Brominated flame retardants and

perfluoroalkyl acids in Swedish indoor

microenvironments

Implications for human exposure

Justina Björklund

Doctoral thesis in Applied Environmental Science

Department of Applied Environmental Science

Stockholm University

2011

ii

Doctoral Thesis, 2011

Justina Björklund

Department of Applied Environmental Science (ITM)

Stockholm University

SE-106 91

©Justina Björklund, Stockholm 2011

ISBN 978-91-7447-393-3, pp i - x, 1 - 47

Printed in Sweden by US-AB, Stockholm 2011

Distributor: Department of Applied Environmental Science

Cover graphic: courtesy Kaj Thuresson, modified Justina Björklund

iii

Dedication

This thesis is dedicated to my boys Hugo and Percy.

It is also dedicated to my parents Mathias and Christiana Awasum, who

taught me that the best kind of knowledge to have is that which is learned for

its own sake and that with patience, even a hot plate of soup can be licked.

iv

v

Abstract

Humans are exposed to persistent organic pollutants (POPs) such as bromi-

nated flame retardants (BFRs, specifically polybrominated diphenyl ethers

(PBDEs) and hexabromocyclododecane (HBCD)) and perfluoroalkyl acids

(PFAAs, specifically perfluoroalkane sulfonate (PFOS) and perfluoroocta-

noic acid (PFOA)). They are used in consumer products found in cars, offic-

es, homes and day care centers. Diet was earlier thought to be a major hu-

man exposure route for legacy POPs, but does not account for body burdens

found for many new POPs and indoor exposure from air and dust has been

hypothesized as also important.

In this thesis, BFRs in air and dust, and PFAAs in dust from different indoor

microenvironments in Sweden were analysed, and the results used to esti-

mate human exposure. BFRs and PFAAs were detected in dust from all mi-

croenvironments and PBDEs in all air samples. BFR and PFAA exposure

occurs mostly in peoples’ homes with toddlers having higher intakes from

dust ingestion than adults. Inhalation and dust ingestion play minor roles

compared to diet for humans with median exposures, but in worst case sce-

narios, dust ingestion may be significant for a small part of the Swedish

population. Sampling using home vacuum cleaner bag dust and researcher-

collected above floor dust was compared. Correlations were seen for

∑OctaBDE and ∑DecaBDE but not for ∑PentaBDE and HBCD. Higher

PBDE concentrations were found in above floor dust but higher HBCD con-

centrations were found in vacuum cleaner bag dust. BDE-47 concentrations

were correlated between vacuum cleaner bag dust and breast milk, indicating

exposure through dust ingestion.

Similar concentrations of PBDEs were measured in indoor and outgoing air

from day care centers, apartment and office buildings. Indoor air explained

54-92% of ∑PentaBDE and 24-86% of BDE-209 total emissions to outdoor

air in Sweden, supporting the hypothesis that the indoor environment is pol-

luting ambient air via ventilation systems.

vi

Svensk sammanfattning

Människor exponeras för långlivade organiska miljögifter (POPs), såsom

bromerade flamskyddsmedel, (BFR, specifikt polybromerade difenyletrar

(PBDE) och hexabromcyklododekan (HBCD)) och perfluoralkylsyror

(PFAA, specifikt perfluoroktansulfonat (PFOS) och perfluoroktansyra

(PFOA)). Dessa används i konsumentprodukter i bilar, kontor, hem och

förskolor. Tidigare har kosten antagits vara den huvudsakliga

exponeringskällan för POPs, men för ett flertal nya POPs kan inte de

uppmätta nivåerna i kroppen förklaras med kostintag. En hypotes är att

exponeringen från luft och damm inomhus också kan vara viktig.

I denna avhandling har BFR i luft och damm och PFAA i damm från olika

inomhusmiljöer i Sverige analyserats och resultaten har använts för att

uppskatta människors exponering. BFR och PFAA detekterades i damm och

PBDE i alla luftprover från alla inomhusmiljöer. Exponering från BFR och

PFAA sker oftast i människors hem, där småbarn har högre estimerade intag

från damm än vuxna. Uppskattningarna visade att inandning och dammintag

spelade mindre roll i förhållande till kosten för människor med måttlig

exponering, men i vissa fall kan dammintag ha betydelse för en liten del av

den svenska befolkningen.

Dammprover som togs direkt från dammsugarpåsar jämfördes med prover

som samlades av forskare från ytor en meter ovanför golvet. Korrelationer

sågs för ΣOctaBDE och ΣDecaBDE men inte för ΣPentaBDE och HBCD.

Högre PBDE halterna fanns i dammproverna som togs av förskarna men

högre HBCD halterna fanns i dammproverna från dammsugarpåsen. Det

fanns en korrelation mellan BDE-47 koncentrationerna i damm från

dammsugarpåsen och bröstmjölk, vilket kan indikera exponering genom

dammintag.

Liknande halter av PBDE uppmättes i inomhus- och utgående luft från

daghem, lägenheter och kontorsbyggnader. Inomhusluft förklarade 54-92%

av ΣPentaBDEs och 24-86% av BDE-209s totala utsläpp till utomhusluft i

Sverige. Detta stöder hypotesen att inomhusmiljön förorenar utomhusluft

genom ventilationssystem.

vii

List of Papers and statement of responsibility

Paper I: Björklund, J. A., Thuresson, K. and de Wit, C. A. (2009)

Perfluoroalkyl compounds (PFCs) in Indoor dust: concentrations, human

exposure estimates, and sources. Environ. Sci. Technol., 43, 2276-2281.

Paper II: Thuresson, K., Björklund, J. A. and de Wit, C.A. (2011) Tri-

decabrominated diphenyl ethers and hexabromocyclododecane in indoor

air and dust from Stockholm microenvironments 1: Levels and profiles.

Sci. Total Environ., In press.

Paper III: Björklund, J. A.; Sellström, U.: de Wit, C. A.; Aune, M.; Lig-

nell,S.; Darnerud, P. O. (2011) Comparisons of PBDE and HBCD con-

centrations in dust collected with two sampling methods and matched

breast milk samples Indoor Air. Accepted for publication.

Paper IV: Björklund, J. A., Thuresson, K., Palm Cousins, A., Sellström, U.

and de Wit, C. A. (2011) Indoor air is a significant source of tri-

decabrominated diphenyl ethers to outdoor air via ventilation systems.

Manuscript.

Paper I is reproduced with permission from Environmental Science and

Technology.

My contributions to the papers included in this thesis were:

Paper I: I performed all laboratory work including the development and

evaluation of the method, with exception of the HPLC analysis. I was re-

sponsible for data acquisition, interpreting the data and took the lead in writ-

ing the paper.

Paper II: I was involved in extraction and analysis of the samples, data ac-

quisition and assisted in writing the manuscript.

Paper III: I contributed to the planning of this study and was responsible for

chemical analysis of the dust samples, data acquisition, statistical analysis

and interpretation. I wrote the drafts of the paper and finalized it with com-

ments and contributions from the co-authors.

Paper IV: I did the extraction and analysis of the samples, data acquisition,

statistical analysis and contributed to interpretation and I had the main re-

sponsibility for writing the paper with the exception of the modeling section.

viii

Abbreviations

AFSD Above floor settled dust

BDE-209 Decabromodiphenyl ether

BFR Brominated flame retardant

BSEF Bromine Science and Environmental Forum

ECNI Electron capture negative ionisation

EFSA European Food Standard Agency

EPA Environmental Protection Agency

EPS Expanded polystyrene

GC Gas chromatography

HBCD Hexabromocyclododecane

HIPS High impact polystyrenes.

HPLC High performance liquid chromatography

Kow n-octanol/water partition coefficient.

LOD Limit of detection

LOQ Limit of quantification

PBDE Polybrominated diphenyl ethers

PFAAs Perfluoroalkyl acids

PFOS Perfluoroalkane sulfonate

PFOA Perfluorooctanoic acid

POPs Persistent organic pollutants

RfD Reference dose

SVOC Semivolatile organic compounds

TDI Tolerable daily intake

VCBD Vacuum cleaner bag dust

ix

Contents

Dedication ..................................................................................................... iii

Abstract ........................................................................................................... v

Sammanfattning ............................................................................................. vi

List of Papers and statement of responsibility ..............................................vii

Abbreviations .............................................................................................. viii

Contents ......................................................................................................... ix

1. Background ............................................................................................ 1

1.1 Perfluoroalkyl acids .......................................................................... 1

1.1.1 Health effects of PFAAs .......................................................... 2

1.1.2 Routes of human exposure to PFAAs ...................................... 3

1.2 PBDEs and HBCD ............................................................................ 3

1.2.1 Health effects of PBDEs and HBCD ....................................... 5

1.2.2 Routes of human exposure to PBDEs and HBCD ................... 6

1.3 Thesis Overview ............................................................................... 8

2. Objectives and Hypothesis .................................................................... 9

3. Experimental Methods ......................................................................... 10

3.1 Sample collection ............................................................................ 10

3.1.1 Sampling sites ........................................................................ 10

3.1.2 Indoor air ............................................................................... 10

3.1.3 Indoor Dust ............................................................................ 11

3.1.4 Breast milk sampling ............................................................. 12

3.2 Extraction and chemical analysis .................................................... 13

3.2.1 Extraction and cleanup........................................................... 13

3.2.2 Instrumental analysis ............................................................. 13

3.3 Quality assurance/Quality control ................................................... 14

3.4 Statistical analyses ......................................................................... 14

4. Results and discussion ......................................................................... 15

4.1 PFAAs in indoor dust ...................................................................... 15

4.1.1 Levels ..................................................................................... 15

4.2 BFRs in indoor dust and air ............................................................ 18

4.2.1 Levels ..................................................................................... 18

4.2.2 Estimated human exposure to PFAAs and BFRs .................. 22

4.2.3 Tolerable daily intakes ........................................................... 26

4.2.4 BFR levels in breast milk....................................................... 26

x

4.2.5 Comparison of dust sampling methods .................................. 26

4.2.6 Association between breast milk and dust levels ................... 27

4.2.7 Levels of PBDEs in indoor and outgoing air ......................... 28

4.2.8 Transport of BDE-209 to ambient air .................................... 28

4.2.9 Significance of ∑PentaBDE and BDE-209 emissions to

ambient air ........................................................................................... 29

5. Conclusions ......................................................................................... 30

5.1 Knowledge gaps and future perspectives. ....................................... 31

6. Acknowledgements ............................................................................. 32

7. Reference List ...................................................................................... 33

1

1. Background

A large number of chemicals off-gas or leach into indoor environments from

sources such as consumer products, household products, furniture, textiles

and building materials. The indoor environments can include private homes

and public buildings, e.g. schools, day care centers, offices, places of leisure

and transport vehicles. Characterizing and understanding the different path-

ways of chemicals from sources to human exposure and effects is vital in

order to implement control strategies and lower exposure.

Because of the extensive use of chemicals in commercial and household

products, many semivolatile organic compounds (SVOCs) have been meas-

ured in higher concentrations indoors than outdoors. Because of their slow

rate of release from sources and their propensity to partition and sorb to sur-

faces, SVOCs can persist indoors for years after they are introduced

(Weschler and Nazaroff, 2008).

Two such groups of chemicals are the BFRs (specifically the PBDEs and

HBCD), and the PFAAs. These chemicals are high production volume

chemicals. A few previous studies (Shoeib et al., 2004; Wilford et al., 2004)

had shown higher concentrations of these chemicals indoors than outdoors.

Coupled with humans spending much more than 90% of their time indoors

(Harrad et al., 2010), indoor exposure to these compounds was of interest.

1.1 Perfluoroalkyl acids

PFAAs are synthetic chemicals, whose structures are characterized by a fully

fluorinated carbon tail attached to a hydrophilic head, the functional group.

Depending on the nature of the functional group, PFAAs can be grouped as

sulfonates (e.g. PFOS), and carboxylates (e.g. PFOA). PFOS and PFOA are

the PFAAs that are most abundant in humans (Vestergren and Cousins,

2009) and the most studied. Their structures are shown in Figure 1.1.

F F

F

F F

F

F

n

S

O

O

O

PFOS, n = 7

FF

F

n

C

O

O

PFOA, n = 6

Figure 1.1. Structures of PFOS and PFOA.

2

Due to the hydrophobic properties of the tail and the hydrophilic nature of

the head, PFAAs have been used as surfactants and surface protectors in

many consumer products and industrial applications such as textiles, leather

and fire-fighting foam. PFAA-coated products may include non-stick cook-

ware, Teflon®, GORE-TEX®, waterproof clothing, fast food wrappers, piz-

za boxes, popcorn bags, stain-resistant carpet, paint, and windshield washer

fluid (Kissa, 2001).

PFAAs have been manufactured through two processes, electrochemical

fluorination and telomerization. Electrochemical fluorination produces a

mixture of linear and branched isomers, while telomerization yields only

linear products (De Silva et al., 2009). PFOS is also a residual formed in the

production of perfluorooctanesulfonyl fluoride (POSF)-based fluorochemi-

cals that have been produced for over 40 years by electrochemical fluorina-

tion (Olsen et al., 2005). POSF was used to produce surfactants and fluori-

nated side chain polymers which were used in paper and packaging treat-

ments, as well as carpet and upholstery surface protectants (Buck et al.,

2011). PFOS is also a metabolite of many other POSF-related compounds

such as perfluorooctane sulfonamide and N-methyl perfluorooctane sulfo-

namidoalcohol (e.g N-MeFOSE) (Seacat et al., 2002; Tomy et al., 2003).

PFOA has been produced for over 50 years as an emulsifying agent in the

production of fluoropolymers such as polytetrafluoroethylene used for many

purposes including non-stick cookware (Prevedouros et al. 2006). PFOA can

also be formed from degradation/transformation of fluorotelomer alcohols

(FTOH) (Ellis et al., 2004).

Based on recommendations from the European Commission’s Scientific

Committee on Health and Environmental Risks (SCHER), which has classi-

fied PFOS as very persistent, very bioaccumulative and toxic, the European

Union has restricted its use and is also considering the risks of PFOA expo-

sure (European Union, 2006). In May 2009, PFOS was banned under the

United Nations Environmental Program (UNEP) Stockholm Convention on

Persistent Organic Pollutants (POPs), which aims to protect human health

and the environment from POPs, albeit with exemptions for certain uses

(UNEP Stockholm Convention on POPs, 2009c).

1.1.1 Health effects of PFAAs

PFOS and PFOA are ubiquitous in the environment, have long half-lives in

humans (5.4 and 3.8 years, respectively) (Olsen et al., 2007), and have been

detected in human breast milk and human blood (Kannan et al., 2004;

Karrman et al., 2007). Despite this, the implications of their presence in hu-

mans are still not fully understood. However, some animal studies provide

some indications that PFOA exposure causes liver toxicity, carcinogenicity,

3

increased liver weight and developmental toxicity (Beigel et al., 2001;

Butenhoff et al., 2002; Butenhoff et al., 2004; White et al., 2007). PFOS

exposure in cynomolgous monkeys led to reduced body weight, increased

liver weight, decreased total serum cholesterol and changes in levels of thy-

roid stimulating hormone (Seacat et al., 2002). Apelberg et al. (2007) found

levels of PFOA and PFOS in umbilical cord blood to be inversely related to

birth weight.

In human studies, women with higher serum levels of PFOA and PFOS had

increased risk of infertility and were also more likely to have irregular men-

strual cycles (Fei et al., 2009). A Danish study reported that young men with

high combined levels of PFOS and PFOA had less than half the number of

normal sperm than men with low levels of these chemicals (Joensen et al.,

2009). Elevated PFOS and PFOA exposure have been associated with in-

creased cholesterol levels (Nelson et al., 2009).

1.1.2 Routes of human exposure to PFAAs

The major sources of PFAA exposure in humans are not well understood.

Modeling studies indicate diet as a major exposure route for PFOS and

PFOA (Egeghy and Lorber, 2011; Trudel et al., 2008; Vestergren et al.,

2008; Vestergren & Cousins, 2009), with fish, dairy and beef products

representing the most significant sources (Ericson et al., 2008; Noorlander et

al., 2011; Ostertag et al., 2009; Tittlemier et al., 2007). Additional dietary

intake of PFAAs may also originate from grease- and water-resistant coat-

ings in food packing materials often used for fast food (Begley et al., 2008)

and drinking water (Skutlarek et al., 2006).

1.2 PBDEs and HBCD

PBDEs and HBCD (Figure 1.2) are additive flame retardants and a sub-

group of BFRs, used in commercial products to reduce flammability, giving

people more time to extinguish or escape the fire. When fire occurs, the

BFRs utilize vapor phase chemical reactions that interfere with the combus-

tion process, thus delaying ignition and inhibiting the spread of fire. These

characteristics promote their use in textiles, flexible polyurethane foams

used in upholstery stuffing for furniture and car seats, electronic and elec-

trical components, and plastics used in the casings of televisions, personal

computers, and other electronic equipment. As additive rather than reactive

flame retardants, PBDEs and HBCD are likely to be released from the prod-

ucts to which they are added (Hutzinger and Thoma, 1987; Stapleton et al.,

2008).

4

PBDEs have a backbone structure of a brominated diphenyl ether molecule

(Figure 1.2) that may have from 1 to 10 bromine atoms attached. Depending

on the location and number of bromine atoms, there are 209 possible confi-

gurations or congeners and each has been assigned a unique brominated

diphenyl ether (BDE) number.

PBDEs have been marketed for use in commercial products as three technic-

al mixtures: Penta-, Octa- and DecaBDE. The PentaBDE mixture consists

primarily of BDE-47 and -99 (about 37% of each) along with smaller

amounts of other tri-hexaBDEs (primarily BDE-28, -100, -153, -154). Oc-

taBDE is a mixture of hexa- (10−12%), hepta- (44−46%), octa- (33−35%),

and nonaBDEs (10−11%). The major congeners are BDE-183 and, to a mi-

nor extent, BDE-197 and -203. The DecaBDE mixture consists predominant-

ly of BDE-209 (98%) and nonaBDEs (2%) (La Guardia et al., 2006;

McDonald, 2002) .

Br

Br

Br

Br

Br

Br

HBCD

Br1-10

O

PBDE

Figure 1.2. General structures of PBDE and HBCD

The PentaBDE technical product was primarily used in polyurethane foam

for furniture, but has also been used in computer circuit boards and textiles.

The Octa- and DecaBDE products are mainly found in plastic housings for

electronics and electrical appliances and back coatings of textiles and carpets

(de Wit, 2002).

The technical HBCD (Figure 1.2) mixture consists of mainly three stereo-

isomers, α-, β-, and γ-HBCD, with the γ isomer being the predominant one

(Alaee et al., 2003). The primary use of HBCD was to flame retard

expanded and extruded polystyrene foams used for thermal insulation in

buildings and in the construction industry (Morose, 2006). HBCD has also

been used for back-coating of textiles and in high impact polystyrene used

in electronic equipment like TV sets and computers (Covaci et al., 2006).

5

Due to their physical chemical properties, 20% of BDE-47, 60-90% of

penta-heptaBDEs and almost 100% of BDE-209 are expected to partition to

particles at room temperature, when released to air (Shoeib et al., 2004).

HBCD is predicted to behave similarly to the penta-hexaBDEs (Meyer and

Wania, 2004). The production, import and usage of PBDEs and HBCD vary

greatly geographically, e. g. the strict fire regulations in the US have led to

higher usage of PentaBDE than in the EU. Likewise the stricter fire safety

standards in the UK have resulted in the highest measured dust concentra-

tions of BDE-209 and HBCD in Europe (Harrad et al., 2008).

PBDEs and HBCD are lipophilic compounds. PBDEs (tri- heptaBDE) and

HBCD are persistent, bioaccumulative and toxic and undergo long-range

atmospheric transport, all typical behaviors for POPs (Birnbaum and Staskal,

2004; Darnerud, 2003; de Wit, 2002; Muir and Howard, 2006). Although

BDE-209, the major constituent of DecaBDE is not considered to be a clas-

sical POP, studies have shown that it can degrade to lower brominated

PBDEs upon exposure to sunlight (Söderström et al., 2004; Stapleton and

Dodder, 2008).

Penta- and OctaBDE technical mixtures were banned within the EU in 2004

and banned globally under the Stockholm Convention in 2009 (Cox and

Efthymiou, 2003; UNEP Stockholm Convention on POPs, 2009b).DecaBDE

was banned in Sweden in 2006 but in May 2008 the ban was reversed by the

Swedish government (Regeringskansliet, 2008). Since June 2008, DecaBDE

is prohibited in electronic and electrical equipment according to the EU’s

Restriction of Hazardous Substances Directive (RoHS). DecaBDE produc-

tion by major producers in the US will be voluntarily discontinued by 2013

(Hess G., 2009).

The EU announced in February 2011 that a ban has been placed on HBCD,

to take effect by mid-2015 and to be implemented through the EU’s REACH

program (Registration, Evaluation, Authorization and Restriction of Chemi-

cals). The goal of this program is to protect human health and environment

from the risks posed by chemicals. HBCD is currently proposed to be re-

viewed under the Stockholm Convention (UNEP Stockholm Convention on

POPs, 2009a). Although PBDEs and HBCD have been banned, phased out

or will be phased out worldwide, different consumer products (reservoirs)

containing these substances will still be in use for decades to come, resulting

in continuing releases (Harrad and Diamond, 2006).

1.2.1 Health effects of PBDEs and HBCD

Toxicological effects of PBDEs seen in laboratory animals include endocrine

disruption, neurodevelopmental and behavioral outcomes, hepatic abnormal-

ities, and possibly cancer (Birnbaum & Staskal, 2004; Costa and Giordano,

6

2007; Darnerud, 2008; Zhou et al., 2001). The half-lives in humans of the

main congeners in the PentaBDE mixture have been estimated to be 2-3

years (BDE-47, 99, 100, and 154) and 4-6 years (BDE-153). Hepta-

decaBDEs have relatively short half-lives (16 days) in humans (Geyer et al.,

2004; Thuresson et al., 2006). The half-life of HBCD in human adipose tis-

sue was observed to be 64 days (Geyer et al., 2004). The long estimated

half-lives for PentaBDE congeners in humans raises concern about their

long-term effects on human health.

Although little human epidemiology has yet been done, research findings for

PentaBDE congeners are consistent with animal studies. Epidemiological

studies have reported associations between exposure to PentaBDE congeners

and effects on reproduction (Akutsu et al., 2008), neurodevelopmental

effects (Chao et al., 2007; Herbstman et al., 2010), cryptorchidism (Main et

al., 2007), testicular cancer (Hardell et al., 2006), thyroid function ((Chevrier

et al., 2010; Turyk et al., 2008) and endocrine disruption (Meeker et al.,

2009).

For HBCD, rodent studies have shown effects on neurobehavioral function

(Lilienthal et al., 2009), thyroid dysfunction (Darnerud, 2003) and endocrine

disruption (Darnerud, 2003; Vonderheide et al., 2008). No human effect

studies for HBCD are available yet.

1.2.2 Routes of human exposure to PBDEs and HBCD

The increasing number of studies showing measurable body burdens of

BFRs coupled to the growing epidemiological evidence of human health

effects raises the question of how humans are exposed to these compounds.

Dietary intake has been assumed to be the major exposure pathway for these

compounds, just as for legacy POPS, with dust ingestion, inhalation and

dermal exposure playing minor roles. In Europe, fish, followed by meat and

dairy products are the largest contributors to PBDEs and HBCD in the diet

(Darnerud et al., 2006; Roosens et al., 2009b; Voorspoels et al., 2007), while

in the US meat is the main contributor (Schecter et al., 2010b). Concentra-

tions of PBDEs and HBCD measured in dietary studies from the US and

several European countries indicated comparable dietary intakes (Darnerud

et al., 2006; Fromme et al., 2009; Roosens et al., 2009a; Schecter et al.,

2010b). However, the body burdens of PentaBDE in Americans were much

higher than the burdens reported for populations in other parts of the world

(Hites, 2004; Johnson-Restrepo et al., 2005; Lorber, 2008). This implied that

food alone could not account for the total human body burden in the US

(Johnson-Restrepo and Kannan, 2009). PBDE levels measured in human

studies including in Sweden (Glynn et al., 2011; Lignell et al., 2009) are

highly skewed, with a few individuals having 10-100 times higher levels

7

than the mean, also suggesting that there may be other significant exposure

pathways than diet.

PBDEs and HBCDs have been detected in indoor air (Abdallah et al.,

2008a; Allen et al., 2007; Shoeib et al., 2004; Wilford et al., 2004) and dust

(Abdallah et al., 2008b; Allen et al., 2008; Harrad et al., 2006; Roosens et

al., 2009a; Wilford et al., 2005). Levels of PBDEs in indoor air are 10-50

times higher than those outdoors, with concentrations decreasing from urban

areas and cities to rural and background areas (Harrad et al., 2010). The

highest PentaBDE concentrations in dust have been measured in the US.

These results led to speculation that emissions from flame-retarded products

may lead to elevated PBDE levels indoors. Furthermore, it was hypothesized

that other exposure routes such as dust ingestion and air inhalation are im-

portant human exposure pathways. In support of this, modeling studies

showed dust ingestion could be the most important exposure pathway to

PentaBDE for most North Americans (Jones-Otazo et al., 2005; Lorber,

2008). Thus, the sources of human exposure to BFRs could not only be diet

but also from the indoor environment (inhalation and dust ingestion) and

possibly also from dermal contact (Allen et al., 2008; Harrad et al., 2006;

Jones-Otazo et al., 2005; Lorber, 2008; Wilford et al., 2004).

Large differences in PBDE levels in indoor dust and body burden have been

observed in the US compared to Europe (Frederiksen et al., 2009). Recent

European studies from Belgium and the UK indicate dietary intake may be

more important as a route of human exposure to PentaBDE and HBCD than

in the US (Harrad et al., 2004; Roosens et al., 2009b). Within the EU, there

is relatively little geographic variation in adult human PBDE levels except

for the UK. Similar to observations for the US population, Knutsen et al.

(2008) and Thomsen et al. (2008) also found that dietary intake alone was

insufficient to explain body burdens of PBDEs and HBCD in a Norwegian

cohort. This implies that other exposure routes than diet are also significant

for European exposure. Karlsson et al. (2007) found a correlation between

levels in serum and levels of PBDEs in indoor dust while Fromme et al.

(2009) found no direct correlation between diet or dust PBDE concentrations

and concentrations in blood. Regional differences in exposure pathways may

also be influenced by life style, product usage, policy regulations and cultur-

al differences (Frederiksen et al., 2009). Age may also play a role as meas-

ured concentrations of PBDEs and HBCD in children have been found to be

2 - 5-fold higher than the concentrations in adults (Lunder et al., 2010;

Thomsen et al., 2002; Toms et al., 2009b) possibly due to the higher dust

ingestion rates of young children (Jones-Otazo et al., 2005).

8

1.3 Thesis Overview

At the start of this project in 2006, no data existed on PFAA concentrations

in dust in Europe nor were there indicative values of PFAAs in the standard

reference material SRM 2585 (house dust). The few studies that existed from

Japan, Canada and the US indicated the presence of PFOS and PFOA in

indoor dust from homes but no data for other microenvironments existed

(Kubwabo et al., 2005; Moriwaki et al., 2003; Strynar and Lindstrom, 2008).

Limited Swedish data existed for the presence of BFRs in non-occupational

microenvironments. The only existing data from non-occupational settings

were for PBDEs in indoor air and dust from five homes in Örebro (Karlsson

et al., 2007) and indoor air from two offices in Stockholm (Sjödin et al.,

2001). A few studies from the UK, Canada and the USA had assessed human

exposure to tri-heptaBDEs via dust ingestion and inhalation from samples

taken from home environments only (Harrad et al., 2004; Harrad et al., 2006;

Stapleton et al., 2005; Wilford et al., 2004; Wilford et al., 2005). No human

exposure assessment existed for HBCD and BDE-209, both BFRs with high

global production volumes according to the 2001 global production figures

(BSEF, 2003) and both still on the market. Data for BFRs in some common-

ly frequented indoor microenvironments like cars, offices and day care cen-

ters were lacking.

Swedish citizens spend a disproportionate fraction of time indoors (typically

in excess of 90%) where these compounds are predominantly used and re-

cent findings showed children had higher concentrations of these compounds

than adults (Calafat et al., 2006; Toms et al., 2009b). Thus, there was cause

for concern and a need to quantify these compounds in the indoors in order

to better assess exposure.

Studies have linked indoor dust exposure to body burden but the relevance

of different dust sampling methods is not clear. While studies have shown

associations between PentaBDE concentrations in vacuum cleaner bag dust

and concentrations in breast milk and blood (Johnson et al., 2010; Wu et al.,

2007), and HBCD concentrations in researcher-collected floor dust corre-

lated with those in human blood (Roosens et al., 2009a), no study has been

performed to check the conformity of BFR concentrations between vacuum

cleaner floor dust and above-floor settled dust, nor the associations between

above-floor settled dust and breast milk concentrations.

Elevated concentrations of PBDEs have been measured in indoor air com-

pared to outdoor air, with the presence of gradients from cities to rural areas

(Bohlin et al., 2008; Butt et al., 2003; Cetin and Odabasi, 2011; Rudel et al.,

2010; Rudel and Perovich, 2009; Shoeib et al., 2004; Wilford et al., 2004).

9

This has led to speculation that indoor air is a source to outdoor air (Gouin et

al., 2006; Harrad and Hunter, 2006; Zhang et al., 2011). One possible me-

chanism for this was hypothesized to be via ventilation systems.

2. Objectives and Hypothesis

The major objectives of this study were to:

1) Obtain further knowledge of the significance of indoor air and dust

as pathways of human exposure to PFOS, PFOA, PBDEs and HBCD.

2) Extend the range of microenvironments studied from homes to in-

clude apartments, offices, day care centers and cars in Sweden. This in-

cluded:

Quantification of tri-decaBDEs and HBCD in indoor air and dust, and

PFOS and PFOA in dust from homes, offices, day care centers, cars

and apartments.

Estimation of exposure to PFOS and PFOA from ingestion of dust.

3) Determine the internal consistency of two dust sampling methods and to

examine the relationship between indoor exposure via dust and human

body burden by:

Measuring and comparing concentrations of PBDEs and HBCD in

matched dust samples collected using both methods

Comparing concentrations of PBDEs and HBCD in dust samples col-

lected using both methods to concentrations in matched breast milk

samples.

4) Study the role of indoor air as a source of PBDEs to the outdoor environ-

ment.

The hypotheses were:

Indoor air and dust are significant exposure pathways for PFAAs and

BFRs for the Swedish population, particularly for toddlers.

Indoor air is a significant source of tri-decaBDEs to outdoor air.

10

3. Experimental Methods

3.1 Sample collection

3.1.1 Sampling sites

Indoor air (gas and particle phase) and dust samples were collected from 54

homes (10 houses, 44 apartments (dust for 34)), 10 day care centers and 10

offices (different buildings) from the Stockholm City area in 2006 (paper I,

II). Buildings were chosen to represent different construction years and dif-

ferent parts of the city. Attempts were made to sample four apartments on

different floors in each apartment building chosen. Air samples were also

collected from ventilation systems of eleven apartment buildings, five day

care centers and nine office buildings (paper IV). In addition, seventeen cars

from seven different manufacturers were sampled with windows closed (pa-

per II). Five cars from two different manufacturers were sampled twice,

once while indoors at room temperature, and once standing outside in the

sunshine in the summer so that a higher indoor temperature was reached.

The remaining twelve cars represented five other manufacturers and were

sampled indoors only. Together, the seven manufacturers represent the major

car models sold in Sweden. All cars were new and were sampled indoors in

dealership halls with the car fan on to simulate ventilation during driving

conditions. In addition, indoor air samples were taken in two of the dealer-

ship halls. In paper III, dust sampling was done from 19 homes in Uppsala

and breast milk samples from women residing in these homes were also col-

lected.

3.1.2 Indoor air

The standard techniques used to sample PBDEs in indoor air, are the passive

(diffusive) and active (low and high volume) air sampling techniques. Al-

though the passive sampler has the advantage that it is easy to use and dep-

loy (facilitates simultaneous deployment in a large number of locations),

inexpensive and does not require electricity, it samples effectively only the

gas phase (Shoeib and Harner, 2002) which renders it inappropriate for mon-

itoring airborne levels of contaminants which are preferentially associated

with particles such as BDE-209 (Hoh and Hites, 2005). In active sampling, a

defined volume of air is pumped through a sampling train consisting of two

components (Figure 3.1): a filter (made of e.g. teflon, glass fiber (GFF) or

quartz fiber) and an adsorbent (e.g. polyurethane foam (PUF), XAD-2 or

Tenax) where pollutants (particle-associated and gaseous respectively) are

retained (Ras et al., 2009). Because of the large size and noise level of high

volume active samplers, low volume active samplers were used in this study.

11

The sampling train consisted of a GFF and two PUFs. To increase the mass

of sample, four sampling trains were placed in parallel on one pump (Figure

3. 1) with a flow rate of 12 L/min (3 L/min per sampler). The sampler was

hung on a stand or otherwise suspended at least 1 m above the floor with the

sampler train (filter end) pointing downwards. Air from buildings was sam-

pled for 8 (offices, day care centers) or 24 hours (houses, apartments) during

the winter half of the year (heating season) when windows and doors are

more often closed. The houses, day care centers and offices were sampled in

March-April 2006, the cars in June 2006 and the apartments in October-

December 2006 (paper I & II). Air samples were also collected from the

ventilation system in 11 apartment buildings, 5 day care centers and 9 office

buildings (paper IV). The sampling train was positioned in the ventilation

duct, after the ventilation fans just before the final exit point to outdoors.

The ventilation sample and the indoor air sample were taken simultaneously.

PFOS and PFOA were not analysed in air samples.

PUFs

Figure 3.1. Low volume active air sampler used to collect BFRs.

3.1.3 Indoor Dust

Indoor dust consists of a variety of particles of different sizes having both

inorganic and organic sources. These include combustion products, frag-

ments of fibers and hair, insect remains, dandruff, sand, pollen grains and

fungal spores (Lioy et al., 2002). These particles can sorb semivolatile or-

12

ganic compounds (SVOCs) present in the environment and later re-emit

them by desorption. Dust sampling methods used in other studies include

sampling directly from vacuum cleaner bags, or using a filter either inside

the vacuum cleaner or at the intake nozzle of the vacuum cleaner (Harrad et

al., 2010). Most often floor dust is sampled. Although using vacuum cleaner

bag dust is cost-effective, easy to use and also provides an integrated meas-

ure of contamination in the home, such integrated samples may not reflect

the levels of contamination in different rooms, which may affect exposure

assessments. There is also a potential for contamination of the dust from the

vacuum cleaner itself when taken from vacuum cleaner bags or from filters

inside the vacuum cleaner. It is preferable to use a sampler that catches the

dust before it enters the vacuum cleaner.

In this study, dust samples were collected after the air sampling was com-

plete. Sampling was done in the living room using an industrial strength

vacuum cleaner equipped with a forensic nozzle containing a cellulose filter

(Figure. 3.2). Sampling was done from surfaces at least one meter above the

floor (above floor settled dust (AFSD)) in order to eliminate dirt, gravel and

sand. For comparison vacuum cleaner bag dust (VCBD) was also used in

paper III. Both air (PUFs and filters) and dust samples were wrapped in

aluminum foil, sealed in separate plastic bags and kept frozen at -20 °C.

Figure 3.2. Indoor dust sampling device used in this study.

3.1.4 Breast milk sampling

Breast milk samples were collected by the mothers during the third week

after delivery (starting approximately on day 15 postpartum) using a manual

breast milk pump (paper III). Small milk volumes were collected at the

beginning and end of the breast-feeding sessions, and the milk was stored in

acetone-washed glass bottles in the home freezer during the sampling week.

13

3.2 Extraction and chemical analysis

3.2.1 Extraction and cleanup

Ultrasonication, a rapid extraction method which uses small amounts of

solvents, was used for extracting target compounds from air and dust

samples. Briefly for PFAAs, dust samples spiked with surrogate standards of 13

C-PFOA and 18

O-PFOS were extracted with methanol using an integrated

extraction and clean-up method based on Powley et al. (2005) with some

modification. The method integrated a clean-up step side-by-side with

extraction as Envi-Carb was added directly to the dust samples before

extraction. Injection standards (BDE-77 for BFRs and 3,5-

bis(trifluoromethyl) phenyl acetic acid for PFAAs) were added prior to

instrumental analysis to calculate the recovery of the surrogate standard. The

filter and both PUFs from the four sampling trains were combined and ex-

tracted together. For BFRs, dust and air (PUFs and filters) samples were

spiked with surrogate standards (13

C-BDE-209 and Dechlorane 603) and

extracted with dichloromethane in an ultrasonic bath. Clean-up of sample

extracts was done on H2SO4/SiO2-gel (1:2, w/w).

3.2.2 Instrumental analysis

PFOS and PFOA were measured using LC/MS in selected reaction monitor-

ing (SRM) mode. Ion transitions monitored were m/z 499>80 for PFOS and

m/z 413>369 for PFOA (paper I). PFAAs were quantified by single point

calibration versus an external standard. The linearity of the response was

checked by injection of external standards at different concentrations.

BFRs were analysed using MS in the electron capture negative ion chemical

ionization (ECNI) mode (paper II, III & IV) employing selected ion moni-

toring (SIM). The ions monitored were m/z 237 and 239 for the surrogate

standard Dechlorane and 494.3 and 496.3 for 13

C-BDE-209. The analyses of

nonaBDE and BDE-209 were done by monitoring m/z 484.2 and 486.2. The

remaining BDE congeners and HBCD were analysed by monitoring m/z 79

and 81. Identification of BFRs in the sample was based on comparison of the

retention times of these compounds with those of the authentic reference

substances. Quantification of all BFRs was performed by GC/MS (ECNI)

using calibration curves prepared at 10 to 14 levels. Since there was no

available authentic reference standard for BDE-208, the response factor for

BDE-207 was applied with the assumption that the difference in responses

was small.

14

3.3 Quality assurance/Quality control

Isotope labeled internal standards were added before the extraction to in-

crease the accuracy and precision of analytical methods, and authentic stan-

dards were used for the quantification and identification. Injection standards

were added to sample extracts before instrumental analysis to measure the

overall method recovery. All samples were analysed in sets. For each set of

samples, three laboratory (solvent) blanks, three quality control samples

(SRM 2585, house dust reference material, National Institute of Standards

and Technology, Department of Commerce, USA) and 3 reference standards

were also included. Field blanks were analysed as samples. Control charts

with defined warning limits and action limits were established to detect dev-

iation over time using values from the SRM dust. General precautions to

minimize contamination of samples or degradation of analytes are described

in detail in the papers.

3.4 Statistical analyses

For details see the different papers.

15

4. Results and discussion

This thesis offers an overview on the levels of BFRs and PFAAs found in

Swedish indoor microenvironments as well as assesses indoor air as a path-

way for human exposure to BFRs. It also examines the importance of dust

ingestion as a potentially substantial exposure pathway for PFAAs and BFRs

to the Swedish population. Paper I reports on the levels of PFAAs (paper

I) in dust from homes, offices, apartment buildings, cars and day care centers

from Stockholm. Paper I also reports on the estimated daily intake of PFOS

and PFOA from dust ingestion as well as the first indicative values of PFOS

and PFOA in the NIST SRM 2585 house dust.

The levels of BFRs in indoor dust and air from these microenvironments are

presented in Paper II.

In Paper III, concentrations of tri-decaBDEs and total HBCD in paired dust

samples from household vacuum cleaner bags and in researcher-collected,

above-floor settled dust from the same homes were measured and compared

in order to determine if the methods are comparable for exposure studies.

The concentrations of BFRs in dust from both methods were also compared

to the concentrations in matched human milk samples donated by primipar-

ous women, resident in these homes to determine if either or both methods

were linked to body burdens.

Paper IV contributes toward the understanding of the mechanism of how

PBDEs, and BDE-209 in particular, in indoor air reach outdoor air.

4.1 PFAAs in indoor dust

4.1.1 Levels

The average PFAA concentrations measured in the NIST SRM 2585 dust

(paper I) were 1990 ± 78 ng/g dry weight (dw) for PFOS and 673 ± 26 ng/g

dw for PFOA (n=19). These values are in agreement with those later re-

ported by Goosey and Harrad (2011).

PFOS, PFOA, tri-decaBDEs and HBCD were detected in dust samples from

all microenvironments (paper I & II). Detailed information on descriptive

statistics for BFR concentrations in indoor air and dust and PFAA concentra-

tions in indoor dust are provided in paper I & II. The distributions of

PFAAs in dust from all microenvironment were positively skewed with a

few samples having higher concentrations. For PFAAs, the highest variation

16

in concentrations in dust was seen in apartments while houses and day care

centres had much less variability (Figure 4.1).

1

10

100

1000

10000

PF

OS

PF

OA

PF

OS

PF

OA

PF

OS

PF

OA

PF

OS

PF

OA

PF

OS

PF

OA

Apartments Offices Homes Day care Cars

ng/

g d

w

PFOS and PFOA in dust

Figure 4.1. Box-whisker plot (Log scale) of median, 25th and 75th quartiles and range for

PFOS and PFOA from different Swedish microenvironments

The median concentrations in the different microenvironments are within

one order of magnitude of each other. Highest median PFOS concentrations

were seen in offices (110 ng/g dw), and the lowest concentrations were seen

in apartments (19 ng/g dw) and cars (11 ng/g dw). For PFOA, the median

concentrations were more similar between the different microenvironments,

with highest concentrations found in apartments (78 ng/g dw) and offices (70

ng/g dw). Offices had higher median PFOS concentrations than PFOA,

while houses, apartments, day care centers and cars had higher median

PFOA concentrations than PFOS. Offices where large volumes of papers

were present had the highest PFOS and PFOA concentrations. This is not

surprising since PFOS is an essential degradation product from different

fluoropolymers that have been used as surfactants, water and dirt repellents,

electrostatic charge and friction control agents for mixtures used in coatings

applied to papers and printing plates (DEFRA, 2004). PFOA has been used

as a non-reactive polymerization aid in the production of fluorotelomer alco-

hol (FTOH) polymers. FTOH derived polymers have been used to impreg-

nate paper/cardboard to make them grease and water repellant (GreenPeace,

2006).

In recent studies where other microenvironments than homes were studied,

offices have also had the highest concentrations of PFOS but also PFOA in

dust (D'Hollander et al., 2010a; Goosey and Harrad, 2011; Zhang et al.,

2010). The median PFOS and PFOA concentrations in Swedish microenvi-

ronments are compared to those from other countries in Figure 4.2. PFOS

17

concentrations in dust from Swedish homes were similar (houses, 39 ng/g

dw) or higher (apartments, 85 ng/g dw) than those from Japan (Moriwaki et

al., 2003), Kazakhstan, Thailand (Goosey & Harrad, 2011) and Canada

(Kubwabo et al., 2005; Shoeib et al., 2011), higher than in Belgium

(D'Hollander et al., 2010a) and Norway (Haug et al., 2011) but four to five

times lower than those of the UK, Australia, France, Germany (Goosey &

Harrad, 2011) and the USA (Strynar & Lindstrom, 2008).

For PFOA, the concentrations in dust from Swedish homes (houses 54 ng/g

dw, apartments 93 ng/g dw) were similar to those from France and Thailand,

(Goosey & Harrad, 2011) lower than in dust from the UK, Australia, Ger-

many (Goosey & Harrad, 2011), USA (Strynar & Lindstrom, 2008), and

Japan (Moriwaki et al., 2003), but higher than in Belgium (D'Hollander et

al., 2010a), Canada, (Kubwabo et al., 2005; Shoeib et al., 2011) Kazakhstan

(Goosey & Harrad, 2011) and Norway (Haug et al., 2011) (Figure 4.2.). For

Swedish offices, both median PFOS and PFOA concentrations in dust were

in between concentrations found in the UK and Belgium (Figure 4.2.). Me-

dian concentrations of PFOS and PFOA in Swedish day care centers and

cars were lower than those from the UK (Goosey & Harrad, 2011).

0

200

400

600

800

1000

Germ

an

y

Ja

pa

n

US

A

Au

stra

lia

UK

Sw

ed

en

Fra

nce

Ca

na

da

No

rw

ay

Belg

ium

Ka

za

hk

sta

n

Th

ail

an

d

UK

Sw

ed

en

Belg

ium

UK

Sw

ed

en

UK

Sw

ed

en

Homes Offices Cars day care

ng

/g d

w

PFOS

PFOA

Figure 4.2. Comparison of median PFOS and PFOA concentrations in dust from

different microenvironments worldwide.

The geographical variation in PFAA concentrations found amongst the stud-

ies listed above can be due to differences in the usage of products containing

PFAAs, but different dust sampling methods may also play a role. For

example the extensive use of carpeting in the UK and US may be reflected in

the higher concentrations of PFAAs measured in dust. Gerwurtz et al. (2009)

found concentrations of PFAAs to be higher in carpeted homes than in non-

carpeted homes.

18

A statistically significant correlation between PFOS and PFOA concentra-

tions in dust was found for data from all Swedish microenvironments and

also for houses, apartments and offices studied separately (r = 0.607, p <

0.05). This finding is consistent with other studies (D'Hollander et al., 2010a;

Goosey & Harrad, 2011; Haug et al., 2011; Shoeib et al., 2011; Strynar &

Lindstrom, 2008). The significant correlation between PFOS and PFOA may

suggest that these PFAAs come from a common source in the different mi-

croenvironments or may originate from the same precursor compound.

4.2 BFRs in indoor dust and air

4.2.1 Levels

Concentrations of BFRs were positively skewed in both air and dust. There

was variability in both the levels of BFR contamination and the proportions

of the different BFR technical products both within and between the differ-

ent microenvironments. This suggests that the contents of the rooms were a

major source of the BFRs. There were differences between BFR concentra-

tions between car makes and in different cars of the same make (Paper II,

Table S3). No differences in concentrations in air were seen when cars were

standing inside the dealership halls or outside in the sun. The detection fre-

quencies of BFRs in air and dust were high except for BDE-206, which was

detected in only a few dust samples and a few air samples from apartments,

and HBCD, which was detected in only some air samples but in most dust

samples. For all microenvironments, BDE-209 was the most predominant

congener contributing on average, approximately 67% and 78% of the total

PBDE concentrations for air and dust samples, respectively. Median BDE-

209 concentrations in dust were highest in cars (1300 ng/g) and apartments

(1100 ng/g). Day care centers had the highest median ∑PentaBDE (240

ng/g) and HBCD (340 ng/g) concentrations. Offices also had higher median

HBCD (300 ng/g) and the highest median ∑OctaBDE (84 ng/g) (Figure

4.4).

19

1

10

100

1000

10000

100000

1000000

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

Houses Day care Offices Cars Apartments

ng

/g d

wBFRs in dust

0

1

10

100

1000

10000

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

∑PentaBDE

BD

E-2

09

∑PBDE

HB

CD

Houses Day care Offices Cars Apartments

pg

/m3

BFRs in air

Figure 4.4. Box-whisker plot (Log scale) of median, 25

th and 75

th quartiles and

range

for ∑PentaBDE, BDE-209, ∑PBDEs and total HBCD, in air and dust samples.from

different Swedish microenvironments. – means not detected .

Concentrations of BDE-209 in dust from apartments were very positively

skewed (range: <50 - 100000 ng/g dw), due to an extremely high BDE-209

concentration (100000 ng/g dw) in one sample. A few dust samples from

offices also had high BDE-209 concentrations. Webster et al. (2009) have

found elevated concentrations of BDE-209 to be associated with discrete,

highly contaminated particles in dust. They state that inhomogeneous distri-

bution of BDE-209 in dust may lead to high exposure. The high BDE-209

concentration in cars in this study may be due to the fact that samples were

collected directly from car seats and the textile may be a source of this com-

pound or that DecaBDE technical mixture is used extensively in cars.

20

Wilford et al. (2005) indicated that vacuuming close to PBDE-emitting point

sources could lead to exceptionally high values. In air samples for this study,

offices had the highest median ∑PentaBDE (560 pg/m3), ∑OctaBDE (280

pg/m3) and BDE-209 (2000 pg/m

3) concentrations while HBCD was found

in only a few air samples (median range, <1.6 – 2.0 pg/m3). The reason may

be that the sampling volumes used were too small to detect HBCD.

The predominance of BDE-209 in air and dust is in conformity with other

studies (Allen et al., 2008; D'Hollander et al., 2010a; Harrad et al., 2008;

Vorkamp et al., 2011). The median BDE-209 concentrations in offices (pa-

per II) were higher than those reported previously for two offices in Stock-

holm (83 pg/m3) (Sjödin et al., 2001). The median ∑PentaBDE concentra-

tion in dust and BDE-209 concentration in air from homes in this thesis were

similar to those found in homes in Örebro (Karlsson et al., 2007). However,

the median BDE-209 concentration in dust from homes in Stockholm (Pa-

per II) was higher and the median ∑PentaBDE concentration in air was

lower than those found in Örebro homes (Figure 4.5).

0

50

100

150

200

250

300

350

Örebro Stockholm

Med

ian

co

nce

ntr

ati

on

∑PentaBDE dust

BDE 209 dust

∑PentaBDE air

BDE 209 air

Figure 4.5. Concentrations in air (pg/m

3) and dust (ng/g) from homes in Stockholm

(paper II) and Örebro (Karlsson et al. 2007).

Compared with recently published studies, median ∑PentaBDE concentra-

tions in dust (paper II) were within the lower range of concentrations meas-

ured in home, office and car samples from Europe (Figure 4.6) but lower

than those for the US. The results for BDE-209 reported in this thesis were

within the range of concentrations measured in home, office and car samples

from Europe (except for the UK) but lower than in the US (Figure 4.6).

(Cunha et al., 2010; D'Hollander et al., 2010b; Fromme et al., 2009;

Vorkamp et al., 2011).

21

0

2000

4000

6000

8000

Offices

BDE-209

∑Penta

HBCD

0

1000

2000

3000

4000

5000

UK Sweden

Day care centers/schools

0

20000

40000

60000

80000

100000

UK US Portugal Sweden

Cars

Co

nce

ntr

ati

on

(n

g/g

dw

)

0

1000

2000

3000

Homes

Figure 4.6. Comparison of median ∑PentaBDE, BDE-209 and HBCD concentra-

tions in dust from different microenvironments worldwide.

HBCD concentrations in dust from homes and offices reported in Paper II

were similar to those in Belgium (D'Hollander et al., 2010a), but lower than

those reported for homes and public environments in the UK (Abdallah et

al., 2008a; 2008b) and homes in the US and Canada (Abdallah et al.,

2008b). The generally higher levels of PBDEs in dust observed in North

America compared to Europe, as well as the higher BDE-209 and HBCD

levels seen in the UK than other European countries, may be as a result of a

higher use of flame retardants due to the more stringent fire regulations in

these countries.

There is limited data published for BFR concentrations in air. Most of the

few existing studies have used passive sampling techniques which do not

sample particle-bound BDEs such as BDE-209. Thus, results from paper II

help to increase the data base on BDE-209 in air from homes. BDE-209 has

not been measured in air from offices and day care centers previously, so

results from this thesis are the first to show levels of BDE-209 in air from

these microenvironments. The median BDE-209 concentration in air meas-

ured in homes and apartments (paper II) were higher than those measured in

the US (Allen et al., 2007), Japan (Takigami et al., 2009), Denmark

(Vorkamp et al., 2011) and Germany (Fromme et al., 2009). Median BDE-

209 concentrations in air from cars (paper II) were higher than in cars from

Greece (Mandalakis et al., 2008).

22

The median ∑PentaBDE concentrations in air from Swedish homes (paper

II) fall within the range of reported data for ∑PentaBDE in indoor air from

other European studies but are lower than concentrations measured in the

US (Figure 4.7) (Fromme et al., 2009; Harrad et al., 2006; Johnson-Restrepo

& Kannan, 2009; Shoeib et al., 2004; Vorkamp et al., 2011). For offices, the

median concentration of ∑PentaBDE in this study was higher than found in

the UK but lower than those measured in the USA (Batterman et al., 2010;

Harrad et al., 2006).

0

200

400

600

800

1000

US

Sw

eden

UK

US

Ca

na

da

Den

ma

rk

UK

Ger

ma

ny

Sw

eden

Offices Homes

pg

/m3

∑Penta

Figure 4.7. Comparison of median ∑PentaBDE concentrations in air from homes

and offices from relevant studies worldwide.

4.2.2 Estimated human exposure to PFAAs and BFRs

To assess the impact of the indoor environment on human exposure to

PFAAs and BFRs, two exposure scenarios were used based on mean and

high dust ingestion. Details on assumptions and the total estimates for adult

and toddler exposures to PFOS and PFOA via dust ingestion are summarized

in paper I. Estimated total intakes of PFAAs were calculated using mean

and high dust ingestion rates of 4.16 and 50 mg/day for adults and 20 and

100 mg/day for toddlers (12-24 months of age) (USEPA, 2002). Updated

mean ingestion rates of 50 (adult) and 60 mg/day (toddler) and inhalation

rates of 16 (adult) and 8 m3/day (toddler) were used for BFRs (Paper II)

(USEPA, 2008; USEPA, 2009). Total BFR ingestion was also estimated in

relation to the estimated fraction of time spent in the studied microenviron-

ments based on data from the UK (Harrad et al., 2006) (paper II; (de Wit et

al., 2011).

23

Based on these estimates of dust ingestion and inhalation, the major expo-

sures for PFOS, PFOA, ∑PentaBDE and ∑DecaBDE using median concen-

trations were found to occur from dust ingestion from homes (Figure 4.8).

For ∑OctaBDE, dust ingestion from homes are the most important for tod-

dlers, but for adults, the major exposure was from dust ingestion and inhala-

tion in offices. HBCD exposure from dust ingestion was slightly higher from

offices/day care centers than homes. Toddlers were found to have higher

ingestion of PFOS, PFOA, ∑PentaBDE, ∑DecaBDE and HBCD from dust

than adults in all scenarios.

0

2

4

6

8

10

Adultinhalation

Adult dustingestion

Toddlerinhalation

Toddler dustingestion

PentaBDE

0

0,4

0,8

1,2

1,6

Adultinhalation

Adult dustingestion

Toddlerinhalation

Toddler dustingestion

OctaBDE

0

10

20

30

40

50

Adultinhalation

Adult dustingestion

Toddlerinhalation

Toddlerdust

ingestion

DecaBDE

0

2

4

6

8

10

Adultinhalation

Adult dustingestion

Toddlerinhalation

Toddlerdust

ingestion

HBCD

Inta

ke

(ng

/da

y) 0

2

4

6

Adults dustingestion

Toddler dustingestion

PFOS

0

2

4

6

8

10

12

Adults dustingestion

Toddler dustingestion

PFOA

Cars

Offices/day care

Homes

Figure 4.8. Mean estimated intakes (ng/day) of PFAAs and BFRs for adults and

toddlers from air inhalation or dust ingestion from different microenvironments

based on the estimated fraction of time spent in each and median concentrations in

air and dust (paper I and II).

The estimated intakes of PFOS and PFOA through dust ingestion for this

study (0.3 ng/d for adults and 3 ng/d for toddlers) were similar to those

found in the UK (Goosey & Harrad, 2011), and Norway (Haug et al., 2011)

but lower than for Belgium (D'Hollander et al., 2010a).

The estimated intakes for ∑PentaBDE, ∑DecaBDE and HBCD for adults

and toddlers from mean dust ingestion and inhalation are compared to those

reported for Belgium, the US and the UK (Table 1) (Abdallah et al., 2008a;

Allen et al., 2007; Harrad et al., 2008; Harrad et al., 2006; Johnson-Restrepo

& Kannan, 2009; Roosens et al., 2010a).

24

Table 1. Estimated daily intake (ng/day) of ∑PentaBDE, ∑DecaBDE and HBCD via mean dust ingestion and inhalation by country.

Dust ingestion Inhalation

Country PBDE congener Adults Toddlers Adults Toddlers

Sweden ∑PentaBDE 5.8 7.8 2.3 0.39

∑DecaBDE 38 43

HBCD 6 7.6

Belgium ∑PentaBDE 0.35 1.2 0.17 0.98

∑DecaBDE 2.5 18 0.11 0.06

HBCD 3.7 8.1 0.17 0.98

UK ∑PentaBDE 1.3 2.6 0.82 0.16

∑DecaBDE 230 610

HBCD 15 37 3.9 0.8

US ∑PentaBDE 44 80 5.6 3.2

∑DecaBDE 41 70 3.5 2

HBCD

There are no published studies of the dietary intake of PFAAs in Sweden.

Comparing air and dust intake estimates using median concentrations from

Paper I and II and dietary estimates for PFAAs from Holland (Noorlander

et al., 2011) and Swedish market basket surveys for ∑PentaBDE (Darnerud

et al., 2006) and HBCD (Tornkvist et al., 2011), dust ingestion accounted for

1, 2, 11 and 37% of adult intake for PFOS, PFOA, ∑PentaBDE and HBCD,

respectively, and 9, 28, 23 and 57% of toddlers total intake of PFOS, PFOA,

∑PentaBDE and HBCD respectively (Table 2). In comparison, Egeghy and

Lorber (Egeghy & Lorber, 2011) recently estimated that dust contributes as

much as diet for PFAA exposure in 2-year-olds in the US. When maximum

air and dust concentrations from this thesis were used, dust ingestion ac-

counted for up to 71, 82, 77 and 95% of total toddler intake for PFOS,

PFOA, ∑PentaBDE and HBCD, respectively. Similar results were found in

paper I using Spanish (Ericson et al., 2008) and Canadian (Tittlemier et al.,

2007) dietary intake estimates for PFOS and PFOA.

These estimates of the contributions to total intake of PFAAs and BFRs from

dust ingestion are limited by the uncertainty in dust ingestion rates and me-

thods for measuring dust ingestion. In order to determine which of the expo-

sure scenarios is the most relevant, studies of body burdens of BFRs and

PFAAs in adults and in toddlers in conjunction with dust sampling are

25

needed as well as studies of the dietary intake of PFAAs from foodstuffs in

Sweden. Also, the contribution of drinking water (Vestergren et al., 2008)

and inhalation to total PFAA exposure should be assessed. A recent study

from Canada found inhalation to be a more important exposure pathway for

PFOS and PFOA in adults (Shoeib et al., 2011).

Table 2. Estimates of exposure (ng/day) for adults and toddlers to ∑PentaBDE, HBCD, PFOS and PFOA via the diet (Darnerud et al., 2006; Noorlander et al., 2011; Tornkvist et al., 2011), dust ingestion (mean dust ingestion rates used) and inhala-tion as well as the relative significance of each pathway.

Adult Toddler

Median Maximum Median Maximum

∑PentaBDE House Apt House Apt House Apt House Apt

Air 2.4 2.2 5.8 6.5 0.39 0.39 1.8 1.8

Dust 5.8 5.8 27 74 7.8 7.8 36 92

Food 44

52

44 44 44 25 25 25 25

Total 52 77 124 33 33 63 118

% contribution

Air 4.6 4.3 7.6 5.2 1.2 1.2 2.8 1.5 Dust 11 11 35 59 23 23 57 77

Food 84 85 57 35 75 75 40 21

HBCD

Air 0 0 0.35 0.14 0 0 0 0

Dust 7.0 5.0 97 150 8.8 6.4 64 120

Food 10 10 10 10 5.7 5.7 5.7 5.7

Total 17 15 108 160 15 12 70 126

% contribution

Air 0 0 0.32 0.088 0 0 0.38 0.10

Dust 41 33 90.4 93.5 61 53 91.4 95.4

Food 59 67 9.3 6.4 39 47 8.2 4.5

PFOS

Dust 0.3 0.4 0.8 3.1 1.8 1.2 5 35

Food 19.2 19.2 19.2 19.2 15 15 15 15 Total 19

20

20

22

16 16

20

50 % contribution

Dust 1.3 2.0 4.0 14 11 7.4 26 71

Food 99 98 96 86 89 93 74 29

PFOA

Dust 0.2 0.3 0.8 3.1 2.6 3.5 5 36

Food 17 17 17 17 7.9 7.9 7.9 7.9 Total 17

17

18

20

11

11

13

44

% contribution

Dust 1.2 1.6 4.2 15 25 31 39 82

Food 99 98 96 85 75 69 61 18

26

4.2.3 Tolerable daily intakes

Using tolerable daily intakes (TDI) established by the European Food Agen-

cy (EFSA) for PFOS (150 ng/kg bw/ day) and PFOA (1500 ng/kg bw/ day)

(EFSA, 2008) and an average body weight of 70 kg for an adult and 10 kg

for a toddler, a tolerable daily intake for this study was estimated. The max-

imum daily intakes of PFOS and PFOA from dust for adults (0.028 ng/kg

bw/ day) and toddlers (2 ng/kg bw/ day) are well below the tolerable daily

intakes. No health based standards exist for Europe for BFRs. Median

∑PentaBDE and ∑DecaBDE (BDE-209 is approximately 70-80% of

∑DecaBDE) doses calculated for adults (0.08 and 0.54 ng/kg/day) and tod-

dlers (0.78 and 4.3 ng/kg/day) were also well below the references doses

(RfDs) for BDE-47 (100 ng/kg/day), BDE-99 (100 ng/kg/day), BDE-153

(200 ng/kg/day) and BDE-209 (7000 ng/kg/day) determined by the US EPA

(IRIS, 2007). The median adult and toddler intakes of HBCD (0.086 and

0.76 ng/kg/day) are also below the US National Research Council’s RfD of

200 ng/kg/day (NRC, 2000). The doses calculated using maximum intakes

were also still below levels of concern.

4.2.4 BFR levels in breast milk

Detailed information on the concentrations, ranges and detection frequencies

of BFRs in breast milk are given in paper III. Only tri-heptaBDEs and

HBCD were measured. The predominant congeners found were BDE-47 and

BDE-153. The results from this study were similar to those found in other

Swedish studies (Fängström et al., 2008; Glynn et al., 2011; Lignell et al.,

2009). Thus, BFR exposures in this study are likely representative of expo-

sures in urban areas in Sweden. The concentration of BDE-47 and -153

measured in breast milk from this study were similar to those measured in

other European countries (Frederiksen et al., 2009) and Ghana (Asante et al.,

2011), a little lower than those measured in the UK (Kalantzi et al., 2004)

and Australia (Toms et al., 2009a) but almost ten times lower than those

measured in the US (Dunn et al., 2010; Schecter et al., 2010a). The concen-

trations of HBCD are similar to those reported for human milk samples from

the USA (Ryan et al., 2006), Ghana (Asante et al., 2011) and Belgium

(Roosens et al., 2010b) but lower than those reported from Norway

(Thomsen et al., 2010), Canada (Ryan et al., 2006) and the UK (Abdallah

and Harrad, 2011). These reported differences in BFR concentrations may be

from variations in personal exposure of the participants in the different stu-

dies (Abdallah & Harrad, 2011).

4.2.5 Comparison of dust sampling methods

The median concentrations of PBDEs in researcher collected above-floor

settled dust were statistically significantly (p<0.001-0.05) higher than in

27

vacuum cleaner bag dust. This is consistent with what Allen et al. (Allen et

al., 2008) observed when comparing vacuum cleaner bag dust and researcher

collected floor dust. In contrast, the median concentrations of total HBCD

were 2 to over 1000 times higher in dust from the vacuum cleaner bags

compared to researcher collected above-floor settled dust (paper III). Con-

centrations of BFRs in dust from Uppsala homes (paper III) were similar to

those found in Stockholm (paper II). Despite the differences in concentra-

tion between the two sampling methods, statistically significant correlations

for ∑OctaBDE (r = 0.595, p < 0.05), ∑DecaBDE (r = 0.649, p < 0.05) and

HBCD (r = 0.613, p < 0.05) but not for ∑PentaBDE (r = 0.382, p > 0.05)

where found between matched pairs of researcher-collected above floor dust

and vacuum cleaner bag floor dust. The correlation for HBCD (r = 0.237, p

> 0.05) and ∑DecaBDE were influenced by one high value and were not

significant when this was removed. These correlations suggest that ∑Octa-

and ∑DecaBDE in dust sampled by both sampling methods originated from

flame retarded items containing these PBDE technical mixtures within the

home. The lack of correlation for ∑PentaBDE and HBCD may be due to the

limited sample size.

4.2.6 Association between breast milk and dust levels

To determine which dust sampling method best relates to body burden, the

concentrations of BFRs in dust from both methods were also compared to

the concentrations in matched human milk samples donated by primiparous

women, resident in these homes. Due to the low detection frequencies of the

other PBDE congeners and HBCD, a correlation could only be performed

for BDE-47 and -153.

A statistically significant positive correlation was seen only for BDE-47

concentrations in matched breast milk samples and vacuum cleaner bag dust

(r = 0.514, p = 0.029). These results may indicate that one possible route of

exposure to BDE-47 is via dust ingestion. No associations were observed

between researcher-collected above floor settled dust and breast milk sam-

ples for BDE-47, or between dust from either method and BDE-153 concen-

trations in breast milk. The lack of correlations between BDE-153 concentra-

tions in dust from both methods and breast milk as well as between BDE-47

in AFSD and breast milk may be due to the limited sample size of this study

or that diet may be more important for BDE-153 exposure. Other possible

explanations could be that dust ingestion was not a significant vector of ex-

posure for these women or that the dust sampled was not representative of

the dust they had ingested over a long time period that would be needed to

be reflected in breast milk concentrations.

28

4.2.7 Levels of PBDEs in indoor and outgoing air

Tri-decaBDEs were detected in all indoor and outgoing air samples from the

different microenvironments (paper IV). The ∑10PBDEs (the sum of BDE-

28, -47, -99, -153, -183, -197, -206, -207, -208, -209) concentrations for

indoor air ranged from 18 – 7300 pg/m3 and 17 to 8500 pg/m

3 for outgoing

air for all microenvironments (paper IV). The most predominant PBDE

congener was BDE-209 which constituted up to 80% of the total BDEs for

indoor air and 51% for outgoing air. Of the indoor microenvironments stud-

ied, indoor and outgoing air from offices had the highest median

∑PentaBDE and BDE-209 concentrations (paper IV, Table 2). Comparison

of matched indoor air/outgoing air samples from the same building showed

no statistically significant differences in individual BDE congener concen-

trations except for BDE-28 (p = 0.028) in apartments and BDE-183 (p =

0.040) in day care centers. Also, statistically significant correlations (r =

0.358 – 0.729, p < 0.05) were seen between indoor air and outgoing air con-

centrations for all congeners except BDE-28 (r =0.240, p = 0.178), indicating

a common source.

The median concentrations of ∑PentaBDE (35-630 pg/m3) and BDE-209

(22-1900 pg/m3) in outgoing air from the three types of microenvironments

are up to several orders of magnitude higher than seen in background out-

door air from Sweden (∑PentaBDE 1.6-3.7 pg/m3; BDE-209 6.1-6.5 pg/m

3)

(Agrell et al., 2004; Jaward et al., 2004; Ter Schure et al., 2004).

4.2.8 Transport of BDE-209 to ambient air

At room temperature 80% of BDE-47 and 10-40% of penta-heptaBDEs are

expected to be in the gas phase while BDE-209 will predominantly be found

in the particulate phase (Shoeib et al., 2004). The most predominant conge-

ner in both indoor and outgoing air, with no significant difference in concen-

tration, was BDE-209. This indicates that particle-bound contaminants are

also transported through ventilation systems to ambient air. Alternatively, a

portion of BDE-209 may be bound to very fine particles that are able to pass

through filters and/or be transported through the ventilation systems and/or it

is present in the gas phase in indoor air. The results from outgoing air to-

gether with evidence of lower PBDE concentrations in outdoor air support

the hypothesis that PBDEs (including BDE-209) in the indoor air enter ven-

tilation systems, are vented unchanged to the outdoors where they later un-

dergo dilution through atmospheric mixing leading to the lower concentra-

tions found outdoors than indoors. Thus ventilation is a likely conduit of

PBDEs from indoor sources to the outdoors.

29

4.2.9 Significance of ∑PentaBDE and BDE-209 emissions to ambient air

The calculated emission rates from buildings from Paper IV were 0.013-2.0

ng/h/m2 and 0.0088-5.0 ng/h/m

2 for ∑PentaBDE and BDE-209 respectively

(paper IV).This calculated emission rate for ∑PentaBDE in (paper IV) is

similar to 1 and 7 ng/h/m2 estimated for an office in the UK, but an order of

magnitude lower than those estimated for US homes (20 ng/h/m2), a new

office building (22 ng/h/m2) and a Canadian office (6-40 ng/h/m

2)

(Batterman et al., 2009; Batterman et al., 2010; Zhang et al., 2009; Zhang et

al., 2011). The higher emission rates in North America probably reflect the

higher use of technical PentaBDE products there. To the best of our knowl-

edge this is the first data for BDE-209 emission rates to outdoor air. Total

emissions of ∑PentaBDE and BDE-209 to outdoor air from all sources, in-

cluding metal and plastics manufacturing, waste incineration, electronics

recycling, e-waste and landfill fires and indoor air, were estimated (paper

IV). The estimated emissions from indoor air accounted for 50-93% of the

total emissions of ∑PentaBDE and 25-86% of BDE-209 to the outdoor air.

These results thus support the previous hypothesis that indoor air is a signifi-

cant source of PBDEs to outdoor air which may eventually lead to contami-

nation of food and dietary exposure (Harrad & Diamond, 2006).

30

5. Conclusions

PFOS, PFOA, PBDEs and HBCD were found in dust from all of the micro-

environments studied. PBDEs were found in air from the different microen-

vironments but HBCD was detected in only a few air sam-

ples.Concentrations of PFOS and ∑OctaBDE in office dust were significant-

ly higher (p<0.05) than in the other microenvironments while ∑PentaBDE

and HBCD were significantly higher in offices and day care centers com-

pared to the other microenvironments studied. Car dust had higher median

BDE-209 concentrations than homes, apartments and day care centers. Sig-

nificantly higher concentrations of tri-decaBDEs were detected in air from

offices compared to homes, daycare centers and cars, while BDE-209 con-

centrations in cars were significantly higher than in other microenviron-

ments. The presence of congeners from the PentaBDE and OctaBDE tech-

nical mixtures, which have been banned since 2004, in air and dust samples

confirms that Swedish indoor environments still contain flame-retarded

products that are reservoirs for these POPs. BDE-209 was the most predo-

minant congener in air and dust reflecting its continued usage.

Concentrations of PBDEs from homes in this study are in line with those

earlier reported for Sweden and other European countries other than the UK.

Compared to total exposure from diet, inhalation and dust ingestion are mi-

nor pathways for exposure, but in worst case scenarios dust ingestion may be

a dominant pathway.

The major exposures from indoor microenvironments for adults and toddlers

to PFAAs and PBDEs occur from dust ingestion in homes with inhalation

playing a minor role. The exception was for ∑OctaBDE for which inhalation

from offices was also important for adults’ exposure. Major exposure to

HBCD occurred in offices/day care centers.

Toddlers have higher estimated intakes of all the studied PFAAs and BFRs

from dust ingestion than adults. The estimated doses from dust ingestion

compared to the TDI for PFAAs and RfDs for BFRs from this study are

below levels of concern. The estimated contributions to total intake from

inhalation and dust ingestion are limited by the uncertainty in dust ingestion

rates. In order to determine which of the exposure scenarios is the most rele-

vant, body burdens of PFAAs and BFRs in adults and in toddlers in conjunc-

tion with air and dust sampling should be studied.

Vacuum cleaner bag dust and above floor settled dust correlated signifi-

cantly for ∑Octa-, and ∑DecaBDE, suggesting that both methods may be

relevant for studying these contaminants but it is still unclear as to which

method is best for studying ∑PentaBDE and HBCD. A correlation was seen

31

between concentrations of BDE-47 and vacuum cleaner bag dust indicating

that the indoor (dust ingestion) pathway may play a role in exposure. How-

ever, no firm conclusions could be drawn from this study due to the limited

sample size and the low detection of many PBDE congeners and HBCD.

PBDEs were detected in all outgoing air samples, with BDE-209 constituting

an average of 31-51% of the total BDE concentration from the different mi-

croenvironments. No statistically significant differences were seen for con-

centrations of PBDEs in indoor and outgoing air. There was a significant

correlation between concentrations of BDEs in indoor and outgoing air con-

firming the indoors as a source of PBDEs to the outdoors. The estimated

emissions of ∑PentaBDE and BDE-209 contributed significantly to outdoor

concentrations in Sweden.

5.1 Knowledge gaps and future perspectives.

Elucidate the pathways through which less volatile congeners like

BDE-209 migrate from treated products to air and dust and a better

understanding of the partitioning of BDE-209 in the indoor envi-

ronment.

Improve analytical methods to better quantify higher brominated

PBDEs in breast milk and to continuously monitor replacement

products for the banned PBDEs in indoor environments and the con-

sequent exposure from such products.

Improve knowledge of the potential adverse effects of BFRs on human

health to enable the determination of a tolerable daily intake (TDI)

for these compounds.

Further studies to determine if vacuum cleaners could be the source of

elevated HBCD in vacuum cleaner bag dust.

Better understanding of the different dust sampling methods in rela-

tionship to body burden.

32

6. Acknowledgements

I would like to return thanks and appreciation to my supervisor; Prof. Cyn-

thia de Wit for believing in me and whose scientific advice, knowledge and

encouragement has helped me in completing this project especially through

the highs and lows of this project. Thanks for linking me into an amazing

network of researchers on indoor POPS. I have benefited a lot from these

relations.

I would also like to thank my co-supervisor Ulla Sellström and Kaj Thu-

resson and Urs Berger for the valuable analytical knowledge they shared

with me. Kaj I am grateful for all the moments we had in the lab, the de-

bates, daily life stories and laughter will never be forgotten, you are a true

Team player. Gratitude is also expressed to all those who aided in the sample

collection, a lot of which would have been unattainable without their help:

Karin S, Thorvald, Ulrika Friden and Caroline. Huge thanks to Micke

and Ulla E for doing their best in keeping the instruments intact and to

Michael McLachlan for reading and commenting on this thesis.

Thanks to my co-authors Per Ola Danerud, Sanna Lignell, Marie Aune

and Anna Palm Cousins for your important contributions, good discussions

and helpful suggestions.

Thanks to all my Colleagues at Mo for all the good memories, fun and

friendships made and to all those who passed by room V507 for idle chat,

intellectual conversations and impassioned discussions.

I wish to express my love and thanks to my family for their indomitable en-

couragement and support from the day I was born to this moment. My par-

ents and siblings (Emmanuel and Margaret) have been an ancillary in the

achievement of this PhD, with their immutable love and belief in me, enabl-

ing me to achieve my aspirations in life. Thank you, Mrs. Gwendoline

Burnley and Joseph Che for all the support and for always reminding me

that even the largest task can be accomplished if it is done one step at a time.

Thanks to all my friends, the Björklunds and Tengberts for helping me

see life through another perspective than research. Tack så hemsk mycket

Lars och Ulla-Britt Björklund för all hjälp med att få livet att gå ihop.

Tack för att ni trots sporadisk kontakt alltid ställer up som barnvakt.

Utmost, for my children Hugo, Percy and Eposi I wish to express my grati-

tude for your love and all the crazy things you do that put smiles on my face

every day, you are all so precious. Tack Magnus, för att du finns alltid där

för mig , och stödjar mig till 100%. Tack för ert tålamod under den hektiska

slutfasen.

33

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