Brominated flame retardants and perfluoroalkyl acids in Swedish indoor microenvironments

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
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
ii
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.
iii
iv
Abstract
Humans are exposed to persistent organic pollutants (POPs) such as brominated flame retardants (BFRs, specifically polybrominated diphenyl ethers
(PBDEs) and hexabromocyclododecane (HBCD)) and perfluoroalkyl acids
(PFAAs, specifically perfluoroalkane sulfonate (PFOS) and perfluorooctanoic acid (PFOA)). They are used in consumer products found in cars, offices, homes and day care centers. Diet was earlier thought to be a major human 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 estimate human exposure. BFRs and PFAAs were detected in dust from all microenvironments 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 scenarios, dust ingestion may be significant for a small part of the Swedish
population. Sampling using home vacuum cleaner bag dust and researchercollected 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 concentrations 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 polluting ambient air via ventilation systems.
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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.
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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) Tridecabrominated 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.; Lignell,S.; Darnerud, P. O. (2011) Comparisons of PBDE and HBCD concentrations 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 tridecabrominated 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 responsible for data acquisition, interpreting the data and took the lead in writing the paper.
Paper II: I was involved in extraction and analysis of the samples, data acquisition 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 comments 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 responsibility for writing the paper with the exception of the modeling section.
vii
Abbreviations
AFSD
BDE-209
BFR
BSEF
ECNI
EFSA
EPA
EPS
GC
HBCD
HIPS
HPLC
Kow
LOD
LOQ
PBDE
PFAAs
PFOS
PFOA
POPs
RfD
SVOC
TDI
VCBD
viii
Above floor settled dust
Decabromodiphenyl ether
Brominated flame retardant
Bromine Science and Environmental Forum
Electron capture negative ionisation
European Food Standard Agency
Environmental Protection Agency
Expanded polystyrene
Gas chromatography
Hexabromocyclododecane
High impact polystyrenes.
High performance liquid chromatography
n-octanol/water partition coefficient.
Limit of detection
Limit of quantification
Polybrominated diphenyl ethers
Perfluoroalkyl acids
Perfluoroalkane sulfonate
Perfluorooctanoic acid
Persistent organic pollutants
Reference dose
Semivolatile organic compounds
Tolerable daily intake
Vacuum cleaner bag dust
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
ix
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
x
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 pathways 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 measured in higher concentrations indoors than outdoors. Because of their slow
rate of release from sources and their propensity to partition and sorb to surfaces, 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.
PFOA, n = 6
PFOS, n = 7
F
F
F
O
S
F
n
F
F
F
O
O
C
F
n
O
F
O
F
Figure 1.1. Structures of PFOS and PFOA.
1
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 cookware, Teflon®, GORE-TEX®, waterproof clothing, fast food wrappers, pizza 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 fluorochemicals that have been produced for over 40 years by electrochemical fluorination (Olsen et al., 2005). POSF was used to produce surfactants and fluorinated side chain polymers which were used in paper and packaging treatments, 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 sulfonamidoalcohol (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 classified PFOS as very persistent, very bioaccumulative and toxic, the European
Union has restricted its use and is also considering the risks of PFOA exposure (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 humans are still not fully understood. However, some animal studies provide
some indications that PFOA exposure causes liver toxicity, carcinogenicity,
2
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 thyroid 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 menstrual 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 increased 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 coatings 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 subgroup 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 combustion 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 electrical 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 products to which they are added (Hutzinger and Thoma, 1987; Stapleton et al.,
2008).
3
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 configurations 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 technical 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). OctaBDE 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 minor extent, BDE-197 and -203. The DecaBDE mixture consists predominantly of BDE-209 (98%) and nonaBDEs (2%) (La Guardia et al., 2006;
McDonald, 2002) .
Br
Br
PBDE
HBCD
O
Br
Br1-10
Br
Br
Br
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 stereoisomers, 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).
4
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 concentrations 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 classical 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 production 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 Chemicals). The goal of this program is to protect human health and environment
from the risks posed by chemicals. HBCD is currently proposed to be reviewed 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 abnormalities, and possibly cancer (Birnbaum & Staskal, 2004; Costa and Giordano,
5
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). HeptadecaBDEs 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 tissue 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). Concentrations 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
6
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 important 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 cultural differences (Frederiksen et al., 2009). Age may also play a role as measured 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).
7
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 commonly frequented indoor microenvironments like cars, offices and day care centers were lacking.
Swedish citizens spend a disproportionate fraction of time indoors (typically
in excess of 90%) where these compounds are predominantly used and recent 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 correlated 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 compared 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).
8
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 mechanism 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 include apartments, offices, day care centers and cars in Sweden. This included:
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 collected 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 environment.
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.
9
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 different 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 (paper 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 dealership 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 collected.
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. Although the passive sampler has the advantage that it is easy to use and deploy (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 monitoring 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.
10
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 sampled 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 OctoberDecember 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, fragments of fibers and hair, insect remains, dandruff, sand, pollen grains and
fungal spores (Lioy et al., 2002). These particles can sorb semivolatile or11
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 measure 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 complete. 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.
12
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 18O-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 3,5bis(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 extracted together. For BFRs, dust and air (PUFs and filters) samples were
spiked with surrogate standards (13C-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 monitoring (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 monitoring (SIM). The ions monitored were m/z 237 and 239 for the surrogate
standard Dechlorane and 494.3 and 496.3 for 13C-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.
13
3.3 Quality assurance/Quality control
Isotope labeled internal standards were added before the extraction to increase the accuracy and precision of analytical methods, and authentic standards 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 deviation 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.
14
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 pathway 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 primiparous 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 reported 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 concentrations 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
15
in concentrations in dust was seen in apartments while houses and day care
centres had much less variability (Figure 4.1).
PFOS and PFOA in dust
10000
ng/g dw
1000
100
10
Apartments
Offices
Homes
Day care
PFOA
PFOS
PFOA
PFOS
PFOA
PFOS
PFOA
PFOS
PFOA
PFOS
1
Cars
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 alcohol (FTOH) polymers. FTOH derived polymers have been used to impregnate 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 microenvironments are compared to those from other countries in Figure 4.2. PFOS
16
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, Germany (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.). Median concentrations of PFOS and PFOA in Swedish day care centers and
cars were lower than those from the UK (Goosey & Harrad, 2011).
PFOS
800
PFOA
ng/g dw
1000
600
400
Homes
Offices
Cars
Sweden
UK
Sweden
UK
Belgium
Sweden
UK
Thailand
Belgium
Kazahkstan
Norway
Canada
France
Sweden
UK
Australia
USA
Japan
0
Germany
200
day care
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 studies 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 noncarpeted homes.
17
A statistically significant correlation between PFOS and PFOA concentrations 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 microenvironments 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 different microenvironments. This suggests that the contents of the rooms were a
major source of the BFRs. There were differences between BFR concentrations 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 frequencies 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 BDE209 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).
18
Houses
Day care
Offices
∑PBDE
HBCD
BDE-209
HBCD
Cars
∑PentaBDE
∑PBDE
BDE-209
HBCD
∑PentaBDE
∑PBDE
BDE-209
∑PentaBDE
HBCD
∑PBDE
∑PentaBDE
BDE-209
HBCD
∑PBDE
BDE-209
∑PentaBDE
ng/g dw
BFRs in dust
1000000
100000
10000
1000
100
10
1
Apartments
BFRs in air
10000
pg/m3
1000
100
10
Houses
Day care
Offices
Cars
HBCD
∑PBDE
BDE-209
HBCD
∑PentaBDE
∑PBDE
BDE-209
∑PentaBDE
HBCD
∑PBDE
BDE-209
∑PentaBDE
HBCD
∑PBDE
BDE-209
∑PentaBDE
HBCD
∑PBDE
BDE-209
0
∑PentaBDE
1
Apartments
Figure 4.4. Box-whisker plot (Log scale) of median, 25 th and 75th 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 distribution 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 compound or that DecaBDE technical mixture is used extensively in cars.
19
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/m3) 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 (paper II) were higher than those reported previously for two offices in Stockholm (83 pg/m3) (Sjödin et al., 2001). The median ∑PentaBDE concentration 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 (Paper II) was higher and the median ∑PentaBDE concentration in air was
lower than those found in Örebro homes (Figure 4.5).
Median concentration
350
300
250
∑PentaBDE dust
200
BDE 209 dust
150
∑PentaBDE air
100
BDE 209 air
50
0
Örebro
Stockholm
3
Figure 4.5. Concentrations in air (pg/m ) and dust (ng/g) from homes in Stockholm
(paper II) and Örebro (Karlsson et al. 2007).
Compared with recently published studies, median ∑PentaBDE concentrations in dust (paper II) were within the lower range of concentrations measured 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).
20
Offices
Day care centers/schools
8000
5000
BDE-209
HBCD
3000
Concentration (ng/g dw)
∑Penta
6000
4000
4000
2000
2000
1000
0
0
UK
Sweden
Homes
Cars
100000
3000
80000
2000
60000
40000
1000
20000
0
0
UK
US
Portugal
Sweden
Figure 4.6. Comparison of median ∑PentaBDE, BDE-209 and HBCD concentrations 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 measured 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 BDE209 concentrations in air from cars (paper II) were higher than in cars from
Greece (Mandalakis et al., 2008).
21
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).
∑Penta
1000
pg/m3
800
600
400
Offices
Sweden
Germany
UK
Denmark
Canada
US
UK
Sweden
0
US
200
Homes
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 microenvironments based on data from the UK (Harrad et al., 2006) (paper II; (de Wit et
al., 2011).
22
Based on these estimates of dust ingestion and inhalation, the major exposures for PFOS, PFOA, ∑PentaBDE and ∑DecaBDE using median concentrations were found to occur from dust ingestion from homes (Figure 4.8).
For ∑OctaBDE, dust ingestion from homes are the most important for toddlers, but for adults, the major exposure was from dust ingestion and inhalation 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.
10
PentaBDE
50
8
6
DecaBDE
PFOS
40
6
4
30
4
20
Intake (ng/day)
2
2
10
0
0
0
Adult
inhalation
Adult dust
ingestion
Toddler Toddler dust
inhalation
ingestion
Adult
Adult dust Toddler
inhalation ingestion inhalation
Adults dust
ingestion
Toddler
dust
ingestion
12
1,6
OctaBDE
10
1,2
HBCD
10
Toddler dust
ingestion
PFOA
Cars
Offices/day care
8
8
Homes
0,8
6
6
4
4
2
2
0,4
0
Adult
inhalation
Adult dust
ingestion
Toddler Toddler dust
inhalation ingestion
0
0
Adult
Adult dust Toddler
inhalation ingestion inhalation
Toddler
dust
ingestion
Adults dust
ingestion
Toddler dust
ingestion
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).
23
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
∑DecaBDE
HBCD
5.8
38
6
7.8
43
7.6
2.3
0.39
Belgium
∑PentaBDE
∑DecaBDE
HBCD
0.35
2.5
3.7
1.2
18
8.1
0.17
0.11
0.17
0.98
0.06
0.98
UK
∑PentaBDE
∑DecaBDE
HBCD
1.3
230
15
2.6
610
37
0.82
0.16
3.9
0.8
∑PentaBDE
∑DecaBDE
HBCD
44
41
80
70
5.6
3.5
3.2
2
US
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 accounted 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 methods for measuring dust ingestion. In order to determine which of the exposure scenarios is the most relevant, studies of body burdens of BFRs and
PFAAs in adults and in toddlers in conjunction with dust sampling are
24
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 inhalation as well as the relative significance of each pathway.
∑PentaBDE
Air
Dust
Food
Total
% contribution
Air
Dust
Food
HBCD
Air
Dust
Food
Total
% contribution
Air
Dust
Food
PFOS
Dust
Food
Total
% contribution
Dust
Food
PFOA
Dust
Food
Total
% contribution
Dust
Food
Adult
Median
House
2.4
5.8
44
52
Apt
2.2
5.8
44
52
Maximum
House
5.8
27
44
77
Apt
6.5
74
44
124
4.6
11
84
4.3
11
85
7.6
35
57
5.2
59
35
0
7.0
10
17
0
5.0
10
15
0.35
97
10
108
0
41
59
0
33
67
0.3
19.2
19
Toddler
Median
House
0.39
7.8
25
33
Apt
0.39
7.8
25
33
Maximum
House
1.8
36
25
63
Apt
1.8
92
25
118
1.2
23
75
1.2
23
75
2.8
57
40
1.5
77
21
0.14
150
10
160
0
8.8
5.7
15
0
6.4
5.7
12
0
64
5.7
70
0
120
5.7
126
0.32
90.4
9.3
0.088
93.5
6.4
0
61
39
0
53
47
0.38
91.4
8.2
0.10
95.4
4.5
0.4
19.2
20
0.8
19.2
20
3.1
19.2
22
1.8
15
16
1.2
15
16
5
15
20
35
15
50
1.3
99
2.0
98
4.0
96
14
86
89
7.4
93
26
74
71
29
0.2
17
17
0.3
17
17
0.8
17
18
3.1
17
20
2.6
7.9
11
3.5
7.9
11
5
7.9
13
36
7.9
44
1.2
99
1.6
98
4.2
96
15
85
25
75
31
69
39
61
82
18
11
25
4.2.3 Tolerable daily intakes
Using tolerable daily intakes (TDI) established by the European Food Agency (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 maximum 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 toddlers (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 exposures 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 concentrations 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 studies (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
26
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). Concentrations of BFRs in dust from Uppsala homes (paper III) were similar to
those found in Stockholm (paper II). Despite the differences in concentration 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 ∑Octaand ∑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 samples for BDE-47, or between dust from either method and BDE-153 concentrations in breast milk. The lack of correlations between BDE-153 concentrations 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 exposure 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.
27
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 BDE28, -47, -99, -153, -183, -197, -206, -207, -208, -209) concentrations for
indoor air ranged from 18 – 7300 pg/m3 and 17 to 8500 pg/m3 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 studied, 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 concentrations 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 concentrations 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 outdoor air from Sweden (∑PentaBDE 1.6-3.7 pg/m3; BDE-209 6.1-6.5 pg/m3)
(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 congener in both indoor and outgoing air, with no significant difference in concentration, 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 together with evidence of lower PBDE concentrations in outdoor air support
the hypothesis that PBDEs (including BDE-209) in the indoor air enter ventilation systems, are vented unchanged to the outdoors where they later undergo dilution through atmospheric mixing leading to the lower concentrations found outdoors than indoors. Thus ventilation is a likely conduit of
PBDEs from indoor sources to the outdoors.
28
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/m2 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/m2)
(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 knowledge 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, including 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 significant source of PBDEs to outdoor air which may eventually lead to contamination of food and dietary exposure (Harrad & Diamond, 2006).
29
5. Conclusions
PFOS, PFOA, PBDEs and HBCD were found in dust from all of the microenvironments studied. PBDEs were found in air from the different microenvironments but HBCD was detected in only a few air samples.Concentrations of PFOS and ∑OctaBDE in office dust were significantly higher (p<0.05) than in the other microenvironments while ∑PentaBDE
and HBCD were significantly higher in offices and day care centers compared to the other microenvironments studied. Car dust had higher median
BDE-209 concentrations than homes, apartments and day care centers. Significantly higher concentrations of tri-decaBDEs were detected in air from
offices compared to homes, daycare centers and cars, while BDE-209 concentrations in cars were significantly higher than in other microenvironments. The presence of congeners from the PentaBDE and OctaBDE technical 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 predominant 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 minor 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 relevant, body burdens of PFAAs and BFRs in adults and in toddlers in conjunction with air and dust sampling should be studied.
Vacuum cleaner bag dust and above floor settled dust correlated significantly 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
30
between concentrations of BDE-47 and vacuum cleaner bag dust indicating
that the indoor (dust ingestion) pathway may play a role in exposure. However, 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 microenvironments. No statistically significant differences were seen for concentrations of PBDEs in indoor and outgoing air. There was a significant
correlation between concentrations of BDEs in indoor and outgoing air confirming 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 environment.
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 consequent 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 relationship to body burden.
31
6. Acknowledgements
I would like to return thanks and appreciation to my supervisor; Prof. Cynthia 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 Thuresson 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 debates, 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 encouragement and support from the day I was born to this moment. My parents and siblings (Emmanuel and Margaret) have been an ancillary in the
achievement of this PhD, with their immutable love and belief in me, enabling 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 gratitude 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.
32
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