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Please cite this article as: Anna S. Young, Environment International, https://doi.org/10.1016/j.envint.2020.106151
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Impact of “healthier” materials interventions on dust concentrations of per-
and polyuoroalkyl substances, polybrominated diphenyl ethers, and
organophosphate esters
Anna S. Young
a
,
b
,
*
, Russ Hauser
a
, Tamarra M. James-Todd
a
, Brent A. Coull
a
,
c
, Hongkai Zhu
d
,
Kurunthachalam Kannan
d
, Aaron J. Specht
a
, Maya S. Bliss
a
, Joseph G. Allen
a
a
Department of Environmental Health, Harvard T.H. Chan School of Public Health, Boston, MA, USA
b
Department of Population Health Sciences, Harvard Graduate School of Arts and Sciences, Cambridge, MA, USA
c
Department of Biostatistics, Harvard T.H. Chan School of Public Health, Boston, MA, USA
d
Department of Pediatrics and Department of Environmental Medicine, New York University School of Medicine, New York, NY, USA
ARTICLE INFO
Handling Editor: Olga Kalantzi
Keywords:
Buildings
Products
Chemicals
Flame retardants
Dust
Ofces
ABSTRACT
Per- and polyuoroalkyl substances (PFAS), polybrominated diphenyl ethers (PBDEs), and organophosphate
esters (OPEs) are found in building materials and associated with thyroid disease, infertility, and impaired
development. This study’s objectives were to (1) compare levels of PFAS, PBDEs, and OPEs in dust from spaces
with conventional versus “healthier” furniture and carpet, and (2) identify other product sources of ame re-
tardants in situ. We measured 15 PFAS, 8 PBDEs, and 19 OPEs in dust from ofces, common areas, and class-
rooms having undergone either no intervention (conventional rooms in older buildings meeting strict re codes;
n =12), full “healthier” materials interventions (rooms with “healthier” materials in buildings constructed more
recently or gut-renovated; n =7), or partial interventions (other rooms with at least “healthier” foam furniture
but more potential building contamination; n =28). We also scanned all materials for bromine and phosphorus
as surrogates of PBDEs and OPEs respectively, using x-ray uorescence. In multilevel regression models, rooms
with full “healthier” materials interventions had 78% lower dust levels of PFAS than rooms with no intervention
(p <0.01). Rooms with full “healthier” interventions also had 65% lower OPE levels in dust than rooms with no
intervention (p <0.01) and 45% lower PBDEs than rooms with only partial interventions (p <0.10), adjusted for
covariates related to insulation, electronics, and furniture. Bromine loadings from electronics in rooms were
associated with PBDE concentrations in dust (p <0.05), and the presence of exposed insulation was associated
with OPE dust concentrations (p <0.001). Full “healthier” materials renovations successfully reduced chemical
classes in dust. Future interventions should address electronics, insulation, and building cross-contamination.
1. Introduction
The products that furnish buildings contain complex mixtures of
chemicals (Lucattini et al., 2018). Many chemicals can migrate out of
products and into indoor environments (Mitro et al., 2016; Rauert et al.,
2014; Tokranov et al., 2018), exposing people via dust ingestion, inha-
lation, and dermal sorption (Johnson-Restrepo and Kannan, 2009; Kim
et al., 2019; Poothong et al., 2020). Several studies have linked the
concentrations in dust and air to higher body burdens of these chemicals
(Cequier et al., 2015; Frederiksen et al., 2010; Koponen et al., 2018;
Watkins et al., 2011). Thus, the pathway from source to environmental
media, exposure, and body burden has been well established for
chemicals used in building products. An upstream intervention on the
product sources of harmful chemical classes presents a public health
opportunity to reduce exposure. Recently, some organizations have
incorporated the science on toxic chemical classes into their product
procurement practices and moved towards purchasing “healthier” ma-
terials, but there is still a lack of peer-reviewed evidence on the exposure
reduction benets of using these materials. Per- and polyuoroalkyl
substances (PFAS) and chemical ame retardants (FRs) are two classes
of chemicals found in building furnishings that are of particular concern.
PFAS are a class of extremely persistent, highly uorinated aliphatic
* Corresponding author at: 401 Park Dr, 4W Suite 405, Boston, MA 02215, USA.
E-mail address: ayoung@mail.harvard.edu (A.S. Young).
Contents lists available at ScienceDirect
Environment International
journal homepage: www.elsevier.com/locate/envint
https://doi.org/10.1016/j.envint.2020.106151
Received 16 June 2020; Received in revised form 31 August 2020; Accepted 17 September 2020
Environment International xxx (xxxx) xxx
2
chemicals widely used as stain- and water-repellant coatings in furni-
ture, carpet, clothing, disposable food packaging, non-stick pans, and
building materials (Sunderland et al., 2018). Research has found PFAS
exposure to be associated with thyroid disease, impairment of fetal
development, immune system suppression, high cholesterol, and
possibly obesity, diabetes, kidney cancer, and testicular cancer (Barry
et al., 2013; Liew et al., 2018; Lin et al., 2019; Rappazzo et al., 2017;
Stanifer et al., 2018; Sunderland et al., 2018; Xiao et al., 2019).
Polybrominated diphenyl ethers (PBDEs) are a class of 209 conge-
ners used as chemical ame retardants in foam furniture, carpet, elec-
tronics, and insulation to comply with re codes (Cooper et al., 2016;
Jinhui et al., 2017; Kemmlein et al., 2003). The increasing use of PBDEs
in foam products across the U.S. market was catalyzed by California’s
rst ammability standard for interior foam lling in upholstered
furniture (Technical Bulletin [TB] 117) passed in 1975 (Dodson et al.,
2017; Petreas et al., 2016). In 1991, a new standard, TB 133, required
even stricter ammability tests for furniture in publicly occupied spaces
(Dodson et al., 2017). Human exposure to PBDEs is associated with
adverse health effects on thyroid function, reproductive success, and
reproductive and brain development (Allen et al., 2016; Boas et al.,
2012; Choi et al., 2019; Czerska et al., 2013; Johnson et al., 2013;
Linares et al., 2015; Mumford et al., 2015; Vuong et al., 2018).
Even when certain chemicals in the PFAS and PBDE classes were
found to be harmful and were largely removed from production, they
have often been replaced with other similar chemicals with human
health concerns. For example, two of the most well-known PFAS, per-
uorooctanoate (PFOA) and peruorooctane sulfonate (PFOS), were
voluntarily phased out by manufacturers in the early 2000s in the U.S.
(Wang et al., 2015a). However, new replacement PFAS, including short-
chain alternatives and precursors that break down into legacy PFAS,
may be just as concerning for health (Wang et al., 2015b, 2013; Z. Wang
et al., 2017). In fact, there are over 4700 different PFAS currently
available (OECD, 2018). Due to the health concerns of PBDEs, two
commercial ame retardant mixtures, penta-BDE and octa-BDE, were
also voluntarily phased out in the U.S. in 2004, followed by deca-BDE in
2013 (Dodson et al., 2012). Similar to the substitution of legacy PFAS,
PBDEs were often simply replaced with other organohalogenated ame
retardants such as organophosphate esters (OPEs), which have been
found in recent research to be associated with similar adverse effects on
thyroid function, pregnancy outcomes and fertility, and development
(Carignan et al., 2018, 2017; Doherty et al., 2019a, 2019b; Meeker and
Stapleton, 2010; Messerlian et al., 2018; Preston et al., 2017; Wang
et al., 2019).
PFAS, PBDEs, and OPEs have been used ubiquitously in building
furnishings and have been found in the blood or urine of over 90% of
Americans as a result (Calafat et al., 2007; Ospina et al., 2018; Sj¨
odin
et al., 2008). Despite widespread concerns about these chemical classes,
to our knowledge there have been no studies assessing interventions to
reduce PFAS indoors. Only a few research studies have been able to
evaluate interventions on ame retardants. Stubbings et al. found that
several OPEs were signicantly reduced in dust after swapping in FR-
free nap mats in childcare centers (Stubbings et al., 2018). Another
study reported a signicant decline in FRs on hand wipes of gymnasts
after replacing pit foam with FR-free alternatives (Dembsey et al., 2019).
Recently, the loosening of TB 117 to the less strict TB 117-2013 (effec-
tive as of 2014) and the repeal of TB 133 in 2019 have laid the foun-
dation to enable reductions in chemical ame retardant use in
upholstered furniture (State of California, 2019, 2013).
One of the limitations with previous studies of interventions was the
inability to non-destructively screen chemicals in products in situ in
order to identify important driving product sources and to inform future
interventions. Research has shown that handheld x-ray uorescence
(XRF) instruments can measure elemental bromine as a reliable way to
screen products for the total potential content of PBDEs and other
brominated FRs (Allen et al., 2008a; Gallen et al., 2014; Petreas et al.,
2016; Stapleton et al., 2011). In fact, XRF-measured bromine levels in
furniture and/or electronics have been signicantly associated with
PBDE concentrations in house dust (Allen et al., 2008a) and human
blood (Imm et al., 2009). Similarly, XRF-measured phosphorus may also
serve as a useful surrogate for total OPE levels in products, although
there has been less research and validation of this method (Petreas et al.,
2016).
For this study, we evaluated a chemical class-based “healthier” ma-
terials intervention at a university which, since 2017, has renovated
over a dozen buildings with furniture, carpet, and other products spec-
ied by manufacturers as free of the entire classes of PFAS and ame
retardants (often customized as so for the rst time). The objectives of
this study were to: (1) measure levels of 42 PFAS, PBDEs, and OPEs in
indoor dust from spaces with “healthier” materials compared to analo-
gous samples from conventional spaces, and (2) identify important
sources of ame retardants in the buildings using XRF product screening
for bromine and phosphorus.
2. Methods
2.1. Study design
We sampled indoor dust in 47 rooms from 21 buildings at a uni-
versity in the northeastern U.S. Specically, we studied six ofce suites,
23 common rooms, and 18 classrooms.
We rst selected as many rooms as possible that were renovated with
“healthier” materials (n =22), including furniture and carpet specied
by manufacturers in product purchasing agreements as free of all PFAS
and chemical FRs. “Healthier” furniture included most non-xed fur-
nishings, except electronics or any existing furniture that was kept.
Trace contamination thresholds for “healthier” product procurement
agreements were 100 ppm for PFAS and 1000 ppm for FRs by weight
(the same FR threshold as TB 117-2013). The furniture also could not
contain over 1% polyvinyl chloride. The interventions were conducted
in 2017 through 2019, mostly in pre-existing buildings (86%).
We selected an equivalent set of carpeted room types that did not
undergo “healthier” materials interventions and were refurnished as
recently as possible (n =25). In ve cases, we were able to sample a
conventional room in a building that had a “healthier” room on a
different oor in order to maximize comparability. Occasionally sam-
pling multiple rooms within a building, if different enough in function or
characteristics, also helped increase the sample size of available
“healthier” rooms and recently renovated conventional spaces.
Construction years were similar for “healthier” rooms (median 1970;
range 1863–2018) and conventional rooms (median 1965; range
1863–2017). Years of last refurnishing ranged from 2017 to 2019 (me-
dian 2018) for “healthier” and 2001 to 2019 (median 2016) for con-
ventional rooms. Rooms were vacuumed at least twice weekly and never
had stain-repellant coatings applied to carpets. Only two conventional
rooms were naturally, not mechanically, ventilated.
2.2. “Healthier” materials intervention classications
Given the spectra of “healthier” materials statuses in the studied
rooms, we decided to categorize the spaces into three groups: no inter-
vention, partial intervention, and full “healthier” materials intervention.
We avoided more than three categories to ensure reasonable sample
sizes within each group and parsimony for statistical modeling.
•“No intervention” =rooms in older buildings constructed before the
2004 phase-out of most PBDEs and that had foam furniture meeting
historically stringent ammability standards (TB 117 or 133).
•“Full ‘healthier’ materials intervention” =rooms that were (1)
renovated with furniture and carpet specied as free of PFAS and
FRs, and (2) in buildings built after the 2004 PBDE phase-out or that
had renovated the entire oor to be “healthier” (i.e. no adjacent
contamination from conventional spaces).
A.S. Young et al.
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•“Partial intervention” =All other rooms, including (1) conventional
spaces built after the 2004 PBDE phase-out or likely with fewer FRs
in foam furniture (under the TB 117-2013 option), and (2) rooms
with “healthier” materials with potential cross contamination from
the old building and connection to adjacent conventional spaces.
Note than none of the buildings were constructed between 2002 and
2006, so any transition period after the 2004 phase-out for manufac-
turers to sell old stocks of products would likely not impact our classi-
cations of buildings.
2.3. Dust sample collection
We collected a dust sample in each room between January and
March 2019, at least two months after any intervention. We asked
custodial crews to leave the space unvacuumed for 2–3 days to increase
dust mass. For each sample, we vacuumed oor dust (including under-
neath furniture) for 10 min using a vacuum cleaner (Dyson CY18) with
an attached crevice tool that housed a cellulose extraction thimble
(secured with a nitrile rubber o-ring) to collect dust. Dust only came into
contact with crevice tools, which were cleaned with isopropyl alcohol
and tap water between samples. Our sampling followed published pro-
tocols (Allen et al., 2008b; Fraser et al., 2013). After vacuuming, the
thimbles were stored in polypropylene centrifuge tubes in polyethylene
bags in a −13 ◦C freezer. We collected ve eld blanks by carrying
unopened centrifuge tubes on multiple different sampling days.
2.4. PFAS and FRs in dust
Dust samples and eld blanks were analyzed for 15 PFAS, eight
PBDEs, and 19 OPEs, following previously published methods (Johnson-
Restrepo and Kannan, 2009; Kim and Kannan, 2007; Kim et al., 2019,
2017). The PFAS analytes included PFOS, PFOA, peruorohexanoate
(PFHxA), peruorohexane sulfonate (PFHxS), peruorooctane sulfon-
amide (FOSA), peruoroheptanoate (PFHpA), peruoropentanoate
(PFPeA), peruorononanoate (PFNA), peruorobutane sulfonate
(PFBS), peruorodecane sulfonate (PFDS), peruorobutanoate (PFBA),
peruorodecanoate (PFDA), peruoroundecanoate (PFUnDA), per-
uorododecanoate (PFDoDA), and n-methyl peruorooctane sulfona-
midoacetic acid (N-MeFOSAA). The PBDE analytes included congeners
28, 47, 99, 100, 153, 154, 183, and 209. The OPE analytes included tris
(2-butoxyethyl) phosphate (TBOEP), tris(1-chloro-2-propyl) phosphate
(TCIPP), tris(1,3-dichloro-2-propyl) phosphate (TDCIPP), triphenyl
phosphate (TPHP), tris(2-chloroethyl) phosphate (TCEP), 2-ethylhexyl
diphenyl phosphate (EHDPP), isodecyl diphenyl phosphate (IDDP), tri-
iso-butyl phosphate (TIBP), tripropyl phosphate (TPP), cresyl diphenyl
phosphate (CDPP), tert-butylphenyl diphenyl phosphate (BPDP), tri-n-
butyl phosphate (TNBP), tetrakis(2-chloroethyl) dichloroisopentyl
diphosphate (V6), bisphenol a bis(diphenyl phosphate) (BDP), resor-
cinol bis(diphenyl phosphate) (RDP), tris(2-ethylhexyl) phosphate
(TEHP), tris(methylphenyl) phosphate (TMPP), triethyl phosphate
(TEP), and tris(p-tert-butylphenyl) phosphate (TBPHP).
PFAS concentrations were measured using high-performance liquid
chromatography (HPLC) coupled with electrospray triple quadrupole
tandem mass spectrometry (ESI-MS/MS) and monitored by multiple
reaction monitoring mode under negative ionization. OPEs were
measured with HPLC coupled with ESI-MS/MS, using electrospray
positive ionization multiple reaction monitoring. PBDEs were measured
with a gas chromatographer coupled with a mass spectrometer (GC–MS)
using electronic impact ionization mode. The limits of detection (LODs)
ranged from 0.1–0.8 ng/g for OPEs, 0.09–4.5 ng/g for PBDEs, and
0.06–1.5 ng/g for PFAS.
Before instrumental analysis, the dust samples were rst sieved
through a 150-µm stainless steel mesh. The samples (with masses
ranging from 0.2 to 0.5 g) were then spiked with 30 ng each of labeled
surrogate standard mixture. The samples were extracted with methanol
(3 mL) under mechanical oscillation (1 h) followed by ultrasonication
(30 min). Then, the dust extracts were centrifuged (3500g, 10 min) and
transferred into new polypropylene tubes. The extraction was repeated
twice with acetonitrile (3 mL) and ethyl acetate (3 mL), after which the
extracts were combined and evaporated to 3 mL under a gentle stream of
nitrogen. The extracts were split into three aliquots for the measurement
of PFAS, PBDEs, and OPEs. The extracts evaporated to near dryness and
were then reconstituted with 200 µL of solvents: methanol for PFAS,
hexane for PBDEs, and water/methanol (4/6; v/v) for OPEs. Before
instrumental analysis, the sample extracts were ltered through a 0.2
µm nylon lter into glass vials.
All eld blanks either had values below the LOD or detected con-
centrations well below the levels in samples. The spiked recovery results
were 67.4–104% for PFAS, 59.0–128% for OPEs, and 101–115% for
PBDEs, respectively. We split and analyzed duplicates of seven dust
samples in the same batch using the same method for quality assurance
and quality control (QA/QC) and calculated median relative percent
differences of 0% (range: −62 to 190%) for PFAS, −3.2% (range: −50 to
80%) for PBDEs, and 0% (range: −96 to 52%) for OPEs. There was some
variability in chemical levels in the duplicates, likely due to the natural
heterogeneity of dust. Detailed QA/QC data and typical chromatograms
of target analytes are provided in Tables S1–S3 and Figs. S1–S3 in the
Supplemental Materials.
2.5. Screening products for bromine and phosphorus
We returned to all rooms between September and October 2019 to
characterize potential sources of FRs and corroborate differences in
content of conventional versus “healthier” products. Only two of the 47
rooms had a few missing furniture pieces since the earlier dust sampling.
We noted the counts and surface areas of all product types. For uphol-
stered foam furniture, we recorded its compliance to a particular am-
mability standard (if tag was present).
We screened the materials within each room for bromine and
phosphorus content (as surrogates for PBDEs and OPEs, respectively)
using two handheld, non-destructive XRF instruments. XRF has been
reliably used in numerous previous studies to measure bromine as a
surrogate for PBDEs (Allen et al., 2008a; Gallen et al., 2014; Imm et al.,
2009; Petreas et al., 2016; Stapleton et al., 2011), By contrast, XRF-
measured phosphorus as a surrogate for OPEs in products has been
less studied and should be interpreted with more caution. One study
analyzed phosphorus in furniture using XRF and advised against clas-
sifying a product as OPE-free based on XRF data alone (Petreas et al.,
2016). However, phosphorus measurements should be a useful surro-
gate to statistically model patterns in sources of OPEs because exposure
misclassication issues arise from models relying on raw counts of
furniture and electronics in rooms (Allen et al., 2008a). Fluorine (a PFAS
surrogate) was too light an element to be detectable.
The Olympus Innov-x XRF with a consumer product (RoHS) testing
mode was used to measure bromine but could not detect phosphorus. We
took minimum 15-second measurements with this instrument by hold-
ing it still against the product. The more sensitive Thermo Fisher Sci-
entic Niton xl3t GOLDD+XRF was used to measure phosphorus. We
could only use the TestAllGeo mode (hereby referred to as “geo”)
designed for geological samples. Spectral data from this instrument can
also be analyzed for bromine, and post-hoc we found extremely strong
correlations between the two instruments/modes, even when measured
on different spots of the products, which suggests the geo mode did not
produce signicant interference in modeling (foam furniture: Spearman
r =0.91, p <0.0001; all products: r =0.65, p <0.0001). We pro-
grammed the Niton XRF to collect measurements with 15 s in the main
element range and 30 s in the light element range (an advanced feature
to capture phosphorus). The XRF measurements penetrate deep enough
to capture concentrations inside foam lling of furniture. The LODs were
expected to be approximately 10 ppm for bromine and 20 ppm for
phosphorus.
A.S. Young et al.
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We took at least one measurement (on each XRF) for most products
in the rooms, but focused on foam furniture, carpet, insulation, and
electronics as FR sources. For carpet and televisions, we measured two
different spots for up to two different items of each model. Measure-
ments of electronics were taken on the external plastic housing (when
possible). For foam furniture, we also measured two different spots (on
both the seat cushion and seatback) for up to two items. Some products
on ceilings (e.g. insulation, routers, speakers) could not be reached for
scanning.
2.6. Statistical analyses
We blank-corrected the concentrations of chemicals in dust by sub-
tracting the average eld blank value. Concentrations below the LOD in
dust were substituted with half the LOD (Hornung and Reed, 1990). For
statistical models, we log-transformed the concentration data (due to
non-normal distributions in histograms and Shapiro-Wilk tests). First,
we performed principal component analysis on scaled and centered
PBDE concentrations to identify congener groupings according to com-
mercial ame-retardant mixtures. We evaluated principal components
together explaining at least 70% of the variance.
For modeling covariates, we calculated room loadings of bromine
and phosphorus from foam furniture and electronics separately by tak-
ing the sum of each unique product type’s mean element concentration
multiplied by its surface area and count, following Eq. (1) and a previous
protocol (Allen et al., 2008a). For foam furniture loadings, we included
chairs, armchairs, couches, foam ottomans, pillows, and foam drawer
tops. Electronics included televisions, computers, projector systems and
screens, routers, keyboards, computer mice, telephones, speakers, TV
remotes, tablets, power strips, plug ports, printers, portable scanners,
oor outlets, server kits, audio radiators, and DVD or video game
players. Non-detect XRF measurements were conservatively substituted
with zero for loading calculations, but we added one to all concentra-
tions so we could log-transform the non-normally distributed data for
modeling. Occasionally when some ceiling electronics (such as routers)
were not able to be measured, we borrowed the result for an identical
product elsewhere in the building.
Loading =∑
products
(mean concentration*surface area*count)(1)
We conducted multilevel models for the chemical classes in dust,
with a random intercept for the building because some rooms were
sampled within the same building. As covariates, we included the three-
category “healthier” materials intervention status (conducted with the
reference as “none” and as “partial” based on interpretability and largest
sample size, respectively), a binary variable for exposed insulation in or
immediately adjacent to the room (reference =no), tertiles of element
loadings from foam furniture (reference =low), and tertiles of element
loadings from electronics (reference =low). We did not use an element
loading variable for insulation because insulated pipes on ceilings were
not always accessible for measurement. To present model results, we
transformed estimates back to the linear scale and reported them as
percent differences (since the dependent variables were log-
transformed).
To model the association between furniture characteristics and
element concentrations in products, we conducted multilevel models
with two random intercepts for the room and specic product (to ac-
count for expected correlation between products within a room and
between replicate items or cushions of the same product). We included
three categorical covariates: the ammability standard the furniture was
labeled to comply with (reference =“healthier”; TB 117-2013 with no
added FRs; TB 117 or 117-2013 with added FRs; TB 133; and unlabeled),
type of foam furniture (reference =chair; fully upholstered armchair;
couch; ottoman), and measurement spot (reference =seat; seatback
cushion). We evaluated statistical signicance at
α
=0.05 and sugges-
tive evidence at
α
=0.10. All statistical analyses were performed in R
(version 3.3.1).
3. Results
3.1. Per- and polyuoroalkyl substances in dust
PFAS were detected in 100% of our dust samples (n =47) from of-
ces, common areas, and classrooms (Table S4). The most frequently
detected PFAS were PFHxA at 98%, PFOS at 98%, PFOA at 75%, PFHxS
at 64%, FOSA at 60%, and PFHpA at 51%. Tables S4 and S5 present
summary statistics by intervention and room type.
The total concentrations of PFAS in dust tended to be the lowest in
the rooms with full “healthier” materials interventions and highest in
rooms with no interventions (Fig. 1). The geometric mean total Σ15PFAS
concentrations were 262 ng/g (range: 18.1–8310 ng/g) overall, 481 ng/
g (225–1140 ng/g) in rooms with no interventions, 252 ng/g
(18.1–8310 ng/g) in rooms with partial “healthier” materials in-
terventions, and 108 ng/g (43.6–243 ng/g) in rooms with full
Fig. 1. Boxplots of concentrations (ng/g) of Σ
15PFAS, Σ8PBDEs, and Σ19 OPEs in indoor dust samples by “healthier” materials intervention status, with outliers
excluded to obtain tighter y-axis scales (n =47).
A.S. Young et al.
Environment International xxx (xxxx) xxx
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“healthier” materials interventions (Table S4). PFHxA, PFOS, and PFOA
dominated the geometric mean proles of PFAS analytes in dust across
intervention statuses, although concentrations were much lower in the
“healthier” rooms (Fig. 2).
In a multilevel regression model, rooms with full “healthier” mate-
rials interventions had 78% (95% CI: 38–92%) lower Σ15 PFAS concen-
trations in dust than rooms with no interventions (p =0.006) (Table 1).
The spaces with full “healthier” materials interventions even had 68%
(95% CI: 15–88%) lower Σ15 PFAS concentrations in dust than rooms
with partial interventions (p =0.02). About 71% of the variability in log
PFAS concentrations was attributable to differences between rooms
within a building as opposed to differences across buildings.
3.2. Polybrominated diphenyl ethers in dust
BDE-209, BDE-100, BDE-99, and BDE-47 were each detected in
100% of our dust samples (Table S4). The other four PBDE congeners
were detected in at least 80.9% of samples. The PBDE congeners
grouped into three principal components that aligned well with com-
mercial ame retardant mixture formulas: penta-BDE (congeners 28, 47,
99, 100, and 153), octa-BDE (congeners 154 and 183), and deca-BDE
(congener 209) (Fig. S4). The geometric mean proles of PBDEs in
Fig. 2. Geometric mean concentrations (ng/g) of each main PFAS, PBDE, and OPE analyte in indoor dust samples (n =47) by “healthier” materials intervention
status. Note: chemicals that had relatively very small geometric mean concentrations were shaded grey and not distinguished.
Table 1
Results from multilevel models of the impact of “healthier” materials intervention status, presence of exposed insulated pipes, ame retardant-related element loadings
in furniture, and element loadings in electronics (bromine [Br] for PBDEs; phosphorus [P] for OPEs) on total concentrations (ng/g) of 15 PFAS, 8 PBDEs, and 19 OPEs in
indoor dust samples (n =47).
% Change (p value)
a
Covariate n Sum of 15 PFAS Sum of 8 PBDEs Sum of 19 OPEs
“Healthier” Materials Intervention Status
No intervention 12 Reference 47% (0.3) Reference −33.7% (0.2) Reference 7.93% (0.7)
Partial intervention 28 −32% (0.3) Reference
b
50.7% (0.2) Reference
b
−7.34% (0.7) Reference
b
Full intervention 7 ¡77.9% (0.006)** ¡67.5% (0.02)* −17.4% (0.6) ¡45.2% (0.09) . ¡65.0% (<0.001)*** ¡62.2% (<0.001)***
Exposed Insulation in Room
No 41 Not included Reference Reference
Yes 6 Not included 24.8% (0.5) 176% (<0.001)***
Br or P Loading
c
from Foam Furniture For bromine: For phosphorus:
Low 16 Not included Reference Reference
Medium 15 Not included −12.6% (0.6) −26.8% (0.2)
High 16 Not included 86.2% (0.04)* 0.393% (1.0)
Br or P Loading
c
from Electronics
Low 16 Not included Reference Reference
Medium 15 Not included 42.1% (0.2) 43.9% (0.09) .
High 16 Not included 169% (<0.001)*** −3.32% (0.9)
PFAS =per- and polyuoroalkyl substances; PBDEs =polybrominated diphenyl ethers; OPEs =organophosphate esters.
. p <0.10.
* p <0.05.
** p <0.01.
*** p <0.001.
a
Chemical concentrations were log-transformed in the multilevel models, but estimates are presented as the percent change on the linear scale.
b
The models were conducted a second time with ‘partial intervention’ as the reference category in order to assess any improvements of the full intervention over the
partial one and to increase statistical power by using the group with the largest sample size.
c
Loadings were calculated as the sum across foam furniture or electronics products of the average element concentration of each unique product type multiplied by
its surface area and count in the space.
A.S. Young et al.
Environment International xxx (xxxx) xxx
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dust were mostly inuenced by deca-BDE, followed by congeners BDE-
47 and BDE-99 commonly used in penta-BDE (Fig. 2).
Total PBDE dust concentrations were similar between rooms with no
interventions and rooms with partial interventions, but the full
“healthier” rooms had lower levels (Fig. 1). Geometric mean concen-
trations of Σ8PBDEs in dust were 1360 ng/g (range: 452–5930 ng/g) in
rooms with no interventions, 1390 ng/g (179–14,200 ng/g) in rooms
with partial “healthier” interventions, and 839 ng/g (414–1570 ng/g) in
rooms with full “healthier” materials interventions (Table S4).
In the multilevel model adjusted for exposed insulation, bromine
loading from foam furniture, and bromine loading from electronics
(Table 1), there was suggestive evidence that rooms with full “healthier”
materials interventions had 45% (95% CI: −1.7 to 71%) lower Σ8PBDEs
concentrations compared to rooms with only partial interventions (p =
0.09). The rooms with full interventions also had 17% lower Σ8PBDEs
concentrations in dust than rooms with no interventions, but this did not
reach statistical signicance (p =0.68), perhaps because of the lower
sample size of the no-intervention reference group. The rooms with
partial versus no interventions also did not have signicantly different
Σ8PBDE dust levels (p =0.16). The PBDE concentrations in dust were
signicantly associated with bromine loadings from both electronics and
from foam furniture. Rooms with high bromine loadings in electronics
had 169% (95% CI: 51–362%) higher levels of Σ8PBDEs in dust than
rooms with low electronic loadings (p =0.0008). Rooms with high
bromine loadings in foam furniture had 86% (95% CI: 2–226) higher
Σ8PBDEs concentrations in dust than rooms with low loadings (p =
0.04). Differences between rooms within a building (as opposed to be-
tween buildings) explained 83% of the variability in log PBDE concen-
trations in dust.
Two buildings in the study were built after the 2013 voluntary phase-
out of BDE-209. In the new building with full “healthier” materials in-
terventions, three sampled rooms had detectable levels of BDE-209 in
dust between 52.9 and 1340 ng/g. In the new building with partial in-
terventions, three rooms had dust concentrations of BDE-209 between
369 and 2680 ng/g.
3.3. Organophosphate esters in dust
Nine OPEs were detected in 100% of dust samples, and another eight
were detected in at least 91% of samples (Table S4). TBOEP, TCIPP,
TDCIPP, and TPHP made up the vast majority of geometric mean pro-
les of OPEs (Fig. 2).
The OPE concentrations were generally lowest in dust from the full
“healthier” rooms, while the rooms with no or partial interventions had
similar concentrations to each other (Fig. 1). Geometric mean concen-
trations of Σ19 OPEs in dust were 30,400 ng/g (range: 13,400–60,300
ng/g) in rooms with no interventions, 37,500 ng/g (10,300–182,000
ng/g) in rooms with partial interventions, and 14,000 ng/g
(6760–30,900 ng/g) in rooms with full “healthier” materials in-
terventions (Table S4).
In a multilevel model controlling for exposed insulation, furniture
phosphorus loading, and electronics phosphorus loading, rooms with
full “healthier” materials interventions had 65% (95% CI: 37–80%)
lower total levels of OPEs in dust than rooms with no interventions (p =
0.0006) (Table 1). Rooms with only partial interventions did not have
signicantly lower Σ19OPEs concentrations in dust than rooms with no
interventions (p =0.7). Rooms with exposed insulation had 176%
higher Σ19 OPEs levels in dust than rooms without visible insulation
(95% CI: 55–396%; p =0.0005), adjusted for intervention status,
furniture loading, and electronics loading. There was suggestive evi-
dence that rooms with medium phosphorus loadings from electronics,
compared to rooms with low loadings, had 43.9% higher Σ19OPEs
concentrations in dust (95% CI: −7 to 118%; p =0.09). Phosphorus
loadings in foam furniture were not statistically signicant predictors of
Σ19OPEs concentrations. All (100%) of the variability in log OPE con-
centrations could be attributed to differences between rooms within a
Table 2
Summary of concentrations (µg/g) of bromine and phosphorus in different
product types in the 47 studied spaces, as measured using portable x-ray uo-
rescence (XRF).
Bromine (µg/g) Phosphorus (µg/g)
Product Type n Median
[Range]
n Median [Range]
Foam furniture
a
293 21.5 [ND,
79,240]
284 585.8 [ND,
22,880]
By ammability standard:
“Healthier”: no FRs or
PFAS
122 10.05 [ND,
4820]
114 457.4 [ND,
4827]
Conv.: TB 117-2013
with no FRs
39 7.7 [ND, 848] 40 732 [ND,
18,430]
Conv.: FRs at TB 117 or
117-2013
25 27.2 [5.9,
46,120]
25 915.9 [ND,
2416]
Conv.: FRs at TB 133 30 917 [2.6,
65,910]
30 493 [ND,
22,880]
Conv,: unlabeled 77 141 [2.2,
79,240]
75 1124 [ND,
13,140]
By product type:
Chair with cushioned
seat
75 25.8 [ND,
38,440]
73 491 [ND, 4450]
Armchair 126 14.95 [ND,
76,270]
120 864.1 [ND,
18,430]
Couch 62 94.6 [2,
79,240]
61 557.7 [ND,
22,880]
Ottoman 30 9.9 [2, 55,510] 30 512.6 [ND,
13,010]
Carpet 91 10.6 [0.2, 305] 91 ND [ND, 1840]
“Healthier”: no FRs 44 7.6 [0.2,
153.1]
44 99.12 [ND,
1840]
Conv. 47 21.2 [1.9, 305] 47 ND [ND, 1835]
Electronics
b
240 20.35 [ND,
147,800]
229 719 [ND,
128,800]
Televisions 51 26 [ND,
133,500]
50 2025 [ND,
42,680]
Computer monitors 30 57.2 [ND, 965] 28 917.2 [ND,
128,800]
Projector systems 13 2.4 [0.4, 5285] 12 25,230 [887.8,
34,740]
Keyboards 32 381.5 [ND,
2931]
32 202.6 [ND, 878]
Mouses 24 247.4 [0.9,
821]
23 426.9 [ND,
5941]
Telephones 12 9.05 [1.1, 395] 12 412 [239.9,
1183]
Audio/video devices 15 2.3 [ND,
147,800]
13 711.8 [154.8,
44,780]
Printers/copiers 5 7.3 [1, 224] 6 39,790 [17,030,
43,250]
Routers/modems 16 0.5 [ND,
108,100]
14 25,020 [ND,
39,450]
Floor outlets 11 4.7 [ND,
23,820]
9 668.6 [ND,
1458]
Projector screens 3 ND [ND, ND] 3 11,490 [ND,
48,350]
Foam pillows 12 5.45 [0.7, 106] 13 418 [ND,
10,240]
Foam drawer tops 5 10.2 [1.7,
12.8]
6 652.1 [551,
883.6]
Exposed insulation 10 16,080 [5.4,
21,030]
10 ND [ND, 1308]
ND =not detected; Conv =conventional; n =number of XRF measurements,
including duplicates on same product. The number of unique items were 105
upholstered foam furniture pieces, 54 carpets, and 170 electronics.
a
Including only chairs, armchairs, couches, and ottomans with foam lling.
b
Including the mentioned product types, DVD players, tablets, power strips,
scanners, TV remotes, audio radiators, plug ports, monitor-free computer sys-
tems, and a server kit.
A.S. Young et al.
Environment International xxx (xxxx) xxx
7
building, so between-building differences did not play a signicant role.
3.4. Product sources of bromine and phosphorus
Foam furniture, carpet, electronics, and exposed insulation (mostly
hot water pipes) were four primary product sources of XRF-measured
bromine and phosphorus in the rooms, and the “healthier” materials
usually had lower levels than conventional materials. Table 2 provides
summary statistics of element concentrations by intervention status and
product type, with additional products summarized in Table S6.
In an unadjusted multilevel regression model, conventional carpet
had 210% (95% CI: 65–484%) higher bromine levels than “healthier”
carpet specied as free of chemical ame retardants (p =0.001).
Phosphorus concentrations in conventional and “healthier” carpets were
not signicantly different.
Foam seating furniture that met the most stringent ammability
standard, TB 133, had 2940% (95% CI: 458–16,200%) higher levels of
bromine than “healthier” furniture specied by manufacturers as free of
all chemical ame retardants (p =0.0002), adjusted for furniture type
and measurement spot (Table 3). Foam seating furniture that met other
stringent ammability standards (TB 117 or TB 117-2013 with added
chemical FRs) had 922% (95% CI: 45–7210%) higher bromine levels
than “healthier” furniture (p =0.026). As expected, conventional
furniture labeled under the newest TB 117-2013 standard to not contain
chemical FRs above 1000 ppm did not have signicantly different
bromine content compared to the university’s “healthier” furniture that
followed the same FR restrictions. The furniture product with the
highest bromine level under the “healthier” category was labeled in the
manufacturer’s specication to be “FR-free” and “bromine-free” but to
contain recycled content.
For phosphorus, median concentrations were lowest in “healthier”
foam seating furniture items, followed by items labeled under the his-
torical, stringent standard TB 133 (Table 2). Conventional furniture with
no ammability standard tag or labeled to have no added ame re-
tardants under the TB 117-2013 option had 495% (95% CI: 48–2300%;
p =0.02) or 2200% (95% CI: 309–12,800%; p =0.001) higher phos-
phorus levels than “healthier” furniture, respectively (Table 3). About
60% of the variability in phosphorus levels and 83% in bromine levels
for foam furniture were attributable to differences between product
types across rooms as opposed to differences in repeated measurements
on different spots or items of a product type in the room.
The 1000 µg/g limit for PBDEs would be equivalent to at most 833
µg/g of bromine (based conservatively on the most brominated
congener, BDE-209), so the bromine concentrations in TB 117-2013
furniture with ‘no added FRs’ seem to mostly be below that regulatory
limit, unlike the furniture meeting stricter regulations (Table 2). The
1000 µg/g limit would be equivalent to up to 170 µg/g phosphorus for
our OPE analytes (specically TEP). Geometric mean concentrations of
phosphorus for all furniture standards were above that limit, although
organophosphate chemicals may also be present in the products as
plasticizers.
4. Discussion
Among 47 university ofces, classrooms, and common rooms,
“healthier” furniture and carpet interventions were associated with
substantially lower total levels of PFAS, PBDEs, and OPEs in dust relative
to conventional spaces.
4.1. Per- and polyuoroalkyl substances
Compared to rooms with no intervention, PFAS were 78% lower in
dust in rooms with full “healthier” materials interventions, which
included carpet and furniture specied by manufacturers to be free of all
PFAS and chemical ame retardants. The specic PFAS detected in dust
mostly included PFHxA, PFOS, and PFOA. These three chemicals have
all been used as coatings on furniture and carpet, so their substantial
reductions in dust align with the “healthier” materials that were inter-
vened on.
Dust concentrations of two historically widely used legacy chem-
icals, PFOS and PFOA, were mostly much lower than in previous studies
of indoor dust in the U.S. (Fig. 3). Some residual contamination of these
two legacy PFAS still persists in the studied buildings despite their
voluntary phase-out by major manufacturers in the early 2000s (Wang
et al., 2015a, 2013), suggesting that products with long life spans as well
as materials inherent to the building continue to be used even after
phase-outs.
At the same time, the newer, short-chain replacement chemical
PFHxA tended to be found at higher dust concentrations in our study
(even in “healthier” rooms) than have been previously measured in dust
collected from homes or ofces in the U.S. between 2000 and 2013
(Fig. 3). The higher levels are likely due to the increasing use of PFHxA
over time and our sampling of mostly recently refurnished buildings
(unlike the case for many homes). It’s possible our studied university
buildings also had higher densities of furniture and carpeting than
homes. PFHxA contamination was still found in the rooms with
“healthier” furniture and carpet, likely from other types of building
materials that were not intervened on, furnishings from any adjacent
conventional spaces, and consumer products that people carry in. Other
than upholstered furniture and carpet, PFHxA (and/or PFOA and PFOS)
have been found in clothing, disposable food packaging, oor waxes,
wood sealants, paints, and other products with water-repellant or stain-
repellant coatings (Herzke et al., 2012; Janousek et al., 2019; Kotthoff
et al., 2015; Liu et al., 2014; Tokranov et al., 2018). Overall, the com-
parisons of our study to previous studies reect the phase-out of legacy
PFAS and the emerging substitution of other PFAS like PFHxA.
Table 3
Results from multilevel models of the impact of ammability standard, furniture
type, and type of cushion measured on concentrations (µg/g) of bromine and
phosphorus in foam furniture in 47 studied spaces as measured with a portable x-
ray uorescence (XRF) instrument.
% Change (p value)
a
Covariate n Bromine Phosphorus
Intervention Status
“Healthier”: no FRs (or
PFAS)
125 Reference Reference
Conv.: no FRs at TB 117-
2013
40 −25.2% (0.7) 2200% (0.001)
**
Conv.: FRs at TB 117 (-2013) 25 922% (0.03)* 566% (0.1)
Conv.: FRs at TB 133 30 2940%
(<.001)
***
103% (0.5)
Conv.: unlabeled 77 1510%
(<.001)
***
495% (0.02)*
Foam Furniture
Chair 75 Reference Reference
Armchair 129 −17.2% (0.7) 28.3% (0.7)
Couch 63 632% (0.004)
**
58.3% (0.6)
Ottoman 30 56.1% (0.5) −45.3% (0.5)
Spot Measured
Seat Cushion 199 Reference Reference
Seatback Cushion 98 −11.9% (0.4) ¡59.9%
(<.001)
***
FRs =added chemical ame retardants; TB =technical bulletin of California
furniture ammability standard; Conv =conventional.
a
Element concentrations were log-transformed in the multilevel models, but
estimates are presented as the percent change on the linear scale.
*
p <0.05.
**
p <0.01.
***
p <0.00.
A.S. Young et al.
Environment International xxx (xxxx) xxx
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4.2. Polybrominated diphenyl ethers
We found suggestive evidence that PBDE levels in dust were 45%
lower in rooms renovated with full “healthier” materials interventions
that those with only partial interventions. BDE-209 was the congener
detected at the highest geometric mean concentrations by far and with
the highest absolute reduction due to the full “healthier” materials
intervention, which reects its relatively recent phase-out by manufac-
turers a decade after the other measured PBDEs.
The PBDE dust concentrations were signicantly higher in rooms
with high bromine loadings from foam furniture compared to rooms
with low loadings, indicating that upholstered foam furniture is indeed
an important source of FRs. The XRF measurements of products in the
rooms veried that there were very low levels of bromine, a surrogate
for PBDEs, in most “healthier” furniture specied by manufacturers to be
free of chemical FRs. In fact, foam furniture meeting the historic, most
stringent ammability standard (TB 133) had substantially higher
bromine levels than “healthier” furniture. Similarly, “healthier” carpet
specied as FR-free had signicantly lower bromine levels than con-
ventional carpet. The mean bromine concentrations in “healthier” foam
furniture and carpet in this study were lower than a few previous studies
that employed XRF measurements (Abbasi et al., 2016; Allen et al.,
2008a; Gallen et al., 2014; Imm et al., 2009). The intervention on
“healthier” furniture and carpet was associated with lower levels of a
PBDE surrogate in the materials as well as lower resulting levels of
PBDEs in dust.
Other sources besides furniture and carpet likely contributed to
PBDEs in dust, which motivates the need to continue developing in-
terventions on other product categories. For one, plastic electronics like
televisions in the rooms often had very high bromine levels compared to
foam furniture. In addition, rooms with high bromine loadings from
electronics had higher PBDE levels in dust than those with low loadings.
Since PBDEs have been added to electronics as additive (not reactive)
FRs, they are not covalently bound to the polymer and can migrate out of
products during normal use (Rauert et al., 2014; Rauert and Harrad,
2015; Webster et al., 2009).
The building itself, and the materials behind its walls, may be an
important source of PBDE contamination, not just the furnishings inside
it. PBDEs, especially BDE-209, have been historically used in poly-
urethane foam wall insulation, insulated hot water pipes, and other
construction materials (Duan et al., 2016; Vojta et al., 2017). Although
we may not have had enough statistical power to evaluate the impact of
visibly exposed insulation in six of the studied rooms on dust levels of
PBDEs, we did measure high levels of bromine in the insulation that
were comparable to concentrations in conventional foam furniture. In
addition, we did not see a signicant difference between rooms with no
interventions and rooms with only partial interventions; PBDEs were
only lower in dust in the spaces with a full intervention (“healthier”
materials and either located in a newer building or not connected to
conventional rooms). Furthermore, we found that 17% of the variability
in log PBDE concentrations could be explained by differences across
buildings as opposed to between different rooms within a building,
suggesting that the building does play a role. By contrast, 0% was
explained by between-building differences for OPEs, which are rela-
tively newer substitute ame retardants as well as plasticizers.
The geometric mean PBDE levels in dust samples from this study’s
recently refurnished conventional or “healthier” rooms were substan-
tially lower than most previous studies of homes or ofces in the U.S.
(sampled between 2000 and 2015), reecting voluntarily reductions in
manufacturing of these chemicals (Fig. 3). However, in two newly
constructed buildings in our study that opened a few years after the
2013 phase-out of BDE-209, the dust concentrations were still detectable
at up to 2680 ng/g. So, there were residual PBDEs in even the newest
buildings that theoretically should not be contaminated. Despite phase-
outs, PBDEs may take some time to be replaced with new products in
warehouse stocks, can persist in older materials and buildings, and may
be carried over into new materials through the recycling of older plastics
(Abbasi et al., 2015; Y. Li et al., 2019; Turner and Filella, 2017; Vojta
Fig. 3. Central tendency concentrations (medians or geometric means) of select PFAS, PBDE, and OPE analytes (ng/g) in indoor dust in this study (the two bars on
the left represent full “healthier” spaces versus other spaces), compared to select previous studies of dust in the United States, sorted chronologically by sampling
year. Note: Numbers on top of stacked bars indicate the number of the selected analytes that were measured in the study (a few studies did not measure all visualized chemicals).
References: Allen et al., 2008a; Batterman et al., 2010; Bennett et al., 2015; Bi et al., 2018; Castorina et al., 2017; Dodson et al., 2012, 2017; Fraser et al., 2013; Goosey
and Harrad, 2011; Hammel et al., 2017; Harrad et al., 2008; Hoffman et al., 2015; Karaskova et al., 2016; Knobeloch et al., 2012; Li et al., 2019a; Percy et al., 2020;
Scher et al., 2019; Schreder and La Guardia, 2014; Stapleton et al., 2005, 2009, 2014; Strynar and Lindstrom, 2008; Vykoukalov´
a et al., 2017; Watkins et al., 2013;
Whitehead et al., 2013; Wu et al., 2015.
A.S. Young et al.
Environment International xxx (xxxx) xxx
9
et al., 2017). Thus, these ame retardants will likely contaminate
buildings at some level for years to come.
4.3. Organophosphate esters
OPEs in dust were signicantly lower (65%) in spaces with full
“healthier” materials interventions than rooms with none. The presence
of exposed insulation in the rooms was also signicantly associated with
OPE contamination in dust, and phosphorus-containing electronics in
the rooms may be another contributing source too. Our ndings are
supported by what we know of product sources of these chemicals. The
main OPE chemicals present in the dust samples were TBOEP, TCIPP,
TDCIPP, and TPHP, which are used as ame retardants in upholstered
foam furniture, carpet, electronics, and/or building insulation (Abbasi
et al., 2016; Bello et al., 2018; Kajiwara et al., 2011; Kemmlein et al.,
2003; Stapleton et al., 2012; van der Veen and de Boer, 2012; Zheng
et al., 2017). Rooms with only partial interventions did not signicantly
differ from those with no intervention, demonstrating the importance of
cross-contamination of “healthier” single rooms from adjacent conven-
tional spaces and/or the contributions from other materials besides FR-
free foam furniture in partially renovated rooms.
Phosphorus loadings from foam furniture in the rooms were not
statistically signicant predictors of OPE levels in dust. However, the
intervention on “healthier” furniture (and carpet) was associated with a
signicant reduction in OPEs in dust. In addition, XRF product mea-
surements did conrm that the “healthier” furniture items had relatively
low phosphorus levels, although XRF-measured phosphorus as a surro-
gate for OPEs has not been as studied or well-established as bromine. In
fact, the conventional foam furniture items meeting TB 117-2013
without added FRs still had signicantly higher levels of phosphorus
than the “healthier” furniture specied by manufacturers to not contain
FRs. This difference could arise from the additional avoidance of poly-
vinyl chloride (and thus plasticizers) to at least below 1% in the
“healthier” furniture. The conventional furniture meeting older, more
stringent ammability standards did not have statistically higher phos-
phorus levels than “healthier” foam furniture (except for the ‘unlabeled’
group), which is likely due to the historic preference towards PBDEs,
which were only more recently substituted with OPEs. Furthermore, the
mean (476 ppm) and maximum (4830 ppm) XRF-measured phosphorus
levels in the “healthier” foam furniture items were lower than were
found in furniture by one previous study using a different instrument
(mean: 2060; max: 8830 ppm) (Petreas et al., 2016).
We also may not have seen a statistically signicant impact of
phosphorus loadings from foam furniture on OPE dust levels because
there are so many more product sources (with added FRs or plasticizers)
than for PBDEs. This may also explain why the studied rooms had
signicantly higher OPE concentrations compared to previous studies of
home indoor dust in the U.S., although we studied non-residential
buildings that may have higher densities of furniture and more recent
renovations than many homes (Fig. 3). Compared to previous research
of homes sampled between 2000 and 2015, our study did have lower
geometric mean or median levels of TDCIPP, which has been suggested
to be phased out in favor of TCIPP in foam products after TDCIPP was
recently added as a carcinogen to California’s Proposition 65 list
(Cooper et al., 2016). TBOEP was the most dominant analyte in our
study, and it may be a ame-retardant substitute (or plasticizer)
increasing in use, too. Although many prior studies did not measure
TBOEP, median levels of TBOEP in dust in our studied rooms, excluding
the full “healthier” rooms, were higher than three previous studies of
home dust collected between 2011 and 2015 (Fig. 3) (Bi et al., 2018;
Dodson et al., 2012; W. Li et al., 2019). TBOEP is used in furniture, oor
nishes, rubber, plastic, lacquers/paints, and wallpaper (van der Veen
and de Boer, 2012; Y. Wang et al., 2017).
4.4. Alternative strategies to PFAS and ame retardants
Because PFAS and ame-retardant chemicals are usually non cova-
lently bound additives, they can be removed as ingredients without
sacricing the functional integrity of the product. In addition, the costs
of the furniture to the university were reported to be lower when
chemical ame retardants were not added. There is even evidence that
disputes the hypothesized effectiveness of using chemical ame re-
tardants in furniture foam lling to protect against residential res
(Rodgers et al., 2019; Shaw et al., 2010; US Consumer Product Safety
Commission, 2012). At the building level, public spaces often have
extensive measures to protect against res, including smoke detectors,
smoking prohibitions, and sprinkler systems. At the ignition source
level, self-extinguishing cigarettes and re-safe candles can help prevent
res from starting (Rodgers et al., 2019; Shaw et al., 2010). At the
furniture level, safer ame retardants, other lling materials, naturally
re-resistant fabrics (such as wool), and re barriers between the fabric
and foam lling have been investigated as alternative strategies (Nazar´
e
and Davis, 2012; Shaw et al., 2010; US Consumer Product Safety Com-
mission, 2012). One carpet manufacturer reported to use aluminum
hydroxide (ATH) as a non-halogenated mineral ller ame retardant to
meet re codes. ATH is generally thought to be safer than chemical
ame retardants, although there are limited studies (National Research
Council, 2000).
A carpet manufacturer also described to us their approach to alter the
ionic charge and microscopic geometry of the nylon bers to achieve
long-term stain repellence without the need for PFAS. Water-repellant
alternatives on the market may include parafn, silicone, dendrimer,
and polyurethane chemistries (Danish Ministry of the Environment,
2015). Before implementing any potential alternative solution, espe-
cially a substitute chemical, health risks should be carefully studied and
balanced.
4.5. Strengths and limitations
A key, unique strength was the university’s holistic, chemical class-
based “healthier” building materials intervention to remove three clas-
ses of chemicals in a practical, scalable solution. Our thousands of
portable XRF measurements enabled us to assess the importance of
different product categories and to characterize sources of phosphorus
for the rst time.
There were a few study limitations. First, the “healthier” materials
interventions did not address all materials in a given room and were
often conducted in rooms next to conventional spaces. Our categoriza-
tion of rooms into no, partial, or full interventions captured some of
these nuances. Second, the sample size was relatively small, although we
sampled as many “healthier” rooms as available and possible. This
limited our statistical power and prevented us from developing more
complex models. We also could not feasibly measure all chemicals in
each class, of which some could be unknown substitutes although
manufacturers specied the “healthier” materials as free of the entire
classes. Another limitation is that some OPEs are used as plasticizers, so
we could not disentangle these separate OPE functions in buildings.
However, we did observe signicant reductions in OPEs in dust due to
the intervention that focused mainly on their elimination as FRs. In
addition, the natural heterogeneity of dust could have limited our sta-
tistical power to detect certain effects (Rauert and Harrad, 2015). For
XRF product scanning, phosphorus is not a perfect indicator of OPEs in a
particular product, however, phosphorus loadings were useful for
evaluating differences in furniture by ammability standard and
assessing overall product sources in statistical models to reduce expo-
sure misclassication that arises from using raw counts of products in
rooms. Bromine and phosphorus have an advantage of reecting even
unknown chemicals in a class, but this may have limited statistical
power to relate elements in products to the individual chemical analytes
in dust. Other sources of phosphorus besides OPEs include soil,
A.S. Young et al.
Environment International xxx (xxxx) xxx
10
pesticides, fertilizers, nerve agents, pharmaceuticals, industrial solvents,
and fuel additives, but we do not expect these chemicals to differ in
furniture products or by intervention status in dust (Liu et al., 2020;
Naughton and Terry, 2018). Similarly, other sources of bromine besides
PBDEs could include other brominated FR (which should also be
reduced in the intervention), biocides, fuel additives, pharmaceuticals,
polymers, halons, rubber, and dyes (Ioffe and Frim, 2011; Saikia et al.,
2016), but these should not signicantly interfere with our results, and
bromine has been shown to highly correlate with FRs for indoor sources
(Allen et al., 2008a).
5. Conclusions
This study demonstrates the effectiveness of a chemical class-based
“healthier” furniture and carpet intervention on reducing levels of
PFAS and ame retardants in dust indoors. The scanning of products in
situ for element surrogates using XRF proved helpful to identify elec-
tronics and exposed insulation as additional important sources of FRs in
buildings. Future interventions should target more product categories
and consider strategies to limit the use of, or reduce occupant exposures
to, legacy materials of the building. Overall, we observed signicant
indoor chemical reduction benets from the use of “healthier” materials
in buildings.
CRediT authorship contribution statement
Anna S. Young: Conceptualization, Methodology, Investigation,
Software, Formal analysis, Writing - original draft, Writing - review &
editing, Visualization, Project administration, Funding acquisition. Russ
Hauser: Conceptualization, Methodology, Writing - review & editing.
Tamarra James-Todd: Conceptualization, Methodology, Writing - re-
view & editing. Brent A. Coull: Conceptualization, Methodology,
Writing - review & editing. Hongkai Zhu: Methodology, Investigation,
Writing - review & editing. Kurunthachalam Kannan: Methodology,
Investigation, Writing - review & editing. Aaron Specht: Methodology,
Investigation, Writing - review & editing. Maya S. Bliss: Methodology,
Investigation. Joseph G. Allen: Conceptualization, Methodology,
Formal analysis, Writing - review & editing, Funding acquisition,
Supervision.
Declaration of Competing Interest
The authors declare that they have no known competing nancial
interests or personal relationships that could have appeared to inuence
the work reported in this paper.
Acknowledgments
We would like to thank Heather Henriksen, John Ullman, and
Brandon Geller from the Harvard Ofce for Sustainability for their
support. We are grateful to Jose Vallarino, Emily Jones, Marianne
Lahaie Luna, Erika Eitland, Maya Bliss, Parichehr Salimifard, Ben Jones,
and Sydney Robinson for their help with eld sampling. This research
was made possible by the Harvard Campus Sustainability Innovation
Fund, NIH Grant P30ES000002, NIOSH Grant T42 OH008416, and
NIOSH Grant 1K01 OH011648 (United States).
Appendix A. Supplementary material
Supplementary data to this article can be found online at https://doi.
org/10.1016/j.envint.2020.106151.
References
Abbasi, G., Buser, A.M., Soehl, A., Murray, M.W., Diamond, M.L., 2015. Stocks and Flows
of PBDEs in Products from Use to Waste in the U.S. and Canada from 1970 to 2020.
Environ. Sci. Technol. 49, 1521–1528. https://doi.org/10.1021/es504007v.
Abbasi, G., Saini, A., Goosey, E., Diamond, M.L., 2016. Product screening for sources of
halogenated ame retardants in Canadian house and ofce dust. Sci. Total Environ.
545–546, 299–307. https://doi.org/10.1016/j.scitotenv.2015.12.028.
Allen, J.G., Gale, S., Zoeller, R.T., Spengler, J.D., Birnbaum, L., McNeely, E., 2016. PBDE
ame retardants, thyroid disease, and menopausal status in U.S. women. Environ.
Health 15, 60. https://doi.org/10.1186/s12940-016-0141-0.
Allen, J.G., McClean, M.D., Stapleton, H.M., Webster, T.F., 2008a. Linking PBDEs in
House Dust to Consumer Products using X-ray Fluorescence. Environ. Sci. Technol.
42, 4222–4228. https://doi.org/10.1021/es702964a.
Allen, J.G., McClean, M.D., Stapleton, H.M., Webster, T.F., 2008b. Critical factors in
assessing exposure to PBDEs via house dust. Environ. Int. 34, 1085–1091. https://
doi.org/10.1016/j.envint.2008.03.006.
Barry, V., Winquist, A., Steenland, K., 2013. Peruorooctanoic acid (PFOA) exposures
and incident cancers among adults living near a chemical plant. Environ. Health
Perspect. 121, 1313–1318. https://doi.org/10.1289/ehp.1306615.
Batterman, S., Godwin, C., Chernyak, S., Jia, C., Charles, S., 2010. Brominated ame
retardants in ofces in Michigan. USA. Environ. Int. 36, 548–556. https://doi.org/
10.1016/j.envint.2010.04.008.
Bello, A., Carignan, C.C., Xue, Y., Stapleton, H.M., Bello, D., 2018. Exposure to
organophosphate ame retardants in spray polyurethane foam applicators: Role of
dermal exposure. Environ. Int. 113, 55–65. https://doi.org/10.1016/j.
envint.2018.01.020.
Bennett, D.H., Moran, R.E., Wu, X.M., Tulve, N.S., Clifton, M.S., Colon, M., Weathers, W.,
Sjodin, A., Jones, R., Hertz-Picciotto, I., 2015. Polybrominated diphenyl ether
(PBDE) concentrations and resulting exposure in homes in California: relationships
among passive air, surface wipe and dust concentrations, and temporal variability.
Indoor Air 25, 220–229. https://doi.org/10.1111/ina.12130.
Bi, C., Maestre, J.P., Li, H., Zhang, G., Givehchi, R., Mahdavi, A., Kinney, K.A., Siegel, J.,
Horner, S.D., Xu, Y., 2018. Phthalates and organophosphates in settled dust and
HVAC lter dust of U.S. low-income homes: Association with season, building
characteristics, and childhood asthma. Environ. Int. 121, 916–930. https://doi.org/
10.1016/j.envint.2018.09.013.
Boas, M., Feldt-Rasmussen, U., Main, K.M., 2012. Thyroid effects of endocrine disrupting
chemicals. Mol. Cell. Endocrinol. 355, 240–248. https://doi.org/10.1016/j.
mce.2011.09.005.
Calafat, A.M., Wong, L.Y., Kuklenyik, Z., Reidy, J.A., Needham, L.L., 2007.
Polyuoroalkyl chemicals in the U.S. population: Data from the national health and
nutrition examination survey (NHANES) 2003–2004 and comparisons with NHANES
1999–2000. Environ. Health Perspect. 115, 1596–1602. https://doi.org/10.1289/
ehp.10598.
Carignan, C.C., Mínguez-Alarc´
on, L., Butt, C.M., Williams, P.L., Meeker, J.D.,
Stapleton, H.M., Toth, T.L., Ford, J.B., Hauser, R., 2017. Urinary concentrations of
organophosphate ame retardant metabolites and pregnancy outcomes among
women undergoing in vitro fertilization for the EARTH study team. Environ. Health
Perspect. 125, 8. https://doi.org/10.1289/EHP1021.
Carignan, C.C., Minguez-Alarcon, L., Williams, P.L., Meeker, J.D., Stapleton, H.M.,
Butt, C.M., Toth, T.L., Ford, J.B., Hauser, R., 2018. Paternal urinary concentrations
of organophosphate ame retardant metabolites, fertility measures, and pregnancy
outcomes among couples undergoing in vitro fertilization. Environ. Int. 111,
232–238. https://doi.org/10.1016/j.envint.2017.12.005.
Castorina, R., Butt, C., Stapleton, H.M., Avery, D., Harley, K.G., Holland, N., Eskenazi, B.,
Bradman, A., 2017. Flame retardants and their metabolites in the homes and urine of
pregnant women residing in California (the CHAMACOS cohort). Chemosphere 179,
159–166. https://doi.org/10.1016/j.chemosphere.2017.03.076.
Cequier, E., Sakhi, A.K., Marce, R.M., Becher, G., Thomsen, C., 2015. Human exposure
pathways to organophosphate triesters - a biomonitoring study of mother-child pairs.
Environ. Int. 75, 159–165. https://doi.org/10.1016/j.envint.2014.11.009.
Choi, G., Wang, Y.-B., Sundaram, R., Chen, Z., Barr, D.B., Buck Louis, G.M., Smarr, M.M.,
2019. Polybrominated diphenyl ethers and incident pregnancy loss: The LIFE Study.
Environ. Res. 168, 375–381. https://doi.org/10.1016/j.envres.2018.09.018.
Cooper, E.M., Kroeger, G., Davis, K., Clark, C.R., Ferguson, P.L., Stapleton, H.M., 2016.
Results from Screening Polyurethane Foam Based Consumer Products for Flame
Retardant Chemicals: Assessing Impacts on the Change in the Furniture
Flammability Standards. Environ. Sci. Technol. 50, 10653–10660. https://doi.org/
10.1021/acs.est.6b01602.
Czerska, M., Zielinski, M., Kaminska, J., Ligocka, D., 2013. Effects of polybrominated
diphenyl ethers on thyroid hormone, neurodevelopment and fertility in rodents and
humans. Int. J. Occup. Med. Environ. Health 26, 498–510. https://doi.org/10.2478/
s13382-013-0138-7.
Danish Ministry of the Environment, 2015. Alternatives to peruoroalkyl and
polyuoroalkyl substances (PFAS) in textiles.
Dembsey, N.A., Brokaw, F.M., Stapleton, H.M., Dodson, R.E., Onasch, J., Jazan, E.,
Carignan, C.C., 2019. Intervention to reduce gymnast exposure to ame retardants
from pit foam: A case study. Environ. Int. 127, 868–875. https://doi.org/10.1016/j.
envint.2019.01.084.
Dodson, R.E., Perovich, L.J., Covaci, A., Van den Eede, N., Ionas, A.C., Dirtu, A.C.,
Brody, J.G., Rudel, R.A., 2012. After the PBDE Phase-Out: A Broad Suite of Flame
Retardants in Repeat House Dust Samples from California. Environ. Sci. Technol. 46,
13056–13066. https://doi.org/10.1021/es303879n.
Dodson, R.E., Rodgers, K.M., Carey, G., Cedeno Laurent, J.G., Covaci, A., Poma, G.,
Malarvannan, G., Spengler, J.D., Rudel, R.A., Allen, J.G., 2017. Flame Retardant
A.S. Young et al.
Environment International xxx (xxxx) xxx
11
Chemicals in College Dormitories: Flammability Standards Inuence Dust
Concentrations. Environ. Sci. Technol. 51, 4860–4869. https://doi.org/10.1021/acs.
est.7b00429.
Doherty, B.T., Hoffman, K., Keil, A.P., Engel, S.M., Stapleton, H.M., Goldman, B.D.,
Olshan, A.F., Daniels, J.L., 2019a. Prenatal exposure to organophosphate esters and
cognitive development in young children in the Pregnancy, Infection, and Nutrition
Study. Environ. Res. 169, 33–40. https://doi.org/10.1016/j.envres.2018.10.033.
Doherty, B.T., Hoffman, K., Keil, A.P., Engel, S.M., Stapleton, H.M., Goldman, B.D.,
Olshan, A.F., Daniels, J.L., 2019b. Prenatal exposure to organophosphate esters and
behavioral development in young children in the Pregnancy, Infection, and Nutrition
Study. Neurotoxicology 73, 150–160. https://doi.org/10.1016/j.
neuro.2019.03.007.
Duan, H., Yu, D., Zuo, J., Yang, B., Zhang, Y., Niu, Y., 2016. Characterization of
brominated ame retardants in construction and demolition waste components:
HBCD and PBDEs. Sci. Total Environ. 572, 77–85. https://doi.org/10.1016/j.
scitotenv.2016.07.165.
Fraser, A.J., Webster, T.F., Watkins, D.J., Strynar, M.J., Kato, K., Calafat, A.M., Vieira, V.
M., McClean, M.D., 2013. Polyuorinated compounds in dust from homes, ofces,
and vehicles as predictors of concentrations in ofce workers’ serum. Environ. Int.
60, 128–136. https://doi.org/10.1016/j.envint.2013.08.012.
Frederiksen, M., Thomsen, C., Froshaug, M., Vorkamp, K., Thomsen, M., Becher, G.,
Knudsen, L.E., 2010. Polybrominated diphenyl ethers in paired samples of maternal
and umbilical cord blood plasma and associations with house dust in a Danish
cohort. Int. J. Hyg. Environ. Health 213, 233–242. https://doi.org/10.1016/j.
ijheh.2010.04.008.
Gallen, C., Banks, A., Brandsma, S., Baduel, C., Thai, P., Eaglesham, G., Heffernan, A.,
Leonards, P., Bainton, P., Mueller, J.F., 2014. Towards development of a rapid and
effective non-destructive testing strategy to identify brominated ame retardants in
the plastics of consumer products. Sci. Total Environ. 491–492, 255–265. https://
doi.org/10.1016/j.scitotenv.2014.01.074.
Goosey, E., Harrad, S., 2011. Peruoroalkyl compounds in dust from Asian, Australian,
European, and North American homes and UK cars, classrooms, and ofces. Environ.
Int. 37, 86–92. https://doi.org/10.1016/j.envint.2010.08.001.
Hammel, S.C., Hoffman, K., Lorenzo, A.M., Chen, A., Phillips, A.L., Butt, C.M., Sosa, J.A.,
Webster, T.F., Stapleton, H.M., 2017. Associations between ame retardant
applications in furniture foam, house dust levels, and residents’ serum levels.
Environ. Int. 107, 181–189. https://doi.org/10.1016/j.envint.2017.07.015.
Harrad, S., Ibarra, C., Diamond, M., Melymuk, L., Robson, M., Douwes, J., Roosens, L.,
Dirtu, A.C., Covaci, A., 2008. Polybrominated diphenyl ethers in domestic indoor
dust from Canada, New Zealand, United Kingdom and United States. Environ. Int.
34, 232–238. https://doi.org/10.1016/j.envint.2007.08.008.
Herzke, D., Olsson, E., Posner, S., 2012. Peruoroalkyl and polyuoroalkyl substances
(PFASs) in consumer products in Norway - a pilot study. Chemosphere 88, 980–987.
https://doi.org/10.1016/j.chemosphere.2012.03.035.
Hoffman, K., Garantziotis, S., Birnbaum, L.S., Stapleton, H.M., 2015. Monitoring indoor
exposure to organophosphate ame retardants: hand wipes and house dust. Environ.
Health Perspect. 123, 160–165. https://doi.org/10.1289/ehp.1408669.
Hornung, R.W., Reed, L.D., 1990. Estimation of Average Concentration in the Presence of
Nondetectable Values. Appl. Occup. Environ. Hyg. 5, 46–51. https://doi.org/
10.1080/1047322X.1990.10389587.
Imm, P., Knobeloch, L., Buelow, C., Anderson, H.A., 2009. Household exposures to
polybrominated diphenyl ethers (PBDEs) in a Wisconsin Cohort. Environ. Health
Perspect. 117, 1890–1895. https://doi.org/10.1289/ehp.0900839.
Ioffe, D., Frim, R., 2011. Bromine, Organic Compounds. Kirk-Othmer Encycl. Chem.
Technol., Major Reference Works. https://doi.org/10.1002/0471238961.021815
1325150606.a01.pub2.
Janousek, R.M., Lebertz, S., Knepper, T.P., 2019. Previously unidentied sources of
peruoroalkyl and polyuoroalkyl substances from building materials and industrial
fabrics. Environ. Sci. Process. Impacts. https://doi.org/10.1039/c9em00091g.
Jinhui, L., Yuan, C., Wenjing, X., 2017. Polybrominated diphenyl ethers in articles: a
review of its applications and legislation. Environ. Sci. Pollut. Res. Int. 24,
4312–4321. https://doi.org/10.1007/s11356-015-4515-6.
Johnson-Restrepo, B., Kannan, K., 2009. An assessment of sources and pathways of
human exposure to polybrominated diphenyl ethers in the United States.
Chemosphere 76, 542–548. https://doi.org/10.1016/j.chemosphere.2009.02.068.
Johnson, P.I., Stapleton, H.M., Mukherjee, B., Hauser, R., Meeker, J.D., 2013.
Associations between brominated ame retardants in house dust and hormone levels
in men. Sci. Total Environ. 445–446, 177–184. https://doi.org/10.1016/j.
scitotenv.2012.12.017.
Kajiwara, N., Noma, Y., Takigami, H., 2011. Brominated and organophosphate ame
retardants in selected consumer products on the Japanese market in 2008. J. Hazard.
Mater. 192, 1250–1259. https://doi.org/10.1016/j.jhazmat.2011.06.043.
Karaskova, P., Venier, M., Melymuk, L., Becanova, J., Vojta, S., Prokes, R., Diamond, M.
L., Klanova, J., 2016. Peruorinated alkyl substances (PFASs) in household dust in
Central Europe and North America. Environ. Int. 94, 315–324. https://doi.org/
10.1016/j.envint.2016.05.031.
Kemmlein, S., Hahn, O., Jann, O., 2003. Emissions of organophosphate and brominated
ame retardants from selected consumer products and building materials. Atmos.
Environ. 37, 5485–5493. https://doi.org/10.1016/j.atmosenv.2003.09.025.
Kim, S.-K., Kannan, K., 2007. Peruorinated acids in air, rain, snow, surface runoff, and
lakes: relative importance of pathways to contamination of urban lakes. Environ. Sci.
Technol. 41, 8328–8334. https://doi.org/10.1021/es072107t.
Kim, U.-J., Wang, Y., Li, W., Kannan, K., 2019. Occurrence of and human exposure to
organophosphate ame retardants/plasticizers in indoor air and dust from various
microenvironments in the United States. Environ. Int. 125, 342–349. https://doi.
org/10.1016/j.envint.2019.01.065.
Kim, U.J., Oh, J.K., Kannan, K., 2017. Occurrence, removal, and environmental emission
of organophosphate ame retardants/plasticizers in a wastewater treatment plant in
New York State. Environ. Sci. Technol. 51, 7872–7880. https://doi.org/10.1021/acs.
est.7b02035.
Knobeloch, L., Imm, P., Anderson, H., 2012. Peruoroalkyl chemicals in vacuum cleaner
dust from 39 Wisconsin homes. Chemosphere 88, 779–783. https://doi.org/
10.1016/j.chemosphere.2012.03.082.
Koponen, J., Winkens, K., Airaksinen, R., Berger, U., Vestergren, R., Cousins, I.T.,
Karvonen, A.M., Pekkanen, J., Kiviranta, H., 2018. Longitudinal trends of per- and
polyuoroalkyl substances in children’s serum. Environ. Int. 121, 591–599. https://
doi.org/10.1016/j.envint.2018.09.006.
Kotthoff, M., Muller, J., Jurling, H., Schlummer, M., Fiedler, D., 2015. Peruoroalkyl and
polyuoroalkyl substances in consumer products. Environ. Sci. Pollut. Res. Int. 22,
14546–14559. https://doi.org/10.1007/s11356-015-4202-7.
Li, W., Wang, Y., Asimakopoulos, A.G., Covaci, A., Gevao, B., Johnson-Restrepo, B.,
Kumosani, T.A., Malarvannan, G., Moon, H.-B., Nakata, H., Sinha, R.K., Tran, T.M.,
Kannan, K., 2019a. Organophosphate esters in indoor dust from 12 countries:
Concentrations, composition proles, and human exposure. Environ. Int. 133,
105178 https://doi.org/10.1016/j.envint.2019.105178.
Li, Y., Chang, Q., Duan, H., Liu, Y., Zhang, J., Li, J., 2019b. Occurrence, levels and
proles of brominated ame retardants in daily-use consumer products on the
Chinese market. Environ. Sci. Process. Impacts 21, 446–455. https://doi.org/
10.1039/c8em00483h.
Liew, Z., Goudarzi, H., Oulhote, Y., 2018. Developmental Exposures to Peruoroalkyl
Substances (PFASs): An Update of Associated Health Outcomes. Curr. Environ. Heal.
reports 5, 1–19. https://doi.org/10.1007/s40572-018-0173-4.
Lin, P.-I.-D., Cardenas, A., Hauser, R., Gold, D.R., Kleinman, K.P., Hivert, M.-F.,
Fleisch, A.F., Calafat, A.M., Webster, T.F., Horton, E.S., Oken, E., 2019. Per- and
polyuoroalkyl substances and blood lipid levels in pre-diabetic adults-longitudinal
analysis of the diabetes prevention program outcomes study. Environ. Int. 129,
343–353. https://doi.org/10.1016/j.envint.2019.05.027.
Linares, V., Belles, M., Domingo, J.L., 2015. Human exposure to PBDE and critical
evaluation of health hazards. Arch. Toxicol. 89, 335–356. https://doi.org/10.1007/
s00204-015-1457-1.
Liu, D., Zheng, C., Qiu, Q., Tang, J., Xu, Y., 2020. Global pattern of studies on phosphorus
at watershed scale. Environ. Sci. Pollut. Res. https://doi.org/10.1007/s11356-020-
07771-y.
Liu, X., Guo, Z., Krebs, K.A., Pope, R.H., Roache, N.F., 2014. Concentrations and trends of
peruorinated chemicals in potential indoor sources from 2007 through 2011 in the
US. Chemosphere 98, 51–57. https://doi.org/10.1016/j.chemosphere.2013.10.001.
Lucattini, L., Poma, G., Covaci, A., de Boer, J., Lamoree, M.H., Leonards, P.E.G., 2018.
A review of semi-volatile organic compounds (SVOCs) in the indoor environment:
occurrence in consumer products, indoor air and dust. Chemosphere 201, 466–482.
https://doi.org/10.1016/j.chemosphere.2018.02.161.
Meeker, J.D., Stapleton, H.M., 2010. House dust concentrations of organophosphate
ame retardants in relation to hormone levels and semen quality parameters.
Environ. Health Perspect. 118, 318–323. https://doi.org/10.1289/ehp.0901332.
Messerlian, C., Williams, P.L., Minguez-Alarcon, L., Carignan, C.C., Ford, J.B., Butt, C.M.,
Meeker, J.D., Stapleton, H.M., Souter, I., Hauser, R., 2018. Organophosphate ame-
retardant metabolite concentrations and pregnancy loss among women conceiving
with assisted reproductive technology. Fertil. Steril. 110, 1137–1144.e1. https://doi.
org/10.1016/j.fertnstert.2018.06.045.
Mitro, S.D., Dodson, R.E., Singla, V., Adamkiewicz, G., Elmi, A.F., Tilly, M.K., Zota, A.R.,
2016. Consumer Product Chemicals in Indoor Dust: A Quantitative Meta-analysis of
U.S. Studies. Environ. Sci. Technol. 50, 10661–10672. https://doi.org/10.1021/acs.
est.6b02023.
Mumford, S.L., Kim, S., Chen, Z., Gore-Langton, R.E., Boyd Barr, D., Buck Louis, G.M.,
2015. Persistent organic pollutants and semen quality: The LIFE Study. Chemosphere
135, 427–435. https://doi.org/10.1016/j.chemosphere.2014.11.015.
National Research Council, 2000. Toxicological risks of selected ame-retardant
chemicals. National Academies Press.
Naughton, S.X., Terry, A.V.J., 2018. Neurotoxicity in acute and repeated
organophosphate exposure. Toxicology 408, 101–112. https://doi.org/10.1016/j.
tox.2018.08.011.
Nazar´
e, S., Davis, R.D., 2012. A review of re blocking technologies for soft furnishings.
Fire Sci. Rev. 1, 1. https://doi.org/10.1186/2193-0414-1-1.
OECD, 2018. Toward a New Comprehensive Global Database of Per- and Polyuoroalkyl
Substances (PFASs).
Ospina, M., Jayatilaka, N.K., Wong, L.-Y., Restrepo, P., Calafat, A.M., 2018. Exposure to
organophosphate ame retardant chemicals in the U.S. general population: Data
from the 2013–2014 National Health and Nutrition Examination Survey. Environ.
Int. 110, 32–41. https://doi.org/10.1016/j.envint.2017.10.001.
Percy, Z., La Guardia, M.J., Xu, Y., Hale, R.C., Dietrich, K.N., Lanphear, B.P., Yolton, K.,
Vuong, A.M., Cecil, K.M., Braun, J.M., Xie, C., Chen, A., 2020. Concentrations and
loadings of organophosphate and replacement brominated ame retardants in house
dust from the home study during the PBDE phase-out. Chemosphere 239, 124701.
https://doi.org/10.1016/j.chemosphere.2019.124701.
Petreas, M., Gill, R., Takaku-Pugh, S., Lytle, E., Parry, E., Wang, M., Quinn, J., Park, J.-S.,
2016. Rapid methodology to screen ame retardants in upholstered furniture for
compliance with new California labeling law (SB 1019). Chemosphere 152,
353–359. https://doi.org/10.1016/j.chemosphere.2016.02.102.
Poothong, S., Papadopoulou, E., Padilla-S´
anchez, J.A., Thomsen, C., Haug, L.S., 2020.
Multiple pathways of human exposure to poly- and peruoroalkyl substances
(PFASs): From external exposure to human blood. Environ. Int. 134, 105244 https://
doi.org/10.1016/j.envint.2019.105244.
A.S. Young et al.
Environment International xxx (xxxx) xxx
12
Preston, E.V., McClean, M.D., Claus Henn, B., Stapleton, H.M., Braverman, L.E.,
Pearce, E.N., Makey, C.M., Webster, T.F., 2017. Associations between urinary
diphenyl phosphate and thyroid function. Environ. Int. 101, 158–164. https://doi.
org/10.1016/j.envint.2017.01.020.
Rappazzo, K.M., Coffman, E., Hines, E.P., 2017. Exposure to Peruorinated Alkyl
Substances and Health Outcomes in Children: A Systematic Review of the
Epidemiologic Literature. Int. J. Environ. Res. Public Health 14. https://doi.org/
10.3390/ijerph14070691.
Rauert, C., Harrad, S., 2015. Mass transfer of PBDEs from plastic TV casing to indoor dust
via three migration pathways–A test chamber investigation. Sci. Total Environ. 536,
568–574. https://doi.org/10.1016/j.scitotenv.2015.07.050.
Rauert, C., Lazarov, B., Harrad, S., Covaci, A., Stranger, M., 2014. A review of chamber
experiments for determining specic emission rates and investigating migration
pathways of ame retardants. Atmos. Environ. 82, 44–55. https://doi.org/10.1016/
j.atmosenv.2013.10.003.
Rodgers, K.M., Swetschinski, L.R., Dodson, R.E., Alpert, H.R., Fleming, J.M., Rudel, R.A.,
2019. Health Toll From Open Flame and Cigarette-Started Fires on Flame-Retardant
Furniture in Massachusetts, 2003–2016. Am. J. Public Health 109, 1205–1211.
https://doi.org/10.2105/AJPH.2019.305157.
Saikia, I., Borah, A.J., Phukan, P., 2016. Use of Bromine and Bromo-Organic Compounds
in Organic Synthesis. Chem. Rev. 116, 6837–7042. https://doi.org/10.1021/acs.
chemrev.5b00400.
Scher, D.P., Kelly, J.E., Huset, C.A., Barry, K.M., Yingling, V.L., 2019. Does soil track-in
contribute to house dust concentrations of peruoroalkyl acids (PFAAs) in areas
affected by soil or water contamination? J. Expo. Sci. Environ. Epidemiol. 29,
218–226. https://doi.org/10.1038/s41370-018-0101-6.
Schreder, E.D., La Guardia, M.J., 2014. Flame retardant transfers from U.S. households
(dust and laundry wastewater) to the aquatic environment. Environ. Sci. Technol.
48, 11575–11583. https://doi.org/10.1021/es502227h.
Shaw, S.D., Blum, A., Weber, R., Kannan, K., Rich, D., Lucas, D., Koshland, C.P.,
Dobraca, D., Hanson, S., Birnbaum, L.S., 2010. Halogenated ame retardants: do the
re safety benets justify the risks? Rev. Environ. Health 25, 261–305.
Sj¨
odin, A., Wong, L.-Y., Jones, R.S., Park, A., Zhang, Y., Hodge, C., DiPietro, E.,
McClure, C., Turner, W., Needham, L.L., Patterson, D.G., 2008. Serum
Concentrations of Polybrominated Diphenyl Ethers (PBDEs) and Polybrominated
Biphenyl (PBB) in the United States Population: 2003–2004. Environ. Sci. Technol.
42, 1377–1384. https://doi.org/10.1021/es702451p.
Stanifer, J.W., Stapleton, H.M., Souma, T., Wittmer, A., Zhao, X., Boulware, L.E., 2018.
Peruorinated Chemicals as Emerging Environmental Threats to Kidney Health: A
Scoping Review. Clin. J. Am. Soc. Nephrol. 13, 1479–1492. https://doi.org/
10.2215/CJN.04670418.
Stapleton, H.M., Dodder, N.G., Offenberg, J.H., Schantz, M.M., Wise, S.A., 2005.
Polybrominated diphenyl ethers in house dust and clothes dryer lint. Environ. Sci.
Technol. 39, 925–931. https://doi.org/10.1021/es0486824.
Stapleton, H.M., Klosterhaus, S., Eagle, S., Fuh, J., Meeker, J.D., Blum, A., Webster, T.F.,
2009. Detection of Organophosphate Flame Retardants in Furniture Foam and U.S.
House Dust. Environ. Sci. Technol. 43, 7490–7495. https://doi.org/10.1021/
es9014019.
Stapleton, H.M., Klosterhaus, S., Keller, A., Ferguson, P.L., van Bergen, S., Cooper, E.,
Webster, T.F., Blum, A., 2011. Identication of ame retardants in polyurethane
foam collected from baby products. Environ. Sci. Technol. 45, 5323–5331. https://
doi.org/10.1021/es2007462.
Stapleton, H.M., Misenheimer, J., Hoffman, K., Webster, T.F., 2014. Flame retardant
associations between children’s handwipes and house dust. Chemosphere 116,
54–60. https://doi.org/10.1016/j.chemosphere.2013.12.100.
Stapleton, H.M., Sharma, S., Getzinger, G., Ferguson, P.L., Gabriel, M., Webster, T.F.,
Blum, A., 2012. Novel and High Volume Use Flame Retardants in US Couches
Reective of the 2005 PentaBDE Phase Out. Environ. Sci. Technol. 46,
13432–13439. https://doi.org/10.1021/es303471d.
State of California, 2019. In re: Bureau of Electronic and Appliance Repair. Amended
notice of approval of regulatory action, Home Furnishings and Thermal Insulation.
State of California, 2013. Technical Bulletin 117-2013: Requirements, Test Procedure
and Apparatus for Testing the Smolder Resistance of Materials Used in Upholstered
Furniture. Department of Consumer Affairs, Sacramento, CA.
Strynar, M.J., Lindstrom, A.B., 2008. Peruorinated compounds in house dust from Ohio
and North Carolina. USA. Environ. Sci. Technol. 42, 3751–3756. https://doi.org/
10.1021/es7032058.
Stubbings, W.A., Schreder, E.D., Thomas, M.B., Romanak, K., Venier, M., Salamova, A.,
2018. Exposure to brominated and organophosphate ester ame retardants in U.S.
childcare environments: Effect of removal of ame-retarded nap mats on indoor
levels. Environ. Pollut. 238, 1056–1068. https://doi.org/10.1016/j.
envpol.2018.03.083.
Sunderland, E.M., Hu, X.C., Dassuncao, C., Tokranov, A.K., Wagner, C.C., Allen, J.G.,
2018. A review of the pathways of human exposure to poly- and peruoroalkyl
substances (PFASs) and present understanding of health effects. J. Expo. Sci.
Environ. Epidemiol. https://doi.org/10.1038/s41370-018-0094-1.
Tokranov, A.K., Nishizawa, N., Amadei, C.A., Zenobio, J.E., Pickard, H.M., Allen, J.G.,
Vecitis, C.D., Sunderland, E.M., 2018. How Do We Measure Poly- and Peruoroalkyl
Substances (PFASs) at the Surface of Consumer Products? Environ. Sci. Technol. Lett.
6 https://doi.org/10.1021/acs.estlett.8b00600.
Turner, A., Filella, M., 2017. Bromine in plastic consumer products - Evidence for the
widespread recycling of electronic waste. Sci. Total Environ. 601–602, 374–379.
https://doi.org/10.1016/j.scitotenv.2017.05.173.
US Consumer Product Safety Commission, 2012. Upholstered furniture full scale chair
tests – open ame ignition resluts and analysis. Bethesda, MD.
van der Veen, I., de Boer, J., 2012. Phosphorus ame retardants: Properties, production,
environmental occurrence, toxicity and analysis. Chemosphere 88, 1119–1153.
https://doi.org/10.1016/j.chemosphere.2012.03.067.
Vojta, ˇ
S., Beˇ
canov´
a, J., Melymuk, L., Komprdov´
a, K., Kohoutek, J., Kukuˇ
cka, P.,
Kl´
anov´
a, J., 2017. Screening for halogenated ame retardants in European consumer
products, building materials and wastes. Chemosphere 168, 457–466. https://doi.
org/10.1016/j.chemosphere.2016.11.032.
Vuong, A.M., Yolton, K., Dietrich, K.N., Braun, J.M., Lanphear, B.P., Chen, A., 2018.
Exposure to polybrominated diphenyl ethers (PBDEs) and child behavior: Current
ndings and future directions. Horm. Behav. 101, 94–104. https://doi.org/10.1016/
j.yhbeh.2017.11.008.
Vykoukalov´
a, M., Venier, M., Vojta, ˇ
S., Melymuk, L., Beˇ
canov´
a, J., Romanak, K.,
Prokeˇ
s, R., Okeme, J.O., Saini, A., Diamond, M.L., Kl´
anov´
a, J., 2017.
Organophosphate esters ame retardants in the indoor environment. Environ. Int.
106, 97–104. https://doi.org/10.1016/j.envint.2017.05.020.
Wang, S., Romanak, K.A., Hendryx, M., Salamova, A., Venier, M., 2019. Association
between Thyroid Function and Exposures to Brominated and Organophosphate
Flame Retardants in Rural Central Appalachia. Environ. Sci. Technol. https://doi.
org/10.1021/acs.est.9b04892.
Wang, Y., Hou, M., Zhang, Q., Wu, X., Zhao, H., Xie, Q., Chen, J., 2017a.
Organophosphorus Flame Retardants and Plasticizers in Building and Decoration
Materials and Their Potential Burdens in Newly Decorated Houses in China. Environ.
Sci. Technol. 51, 10991–10999. https://doi.org/10.1021/acs.est.7b03367.
Wang, Z., Cousins, I.T., Scheringer, M., Hungerbuehler, K., 2015. Hazard assessment of
uorinated alternatives to long-chain peruoroalkyl acids (PFAAs) and their
precursors: Status quo, ongoing challenges and possible solutions. Environ. Int. 75,
172–179. https://doi.org/10.1016/j.envint.2014.11.013.
Wang, Z., Cousins, I.T., Scheringer, M., Hungerbühler, K., 2013. Fluorinated alternatives
to long-chain peruoroalkyl carboxylic acids (PFCAs), peruoroalkane sulfonic acids
(PFSAs) and their potential precursors. Environ. Int. 60, 242–248. https://doi.org/
10.1016/j.envint.2013.08.021.
Wang, Z., Dewitt, J.C., Higgins, C.P., Cousins, I.T., 2017b. A Never-Ending Story of Per-
and Polyuoroalkyl Substances (PFASs)? Environ. Sci. Technol. 51, 2508–2518.
https://doi.org/10.1021/acs.est.6b04806.
Watkins, D.J., McClean, M.D., Fraser, A.J., Weinberg, J., Stapleton, H.M., Sjodin, A.,
Webster, T.F., 2011. Exposure to PBDEs in the ofce environment: evaluating the
relationships between dust, handwipes, and serum. Environ. Health Perspect. 119,
1247–1252. https://doi.org/10.1289/ehp.1003271.
Watkins, D.J., McClean, M.D., Fraser, A.J., Weinberg, J., Stapleton, H.M., Webster, T.F.,
2013. Associations between PBDEs in ofce air, dust, and surface wipes. Environ.
Int. 59, 124–132. https://doi.org/10.1016/j.envint.2013.06.001.
Webster, T.F., Harrad, S., Millette, J.R., Holbrook, R.D., Davis, J.M., Stapleton, H.M.,
Allen, J.G., McClean, M.D., Ibarra, C., Abdallah, M.-A.-E., Covaci, A., 2009.
Identifying Transfer Mechanisms and Sources of Decabromodiphenyl Ether (BDE
209) in Indoor Environments Using Environmental Forensic Microscopy. Environ.
Sci. Technol. 43, 3067–3072. https://doi.org/10.1021/es803139w.
Whitehead, T.P., Brown, F.R., Metayer, C., Park, J.-S., Does, M., Petreas, M.X., Bufer, P.
A., Rappaport, S.M., 2013. Polybrominated diphenyl ethers in residential dust:
sources of variability. Environ. Int. 57–58, 11–24. https://doi.org/10.1016/j.
envint.2013.03.003.
Wu, X.M., Bennett, D.H., Calafat, A.M., Kato, K., Strynar, M., Andersen, E., Moran, R.E.,
Tancredi, D.J., Tulve, N.S., Hertz-Picciotto, I., 2015. Serum concentrations of
peruorinated compounds (PFC) among selected populations of children and adults
in California. Environ. Res. 136, 264–273. https://doi.org/10.1016/j.
envres.2014.09.026.
Xiao, C., Grandjean, P., Valvi, D., Nielsen, F., Jensen, T.K., Weihe, P., Oulhote, Y., 2019.
Associations of exposure to peruoroalkyl substances with thyroid hormone
concentrations and birth size. J. Clin. Endocrinol. Metab. https://doi.org/10.1210/
clinem/dgz147.
Zheng, X., Sun, R., Qiao, L., Guo, H., Zheng, J., Mai, B., 2017. Flame retardants on the
surface of phones and personal computers. Sci. Total Environ. 609, 541–545.
https://doi.org/10.1016/j.scitotenv.2017.07.202.
A.S. Young et al.