Indoor emissions and fate of flame retardants A modelling approach

Ioannis Liagkouridis

©Ioannis Liagkouridis, Stockholm University 2016 Cover graphic: Evi Markou and ‘Freepik’ ISBN 978-91-7649-341-0 Printed in Sweden by Holmbergs, Malmö 2016 Distributor: Department of Environmental Science and Analytical Chemistry (ACES)

To my family

Στην οικογένειά μου

i

Abstract

A significant number of consumer goods and building materials act as

emission sources of flame retardants (FRs) in the indoor environment. As a

result, FRs have become ubiquitous indoors raising concerns about human

exposure and possible health implications. Once released indoors, FRs can

escape to the outdoors where they can persist, be transported over long

distances and present a threat to the environment. Despite the increasing

number of studies reporting the occurrence of FRs in the indoor

environment, the understanding of i) how and to what extent these

chemicals are released from indoor sources, and ii) their subsequent fate

indoors remains limited. The overarching objective of this thesis was to

improve this understanding by assessing the indoor emissions and fate of FRs

using a combination of multimedia modelling strategies and

experimental/empirical approaches. Paper I identifies a number of

knowledge gaps and limitations regarding indoor emissions and fate of FRs

and the available modelling approaches. These include a limited

understanding of the key emission mechanisms for low volatility FRs,

uncertainties regarding indoor air/surface partitioning, poor characterization

of dust and film dynamics and a significant lack of knowledge regarding

indoor reaction/degradation processes. In Paper II we highlighted the serious

scarcity in physicochemical property data for the alternative FRs and

demonstrated the applicability of a simple QSPR technique for selecting

reliable property estimates for chemical assessments. A modelling fate

assessment indicated a strong partitioning to indoor surfaces and dust for

most of the alternative FRs. Indications for POP (persistent organic

pollutant)-like persistence and LRT (long-range transport) and

bioaccumulative potential in the outdoor environment were also identified

for many alternative FRs. Using an inverse modelling approach in Paper III

we estimated 2 to 3 orders of magnitude higher emissions of

organophosphate FRs (0.52 and 0.32 ng.h-1) than brominated FRs (0.083 μg.h-

1 and 0.41 μg.h-1) in Norwegian households. An emission-to-dust signal was

also identified for organophosphate FRs suggesting that direct migration to

dust may be a key fate process indoors. No evidence of a direct source-to-

dust transfer mechanism was seen in Paper IV where the chemical transfer

between a product treated with an organophosphate FR and dust in direct

contact was experimentally investigated. It was concluded though that direct

contact between an FR source and dust can result in contamination hotspots

indoors.

ii

Sammanfattning

Många konsumentvaror och byggmaterial fungerar som utsläppskällor till flamskyddsmedel (FRs) i inomhusmiljön. Av denna anledning är FRs vanligt förekommande föroreningar i inomhusmiljön, vilket har ökat farhågorna för att människors exponering för dessa ämnen kan leda till negativa hälsoeffekter. När FRs släpper från material i inomhusmiljön, kan de transporteras vidare till utomhusmiljön där de sedan kan finnas kvar, transporteras över långa avstånd och utgöra ett hot mot miljön på grund av sin svårnedbrytbarhet. Trots ett ökande antal studier med avseende på förekomst av FR i inomhusmiljön är förståelsen av i) hur och i vilken omfattning dessa substanser frigörs från inomhuskällor, och ii) deras efterföljande öde inomhus, fortfarande begränsad. Det övergripande syftet med denna avhandling var att förbättra denna förståelse genom att bedöma inomhus utsläpp och öde FR med hjälp av en kombination av modelleringsstrategier och empiriska metoder. I Paper I identifieras ett antal kunskapsluckor och begränsningar när det gäller utsläpp och öde och tillgängliga modelleringsmetoder av flamskyddsmedel inomhus. Viktiga luckor var bland annat en begränsad förståelse av de viktigaste utsläppsmekanismerna för FR med låg flyktighet, osäkerhet gällande fördelningen mellan luft och ytor, bristande karaktärisering av dynamiken av damm och ytfilmer samt en betydande brist på kunskap om reaktions- och nedbrytningsprocesser inomhus. I Paper II betonade vi den allvarliga bristen i fysikalisk-kemiska egenskapsdata för alternativa FR och visade tillämpligheten av en enkel QSPR teknik för att välja tillförlitliga uppskattningar av sådana egenskaper. Med hjälp av dessa egenskapsdata kunde vi genom modellering visa att de flesta alternativa FRs har en tydlig affinitet för damm och inomhusytor. Ett flertal av de alternativa flamskyddsmedlen uppvisade i likhet med klassiska långlivade organiska föroreningar egenskaper såsom persistens, och potential för bioackumulation och långväga transport. I Paper III uppskattades utsläppen av de fosforbaserade flamskyddsmedel i norska hushåll till 2-3 tiopotenser högre (0.52 och 0.32 ng.h-1) än utsläppen av bromerade FR (0.083 µg.h-1 och 0.41 µg.h-1). En ”utsläpp-till-damm” signal identifierades också för organofosfater vilket tyder på att direkt övergång till damm från flamskyddade material kan vara en viktig process inomhus. Detta undersöktes empiriskt med hjälp av kammarförsök där överföringen mellan en FR-behandladprodukt och damm i direkt kontakt med produkten studerades (Paper IV). Inga tydliga tecken på denna överföringsmekanism kunde observeras, men det kunde konstateras att direktkontakt mellan en FR-behandlad produkt och damm kan leda till s.k. ”föroreningshotspots” inomhus.

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List of Papers

Paper I

Liagkouridis I, Cousins IT, Cousins AP. Emissions and fate of brominated

flame retardants in the indoor environment: A critical review of modelling

approaches. Science of The Total Environment 2014, 491–492, 87-99.

Paper II

Liagkouridis I, Cousins AP, Cousins IT. Physical-chemical properties and evaluative fate modelling of 'emerging' and 'novel' brominated and organophosphorus flame retardants in the indoor and outdoor environment. Science of Total Environment 2015, 524-525, 416-426.

Paper III

Liagkouridis I, Cequier E, Lazarov B, Cousins AP, Thomsen C, Stranger M, Cousins IT. Relationships between estimated flame retardant emissions and levels in indoor air and house dust. Submitted to Indoor Air

Paper IV

Liagkouridis I, Lazarov B, Giovanoulis G, Cousins IT. Chemical mass transfer

of an organophosphate flame retardant between product source and dust

in direct contact. Manuscript

iv

Statement of contribution

Paper I

I was responsible for reviewing the literature, and extracting and evaluating

the critical information. I took the lead in writing the paper.

Paper II

I was responsible for compiling the physicochemical property data, applying the QSPR data evaluation technique, performing the modelling assesments and analysing results. I took the lead in writing the paper.

Paper III

I was involved in planning the study and primarly responsible for compiling the experimental emission factors. I modified and ran the model, derived the empirical emission estimates and took the lead in interpeting the results and writing the paper.

Paper IV

I took the lead role in planning the study and was responsible for analysing

the experimental results. I took the lead role in authoring the paper.

v

Contents

Abstract .................................................................................................. i

Sammanfattning ..................................................................................... ii

List of Papers ......................................................................................... iii

Statement of contribution ..................................................................... iv

List of abbreviations .............................................................................. vi

1. Introduction........................................................................................ 1

1.1. FRs in the indoor environment: cause for concern? ....................................... 2

1.2. Sources, emissions and indoor fate ................................................................ 3

1.3. Modelling approaches for emissions and indoor fate .................................... 4

1.4. Challenges in assessing the emissions and indoor fate of FRs ........................ 5

2. Objectives ........................................................................................... 7

3. Methods ............................................................................................. 8

3.1. Study chemicals .............................................................................................. 8

3.2. Physicochemical properties estimation methods ......................................... 10

3.3. Indoor emissions and fate of FRs .................................................................. 11

3.3.1. Multimedia modelling tools .................................................................. 11

3.3.2. Chamber experiment investigating chemical transfer from source to

dust in direct contact ...................................................................................... 12

3.4. Outdoor fate ................................................................................................. 12

3.5. Sensitivity and uncertainty analysis tools ..................................................... 13

4. Results and discussion ....................................................................... 14

4.1. Physicochemical properties .......................................................................... 14

4.2. Indoor emissions ........................................................................................... 16

4.2.1 Estimation of indoor emissions ............................................................. 16

4.2.2 Indication of direct migration to dust and experimental investigation . 17

4.3 Indoor fate and exposure .............................................................................. 19

4.3.1 Evaluative fate assessment .................................................................... 19

4.4 Outdoor fate: persistence and long-range transport .................................... 20

5. Conclusions and future perspectives .................................................. 22

Acknowledgements .............................................................................. 26

References ........................................................................................... 28

vi

List of abbreviations

BCMP-BCEP 2,2-bis(chloromethyl)trimethylene bis(bis(2-chloroethyl)phosphate)

BDE 28 2,4,4'-tribromodiphenyl ether

BDE 47 2,2',4,4'-tetrabromodiphenyl ether

BDE 99 2,2',4,4',5-pentabromodiphenyl ether

BDE 100 2,2',4,4',6-pentabromodiphenyl ether

BDE 153 2,2',4,4',5,5'-hexabromodiphenyl ether

BDE 154 2,2',4,4',5,6'-hexabromodiphenyl ether

BDE 183 2,2',3,4,4',5',6-heptabromodiphenyl ether

BDE 209 Decabromodiphenyl ether

BEH-TEBP (TBPH) Bis(2-ethylhexyl) tetrabromophthalate

BFR Brominated flame retardant

BPA-BDPP (BADP) Bisphenol A bis(diphenyl phosphate)

BTBPE 1,2-Bis(2,4,6-tribromophenoxy)ethane

Co Material surface concentration

CTD Characteristic travel distance

µ-CTE Micro-chamber/thermal extractor

DBDBE (BDBE-209) Decabromodibenzyl ether

DBDPE (BDPE-209) Decabromodiphenyl ethane

DBE-DBCH (TBECH) 4-(1,2-Dibromoethyl)-1,2-dibromocyclohexane

DBHCTD 5,6-Dibromo-1,10,11,12,13,13-hexachloro-11-tricyclo[8.2.1.02,9]tridecene

DBNPG Dibromoneopentyl glycol

DBP 2,4-Dibromophenol

DBS Dibromostyrene

DCP Diphenyl cresyl phosphate

Deca-BDE Decabrominated diphenyl ether, commercial mixture

Dm Solid-phase diffusion coefficient

DOPO 9,10-Dihydro-9-oxa-10-phosphaphenanthrene-10-oxide

EBTEBPI N,N′-Ethylenebis(tetrabromophthalimide)

vii

EH-TBB 2-Ethylhexyl 2,3,4,5-tetrabromobenzoate

EPS Expanded polystyrene

EU European Union

FRs Flame retardants

HBB Hexabromobenzene

HBCDD (HBCD) 1,2,5,6,9,10-Hexabromocyclododecane

HBCYD Hexabromocyclodecane

HCTBPH 1,2,3,4,7,7-hexachloro-5-(2,3,4,5-tetrabromophenyl)-Bicyclo[2.2.1]hept-2-ene

HEEHP-TEBP 2-(2-hydroxyethoxy)ethyl 2-hydroxypropyl 3,4,5,6-tetrabromophthalate

HOPFRs Halogenated organophosphorus flame retardants

hm Convective mass transfer coefficient

ICECRM Indoor chemical exposure classification/ranking model

KAW Air-water partition coefficient

KM/A Material-air partition coefficient

KOA Octanol-air partition coefficient

KOW Octanol-water partition coefficient

Ks Surface/air partition coefficient

LRTP Long-range transport potential

MV Molar volume

MW Molecular weight

NHOPFRs Non-halogenated organophosphorus flame retardants

Octa-BDE Octabrominated diphenyl ether, commercial mixture

OPFRs Organophosphorus flame retardants

PBB-Acr Pentabromobenzyl acrylate

PBBB Pentabromobenzyl bromide

PBBC Pentabromobenzyl chloride

PBDEs Polybrominated diphenyl ethers

PBDMPP Tetrakis(2,6-dimethylphenyl)-m-phenylene biphosphate

PBDPP (RDP) Resorcinol bis(diphenyl phosphate)

PBP Pentabromophenol

viii

PBP-AE Pentabromophenol allyl ether

PBT 2,3,4,5,6-Pentabromotoluene

PBT Persistence, bioaccumulation, toxicity

Penta-BDE Pentabrominated diphenyl ether, commercial mixture

PIR Polyisocyanurate

POP Persistent organic pollutant

POV Overall persistence

PUF Polyurethane foam

QSPR Quantitative structure-property relationship

REACH Registration, Evaluation, Authorisation and restriction of Chemicals

RoHS Restriction of Hazardous Substances Directive

SMURF Stockholm multimedia urban fate model

SVOCs Semi-volatile organic compounds

t1/2,BODY Human body biotransformation half-life

t1/2,air Degradation half-life in air

t1/2,soil Degradation half-life in soil

t1/2,water Degradation half-life in water

TBBPA Tetrabromobisphenol A

TBBPA-BAE Tetrabromobisphenol A bis(allyl) ether

TBBPA-BDBPE Tetrabromobisphenol A bis(2,3-dibromopropyl) ether

TBBPA-BHEE Tetrabromobisphenol A bis(2-hydroxyethyl) ether

TBBPA-BME Tetrabromobisphenol A bismethyl ether

TBBPS Tetrabromobisphenol S

TBCO 1,2,5,6-Tetrabromocyclooctane

TBNPA Tribromoneopentyl alcohol

TBOEP (TBEP) Tris(2-butoxyethyl) phosphate

TBP 2,4,6-Tribromophenol

TBP-AE (ATE) 2,4,6-Tribromophenyl allyl ether

TBP-DBPE 2,4,6-Tribromophenyl 2,3-dibromopropyl ether

TBX 1,2,4,5-Tetrabromo-3,6-dimethylbenzene

TCEP Tris (2-chlorethyl) phosphate

ix

TCIPP (TCPP) Tris(2-chloroisopropyl) phosphate

TDBPP Tris (2,3 dibromopropyl) phosphate

TDBP-TAZTO 1,3,5-Tris(2,3-dibromopropyl)-1,3,5-triazine-2,4,6- trione

TDCIPP (TDCPP) Tris[2-chloro-1-(chloromethyl)ethyl] phosphate

TDCPP Tris(2,3-dichloropropyl) phosphate

TE Transfer efficiency

TEHP Tris(2-ethylhexyl) phosphate

TEP Triethyl phosphate

TIBP Triisobutyl phosphate

TIPPP Tris(4-isopropylphenyl) phosphate

TMP Trimethyl phosphate

TMPP (TCP) Tricresyl phosphate

TNBP (TBP) Tri-n-butyl phosphate

TPHP (TPhP) Triphenyl phosphate

TPP Tri-n-propyl phosphate

TTBNPP Tri[3-bromo-2,2-bis(bromomethyl)propyl]phosphate

TTBP-TAZ 2,4,6-Tris(2,4,6-tribromophenoxy)-1,3,5-triazine

USD United States Dollar

USEPA United States Environmental Protection Agency

VOCs Volatile organic compounds

XPS Extruded polystyrene

y Air room concentration

yo Near surface, boundary-layer air concentration

x

1

1. Introduction

As a result of the technological advances and socioeconomic development in

the last century a wide range of chemicals and materials used in technical

applications and consumer products have become available. For example,

flame retardants (FRs), which include several classes of semivolatile organic

chemicals (SVOCs), have seen a drastic rise in production and use over the

past several decades corresponding with an increasing use of polymeric

materials (Alaee et al., 2003). FRs are applied to a wide range of plastics,

furniture and textiles, construction materials and electric and electronic

equipment to meet fire safety standards. Depending on the mode of their

incorporation FRs are divided into additive and reactive compounds (EFRA,

2016). Use in plastics accounts for approximately 85% of the total use of FRs

with the rest being used mostly in textile and rubber products (Beard and

Klimes, 2013). Indicative of the market size of FRs are the rough estimations

of 1.5 and 2 million metric tonnes global consumption for 2005 and 2011,

corresponding to a 2.9 and 5 billion USD market value, respectively (Beard

and Klimes, 2013; Harju et al., 2009).

Overall, four major groups of FRs are identified, namely: inorganic, organic

halogenated (brominated and chlorinated), organophosphorus and nitrogen-

based (Alaee and Wenning, 2002). This thesis focuses on brominated and

organophosphorus flame retardants (BFRs and OPFRs, respectively) because

these are the classes of FRs for which there has been most concern regarding

environmental effects and human health. The global market share of BFRs in

2005 and 2011 was estimated at around 20-21% while for OPFRs this was 14-

15% (Beard and Klimes, 2013; Harju et al., 2009). In Europe, BFRs and OPFRs

accounted for 10% and 20% of the total consumption in 2006, respectively

(van der Veen and de Boer, 2012).

The most widely produced and used BFRs, the polybrominated diphenyl

ethers (PBDEs), hexabromocyclododecane (HBCDD) and

tetrabromobisphenol A (TBBPA) received particular attention due to their

ubiquitous presence in the environment (de Wit, 2002; 2010; Hites, 2004;

Law et al., 2006; 2014; Wang et al., 2007) and their potential adverse effects

for wildlife and human health (Darnerud, 2003; 2008; Eskenazi et al., 2013;

Gascon et al., 2011). In light of strong evidence about their persistence,

bioaccumulation, toxicity (PBT) and long-range transport potential (LRTP),

penta- and octa-BDE technical formulations were listed as persistent organic

pollutants (POPs) under the Stockholm Convention on POPs (UNEP, 2009)

and their production and use was discontinued. Deca-BDE was recently

2

recommended to be included in Annex A of the Convention as it fulfils the

PBT and LRTP criteria (UNEP, 2014), with the final decision to be taken in

2017. Meanwhile, its use in electronic and electrical equipment has been

prohibited in the EU since 2008 under the EU Restriction of Hazardous

Substances Directive (RoHS) (BSEF, 2014). HBCDD was also listed as a POP

substance in 2013 with specific exemptions for use in expanded and extruded

polystyrene (EPS & XPS) in buildings (UNEP, 2013) . A ‘sunset day’ set for mid-

2015 was announced for HBCDD by the European Commission under the

REACH Regulation.

To meet the continuous demand for FRs following the bans and

restrictions on the production and use of PBDEs and HBCDD, there has been

a shift towards alternative FRs. These include the ‘novel’ and ‘emerging’ BFRs

and the OPFRs. Some of these chemicals have been used for decades,

however they have simply been out of scientific and political focus. Their use

has increased since the legislative regulations on PBDEs. The increasingly

frequent detection of these alternative FRs in a variety of environmental

matrices (Covaci et al., 2011; van der Veen and de Boer, 2012) has raised

concerns about potential risks to humans and the environment. Existing

evidence suggests that some of the alternative BFRs and OPFRs exhibit PBT

and LRTP characteristics (Covaci et al., 2011; de Wit et al., 2010; EFSA, 2012;

Moller et al., 2012); yet for most of these chemicals the information on their

production volumes, use patterns and emissions as well as their

environmental fate and toxicity profiles is often limited and inconclusive.

1.1. FRs in the indoor environment: cause for concern?

According to studies on human activity patterns, humans spend on average

more than 90% of their time indoors (Klepeis et al., 2001; Leech et al., 2002;

Schweizer et al., 2007). Therefore, the indoor environment is regarded as

particularly important for potential human exposure to chemical pollutants.

Due to their widespread use in a variety of indoor materials and consumer

goods, FRs can migrate into the indoor environment (Kemmlein et al., 2003;

Rauert and Harrad, 2015). Most of the FRs are used as additives rather than

being chemically bonded to the polymeric material, thus they are more likely

to be released from the source. High levels of FRs are continuously reported

in indoor environments worldwide (i.e. Ali et al., 2012; Bergh et al., 2011;

Cequier et al., 2014; Harrad et al., 2010; Saito et al., 2007; Shoeib et al., 2012;

Sjodin et al., 2008; Wensing et al., 2005); these often exceed outdoor levels

3

suggesting that the indoor environment is a potential source to outdoors

(Björklund et al., 2012; Newton et al., 2015)

The indoor contamination with FRs has raised concerns as to whether and

to what extent it leads to significant human exposure to these chemicals. A

number of studies have demonstrated that indoor exposure to FRs mainly

through dust ingestion may be a significant contributor to body burden,

especially for sensitive age groups; i.e. for PBDEs (Lorber, 2007; Trudel et al.,

2011; Watkins et al., 2012), HBCDD (Roosens et al., 2009) and OPFRs (Cequier

et al., 2015; Fromme et al., 2014). Moreover, epidemiological studies have

identified associations between levels of certain BFRs (PBDEs and alternates)

and OPFRs in indoor dust and human health risks including endocrine

disrupting effects (Araki et al., 2014; Johnson et al., 2013; Meeker et al.,

2009; Meeker and Stapleton, 2010).

1.2. Sources, emissions and indoor fate

In a few cases, FR levels in indoor air and dust have been successfully

correlated with the presence or number of certain FR treated consumer

products (Ali et al., 2012; Allen et al., 2008; de Wit et al., 2012; Harrad et al.,

2004); though, such clear, consistent relationships are often hard to obtain,

as they are also influenced by other microenvironment characteristics.

However, the influence of FR sources has been exemplified by associations

between FR levels and the presence, introduction/removal or the proximity

to sources (or likely sources) (Brandsma et al., 2014; Muenhor and Harrad,

2012; Stuart et al., 2008; Whitehead et al., 2013).

To sufficiently identify the risks arising from the indoor occurrence of FRs

so that the necessary measures are taken to mitigate those risks, a sufficient

understanding of i) how and to what extent these chemicals are released

from indoor sources, and ii) their subsequent fate indoors is required. Gas-

phase emission to the air (volatilisation) has received most attention as the

main chemical release mechanism for SVOCs such as FRs. Once emitted to

the air, FRs may partition to airborne particles, settled dust and other indoor

surfaces including humans or be removed by ventilation. Physicochemical

properties play a key role in SVOC partitioning behaviour indoors with the

octanol-air partition coefficient (KOA) being a strong indicator of the likely

behaviour (Weschler and Nazaroff, 2010). Given the high chemical content

and the relatively low volatilisation rates (compared to VOCs), FR-treated

products may remain continuous emission sources over extended periods of

time (Kemmlein et al., 2003; Wensing et al., 2005). Recent research however

4

has pointed towards alternative emission mechanisms which may lead to

direct, enhanced migration of chemicals into dust (Rauert and Harrad, 2015;

Rauert et al., 2014a; Schripp et al., 2010). These include i) abrasion/physical

weathering of the chemically treated product and ii) chemical transfer from

the chemically treated product to the dust in direct contact with the material

surface. The occurrence of such mechanisms could explain the elevated dust

concentrations of the extremely low volatility FRs, for which volatilisation

doesn’t seem a plausible emission mechanism.

1.3. Modelling approaches for emissions and indoor fate

Models constitute useful tools for elucidating and predicting the emissions,

behavior and fate of SVOCs such as FRs in the indoor environment. Over the

years, a large number of mass transfer models for emissions from indoor

diffusional sources such as building materials and consumer products (Guo,

2013) as well as several multimedia mass-balance indoor fate models have

been developed (see Paper I). From a mechanistic point of view, SVOC

emissions from diffusional sources are governed by the material-air partition

coefficient (KM/A) and the convective mass transfer coefficient (hm) (Xu and

Little, 2006). A simplified emission model for SVOCs present in high

concentrations was suggested by (Xu et al., 2009):

E(t) = hm [yo – y(t)] (1)

where yo – y(t) represents the concentration gradient between the near

surface, boundary-layer air concentration (yo) and the air room

concentration (y). The boundary-layer air concentration is calculated from

the linear equilibrium relationship Co/KM/A, where Co is the material surface

concentration (assumed to remain effectively constant). Several approaches

for the estimation of KM/A and hm exist (Holmgren et al., 2012).

Such an emission model may be coupled with an indoor multimedia

model in order to predict chemical fate (concentrations, residence time, etc.)

indoors. This approach is favourable when an exposure assessment is the

main endpoint of interest. Indoor multimedia models consist of several

compartments representing typical indoor media and different phases within

the same medium (i.e. air, airborne and dust particles, interior

surfaces/organic film, human skin etc.). Indoor chemical fate is simulated

through a number of diffusive and advective mass-transfer as well as

reaction/transformation processes, responsible for the introduction,

intermedia transfer and removal of a chemical. An emission input is assigned

5

and concentrations in different indoor compartments and phases are

calculated solving a system of mass-balance equations. Such an indoor fate

model can also be fitted in order to calculate (back-calculate) emissions

based on measured levels indoors.

1.4. Challenges in assessing the emissions and indoor fate of FRs

From a modelling perspective, an integrated approach for assessing indoor

chemical fate requires information on the 3 following model components;

emissions, physicochemical properties and indoor fate processes. It is

common, especially at the early stages of chemical assessment of emerging

contaminants, that limited information is available regarding their emissions

and physicochemical properties. Emissions and indoor fate were specifically

addressed in Paper I while physicochemical properties were reviewed in

Paper I. The main challenges associated with each of the 3 basic components

are discussed below:

i) Emissions. The emission-to-air magnitude of FRs can be experimentally

measured in controlled chamber environments (Rauert et al., 2014b).

However, such emission factors/studies are scarce due to practical

limitations. A mass-transfer model such as the one presented above (see

Eq. 1) can be used to estimate emissions from sources. This approach

however requires a good characterisation of sources (i.e. FR content,

material composition/geometry) and this information is often not

readily available. Additionally, limited information regarding model

parameterisation (i.e. KM/A) or increased model complexity to capture

more complex sources (i.e. multiple layers) may render this approach

highly uncertain and effectively impractical. Finally, significant

uncertainties remain/exist regarding the occurrence of a direct

migration to dust pathway. The magnitude and the mechanisms behind

direct migration to dust are currently understudied and its impact on

the indoor fate and exposure to SVOCs/FRs widely unexplored.

ii) Physicochemical properties. The unavailability of physicochemical

properties can pose a significant challenge to any chemical fate

assessment. For emerging FRs, measured physicochemical properties

are scarce. In some cases, the extreme properties of FRs exceed the

performance limit of analytical methods for determination. This paucity

can be overcome with the aid of quantitative structure-property

6

relationship (QSPR) methods. Nevertheless, these tools have their own

limitations and results must be interpreted with caution. It is then

advisable that the plausibility of the available physicochemical

properties is assessed.

iii) Indoor fate processes. Key fate processes indoors include air-surface and

air-particle partitioning, air advection/ventilation, particle deposition

and resuspension, dust and organic film removal and

reaction/degradation (see Paper I for a detailed description). Depending

on the physicochemical properties of an organic chemical each of the

above processes is more or less influential for indoor fate. Many of these

processes have not been rigorously evaluated/characterized indoors

(i.e. air-surface partitioning, dust/film removal rates, degradation

mechanisms and rates) and may cause significant uncertainty in model

predictions.

Another challenge to be addressed in modelling the indoor fate of chemicals

is the description of the indoor environment. Indoor fate models with a

simple or more complex indoor description exist; in addition, these models

assume steady-state or dynamic conditions. Although the selection of an

appropriate model usually depends on the endpoint of interest, there is still

an insufficient number of studies where the predictive power of indoor fate

models is evaluated against real-time measurements.

7

2. Objectives

The overall objectives of this thesis were to assess the indoor emissions and

fate of FRs with the ultimate goal of improving our understanding of the key

factors that govern their release from indoor sources and their subsequent

fate in the indoor environment. This was achieved in 4 individual studies

(Papers I-IV) mainly with the aid of multimedia modelling tools but also by

carrying out a specially designed experiment to study one of the key fate

processes identified. The main hypotheses tested in this thesis are:

I. the current understanding of the physicochemical properties,

emissions and fate processes of BFRs is sufficient to accurately model

their emission and fate in the indoor environment (Papers I and II),

II. alternative FRs demonstrate a similar environmental fate behavior to

PBDEs (Paper II),

III. inverse modelling can be used to reliably estimate the indoor

emissions of FRs (Paper III),

IV. direct emission-to-dust mechanisms are important for controlling

the indoor fate of FRs, especially low volatile ones (Papers III and IV)

The major objectives of each paper are presented below.

Paper I aimed to critically explore the current (as of 2013) understanding of

the indoor emissions and fate of BFRs and the available modelling

approaches. The scope of this critical review was to identify key limitations

and provide a roadmap for future experimental and modelling research

needed to improve our understanding of indoor fate and exposure of BFRs.

Paper II’s main objective was to evaluate the environmental fate of

alternative FRs in the indoor and outdoor environment. This study also aimed

at assessing the availability and reliability of the physicochemical properties

required for modelling purposes.

Paper III aimed at quantifying/estimating indoor emissions of FRs as well as

to identify an emission-to-dust ‘signal’ and link it with possible sources.

Paper IV’s main objective was to investigate the magnitude, timescale, and

possible mechanism of chemical mass transfer between an OPFR treated

product and dust in direct contact.

8

3. Methods

3.1. Study chemicals

In total 67 FRs (Table 1) were assessed in this thesis. These are divided into

two main groups; BFRs (n = 45) and OPFRs (n = 22). The selected BFRs include

10 ‘established’ BFRs, namely the PBDEs (8 BDE congeners), HBCDD and

TBBPA, and 35 of the mostly known as ‘novel’ and ‘emerging’ BFRs (herein

referred to as ‘alternative’ FRs). The selected OPFRs can be grouped into two

sets; halogenated OPFRs (HOPFRs, n = 7) and non-halogenated OPFRs

(NHOPFRs, n = 15). Although this thesis focuses on organic FRs, the methods

and models presented here can be conceptually applied to most non-ionic,

organic chemicals.

Table 1. Abbreviations, common names and CAS numbers of the FRs studied in

Papers II-IV (Bergman et al., 2012; van der Veen and de Boer, 2012)

ABBREVIATION COMMON NAME CAS NUMBER PAPER

‘Established’ BFRs

BDE 28 2,4,4'-tribromodiphenyl ether 41318-75-6 II, III

BDE 47 2,2',4,4'-tetrabromodiphenyl ether 5436-43-1 II, III

BDE 99 2,2',4,4',5-pentabromodiphenyl ether 60348-60-9 II, III

BDE 100 2,2',4,4',6-pentabromodiphenyl ether 189084-64-8 II, III

BDE 153 2,2',4,4',5,5'-hexabromodiphenyl ether 68631-49-2 II, III

BDE 154 2,2',4,4',5,6'-hexabromodiphenyl ether 207122-15-4 II, III

BDE 183 2,2',3,4,4',5',6-heptabromodiphenyl ether 207122-16-5 II, III

BDE 209 Decabromodiphenyl ether 1163-19-5 II, III

TBBPA Tetrabromobisphenol A 79-94-7 II

HBCDD (HBCD)

1,2,5,6,9,10-Hexabromocyclododecane 3194-55-6 II

‘Alternative’ BFRs

BEH-TEBP (TBPH)

Bis(2-ethylhexyl) tetrabromophthalate 26040-51-7 II, III

BTBPE 1,2-Bis(2,4,6-tribromophenoxy)ethane 37853-59-1 II, III

DBDBE (BDBE-209)

Decabromodibenzyl ether 497107-13-8 II

DBDPE (BDPE-209)

Decabromodiphenyl ethane 84852-53-9 II, III

DBE-DBCH (TBECH)

4-(1,2-Dibromoethyl)-1,2-dibromocyclohexane

3322-93-8 II, III

DBHCTD 5,6-Dibromo-1,10,11,12,13,13-hexachloro-11-tricyclo[8.2.1.02,9]tridecene

51936-55-1 II

9

DBNPG Dibromoneopentyl glycol 3296-90-0 II

DBP 2,4-Dibromophenol 615-58-7 II

DBS Dibromostyrene 31780-26-4 II

EBTEBPI N,N′-Ethylenebis(tetrabromophthalimide) 32588-76-4 II

EH-TBB 2-Ethylhexyl 2,3,4,5-tetrabromobenzoate 183658-27-7 II, III

HBB Hexabromobenzene 87-82-1 II, III

HBCYD Hexabromocyclodecane 25495-98-1 II

HCTBPH 1,2,3,4,7,7-hexachloro-5-(2,3,4,5-tetrabromophenyl)-Bicyclo[2.2.1]hept-2-ene

34571-16-9 II

HEEHP-TEBP 2-(2-hydroxyethoxy)ethyl 2-hydroxypropyl 3,4,5,6-tetrabromophthalate

20566-35-2 II

PBB-ACR Pentabromobenzyl acrylate 59447-55-1 II

PBBB Pentabromobenzyl bromide 38521-51-6 II

PBBC Pentabromobenzyl chloride 58495-09-3 II

PBEB 2,3,4,5,6-Pentabromoethylbenzene 85-22-3 II, III

PBP Pentabromophenol 608-71-9 II

PBP-AE Pentabromophenol allyl ether 3555-11-1 II, III

PBT 2,3,4,5,6-Pentabromotoluene 87-83-2 II, III

TBBPA-BAE Tetrabromobisphenol A bis(allyl) ether 25327-89-3 II

TBBPA-BDBPE Tetrabromobisphenol A bis(2,3-dibromopropyl) ether

21850-44-2 II

TBBPA-BHEE Tetrabromobisphenol A bis(2-hydroxyethyl) ether

4162-45-2 II

TBBPA-BME Tetrabromobisphenol A bismethyl ether 37853-61-5 II

TBBPS Tetrabromobisphenol S 39635-79-5 II

TBCO 1,2,5,6-Tetrabromocyclooctane 3194-57-8 II

TBNPA Tribromoneopentyl alcohol 1522-92-5 II

TBP 2,4,6-Tribromophenol 118-79-6 II

TBP-AE (ATE) 2,4,6-Tribromophenyl allyl ether 3278-89-5 II, III

TBP-DBPE 2,4,6-Tribromophenyl 2,3-dibromopropyl ether

35109-60-5 II

TBX 1,2,4,5-Tetrabromo-3,6-dimethylbenzene 23488-38-2 II, III

TDBP-TAZTO 1,3,5-Tris(2,3-dibromopropyl)-1,3,5-triazine-2,4,6- trione

52434-90-9 II

TTBP-TAZ 2,4,6-Tris(2,4,6-tribromophenoxy)-1,3,5-triazine

25713-60-4 II

Non-halogenated OPFRs

BPA-BDPP (BADP)

Bisphenol A bis(diphenyl phosphate) 5945-33-5, 181028-79-5

II

DCP Diphenyl cresyl phosphate 26444-49-5 II

10

DOPO 9,10-Dihydro-9-oxa-10-phosphaphenanthrene-10-oxide

35948-25-5 II

PBDMPP Tetrakis(2,6-dimethylphenyl)-m-phenylene biphosphate

139189-30-3 II

PBDPP (RDP) Resorcinol bis(diphenyl phosphate) 57583-54-7, 125997-21-9

II

TBOEP (TBEP) Tris(2-butoxyethyl) phosphate 78-51-3 II, III

TEHP Tris(2-ethylhexyl) phosphate 78-42-2 II

TEP Triethyl phosphate 78-40-0 II

TIBP Triisobutyl phosphate 126-71-6 II

TIPPP Tris(4-isopropylphenyl) phosphate 2502-15-0 II

TMP Trimethyl phosphate 512-56-1 II

TMPP (TCP) Tricresyl phosphate 1330-78-5 II, III

TNBP (TBP) Tri-n-butyl phosphate 126-73-8 II, III

TPHP (TPHP) Triphenyl phosphate 115-86-6 II, III

TPP Tri-n-propyl phosphate 513-08-6 II

Halogenated OPFRs

BCMP-BCEP 2,2-bis(chloromethyl)trimethylene bis(bis(2-chloroethyl)phosphate)

38051-10-4 II

TCEP Tris (2-chlorethyl) phosphate 115-96-8 II, III

TCIPP (TCPP) Tris(2-chloroisopropyl) phosphate 13674-84-5 II, III, IV

TDBPP Tris (2,3 dibromopropyl) phosphate 126-72-7 II

TDCIPP (TDCPP)

Tris[2-chloro-1-(chloromethyl)ethyl] phosphate

13674-87-8 II, III

TDCPP Tris(2,3-dichloropropyl) phosphate 78-43-3 II

TTBNPP Tri[3-bromo-2,2-bis(bromomethyl)propyl]phosphate

19186-97-1 II

3.2. Physicochemical properties estimation methods

Several key physicochemical properties determine the environmental

behaviour and fate of non-ionic, organic chemicals such as the BFRs and

OPFRs (Mackay, 2001). The physicochemical properties required for the

modelling assessments carried out in this thesis (Papers II, III and IV) are the

molecular weight (MW), the octanol-water (KOW), air-water (KAW) and the

octanol-air (KOA) partition coefficients, and the degradation half-lives in air

(t1/2,air), water (t1/2,water) and soil (t1/2,soil). For internal consistency KOA is

calculated as the KOW/KAW ratio. In Paper II a dataset of KOW, KAW and half-lives

including experimental and software estimated values by 2 property

estimation tools, the SPARC On-Line Calculator (Hilal et al., 2004) and the

11

USEPA’s EPISuite platform (USEPA, 2012) was compiled. These estimation

models are based on QSPRs (relationships between structural features and

chemical properties) and used as well-established tools in regulatory risk

assessment.

A QSPR-based technique presented by Stieger et al. (2014) is applied to

evaluate the reliability of the available KOW and KAW property data and assist

in property selection (Paper II). This method is based on an expected linear

relationship between log KOW (and log KAW) and a molecular descriptor such

as molecular weight (MW) or molar volume (MV) for an individual class of

organic compounds (Mackay et al., 2006; Palm et al., 2002; Schenker et al.,

2005). The method is implemented in 3 steps;

i) identification of structurally similar compounds,

ii) plotting of all available log KOW (log KAW) values for the structural

analogues against MW (or MV) to derive a linear relationship and,

iii) exclusion of those values that lie clearly above or below the

regression line (a general rule of 2 log units was applied) and

averaging of the remaining ones (referred to as ‘best guess’

estimates).

3.3. Indoor emissions and fate of FRs

3.3.1. Multimedia modelling tools

A level III fugacity-based chemical fate model was used for estimating the

emissions (Paper III) and the evaluative fate assessment (Papers II) of the

selected FRs in the indoor environment. This model was developed as the

indoor module (Figure 1) of a larger urban chemical fate model, known as

the Stockholm Multimedia URban Fate Model (‘SMURF’) (Cousins, 2012)). It

consists of 3 compartments; indoor air (including particles), vertical and

horizontal surfaces. Both surface types are assumed to be covered by an

organic film layer with an additional thin layer of dust covering horizontal

surface. All indoor chemical fate processes are schematically shown in Figure

1. In its original setup, the indoor ‘SMURF’ solves a steady-state mass balance

to calculate chemical concentrations in the 3 compartments, as well as the

different phases within, given an emission-to-air input flux. In Paper II the

model was run with a unit emission-to-air input to perform an evaluative

indoor fate assessment. In Paper III the model’s mass-balance was re-

arranged so it can (back)-calculate the emissions to indoor air from both air

12

and dust concentrations (air-based and dust based emission estimates,

respectively).

3.3.2. Chamber experiment investigating chemical transfer from source to

dust in direct contact

In Paper IV, an experiment investigating chemical transfer between an OPFR-

treated source and dust in direct contact was performed. The experiment

was carried out in a Micro-Chamber/Thermal ExtractorTM - 120 (μ-CTETM,

Markes International). The μ-Chamber consists of six separate emission test

cells (ETCs) each having a diameter of 4.5 cm and a height of 2.8 cm, resulting

to a total internal surface area of 71 cm2 and a total volume of 44 cm3. The

internal surface of each cell is made of inert-coated stainless steel. To

maintain a constant temperature during the test, the µ-CTE unit was placed

in a laboratory where the temperature was controlled at 23 ± 1°C. A

Polyisocyanurate (PIR) insulation board (1.2 x 0.6 x 0.09 m) was used as the

test source material. The PIR insulation board contained 2.2% w/w TCIPP

which is added to meet fire safety regulations.

3.4. Outdoor fate

The OECD POV (overall persistence) & LRTP (long-range transport potential)

Screening Tool (Wegmann et al., 2009) was used to assess the environmental

fate of the FRs in the outdoor environment (Paper II). The Tool is designed as

a support tool in chemical risk assessment and uses POV and LRTP metrics to

identify possible POP-like chemicals. Τhe Tool’s multimedia fate model

employs a representative global-scale environmental description (Figure 1).

The mode of release can be into either of the three compartments; air,

seawater and soil. The environmental fate is evaluated in terms of the

steady-state mass distribution, POV and LRTP. POV quantifies the timescale of

the degradation loss of a selected chemical in the entire environment. LRTP

is expressed through two individual indicators; the characteristic travel

distance (CTD, km) which is the distance a chemical travels from the point of

release to the point where its concentration has dropped to about 37% and

the transfer efficiency (TE, %) defined as the mass flux into a selected target

compartment divided by the emission mass flux.

13

Figure 1. Schematic of the indoor module of SMURF model (Cousins, 2012) and The

Tool’s unit-world fate model (Wegmann et al., 2009)

3.5. Sensitivity and uncertainty analysis tools

Uncertainty in environmental modeling can often be an impeding factor

(Buser et al., 2012). The limited availability of high quality physicochemical

property data for the alternative FRs as an important source of uncertainty

was emphasised before. A QSPR-based technique was also presented aimed

at reducing uncertainty in physicochemical property input data. With a focus

on physicochemical properties as a major contributor to model uncertainty,

appropriate tools to illustrate the sensitivity of model outputs to input

property data and quantify the propagated uncertainty were applied. In

Paper II a Monte Carlo analysis using the built-in module of the ‘Tool’

applying the default dispersion factors (5 for partition coefficients and 10 for

half-lives) was performed. The uncertainty contribution of each of the five

physicochemical input properties (KOW, KAW, t1/2,air, t1/2,water and t1/2,soil) to Pov

14

and LRTP was monitored. In Paper III a KOA sensitivity and uncertainty analysis

was conducted. The sensitivity of model results to KOA was investigated

according to the principles presented by MacLeod et al. (2002). Min and max

KOA values were used as model input to provide an illustration of the KOA-

associated uncertainty range.

4. Results and discussion

4.1. Physicochemical properties

One of the main goals of Paper II was to assess the availability and plausibility

of the available physicochemical properties. Overall, there is a scarcity of

experimental data for KOW and KAW. The lack of experimental data is especially

pronounced for the alternative BFRs whereas for the ‘established’ BFRs and

the OPFRs data availability is better. Experimental reaction half-lives are

commonly not available. Apart from the limited availability of measured KOW

and KAW which can be addressed with the aid of property estimation tools, a

significant variability among reported values (both measured and software-

calculated) is occasionally seen (see Figure 2 in Paper II). Discrepancies in KOW

are larger for high-molecular weight compounds; similar findings have been

reported for phthalate esters (Cousins and Mackay, 2000) and recently for

PCBs, PBDEs and ‘novel’ FRs (Zhang et al., 2016). Overall, both the limited

availability of experimental physicochemical properties and the variability in

data can be attributed to i) the extreme properties of many FRs (e.g. low

solubilities or high KOW or MW) for which the performance of analytical

methods for determination or models’ application domain are exceeded

and, ii) differences in calculation methods employed by the software

estimation tools (Arp et al., 2010; Arp et al., 2006; Zhang et al., 2010).

The reliability of the available KOW and KAW data was investigated (Paper

II) using the ‘best guess’ estimation method. When applied, a linear

relationship between log KOW and MW was identified for the 3 distinct groups

of FRs (BFRs, NHOPFRs and HOPFRs). A similar linear correlation was found

when MV was used as a molecular descriptor (Figure 2). Based on this linear

trend, the ‘best estimate’ KOW values were obtained after ‘suspected’

erroneous or implausible data points were removed and the remaining

values averaged. In contrast to KOW, a clear linear log KAW - MW (or log KAW -

MV) relationship was not observed for any of the FR groups. This could

possibly be a combined effect of insufficient or poor quality KAW data and

15

structural differences between the selected compounds. Such poor

correlations between KAW and a molecular descriptor have been previously

demonstrated for other organic compound classes (Kuramochi et al., 2014;

Mackay et al., 2006; Palm et al., 2002). A good correlation was obtained only

for PBDEs when these were subdivided from the larger BFR set. Besides being

a more homogenous group of chemicals, KAW data availability for PBDEs is

much better compared to the alternative BFRs.

Figure 2. Log KOW - MW and log KOW - MV linear regressions for the alternative BFRs

(Paper II)

EPISuite estimates of half-lives are widely used since they often are the only

readily available half-life values. The weaknesses of EPISuite’s models in

predicting biodegradation half-lives are well documented, especially for

highly persistent organic chemicals (Aronson et al., 2006; Gouin et al., 2004).

Some strategies to improve the predictive power of BIOWIN model and

approaches to assist in half-life selection have been proposed (Arnot et al.,

2005; Aronson et al., 2006), yet the uncertainty may remain significant and

its effect on model output has to be explored. Due to a lack of experimental

biodegradation half-lives it was not possible to assess the accuracy of

BIOWIN’s estimations for most of the FRs. Even when data exist (i.e. for the

‘established’ BFRs and some of the OPFRs), these are often inconclusive,

highly variable or not reliable, making a direct comparison difficult to

perform. It is therefore suggested that these are used with caution and the

associated uncertainty demonstrated.

y = 0,0106x + 1,1722R² = 0,6977

2

4

6

8

10

12

14

100 300 500 700 900 1100

Log

Ko

w

Molecular weight

y = 0,0274x + 0,6122R² = 0,7004

2

4

6

8

10

12

14

50 150 250 350 450

Molar volume

16

4.2. Indoor emissions

4.2.1 Estimation of indoor emissions

In Paper III we estimated the emission-to-air rates of 26 FRs in indoor

environments from Norway using modelling and empirical methods (the

latter based on experimental emission factors). Overall, modelling results

demonstrated a large range of emission rates spanning over 7 orders of

magnitude (from 0.004 ng.h-1 to 39 μg.h-1). Differences in the emission

strength of FR sources and the high variability environmental conditions

indoors are strong causal factors of the large variation.

An enhanced emission strength of indoor OPFR sources (0.083 and 0.41

μg.h-1 median values for air-based and dust-based emissions, respectively)

compared to BFR ones (0.52 and 0.32 ng.h-1 median values for air-based and

dust-based emissions, respectively) was realised. This may be partly a result

of the increase in the use of OPFRs following the ban of the formerly widely

used PBDEs. The ‘emerging’ BFRs TBECH and DBDPE are among the highest

emitted BFRs, also indicating the shift towards alternative BFRs. TCIPP, TCEP

and TPHP exhibit the highest emissions among all FRs which is indicative of

the widespread use of these chemicals in consumer applications (van der

Veen and de Boer, 2012). Specific applications such as upholstery and

thermal insulation can be significant sources of TCIPP and TCEP due to their

elevated content and large emitting surfaces.

In Figure 3, the emission rates of selected FRs from Paper III and other

modelling studies are listed. Emission rates are area-normalised (ng.m-2.h-1)

to allow for direct comparison. In general, the experimentally-based

emission estimates are in good agreement with modelling results despite the

considerable uncertainty in both. Looking at similar modelling studies from

the US (Batterman et al., 2009; Shin et al., 2014; Zhang et al., 2009; Zhang et

al., 2011) we observe higher emissions compared to Norway. This is most

likely a result of US’s stringent fire safety requirements, however differences

in modelled emission estimates may occasionally arise from different model

parameterisation.

17

Figure 3. Area-specific emission rates (ng.m-2.h-1) estimated using modelling and

empirical methods

4.2.2 Indication of direct migration to dust and experimental investigation

Direct migration to dust has been hypothesised as a possible chemical

release pathway from indoor sources (Suzuki et al., 2009; Takigami et al.,

2008). Two potential transfer mechanisms have been proposed and recent

work by Rauert and coworkers has provided a first indication of their relative

significance (Rauert and Harrad, 2015; Rauert et al., 2014a). Paper III aimed

to investigate the occurrence of such mechanisms using an indirect approach

which is based on modelled emission estimates. Indications of direct

migration to dust were found for 4 of the 7 OPFRs included in the study. The

identification of a consistent emission-to-dust signal (referring to higher

dust-based than air-based emission estimates) for these chemicals is most

likely associated with their presence, often in high concentrations, in certain

consumer products from where they are more prone to be transferred to

dust via abrasion or direct source-dust contact. For example, TCEP and TCIPP

which are added to PVC, PUF applications and textiles, often in high amounts

1,E-04

1,E-03

1,E-02

1,E-01

1,E+00

1,E+01

1,E+02

1,E+03

1,E+04

1,E+05

Pap

er

III

Shin

et

al. 2

014

Pap

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III

Pap

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Shin

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014

Pap

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Zhan

g et

al.

20

09

Zhan

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al.

20

09

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al.

20

11

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11

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Bat

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Pap

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Pap

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TBB TBPH ΣBDEs TCEP TCIPP TDCPP TPHP

Esti

mat

ed e

mis

sio

ns,

ng.

m-2

.h-1

Modelled (air-based)

Modelled (dust-based)

Modelled (air+dust)

Empirical

18

(EC, 2009; Kemmlein et al., 2003) can be released to dust via

abrasion/wearing down of the parent material. Such an emission-to-dust

process is most likely episodic, strongly localized thus, resulting in a weak,

inconsistent signal as demonstrated for the BFRs in the same study. KOA

associated uncertainty may also be responsible for obscuring the

identification of such a signal.

Paper IV, however, did not provide experimental evidence of the

occurrence of a direct chemical migration pathway (via solid or liquid-phase

diffusion) between a TCIPP treated source and dust in direct contact with the

source. It was concluded that the rapid and substantial transfer that was

observed after only 8 h of source-dust contact (Figure 4) was a result of gas-

phase diffusion from the insulation board to the dust on its surface and the

surrounding air. TCIPP in the air and dust appeared close to equilibrium (if

the suspected outlier for the 24 h is excluded, see Figure 4) as a result of the

well-mixed conditions in the chamber. In a real room however where such

well-mixed conditions as in the micro-chamber do not generally apply there

might be a gradient of concentrations of TCIPP in air above the surface of a

product. The dust sitting on the product is surrounded by air that has the

highest concentration of any air in the room, and should therefore come to

a higher concentration than dust elsewhere in the room as observed for

DEHP in the experiment of Schripp et al. (2010). This, along with chemical

transfer by abrasion, may result in contamination hotspots indoors and could

explain why high (non-equilibrium) Kda values have been observed even for

relatively volatile OPFRs (with log KOA < 9) (Cequier et al., 2014)

Figure 4. Mean, min and max (error bars) dust concentrations (µg/g) of TCIPP in the

μChamber pre-experiment (0h) and after 8 h, 24 h and 7 d of source-dust contact.

The air concentrations (µg/m3) correspond to air sampled at 24 h and 7 d (Paper IV)

0

1

2

3

4

5

1,E-01

1,E+00

1,E+01

1,E+02

1,E+03

0h 8h 24h 7d

Air

co

nce

ntr

atio

n (

μg/

m3 )

Du

st c

on

cen

trat

ion

(μ

g/g)

Duration of dust exposure

Dust

Air

19

4.3 Indoor fate and exposure

4.3.1 Evaluative fate assessment

The modelling assessment in Paper II demonstrated the influence of

physicochemical properties (as expressed by KOA) on the indoor fate FRs. Due

to their moderate to high hydrophobicity and low to extremely low volatility

(resulting in a moderate to high KOA) most of the 67 FRs assessed exhibit a

preference for the organic phase on indoor surfaces and particles and thus,

once emitted to indoor air, are likely to distribute favourably to indoor

surfaces (note that indoor surfaces are consisted of an organic surface film

and settled dust). Table 2 shows the predicted indoor distribution by indoor

‘SMURF’ based on the KOA of the FR compound. As to the individual FRs

groups, more than 90% of the steady-state total mass indoors of all the

‘established’ BFRs (9 PBDEs, HBCDD and TBPPA), 27 of the 35 alternative

BFRs, 8 of the 15 NHOPFRs and 5 of the 7 HOPFRs is predicted to be present

on indoor surfaces (vertical and horizontal).

Table 2. FR mass distribution indoors (%) as predicted by the ‘indoor SMURF’ model

Due to the model’s assumption for emission to air, ventilation is a significant

removal mechanism over the entire KOA range; even for very high KOA FRs

such as BDE-209 (KOA = 16.5) or TBBPA-BDBPE (KOA = 20.7) as much as 40% of

the emissions is removed as particle-associated mass to outdoors. In line

Indoor medium

KΟΑ range Air Vertical surfaces

Horizontal surfaces

Compound example

<7.2 100% - 80% 0% - 13% 0% - 7% DBS, TEP, TIBP, TMP, TPP, TDBPP

8 - 8.4 40% - 20% 40% - 50% 20% - 30% TBECH, DBNPG, DBP, TBCO,

TBNPA, TBP-AE, DOPO, TNBP

8.4 - 9 20% - 6% 50% - 60% 30% - 34% PBT, TBP, TBX, TPHP, TCEP, TCIPP

9 - 11 6% - 0.1% 60% - 75% 34% - 25% BDE-28, -47, HBB, PBBB, PBBC, PBEB, PBP, PBP-AE, DPTE, DCP,

TDCIPP, TDCPP

11 - 12 0.1% 75% - 60% 25% - 40% BDE-99, -100, EH-TBB, HBCDD,

PBBA, TBOEP, TEHP, TCP, BCMP-BCEP

12 - 12,4 0.1% 60% - 40% 40% - 60% BDE-153

12.4 - 12,9 0.1% 40% - 20% 60% - 80% BDE-154, PBDPP

>13,3 0.1% 10% - 0% 90% - 100% BDE-183, -209, TBPH, BTBPE,

DBDPE, TBBPA, TTBP-TAZ, BADP and 13 more

20

with previous modelling studies (Bennett and Furtaw, 2004; Zhang et al.,

2009), particle- and organic film-associated processes, including removal

ones, become increasingly important for indoor fate with increasing KOA. FRs

with KOA < 8 which are predominantly present in indoor air exhibit residence

times in the range of a few hours as they are removed fast by ventilation. For

higher KOA FRs, residence time increases significantly (several days to

months) as a result of the lower removal rates on indoor surfaces where they

distribute favourably.

A good knowledge about the likely behaviour and fate of organic

contaminants indoors can also assist to ascertain the relevant importance of

likely exposure pathways indoors. Similarly to what has been postulated for

the highly brominated PBDEs (Sahlstrom et al., 2014; Trudel et al., 2011) dust

may be the most significant exposure pathway for a large number of

alternative FRs with an alike indoor fate (see Table 2). A recent exposure

assessment by Zhang et al. (2014) suggested dust ingestion is the dominant

exposure pathway (for adults) for organic chemicals with KOA > 12, while

inhalation and/or dermal permeation (depending on properties other than

KOA) contribute the most to total exposure for chemicals with a KOA in the

range 8 - 11. Estimated adult intake rates by (Cequier et al., 2014) indicated

that air inhalation and dermal uptake are the main routes of exposure for

BDE-28, TBECH, TBP-AE, PBT, PBB, TNBP (KOA: 8.1 - 9.6) and TCEP, TBOEP,

TCP, TDCIPP (KOA: 8.5 - 11.9), respectively. In the same study, dust ingestion

was the main exposure route for higher KOA FRs. For some FRs though, the

relative significance of exposure routes differed when children exposure was

assessed.

4.4 Outdoor fate: persistence and long-range transport

Even though the indoor environment (residential or occupational) may be

the main source for many FRs, these will eventually reach outdoors where

they can persist, be transported over long distances and present a threat to

the environment. Environmental fate outdoors is largely driven by the

physicochemical properties (KOW, KAW, t1/2,air, t1/2,water and t1/2,soil) and the

environmental release mode. Figure 5 presents the Pov and LRTP of the

studied FRs as predicted by ‘The Tool’ (Paper II). Our analysis suggests that

many of the alternative FRs, some introduced or considered as potential

replacements to PBDEs, exhibit similar POP-like behaviour in the

environment (i.e. high persistence and medium to high LRTP). High

21

persistence and hydrophobicity (high KOA) indicates that many of these

alternative BFRs could be bioaccumulative. On that basis they cannot be

viewed as suitable replacements. Although a number of low molecular

weight alternative BFRs and NHOPFRs appear better based on Pov and LRTP

criteria, results must be interpreted with caution given the significant

uncertainty in physicochemical properties as well as field observations that

contradict model predictions (i.e. (Moller et al., 2011; Salamova et al., 2014).

At the same time other criteria such as bioaccumulation potential, toxicity of

both parent compounds and possible metabolites along with functionality (in

an industrial substitution strategy) have to be considered. Nevertheless, the

high persistence and environmental mobility of some of the alternative FRs

can in itself be considered problematic. Even if emissions cease in the future,

it will likely take decades to reverse the global contamination of alternative

FRs (Scheringer, 2002).

Figure 5. POV versus CTD for the studied FRs as predicted by ‘The Tool’. The 2

perpendicular lines designate a region of POP-like chemicals, with high POV and LRTP

(upper left quadrant) and a low environmental concern region, with low POV and

LRTP. The thick black line defines the maximum LRTP that is possible for a given POV.

1,0E+00

1,0E+01

1,0E+02

1,0E+03

1,0E+04

1,0E+05

1,0E+06

1,0E+07

1,E-01 1,E+00 1,E+01 1,E+02 1,E+03 1,E+04 1,E+05 1,E+06

LRTP

(C

har

acte

rist

ic T

rave

l Dis

tan

ce,

km)

POV (days)

BDE-47 to -209, BTBPE,

TBBPA, DBDPE, PBEB,

PBP, TTBNPP, TTBP-TAZ

BCMP-BCEP, DBHCTD,

HBCYD, PBBA, PBP-AE,

TBBPS, DTPE, TDCIPP

PBT, HBB, TBX

DBNPG, TBOEP,

TEHP, TEP, TIBP,

TMP, TNBP

HEEHP-TEBP, TBP-AE,

TCEP, TCIPP, TDBPP

BDE-28, TBPH, BADP,

TBECH, DBP, DBS, DCP,

DOPO, EH-TBB, HBCDD,

RDP, TBCO, TBNPA, TBP,

TDB-TAZTO, TIPPP, TCP,

TPHP

22

5. Conclusions and future perspectives

Over the last decade an increasing number of studies have reported the

occurrence of FRs in the indoor environment. Although research initially

focused on the PBDEs, attention has gradually shifted to alternative FRs,

whose use has increased as a result of legislative actions on PBDEs.

Understanding how and to what extent FRs are released from indoor sources,

and their resulting fate and exposure behaviour indoors is essential for

predicting and evaluating the risks associated with their indoor occurrence.

An improved understanding will also provide vital information on the

necessary measures required to minimise risks associated with FRs.

In Paper I we identified a number of data gaps and limitations in the

current (as of early 2013) understanding and the available modelling

approaches of indoor emissions and fate of FRs (Hypothesis I). These include

a limited understanding of the key emission mechanisms for low volatility

FRs, uncertainties regarding indoor air/surface partitioning, poor

characterization of dust and film dynamics and a significant lack of

knowledge regarding indoor reaction/degradation processes. In addition, in

Paper II we demonstrated a serious scarcity in physicochemical property

data for the alternative FRs which can hinder the implementation of

comprehensive fate and exposure assessments for these alternatives

(Hypothesis I).

The most important contributions of this thesis in addressing the above

limitations are:

The ‘best guess’ estimation technique can be applied to evaluate the

plausibility of available physicochemical properties and select

reliable estimates required for chemical assessments, especially

when limited or no high quality measured property data are available

(Paper II).

An indication of the likely fate indoors of a large number of ‘novel’

and ‘emerging’ BFRs and OPFRs was provided for the first time. With

a log KOA > 8.5, most of the alternative FRs (mainly the alternative

BFRs and the HOPFRs) exhibit a strong partitioning to indoor surfaces

and dust, similarly to the PBDEs (Hypothesis II) (Paper II).

23

More than half of the alternative BFRs and NHOPFRs exhibit high POV

and medium to high LRTP outdoors. Their high estimated KOW in

combination with the high persistence also indicates a

bioaccumulative potential further increasing concern. On that basis

these alternatives cannot be regarded as suitable replacements to

PBDEs (Hypothesis II) (Paper II).

An inverse modelling approach is a practical and effective way to

estimate indoor emissions and identify indications of emission-to-

dust (Hypothesis III). Rough estimations of the emissions can also be

made based on experimental emission factors; although, a much

higher uncertainty is to be expected for these empirical emission

estimates they compared well to model-estimated emission rates

(Paper III).

Median estimated emissions of OPFRs in Norwegian households are

0.083 μg.h-1 (air-based) and 0.41 μg.h-1 (dust-based) which is 2 to 3

orders of magnitude higher than corresponding BFR emissions (0.52

and 0.32 ng.h-1, respectively). The difference between air- and dust-

based estimates for OPFRs may be evidence for a direct migration-

to-dust pathway (Hypothesis IV) (Paper III).

A direct emission-to-dust mechanism may be important for

controlling the indoor fate of FRs regardless of their relative volatility

and this mechanism is more likely abrasion rather than solid or

liquid-phase diffusion between source and dust in direct contact.

(Papers III & IV) (Hypothesis IV). Direct source-dust contact may,

however, result in higher concentrations in dust sitting on a source

due to an enhanced partitioning from the gas-phase (Paper IV).

The knowledge gained in this research regarding the indoor emissions and

fate of FRs provides useful perspectives regarding the direction of future

research in this area. First and foremost, there is an imperative need for

reliable physicochemical property measurements for many of the alternative

FRs. These property data, in combination with QSPR methods for

reliability/consistency checks (i.e. the ‘best guess’ estimation (Stieger et al.,

2014) or the ‘three solubility’ approach (Cousins and Mackay, 2000)) will

assist in the implementation of more solid environmental fate and exposure

assessments for those chemicals.

24

Another major challenge to be addressed in future studies is migration from

FR sources to dust through abrasion and direct source-dust contact, which

has only recently been investigated including in this thesis. Despite the

findings of Paper IV, results from other chamber and indoor monitoring

studies suggest that those are important chemical release pathways,

meaning they have to be studied more systematically and the underlying

mechanism(s) elucidated. As to chemical transfer between a treated product

and dust in direct contact, this can be further investigated in specially

designed test chamber experiments using multiple time points, air flow rates,

different dust properties test chemicals with a range of physicochemical

properties, and multiple products. A mechanistic description of the transfer

mechanism could be possibly added to an indoor fate model to account for

its effect on indoor fate and exposure. Developing a realistic emission

scenario, however, could be challenging given the presence of multiple point

sources and the inhomogeneity of dust indoors.

Since the beginning of this research, other valuable contributions towards

a better understanding of the indoor fate of SVOCs have been made. Some

of the gaps and limitations identified in Paper I have been also addressed.

For example, the magnitude and kinetics of gas-phase/surface partitioning of

SVOCs for a variety of common indoor materials (i.e. dust, plastics, concrete,

wood, glass, plates, cotton clothing) was experimentally studied in a recent

series of test-chamber/house studies (Bi et al., 2015; Cao et al., 2016; Liang

and Xu, 2015; Liu et al., 2014). The sorption related parameters, surface/air

partition coefficients (Ks) and solid-phase diffusion coefficients (Dm) that

were determined in those studies can be useful for modelling indoor fate and

exposure. Updated indoor models have also been developed. Shin et al.

(2012) incorporated a particle mass-balance in their model, whereas Zhang

et al. (2014) coupled a human exposure module to an indoor fate model and

accounted for the effect of human intake on the chemical mass balance.

Increasing model complexity to capture the complexity of indoor

environmental characteristics/conditions is not per se problematic.

However, it always needs to take into account the resulting uncertainty. This

is extremely important given that some of the key (sensitive) model

parameters remain highly uncertain or variable (i.e. reaction rates on indoor

surfaces, dust removal rates, half-lives in humans). Finally and most

importantly, current indoor fate and exposure models need to be evaluated

systematically against real-time measurements. The inverse modelling

approach adopted in Paper III is only a first step towards testing the

applicability of indoor fate models. A possible further second step would be

25

to couple a human exposure model to an indoor fate model to determine the

contribution of indoor exposure to the total FR exposure. This could be

achieved by comparing model calculated body burdens to human

biomonitoring data.

In a broader context, this thesis raises concerns about the substitution of

the problematic PBDEs with newer substances that exhibit similar POP-like

properties. Unfortunately, scientific research lags behind industry when it

comes to the introduction of new chemicals since many years or even

decades elapse until the weight of evidence for harmful effects is enough to

warrant legislative action. The manufacturing industry is required under the

European Union REACH regulations to provide hazard information on the

chemicals that they produce, but we found that there are still large data gaps

on many emerging and novel FRs that need to be filled before their risks can

be properly evaluated. By the time these data gaps are filled, however, it

could be too late as the contamination may not be readily reversible in the

environment (Scheringer, 2002). Ideally the manufacturing industry should

provide more detailed data on emerging and novel FRs to the public and

strive to publish the information in peer-reviewed journals. If this occurred

the fate, exposure, effects and risks could be properly assessed at an earlier

stage and help avoid the pitfalls of the past. Green Chemistry principles (e.g.

“benign by design”) should also be adopted by the manufacturing industry

(Anastas and Warner, 2000), although the required functionality (e.g. low

reactivity of a chemical in the product) sometimes hinders the adoption of

these approaches. A proactive hazard identification strategy along with the

precautionary principle would provide good management practice for

confronting the threats from chemical pollution (MacLeod et al., 2014;

Martuzzi and Joel, 2004; Persson et al., 2013).

26

Acknowledgements

Many people have helped me along the way with their scholarship and friendship to make this journey happen.

To start with, this thesis could not have been concluded without the invaluable support of my supervisors, Ian Cousins and Anna Palm Cousins. Thank you for the scientific discussions, the continuous encouragement and prompt feedback; and for putting me back on the right track when I was disorientated.

My gratitude also goes to Matthew MacLeod and Peter Tunved, for their insightful comments on the thesis' draft that have strengthened my arguments.

Being a part of the INFLAME project (EU 7th Framework Programme under grant agreement No. 264600) and its people has been a unique experience. I feel honoured to have had the opportunity to meet with leading academics in the field and talented researchers all over the world, with whom I have shared memorable moments.

I feel particularly lucky to have collaborated with Boris Lazarov in Paper III & IV, and Enrique Cequier in Paper III, who are two intelligent and fine colleagues that made work enjoyable. I would also like to thank their supervisors, Catherine Thomsen (Paper III), Marianne Stranger (Paper III), for their engagement in co-authoring these papers.

During my time at the Swedish Environmental Research Institute IVL, I got to meet some great people, who supported me with their knowledge and kindness. Special thanks goes to my good friend and invaluable colleague, Giorgos Giovanoulis, with whom I also co-authored the stressful Paper IV, and Thuy Buy, for being a friend and sports buddy.

The magic former and current ACES/ITM team: Stathis, Marko, Seth, Melissa, Damien, Deguo, Raed, Steffen, Kim, Robin, Kerstin, Li and everyone else I might be forgetting thank you all for the discussions (scientific or other), the laughs and the awesome parties. What a pleasure it has been to work in such a great environment! Xu, thank you for being an excellent host during our stay in China and for your legendary quotes. Stella, our long corridor conversations are definitely to be remembered!

Words are not enough to describe great friendships as the one we shared with Fiona, Wouter (Akinori), Dimitri and Hongyan (aka 'The Six Malakeers'). Not only have I found in you amazing colleagues, but also a second family. Thank you for the amazing road trips and card games; the cooking sessions

27

and movie nights; and the countless laughing moments. And thank you all, Gotland team, for our last (but not final) great adventure!

Thank you Evi for being by my side through the good and the rough times. Your invaluable support and experience (it hasn’t been long since you tamed your own dragon) made this journey easier.

Finally, I would like to thank my parents for their unconditional love and continuous encouragement and support.

28

References

Alaee M, Arias P, Sjodin A, Bergman A. An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 2003; 29: 683-9.

Alaee M, Wenning RJ. The significance of brominated flame retardants in the environment: current understanding, issues and challenges. Chemosphere 2002; 46: 579-82.

Ali N, Dirtu AC, Van den Eede N, Goosey E, Harrad S, Neels H, et al. Occurrence of alternative flame retardants in indoor dust from New Zealand: indoor sources and human exposure assessment. Chemosphere 2012; 88: 1276-82.

Allen JG, McClean MD, Stapleton HM, Webster TF. Linking PBDEs in House Dust to Consumer Products using X-ray Fluorescence. Environ. Sc.i Technol. 2008; 42: 4222-4228.

Anastas TP, Warner CJ. Green Chemistry: Theory and Practice: Oxford University Press, 2000.

Araki A, Saito I, Kanazawa A, Morimoto K, Nakayama K, Shibata E, et al. Phosphorus flame retardants in indoor dust and their relation to asthma and allergies of inhabitants. Indoor Air 2014; 24: 3-15.

Arnot J, Gouin T, Mackay D. Development and Application of Models of Chemical Fate in Canada Practical Methods for Estimating Environmental Biodegradation Rates. Report No. 200503. Canadian Environmental Modelling Network Trent University, Canada, 2005.

Aronson D, Boethling R, Howard P, Stiteler W. Estimating biodegradation half-lives for use in chemical screening. Chemosphere 2006; 63: 1953-1960.

Arp HPH, Droge STJ, Endo S, Goss K-U, Giger W, Hawthorne SB, et al. Response to Comment on “More of EPA’s SPARC Online Calculator—The Need for High Quality Predictions of Chemical Properties”. Environ. Sci. Technol. 2010; 44: 7746-7747.

Arp HPH, Niederer C, Goss K-U. Predicting the Partitioning Behavior of Various Highly Fluorinated Compounds. Environ. Sci. Technol. 2006; 40: 7298-7304.

Batterman SA, Chernyak S, Jia C, Godwin C, Charles S. Concentrations and Emissions of Polybrominated Diphenyl Ethers from U.S. Houses and Garages. Environ. Sci. Technol. 2009; 43: 2693-2700.

Beard A, Klimes M. Searching for safe Flame Retardants – an update on regulatory status and environmental assessments. Safer Products Summit, April 2013, San Francisco, California, USA.

29

Bennett DH, Furtaw EJ, Jr. Fugacity-based indoor residential pesticide fate model. Environ. Sci. Technol. 2004; 38: 2142-52.

Bergh C, Torgrip R, Emenius G, Ostman C. Organophosphate and phthalate esters in air and settled dust - a multi-location indoor study. Indoor Air 2011; 21: 67-76.

Bergman A, Ryden A, Law RJ, de Boer J, Covaci A, Alaee M, et al. A novel abbreviation standard for organobromine, organochlorine and organophosphorus flame retardants and some characteristics of the chemicals. Environ. Int. 2012; 49: 57-82.

Bi C, Liang Y, Xu Y. Fate and Transport of Phthalates in Indoor Environments and the Influence of Temperature: A Case Study in a Test House. Environ. Sci. Technol. 2015; 49: 9674-9681.

Björklund JA, Thuresson K, Cousins AP, Sellström U, Emenius G, de Wit CA. Indoor Air Is a Significant Source of Tri-decabrominated Diphenyl Ethers to Outdoor Air via Ventilation Systems. Environ. Sci. Technol. 2012; 46: 5876-5884.

Brandsma SH, de Boer J, van Velzen MJ, Leonards PE. Organophosphorus flame retardants (PFRs) and plasticizers in house and car dust and the influence of electronic equipment. Chemosphere 2014; 116: 3-9.

BSEF, at http://www.bsef.com (accessed in March 2014). Buser AM, MacLeod M, Scheringer M, Mackay D, Bonnell M, Russell MH, et

al. Good modeling practice guidelines for applying multimedia models in chemical assessments. Integrated Environmental Assessment and Management 2012; 8: 703-708.

Cao J, Weschler CJ, Luo J, Zhang Y. Cm-History Method, a Novel Approach to Simultaneously Measure Source and Sink Parameters Important for Estimating Indoor Exposures to Phthalates. Environ. Sci. Technol. 2016; 50: 825-834.

Cequier E, Ionas AC, Covaci A, Maria Marce R, Becher G, Thomsen C. Occurrence of a Broad Range of Legacy and Emerging Flame Retardants in Indoor Environments in Norway. Environ. Sci. Technol. 2014; 48: 6827-6835.

Cequier E, Sakhi AK, Marce RM, Becher G, Thomsen C. Human exposure pathways to organophosphate triesters - A biomonitoring study of mother-child pairs. Environ. Int. 2015; 75: 159-165.

Cousins AP. The effect of the indoor environment on the fate of organic chemicals in the urban landscape. Sci. Total Environ. 2012; 438: 233-41.

Cousins I, Mackay D. Correlating the physical–chemical properties of phthalate esters using the `three solubility' approach. Chemosphere 2000; 41: 1389-1399.

30

Covaci A, Harrad S, Abdallah MA, Ali N, Law RJ, Herzke D, et al. Novel brominated flame retardants: a review of their analysis, environmental fate and behaviour. Environ. Int. 2011; 37: 532-56.

Darnerud PO. Toxic effects of brominated flame retardants in man and in wildlife. Environ. Int. 2003; 29: 841-53.

Darnerud PO. Brominated flame retardants as possible endocrine disrupters. Int. J. Androl. 2008; 31: 152-60.

de Wit CA. An overview of brominated flame retardants in the environment. Chemosphere 2002; 46: 583-624.

de Wit CA, Bjorklund JA, Thuresson K. Tri-decabrominated diphenyl ethers and hexabromocyclododecane in indoor air and dust from Stockholm microenvironments 2: indoor sources and human exposure. Environ. Int. 2012; 39: 141-7.

de Wit CA, Herzke D, Vorkamp K. Brominated flame retardants in the Arctic environment--trends and new candidates. Sci. Total Environ. 2010; 408: 2885-918.

EC. European Union Risk Assessment Report, TRIS (2-CHLOROETHYL) PHOSPHATE, TCEP, 2009.

EFRA, at http://www.cefic-efra.com/ (accessed in January 2016). EFSA. Scientific Opinion on Emerging and Novel Brominated Flame

Retardants (BFRs) in Food. EFSA Journal 2012; 10: 2908. Eskenazi B, Chevrier J, Rauch SA, Kogut K, Harley KG, Johnson C, et al. In utero

and childhood polybrominated diphenyl ether (PBDE) exposures and neurodevelopment in the CHAMACOS study. Environ. Health Perspect. 2013; 121: 257-62.

Fromme H, Lahrz T, Kraft M, Fembacher L, Mach C, Dietrich S, et al. Organophosphate flame retardants and plasticizers in the air and dust in German daycare centers and human biomonitoring in visiting children (LUPE 3). Environ. Int. 2014; 71: 158-63.

Gascon M, Vrijheid M, Martinez D, Forns J, Grimalt JO, Torrent M, et al. Effects of pre and postnatal exposure to low levels of polybromodiphenyl ethers on neurodevelopment and thyroid hormone levels at 4 years of age. Environ. Int. 2011; 37: 605-11.

Gouin T, Cousins I, Mackay D. Comparison of two methods for obtaining degradation half-lives. Chemosphere 2004; 56: 531-535.

Guo Z. A framework for modelling non-steady-state concentrations of semivolatile organic compounds indoors – I: Emissions from diffusional sources and sorption by interior surfaces. Indoor and Built Environment 2013.

Harju M, Heimstad E, Herzke D, Sandanger T, Posner S, Wania F. Current state of knowledge and monitoring requirements -Emerging “new”

31

Brominated flame retardants in flame retarded products and the environment. Norwegian Pollution Control Authority, Norway, 2009.

Harrad S, de Wit CA, Abdallah MA-E, Bergh C, Björklund JA, Covaci A, et al. Indoor Contamination with Hexabromocyclododecanes, Polybrominated Diphenyl Ethers, and Perfluoroalkyl Compounds: An Important Exposure Pathway for People? Environ. Sci. Technol. 2010; 44: 3221-3231.

Harrad S, Wijesekera R, Hunter S, Halliwell C, Baker R. Preliminary Assessment of U.K. Human Dietary and Inhalation Exposure to Polybrominated Diphenyl Ethers. Environ. Sci. Technol. 2004; 38: 2345-2350.

Hilal SH, Karickhoff SW, Carreira LA. Prediction of the Solubility, Activity Coefficient and Liquid/Liquid Partition Coefficient of Organic Compounds. QSAR Comb. Sci. 2004; 23: 709-720.

Hites RA. Polybrominated Diphenyl Ethers in the Environment and in People:  A Meta-Analysis of Concentrations. Environ. Sci. Technol. 2004; 38: 945-956.

Holmgren T, Persson L, Andersson PL, Haglund P. A generic emission model to predict release of organic substances from materials in consumer goods. Sci. Total Environ. 2012; 437: 306-14.

Johnson PI, Stapleton HM, Mukherjee B, Hauser R, Meeker JD. Associations between brominated flame retardants in house dust and hormone levels in men. Sci. Total Environ. 2013; 445-446: 177-84.

Kemmlein S, Hahn O, Jann O. Emissions of organophosphate and brominated flame retardants from selected consumer products and building materials. Atmos. Environ. 2003; 37: 5485-5493.

Klepeis NE, Nelson WC, Ott WR, Robinson JP, Tsang AM, Switzer P, et al. The National Human Activity Pattern Survey (NHAPS): a resource for assessing exposure to environmental pollutants. J. Expo. Anal. Environ Epidemiol. 2001; 11: 231-52.

Kuramochi H, Takigami H, Scheringer M, Sakai S. Estimation of physicochemical properties of 52 non-PBDE brominated flame retardants and evaluation of their overall persistence and long-range transport potential. Sci. Total Environ. 2014; 491-492: 108-17.

Law RJ, Allchin CR, de Boer J, Covaci A, Herzke D, Lepom P, et al. Levels and trends of brominated flame retardants in the European environment. Chemosphere 2006; 64: 187-208.

Law RJ, Covaci A, Harrad S, Herzke D, Abdallah MA, Fernie K, et al. Levels and trends of PBDEs and HBCDs in the global environment: status at the end of 2012. Environ. Int. 2014; 65: 147-58.

Leech JA, Nelson WC, Burnett RT, Aaron S, Raizenne ME. It's about time: A comparison of Canadian and American time–activity patterns.

32

Journal of Exposure Analysis & Environmental Epidemiology 2002; 12: 427.

Liang Y, Xu Y. The influence of surface sorption and air flow rate on phthalate emissions from vinyl flooring: Measurement and modeling. Atmos. Environ. 2015; 103: 147-155.

Liu X, Guo Z, Roache NF. Experimental method development for estimating solid-phase diffusion coefficients and material/air partition coefficients of SVOCs. Atmos. Environ. 2014; 89: 76-84.

Lorber M. Exposure of Americans to polybrominated diphenyl ethers. J. Expos. Sci. Environ. Epidemiol. 2007; 18: 2-19.

Mackay D. Multimedia environmental models: the fugacity approach: CRC press, 2001.

Mackay D, Shiu W-Y, Ma K-C, Lee SC. Handbook of physical-chemical properties and environmental fate for organic chemicals: CRC press, 2006.

MacLeod M, Breitholtz M, Cousins IT, de Wit CA, Persson LM, Rudén C, et al. Identifying Chemicals That Are Planetary Boundary Threats. Environ. Sci. Technol. 2014.

MacLeod M, Fraser AJ, Mackay D. Evaluating and expressing the propagation of uncertainty in chemical fate and bioaccumulation models. Environ. Toxicol. Chem. 2002; 21: 700-9.

Martuzzi M, Joel TA. The precautionary principle: protecting public health, the environment and the future of our children. WHO, Copenhagen, 2004.

Meeker JD, Johnson PI, Camann D, Hauser R. Polybrominated diphenyl ether (PBDE) concentrations in house dust are related to hormone levels in men. Sci. Total Environ. 2009; 407: 3425-3429.

Meeker JD, Stapleton HM. House dust concentrations of organophosphate flame retardants in relation to hormone levels and semen quality parameters. Environ. Health Perspect. 2010; 118: 318-23.

Moller A, Sturm R, Xie Z, Cai M, He J, Ebinghaus R. Organophosphorus flame retardants and plasticizers in airborne particles over the Northern Pacific and Indian Ocean toward the Polar Regions: evidence for global occurrence. Environ. Sci. Technol. 2012; 46: 3127-34.

Moller A, Xie Z, Caba A, Sturm R, Ebinghaus R. Organophosphorus flame retardants and plasticizers in the atmosphere of the North Sea. Environ. Pollut. 2011; 159: 3660-5.

Muenhor D, Harrad S. Within-room and within-building temporal and spatial variations in concentrations of polybrominated diphenyl ethers (PBDEs) in indoor dust. Environ. Int. 2012; 47: 23-27.

33

Newton S, Sellstrom U, de Wit CA. Emerging flame retardants, PBDEs, and HBCDDs in indoor and outdoor media in Stockholm, Sweden. Environ. Sci. Technol. 2015; 49: 2912-20.

Palm A, Cousins IT, Mackay D, Tysklind M, Metcalfe C, Alaee M. Assessing the environmental fate of chemicals of emerging concern: a case study of the polybrominated diphenyl ethers. Environ. Pollut. 2002; 117: 195-213.

Persson LM, Breitholtz M, Cousins IT, de Wit CA, MacLeod M, McLachlan MS. Confronting Unknown Planetary Boundary Threats from Chemical Pollution. Environ. Sci. Technol. 2013; 47: 12619-12622.

Rauert C, Harrad S. Mass transfer of PBDEs from plastic TV casing to indoor dust via three migration pathways — A test chamber investigation. Sci. Total Environ. 2015; 536: 568-574.

Rauert C, Harrad S, Suzuki G, Takigami H, Uchida N, Takata K. Test chamber and forensic microscopy investigation of the transfer of brominated flame retardants into indoor dust via abrasion of source materials. Sci. Total Environ. 2014a; 493: 639-648.

Rauert C, Lazarov B, Harrad S, Covaci A, Stranger M. A review of chamber experiments for determining specific emission rates and investigating migration pathways of flame retardants. Atmos. Environ. 2014b; 82: 44-55.

Roosens L, Abdallah MA-E, Harrad S, Neels H, Covaci A. Exposure to Hexabromocyclododecanes (HBCDs) via Dust Ingestion, but Not Diet, Correlates with Concentrations in Human Serum: Preliminary Results. Environ. Health Perspect. 2009; 117: 1707-1712.

Sahlstrom LMO, Sellstrom U, de Wit CA, Lignell S, Darnerud PO. Brominated Flame Retardants in Matched Serum Samples from Swedish First-Time Mothers and Their Toddlers. Environ. Sci. Technol. 2014; 48: 7584-7592.

Saito I, Onuki A, Seto H. Indoor organophosphate and polybrominated flame retardants in Tokyo. Indoor Air 2007; 17: 28-36.

Salamova A, Hermanson MH, Hites RA. Organophosphate and Halogenated Flame Retardants in Atmospheric Particles from a European Arctic Site. Environ. Sci. Technol. 2014; 48: 6133-6140.

Schenker U, MacLeod M, Scheringer M, Hungerbühler K. Improving Data Quality for Environmental Fate Models:  A Least-Squares Adjustment Procedure for Harmonizing Physicochemical Properties of Organic Compounds. Environ. Sci. Technol. 2005; 39: 8434-8441.

Scheringer M. Persistence and Spatial Range of Environmental Chemicals: New Ethical and Scientific Concepts for Risk Assessment: Wiley-VCH, 2002.

34

Schripp T, Fauck C, Salthammer T. Chamber studies on mass-transfer of di(2-ethylhexyl)phthalate (DEHP) and di-n-butylphthalate (DnBP) from emission sources into house dust. Atmos. Environ. 2010; 44: 2840-2845.

Schweizer C, Edwards RD, Bayer-Oglesby L, Gauderman WJ, Ilacqua V, Juhani Jantunen M, et al. Indoor time–microenvironment–activity patterns in seven regions of Europe. J. Expo. Sci. Environ. Epidemiol. 2007; 17: 170-181.

Shin H-M, McKone TE, Tulve NS, Clifton MS, Bennett DH. Indoor Residence Times of Semivolatile Organic Compounds: Model Estimation and Field Evaluation. Environ. Sci. Technol. 2012; 47: 859-867.

Shin HM, McKone TE, Nishioka MG, Fallin MD, Croen LA, Hertz-Picciotto I, et al. Determining source strength of semivolatile organic compounds using measured concentrations in indoor dust. Indoor Air 2014; 24: 260-71.

Shoeib M, Harner T, Webster GM, Sverko E, Cheng Y. Legacy and current-use flame retardants in house dust from Vancouver, Canada. Environ. Pollut. 2012; 169: 175-82.

Sjodin A, Papke O, McGahee E, Focant JF, Jones RS, Pless-Mulloli T, et al. Concentration of polybrominated diphenyl ethers (PBDEs) in household dust from various countries. Chemosphere 2008; 73: S131-6.

Stieger G, Scheringer M, Ng CA, Hungerbuhler K. Assessing the persistence, bioaccumulation potential and toxicity of brominated flame retardants: Data availability and quality for 36 alternative brominated flame retardants. Chemosphere 2014.

Stuart H, Ibarra C, Abdallah MA, Boon R, Neels H, Covaci A. Concentrations of brominated flame retardants in dust from United Kingdom cars, homes, and offices: causes of variability and implications for human exposure. Environ. Int. 2008; 34: 1170-5.

Suzuki G, Kida A, Sakai S-i, Takigami H. Existence State of Bromine as an Indicator of the Source of Brominated Flame Retardants in Indoor Dust. Environ. Sci. Technol. 2009; 43: 1437-1442.

Takigami H, Suzuki G, Hirai Y, Sakai S. Transfer of brominated flame retardants from components into dust inside television cabinets. Chemosphere 2008; 73: 161-9.

Trudel D, Scheringer M, von Goetz N, Hungerbuhler K. Total consumer exposure to polybrominated diphenyl ethers in North America and Europe. Environ. Sci. Technol. 2011; 45: 2391-7.

UNEP. Report of the Conference of the Parties of the Stockholm Convention on Persistent Organic Pollutants on the work of its fourth meeting, 2009.

35

UNEP. Report of the Conference of the Parties to the Stockholm Convention on Persistent Organic Pollutants on the work of its sixth meeting, 2013.

UNEP. Report of the Persistent Organic Pollutants Review Committee on the work of its tenth meeting. Risk profile on decabromodiphenyl ether (commercial mixture, c-decaBDE), 2014.

USEPA. Estimation Programs Interface Suite™ for Microsoft® Windows, v.4.1. United States Environmental Protection Agency, Washington, DC, USA, 2012.

van der Veen I, de Boer J. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis. Chemosphere 2012; 88: 1119-53.

Wang Y, Jiang G, Lam PK, Li A. Polybrominated diphenyl ether in the East Asian environment: a critical review. Environ. Int. 2007; 33: 963-73.

Watkins DJ, McClean MD, Fraser AJ, Weinberg J, Stapleton HM, Sjodin A, et al. Impact of dust from multiple microenvironments and diet on PentaBDE body burden. Environ. Sci. Technol. 2012; 46: 1192-200.

Wegmann F, Cavin L, MacLeod M, Scheringer M, Hungerbühler K. The OECD software tool for screening chemicals for persistence and long-range transport potential. Environ. Model Softw. 2009; 24: 228-237.

Wensing M, Uhde E, Salthammer T. Plastics additives in the indoor environment—flame retardants and plasticizers. Sci. Total Environ. 2005; 339: 19-40.

Weschler CJ, Nazaroff WW. SVOC partitioning between the gas phase and settled dust indoors. Atmos. Environ. 2010; 44: 3609-3620.

Whitehead TP, Brown FR, Metayer C, Park J-S, Does M, Petreas MX, et al. Polybrominated diphenyl ethers in residential dust: Sources of variability. Environ. Int. 2013; 57–58: 11-24.

Xu Y, Hubal EA, Clausen PA, Little JC. Predicting residential exposure to phthalate plasticizer emitted from vinyl flooring: a mechanistic analysis. Environ. Sci. Technol. 2009; 43: 2374-80.

Xu Y, Little JC. Predicting emissions of SVOCs from polymeric materials and their interaction with airborne particles. Environ. Sci. Technol. 2006; 40: 456-61.

Zhang X, Arnot JA, Wania F. Model for screening-level assessment of near-field human exposure to neutral organic chemicals released indoors. Environ. Sci. Technol. 2014; 48: 12312-9.

Zhang X, Brown TN, Wania F, Heimstad ES, Goss KU. Assessment of chemical screening outcomes based on different partitioning property estimation methods. Environ. Int. 2010; 36: 514-20.

36

Zhang X, Diamond ML, Ibarra C, Harrad S. Multimedia Modeling of Polybrominated Diphenyl Ether Emissions and Fate Indoors. Environ. Sci. Technol. 2009; 43: 2845-2850.

Zhang X, Diamond ML, Robson M, Harrad S. Sources, Emissions, and Fate of Polybrominated Diphenyl Ethers and Polychlorinated Biphenyls Indoors in Toronto, Canada. Environ. Sci. Technol. 2011; 45: 3268-3274.

Zhang X, Suhring R, Serodio D, Bonnell M, Sundin N, Diamond ML. Novel flame retardants: Estimating the physical-chemical properties and environmental fate of 94 halogenated and organophosphate PBDE replacements. Chemosphere 2016; 144: 2401-7.