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MSc Analytical Sciences

Chemistry

Master Thesis

Determination of aryl-PFRs in indoor dust from different

microenvironments in Spain and the Netherlands and assessment of

human exposure

By

Maria K. Björnsdotter

UvA: 11166835

VU: 2574756

June 2017

42 EC

Period 2 to 6

Supervisor/Examiner: Examiner:

Dr. A. Ballesteros-Gómez Dr. W.T. Kok

Dr. H. Lingeman

Department of Analytical Chemistry

Institute of Fine Chemistry and Nanochemistry

University of Córdoba, Spain

Abstract

Phosphate flame retardants (PFRs) are ubiquitous chemicals in the indoor environment. Among

them, aryl-PFRs, such as triphenyl phosphate (TPHP), are frequently detected in indoor dust,

which is an important route for human exposure to these contaminants. TPHP is an aryl-PFR

and a plasticizer that is widely used in electric and electronic equipment. It has been shown to

migrate from materials resulting in environmental contamination and it has been detected in

indoor dust worldwide. Diphenyl phosphate (DPHP), the hydrolyzed metabolite of TPHP, has

been used as a biomarker for monitoring the human exposure to TPHP. However, a lack of

correlation between the levels of TPHP in indoor dust and DPHP in urine has been observed

up to date. The high urinary concentrations of DPHP suggests additional sources of TPHP and

DPHP and/or other aryl-PFRs that could also be metabolized into DPHP. In this study, DPHP

(and TPHP) are measured in indoor dust in samples collected in the Netherlands (n=23) in June

2016 and in Spain (n=57) in March and April 2017 using liquid chromatography coupled with

triple-quadrupole mass spectrometry (LC-QQQ-MS/MS). A suitable extraction/clean-up

method based on salting-out extraction followed by dispersive solid phase extraction (SPE)

was optimized and employed for this purpose. Additionally, the presence of other emerging

aryl-PFRs was monitored by target screening of the samples.

TPHP and DPHP were present in all samples analyzed from Spain and the Netherlands (n=80)

in the range 169-142459 ng/g and 106-79661 ng/g, respectively. The DPHP concentrations

were strongly correlated to the TPHP concentrations (r=0.90, p<0.01). Estimated exposures for

adults and toddlers in Spain to TPHP via dust ingestion were much lower than the reference

dose, indicating no current health risk to the Spanish population. The estimated urinary DPHP

levels for adults and toddlers in Spain as a result of exposure to TPHP And DPHP via indoor

dust ingestion were too low to significantly contribute to the high urinary DPHP concentrations

reported in the literature, indicating that other sources of DPHP may play an essential role in

the urinary levels of DPHP. Other aryl-PFRs, namely Cresyl diphenyl phosphate (CDP),

resorcinol bis(diphenyl phosphate) (RDP), 2-Ethylhexyl diphenyl phosphate (EDP), Isodecyl

diphenyl phosphate (IDP) and Bisphenol A bis(diphenyl phosphate) (BADP), were all detected

in indoor dust, however, with lower frequency.

Table of Contents

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

1.1. Aryl-phosphate flame retardants (aryl-PFRs) ....................................................................................... 1

1.2. Triphenyl phosphate .............................................................................................................................. 2

1.3. Toxicity and environmental concern of triphenyl phosphate ................................................................ 3

1.4. Exposure sources and pathways ............................................................................................................ 3

1.5. Monitoring human exposure ................................................................................................................. 4

2. Experimental section ..................................................................................................................................... 5

2.1. Chemicals and reagents ......................................................................................................................... 5

2.2. Method optimization ............................................................................................................................. 5

2.3. Apparatus and sample analysis ............................................................................................................. 6

2.4. Sample collection and preparation ........................................................................................................ 7

2.5. Data processing ..................................................................................................................................... 8

Quantification of TPHP and DPHP in indoor dust ........................................................................................ 8

Statistics ........................................................................................................................................................ 8

Screening of aryl-phosphate flame retardants ............................................................................................... 9

3. Results and discussion .................................................................................................................................. 9

3.1. Method optimization ................................................................................................................................... 9

Salting-out phase separation ......................................................................................................................... 9

Sample preparation recoveries .................................................................................................................... 10

Column and LC gradient ............................................................................................................................. 12

3.2. TPHP and DPHP concentrations in indoor dust .................................................................................. 13

3.3. Correlation between TPHP and DPHP concentrations in indoor dust ................................................ 20

3.4. Estimated exposure to TPHP and DPHP in indoor dust ...................................................................... 22

3.5. Estimated urinary levels of DPHP ...................................................................................................... 25

3.6. Screening of aryl-phosphate flame retardants ..................................................................................... 26

4. Conclusions ................................................................................................................................................. 28

Acknowledgments ................................................................................................................................................ 30

References ............................................................................................................................................................ 31

1

1. Introduction

Flame retardants (FRs) are widespread in the environment due to their wide use in materials,

such as furniture, electronics and textiles, in order to prevent ignition and to slow down the

spread of an already initiated fire (EFRA, 2007). Concern has been raised considering their

migration from materials as it affects the indoor air quality and is a route for human exposure

(Kemmlein et al., 2003). Polybrominated diphenyl ethers (PBDEs) have been commonly used

FRs until recently, when their use in electrical and electronic equipment was restricted due to

their known toxicity, persistence and bioaccumulative properties (U.S. EPA, 2009). The

European Union has banned the use of pentaBDE and octaBDE in 2004 (Directive

2002/96/EC) and the use of decaBDE in electric and electronic equipment in 2009 (European

Court of Justice, 2008). This regulation has led to a phase-out of PBDEs in materials resulting

in an introduction of alternatives, such as aryl-phosphate flame retardants (aryl-PFRs), onto the

market. Studies have demonstrated an increase in the presence of alternative FRs in indoor

dust, for which toxicity is still uncharacterized, in conjunction with the decrease of PBDE

(Dodson et al., 2012; Tao et al., 2016).

1.1. Aryl-phosphate flame retardants (aryl-PFRs)

Phosphorus FRs (PFRs) include both inorganic and organic compounds and are widely used in

plastics, polyurethane foams, thermosets, coatings and textiles (EFRA, 2007). Organic PFRs,

commonly known as organophosphate FRs (OPFRs) mainly act by forming a polymeric

structure of phosphoric acid formed from the reaction of phosphates under high heat. This

layer, known as a char layer, shields from oxygen and prevents the formation of flammable

gases and thereby lowers the risk of an initiated fire to spread (Schmitt, 2007). One of the most

important groups of OPFRs are phosphate esters, which are mainly used as additive FRs in

polyvinyl chloride and engineering plastics commonly used is electronic equipment (EFRA,

2007). Phosphate esters are derivatives from phosphoric acid with one, two, or three substituted

groups (ATSDR) (Figure 1). The substituents might be aliphatic (alkyl-PFRs) or aromatic

(aryl-PFRs) and are in many cases identical, which is the case for triphenyl phosphate (TPHP),

a phosphate triester with three phenyl groups.

2

Figure 1. Phosphate esters, mono-, di-, and tri-substituted.

1.2. Triphenyl phosphate

Triphenyl phosphate (TPHP; CAS no. 115-86-6) (Figure 2) is an aryl-PFR mainly used as an

additive in polymer mixtures used in electronic enclosure applications (LCSP, 2005). The use

of TPHP has resulted in environmental contamination due to its migration from materials

(Kemmlein et al., 2003).

Figure 2. Molecular structure of triphenyl phosphate (TPHP).

TPHP has been reported in the indoor environment in indoor dust collected from the floors of

residences (<2-1798000 ng/g) (Garcia et al., 2007; Stapleton et al., 2009; Kanazawa et al.,

2010; Bergh et al., 2011; Van den Eede et al., 2011; Ali et al., 2012a; Ali et al., 2012b; Dirtu

et al., 2012; Dodson et al., 2012; Ali et al., 2013; Kim et al., 2013; Abdallah and Covaci, 2014;

Araki et al., 2014; Brandsma et al., 2014; Cequier et al., 2014; Fan et al., 2014; Tajima et al.,

2014; Brommer and Harrad, 2015; Hoffman et al., 2015; Mizouchi et al., 2015; Zheng et al.,

2015; Ali et al., 2016; Canbaz et al., 2016; Cristale et al., 2016; Harrad et al., 2016; He et al.,

2016; Wu et al., 2016; Kademoglou et al., 2017), in indoor dust from offices (11-50000 ng/g)

(Bergh et al., 2011; Abdallah and Covaci, 2014; Brommer and Harrad, 2015; Cristale et al.,

2016; Harrad et al., 2016; He et al., 2016; Wu et al., 2016; Kademoglou et al., 2017), in indoor

dust from schools and daycare centers (10-90000 ng/g) (Bergh et al., 2011; Cequier et al., 2014;

Fromme et al., 2014; Brommer and Harrad, 2015; Mizouchi et al., 2015; Cristale et al., 2016;

Wu et al., 2016). TPHP has also been reported in dust from cars (<2-170000 ng/g) (Ali et al.,

2013; Abdallah and Covaci, 2014; Brandsma et al., 2014; Brommer and Harrad, 2015; Ali et

P

O

OHR1

R2

P

O

OHR

OH

P

O

R3

R1

R2

P

O

OHOH

OH

Phosphoric acid Phosphate monoester Phosphate diester Phosphate triester

P

O

O

OO

TPHP

3

al., 2016; Harrad et al., 2016), and from dust collected from public microenvironments such as

shops, restaurants and supermarkets (14-34200 ng/g) (Van den Eede et al., 2011; Ali et al.,

2012b; Abdallah and Covaci, 2014; Cristale et al., 2016; He et al., 2016). TPHP has also been

reported in indoor air (0.19-5.7 ng/m3) (Björklund et al., 2004; Hartmann et al., 2004), in

outdoor air (0.003 ng/m3) (Wolschke et al., 2016), sewage water influent (76-290 ng/L) and

effluent (41-130 ng/L) and sewage sludge (52-320 ng/g dw) (Marklund et al., 2006), surface

water (<LOD-10.3 ng/L) (Bollmann et al., 2012), sediment (5.6-253 ng/g) (Giulivo et al., 2016;

Tan et al., 2016) and in fish (43-230 ng/g lw) (Giulivo et al., 2016; Matsukami et al., 2016).

Furthermore, TPHP has been reported associated with airborne particles over the global oceans

indicating a potential for long-range atmospheric transport towards the polar regions (Möller

et al., 2012).

1.3. Toxicity and environmental concern of triphenyl phosphate

The widespread occurrence of TPHP in the indoor- and outdoor environment has led to concern

regarding human health and the environment. The human toxicity of TPHP is considered low

to high according to a recent alternatives assessment report (U.S. EPA, 2014). This is based on

OncoLogic modeling studies showing a moderate potential for carcinogenicity and in vivo

studies indicating a high potential for repeated dose effects based on reduced body weight in

male rats administered TPHP in diet during four weeks (U.S. EPA, 2014). Furthermore, PFRs

including TPHP may be associated with altered hormone levels and decreased semen quality

in men (Meeker and Stapleton, 2010). The aquatic toxicity of TPHP is considered very high

(Fish 96 h EC50=0.4 mg/L, fish 30-day LOEC=0.037 mg/L) and may cause long-term adverse

effects in the aquatic environment (U.S. EPA, 2014). The environmental persistence is

considered low, although there is a moderate potential for bioaccumulation (U.S. EPA, 2014).

1.4. Exposure sources and pathways

FRs are commonly used as additives in consumer products such as furniture, electronics and

textiles, i.e. they are not necessarily covalently bound in materials and tend to migrate into the

surrounding environment (Kemmlein et al., 2003). Human exposure to FRs as well as other

contaminants has been associated with inhalation and ingestion of contaminated indoor dust

(Covaci et al., 2012). High levels of contaminants in indoor dust is posing a risk to human

health, particularly vulnerable groups such as toddlers, which are especially prone to exposure

to contaminants associated with dust since they encounter it more when crawling and playing

on the floor as well as when they put items in their mouth (WHO, 2011). We spend most of

4

our time indoor in homes and offices and are continuously exposed to contaminants in indoor

dust. Indoor dust has been considered one of the most important pathways for exposure to FRs

(de Boer et al., 2016) and measuring levels of FRs in indoor dust therefore is considered a

suitable approach for monitoring chronic exposure.

1.5. Monitoring human exposure

Recent research has been focusing on characterizing human exposure to aryl-PFRs by

investigating correlations between aryl-PFRs in indoor dust and their metabolites in urine. For

this purpose, several metabolites may be of interest, including diphenyl phosphate (DPHP),

which is the hydrolyzed metabolite of TPHP (Figure 3) (U.S. EPA, 2014). DPHP has been used

as a biomarker for assessing exposure to TPHP in indoor dust and has been widely reported in

urine in the range <0.13-727 ng/mL (Cooper et al., 2011; Meeker et al., 2013; Van den Eede

et al., 2013b; Hoffman et al., 2014; Hoffman et al., 2015; Van den Eede et al., 2015; Kosarac

et al., 2016).

Figure 3. Hydrolysis of TPHP into DPHP.

Recent research, however, have found that the urinary levels of DPHP are strongly uncorrelated

to TPHP concentrations in indoor dust (rS=0.04, (Meeker et al., 2013); rS=0.15, (Hoffman et

al., 2015)), indicating that TPHP in dust is not the only source for human urinary levels of

DPHP. A possible additional source to explain the high urinary concentrations of DPHP could

be the direct exposure to DPHP itself as it is used in other applications (Makiguchi et al., 2011;

Zhao and Hadjichristidis, 2015) or direct exposure to DPHP via indoor dust ingestion as it may

be present as an impurity and/or as a degradation product as a result of spontaneous or

microbial hydrolysis of TPHP and/or other aryl-PFRs. DPHP has been reported in plastics from

electrical and electronic equipment that contain high levels of resorcinol bis(diphenyl

phosphate) (RDP) (Ballesteros-Gomez et al., 2016a; Ballesteros-Gomez et al., 2016b).

Moreover, DPHP has been demonstrated to be a metabolite to some other aryl-PFRs such as

2-Ethylhexyl diphenyl phosphate (EDP) (Nishimaki-Mogami et al., 1988; Ballesteros-Gomez

et al., 2015a), RDP (Ballesteros-Gomez et al., 2015b) and tert-Butylphenyl diphenyl phosphate

P

O

O

OO

TPHP DPHP

P

O

O O

OH

5

(BPDP) (Heitkamp et al., 1985). There is almost no data available about the presence of DPHP

in the indoor environment and determining levels of DPHP in indoor dust could play an

essential role in the understanding of the exposure sources and routes to TPHP and other aryl-

PFRs as well as DPHP itself. To the best of our knowledge only one study has reported levels

of DPHP (75-190 ng/g) in 4 dust samples collected in Australia (Van den Eede et al., 2015).

In this study, a method for the quantitation of TPHP and DPHP in indoor dust was developed

using liquid chromatography coupled with a triple-quadrupole mass spectrometer (LC-QQQ-

MS/MS). The developed method was applied on indoor dust samples collected from

households in the Netherlands in June 2016 (n=23) and in Spain in March and April 2017

(n=57) for the quantification of TPHP and DPHP. The levels of TPHP and DPHP were

compared between different microenvironments and countries and the correlation between

TPHP and DPHP concentrations were investigated. Human exposure to TPHP and DPHP via

indoor dust ingestion was estimated using different exposure scenarios. Furthermore, to gain

knowledge about the presence of other aryl-PFRs in indoor dust that could also contribute to

the formation of DPHP, an in-house developed database was used for target screening of other

emerging aryl-PFRs.

2. Experimental section

2.1. Chemicals and reagents

Acetonitrile and Methanol were acquired from VWR chemicals (Llinars del Vallès, Barcelona,

Spain). Ammonium acetate was from Sigma Aldrich (Zwijndrecht, the Netherlands). Ultra-

high-quality water was obtained from a Milli-Q water purification system (Millipore, Madrid,

Spain). Standard reference material (SRM) 2585 (organic contaminants in house dust) were

provided by the National Institute of Standards and Technology (NIST). TPHP, DPHP, TPHP-

d15 and DPHP-d10 were obtained from Sigma Aldrich Chemie B.V. (Zwijndrecht, the

Netherlands). Cresyl diphenyl phosphate (CDP), Isodecyl diphenyl phosphate (IDP), EDP,

RDP and Bisphenol A bis(diphenyl phosphate) (BADP) analytical standards were obtained

from AccuStandard (New Haven, CT).

2.2. Method optimization

The method optimization was done by using the indoor dust reference material SRM 2585 (50

mg). Since the material already contained DPHP and TPHP at relatively high concentrations,

the deuterated internal standards were employed for recovery optimization. The observed

6

average concentration of TPHP and DPHP in SRM 2585 (n=3) was 1075±151 ng/g and

4967±129 ng/g, respectively, which for TPHP is in accordance with previously reported

concentrations by other authors ranging 980±60 (Harrad et al., 2016) to 1110±48 (Brandsma

et al., 2013).

The reference material was spiked with 0.1 μg internal standard (IS) (TPHP-d15 and DPHP-

d10). The spiking was done before extraction, before clean-up or at the final reconstitution step

in order to assess the extraction efficiency, clean-up losses, matrix effects and total recoveries.

When spiked before extraction, the SRM was left stand for 2 h prior to extraction to allow the

solvent to evaporate in order to mimic as much as possible the interaction of the compound

with the dust matrix.

The extraction of DPHP and TPHP in indoor dust was performed by salting-out extraction with

acetonitrile and aqueous ammonium acetate (NH4Ac). A two-phase system was used to reduce

co-extraction of unwanted matrix components and thus achieve cleaner extracts. Due to the

high polarity of DPHP it is expected to partly remain in the aqueous phase. Therefore, different

salt concentrations were evaluated to increase the partition into the acetonitrile phase. After

extraction, due to the complexity of the dust matrix, a clean-up step with QuEChERS (75 mg

MgSO4, 25 mg PSA, 25 mg C18, 25 mg GCB) was assessed. Finally, two different LC columns

were evaluated to reduce interferences from co-eluting compounds, namely a Phenomenex

Luna® C18 column (2.0 mm i.d., 100 mm length, 3.0 m particle size) and a Phenomenex

Luna® phenyl-hexyl column (2.0 mm i.d., 100 mm length, 3.0 m particle size) equipped with

a Phenomenex SecurityGuardTM C18 guard column (2.0 mm i.d., 4.0 mm length). The LC

gradient and the MRM transitions were optimized.

An in-house database was developed for the targeted screening of CDP, IDP, EDP, RDP and

BADP in indoor dust. MS/MS transitions and parameters were optimized.

2.3. Apparatus and sample analysis

The LC system used was an Agilent Technologies 1200 LC. A Phenomenex Luna® C18 column

(2.0 mm i.d., 100 mm length, 3.0 m particle size) was used for separation. The mobile phase

consisted of 5 mM aqueous ammonium acetate (A) and methanol (B) at a flow rate of 0.25

mL/min. The gradient was as follows: initial 20% B, increased to 95% in 7.5 min and hold for

3 min and finally re-conditioning for 7 min. The MS/MS system was an Agilent Technologies

6420 Triple Quadrupole mass spectrometer equipped with LC-electrospray ionization (ESI)

source. The source parameters were set as following: Gas temperature, 320°C; gas flow, 12.0

7

L/min; nebulizer, 50 psi; capillary voltage, +/-4000 V; MS1 heater, 100°C; MS2 heater, 100°C.

The MRN transitions for target masses are given in Table 1. TPHP, BADP, RDP, IDP, EDP

and CDP were analyzed in positive ionization mode and DPHP was analyzed in both negative

and positive ionization mode.

Table 1. MRM transitions, dwell time, fragmentor voltage and collision energy. Quantifiers

for TPHP and DPHP are indicated in bold. Quantification of DPHP was performed with data

acquired using negative ionization.

Compound Precursor ion

(m/z)

Product ion

(m/z)

Dwell time (ms) Fragmentor (V) Collision energy

(eV)

Polarity

TPHP 327.1 77.1 150 150 40 Positive

TPHP 327.1 215.0 150 135 30 Positive

TPHP-d15 342.2 82.2 150 135 30 Positive

TPHP-d15 342.2 222.1 150 135 30 Positive

DPHP 251.0 77.1 150 135 30 Positive

DPHP 251.0 233.1 150 120 20 Positive

DPHP-d10 261.1 81.1 150 135 30 Positive

DPHP-d10 261.1 161.0 150 120 20 Positive

DPHP 249.0 93.0 150 135 30 Negative

DPHP 249.0 155.0 150 120 20 Negative

DPHP-d10 259.1 97.9 150 135 30 Negative

DPHP-d10 259.1 158.6 150 120 20 Negative

CDP 341.1 65.2 150 116 89 Positive

CDP 341.1 91.2 150 116 53 Positive

IDP 391.2 251.1 150 81 15 Positive

IDP 391.2 77.1 150 81 73 Positive

EDP 363.1 251.1 150 71 11 Positive

EDP 363.1 77.1 150 71 93 Positive

RDP 575.1 77.2 150 151 115 Positive

RDP 575.1 152.2 150 151 79 Positive

BADP 693.2 367.2 150 156 47 Positive

BADP 693.2 115.2 150 156 101 Positive

2.4. Sample collection and preparation

Sampling was performed using a filter (40 m) mounted in a nozzle adapted to a vacuum

cleaner and were not further sieved. Dust samples were collected from residences in the

Netherlands in June 2016 from floors (n=12) and from the surface of electrical equipment

(n=11) and in Spain in March and April 2017 from the floors of living rooms (n=9), bedrooms

(n=9) and offices (n=4), from the surface of electrical equipment (n=13), from cars (n=15) and

8

from public microenvironments (PMEs) (n=7) (two electronic shops, two clothing shops, one

sport clothing shop, one decoration shop and one cafeteria). Approximately 50 mg dust were

accurately weighed in 15 mL glass tubes and spiked with IS (TPHP-d15 and DPHP-d10, 0.1 g

each) prior to extraction. Due to the limitation of dust on top of electrical equipment, these

samples were in the size of approximately 10-30 mg.

Salting-out extraction with acetonitrile was performed with aqueous NH4Ac (3 M):acetonitrile

(1:1 v/v) by vortex for 2 min followed by centrifugation at 3000 rpm for 5 min. After phase-

separation, the acetonitrile layer was collected and transferred to a glass tube. The extraction

was repeated 2 times and the acetonitrile layers (~ 6 mL) were combined and evaporated to

approximately 1.5 mL (N2, 50°C). Sample clean-up was performed with QuEChERS (75 mg

MgSO4, 25 mg PSA, 25 mg C18, 25 mg GCB) by vortex for 2 min followed by

ultracentrifugation at 10 000 rpm for 5 min. The extract was then evaporated to near dryness

(N2, 50°C) and reconstituted in 200 L MilliQ:acetonitrile (1:1 v/v) by vortex for 30 s followed

by ultracentrifugation at 10 000 rpm for 5 min. Extracts were transferred to LC vials and

aliquots of 5 L were injected into the LC-MS/MS system.

2.5. Data processing

Quantification of TPHP and DPHP in indoor dust

Quantification of TPHP and DPHP in indoor dust was performed using the quantitative

analysis MassHunter workstation software from Agilent Technologies. Linear calibration (1/x

weighing, origin included) was employed. The method was evaluated based on extraction

efficiency, clean-up losses, matrix effects and reproducibility. Method limits of detection

(LOD) and quantification (LOQ) (ng/g) were estimated based on a signal-to-noise ratio of 3

and 10, respectively, taking into account the concentration factor of the method (sample size

of 50 mg and final extract volume of 200 μL) and the actual total recoveries.

Statistics

One-way ANOVA was employed to investigate if the TPHP and DPHP concentrations were

significantly different in dust collected in Spain and the Netherlands as well as in dust collected

from different microenvironments. Pearson correlation was performed in order to investigate

the correlation between DPHP and TPHP in indoor dust. For the statistical calculations, the

microenvironments were divided into four groups: floor dust (bedrooms, living rooms and

offices), dust collected from the surface of electronic equipment, car dust, and dust from the

floors of PMEs.

9

Screening of aryl-phosphate flame retardants

Targeted screening of aryl-PFRs was performed using the quantitative analysis MassHunter

workstation software from Agilent Technologies. An in-house database was built containing

the masses of the [M+H]+ ion as well as two abundant fragment ions for each target compound.

Criteria used for positives were: i) signal-to-noise ratio above 3, ii) qualifier ratio within 80-

120% range of the ratio observed from injected authentic standards and iii) an absolute peak

area larger than the area obtained from the lowest concentration authentic standard yielding a

defined peak.

3. Results and discussion

3.1. Method optimization

The method for quantification of TPHP and DPHP in indoor dust was evaluated based on

extraction recovery (%), clean-up recoveries (%), matrix effects (%), and reproducibility

(RSD%).

Salting-out phase separation

Three different concentrations of NH4Ac (2, 3 and 4 M) were evaluated in order to increase the

extraction efficiency of DPHP and TPHP. These initial experiments were carried out without

the presence of the dust matrix.

Extraction recovery (%) and related SD and RSD (%) for DPHP and TPHP are listed in Table

2 and illustrated in Figure 4. The extraction recoveries for DPHP were between 73% and 82%

with the highest recovery obtained using 3 M NH4Ac. For TPHP, extraction recoveries were

in the range 75% - 87% with the highest recovery obtained using 3 M NH4Ac. For both DPHP

and TPHP, standard deviations between replicates were higher at 2 M (5% and 19% RSD for

DPHP and TPHP, respectively) and 4 M (8% and 21% RSD for DPHP and TPHP respectively)

NH4Ac compared to 3 M (3% and 1% RSD for DPHP and TPHP, respectively).

As expected, the salting-out phase separation was more efficient at higher salt concentrations

resulting in extracts containing less water and so reducing the time required for sample

evaporation in the pre-concentration step. The optimum salt concentration was considered 3

M, at this concentration the amount of remaining water in the organic phase was minimal and

did not affect the evaporation step.

10

Table 2. Extraction recovery (%) and related SD and RSD (%) for DPHP and TPHP when

extracted with ACN:NH4Ac at different salt concentrations (2, 3 and 4 M).

NH4Ac concentration (M)

(ionization mode)

Extraction recovery (%) RSD (%)

DPHP

2 (neg) 76 ± 4 5

2 (pos) 73 ± 4 5

3 (neg) 82 ± 2 3

3 (pos) 80 ± 2 2

4 (neg) 78 ± 7 8

4 (pos) 78 ± 7 9

TPHP

2 (pos) 83 ± 16 19

3 (pos) 87 ± 1 1

4 (pos) 75 ± 16 21

Figure 4. Extraction recovery (%) of DPHP and TPHP when extracted with ACN:NH4Ac at 2, 3 and 4 M.

Sample preparation recoveries

Sample preparation consisted in salting-out extraction, clean-up with dispersive SPE and an

evaporation/reconstitution step. In order to assess losses of the target compounds during the

procedure, the recoveries were estimated at the different sample preparation steps, namely

extraction efficiency, clean-up recovery, matrix effects, and total recovery (extraction + clean-

up + matrix effects).

For the extraction efficiency, recoveries were calculated by spiking with IS before extraction

and comparing the signals with those obtained when spiking with IS after the extraction. For

the clean-up step, sample extracts were spiked before the clean-up and signals were compared

to those obtained when spiking with IS at the reconstitution step. For assessing the matrix

7682 78

7380

78

8387 75

0

10

20

30

40

50

60

70

80

90

100

NH4Ac 2M NH4Ac 3M NH4Ac 4M NH4Ac 2M NH4Ac 3M NH4Ac 4M NH4Ac 2M NH4Ac 3M NH4Ac 4M

Negative ionization Positive ionization Positive ionization

DPHP TPHP

Extraction recovery (%)

11

effects, samples were spiked at the reconstitution step and IS signals were compared with those

obtained from spiked injection solvent. Finally, for total recoveries, samples were spiked

before extraction and IS signals were compared with those obtained from spiked injection

solvent. Experiments were done in replicates. The results are listed in Table 3 and illustrated

in Figure 5.

Extraction recoveries for DPHP at the optimal salt concentration did not vary with the presence

of dust (82±2 without dust and 80±3 with dust). The extraction recovery for TPHP was

somewhat higher with the presence of dust (99±8) compared to without dust (87±1).

Regarding clean-up recoveries, values were 100±7% and 91±1% for DPHP and TPHP,

respectively, so that losses in this step were minimal. For DPHP, matrix effects were not

improved by the clean-up step. When operating in negative ionization mode, signal suppression

was even slightly higher with the clean-up (87±2 and 94±6 with and without clean-up,

respectively). However, when using positive ionization mode, a slightly larger suppression was

observed without the clean-up step (56±3% against 62±0%). In general, it can be concluded

that the clean-up step did not affect significantly the matrix effects for the analysis of DPHP.

However, the negative ionization mode was clearly more selective and the signal less affected

by matrix components than the positive ionization mode and was selected for the quantification

of DPHP, despite being less sensitive (around 6 times less sensitive than positive ionization

mode). For TPHP, the matrix effects were significant and improved when including clean-up

(29±5%) compared to when the clean-up step was not included (13±2%). Although the

improvement introduced by the clean-up step in terms of matrix effects was not optimal for the

target compounds, extracts were clearly cleaner, which is beneficial for a good performance of

the LC column and the MS source. The reproducibility in the TPHP analysis in terms of

standard deviation (SD) was also improved by the clean-up (deviations were higher ranging

18% to 71% when not using clean-up and 7% to 20% when including clean-up).

Total recoveries (extraction + clean-up + matrix effects) for DPHP were 69±2% (with clean-

up, negative mode). As explained before, losses were mainly due to matrix effects and to

extraction efficiency. Total recoveries for TPHP were 24±5% (with clean-up, positive mode),

due mainly to strong matrix effects.

Table 3. Extraction recovery (%), clean-up recovery (%), matrix effects (%) and total recovery

(%) and related SD and RSD (%) for DPHP and TPHP obtained with and without clean-up.

12

Extraction

recovery (%)

RSD

(%)

Clean-up

recovery (%)

RSD

(%)

Matrix

effects (%)

RSD

(%)

Total

recovery (%)

RSD

(%)

DPHP

Without clean-

up (neg)

80 ± 3 3 - - 94 ± 6 6 75 ± 2 3

Without clean-

up (pos) 77 ± 5 6 - - 56 ±3 5 43 ± 3 6

With clean-up

(neg)

79 ± 3 4 100 ± 7 7 87 ± 2 2 69 ± 2 4

With clean-up

(pos)

80 ± 2 2 94 ± 1 1 62 ± 0 0 47 ± 1 2

TPHP

Without clean-

up (pos) 99 ± 8 18 - - 13 ± 2 71 20 ± 4 18

With clean-up

(pos)

92 ± 19 20 91 ± 6 7 29 ± 5 16 24 ± 5 20

Figure 5. Extraction recovery (%), clean-up recovery (%), matrix effects (%) and total recovery (%) and related SD for DPHP and TPHP

obtained with and without clean-up.

Column and LC gradient

In order to reduce the signal suppression of TPHP, two different columns were evaluated (C18

and phenyl-hexyl) and the samples were analyzed using two different chromatographic

gradients (Table 4).

Table 4. LC gradients.

Short gradient Long gradient

Time (min) %B Time (min) %B

0 20 0 10

0.5 20 0.5 10

8 95 20 95

11 95 23 95

11.10 20 23.10 10

18.10 20 30.10 10

80 79 77 80

99 9210094 9194

87

5662

1329

7569

43 47

20 24

0

20

40

60

80

100

120

Without clean-up With clean-up Without clean-up With clean-up Without clean-up With clean-up

Negative ionization Positive ionization Positive ionization

DPHP TPHP

Extraction recovery (%) Clean-up recovery (%) Matrix effects ( %) Total recovery (%)

13

The matrix effects for DPHP were not improved when using a phenyl-hexyl column compared

to a C18 column. However, the matrix effects were improved when using a longer gradient

(101±1 with long gradient and 83±1 with short gradient). For TPHP, the matrix effects were

not improved when using a phenyl-hexyl column compared to a C18 column or when using a

longer gradient (Table 5, Figure 6). The C18 column and the short gradient were used for further

experiments in order to save time.

Table 5. Matrix effects (%), total recovery (%) and related SD and RSD (%) for DPHP and

TPHP obtained using two different columns and two different gradients.

Column Gradient (ionization

mode)

Matrix effects (%) RSD (%) Total recovery (%) RSD (%)

DPHP

C18 Short gradient (neg) 87 ± 2 2 69 ± 2 4

C18 Short gradient (pos) 62 ± 0 0 47 ± 1 2

Phenyl-hexyl Short gradient (neg) 83 ± 1 1 63 ± 0 0

Phenyl-hexyl Short gradient (pos) 55 ± 1 2 41 ± 0 1

Phenyl-hexyl Long gradient (neg) 101 ± 1 1 69 ± 1 1

Phenyl-hexyl Long gradient (pos) 74 ± 0 1 53 ± 1 1

TPHP

C18 Short gradient (pos) 29 ± 5 16 24 ± 5 20

Phenyl-hexyl Short gradient (pos) 29 ± 7 23 24 ± 8 31

Phenyl-hexyl Long gradient (pos) 28 ± 7 25 22 ± 5 22

Figure 6. Matrix effects (%) and total recovery (%) for DPHP and TPHP obtained using two different columns and with two different

gradients.

3.2. TPHP and DPHP concentrations in indoor dust

For TPHP and DPHP, the instrument linear range was 0.005-5 µg/mL and 0.005-10 µg/mL,

respectively. The instrument LOD and LOQ (TPHP and DPHP) were 0.0001 µg/mL and 0.005

µg/mL, respectively. Method LOD and LOQ were calculated based on signal-to-noise ratio of

3 and 10, respectively, considering sample amount, final extract volume, and total recovery.

87

62

83

55

101

74

29 29 28

69

47

63

41

69

53

24 2422

0

20

40

60

80

100

120

Short gradient(neg)

Short grdient(pos)

Short gradient(neg)

Short grdient(pos)

Long gradient(neg)

Long gradient(pos)

Short gradient(pos)

Short gradient(pos)

Long gradient(pos)

C18 Phenyl-hexyl C18 Phenyl-hexyl

DPHP TPHP

Matrix effects (%) Total recovery (%)

14

The method LOD and LOQ for TPHP were 1.54 ng/g and 73.96 ng/g, respectively. For DPHP,

the method LOD and LOQ were 0.38 ng/g and 19.23 ng/g, respectively. The total recovery in

the real dust samples were 26 ± 14 and 104 ± 28 for TPHP and DPHP, respectively (calculated

based on IS signal in samples (n=80) spiked prior to extraction and the average IS signal in

spiked blanks (n=3)).

TPHP and DPHP were detected at high concentrations in all samples analyzed from the

Netherlands and from Spain (Table 6, Figure 7). The highest concentrations of both TPHP and

DPHP were observed in dust samples collected from the seats and dashboards of cars (142459

ng/g and 79661 ng/g for TPHP and DPHP, respectively) followed by dust collected from on

top of electronic equipment (45330 ng/g and 21899 ng/g for TPHP and DPHP, respectively).

To the best of our knowledge, only one study has reported DPHP in indoor dust in the range

75-190 ng/g (Van den Eede et al., 2015). In general, the TPHP concentration was higher than

the DPHP concentration, commonly 2-3 times higher, in some cases up to 90 times higher.

However, in some samples the DPHP concentration were up to 2 times higher than the TPHP

concentration.

In dust collected from the Netherlands, TPHP concentration ranged 172-12853 ng/g and 285-

45330 ng/g in dust collected from the floors of homes and offices and from on top of electronic

equipment, respectively. The DPHP concentration ranged 151-4189 ng/g (homes and offices)

and 218-6588 ng/g (on top of electronics). In samples collected from Spain, the TPHP

concentrations ranged 265-18912 ng/g (living rooms), 211-1094 ng/g (bedrooms), 412-1353

ng/g (offices), 1270-26210 ng/g (on top of electronic equipment), 762-142459 ng/g (cars), and

169-1004 ng/g (PMEs). The DPHP concentrations ranged 111-461 ng/g (living rooms), 106-

1031 ng/g (bedrooms), 408-1251 ng/g (offices), 299-21899 ng/g (on top of electronic

equipment), 923-79661 ng/g (cars), and 263-556 ng/g (PMEs). The high concentrations of

TPHP and DPHP found on top of electronic equipment in comparison to concentrations

observed in dust collected from the floor in the same room (Figure 8) suggest that electronic

equipment is a relevant source of TPHP and DPHP in the indoor environment. However, no

correlation was observed between concentrations found in floor dust and in dust collected from

the surface of electronic equipment (TPHP, r=0.18; DPHP, r=0.04).

One-way ANOVA revealed that there was no statistically significant difference in TPHP and

DPHP levels in dust collected in Spain and in the Netherlands (TPHP, p=0.94, DPHP, p=0.62).

The microenvironments were divided into four groups: floor dust (bedrooms, living rooms and

15

offices), dust collected on top of electronic equipment, car dust and dust collected from the

floors of PMEs. Among these groups, no statistically significant difference in TPHP and DPHP

levels were revealed except between car dust and floor dust. The concentration TPHP and

DPHP in car dust where significantly higher than in floor dust (p<0.05), which could be

explained by a high use of flame retardants in the manufacturing of car seats and dashboards

and/or less frequently cleaning of cars in comparison to houses.

Table 6. TPHP and DPHP detection frequency (DF) and concentrations (ng/g) in indoor dust

from different microenvironments in Spain and the Netherlands.

DF (%) Mean ± SD Median Minimum Maximum

TPHP

(Spain)

Living rooms (n=9) 100 3161 ± 6051 944 265 18912

Bedrooms (n=9) 100 674 ± 297 734 211 1094

Offices (n=4) 100 760 ± 413 637 412 1353

On top of electronics (n=13) 100 5900 ± 7105 2416 1270 26210

Cars (n=15) 100 18305 ± 36362 4441 762 142459

PMEs (n=7) 100 665 ± 281 687 169 1004

DPHP

(Spain)

Living rooms (n=9) 100 241 ± 127 211 111 461

Bedrooms (n=9) 100 314 ± 284 197 106 1031

Offices (n=4) 100 771 ± 354 712 408 1251


Cars (n=15) 100 8294 ± 19897 2311 923 79661

PMEs (n=7) 100 371 ± 103 357 263 556

TPHP

(The Netherlands)

Homes and offices (n=12) 100 3073 ± 3789 1438 172 12853


DPHP

(The Netherlands)

Homes and offices (n=12) 100 1199 ± 1227 742 151 4189


16

Figure 7. Median concentration (ng/g) TPHP and DPHP in indoor dust from different microenvironments in Spain and in the

Netherlands.

Figure 8. TPHP (A) and DPHP (B) concentrations (ng/g) in dust collected from on top of electronic equipment and from the floor in the

same room

The TPHP concentrations in indoor dust from homes in Spain and in the Netherlands are in

line with those reported elsewhere (Table 7, Figure 9) (Garcia et al., 2007; Stapleton et al.,

2009; Kanazawa et al., 2010; Van den Eede et al., 2011; Ali et al., 2012a; Ali et al., 2012b;

Dirtu et al., 2012; Dodson et al., 2012; Ali et al., 2013; Kim et al., 2013; Abdallah and Covaci,

2014; Araki et al., 2014; Cequier et al., 2014; Fan et al., 2014; Tajima et al., 2014; Brommer

and Harrad, 2015; Hoffman et al., 2015; Mizouchi et al., 2015; Zheng et al., 2015; Ali et al.,

2016; Cristale et al., 2016; Harrad et al., 2016; He et al., 2016; Wu et al., 2016; Kademoglou

et al., 2017). Same accounts for TPHP concentrations in dust collected from on top of electronic

equipment as well as from floors of offices and PMEs (Figure 1, Table S-2) (Kanazawa et al.,

2010; Bergh et al., 2011; Van den Eede et al., 2011; Ali et al., 2012b; Ali et al., 2013; Abdallah

and Covaci, 2014; Araki et al., 2014; Brandsma et al., 2014; Tajima et al., 2014; Brommer and

100

1000

10000

Living rooms(n=9)

Bedrooms(n=9)

Offices (n=4) On top ofelectronics

(n=13)

Cars (n=15) PMEs (n=7) Homes andoffices (n=12)

On top ofelectronics

(n=11)

Spain The Netherlands

log

(ng/

g)

TPHP DPHP

100

1000

10000

100000

1 2 3 4 5 6 7 8 9 10 11 12 13

log

TPH

P (n

g/g)

Room

On top of electronic equipment Floor

100

1000

10000

100000

1 2 3 4 5 6 7 8 9 10 11 12 13

log

DPH

P (n

g/g)

Room

On top of electronic equipment Floor

(A) (B)

17

Harrad, 2015; Ali et al., 2016; Ballesteros-Gomez et al., 2016a; Cristale et al., 2016; Harrad et

al., 2016; He et al., 2016; Wu et al., 2016; Kademoglou et al., 2017). The median TPHP

concentration observed in car dust (4441 ng/g) was however somewhat higher than reported

before (135-3700 ng/g). Reported TPHP concentrations in house dust as well as in dust from

other microenvironments span over a wide concentration range (<2-1798000 ng/g) with the

highest concentration reported being observed in house dust from the U.S. (Stapleton et al.,

2009). The lowest concentration was observed in house dust from Pakistan (Ali et al., 2012b;

Ali et al., 2013) and in car dust from Kuwait (Ali et al., 2013). This high variation in TPHP

concentrations, spanning several orders of magnitude, may be explained by different fire-safety

regulations in different countries as well as different regulations regarding the production and

use of PBDEs. Abdallah and Covaci (2014) reported levels of PFRs in house dust from Egypt

which are among the lowest reported PFR levels worldwide (maximum concentration TPHP

was 289 ng/g), which may be explained by a higher use of PBDEs and/or less strict fire-safety

regulations in Egypt.

Table 7. Summary of TPHP concentration (ng/g) in dust from different microenvironments in

Spain compared to concentrations reported elsewhere.

Microenvironment n DF (%) Mean Median Minimum Maximum Country Reference

Houses (floors)

18 100 1918 782 211 18912 Spain Present study

12 100 3073 1438 172 12853 The Netherlands Present studya

8 100 2600 1850 290 9500 Spain García et al., 2007

5 100 1300 1102 580 2633 Spain Cristale et al., 2016

22 100 1171 230 70 18000 Germany Harrad et al., 2016

33 100 2020 500 40 29800 Belgium Van den Eede et al., 2011

32 - 10000 3300 490 110000 UK Brommer & Harrad, 2015

10 100 2737 1509 190 9549 UK Kademoglou et al., 2017

10 100 931 830 202 2922 Norway Kademoglou et al., 2017

48 100 1240 981 - 4850 Norway Cequier et al., 2014

47 96 1600 500 <20 22600 Eastern Romania Dirtu et al., 2012

20 100 101 67 8 289 Egypt Abdallah & Covaci 2014

15 100 1080 430 44 6890 Kuwait Ali et al., 2013

9 100 4511 3800 1200 9200 Kazakhstan Harrad et al., 2016

15 100 310 230 65 1200 Saudi Arabia Ali et al., 2016


31 100 107 94 <2 630 Pakistan Ali et al., 2012b

15 87 155 175 <2 330 Pakistan Ali et al., 2013

17 100 110 89 8.5 2500 Philipines Kim et al., 2013

20 100 73 71 13 440 Philipines Kim et al., 2013

9 - - 1110 86 15800 China Zheng et al., 2015

18

7 - - 2500 122 16500 China Zheng et al., 2015

13 - - 3320 119 6030 China Zheng et al., 2015

13 - - 1740 31 6660 China Zheng et al., 2015

14 - - 9810 371 332000 China Zheng et al., 2015

6 100 560 600 150 1030 China He et al., 2016

21 100 540 376 122 1829 China Wu et al., 2016

8 63 160 150 ND 390 China He et al., 2016

41 76 - 5400 <1600 78400 Japan Kanazawa et al., 2010

48 60 - 870 - 2335 Japan Tajima et al., 2014

10 100 1400 820 230 6700 Japan Mizouchi et al., 2015

148 89 - 4510 <1600 245080 Japan Araki et al., 2014

32 100 10145 1200 490 110000 Australia Harrad et al., 2016

34 100 590 565 20 7510 New Zealand Ali et al., 2012a

14 100 8704 1600 20 37000 Canada Harrad et al., 2016

134 100 - 1700 260 63000 Canada Fan et al., 2014

50 98 7360 5470 <150 1798000 US Stapleton et al., 2009

53 100 1020 - 100 40350 US Hoffman et al., 2015

16 100 7999 2797 786 36463 US Dodson et al., 2012

Houses (on top of

and around

electronic

equipment as well as

on surfaces e.g.

tables, door frames

etc.)


11 100 10353 9786 285 45330 The Netherlands Present study

8 - - 6500 1600 21000 The Netherlands Brandsma et al., 2014

30 100 7962 3721 222 50728 The Netherlands Ballesteros-Gómez et al.,

2016a


10 100 1600 1200 100 4200 Sweden Bergh et al., 2011

128 95 - 3130 - 27470 Japan Tajima et al., 2014

41 98 - 14300 <1600 175000 Japan Kanazawa et al., 2010

120 94 - 11540 <1600 889180 Japan Araki et al., 2014

Houses (coaches and

mattresses)

10 100 2985 2350 180 8400 UK Harrad et al., 2016

220 >81 - 419 96 >95000 Sweden Canbaz et al., 2016


16 100 465 240 20 35190 New Zealand Ali et al., 2012a

Offices





12 100 8834 5752 1331 38094 UK Kademoglou et al., 2017



9 100 15708 5300 390 48000 Kazakhstan Harrad et al., 2016

23 100 4136 1928 31 38646 China Wu et al., 2016

12 100 1050 900 330 2380 China He et al., 2016

19


63 76 2560 500 <300 64500 Germany Fromme et al., 2014

Daycare centers and

schools


6 100 2400 1540 - 6150 Norway Cequier et al., 2014


16 100 868 531 41 3514 China Wu et al., 2016

9 100 269 140 10 1023 China Wu et al., 2016

18 100 6200 2200 350 62000 Japan Mizouchi et al., 2015

Cars







15 87 2165 1760 <2 7415 Kuwait Ali et al., 2013


15 100 665 245 2 4800 Pakistan Ali et al., 2013


Public

microenvironments

(PMEs)




15 100 4700 1970 150 34200 Belgium Van den Eede et al., 2011


12 100 101 109 13.5 185 Pakistan Ali et al., 2012b

7 100 520 220 70 1840 China He et al., 2016

a pooled floor dust from homes and offices.

20

Figure 9. Reported median concentration TPHP (ng/g) in indoor dust from houses in different countries.

Figure 10. Reported median concentration TPHP (ng/g) in indoor dust from different microenvironments in different countries

3.3. Correlation between TPHP and DPHP concentrations in indoor dust

Pearson correlation was performed to investigate the correlation between TPHP and DPHP

concentrations in indoor dust. Pearson correlation was calculated using all samples collected

from the Netherlands and from Spain (n=80). The result indicates a strong and statistically

significant positive correlation between the concentration of TPHP and DPHP in indoor dust

(r=0.90, p<0.01) (Figure 11). Pearson correlation was also performed for individual

microenvironments (Figure 12). Observations with standardized residuals larger than 2

(absolute value) were considered outliers and were excluded prior to regression (indicated in

orange). Statistically significant positive correlations were observed in dust collected from

floors of houses and offices (r=0.46, p<0.05) (Figure 12A), on top of electronic equipment

(r=0.60, p<0.01) (Figure 12B) and cars (r=0.99, p<0.01) (Figure 12C). A strong correlation

were also observed in dust collected from the floors of PMEs (r=0.72) (Figure 12D), however,

not statistically significant (p=0.07). These findings indicate that the presence of DPHP in

indoor dust is to a considerable extent related to the presence of TPHP suggesting that DPHP

in indoor dust is mainly present as an impurity and/or a degradation product of TPHP.

7821438

18501102

230 500

3300

1509830 981

50067

430

3800

230600

94 175 89 71

1110

2500

3320

1740

9810

600 376 150

5400

870 820

4510

1200565

1600 1700

5470

2797

0

2000

4000

6000

8000

10000

12000

2416

9786

6500

3721

820 1200

3130

14300

11540

637 7401500

4300

5752 5300

73

5300

1928900

4441

18002400

1700

3300

135

1760

470 245

3700

687

4010

179

1970

629109 220

0

2000

4000

6000

8000

10000

12000

14000

16000

On top of electronic equipment Offices Cars PMEs

21

However, it cannot be ruled out that the presence of DPHP in indoor dust might also be a result

of degradation of other aryl-PFRs.

Figure 11. Correlation between TPHP and DPHP concentration in indoor dust from Spain and the Netherlands.

1.00

1.50

2.00

2.50

3.00

3.50

4.00

4.50

5.00

5.50

1.00 1.50 2.00 2.50 3.00 3.50 4.00 4.50 5.00 5.50

log

DPH

P (n

g/g)

log TPHP (ng/g)

22

Figure 12. Correlation between TPHP and DPHP concentration in indoor dust in different microenvironments (A) floor dust (houses and

offices), (B) on top of electronic equipment, (C) cars, and (D) PMEs.

3.4. Estimated exposure to TPHP and DPHP in indoor dust

Human exposure scenarios to TPHP and DPHP via dust ingestion in Spain were estimated

using a method based on that described by Abdallah and Covaci (2014). Briefly, average and

high dust ingestion rates (95th percentile) for adults (2.6 mg/day and 8.6 mg/day, respectively)

and toddlers (41 mg/day and 140 mg/day, respectively) (Wilson et al., 2013) were used to

calculate an average and a worst-case scenario exposure to TPHP and DPHP via indoor dust.

Estimated exposure scenarios were calculated based on median and maximum concentrations

in indoor dust in homes (bedrooms and living rooms), offices, cars and PMEs (different stores

and one cafeteria) in Spain, taking into account the time spent in each environment according

to the typical human activity patterns described by Abdallah and Covaci (2014) (i.e. 63.8%

home, 22.3% office, 5.1% PMEs, 4.1% car and 4.7% outdoors, for adults, and 86.1% home,

5.1% PMEs, 4.1% car and 4.7% outdoors, for toddlers). Occupational exposure of drivers (e.g.

taxi drivers and truck drivers) were estimated by using the concentrations in cars as

representative concentrations for the working environment (i.e. time fraction spent in car was

4.1% + 22.3%). Following equation was used to calculate estimated exposure scenarios:

1.00

1.50

2.00

2.50

3.00

3.50

4.00

1.00 1.50 2.00 2.50 3.00 3.50 4.00 4.50

log

DPH

P (n

g/g)

log TPHP (ng/g)

1.00

1.50

2.00

2.50

3.00

3.50

4.00

4.50

5.00

5.50

1.00 2.00 3.00 4.00 5.00 6.00

log

DPH

P (n

g/g)

log TPHP (ng/g)

1.00

1.50

2.00

2.50

3.00

3.50

4.00

4.50

5.00

1.00 2.00 3.00 4.00 5.00

log

DPH

P (n

g/g)

log TPHP (ng/g)

1.00

1.20

1.40

1.60

1.80

2.00

2.20

2.40

2.60

2.80

3.00

1.00 1.50 2.00 2.50 3.00 3.50

log

DPH

P (n

g/g)

log TPHP (ng/g)

(D)(C)

(B)(A)

23

𝐸𝑠𝑡𝑖𝑚𝑎𝑡𝑒𝑑 𝑑𝑎𝑖𝑙𝑦 𝑒𝑥𝑝𝑜𝑠𝑢𝑟𝑒 (𝑛𝑔/𝑑𝑎𝑦) = 𝐼𝑅× ∑ 𝐶𝑖×𝐹𝑖

where IR is the dust ingestion rate (g/day), Ci the concentration TPHP or DPHP in dust in

microenvironment i (ng/g) and Fi is the time fraction spent in microenvironment i.

The estimated exposure to TPHP and DPHP for different exposure scenarios including workers

(offices), drivers, non-workers and stay-home toddlers are shown in Table 8. The estimated

daily exposure to TPHP via indoor dust ingestion in Spain (based on average dust ingestion

rates and median concentrations) were 2.2 ng/day, 4.4 ng/day, 2.3 ng/day and 36.5 ng/day for

adult workers, drivers, non-workers and stay-home toddlers, respectively. These exposure

scenarios are in line with those reported elsewhere (based on average dust ingestion rate and

median concentration) which are in the range 0.9-58.5 ng/day, 7.0-30.2 ng/day and 3-75.4

ng/day for adult workers, non-workers and stay-home toddlers, respectively (Table 9) (Van

den Eede et al., 2011; Ali et al., 2012a; Dirtu et al., 2012; Kim et al., 2013; Abdallah and

Covaci, 2014; He et al., 2016; Wu et al., 2016; Kademoglou et al., 2017).

Worst-case scenario estimated daily exposure to TPHP via indoor dust ingestion (based on

high dust ingestion rates and maximum concentrations) were 157.0 ng/day, 427.6 ng/day, 190.7

ng/day and 3104.5 ng/day for adult workers, drivers, non-workers and stay-home toddlers,

respectively (Table 8). For adults, the calculated exposure estimates are in line with those

reported elsewhere (based on high dust ingestion rate and 95th percentile or maximum

concentration) which are in the range 13.0-953.2 ng/day and 70.0-506.1 ng/day for workers

and non-workers, respectively. However, the worst-case scenario estimated daily exposure to

TPHP via dust ingestion for stay-home toddlers were 3104.5 ng/day, higher than reported in

previous studies (Table 9) (Van den Eede et al., 2011; Ali et al., 2012a; Dirtu et al., 2012; Kim

et al., 2013; Abdallah and Covaci, 2014; He et al., 2016; Wu et al., 2016; Kademoglou et al.,

2017). Despite the high estimated daily exposure for toddlers, all calculated exposure estimates

for different scenarios are far below the reference dose of 164500 ng/day (adults) and 28905

ng/day (toddlers) calculated from the lowest reported chronic NOAEL, 23.5 mg/kg/day (U.S.

EPA, 2015) divided by a safety factor of 10000 assuming body weights of 70 kg and 12.3 kg

for adults and toddlers, respectively (U.S. EPA, 2008).

The estimated daily exposure to DPHP via indoor dust ingestion in Spain (based on average

dust ingestion rates and median concentrations) were 1.0 ng/day, 2.0 ng/day, 0.8 ng/day and

11.8 ng/day for adult workers, drivers, non-workers and stay-home toddlers, respectively

(Table 8). Worst-case scenario estimated daily exposure to DPHP via indoor dust ingestion

24

(based on high dust ingestion rates and maximum concentrations) were 36.4 ng/day, 186.8

ng/day, 36.0 ng/day and 585.5 ng/day for adult workers, drivers, non-workers and stay-home

toddlers, respectively. To the best of our knowledge, this is the first study to report estimated

daily exposure scenarios to DPHP via indoor dust ingestion.

Table 8. Estimated daily exposure (ng/day) for different exposure scenarios in Spain.

Ingestion

rate

Workers Drivers Non-workers Stay-home toddlers

Median Maximum Median Maximum Median Maximum Median Maximum

TPHP Average 2.2 47.5 4.4 129.3 2.3 57.7 36.5 909.2

High 7.4 157.0 14.7 427.6 7.7 190.7 124.6 3104.5

DPHP Average 1.0 11.0 2.0 56.5 0.8 10.9 11.8 171.5

High 3.5 36.4 6.5 186.8 2.5 36.0 40.4 585.5

Table 9. Estimated daily exposure to TPHP (ng/day) reported elsewhere.

Dust

ingestion

rate

Adults (working) Adults (non-working) Toddlers

Median Maximum Median Maximum Median Maximum Country Reference

Average 1.9 - - - 4.8b -

Egypt Abdallah & Covaci

2014 High 4.8 - - - 19.3b -

Average 1.1 18.0a - - 3.8 59.0a Philipines

(residental area) Kim et al. 2013

High 2.8 44.0a - - 15.0 240.0a

Average 0.9 5.3a - - 3.0 18.0a Philipines

(municipal

dumping area) Kim et al. 2013

High 2.3 13.0a - - 12.0 71.0a

Average 58.5 381.3 30.2 191.0 75.4 477.4 UK

Kademoglou et al.

2017 High 146.2 953.2 75.5 477.5 301.8 1909.8

Average - - 16.6 58.5 41.5 146.1 Norway

Kademoglou et al.

2017 High - - 41.5 146.1 166.0 584.4

Average - - 7.0 28.0a 18.0 69.8a New Zealand Ali et al. 2012a

High - - 18.2 70.0a 71.9 279.6a

Average 7.0 - 14.0 - 24.6 - Belgium

Van den Eede et al.

2011 High 28.0 357.0a 28.0 147.0a 100.9 500.6a

Average 21.0 98.0a - - 21.6 86.4a China Wu et al. 2016

High 49.0 245.0a - - 86.4 344.4a

Average - - - - - - China He et al. 2016

High - 141.5 - - - 195.0a

Average - - 10.5 203.0a 26.3 504.0a Eastern

Romania Dirtu et al. 2012

High - - 26.3 506.1a 105.0 2016.0a

a 95th percentile

b Toddlers going to daycare centers

25

3.5. Estimated urinary levels of DPHP

Estimated urinary levels of DPHP as a result of exposure to DPHP and TPHP via indoor

dust were calculated based on the median and maximum levels of DPHP and TPHP in indoor

dust in Spain as well as the time fractions spent in each microenvironment according to typical

human activity patterns described in previous section. A method based on that described by

Van den Eede et al. (2015) was employed. Briefly, an average and a high dust ingestion rate

(95th percentile) for adults (2.6 mg/day and 8.6 mg/day) and toddlers (41 mg/day and 140

mg/day) (Wilson et al., 2013) was assumed. Other assumptions were the complete absorption

of DPHP and TPHP after dust ingestion as well as the complete excretion of DPHP in urine

and that DPHP is absorbed and excreted unchanged (Sudakin and Stone, 2011). The

assumption that TPHP is metabolized into DPHP by liver enzymes at a rate of 20% was also

included (Van den Eede et al., 2013a). Based on these assumptions and assuming a mean

urinary output of 800 mL/day for adults and 600 mL/day for children, the estimated DPHP

excretion rate (ng/day) and urinary levels (ng/mL) were calculated using following equation:

𝑈𝑟𝑖𝑛𝑎𝑟𝑦 𝑐𝑜𝑛𝑐𝑒𝑛𝑡𝑟𝑎𝑡𝑖𝑜𝑛 (𝑛𝑔

𝑚𝐿) =

𝐼𝑅× ∑ 𝐶𝑖(𝐷𝑃𝐻𝑃)𝐹𝑖 + 0.2×𝐶𝑖(𝑇𝑃𝐻𝑃)𝐹𝑖

𝑈𝑟𝑖𝑛𝑎𝑟𝑦 𝑜𝑢𝑡𝑝𝑢𝑡 (𝑚𝐿𝑑𝑎𝑦

)

Where IR is the dust ingestion rate (g/day), Ci the concentration DPHP and TPHP in dust in

microenvironment i, and Fi is the time fraction spent in microenvironment i.

The estimated urinary DPHP levels for different exposure scenarios in Spain including workers

(offices), drivers, non-workers and stay-home toddlers are shown in Table 10. The estimated

urinary DPHP levels as a result of exposure to TPHP and DPHP via indoor dust ingestion

(based on average dust ingestion rates and median concentrations) were 0.002 ng/mL, 0.004

ng/mL, 0.002 ng/mL and 0.032 ng/mL for adult workers, drivers, non-workers, and stay-home

toddlers, respectively. These estimated urinary DPHP levels as a result of exposure to TPHP

and DPHP via indoor dust ingestion are not high enough to significantly contribute to the high

DPHP urinary levels reported in the literature ranging <0.13-727 ng/mL (Cooper et al., 2011;

Meeker et al., 2013; Van den Eede et al., 2013b; Hoffman et al., 2014; Hoffman et al., 2015;

Van den Eede et al., 2015; Kosarac et al., 2016).

Worst-case scenario estimated urinary DPHP levels for the different exposure scenarios (based

on high dust ingestion rate and maximum concentration in dust) were 0.085 ng/mL, 0.34

ng/mL, 0.094 ng/mL, and 2.011 ng/mL for workers (offices), drivers, non-workers and stay-

26

home toddlers, respectively. The estimated urinary DPHP level in toddlers is 40 times higher

than the worst-case scenario reported previously (0.05 ng/mL) (Van den Eede et al., 2015).

Furthermore, the estimated worst-case scenario urinary DPHP levels are in the same range as

the lower urinary DPHP concentrations reported previously (<0.13 ng/mL) (Kosarac et al.,

2016).

Van den Eede et al. (2016) showed that serum enzymes are involved in the transformation of

TPHP into DPHP and that the amount TPHP that reaches the liver after intake may be strongly

reduced. Therefore, the metabolic transformation rate of TPHP into DPHP (by serum and liver

enzymes) could be higher than 20% resulting in an underestimation of urinary DPHP levels.

Same study also investigated the hydrolysis products of EDP by serum enzymes and results

suggest an additional production of DPHP from EDP, however, at a much lower rate than for

TPHP.

Additional sources of TPHP as well as direct exposure to DPHP from other sources in addition

to exposure to other aryl-PFRs being metabolised into DPHP, may play an essential role in the

high urinary DPHP levels.

It should be noted that the TPHP and DPHP concentrations in indoor dust varies over several

orders of magnitude between different studied environments and between different homes and

that the estimated urinary DPHP urinary levels in the present study therefore not can be

compared directly to reported urinary levels elsewhere without a large degree of uncertainty.

Table 10. Estimated urinary DPHP concentration (ng/mL) for different exposure scenarios in

Spain.

Average ingestion rate High ingestion rate

Median Maximum Median Maximum

Workers 0.002 0.026 0.006 0.085

Drivers 0.004 0.103 0.012 0.340

Non-workers 0.002 0.028 0.005 0.093

Stay-home toddlers 0.032 0.589 0.109 2.011

3.6. Screening of aryl-phosphate flame retardants

TPHP and DPHP were detected in all samples analyzed from Spain (n=57) and the Netherlands

(n=23). The other aryl-PFRs, namely CDP, IDP, EDP, RDP and BADP, were less frequently

detected (Table 11). EDP was the most frequently detected aryl-PFR after TPHP and DPHP

with a detection frequency of 64.9% and 65.2% in Spain and the Netherlands, respectively,

27

followed by IDP (50.9% and 43.5%), BADP (33.3% and 34.8%), CDP (3.5% and 8.7%) and

RDP (0% and 4.3%). Detection frequencies of all aryl-PFRs included in the present study were

similar in samples collected from Spain and the Netherlands. Furthermore, there were no

observed differences in the presence of aryl-PFRs in different microenvironments (Table 12).

However, due to the limited number collected from each microenvironment, these results are

not conclusive.

Table 11. Compound name, CAS, molecular structure, chemical formula, monoisotopic mass

and detection frequency (%) of TPHP, DPHP, CDP, IDP, EDP, RDP, and BADP in indoor dust

from Spain and the Netherlands.

Compound

CAS

Molecular structure Chemical

formula

Monoisotopic

mass (g/mol)

Detection frequency (%)

Spain (n=57) The Netherlands (n=23)

Triphenyl phosphate

(TPHP)

115-86-6

C18H15O4P 326.070801 100 100

Diphenyl phosphate (DPHP)

838-85-7

C12H10O4P 250.039490 100 100

Cresyl diphenyl

phosphate (CDP)

26444-49-5

C19H17O4P 340.086456 3.5 8.7

Isodecyl diphenyl

phosphate (IDP)

29761-21-5

C22H31O4P 390.195984 50.9 43.5

2-Ethylhexyl diphenyl

phosphate (EDP)

1241-94-7

C20H27O4P 362.164703 64.9 65.2

Resorcinol

bis(diphenyl phosphate) (RDP)

57583-54-7

C30H24O8P2 574.094666 0 4.3

O

O

O

O

P

OH

OO

O

P

CH3

O

O

O

OP

CH3

CH3

O

O

O

OP

CH3CH3

O

O

O

OP

O O

O

O

O

O

O OP P

https://pubchem.ncbi.nlm.nih.gov/search/#collection=compounds&query_type=mf&query=C30H24O8P2&sort=mw&sort_dir=asc

28

Bisphenol A bis(diphenyl

phosphate) (BADP)

5945-33-5

C39H34O8P2 692.172913 33.3 34.8

Table 12. Detection frequency (%) of aryl-PFRs in indoor dust from different

microenvironments in Spain and the Netherlands.

Microenvironment TPHP DPHP CDP IDP EDP RDP BADP

Spain

Living rooms (n=9) 100 100 0.0 44.4 44.4 0.0 11.1

Bedrooms (n=9) 100 100 0.0 100 55.6 0.0 33.3

Offices (n=4) 100 100 0.0 50.0 75.0 0.0 0.0

Floor dust (bedroom +

living room + office) (n=22)

100 100 0.0 68.2 54.5 0.0 18.2

On top of electronics (n=13) 100 100 7.7 46.2 92.3 0.0 61.5

Cars (n=15) 100 100 6.7 26.7 46.7 0.0 33.3

PMEs (n=7) 100 100 0.0 57.1 85.7 0.0 28.6

The

Netherlands

Homes and offices (n=12) 100 100 8.3 33.3 66.7 0.0 0.0

On top of electronics (n=11) 100 100 9.1 54.5 63.6 9.1 72.7

4. Conclusions

Salting-out extraction with acetonitrile and 3 M ammonium acetate provided high

extraction efficiencies for TPHP and DPHP in indoor dust. However, TPHP suffered from

severe signal suppression that were somewhat improved when clean-up with QuEChERS was

employed.

TPHP and DPHP were present at high concentrations with 100% detection frequency in all

samples analyzed from Spain and the Netherlands. The highest maximum concentrations of

TPHP and DPHP were observed in dust collected from the seats and dashboards of cars

(142459 ng/g and 79661 ng/g for TPHP and DPHP, respectively), followed by dust collected

from the surface of electronic equipment (45330 ng/g and 21899 ng/g for TPHP and DPHP,

respectively). This suggest a high use of TPHP in the manufacturing of car interiors and

electronic equipment. The lowest concentrations of TPHP (169 ng/g) and DPHP (106 ng/g)

were observed in floor dust collected from PMEs and bedrooms, respectively.

OPO

O

O

O

PO O

O

CH3

CH3

29

TPHP concentrations in house dust in Spain and the Netherlands are in line with those reported

elsewhere. However, the reported concentrations span over several orders of magnitudes with

the lowest concentration being reported in house dust from Pakistan (<2 ng/g) (Ali et al.,

2012b; Ali et al., 2013) and the highest concentration being reported in house dust from the

U.S. (1798000 ng/g) (Stapleton et al., 2009). This wide range of TPHP concentrations being

reported in house dust may be explained by different fire-safety regulations in different

countries and/or regulations regarding the use of PBDEs.

DPHP concentrations were strongly and statistically significantly correlated to TPHP

concentrations in indoor dust from Spain and the Netherlands (r=0.90, p<0.01). The strongest

correlation was observed in dust collected from cars (r=0.99, p<0.01). These findings suggest

that TPHP is a major source for DPHP in indoor dust. However, other possible sources for

DPHP in indoor dust cannot be ruled out since DPHP has been suggested to be an impurity to

and/or a degradation product of RDP (Ballesteros-Gomez et al., 2016a; Ballesteros-Gomez et

al., 2016b) as well as a metabolite of EDP (Nishimaki-Mogami et al., 1988; Ballesteros-Gomez

et al., 2015a), RDP (Ballesteros-Gomez et al., 2015b) and tert-Butylphenyl diphenyl phosphate

(BPDP) (Heitkamp et al., 1985). Furthermore, DPHP is also used as an organocatalyst in

polymerization processes (Makiguchi et al., 2011; Zhao and Hadjichristidis, 2015). More

research would be desirable in order to investigate whether or not other sources of DPHP is

also relevant.

The estimated average daily exposure to TPHP and DPHP in Spain is highest for toddlers (36.5

ng/g and 11.8 for TPHP and DPHP, respectively) followed by drivers (4.4 ng/g and 2.0 for

TPHP and DPHP, respectively), which is far below the reference dose for TPHP of 164500

ng/day (adults) and 28905 ng/day (toddlers). However, the estimated worst-case scenario daily

exposure to TPHP for toddlers were 3104.5 ng/g, which is less than 10 times below the

reference dose for toddlers.

The estimated average urinary DPHP concentrations as a result of exposure to TPHP and DPHP

via indoor dust ingestion is far below and insufficient to explain the DPHP concentrations

reported in urine. Only the estimated worst-case scenario urinary DPHP concentrations are in

the same range as the lower DPHP concentrations reported in urine, but still insufficient to

explain the higher concentrations reported in urine. Other sources of TPHP exposure and/or the

presence of other aryl-PFRs that are degraded and/or metabolised into DPHP may be a relevant

30

source for the high concentrations of DPHP reported in urine and further research is necessary

in order to understand the high concentrations of DPHP in urine.

TPHP and DPHP were detected in all samples analyzed from Spain (n=57) and the Netherlands

(n=23). CDP, IDP, EDP, RDP and BADP were less frequently detected. EDP was most

frequently detected after TPHP and DPHP. In samples collected from Spain, EDP, IDP, BADP,

CDP and RDP were detected in 64.5%, 50.9%, 33.3%, 3.5%, and 0.0% of the samples,

respectively. In samples collected from the Netherlands, EDP, IDP, BADP, CDP and RDP

were detected in 65.2%, 43.5%, 34.8%, 8.7%, and 4.3% of the samples, respectively. The

presence of all aryl-PFRs included in the present study was similar in samples collected from

Spain and the Netherlands and no differences could be observed between different

microenvironments. However, due to the limited number collected from each

microenvironment, these results are not conclusive.

Acknowledgments

A big thank you to everyone in the Supramolecular research group at the Department of

Analytical Chemistry, University of Córdoba, for welcoming me with open arms into their

group and for being patient and answering my many questions.

And many big thanks to my supervisor Ana Ballesteros-Gómez for giving me the opportunity

to perform my research project abroad, for answering my questions and for taking the time of

teaching me but also for giving me the chance of being independent and taking my own

initiatives.

31

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