MSc Analytical Sciences - UvA · male rats administered TPHP in diet during four weeks (U.S. EPA,...
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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
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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.
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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
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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
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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.
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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.
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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
On top of electronics (n=13) 100 3211 ± 5780 1753 299 21899
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
On top of electronics (n=11) 100 10353 ± 12688 9786 285 45330
DPHP
(The Netherlands)
Homes and offices (n=12) 100 1199 ± 1227 742 151 4189
On top of electronics (n=11) 100 2781 ± 2102 2581 218 6588
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
15 100 880 600 120 2500 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.)
13 100 5900 2416 1270 26210 Spain Present study
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
8 - - 820 680 11000 The Netherlands Brandsma et al., 2014
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
41 100 4951 1800 370 29000 Australia Harrad et al., 2016
16 100 465 240 20 35190 New Zealand Ali et al., 2012a
Offices
4 100 760 637 412 1353 Spain Present study
1 100 740 740 740 740 Spain Cristale et al., 2016
25 100 2419 1500 200 8800 Germany Harrad et al., 2016
61 - 8200 4300 560 50000 UK Brommer & Harrad, 2015
12 100 8834 5752 1331 38094 UK Kademoglou et al., 2017
10 100 8800 5300 900 32000 Sweden Bergh et al., 2011
20 100 94 73 11 337 Egypt Abdallah & Covaci 2014
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
4 100 852 604 294 1907 Spain Cristale et al., 2016
63 76 2560 500 <300 64500 Germany Fromme et al., 2014
Daycare centers and
schools
28 - 12000 4100 220 90000 UK Brommer & Harrad, 2015
6 100 2400 1540 - 6150 Norway Cequier et al., 2014
10 100 3500 1900 300 17000 Sweden Bergh et al., 2011
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
15 100 18305 4441 762 142459 Spain Present study
19 100 2490 1800 330 11000 Germany Harrad et al., 2016
8 - - 2400 670 43000 The Netherlands Brandsma et al., 2014
8 - - 1700 360 14000 The Netherlands Brandsma et al., 2014
21 - 15000 3300 270 170000 UK Brommer & Harrad, 2015
20 100 392 135 26 1872 Egypt Abdallah & Covaci 2014
15 87 2165 1760 <2 7415 Kuwait Ali et al., 2013
15 100 786 470 40 4150 Saudi Arabia Ali et al., 2016
15 100 665 245 2 4800 Pakistan Ali et al., 2013
39 100 9137 3700 330 85000 Australia Harrad et al., 2016
Public
microenvironments
(PMEs)
7 100 665 687 169 1004 Spain Present study
3 100 6348 4010 985 14050 Spain Cristale et al., 2016
1 100 179 179 179 179 Spain Cristale et al., 2016
15 100 4700 1970 150 34200 Belgium Van den Eede et al., 2011
11 100 959 629 116 2357 Egypt Abdallah & Covaci 2014
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
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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