envir onmental l y rel evant chemical disruptor s of oxi...

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


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


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].


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].


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].


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.


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.


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


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.


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.


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].


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 bimolecular 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)



Paper IVPlasma/serum from

human, cat, herring and long-tailed duck

DerivatizationDiazomethane

Denaturation/Extraction2-propanol, HCl (6M), c-hx:MTBE (1:1 v/v)



PhenolsNeutrals







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)



PhenolsNeutrals





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


50

100

150

ng

/g lip

id w

eig

ht

PBDE


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


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


50

100

150

200

250

ng

/g lip

id w

eig

ht

PBDEs


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


5

10

15

20

ng

/g f

resh

we

igh

t

PBDE


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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