6-OH-BDE47 has been observed in
zebrafish[72] and japanese medaka[71], which indicate that metabolic de-
methylation of MeO-PBDEs can play a role as source for OH-PBDEs in fish.
OH-PBDE
Askö Ängskärsklubb0
5
10
15
ng
/g lip
id w
eig
ht
MeO-PBDE
Askö Ängskärsklubb0
50
100
150
200
250
ng
/g lip
id w
eig
ht
PBDEs
Askö Ängskärsklubb0
20
40
60
80
ng
/g lip
id w
eig
ht
OH-PBDE
Askö Ängskärsklubb0.0
0.2
0.4
0.6
0.8
1.0
ng
/g f
resh
we
igh
t
MeO-PBDE
Askö Ängskärsklubb0
5
10
15
20
ng
/g f
resh
we
igh
t
PBDE
Askö Ängskärsklubb0
1
2
3
4
ng
/g f
resh
we
igh
t
a)
b)
Figure 15 Geometric mean concentration (95% confidence intervals) of ΣOH-PBDEs, ΣMeO-PBDEs and
ΣPBDEs in Baltic herring from Askö (n=12) and Ängskärsklubb (n=12) presented on a) lipid weight
basis and b) fresh weight basis. Figure adopted from Paper III.
At Ängskärsklubb, the concentration of ΣMeO-PBDEs (GM 150 ng/g l.w.)
was 15 times higher than the concentration of ΣOH-PBDEs, whereas the
concentration of ΣMeO-PBDEs (GM 42 ng/g l.w.) at Askö was lower.
6-MeO-BDE47 dominated the MeO-PBDE congener profile contributing to
64% (Askö) and 77% (Ängskärsklubb) of the sum MeO-PBDEs. The major
source for MeO-PBDEs, as reported herein, is most likely natural produc-
tion. Metabolic conversion of PBDE to OH-PBDE followed by in vivo O-
methylation in herring is a less likely source, as neither OH-PBDEs nor
MeO-PBDEs were found as transformation products of BDE-47 in either
japanese medaka or zebrafish[71,72]. The high ratio between 6-MeO-
46
BDE47 (GM 116 ng/g l.w.) and BDE-47 (GM 10 ng/g l.w.) in herring from
Ängskärsklubb further supports that the MeO-PBDEs originate from natural
sources rather than from PBDEs.
Substantial maternal transfer of 6-OH-BDE47 and 6-MeO-BDE47 to the roe
has also been observed in zebrafish. For 6-OH-BDE47 the ratio between egg
and liver concentrations were 0.8, and an even higher ratio (2.9) was ob-
served for 6-MeO-BDE47[72]. In japanese medaka have similar ratio been
observed for 6-OH-BDE47 (0.59) while the ratio for 6-MeO-BDE47 was
lower (0.74)[71]. Maternal transfer of OH-PBDEs and MeO-PBDEs (as
precursor for OH-PBDEs) may be of concern, due to OH-PBDEs high tox-
icity to the developing fish embryo[8,9]. 6-OH-BDE47 has shown to cause a
wide range of developmental effects such as pericardial edema, yolk sac
deformations and developmental arrest in zebrafish embryo when exposed to
low water concentrations (EC50=25 nM)[9]. In Baltic herring, high mortality
of herring roe has been observed both in field[33] and in controlled experi-
ments when herring egg were exposed to algae[140], now known to contain
OH-PBDEs[69]. The low effect concentration for 6-OH-BDE47 and the
higher production of OH-PBDEs in algae during the summer months, may
explain the observed herring egg mortality.
Furthermore, as one of the consequences of disturbed OXPHOS is altered
metabolism, potentially leading to weight loss [10,93]. More research is
needed to determine if the decreasing trends in condition and fat content
observed in Baltic herring [17] is associated with their exposure to chemicals
with OXPHOS disrupting properties, such as OH-PBDEs.
47
6 Concluding remarks and future perspective
The thesis main results are that naturally produced OH-PBDEs act as potent
disruptors of OXPHOS in vitro and that these compounds act synergistically.
However, further investigations on the in vivo effects at the environmental
levels determined in Baltic biota (algae, cyanobacteria and in wildlife) are
needed. Especially the effect of chronic low dose exposure to OH-PBDEs
needs to be assessed. This should preferentially be done in controlled exper-
iments and with field studies. Efforts should also be made to investigate
potential biomarkers to study OXPHOS in vivo. Biological factors reported
to be significantly altered in fish exposed to PCP are for example; body
weight, plasma inorganic phosphate, blood glucose, lactate and plasma pro-
tein[141].
Given the high toxicity of OH-PBDEs in zebrafish, and the high exposure
seen in Baltic blue mussels during summer, future studies of in vivo effects
in fish and mussels are warranted. The long-tailed duck are constantly ex-
posed to OH-PBDEs via their mussel diet while wintering in the Baltic Sea,
but the low levels observed in long-tailed duck (liver) compared to their
food, indicate that OH-PBDEs and MeO-PBDEs have low retention in
ducks. However, more knowledge regarding OH-PBDEs toxico-kinetics is
needed. If blue mussels exposure to OH-PBDEs affect their nutritional value
as food source for sea ducks also needs to be further assessed. Since we see a
clear seasonal variation of OH-PBDEs and MeO-PBDEs in algae and blue
mussels, with higher levels in summer[69,79], studies of in vivo effects in
mussels should be carried out during summer. Further, studies with mussel
feeding sea ducks should preferentially be focused on species feeding on
blue mussels during the exposure peak of OH-PBDEs and MeO-PBDEs (e.g.
the Baltic breeding common eider).
Further, OH-PBDEs also have potential to disturb the thyroid hormone sys-
tem, which can affect the metabolism. Hypothyroidism (i.e. elevated thyroid
homone levels) stimulates a hypermetabolic state with increased energy ex-
penditure and weight loss[142]. Thiamine deficiency has also been reported
in both fish (i.e. Baltic salmon)[29] and waterfowl (e.g. Common eider) from
the Baltic Sea[27]. Thiamine (Vitamin B1) is in its phosphorylated form an
essential cofactor for several enzymes involved in catabolism and in exten-
sion in cellular energy production. For example is thiamine crucial for nor-
48
mal function of α-ketoglutarate dehydrogenase, an enzyme involved in the
citric acid cycle[143]. As thiamine deficiency, altered thyroid homoestasis
and disturbed OXHOS have potential to affect catabolism and thereby ulti-
mately the energy (ATP) production and ATP expenditure in the organism.
Studies of combined effects of low thiamine levels and OH-PBDE exposure
should be prioritised in the future.
The effect of eutrophication in the Baltic Sea can have affected the produc-
tion of OH-PBDEs as cyanobacteria and macroalgae species are primary
producers of these compounds. The temporal trend of 6-OH-BDE47 and
2’-OH-BDE68 in herring liver tissue, indicate that the exposure has in-
creased from 1980 to 2010 in the southern and the northern Baltic Prop-
er[81]. However, temporal trends from other species such as blue mussel
would be beneficial to better assess the situation. As the concentration in
algae and blue mussels varies seasonally, care should be taken to ensure that
samples have been collected under similar conditions to avoid drawing erro-
neous conclusions.
In addition to the compounds discussed in this thesis, several yet unidenti-
fied halogenated phenols have been observed in algae, blue mussels[78] and
salmon[30] from the Baltic Sea. Hence, further efforts must be made to iden-
tify these compounds and their potential toxicity should be investigated.
This thesis emphasis the fact that naturally produced compounds, in addition
to anthropogenic substances, have the potential to pose a significant stress on
the already sensitive Baltic Sea ecosystem.
49
7 Sammanfattning (summary in Swedish)
Många antropogena halogenerade organiska ämnen har visats vara skadliga
för miljö och djurliv och är därför idag reglerade på grund av dess toxiska,
persistenta och bioaccumulerande egenskaper. Polybromerade difenyletrar
(PBDEer), i form av tekniska blandningar såsom PentaBDE, OctaBDE och
DecaBDE, är en grupp bromerade organiska ämnen som sedan 1970-talet
använts för att flamskydda bland annat textilier och elektronik men vars
användning idag är föbjuden (PentaBDE och OctaBDE) eller begränsad
(DecaBDE) inom EU.
Samtidigt har kunskapen om att halogenerade ämnen även bildas naturligt i
miljön ökat. En stor mängd bromerade organiska ämnen produceras naturligt
i våra hav. Hydroxylerade polybromerade difenyletrar (OH-PBDEer) är en
grupp ämnen som i Östersjön bildas av cyanobakterier och olika arter av
makroalger, men som även kan bildas in vivo via metabolism av PBDEer.
Östersjön är ett av världens största brackvattenhav som på grund av hög
tillförsel av näringsämnen från omgivningen idag lider av utbredd
eutrofiering. Höga halter av näringsämnen främjar tillväxt av fytoplankton
och cyanobakterier men påverkar även makroalgsamhället, genom att tillväxt
av snabbväxande fintrådiga alger gynnas.
Den avhandlingen som här presenteras har haft fokus på förekomsten av
bromerade ämnen av både naturligt och antropogent ursprung i Östersjöns
flora och fauna, med speciellt focus på OH-PBDEer. Avhandlingen omfattar
både toxikologiska och kemiskt analytiska studier med det övergripande
målet att bestämma OH-PBDEers förmåga att störa oxidativ fosforylering
(OXPHOS) in vitro, samt att bedöma exponeringen av dessa ämnen i olika
arter från Östersjön (dvs. i blåmussla, strömming och alfågel).
I Paper I beskrivs två metoder (s.k. bioassays) vilka användes för att studera
18 olika OH-PBDEers potential för OXPHOS-störning in vitro. OXPHOS är
den biokemiska process som producerar huvuddelen av all energi (ATP) i
eukaryota celler under aeroba förhållanden (se Figur 5, sid. 19). OH-
PBDEerna testades enskilt men även i blandningar för att studera eventuella
synergieffekter. Det visade sig att alla de testade ämnena frikopplade
och/eller inhiberade elektrontransportkedjan då isolerade mitokondrier
exponerades. Exponering av zebrafiskceller (PAC2) visade att 13 av de 18
50
OH-PBDEerna störde OXPHOS-processen. Två ämnen, 6-OH-BDE67 och
6-OH-BDE85 befanns vara lika potenta som modell-substanserna FCCP och
SF6847, vilka används vid studier av effekter på OXPHOS. Dessutom
visade sig blandningar av OH-PBDEer ge upphov till synergistiska effekter.
En blandning bestående av sju OH-PBDEer, som adderats i samma
proportioner som de förekommer i blåmusslor från Östersjön, gav effekt
redan vid koncentrationer som endast var 100-1000 gånger högre än vad som
uppmäts i blåmusslorna.
Flera arter av musselätande dyänder i Östersjön, inklusive alfåglar (Clangula
hyemalis), har minskat dramatiskt i antal de senaste 20 åren. Östersjön är ett
av de viktigaste över-vintringsområdena för den europeiska populationen av
Alfåglar. Från oktober till mitten av maj kan man observera alfåglar i
Östersjön, innan de migrerar norrut till sina häckningsplatser i bland annat
västra Sibirien. Då alfåglars kost till övervägande del består av blåmusslor i
Östersjön, som kan innehålla höga halter av OH-PBDEer var det angeläget
att bedöma alfåglars exponering för dessa ämnen.
I Paper II analyserades halterna av bromerade ämnen av både naturlig och
antropogent ursprung i lever från tio alfåglar. Även blåmusslor från platser
där alfåglar födosöker analyserades för att jämföra halter och kongenmönster
mellan arterna. Resultatet visade att alfåglar är exponerade för flera olika
naturligt producerade bromerade ämnen via födan. Men att änderna tycks
utsöndra dessa ämnen snabbt då halterna i levern var betydligt lägre än i
blåmusslorna, trots ett relativt högt estimerat dagligt intag hos änderna. Dock
tycks de mer lipofila PBDEerna ha högre retention i alfågeln än t.ex. MeO-
PBDEer. Även kongenmönstret mellan alfåglar och musslor skiljde sig åt.
OH-PBDEer har i kontrollerade exponeringsförsök visat sig vara mycket
giftiga för zebrafisk (Danio rerio) och zebrafiskembryo vid mycket låga
vatten koncentrationer av OH-PBDE. Det är därför viktig att bedöma
exponeringen av dessa ämnen i fisk från Östersjön.
I Paper III, analyserades strömming från norra egentliga Östersjön (Askö)
och Södra Bottenhavet (Ängskärsklubb). Den naturliga produktionen av OH-
PBDEer i alger varierar under året, men troligen är som störst under
sommarhalvåret. Fiskarna (n=12/lokal) fångades därför i maj/juni strax före
deras reproduktion. Halterna (analyserade som helfiskhomogenat) av ΣOH-
PBDEer skiljde sig inte signifikant åt mellan Askö (9.1 ng/g fett vikt) och
Ängskärsklubb (10 ng/g fettvikt). 6-OH-BDE47 dominerade och utgjorde
87% (Askö) och 91% (Ängskärsklubb) av den total mängden OH-PBDEer. I
Ängskärsklubb uppmättes höga halter (150 ng/g fettvikt) av naturligt
producerade metoxylerade polybromerade difenyletrar (MeO-PBDEer). I
strömming kan dessa ämnen vara en källa till mer giftiga OH-PBDEer då
51
andra fiskarter visats kunna omvandla MeO-PBDEer till OH-PBDEer in
vivo. Även PBDEer uppmättes i strömming från båda lokalerna.
I Paper IV utvärderades en tidigare publicerad extraktions- och
uppreningsmetod för analys av OH-PBDEr i blod från människa (plasma),
katt (serum), strömming (plasma) och alfågel (plasma). Till proverna
tillsattes bestämda mängder av, fyra bromfenoler, nio OH-PBDEer och tre
MeO-PBDEer. Återvinningen av dessa ämnen bestämdes efter extraktion
och upparbetning av proverna. Metoden visade sig fungera väl och ge goda
återvinningar för MeO-PBDEer (dvs. neutrala föreningar) i samtliga matriser
och arter. Även i human plasma och kattserum erhölls goda återvinningar av
bromfenoler och OH-PBDEer. Dock befanns metoden inte vara applicerbar
på plasma från strömming och alfågel då mycket låga återvinnigar av
bromfenoler och OH-PBDEer erhölls.Vidare arbete visade att metyleringen
av OH-PBDEer med det vanligt använda derivatiseringsreagenset
diazometan, var mycket känsligt för samextraktion av hemin/hematin. Det är
därför mycket viktigt att man vid analys av OH-PBDEer undviker prover
med hemolys, då det kan riskera att diskriminera vissa OH-PBDEer i
analysen.
52
8 Acknowledgement
När jag nu skriver de sista raderna på avhandlingen är det med stor GLÄDJE
jag ser tillbaka på mina år som doktorand. För även om arbetet inte alltid har
löpt enkelt och jag ibland har suckat över krånglande analysmetoder, så har
arbetet alltid varit intressant, spännande och givande! Jag är därför
jättetacksam för mina handledare Lillemor Asplund, Åke Bergman och
Göran Marsh som gav mig chansen att bli doktorand i OXPHOS-projektet!
Lillemor, dig vill jag tacka speciellt mycket då du som min
huvudhandledare har varit ett otroligt stöd under alla år! Du har aldrig ryggat
från att delta i fältprovtagning, läsa manuskript och diskutera svåra analys
problem. Du har varit en klippa att luta sig mot, speciellt här på slutet av
avhandlingsskrivandet, Tack för allt!
Jag vill också tacka dig Åke, för att du som inspererande föreläsare fick mig
att fastna för ämnet miljökemi redan under min grundutbildning i kemi. Jag
vill också tacka dig för att du trots ett späckat schema, tog dig tid att ge mig
värdefulla synpunkter både på såväl avhandling som manuskript. Göran, det
blev inte så mycket tid ute på syntes, men jag har upskattat att ha dig som
handledare. Tack!
I am also most thankful to my colleagues and coauthors at IVM, VU Univer-
sity in Amsterdam, The Netherlands. Juliette Legler - for your effort to
arrange for me to come and work in your lab at IVM, I learnt a lot during my
stay! I am also grateful for you for taking time and providing much appreci-
ated comments on paper and thesis. Peter Cenijn – thank you for showing
me the ways around the lab and for showing me the art of cell culturing.
Jessica Legradi – for a nice collaboration with the paper, we did it finally!
Jag vill också tacka mina svenska medförfattare; Vivian Chen Lindberg –
för all din hjälp, men speciellt för att du kämpade på med GPCn i
sommarvärmen, Tack! Kjell Larsson – för prover och värdefulla kunskaper
om alfåglar. Anders Bignert – för hjälp med statistik. Lotta Hovander –
för värdefulla diskussioner. Jessica Norrgran Engdahl – för gott samarbete
under alla år.
53
Jag vill också tacka mina kollegor som gjort att jag trivts så bra!
Jessica och Emelie - för att ni båda har varit fantastiska att dela kontor och
utlandsresor med. Jag kommer att sakna att ha er som arbetskamrater!
Dennis- för alla kluriga diskussioner om metoder och derivatiserings
reagens, Karin– för att du tog så väl hand om mig när jag var ny doktorand.
Anna – för trevliga (sång)stunder på lab, och för hjälp här på slutet, Ge and
Yihui – I whish you all the best, Ioannis – vår klippa, Johan E – för hjälp
med GPCn, Cissi-min vän, Henrik och Jenny – i riskgruppen, Birgit-för
uppmuntrande ord och hjälp här på slutet, Margareta, samt ”gamla”
miljökemister Johan F, Andreas, Hitesh, Maria, Hans, Linda, Lisa, Anna
M, Karin Norström – min examenshandledare, Anita- alltid med ett vänligt
ord Per.
Sist men inte minst Sören – du är en sann inspirationskälla!
Mina älskade föräldrar, familj och vänner som påminner mig om allt i
livet inte är kemi (... bara nästan allt ).
This thesis was financially supported by the Swedish Research Council for
the Environment, Agriculture Sciences and Spatial Planning (FORMAS) and
from Stockholm University’s Strategic Marine Environment Research
Funds, through the Baltic Ecosystem Adaptive Management (BEAM)
project.
54
9 Appendix A
Contribution to Paper I-IV
I. Performed part of the laboratory work and together with my co-
author Jessica Legradi, was responsible for writing the manu-
script.
II. Assisted the laboratory work and was responsible for the data
evaluation and manuscript writing.
III. Performed all laboratory work and was responsible for data
evaluation and writing the manuscript
IV. Performed major part of the laboratory work and together with
my co-author Jessica Norrgran, was responsible for writing the
manuscript.
55
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