Environmental Pollution 270 (2021) 116219
Contents lists available at ScienceDirect
Environmental Pollution
journal homepage: www.elsevier.com/locate/envpol
Perfluoroalkyl substances in the urine and hair of preschool children,
airborne particles in kindergartens, and drinking water in Hong Kong*
Na Li a, b, Guang-Guo Ying c, d, Huachang Hong e, Wen-Jing Deng a, c, d, *
a
Department of Science and Environmental Studies, The Education University of Hong Kong, Tai Po, N.T., Hong Kong SAR, China
Key Laboratory of Drinking Water Science and Technology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing, 100085,
China
c
SCNU Environmental Research Institute, Guangdong Provincial Key Laboratory of Chemical Pollution and Environmental Safety & MOE Key Laboratory of
Theoretical Chemistry of Environment, South China Normal University, Guangzhou, 510006, China
d
School of Environment, South China Normal University, University Town, Guangzhou, 510006, China
e
College of Geography and Environmental Sciences, Zhejiang Normal University, Jinhua, 321004, China
b
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 16 May 2020
Received in revised form
20 November 2020
Accepted 30 November 2020
Available online 7 December 2020
Seven perfluorinated and polyfluorinated substances (PFASs), namely perfluorononanoic acid (PFNA),
perfluorodecanoic acid (PFDA), perfluorooctanoic acid (PFOA), perfluorooctane sulfonic acid (PFOS),
perfluoroheptanoic
acid
(PFHpA),
perfluorohexanoic
acid
(PFHxA),
and
perfluoro-1,10decanedicarboxylic acid (PFDDA), were evaluated in urine and hair samples from children (age: 4e6
years, N ¼ 53), airborne particles sampled at 17 kindergartens, and tap water and bottled water samples.
All samples were collected in Hong Kong. The analytical results suggested widespread PFAS contamination. All target PFASs were detected in at least 32% of urine samples, with geometric mean (GM)
concentrations ranging from 0.18 to 2.97 ng/L, and in 100% of drinking water samples at GM concentrations of 0.18e21.1 ng/L. Although PFOS and PFDDA were not detected in hair or air samples, the other
target PFASs were detected in 48e70% of hair samples (GM concentrations: 2.40e233 pg/g) and 100% of
air samples (GM concentrations: 14.8e536.7 pg/m3). In summary, the highest PFAS concentrations were
detected in airborne particles measured in kindergartens. PFOA was the major PFAS detected in hair,
urine, and drinking water samples, while PFOA, PFDA, and PFHpA were dominant in airborne particles.
Although a significant difference in PFAS concentrations in hair samples was observed between boys and
girls (p < .05), no significant sex-related difference in urinary PFAS or paired PFAS (hair/urine) concentrations was observed.
© 2020 Elsevier Ltd. All rights reserved.
Keywords:
PFASs
Urine
Drinking water
Children
Airborne particle
1. Introduction
Over the past few decades, concerns have been raised regarding
perfluorinated and polyfluorinated substances (PFASs), which are
highly stable carbon fluorine compounds that are processed into
protective coatings, lubricant components, or pharmaceuticals
(Ahrens and Bundschuh, 2014; Route et al., 2014; SznajderKatarzynska et al., 2019). PFASs are highly transportable, persistent, and bioaccumulative, and their widespread use has led to the
frequent detection of contamination in natural environments,
*
€rg Rinklebe.
This paper has been recommended for acceptance by Dr. Jo
* Corresponding author.Department of Science and Environmental Studies, The
Education University of Hong Kong, Tai Po, N.T., SAR, Hong Kong, China.
E-mail address: wdeng@eduhk.hk (W.-J. Deng).
https://doi.org/10.1016/j.envpol.2020.116219
0269-7491/© 2020 Elsevier Ltd. All rights reserved.
wildlife, indoor air, food, drinking water, and the general population (Bjerregaard-Olesen et al., 2016; Dalahmeh et al., 2018;
Domingo and Nadal, 2017; Fromme et al., 2015; Genuis et al., 2013;
Jian et al., 2018; Sharma et al., 2016; Sznajder-Katarzynska et al.,
2019). In many countries, existing bans on the production of perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid
(PFOA) have led to decreases of PFOS concentrations in humans,
rapid replacement with shorter-chain PFASs (Calafat et al., 2007;
Glynn et al., 2012), and the production and increasing environmental levels of numerous alternatives (Gomis et al., 2015; Wang
et al., 2014, 2019).
The adverse effects of PFASs and alternative PFASs have attracted growing attention following the increasing discovery of the
toxicity and high human exposure levels of these chemicals
(Poothong et al., 2020). Previous studies have summarized and
N. Li, G.-G. Ying, H. Hong et al.
Environmental Pollution 270 (2021) 116219
chemicals were purchased from Sinopharm Chemical Reagent Co.,
Ltd. (Shanghai, China).
compared the concentrations and contamination profiles of PFASs
in the biota and human tissues and explored the bioaccumulation
processes (Cui et al., 2018; Shi et al., 2016, 2017). Many in vitro or
in vivo toxicological studies have indicated that PFASs exert toxic
effects such as endocrine disruption, neurotoxicity, immunotoxicity, mitochondrial dysfunction, peroxisome proliferation, and
renal toxicity (Chen et al., 2018; Chou et al., 2017; Han et al., 2018;
Soloff et al., 2017; Tang et al., 2017). Epidemiology studies have
revealed a potential causal association of PFASs with asthma, thyroid diseases, chronic kidney disease, behavioral disorders, and
many other diseases (Anderson et al., 2019; Sunderland et al., 2019).
Increasing evidence from in vitro and in vivo studies suggests that
many alternative PFASs are also toxic (Zeng et al., 2019).
PFASs can affect developing organisms; consequently, exposure
could cause adverse effects in children (Sunderland et al., 2019).
Compared with adults, children have less mature metabolic systems and a larger body surface and make more frequent hand-tomouth contact and thus are more prone to experience the
adverse effects of PFASs (Landrigan and Miodovnik, 2011; Sanders
et al., 2018), which may persist throughout childhood and potentially thereafter even when the exposure level would have few
harmful effects in adults. For example, prenatal PFOS exposure may
impair behavioral regulation and metacognition in school-age
children (Vuong et al., 2016). Children can be exposed to PFASs
through the intake of food and water, inhalation of air in indoor
environments, and absorption from the daily use of materials
(Andersson et al., 2019; Rovira et al., 2019; Zhou et al., 2019).
However, few studies have explored PFAS exposure and the
potentially affected pathways in children relative to adults. Given
the paucity of current surveys on the status of PFAS contamination
in the air and drinking water, it is impossible to ascertain whether
either exposure route may induce harmful effects in children.
Therefore, the exposure levels and possible exposure pathways
must be monitored to minimize exposure to these chemicals as
much as possible.
Our preliminary surveys demonstrated the widespread prevalence of polybrominated diphenyl ethers (PBDEs), organophosphorus flame retardants (PFRs), and bisphenol A (BPA) in the urine
samples of children at kindergartens in Hong Kong and in airborne
particles collected from these kindergartens (Deng et al., 2016; Li
et al., 2019a). However, few studies have explored PFAS exposure
and possible exposure routes among children in Hong Kong. In this
study, we investigated the exposure of children in Hong Kong (age:
4e6 years) by detecting the concentrations of seven target PFASs in
their urine and hair samples via high-performance liquid
chromatographyetandem mass spectrometry (HPLC-MS/MS). We
additionally measured the concentrations of these PFASs in
airborne particles collected at kindergartens and in drinking water
(tap water/commercial bottled water), and evaluated the relationships between the PFASs concentrations in these matrixes. This is
the first report to describe the concentrations of various PFASs in
paired urine/hair and kindergarten air and water samples in Hong
Kong.
2.2. Sample collection
Kindergarten students in Hong Kong (N ¼ 53; age: 4e6 years)
participated in autumn 2017 to spring 2018. Consent to participate
was provided by the subjects’ parents. This study was approved by
the Human Ethical Review Committee of The Education University
of Hong Kong.
Distal hair samples (3e10 cm from the posterior vertex) were
collected by the children’s parents, using scissors precleaned with
acetone. Spot urine samples were collected from the children upon
awakening and stored in clean glass containers (100 mL). The urine
and hair samples were frozen at 20 C prior to analysis. For the
analysis of drinking water, residential tap water samples were
collected by the parents, and three brands of commercial bottle
water were purchased from a supermarket in Hong Kong. For
airborne particle collection, sampling equipment (MiniVol portable
air samplers, Airmetrics, USA) was used to collect PFASs in fine
particulate matter (PM2.5). The sampler was placed 1.0 m above the
floor level and set at a sampling rate of 1.7 L/min. A quartz membrane filter (47 mm; Whatman Inc., USA) was used to collect the
PM2.5, and each filter was conditioned in dry air for 2 days before
sampling. The total PM2.5 sampling volume was 2.5 m3 per site.
2.3. Sample extraction and determination of PFASs
For urine PFAS extraction, 2 mL aliquots of sampled urine were
spiked and then solvent-extracted with acetonitrile (10 mL). The
mixture was vortex-mixed for 10 s and centrifuged at 8000 rpm for
10 min. The acetonitrile extraction procedure was conducted three
times, and the samples were dried by nitrogen gas. Finally, each
sample was adjusted to 300 mL with methanol and acetonitrile (1:1
v/v).
PFASs in the drinking water samples were extracted using a
solid-phase extraction method with HLB cartridges (Oasis® HLB,
60 mg, 3 cc; Waters, USA). The spiked drinking water samples
(adjusted to pH 3.0) were extracted using cartridges that had been
pre-conditioned with methanol and acetonitrile (1:1 v/v). Subsequently, the cartridges were washed with methanol and acetonitrile (1:1 v/v) to obtain the analytes (1 mL/min). Finally, the sample
was centrifuged, concentrated by nitrogen gas, and adjusted to
300 mL.
The hair samples were rinsed with Milli-Q water in an ultrasonic
bath for 10 min and washed twice with acetone to remove surface
chemicals. The dried hair samples were cut into small pieces
(2e3 mm) using precleaned scissors. The filters containing airborne
particles and the blank filters were freeze-dried. Next, 0.1 g of each
hair and airborne particle sample was weighed and homogenized.
Ten milliliters of methanol and acetonitrile (1:1 v/v) were added to
each sample, followed by shaking for 10 min and centrifugation at
15,000 rpm for 10 min. The supernatants were collected in 15 mL
tubes, dried by nitrogen gas, and reconstituted in for chemical
analysis.
HPLCeMS/MS with a C18 column (5 mm, 2.1 mm inner
diameter 50 mm length, 5 mm particle size; Thermo Fisher Scientific, Waltham, MA, USA) was used to analyze the PFASs and the
extraction control standard (Agilent 1260 series and Agilent 6400
Triple Quadrupole, CA, USA) as described previously. (Wang et al.,
2017). Details are shown in S1.
2. Method
2.1. Chemicals
Perfluorononanoic acid (PFNA, 97%), perfluorodecanoic acid
(PFDA, 98%), perfluorooctanoic acid (PFOA, 96%), perfluorooctane sulfonic acid (PFOS, 97%), perfluoroheptanoic acid
(PFHpA, 98%), and perfluorohexanoic acid (PFHxA, 97%) were
purchased from Sigma-Aldrich (St. Louis, MO, USA). 13C8-labeled
PFOA (Wellington Laboratories, Guelph, Canada) was used as an
internal standard. Organic solvents (analytical grade) and other
2.4. Quality control and quality assurance
Standard solutions of each target chemical were prepared at
2
N. Li, G.-G. Ying, H. Hong et al.
Environmental Pollution 270 (2021) 116219
detected in hair samples at GM concentrations of 2.40 pg/g, 6.01 pg/
g, 19.4 pg/g, 24.0 pg/g, and 233 pg/g. The spatial distribution profiles of the PFASs in urine and hair samples are shown in Fig. S1 and
S2 in the Supporting Information.
concentrations ranging from 0.01 to 100 ng/mL. Procedural blanks
(Milli-Q water), instrumental blanks, duplicate samples (every 10
samples), and spiked blanks (Milli-Q water spiked with 13C8labeled PFOA) were prepared using the same procedures. The
target PFASs were not detected in the procedural or instrumental
blanks. The limits of detection (LODs) were set at a signal-to-noise
ratio of 3. The LODs and recoveries for PFASs in the drinking water
samples, urine samples, hair samples, and airborne particles are
listed in the Supporting Information (Table S1). The relative standard deviations (RSD) were all < 10.0%.
3.2. Detection of PFASs in drinking water
All of the target PFASs were detected in tap water samples from
Hong Kong (DF ¼ 100% for all), with GM concentrations ranging
from 0.17 to 18.1 ng/L (Table 3). PFOA was present at much higher
concentrations than the other target chemicals, with a maximum
concentration of 39.7 ng/L. Although the concentrations of PFOS
(GM ¼ 2.21 ng/L), PFHpA (GM ¼ 3.73 ng/L), and PFHxA
(GM ¼ 1.10 ng/L) were lower than that of PFOA (GM ¼ 18.1 ng/L),
they were nearly an order of magnitude higher than those of PFDDA
(GM ¼ 0.17 ng/L) and PFDA (GM ¼ 0.23 ng/L). All of the target PFASs
were also detected in the bottled water samples (DF ¼ 100% for all),
with GM concentrations ranging from 0.18 to 20.4 ng/L. PFOA was
also the dominant target PFAS in bottled water, with a maximum
concentration of 32.6 ng/L (i.e., similar to that in tap water). The
spatial distribution profiles of the target PFASs in water samples are
shown in Fig. S4 in the Supporting Information.
2.5. Evaluation of estimated daily intake of PFASs through drinking
water and airborne particles
To explore the health risks of exposure to these compounds
among children, the estimated daily intake (EDI) value of each
compound was calculated. Details of this method are described in
S2 and S3 in the Supporting Information.
2.6. Statistical analysis
PFAS concentrations below the specific LODs in hair/urine,
drinking water, and airborne particle samples were divided by the
square root of 2. Statistical analyses of the analytes with a detection
frequency (DF) > 50% were conducted using SPSS (version 19.0, IBM
Corp., Armonk, NY, USA). The normality of data distribution was
assessed using the KolmogoroveSmirnov test. The geometric mean
(GM) concentrations of PFASs in different groups were compared
using the ManneWhitney U test. Statistical significance was set at a
p value of < .05.
3.3. Detection of PFASs in airborne particles sampled from
kindergartens
Our analysis of PFASs in airborne particles sampled from 17
kindergartens revealed no detectable concentrations of PFOS and
PFDDA (Table 4). In contrast, the DFs of the other target PFASs were
all 100%, with GM concentrations ranging from 14.8 to 537 pg/m3.
Of the target PFASs, the concentrations of PFHpA (GM ¼ 270 pg/
m3), PFOA (GM ¼ 318 pg/m3), and PFDA (GM ¼ 537 pg/m3) were
higher than those of PFHxA (GM ¼ 14.8 pg/m3) and PFNA
(GM ¼ 17.1 pg/m3). The maximum concentrations of PFOA and
PFDA were as high as 1896 pg/m3 and 1653 pg/m3, respectively.
3. Results
3.1. Detection of PFASs in Children’s urine and hair samples
Our analysis of urine and hair samples from children in Hong
Kong revealed the presence of various PFASs. The DFs and concentrations of these PFASs were calculated and are listed in
Tables 1e4, respectively. All seven target PFASs were widely
detected in the urine samples, with DFs ranging from 32% to 100%
and GM concentrations ranging from 0.18 to 2.97 ng/L. The total
urinary PFAS concentrations reached up to 14.87 ng/mL, and PFOA
was predominant; as for the other target PFASs in urine samples,
PFDA (GM ¼ 0.65 ng/L, 87% DF), PFHpA (GM ¼ 0.88 ng/L, 75% DF),
PFNA (GM ¼ 0.18 ng/L, 68% DF), PFHxA (GM ¼ 0.89 ng/L, 58% DF)
and PFOS (GM ¼ 2.68 ng/L, 32% DF) were detected. PFOS and PFDDA
were not detected in hair samples. However, the other target PFASs
yielded DFs of 48e70% in hair samples, with GM concentrations
ranging from 2.40 to 233 pg/g. PFOA was also the predominant PFAS
in hair samples. PFDA, PFNA, PFHxA, PFHpA, and PFOA were
3.4. Estimated daily intake (EDI) of PFASs through drinking water
and airborne particles
In this study, the EDIs of PFASs through drinking (tap) water
ranged from 0.03 to 2.64 and from 0.04 to 4.38 ng/kg body weight
(bw)/day when the GM and the 95th percentile of the concentration were used, respectively (Table S3 in Supporting Information).
The corresponding EDIs of PFASs through airborne particle inhalation were 0.006e0.247 and 0.01e0.345 ng/kg bw/day (Table S2 in
Supporting Information).
Table 1
Urinary concentrations of PFASs (ng/L).
Chemicals
Mean
Geometric mean
Percentiles
50%
75%
95%
0.82
0.59
2.46
2.19
0.14
0.61
0.18
1.37
0.85
3.81
2.63
0.19
4.10
0.80
2.2
11.0
21.5
10.5
2.18
30.0
1.91
Min
Max
Detection rates%
0.28
0.34
0.45
1.39
5.98E-02
3.57E-02
2.81E-02
2.79
12.9
28.1
33.6
3.16
61.1
3.97
58
75
100
32
68
87
100
Children (n ¼ 53)
PFHxA
PFHpA
PFOA
PFOS
PFNA
PFDA
PFDDA
1.04
2.03
4.66
4.22
0.42
5.09
0.56
0.89
0.88
2.97
2.68
0.18
0.65
0.24
“-” means undetected; Abbreviations: LOD, the limit of detection; n, number of samples.
3
N. Li, G.-G. Ying, H. Hong et al.
Environmental Pollution 270 (2021) 116219
Table 2
Hair concentrations of PFASs (pg/g).
Chemicals
Mean
Geometric mean
Percentiles
50%
75%
95%
19.0
23.5
217
e
5.23
2.26
e
24.8
29.8
310
e
8.41
2.63
e
40.8
74.4
417
e
12.5
3.80
e
Min
Max
Detection rates%
7.85
12.1
105
e
2.99
1.60
e
41.7
80.4
459
e
20.9
4.42
e
59
70
63
0
70
48
0
Min
Max
Detection rates%
0.40
1.54
5.67
0.11
0.37
0.17
6.91E-02
2.20
7.59
39.7
8.63
1.02
0.37
0.85
100
100
100
100
100
100
100
0.62
4.47
8.27
0.73
0.79
0.16
6.67E-02
1.81
4.60
32.6
7.09
0.81
0.19
0.70
100
100
100
100
100
100
100
Children (n ¼ 27)
PFHxA
PFHpA
PFOA
PFOS
PFNA
PFDA
PFDDA
21.0
28.0
248
e
6.88
2.49
e
19.4
24.0
233
e
6.01
2.40
e
“-” means undetected; Abbreviations: LOD, the limit of detection; n, number of samples.
Table 3
Tap water and commercial bottled water concentrations of PFASs (ng/L).
Chemicals
Mean
Geometric mean
Percentiles
50%
75%
95%
1.48
3.78
22.4
3.74
0.71
0.22
0.10
1.70
4.48
28.9
6.57
0.79
0.25
0.42
2.02
6.50
35.1
7.72
1.00
0.33
0.75
Tap Water (n ¼ 12)
PFHxA
PFHpA
PFOA
PFOS
PFNA
PFDA
PFDDA
1.26
4.00
21.1
3.91
0.70
0.23
0.26
1.10
3.73
18.1
2.21
0.67
0.23
0.17
1.39
4.52
24.2
4.90
0.80
0.18
0.48
1.25
4.52
20.4
3.29
0.80
0.18
0.32
Battled water (n ¼ 3)
PFHxA
PFHpA
PFOA
PFOS
PFNA
PFDA
PFDDA
“-” means undetected; Abbreviations: LOD, the limit of detection; n, number of samples.
Table 4
Air particles concentrations of PFASs (pg/m3).
Chemicals
Mean
Geometric mean
Percentiles
50%
75%
95%
14.3
270
288
e
17.0
629
e
19.7
462
368
e
20.6
679
e
27.4
667
545
e
24.0
921
e
Min
Max
Detection rates%
7.98
83
197
e
10.1
0.67
e
34
902
1896
e
28.4
1653
e
100
100
100
e
100
100
0
Kindergartens (n ¼ 17)
PFHxA
PFHpA
PFOA
PFOS
PFNA
PFDA
PFDDA
15.9
326
358
e
17.7
658
e
14.8
270
318
e
17.1
537
e
“-” means undetected; Abbreviations: LOD, the limit of detection; n, number of samples.
et al. (2014a) detected perfluoro-n-pentanoic acid (PFPeA),
PFHpA, PFHxA, and PFBS in urine samples from 120 healthy South
Korean children (age: 5e13 years) and found that PFPeA (2.34 mg/L)
was predominant. In contrast, PFOA was predominant in samples
from children in Hong Kong in this research, consistent with a
previous study in China (Li et al., 2013). The dominance of PFOA in
samples from children in Hong Kong may be related to the shorter
half-life, faster excretion, and higher water-solubility of this
chemical (Sznajder-Katarzynska et al., 2019), as well as its extensive
use in the region. Few studies have evaluated the urinary PFAS
levels of children in China; however, the urinary PFAS concentrations measured in our study sample were lower than those reported for a general population of Chinese adults (mean
4. Discussion
4.1. Detection of PFASs in urine, hair, airborne particle, and drinking
water samples (Fig. 1)
Because it is very difficult to collect blood from young children,
some researchers have attempted to assess human exposure to
PFASs by using urine samples instead of serum (Kim et al., 2014b;
Perez et al., 2012; Zhang et al., 2013a, b). Some previous studies
identified strong positive correlations (p < .01) between the serum
and urine concentrations of PFOS, PFHxA, PFHpA, and PFOA in
adults, suggesting that urine concentrations of PFASs could be used
as biomonitoring biomarkers (Zhang et al., 2013a, b). In 2012, Kim
4
N. Li, G.-G. Ying, H. Hong et al.
Environmental Pollution 270 (2021) 116219
concentrations, 0.42e81 ng/L, >80% DF) (Zhang et al., 2013a, b).
Reports of PFAS detection in human hair have increased in
recent years. In a study of 14 cities in India, the concentrations of 25
PFASs measured in 39 human hair samples ranged from <0.02 to
3.78 ng/g, and PFHxS, PFOS, and PFOA were predominant (Ruan
et al., 2019). In another study conducted in Busan Metropolitan
City (South Korea), the concentrations of 11 PFASs in 47 hair samples ranged from 0.276 to 14.9 ng/g, with a mean value of 4.94 ng/g
(Kim and Oh, 2017). In this study of hair samples from 27 children
in Hong Kong, the seven target PFASs were detected at concentrations of 0.13e0.61 ng/g, which were slightly lower than those
measured in the Indian and South Korean studies (Kim and Oh,
2017; Ruan et al., 2019). Table 2 show that PFOA accounts for a
much larger proportion of PFASs in hair than in urine. Hair might be
a preferential site for PFOA bioaccumulation, as previous studies
have also reported much higher proportions of PFOA in hair samples (27.29%) relative to urine, nail, and serum samples
mez
(0.78e14.15%) (Li et al., 2013; Wang et al., 2018). Rodríguez-Go
et al. (2017) also reported that PFOA was the most commonly
detected among six PFASs analyzed in human hair samples
collected from Granada, Spain (Rodriguez-Gomez et al., 2017).
Drinking water is considered a common source of human PFAS
exposure. In certain cases, exposure through water intake was as
important as through dietary intake (Meng et al., 2019). PFASs are
detected ubiquitously in many countries (Domingo and Nadal,
2019). PFOA concentrations as high as 100 and 800 ng/L have
been detected in tap and well water samples in Japan and Ghana,
respectively, while PFOS concentrations as high as 100 and 40 ng/L
have been detected in Ghana and Spain (Essumang et al., 2017;
Shiwaku et al., 2016). In China, a nationwide survey of drinking
water samples from 79 cities detected 17 priority PFAAs, and the
sum concentrations (S17PFAAs) ranged from 4.49 to 174.93 ng/L,
with a mean value of 35.13 ng/L (Li et al., 2019b). Perfluorobutanoic
acids (PFBAs; median concentration ¼ 17.87 ng/L) were most
abundant, followed by PFOA (median concentration ¼ 0.74 ng/L),
PFNA (median concentration ¼ 0.40 ng/L), and PFOS (median
concentration ¼ 0.25 ng/L) (Li et al., 2019b). Compared with those
studies, the PFAS concentrations detected in drinking water from
Hong Kong in this study were low.
Although studies have increasingly focused on PFAS monitoring
and exposure, these have not raised concerns about airborne PFAS
concentrations in kindergartens and the corresponding exposure
risk. Goosey and Harrad (2012) detected PFASs (including PFOS,
PFHxS, and PFOA) in 20 home air samples from Birmingham (U.K.)
at median concentrations of 11, 23, and 24 pg/m3, respectively,
which were slightly lower than those measured in our study.
Vestergren et al. (2015) observed substantially higher concentrations of airborne PFASs in homes than in offices, and determined
that compounds with low water solubility and high vapor pressure
were dominant in indoor air samples.
4.2. The correlations of PFASs in different matrices
Spearman’s correlation test was used for a comparison among
the concentrations of PFASs with DFs of at least 50% in drinking
water, airborne particles, hair, and urine samples. Significant
Fig. 1. Distribution of PFASs in urine/hair/tap water/air in Hong Kong.
5
N. Li, G.-G. Ying, H. Hong et al.
Environmental Pollution 270 (2021) 116219
important, whereas indoor air inhalation and indoor dust contributed less than 1% and 2% of all PFOS and PFOA exposures, respectively (Ericson Jogsten et al., 2012). Wang et al. (2015) suggested
that residents of China are exposed to PFOA mainly via terrestrial
food (meat) and seafood. However, studies have increasingly illuminated that, in certain cases, drinking water may contribute
significantly to PFOS and PFOA exposure, and one study of 28 cities
in China observed a significant correlation between the PFOA levels
in the blood and drinking water (r ¼ 0.87, n ¼ 13, p < .001) (Zhang
et al., 2019). In our study, although we did not evaluate food intake,
we estimated that PFASs in drinking water may contribute more to
the overall exposure, compared with those in airborne particles,
consistent with the findings of previous research (Ericson Jogsten
et al., 2012).
correlations of PFHxA concentrations were observed with PFHpA,
PFOA, and PFNA in airborne particles; PFOA, PFOS, and PFDDA in tap
water; and PFHpA, PFOA, PFOS, PFDA, PFDDA, and PFNA in urine
samples (p < .01 or 0.05). Significant correlations of PFHpA were
observed with PFNA in airborne particles and hair and with PFOA
and PFNA in urine (p < .01 or 0.05). Significant correlations of PFOA
were observed with PFNA and PFHxA in airborne particles; PFOS,
PFNA, PFDDA, and PFHxA in tap water; and PFNA, PFHxA, and
PFHpA in urine (p < .01 or 0.05). Significant correlations of PFOS
were observed with PFHxA, PFOA, PFNA, and PFDDA in tap water
and with PFHxA in urine (p < .01). In particular, PFOA concentrations were strongly correlated with PFNA and PFHxA in all of the
matrices except hair, indicating the presence of common sources of
PFASs. In a study of office workers, serum PFOA levels were
significantly correlated with the office air levels (Fraser et al., 2012).
We found no significant correlations for either the individual or
total PFAS concentrations between the paired urine and hair samples in this study. The PFAS concentrations in hair were not
consistent with those in urine samples; PFOA was dominant in hair,
whereas PFOS, PFOA, and PFDA were dominant in urine. We also
analyzed sex-related differences in the PFAS concentrations in hair
and urine samples and identified a significant difference
(R2 ¼ 0.522, p ¼ .015) in the concentration of the PFOA in the former
group, with a higher value in boys than in girls. Regarding urine, a
previous study reported urinary PFNA elimination rates of 2% and
51% in male and female rats, respectively, within 120 h (Kudo et al.,
2001), suggesting a potential sex-related difference in urinary PFAS
concentrations. However, we did not observe significant differences
in the urine target PFAS concentrations between boys and girls. In a
study of eight PFAS concentrations in 15 men and 24 women (age:
20e53 years) in Shenzhen, China, no statistical sex-related differences in PFAS concentrations were observed in the urine, hair, and
nail samples (Wang et al., 2018). Therefore, no clear conclusion has
been reached regarding the relationship of PFAS concentrations
with genders, based on the data available so far (Jian et al., 2018).
4.4. Limitations and perspectives
Given the above standards, the current health risks of exposure
to these target chemicals via drinking water and airborne particles
may not be of concern. However, the likelihood of possible PFAS
exposure through drinking water and other sources among preschool children deserves special attention. In the future, more accurate guidelines for PFASs in food/drinking water and
environmental matrixes should be established. It is particularly
concerning that exposure to very low (i.e., picomolar and nanomolar) concentrations of these PFASs could disrupt the human
endocrine system (Zeng et al., 2019). Given the ubiquitous nature of
certain PFASs detected in children in Hong Kong, a detailed study of
the human health risks associated with exposure is warranted.
5. Conclusions
The results of this study suggest that PFASs are present ubiquitously in children, air samples from kindergartens, and drinking
water in Hong Kong. Although we measured lower concentrations
of PFASs in hair and urine samples and drinking water compared
with those in other regions, we measured higher concentrations in
airborne particles sampled from kindergartens. The body burden of
PFASs. Although the health risks associated with the PFAS exposure
encountered by children in Hong Kong through water and air intake
in kindergartens did not appear to be concerning, the health effects
of low-dose and long-term exposure to these chemicals cannot be
neglected.
4.3. Estimation of PFAS exposure through drinking water and
airborne particles
Limited epidemiological, pharmacokinetic, and toxicological
data, including the urinary excretion rates and the reference doses
(RfD) and multiple criteria, are available for many PFASs. However,
more toxicological data are available for PFOS and PFOA. Currently,
many standards exist for PFOS and PFOA and are based on different
toxicity endpoints. For example, T-cell-dependent antibody
response (TDAR) assays have led to no-observed-adverse-effect
levels (NOAELs) of PFOS and PFOA for immunotoxicity of 1.66 and
1 mg/kg bw/day, respectively. In drinking water, a combined
PFOS þ PFOA NOAEL of 70 ng/L was proposed by the United States
Environmental Protection Agency (USEPA) in 2016 (USEPA, 2016).
Still, although we measured maximum PFOA concentrations of only
39.7 and 32.6 ng/L in tap water and bottled water, respectively, we
cannot ignore the potential health risks of this target chemical and
PFOS. In 2018, the EFSA Scientific Panel on Contaminants in the
Food Chain set tolerable week intakes (TWIs) of 6 and 3 ng/kg bw/
week for PFOA and PFOS, respectively (EFSA et al., 2018). For the
children in this study, we calculated EDIs of PFASs of 0.04e4.38 ng/
kg bw/day through tap water when using 95% percent levels, and
EDIs of 0.01e0.345 ng/kg bw/day through airborne particles when
using 95% percent levels, indicating the potential health risks of
PFASs consumed via the drinking water in Hong Kong.
Exposure route identification is an important step in risk management. Food intake, or contamination from food packaging, has
been identified as the main pathway of human exposure to PFASs
(especially PFOS and PFOA). FDrinking water was secondarily
Authors statement
Na Li: Conceptualization, Writing e original draft. Wen-Jing
Deng: Funding acquisition, Supervision, Writing e review & editing. Guang-Guo Ying: Writing e review & editing. Hua-Chang
Hong: Writing e review & editing. Eric Po Keung Tsang: Writing
e review & editing.
Declaration of competing interest
The authors declare that they have no known competing
financial interests or personal relationships that could have
appeared to influence the work reported in this paper.
Acknowledgement
This research was financially supported by General Research
Fund of Hong Kong (No.18300919), FLASS Dean’s Research Fund
(04424&04389, 04547&04535) and Internal Teaching Development Grant (0300P&T0221) of The Education University of Hong
Kong.
6
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Environmental Pollution 270 (2021) 116219
Appendix A. Supplementary data
171e178.
Kim, D.H., Oh, J.E., 2017. Development and validation of an extraction method for
the analysis of perfluoroalkyl substances in human hair. Chemosphere 175,
446e451.
Kim, H.Y., Kim, S.K., Kang, D.M., Hwang, Y.S., Oh, J.E., 2014b. The relationships between sixteen perfluorinated compound concentrations in blood serum and
food, and other parameters, in the general population of South Korea with
proportionate stratified sampling method. Sci. Total Environ. 470e471,
1390e1400.
Y., K. Kudo, N., Suzuki, E., Katakura, M., Ohmori, K., Noshiro, R., 2001. Comparison of
the elimination between perfluorinated fatty acids with different carbon chain
length in rats Chem. Biol. Interact. 203e216.
Landrigan, P.J., Miodovnik, A., 2011. Children’s health and the environment: an
overview. Mt. Sinai J. Med. 78, 1e10.
Li, J., Guo, F., Wang, Y., Zhang, J., Zhong, Y., Zhao, Y., Wu, Y., 2013. Can nail, hair and
urine be used for biomonitoring of human exposure to perfluorooctane sulfonate and perfluorooctanoic acid? Environ. Int. 53, 47e52.
Li, N., Ho, W.K., Wu, R.S.S., Ying, G.G., Wang, Z.J., Jones, K., Deng, W.J., 2019a.
Organophosphate flame retardants and bisphenol A in children’s urine in Hong
Kong: has the burden been underestimated? Ecotoxicol. Environ. Saf. 183.
Li, Y., Li, J., Zhang, L., Huang, Z., Liu, Y., Wu, N., He, J., Zhang, Z., Zhang, Y., Niu, Z.,
2019b. Perfluoroalkyl acids in drinking water of China in 2017: distribution
characteristics, influencing factors and potential risks. Environ. Int. 123, 87e95.
Meng, J., Liu, S., Zhou, Y., Wang, T., 2019. Are perfluoroalkyl substances in water and
fish from drinking water source the major pathways towards human health
risk? Ecotoxicol. Environ. Saf. 181, 194e201.
Perez, F., Llorca, M., Farre, M., Barcelo, D., 2012. Automated analysis of perfluorinated compounds in human hair and urine samples by turbulent flow
chromatography coupled to tandem mass spectrometry. Anal. Bioanal. Chem.
402, 2369e2378.
Poothong, S., Papadopoulou, E., Padilla-Sanchez, J.A., Thomsen, C., Haug, L.S., 2020.
Multiple pathways of human exposure to poly- and perfluoroalkyl substances
(PFASs): from external exposure to human blood. Environ. Int. 134.
Rodriguez-Gomez, R., Martin, J., Zafra-Gomez, A., Alonso, E., Vilchez, J.L., Navalon, A.,
2017. Biomonitoring of 21 endocrine disrupting chemicals in human hair
samples using ultra-high performance liquid chromatography-tandem mass
spectrometry. Chemosphere 168, 676e684.
Route, W.T., Russell, R.E., Lindstrom, A.B., Strynar, M.J., Key, R.L., 2014. Correction to
spatial and temporal patterns in concentrations of perfluorinated compounds
in bald eagle nestlings in the upper midwestern United States. Environ. Sci.
Technol. 48, 9957-9957.
Sharma, R.P., Espuis, T., Nadal, M., Kumar, V.,
Rovira, J., Martínez, M.A.,
Costopoulou, D., Vassiliadou, I., Leondiadis, L., Domingo, J.L., Schuhmacher, M.,
2019. Prenatal exposure to PFOS and PFOA in a pregnant women cohort of
Catalonia, Spain. Environ. Res. 175, 384e392.
Ruan, Y., Lalwani, D., Kwok, K.Y., Yamazaki, E., Taniyasu, S., Kumar, N.J.I., Lam, P.K.S.,
Yamashita, N., 2019. Assessing exposure to legacy and emerging per- and polyfluoroalkyl substances via hair - the first nationwide survey in India. Chemosphere 229, 366e373.
Sanders, A.P., Saland, J.M., Wright, R.O., Satlin, L., 2018. Perinatal and childhood
exposure to environmental chemicals and blood pressure in children: a review
of literature 2007-2017. Pediatr. Res. 84, 165e180.
, J., Kar
Sharma, B.M., Bharat, G.K., Tayal, S., Larssen, T., Be
canova
askov
a, P.,
Whitehead, P.G., Futter, M.N., Butterfield, D., Nizzetto, L., 2016. Perfluoroalkyl
substances (PFAS) in river and ground/drinking water of the Ganges River basin: emissions and implications for human exposure. Environ. Pollut. 208,
704e713.
Shi, G., Cui, Q., Pan, Y., Sheng, N., Sun, S., Guo, Y., Dai, J., 2017. 6:2 Chlorinated
polyfluorinated ether sulfonate, a PFOS alternative, induces embryotoxicity and
disrupts cardiac development in zebrafish embryos. Aquat. Toxicol. 185, 67e75.
Shi, Y., Vestergren, R., Xu, L., Zhou, Z., Li, C., Liang, Y., Cai, Y., 2016. Human exposure
and elimination kinetics of chlorinated polyfluoroalkyl ether sulfonic acids (ClPFESAs). Environ. Sci. Technol. 50, 2396e2404.
Shiwaku, Y., Lee, P., Thepaksorn, P., Zheng, B., Koizumi, A., Harada, K.H., 2016. Spatial
and temporal trends in perfluorooctanoic and perfluorohexanoic acid in well,
surface, and tap water around a fluoropolymer plant in Osaka, Japan. Chemosphere 164, 603e610.
Soloff, A.C., Wolf, B.J., White, N.D., Muir, D., Courtney, S., Hardiman, G., Bossart, G.D.,
Fair, P.A., 2017. Environmental perfluorooctane sulfonate exposure drives T cell
activation in bottlenose dolphins. J. Appl. Toxicol. 37, 1108e1116.
Sunderland, E.M., Hu, X.C., Dassuncao, C., Tokranov, A.K., Wagner, C.C., Allen, J.G.,
2019. A review of the pathways of human exposure to poly- and perfluoroalkyl
substances (PFASs) and present understanding of health effects. J. Expo. Sci.
Environ. Epidemiol. 29, 131e147.
Sznajder-Katarzynska, K., Surma, M., Cieslik, I., 2019. A review of perfluoroalkyl
acids (PFAAs) in terms of sources, applications, human exposure, dietary intake,
toxicity, legal regulation, and methods of determination. J. Chem., 2717528
Tang, L.L., Wang, J.D., Xu, T.T., Zhao, Z., Zheng, J.J., Ge, R.S., Zhu, D.Y., 2017. Mitochondrial toxicity of perfluorooctane sulfonate in mouse embryonic stem cellderived cardiomyocytes. Toxicology 382, 108e116.
USEPA (United States Environmental Protection Agency), 2016. Drinking water
contaminant candidate list 4dfinal. Fed. Regist. 81099e81114.
Vestergren, R., Herzke, D., Wang, T., Cousins, I.T., 2015. Are imported consumer
products an important diffuse source of PFASs to the Norwegian environment?
Environ. Pollut. 198, 223e230.
Supplementary data to this article can be found online at
https://doi.org/10.1016/j.envpol.2020.116219.
References
Ahrens, L., Bundschuh, M., 2014. Fate and effects of poly- and perfluoroalkyl substances in the aquatic environment: a review. Environ. Toxicol. Chem. 33,
1921e1929.
Anderson, J.K., Luz, A.L., Goodrum, P., Durda, J., 2019. Perfluorohexanoic acid
toxicity, part II: application of human health toxicity value for risk characterization. Regul. Toxicol. Pharmacol. 103, 10e20.
Andersson, E.M., Scott, K., Xu, Y., Li, Y., Olsson, D.S., Fletcher, T., Jakobsson, K., 2019.
High exposure to perfluorinated compounds in drinking water and thyroid
disease. A cohort study from Ronneby, Sweden. Environ. Res. 176, 108540.
Bjerregaard-Olesen, C., Bach, C.C., Long, M., Ghisari, M., Bossi, R., Bech, B.H.,
Nohr, E.A., Henriksen, T.B., Olsen, J., Bonefeld-Jørgensen, E.C., 2016. Time trends
of perfluorinated alkyl acids in serum from Danish pregnant women
2008e2013. Environ. Int. 91, 14e21.
Calafat, A.M., Wong, L.Y., Kuklenyik, Z., Reidy, J.A., Needham, L.L., 2007. Polyfluoroalkyl 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, 1596e1602.
Chen, J., Hu, Y., Cai, A., Cheng, T., Wu, Z., Liu, H., Bao, X., Yuan, P., 2018. The
mesopore-elimination treatment and silanol-groups recovery for macroporous
silica microspheres and its application as an efficient support for polystyrene
hydrogenation. Catal. Commun. 111, 75e79.
Chou, H.C., Wen, L.L., Chang, C.C., Lin, C.Y., Jin, L., Juan, S.H., 2017. From the cover: lcarnitine via PPARg- and sirt1-dependent mechanisms attenuates epithelialmesenchymal
transition
and
renal
fibrosis
caused
by
perfluorooctanesulfonate. Toxicol. Sci. 160, 217e229.
Cui, Q., Pan, Y., Zhang, H., Sheng, N., Wang, J., Guo, Y., Dai, J., 2018. Occurrence and
tissue distribution of novel perfluoroether carboxylic and sulfonic acids and
legacy per/polyfluoroalkyl substances in black-spotted frog (pelophylax nigromaculatus). Environ. Sci. Technol. 52, 982e990.
Dalahmeh, S., Tirgani, S., Komakech, A.J., Niwagaba, C.B., Ahrens, L., 2018. Per- and
polyfluoroalkyl substances (PFASs) in water, soil and plants in wetlands and
agricultural areas in Kampala, Uganda. Sci. Total Environ. 631e632, 660e667.
Deng, W.J., Zheng, H.L., Tsui, A.K.Y., Chen, X.W., 2016. Measurement and health risk
assessment of PM2.5, flame retardants, carbonyls and black carbon in indoor
and outdoor air in kindergartens in Hong Kong. Environ. Int. 96, 65e74.
Domingo, J.L., Nadal, M., 2017. Per- and polyfluoroalkyl substances (PFASs) in food
and human dietary intake: a review of the recent scientific literature. J. Agric.
Food Chem. 65, 533e543.
Domingo, J.L., Nadal, M., 2019. Human exposure to per-and polyfluoroalkyl substances (PFAS) through drinking water: a review of the recent scientific literature. Environ. Res. 177.
€m, G., Domingo, J.L., 2012. PerEricson Jogsten, I., Nadal, M., van Bavel, B., Lindstro
and polyfluorinated compounds (PFCs) in house dust and indoor air in Catalonia, Spain: implications for human exposure. Environ. Int. 39, 172e180.
Essumang, D.K., Eshun, A., Hogarh, J.N., Bentum, J.K., Adjei, J.K., Negishi, J.,
Nakamichi, S., Habibullah-Al-Mamun, M., Masunaga, S., 2017. Perfluoroalkyl
acids (PFAAs) in the Pra and Kakum River basins and associated tap water in
Ghana. Sci. Total Environ. 579, 729e735.
Fraser, A.J., Webster, T.F., Watkins, D.J., Nelson, J.W., Stapleton, H.M., Calafat, A.M.,
Kato, K., Shoeib, M., Vieira, V.M., McClean, M.D., 2012. Polyfluorinated compounds in serum linked to indoor air in office environments. Environ. Sci.
Technol. 46, 1209e1215.
€ lkel, W., 2015. Neutral
Fromme, H., Dreyer, A., Dietrich, S., Fembacher, L., Lahrz, T., Vo
polyfluorinated compounds in indoor air in Germany e the LUPE 4 study.
Chemosphere 139, 572e578.
Genuis, S.J., Beesoon, S., Birkholz, D., 2013. Biomonitoring and elimination of perfluorinated compounds and polychlorinated biphenyls through perspiration:
blood, urine, and sweat study. ISRN Toxicol 2013, 483832.
Glynn, A., Berger, U., Bignert, A., Ullah, S., Aune, M., Lignell, S., Darnerud, P.O., 2012.
Perfluorinated alkyl acids in blood serum from primiparous women in Sweden:
serial sampling during pregnancy and nursing, and temporal trends
1996e2010. Environ. Sci. Technol. 46, 9071e9079.
Gomis, M.I., Wang, Z., Scheringer, M., Cousins, I.T., 2015. A modeling assessment of
the physicochemical properties and environmental fate of emerging and novel
per- and polyfluoroalkyl substances. Sci. Total Environ. 505, 981e991.
Goosey, E., Harrad, S., 2012. Perfluoroalkyl substances in UK indoor and outdoor air:
spatial and seasonal variation, and implications for human exposure. Environ.
Int. 45, 86e90.
Han, R., Zhang, F., Wan, C., Liu, L., Zhong, Q., Ding, W., 2018. Effect of perfluorooctane
sulphonate-induced Kupffer cell activation on hepatocyte proliferation through
the NF-kB/TNF-a/IL-6-dependent pathway. Chemosphere 200, 283e294.
Jian, J.M., Chen, D., Han, F.J., Guo, Y., Zeng, L., Lu, X., Wang, F., 2018. A short review on
human exposure to and tissue distribution of per- and polyfluoroalkyl substances (PFASs). Sci. Total Environ. 636, 1058e1069.
Kim, D.H., Lee, M.Y., Oh, J.E., 2014a. Perfluorinated compounds in serum and urine
samples from children aged 5e13 years in South Korea. Environ. Pollut. 192,
7
N. Li, G.-G. Ying, H. Hong et al.
Environmental Pollution 270 (2021) 116219
€din, A., Calafat, A.M., Braun, J.M.,
Vuong, A.M., Yolton, K., Webster, G.M., Sjo
Dietrich, K.N., Lanphear, B.P., Chen, A., 2016. Prenatal polybrominated diphenyl
ether and perfluoroalkyl substance exposures and executive function in schoolage children. Environ. Res. 147, 556e564.
Wang, Q.W., Yang, G.P., Zhang, Z.M., Jian, S., 2017. Perfluoroalkyl acids in surface
sediments of the East China Sea. Environ. Pollut. 231, 59e67.
Wang, T., Wang, P., Meng, J., Liu, S., Lu, Y., Khim, J.S., Giesy, J.P., 2015. A review of
sources, multimedia distribution and health risks of perfluoroalkyl acids
(PFAAs) in China. Chemosphere 129, 87e99.
Wang, Y., Chang, W., Wang, L., Zhang, Y., Zhang, Y., Wang, M., Wang, Y., Li, P., 2019.
A review of sources, multimedia distribution and health risks of novel fluorinated alternatives. Ecotoxicol. Environ. Saf. 182.
Wang, Y., Zhong, Y., Li, J., Zhang, J., Lyu, B., Zhao, Y., Wu, Y., 2018. Occurrence of
perfluoroalkyl substances in matched human serum, urine, hair and nail.
J. Environ. Sci. (China) 67, 191e197.
Wang, Z., Cousins, I.T., Scheringer, M., Buck, R.C., Hungerbühler, K., 2014. Global
emission inventories for C4eC14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, Part I: production and emissions from quantifiable
sources. Environ. Int. 70, 62e75.
Zeng, Z., Song, B., Xiao, R., Zeng, G., Gong, J., Chen, M., Xu, P., Zhang, P., Shen, M.,
Yi, H., 2019. Assessing the human health risks of perfluorooctane sulfonate by
in vivo and in vitro studies. Environ. Int. 126, 598e610.
Zhang, S., Kang, Q., Peng, H., Ding, M., Zhao, F., Zhou, Y., Dong, Z., Zhang, H., Yang, M.,
Tao, S., Hu, J., 2019. Relationship between perfluorooctanoate and perfluorooctane sulfonate blood concentrations in the general population and
routine drinking water exposure. Environ. Int. 126, 54e60.
Zhang, Y., Beesoon, S., Zhu, L., Martin, J.W., 2013a. Biomonitoring of perfluoroalkyl
acids in human urine and estimates of biological half-life. Environ. Sci. Technol.
47, 10619e10627.
Zhang, Y., Beesoon, S., Zhu, L., Martin, J.W., 2013b. Isomers of perfluorooctanesulfonate and perfluorooctanoate and total perfluoroalkyl acids in
human serum from two cities in North China. Environ. Int. 53, 9e17.
Zhou, W., Zhao, S., Tong, C., Chen, L., Yu, X., Yuan, T., Aimuzi, R., Luo, F., Tian, Y.,
Zhang, J., Shanghai Birth Cohort, S., 2019. Dietary intake, drinking water
ingestion and plasma perfluoroalkyl substances concentration in reproductive
aged Chinese women. Environ. Int. 127, 487e494.
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