Chemosphere 129 (2015) 87–99
Contents lists available at ScienceDirect
Chemosphere
journal homepage: www.elsevier.com/locate/chemosphere
A review of sources, multimedia distribution and health risks of
perfluoroalkyl acids (PFAAs) in China
Tieyu Wang a, Pei Wang a,b, Jing Meng a,b, Shijie Liu a,b, Yonglong Lu a,⇑, Jong Seong Khim c, John P. Giesy d
a
State Key Lab of Urban and Regional Ecology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China
University of Chinese Academy of Sciences, Beijing 100049, China
c
School of Earth and Environmental Sciences & Research Institute of Oceanography, Seoul National University, Seoul, Republic of Korea
d
Department of Veterinary Biomedical Sciences and Toxicology Centre, University of Saskatchewan, Saskatoon, Saskatchewan, Canada
b
h i g h l i g h t s
PFOS and PFOA are the predominant compounds found among environmental media.
PFAAs emissions from industrial source are significant higher than domestic source.
Higher levels of PFAAs were detected in more industrialized and urbanized areas.
Regional urbanization is closely related to emission and contamination of PFAAs.
Risk assessment of PFAAs in China has lagged behind those in developed countries.
a r t i c l e
i n f o
Article history:
Received 2 April 2014
Received in revised form 4 September 2014
Accepted 4 September 2014
Available online 26 September 2014
Handling Editor: I. Cousins
Keywords:
PFOS
PFOA
Source emission
Contamination
Risk assessment
China
a b s t r a c t
Perfluoroalkyl acids (PFAAs) have been recognized as emerging pollutants because of their ubiquitous
occurrence in the environment, biota and humans. In order to investigate their sources, fate and environmental effects, a great number of surveys have been carried out over the past several years. In the present
review, we summarized the status of sources and emission, concentration, distribution and risks of PFAAs
in China. Concentrations of PFAAs, especially perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic
acid (PFOA) in various environmental media including water, sediment, soil, rain, snow and organisms, as
well as human tissues are summarized based on the available data. Concentrations of PFAAs in aquatic
systems are higher in relatively more industrialized and urbanized areas than those from the less populated and remote regions in China, indicating that their emission and distribution are closely related to
regional urbanization and industrialization. PFAAs and related products have been widely used over
the past several decades, which have brought about high concentrations detected in environmental
matrixes, biota and even local residents. Ecological risk assessment of PFAAs is still less developed in
China. Most existing studies compared concentrations of PFAAs to guideline values derived for single species to evaluate the risk. In order to reveal the transport, partitioning and degradation of PFAAs in the
environment, further studies on their behavior, fate, bioaccumulation and adverse effects in different trophic levels should be conducted.
Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction
Perfluoroalkyl acids (PFAAs) are of unique and useful chemical
properties including surface activity, thermal and acid resistance,
as well as repellency of water and oil, thus they have been used
worldwide in various applications such as stain repellents, food
packaging and firefighting foams for more than 60 years (Giesy
⇑ Corresponding author. Tel.: +86 10 62849466; fax: +86 10 62918177.
E-mail address: yllu@rcees.ac.cn (Y. Lu).
http://dx.doi.org/10.1016/j.chemosphere.2014.09.021
0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.
and Kannan, 2001; Wang et al., 2009; Blaine et al., 2013). Due to
high energy of carbon–fluorine bonds, PFAAs are resistant to
hydrolysis, photolysis, microbial degradation, and metabolism by
vertebrates (Kissa, 2001). PFAAs, especially perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA), were first
reported to be widespread in the environment (Giesy and
Kannan, 2001, 2002), and they were subsequently detected in
aquatic systems (Fujii et al., 2007; Rayne and Forest, 2009) and
wildlife (Kannan et al., 2002; Houde et al., 2006a; Suja et al.,
2009). Many PFAAs can be accumulated in aquatic system, leading
88
T. Wang et al. / Chemosphere 129 (2015) 87–99
to bioaccumulation and biomagnification through the food chain to
wildlife and humans (Loi et al., 2011; Squadrone et al., 2014). There
is still long-term public concern over adverse effects of PFAAs on
ecosystem and human health as well as secondary release of PFAAs
from the environment (Pico et al., 2011; Lu et al., 2013).
In recent years, great attention has been paid to PFOS and PFOA,
two of the frequently detected predominant PFAAs in the environment. PFOS and related chemicals have recently been added to the
list of ‘‘persistent organic pollutants’’ (POPs) under the Stockholm
Convention. Inclusion in Annex B of the Stockholm Convention
allows limited on-going uses of PFOS, mostly in the semiconductor,
photolithography and metal plating industries (Wang et al., 2009).
However, in order to meet the growing demand for surfactants and
other surface modification applications, China has become one of
the largest countries of PFOS and related chemicals production
and consumption. From 2001 to 2006, the annual production of
PFOS and related chemicals increased from less than 50 t to about
250 t, and kept almost consistently until 2011. The cumulative historical production volume of PFOS and related chemicals were estimated to be 1800 t by 2011 (Xie et al., 2013b).
In China, research on PFAAs was initiated much later when
compared with more industrially developed countries (Fig. 1).
The status of PFAAs in China was first reported by Jin et al.
(2003). From 2003 to 2014, studies about PFAAs in China have been
focused primarily on four aspects, including detection of PFAAs in
different environmental media, risk assessment of PFAAs on environment and ecosystem, toxicology to humans and wildlife, and
estimation of sources and emission (Wang et al., 2010; Cai et al.,
2012; Wu et al., 2012; Xie et al., 2013a,b; Yeung et al., 2009;
Zhang et al., 2011). Previous studies have summarized and compared PFAAs concentration and contamination profiles in aquatic
biota and human tissues, emissions and degradation mechanisms,
bioaccumulation and biotransformation processes (Houde et al.,
2006b, 2011; Lau et al., 2007; Conder et al., 2008; Houde et al.,
2011; Liu and Avendano, 2013). However, historically only very
limited information about the profile of PFAAs in China is available
in the existing review literature. Zhao et al. (2012a) provided a
review of environmental contamination and human exposure for
PFOS. In 2009, we published our first review of spatial and temporal distributions of PFOS and PFOA in China, covering the articles
published before July 2008 (Chen et al., 2009), in which we summarized production, contamination, and exposure of humans to
PFOS and PFOA. However, studies on PFAAs in China and resulting
publications have increased dramatically in recent years, especially
after PFOS was included in the national priority control list in 2009.
We have carried out systematic studies in the coastal regions of the
Bohai and Yellow Seas. The studies are focused on correlation
between varied degrees of industrialization and PFAAs concentrations in environmental media, classification and identification of
PFAAs emissions from major industries, flux of PFAAs emissions
in some wastewater treatment plants, and effects of PFAAs emissions from fluoropolymer facilities on the coastal rivers (Chen
et al., 2011a,b; Wang et al., 2011b, 2012, 2014b; Xie et al.,
2013a). In the present review, we aim to summarize and update
information on PFAAs sources, status and trends in concentrations
in various media such as water, sediment, soil, sludge, rain and
snow, and assessment of risks to humans and wildlife in China.
We have also endeavored to assess gaps in knowledge and understanding and suggest possible future research perspectives.
2. Potential sources and emission estimation
PFAAs have been manufactured and used in the globe for more
than 50 years. At present, although the manufacture and use of
PFAAs-based products, especially PFOS and PFOA, have been
restricted or eliminated in many developed countries, PFAAs and
related substances are still manufactured and used in certain
industries including metal plating, photographic, fire-fighting
foams, semiconductor and aviation in China (Xie et al., 2013b).
China began to produce PFAAs-based products later than
Fig. 1. Temporal trends of research progress on PFAAs in China. Reference: (Jin et al., 2003, 2009; So et al., 2006a,b; Ju et al., 2008; Liu et al., 2009a; Shi et al., 2010; Zhang
et al., 2010a,b, 2012a,b; Guo et al., 2011; Loi et al., 2011; Xie et al., 2013a,b).
T. Wang et al. / Chemosphere 129 (2015) 87–99
industrially developed countries. From 2003 to 2006, the annual
production of PFAAs-related chemicals had been rapidly increased,
due to the sharp increase in both domestic and overseas demands
in various application fields (Footitt et al., 2004; Liu et al., 2008a,c).
As a result, PFAAs-related industrial processes are still the significant sources of PFAAs releases to the environment in China.
Nevertheless, studies on PFAAs emissions are still at an initial
stage compared with those on environmental exposure and toxic
effects. Furthermore, most of the investigations have focused on
estimation of PFOA and PFOS emissions. Sources of PFOS and PFOA
emitted to the environment in China and other countries are illustrated in Table S1. Historically, global emission of PFOA from both
direct and indirect sources was estimated based on the production
data (Armitage et al., 2006, 2009; Prevedouros et al., 2006). Wang
et al. (2014c) estimated emissions of perfluoroalkyl carboxylic
acids (PFCAs) from quantifiable sources from 1951 to 2030. The
results showed that 98–100% of historical (1951–2002) PFOA emissions are attributed to direct releases during the life-cycle of products containing PFOA. In order to track the geographical shift of
industrial sources, the main countries of manufacturing and using
PFOA were divided into two groups: I (Japan, Western Europe and
the US); II (Russia, China, India and Poland). According to estimation from PFOA production sites in plausible scenario (i.e. assuming a use rate of 0.3 wt% PFOA-based products in fluoropolymer
production), the total annual emissions of PFCAs (C4–C14) in country group II were about 2 t a 1 in 1980, but currently, the emissions
exceeded 25 t a 1. PFOA and its derivatives are important additives
in the production of fluoropolymers, which is bound to cause pollution of PFOA. It is noted that domestic demand and production
PTFE in China rapidly increased from 6.6 kt a 1 in 1999 to about
64 kt a 1 in 2012 (Fang, 2004; Cai, 2009; Kälin et al., 2012;
Wang, 2006). Chen et al. (2011b) tried to estimate mass flow of
PFOA in some rivers in China, and the results showed that the highest levels of PFOA discharges were found in the Yangtze River
(39 200 kg a 1) and the Huangpu River (16 000 kg a 1) in 2005.
Information on emission estimation of PFOS can be divided into
two categories. The first category is estimation based on rates of
PFOS used in specific industries, and emission factors. In this category, PFOS emissions are estimated for both the global and various
geographical regions. For example, an estimate of global historical
production and environmental releases of PFOS was performed
(Paul et al., 2009). In the ‘‘Environment risk evaluation report:
PFOS’’ produced by the UK Environment Agency, the PFOS emissions from metal plating, photolithography, photographic industry,
aviation, fire-fighting, fabric treatment, paper treatment and coatings in UK and EU were estimated (Brook et al., 2004). In China,
estimation of PFOS released from chromium plating and semiconductor manufacturing sites and environmental concentrations of
PFOS near the sites was made based on information in the EU technical guidance document and use of the EUSES model. The emission of PFOS from chromium plating site was estimated to be
0.0012 kg d 1 to water and 0.00072 kg d 1 to air, while the emission from semiconductor manufacturing sites was estimated to
be 0.007 kg d 1 to water (Liu et al., 2008a,b). Liu et al. (2008c) also
estimated the PFOS emission from fabric treatment, metal plating,
semiconductor manufacturing, and fire-fighting, using the emission factors in the EU technical guidance document. It was estimated that the total amount of PFOS annual emission in whole
China was 1.4 t to water from fabric treatment, 7.9 t to water
and 0.005 t to air from metal plating, 0.02 t to water from semiconductor manufacturing, 1.6 t to water and 0.08 t to air from firefighting, respectively.
The second category is estimation of production, use and
release of PFOS based on concentrations and water flux in rivers
or wastewater treatment plants (WWTPs). For instance, Pistocchi
and Loos (2009) estimated overall aqueous emission of PFOS from
89
the European Continent by using measured concentrations of PFOS
in rivers. Similar methods were applied by Kim (2012) to estimate
the watershed-based riverine discharge load of PFOS on the Korean
peninsula. Total flux of sewage-derived PFOS from Japan was estimated by use of per capita emission factor, which was derived
from PFOS loads in the wastewater effluents and the population
served by the WWTPs (Murakami et al., 2008). In China, discharges
of PFOS from major rivers, effluent and sludge in several WWTPs
were estimated by Chen et al. (2011b). Total discharges of PFOS
in main rivers from the coastal regions of the North Bohai Sea in
2008 were estimated to be 122 kg a 1, respectively, to which the
Daliao River and Daling River made great contributions. PFOS discharges in the Pearl River Delta and Yangtze River in 2005 were
found to be 4470 and 807 kg a 1, respectively. As mentioned
above, little information is available on PFOS-related industries
and emissions of PFOS in China. In order to reflect the overall PFOS
emission pattern in China, it is necessary to conduct a systematic
assessment of sources and regional emissions of PFOS-related
chemicals.
PFOS emissions from industrial or domestic sources in China
have been undertaken by our group. Emissions of PFOS equivalents
from major industrial sources including manufacture and use in
textile treatment, metal plating, fire-fighting and semiconductor
industry, were estimated at the provincial level (Xie et al.,
2013b). Total emission of PFOS equivalents from industrial sources
was estimated to be 70 t in 2010 (Fig. 2). This result was about 6
times higher than that conducted by Liu et al. in 2008. There might
be two reasons: Firstly, from 2008 to 2010, related industries have
developed very fast in China. For example, the production capacity
of integrated circuit industry (ICI) has increased 56% from 2008 to
2010, and the increasing capacity on related industries would lead
to increasing emission of PFOS; Secondly, Xie et al. provided
detailed information on emission methodology but underestimation still exists, while Liu et al. just gave a very general and primary
results of PFOS emission, which would lead to more underestimation. Industrial emission of PFOS equivalents in the eastern region
of China was remarkably higher than those in central and western
regions. For the eastern coastal provinces, the Gross Domestic
Products (GDP) is similar to those of upper-middle-income economies (with per capita GDP up to 4136–12745 U.S. dollars) (The
World Bank Group, 2014) (Fig. 2). Metal plating contributed the
largest portion of emissions of PFOS equivalents in eastern and
central China, while fire-fighting services were dominant in the
emissions in western China. Estimation of domestic PFOS emission
was based on the assumption that both PFOS and precursors
released to air, dust and water would enter the municipal wastewater system through cleaning, wiping and washing of the products and indoor environment (Xie et al., 2013a). Total emission
from domestic sources in the eastern coastal region of China was
381 kg in 2010. The results showed that emission from domestic
sources was much less than that from industrial sources, and the
most populous and developed regions, the Pearl River Delta, Bohai
coastal region and Yangtze River Delta, were responsible for 74% of
the total domestic emission of PFOS in the eastern and coastal
regions of China (Fig. 3).
3. Concentrations of PFAAs in various environmental media
3.1. PFAAs in freshwater
Concentrations of PFAAs have been determined in seven major
river systems including main stems and tributaries in China, especially during the last few years. PFAAs are widespread in surface
water in China (Fig. 4, original detailed data for histogram were
provided in Table S2). PFAAs from the Pearl, Yangtze and Haihe
90
T. Wang et al. / Chemosphere 129 (2015) 87–99
Fig. 2. PFOS emissions from industrial sources in China in 2010.
Fig. 3. PFOS emissions from domestic sources in the eastern and coastal regions of China in 2010.
Rivers are the primary focus of monitoring studies. Concentrations
of PFOA and PFOS were 0.85–13 and 0.9–99 ng L 1 in surface water
from the lower stream of the Pearl River, while those were 2–260
and nd–14 ng L 1 from the lower stream of the Yangtze River in
2004, respectively (So et al., 2007). Concentrations of PFOA and
PFOS in water from the lower stream of the Pearl River in 2012
were decreased, especially for PFOS, becoming 0.71–8.7 and
0.52–11 ng L 1, respectively (Zhang et al., 2013e). The decreasing
trend may be caused by enforcement of the Stockholm Convention
on PFOS and related chemicals. In 2003, concentrations of PFOA
and PFOS in surface water from the middle stream of the Yangtze
River were 0.2–297.5 and 0.1–37.8 ng L 1, respectively (Jin et al.,
2009). There are numerous industrial zones along these major rivers in China. For example, the Yangtze River flows through the
highly industrialized and commercial cities such as Chongqing
and Wuhan in the middle stream, and Nanjing and Shanghai in
the lower stream, where diversified industrial sources and intensive commercial activities could generate high concentrations of
PFAAs. In addition, the Haihe River, which is located in Northern
China, also exhibits high concentrations of PFOA and PFOS. Studies
of PFAAs in the Haihe River were all concentrated in the vicinity of
the city of Tianjin, which is situated in the central zone of the Bohai
Economic Rim. The ranges of concentrations were similar for PFOA
or PFOS, and relatively higher concentrations were observed in the
Binhai New Economic Development Area, where the highest concentrations of PFOA and PFOS were 42.1 and 10.6 ng L 1, respectively (Li et al., 2011; Pan et al., 2011; Wang et al., 2011b). Less
data are available for the Songhua, Liaohe, Yellow and Huaihe Rivers. In general, PFAAs in these four rivers were detected with lower
concentrations compared with those from the Pearl, Yangtze and
Haihe Rivers. In northeast China, mean concentrations of PFOA
and PFOS in surface water of the Songhua River were 0.17 and
1.21 ng L 1 (Liu et al., 2007); while those in the Liaohe River were
10.9 and 0.33 ng L 1, respectively (Yang et al., 2011). Even though
it is one of the major rivers in China little data is available for the
Yellow River. Concentrations of PFOA and PFOS in water of tributaries of the Yellow River were less than 15 ng L 1 (Wang et al.,
2012). In central China, data is available for only the Jiangsu reach
T. Wang et al. / Chemosphere 129 (2015) 87–99
91
Fig. 4. Mean concentrations of PFOS and PFOA in water and sediment from the major river systems in China.
of the Huaihe River, where the concentrations of PFOA and PFOS
were 18 and 4.7 ng L 1, respectively (Yu et al., 2013).
Several studies have focused on seriously contaminated tributaries of major rivers, such as the Hanjiang, Huangpu and Daling
Rivers (Fig. 4). The Hanjiang River is the largest tributary of the
Yangtze River, and flows through the city of Wuhan. The Huangpu
River is in the downstream portion of the Yangtze River, which
connects Taihu Lake and the East China, and flows through the
major city of Shanghai. Pollutants from industrial and domestic
wastewater have been discharged into these rivers and cause
severe contamination. The mean concentrations of PFOA and PFOS
in the Hanjiang River were 81 and 51.8 ng L 1, while in the
Huangpu River they were 105 and 5.4 ng L 1, respectively (So
et al., 2007; Wang et al., 2013a). Relatively higher mean concentration of PFOA (169.04 ng L 1) was observed in the Daling River,
while lower PFOS (0.42 ng L 1) was detected (Bao et al., 2011).
The Daling River is located in the Northern Bohai coastal region,
which is one of the most developed regions in north China. There
are two fluorine chemical parks along the Daling River, which
release large amount of wastes directly to the river (Wang et al.,
2013c). These two industrial parks were built in 2004 and 2006
respectively, and they currently produce fluoropolymers such as
PTFE. In the city of Shenyang, the capital of Liaoning province, concentrations of PFOA and PFOS in the Hunhe River were at moderate
levels with mean concentrations of 7.32 and 0.51 ng L 1, compared
to those in other rivers in China (Jin et al., 2009; Sun et al., 2011).
Studies on PFAAs in surface waters of lakes or reservoirs have
been mainly conducted in some great lakes and urban lakes
(Fig. 5; original detailed data are provided in Table S2). Concentrations of PFAAs in East and Tangxun Lakes were approximately
100-fold higher than those from other lakes (Chen et al., 2012b;
Zhou et al., 2013). Both East and Tangxun Lakes are urban lakes of
Wuhan in Hubei province. In Tangxun Lake, it was noted that short
chain PFAAs, mainly perfluorobutanesulfonate (PFBS) and
perfluorobutanoic acid (PFBA), were detected with extremely high
concentrations, with mean concentrations of 4770 and 3660 ng L 1,
respectively, while concentrations of PFOA and PFOS were 372 and
357 ng L 1, respectively. The results reveal that C4-PFAAs acting as
substitutes of C8-PFAAs have been extensively produced and
applied, and monitoring of C4-PFAAs in the environment has
become urgent. PFAAs from three great lakes were also reported,
including the Poyang, Tai and Chaohu Lakes. These three lakes are
all distributed along the Yangtze Watershed and influenced by
intensive human activities. Among these three lakes, the Taihu Lake
has been the most contaminated water body. Concentrations of
PFAAs in the Taihu Lake were similarly reported by Yu et al.
(2013) and Yang et al. (2011). Mean concentrations of PFOA and
PFOS in surface water were reported to be 56 and 15 ng L 1, respectively (Yu et al., 2013), and 21.7 and 26.5 ng L 1, respectively (Yang
et al., 2011). In the Chaohu Lake, only PFOS in water was investigated, and the highest concentration was 400 ng L 1 at a location
close to sewage outfall, situated in the industrially intensive district
of Chaohu City (Zhang et al., 2012b). The Poyang Lake was slightly
contaminated by PFAAs, with mean concentrations of PFOS and
PFOA in water of 0.35 and 1.1 ng L 1, respectively (Zhang et al.,
2012a). In general, concentrations of PFAAs are higher in lakes in
the Yangtze Watershed than those in other lakes. Some lakes in
northern and southwestern China have also been studied. Concentrations of PFOA and PFOS in water from the Dianchi Lake, which is
located in Yunnan province in southwest China, were 3.4–35.4 and
1.71–40.9 ng L 1, respectively (Zhang et al., 2012c). Similar to those
in the Dianchi Lake, moderate concentrations of PFOA (1.71–
43.5 ng L 1) and PFOS (0.11–1.48 ng L 1) were detected in the Baiyangdian Lake, which is the largest natural freshwater body in north
China (Shi et al., 2012). Concentrations of PFOA and PFOS in water of
the Guanting Reservoir were less than those in other lakes, with levels of 0.55–2.3 ng L 1 and nd–0.52 ng L 1, respectively (Wang et al.,
2011a).
92
T. Wang et al. / Chemosphere 129 (2015) 87–99
Fig. 5. Mean concentrations of PFOS and PFOA in water and sediment from lakes or reservoirs in China.
Most PFAAs in China are found at the moderate levels compared
with those in other countries or regions such as the USA, Canada,
Japan and Europe. In general, higher concentrations of PFAAs were
detected in mainstreams or tributaries of the Pearl, Yangtze and
Haihe Rivers, and the East, Tangxun and Taihu Lakes in China,
which were comparable with those in 18 rivers throughout whole
Japan (PFOA: 0.76–192 ng L 1; PFOS: nd–191 ng L 1) (Murakami
et al., 2008), and the Great Lakes in North America (PFOA: 27–
50 ng L 1; PFOS: 21–70 ng L 1) (Boulanger et al., 2004), while
lower concentrations of PFAAs in the studied rivers or lakes in
China were similar to those in European rivers, such as the River
Elbe in Germany (PFOA: 2.8–9.6 ng L 1; PFOS: 0.5–2.9 ng L 1)
(Ahrens et al., 2009), the River Seine in France (PFOA: 1.1–
18.0 ng L 1; PFOS: 9.9–39.7 ng L 1) (Labadie and Chevreuil, 2011),
and the River Rhine (PFOA: 0.61–41.4 ng L 1; PFOS: 0.89–
18.6 ng L 1) (Moeller et al., 2010). PFAAs in surface water from
some sites contaminated by fluorine chemical plants, mainly
including East and Tangxun Lakes in Wuhan (Zhou et al., 2013)
and Daling River in Fuxin (Bao et al., 2011), were less than those
from a cove into which 3 M plant waste water was directly discharged (PFOA: 3,600 ng L 1; PFOS: 18 200 ng L 1) (Oliaei et al.,
2013).
3.2. PFAAs in sediments
Data on concentrations of PFAAs in sediments are available
mostly for rivers that also have high concentrations in surface
water (Fig. 4). The highest concentrations of PFOA and PFOS were
detected in the Huangpu River, with 203 and 8.78 ng g 1 dry
weight (dw), respectively (Li et al., 2010). Bao et al. (2010) reported
that concentrations of PFOA and PFOS in sediments of the Huangpu
River were 0.2–0.64 and nd–0.46 ng g 1 dw, which were less than
that reported by Li et al. (2010). The difference between these two
reports might due to the difference in sample collection. In the
study by Li et al., the highest PFAAs were observed around a fluorine chemical plant for producing PTFE which should be responsible for the higher level of PFOA, while in the study by Bao et al.,
sediments were collected from the lower reach of the Huangpu
River where no point source existed. Concentrations of PFOA in
sediments from the Daling River ranged from 0.18 to 48 ng g 1 dw,
while PFOS was not detected (Bao et al., 2011). The greater frequency of detection of PFOA than other PFAAs could be due to
emissions from the local fluorine chemical parks. Concentrations
of PFAAs in sediments from the Liaohe River, adjacent to the Daling
River, were also analyzed (Yang et al., 2011), PFOA and PFOS were
both detected at concentrations less than 0.48 and 0.18 ng g 1 dw,
respectively. Other studies on distribution of PFAAs in sediments of
rivers were mostly conducted in the Haihe River (Li et al., 2011;
Pan et al., 2011; Wang et al., 2011b), and there was no significant
difference among these results. Data on concentrations of PFAAs in
sediments of lakes or reservoirs are limited. Mean concentrations
of PFOA and PFOS in Tangxun Lake were 2.4 and 74 ng g 1 dw,
respectively (Zhou et al., 2013) (Fig. 5). In addition, PFBA and PFBS
were detected with extremely high concentrations of 16.3 and
50.8 ng g 1 dw, respectively, which was consistent with the results
from the surface water. Concentrations of PFAAs in sediments
remained lower levels from other lakes or reservoirs including
Guanting Reservoir, Baiyangdian and Taihu Lakes (Wang et al.,
2011a; Shi et al., 2012; Yang et al., 2011).
It is noted that concentrations of PFOA and PFOS in sediments in
most sites from China are comparable to those from other countries or regions, such as Roter Main River, Germany (PFOA: 0.02–
0.07 ng g 1 dw; PFOS: 0.09–0.35 ng g 1 dw) (Becker et al., 2008),
San Francisco Bay, USA (PFOA: nd–0.40 ng g 1 dw; PFOS: nd–
3.76 ng g 1 dw) (Higgins et al., 2005), several rivers in Japan (PFOA:
nd–3.9 ng g 1 dw; PFOS: nd–11 ng g 1 dw) (Senthilkumar et al.,
2007), and coastal areas of Korea (PFOA and PFOS: less than
2.0 ng g 1 dw) (Naile et al., 2010). Relatively higher concentrations
93
T. Wang et al. / Chemosphere 129 (2015) 87–99
of PFAAs in sediments have been detected in the surroundings of
fluorine chemical plants in the Huangpu River, Daling River and
Tangxun Lake.
3.3. PFAAs in soils and sludge
Less data is available on concentrations of PFAAs in soils
(Table 1). Concentrations of PFAAs in soils from Shanghai were
the highest, and PFOA and PFOS in soils from Shanghai were
detected at all sampling sites with concentrations in the range of
3.28–47.5 and 8.58–10.4 ng g 1 dw, respectively (Li et al., 2010),
which was consistent with the results for concentrations in surface
water and sediments. Apart from frequent detections of PFAAs in
Shanghai, all other studies conducted in Tianjin, Guanting Reservoir in Beijing and Hebei province, north Bohai Sea and Huaihe
Watershed showed slight contamination by PFAAs (Pan et al.,
2011; Wang et al., 2011a,b,c; Meng et al., 2013). In these studies,
PFAAs were not detected in most samples, and the highest concentration of 2.8 ng g 1 dw for PFOA was observed in Guanting Reservoir, while the highest concentration of 9.4 ng g 1 dw for PFOS was
observed in Tianjin Binhai New Economic Development Area. Studies on PFAAs in soils from other countries or regions are also limited. Strynar et al. (2012) collected 60 soil samples from six
countries including USA, China, Japan, Norway, Greece and Mexico,
and found the concentrations of PFOA and PFOS in the range of nd–
31.7 ng g 1 dw and nd–10.1 ng g 1 dw, respectively. The results
showed that relatively higher detections of PFAAs were mostly
observed in USA and Japan among the six countries, and the high
concentrations were comparable to those in soils from Shanghai.
Concentrations of PFAAs have been measured in sludge from
wastewater treatment plants (WWTPs), which are mainly situated
in some developed cities (Table 1). Sludge from WWTPs can be
applied to improve soils. Thus, PFAAs in sludge could affect concentrations in soils to which they are applied. In Shanghai, concentrations of PFOA and PFOS of sludge collected in 2008 from
different WWTPs were 9.21–75.5 and 28.1–135 ng g 1 dw, which
were higher than those in local soils and sediments. In another
study, the highest concentrations of PFOA and PFOS in sludge from
25 WWTPs of Shanghai in 2008 were reported with 298 and
173 ng g 1 dw, respectively (Yan et al., 2012). Concentrations of
PFOA and PFOS in sludge from Dalian, Shanghai and Guangzhou
were higher than those in soils (Chen et al., 2012a). These results
indicated that WWTPs might be important sources of PFAAs if
the sludge was used or stacked in soil or other matrices (Li et al.,
2010). Concentrations of PFAAs in sludge from Shanghai were
highest, followed by those from Guangzhou, and the least concentrations were detected in sludge from Dalian. Big differences were
shown in these three studies on PFOA and PFOS concentrations in
sludge from Shanghai. Sludge source of WWTPs was considered as
Table 1
Concentrations of PFOS and PFOA (ng g
a
3.4. PFAAs in snow and rain
Studies on PFAAs in snow or rain mainly focused on the Bohai
Economic Rim. In 2010, Zhao et al. (2012a) made a detailed investigation into PFAAs in precipitation from eastern and central China
(18 snow samples and 1 rain samples), where concentrations of
PFOA and PFOS were 0.7–88.0 and 0.6–15.6 ng L 1, respectively.
Among 19 different sites, concentrations of PFAAs from Weifang,
Shandong province were the highest (more than 150 ng L 1), and
those from some districts of Tianjin and Changchun were also high
(about 80 ng L 1). There might be strong point sources in or around
Weifang City. In a study investigating 17 PFAAs levels in the coastal
rivers of Shandong Province, large scale of PTFE production was
found located in the nearby city Zibo, which put significant influence to the local environment (Wang et al., 2014b). Besides industry sources, domestic applications are also important sources of
PFAAs. As reported by D’eon and Mabury (2011), 80% of fluorotelomer-based commercial products were in the polymeric form that
applied to carpets and textiles, whereas the other 20% were used in
non-polymeric form to produce fluorinated surfactants that
applied to personal care products, leveling and wetting agents,
and non-stick food packaging. The study about PFAAs in precipitation also reported that, according to detected results in Tianjin,
there was no apparent seasonal variation in concentrations of
PFAAs, but there was a trend of decreasing concentrations in successive samples during a single precipitation interval. This result
indicated that precipitation played a significant role in transportation of PFAAs from air to soil and surface water. In addition, other
scattered studies were conducted in Beijing, Shenyang and Dalian.
In snow samples from Beijing, concentrations of PFOA and PFOS
were not high and in the range of nd–2.96 and 0.04–4.76 ng L 1,
respectively. The lower pollution mainly benefited from emigration of polluting enterprises (Wang et al., 2014a). In the city of
dw) in soils and sludge from China.
Media
Location
PFOS
PFOA
Sampling time
Reference
Soil
Tianjin Binhai New Area
North Bohai Sea
Guanting Reservoir
Shanghai
Huaihe Watershed
Haihe Watershed
Shanghai
Shanghai
Coastal citiesb
Tianjin
11 cities
nd–9.4(1.76)
nd–0.7(0.58)
nd–0.86(0.12)
8.58–10.4(9.54)
nd–0.21(0.05)
0.02–2.36(0.19)a
28.1–135(44.25)
27.6–173
0.5–19.8
42–169
0.8–22.5
nd–0.93(0.2)
nd–0.47(0.21)
nd–2.8(0.4)
3.28–47.5(35.25)
nd–0.2(0.08)
nd–0.51(0.19)a
9.21–75.5(35.83)
23.2–298
0.5–158
12–68
0.6–6.7
2008
2008
2008
2007
2008
2008
2008
2010
2011
2009
2009
Wang et al. (2011b)
Wang et al. (2011c)
Wang et al. (2011a)
Li et al. (2010)
Meng et al. (2013)
Pan et al. (2011)
Li et al. (2010)
Yan et al. (2012)
Chen et al. (2012a)
Sun et al. (2011)
Zhang et al. (2013b)
Sludge
b
1
the most important factor to influence levels of PFAAs. Industrial
sludge tends to contain higher PFAAs than domestic sludge (Li
et al., 2010; Yan et al., 2012; Chen et al., 2012a). In addition, wastewater treatment technologies and different processing stages also
contributed to the variability of PFAAs in sludge. In 2009, sludge
was collected from 28 WWTPs in eleven economically developed
cities (Zhang et al., 2013b). Concentrations of PFOA and PFOS were
0.6–6.7 and 0.8–22.5 ng g 1 dw, respectively, which were not high
compared with the previous studies. Therein, 26 WWTPs just
received domestic wastewater, and only 2 WWTPs received both
domestic and industrial wastewater, which might be the reason
why concentrations of PFAAs in these samples were lower than
those from other studies in China. In general, PFAAs in sludge from
WWTPs are mainly affected by types of influents, such as different
kinds of industrial wastewater or domestic wastewater.
Median value.
Including Shanghai, Guangzhou and Dalian.
94
T. Wang et al. / Chemosphere 129 (2015) 87–99
Shenyang with heavy industry, concentrations of PFOA and PFOS in
snow in 2006 were 1.6–22.4 and 0.4–46.2 ng L 1, respectively (Liu
et al., 2007), while those collected in 2007 were 0.82–13 and nd–
51 ng L 1, respectively (Liu et al., 2009b). Results from the two
studies did not show big difference. Another study reported the
relatively higher concentrations of PFOA (8.08–65.8 ng L 1) and
PFOS (26.9–545 ng L 1) of precipitation from Dalian in 2006 (Liu
et al., 2009c). This result is unexpected since Dalian is a less industrialized city near the coast. Apart from local application of PFAAs,
the coastal climate is considered to be a main reason for the relatively high concentrations, including the evolution of regional seabreeze circulation and exchange of PFAAs at air–sea interface (Tong
et al., 2005; Ju et al., 2008). Sea-land breeze may cause PFAAs
unable to disperse, therefore accumulate within the city, and many
PFAAs acting as surfactants may also accumulate on the air/water
interface. Some factors can promote sea-to-air exchange of PFAAs
with low volatility, such as bursting process of the bubbles generated by breaking waves (Saint-Louis and Pelletier, 2004). And significantly higher concentrations of PFOA in aerosols compared
with those in the water illustrated the transportation from water
into atmosphere (Mcmurdo et al., 2008). More studies need to be
conducted to compare concentrations in air from inland and
coastal regions.
4. Risk assessment
4.1. Risk of PFAAs on biota
4.1.1. Occurence of PFAAs in biota
PFOS is the predominant PFAA in zooplankton and blood serum
of fish from a lake receiving wastewater with mean concentrations
ranging from 5.7 to 64.2 ng mL 1, and biomagnification of PFOS in
the food chain was also observed (Li et al., 2008). For large wild
fish, such as the Chinese Sturgeon, PFOS was the dominant PFAA
in eggs, while longer chain C11 to C14 and C16 PFCAs were more
accumulated in adult than PFOS (Peng et al., 2010). Age and gender
effects were found for certain PFAAs. For example, concentrations
of PFOS and Perfluoro-undecanoic acid (PFUdA) were significantly
higher in male alligators than those in females (Wang et al.,
2013b).
Concentrations of PFOS were higher in piscivorous waterbirds
than those in the fishes in their diet. Concentrations in eggs of
waterbirds were in the range of 14.4–343 ng g 1 dw (Wang et al.,
2008). Even in the remote region of the Qinghai-Tibetan Plateau,
which is one of the least urbanized and industrialized regions in
China, PFOS was detected as the dominant PFAA with mean concentrations ranging from 0.2 to 5.2 ng g 1 dw in fish muscle (Shi
et al., 2010). These results indicated widespread distribution of
PFAAs from direct emission and through long range transport of
volatile precursors with subsequent degradation (Stock et al.,
2007).
4.1.2. Risk criteria for protection of aquatic ecosystems
Recently, criteria for protection of aquatic organisms in China
have been derived for PFOS and PFOA. Data were obtained from
toxicity tests, including chronic toxicity tests using three species
and acute toxicity tests using nine aquatic species, along with
the toxicity information from other studies in China. The criteria
maximum concentration (CMC) and criteria continuous concentration (CCC) for protection of aquatic organisms, were calculated to
be 3.78 and 0.25 mg L 1 for PFOS, respectively. While for PFOA,
the CMC value was 45.54 mg L 1 and the CCC value was
3.52 mg L 1 (Yang et al., 2014). The CMC and CCC values for both
PFOS and PFOA in China were significantly higher than values
derived in North America, especially for PFOS, where CMC and
CCC values were 21 and 5.1 lg L 1 for PFOS, 25 and 2.9 mg L 1
for PFOA, respectively (Giesy et al., 2010). Given the method used
in both studies was the same, the main reason for such significant
differences might come from the different native species, which
showed different toxic responses to PFOS and PFOA. Although the
CMC and CCC values derived in China for PFOS and PFOA are higher
than those derived in North America, the actual levels are opposite
in the aquatic environments. Historically, discharge from 3 M
plants led to high level of PFOS in the aquatic environment. For
example, a cave received 3 M plant wastewater contained PFOS
up to 18.2 lg L 1 which was higher than the CCC value and even
close to the CMC value derived for North America (Oliaei et al.,
2013). With strict control over PFOS production and emission,
the decreasing trend has already been observed in the environment of the developed countries (Myers et al., 2012; Olsen et al.,
2012). While in China, increasing production capacity of fluorinated surfactants and polymers has brought a great amount of
PFAAs into the environment (Bao et al., 2011; Wang et al.,
2014b; Zhou et al., 2013). Even though none of the reported data
has failed the criteria derived for native Chinese species, there is
still much need for monitoring the fluorinated polymer industry
to make sure that the emission will not pose adverse effect on local
aquatic ecosystem in the future.
4.2. Risks of exposure to PFAAs on health of humans
4.2.1. Exposure of humans to PFOS and PFOA
It has been found that many PFAAs, like PFOS and PFOA, mainly
bind to proteins other than lipids (Li et al., 2013). Until now, blood,
usually in whole blood or serum, has been used in biomonitoring of
human exposure to PFAAs. In order to make a comparison, concentrations in whole blood and plasma are converted to those in
serum by multiplying a factor of 1 and 2, respectively (Ehresman
et al., 2007). PFAAs were widely detected in the general population
of Chinese people (Zhao et al., 2012b). The spatial distribution of
PFAAs concentrations in blood indicated that people living in eastern cities exhibited much higher concentrations and frequency of
detection of PFAAs than people living in western cities (Liu et al.,
2012). As one of the most developed coastal areas in China, the
Bohai-Rim Economic Circle (BREC) is the most frequently studied
area, with concentrated fluoropolymers production, textile, paper
making and electroplating industries. Generally, PFOS was the
dominant PFAA with the most detection frequency, the highest
geometric mean (GM) concentrations and the highest maximum
concentrations (Liu et al., 2009a; Guo et al., 2011. Shenyang City
and Shijiazhuang City exhibited significantly higher levels of PFOS
than other cities investigated with GM concentrations of 56.3 and
34.0 ng mL 1, respectively, due to higher emissions (Yeung et al.,
2008; Pan et al., 2010). This observation is consistent with the fact
that there is still a large amount of PFOS production and emission
in BREC, mostly from metal plating industry (Xie et al., 2013b).
Concentrations of PFOA were less than those of PFOS in most cities
studied, such as Beijing, Jinzhou, Shenyang, Yingkou, Dalian, Huludao, Shijiazhuang, Qingdao, Tangshan and Weihai City, except for
Fuxin City and Zouping City. The mean concentration of PFOA in
serum of residents in Fuxin was 7.6 ng mL 1, with a maximum of
15 ng mL 1 (Bao et al., 2011). Two fluorine industry parks in the
city of Fuxin have been built recently and emission of PFOA has
posed a significant impact to local residents and the environment
(Bao et al., 2011; Wang et al., 2013c). Concentrations of PFAAs in
blood of humans varied among gender and age (Fu et al., 2014).
Positive correlations were often observed between age and concentrations of PFOS and PFOA in serum. Generally PFAAs concentrations in blood of males were higher than those of females, the
difference was small for most PFAAs except for PFOA, and females
exhibited higher concentrations on PFOA serum than males in
95
T. Wang et al. / Chemosphere 129 (2015) 87–99
children less than 15 years of age (Liu et al., 2009a; Bao et al., 2011;
Guo et al., 2011; Zhang et al., 2010b). However, inconsistency with
the influence of age and gender did exist in studies by different
researchers. For example, PFAAs levels in blood except for PFOA
in senior people were significantly higher than those in low age
people in the population of Liaoning Province, China; whereas in
the National Health and Nutrition Examination Survey of the United States (U.S. NHANES), no age effects were observed (Liu et al.,
2009a). The reason for the inconsistency might come from multiple
factors, including lifestyle, occupation, diet and place of habitation
(Liu et al., 2009a). Isomers of PFOS and PFOA were also identified in
human serum in Shijiazhuang City and Handan City of north China.
A significant difference was observed between profiles of individual isomers of PFOS and PFOA. On average, linear PFOS accounted
for 48% of the total PFOS, while linear PFOA accounted for as much
as 96% of the total PFOA (Zhang et al., 2013c). The two main processes used to create compounds containing perfluoroalkyl chains
produce different isomers. Telomerization produces 100% linear
isomer, while electrochemical fluorination produces approximately 70–80% linear and 20–30% branched isomers (Paul et al.,
2009). The less proportion of linear PFOS in serum might be used
as a biomaker of exposure to PFOS-precursors (Martin et al.,
2010). The temporal trend of PFAAs concentrations in three cities
of China implied that concentrations of PFOS in serum have
decreased, while concentrations of PFOA have increased in recent
years. Mean concentration of PFOS in blood of residents in Shenyang City decreased from 112.6 in 2004 to 14.38 ng mL 1 in 2007,
while the mean concentration of PFOA increased from nd to
1.96 ng mL 1 (Liu et al., 2009a; Yeung et al., 2008). Similar trends
were also observed in residents of Guiyang City and Shijiazhuang
City (Pan et al., 2010; Zhang et al., 2013c). The decreasing trend
of PFOS observed in China was consistent with those in developed
countries, which was due to the phase-out of PFOS and related
chemicals since 2000. However, there might be time-lag on the
trend in China as production shift from developed countries to
China might take years. For PFOA, the increasing trend might be
due to the increasing demand for fluoropolymers with PFOA as
processing addictive, which has happened in BREC (Wang et al.,
2014b; Xie et al., 2013b).
In recent years, researchers attempted to explore non-invasive
sampling for biomonitoring of PFAAs in humans due to the disadvantages of collecting blood, such as its invasiveness to participants, difficulty in storage and relatively high cost. Almost all
PFAAs except for PFOA had similar fingernail/toenail levels or
higher fingernail levels than toenail levels. Concentration of PFOS
in fingernail was significantly correlated with concentration of
PFOS in serum (Liu et al., 2011b). Another study showed that concentrations of PFOS in nail, hair and urine were significantly correlated with those in serum. A similar relationship was observed
between concentrations of PFOA in nail and serum (Li et al.,
2013). The relative proportions of PFOS and PFOA levels were also
similar between nail (21.2% of PFOA and 78.8% of PFOS) and serum
(13.4% PFOA and 86.6% of PFOS). Therefore, nail was suggested as
the preferred non-invasive biomarker for monitoring human body
loadings of PFAAs (Li et al., 2013). Urine was used for estimation of
renal clearance and considered as the major elimination route for
short-chain PFAAs (Zhang et al., 2013c) (Fig. 6).
4.2.2. Exposure pathway and risk assessment
For the general Chinese population, human exposure to PFAAs is
evaluated based on food consumption and drinking water. Levels
of PFAAs in various foodstuffs were analyzed, including poultry,
livestock and seafood (Wang et al., 2010a; Lu et al., 2011; Chen
et al., 2011c; Zhang et al., 2010a). Mean concentrations of PFOS
and PFOA in foodstuffs collected from 17 cities of 15 provinces in
China were 0.05–1.99 and 0.06–12.5 ng g 1 fresh weight (fwt),
respectively (Zhang et al., 2010a). Fish usually contained higher
concentrations of PFOS and lower concentrations of PFOA compared with the poultry (Lu et al., 2011), which is probably due to
the difference in bioconcentration in aquatic environment or different emission profiles. In different edible parts of poultry and
livestock, the total PFAAs accumulated in a decreasing order as follows: liver > kidney > meat (Wang et al., 2010a). Concentrations of
PFAAs increased with tropic level, especially in seafood (Chen et al.,
2011c). For Chinese adults, the mean total daily intake (TDI) of
PFOS and PFOA were calculated to be 1.19 and 9.83 ng kg 1 bw d 1,
respectively (Zhang et al., 2010a). The TDI was calculated through
multiplication of mean PFOS or PFOA concentrations in certain
foodstuff, drinking water and dust inhalation. The foodstuff
included meat, animal liver, animal blood, eggs, fish and seafood,
which was based on the dietary pattern and nutrition survey in
China. Seafood was the main dietary source of PFOS, accounting
for 78.9% of the TDI, while meat was the main contributor to dietary exposure of PFOA, accounting for 93.2% of the TDI (Zhang
et al., 2010a). Drinking water accounted for as much as 13% of
the TDI of PFOA for residents living near fluorochemical plants
(Bao et al., 2011). Although exposure from indoor dust might contribute a small portion to the TDI of PFOS and PFOA for the nonoccupational population (Zhang et al., 2010a), it might be a
significant source of exposure to PFOS and PFOA or other PFAAs
Fig. 6. Major exposure pathway and biomonitoring of PFAAs for humans in China.
1, 2, 4
Zhang et al., 2010a;
3
Bao et al., 2010.
96
T. Wang et al. / Chemosphere 129 (2015) 87–99
under some extreme conditions, for example, for the workers in
fluoropolymer facilities (Wang et al., 2010b).
For fetuses and infants, two exposure pathways of PFAAs,
including maternal-fetal transmission and breast feeding transmission, have attracted intensive attention in recent years (So et al.,
2006b; Tao et al., 2008; Liu et al., 2010; Glynn et al., 2012; Zhang
et al., 2013a,d; Nøst et al., 2014). For fetuses, significant positive
correlations have been found between concentrations of PFAAs in
paired samples of maternal whole blood and placenta, and
between placenta and cord blood (Zhang et al., 2013a) (Fig. 6).
Effects of carbon chain length on efficiency of transfer of PFAAs
from the female to fetus were also evaluated, and a U-shaped trend
with increasing carbon chain length was found for C7–C12 PFCAs.
The trend could be explained by the opposite ratios between the
transfer efficiencies of individual PFAA (C6–C12) from maternal
blood to placenta and those from placenta to cord blood (Zhang
et al., 2013a). For newborns, a significant correlation was observed
among concentrations of PFAAs in matched maternal serum, cord
serum and breast milk. Mean concentrations of PFOS and PFOA
were 3.2 and 1.66 ng mL 1 in maternal serum, 1.79 and
1.5 ng mL 1 in cord serum, 0.06 and 0.18 ng mL 1 in breast milk,
respectively. So the efficiency of transport through both placental
barrier and lactation was relatively higher for PFOA than PFOS
(Liu et al., 2011a). Concentration of PFOA in breast milk was usually observed at least 10-fold less than that in matched maternal
serum (Liu et al., 2010, 2011a).
The measured TDI values of PFOS and PFOA for adult or infant
were generally far below the criteria values derived by several governmental agencies (So et al., 2006a; Alexander et al., 2008; Roos
et al., 2008; Fromme et al., 2009) (Table S3). However, the highest
TDI of PFOS (30 ng kg 1 d 1) observed in Zhoushan exceeded the
reference dose of 25 ng kg 1 d 1 derived by the environmental
working group of U.S. (So et al., 2006b), and the highest TDI of
PFOA in Shanghai (88.4 ng kg 1 d 1) was close to the criteria value
of 100 ng kg 1 d 1 derived by the German Federal Institute for Risk
Assessment and the Drinking Water Commission (Liu et al., 2010).
These results suggested that there might be a potential risk of PFOS
or PFOA for infants via breast-breeding and also adults via various
pathways like diet or drinking water in both cities. TDIs of PFOS for
infant (usually used as EDIs, which stand for estimated dietary
intakes for infant) in China were less than or comparable to those
from other Asian countries (Tao et al., 2008), but higher than certain western countries like Canada, Germany and Spain (Liu
et al., 2010). Mean TDIs of PFOA for both adult and infant in China
were higher than those from other countries (Tao et al., 2008; Liu
et al., 2010), which implied that Chinese population might be
exposed to higher PFOA concentrations through daily intake
(Table S3).
5. Conclusions and perspectives
PFAAs are released to the environment from a number of point
sources associated with industrial processes in manufacturing or
using PFAAs-related chemicals, or via diffuse widespread consumer use and disposal of PFAAs-containing products. Due to their
significant lower vapor pressure and higher water-solubility,
PFAAs accumulate and disperse in aquatic environment, so studies
on their adverse effects and risks mostly focus on aquatic ecosystems. In general, concentrations of PFAAs especially PFOS and PFOA
in aquatic organisms collected near cities in most coastal regions
are much higher, due to rapid industrialization and urbanization
in recent years, than those in more remote areas in China. This
observation indicated that the emission and pollution of PFAAs
as well as their distribution are closely related to regional urbanization processes. PFOS, PFOA and related products have been
widely used in industrial and commercial areas over the past several decades, with relatively high concentrations detected in environmental matrixes, biota and even local residents. Studies on
PFAAs in China have lagged behind those in more developed countries. Therefore, the foundation for PFAAs studies is frail, especially
in risk assessment. Risk assessment of PFAAs is still less developed
in China, and most studies have just compared concentrations of
PFAAs with the guideline values derived for certain species or
receptors to evaluate the risks. Although there seems to be little
risk of adverse effects of PFAAs on aquatic environments, risks to
organisms at high trophic levels such as piscivorous birds and
mammals may be great due to trophic biomagnification. However,
studies on biomagnification of PFAAs through food chain is still
limited in China.
PFOS and its salts have been included in the Stockholm Convention, but exemption allows its continuous production and use in
China. China has become one of the largest production and consumption countries to meet the growing demand for surfactants
and other surface modification applications. Excitingly, the government and industries have devoted to developing alternatives to
replace PFOS, PFOA and related salts, including PFBS and PFBA
(Holt, 2011). Based on the previous studies on PFAAs in China,
more strategies and legislation for management of PFAAs need to
be further developed and strengthened. A specific agency should
be designated to investigate these pollutant residues in environment matrixes and different trophic levels on a regular basis. With
rapid industrialization and urbanization in China, emission, transport and fate of PFAAs in environment will still be a great concern
for ecosystem and human health. However, data gaps on the identification of various sources, industrial and domestic emission, spatial and temporal distribution of typical PFAAs from sources to the
environment, toxic effect of certain PFAAs to typical aquatic organisms still exist, which should be taken as priority in the future plan
for national actions.
Acknowledgements
This work supported by the National Natural Science Foundation of China (Nos. 41171394 and 41371488), the Key Research
Program of the Chinese Academy of Sciences (No. KZZD-EW-TZ12), and the National Fundamental Field Study Program (No.
2013FY11110). Prof. Giesy was supported by the Canada Research
Chair Program and the Einstein Professor Program of the Chinese
Academy of Sciences.
Appendix A. Supplementary material
Supplementary data associated with this article can be found, in
the online version, at http://dx.doi.org/10.1016/j.chemosphere.
2014.09.021.
References
Ahrens, L., Felizeter, S., Sturm, R., Xie, Z.Y., Ebinghaus, R., 2009. Polyfluorinated
compounds in waste water treatment plant effluents and surface waters along
the River Elbe, Germany. Mar. Pollut. Bull. 58 (9), 1326–1333.
Alexander, J., Auðunsson, G.A., Benford, D., Cockburn, A., Cravedi, J.P., Dogliotti, E.,
Domenico, A.D., Fernández-Cruz, M.L., Fink-Gremmels, J., Fürst, P., Galli, C.,
Grandjean, P., Gzyl, J., Heinemeyer, G., Johansson, N., Mutti, A., Schlatter, J.,
Leeuwen, R.V., Peteghem, C.V., Verger, P., 2008. Perfluorooctane sulfonate
(PFOS), perfluorooctanoic acid (PFOA) and their salts. EFSA J. 653, 1–131.
Armitage, J., Cousins, I.T., Buck, R.C., Prevedouros, K., Russell, M.H., MacLeod, M.,
Korzeniowski, S.H., 2006. Modeling global-scale fate and transport of
perfluorooctanoate emitted from direct sources. Environ. Sci. Technol. 40 (22),
6969–6975.
Armitage, J.M., MacLeod, M., Cousins, I.T., 2009. Modeling the global fate and
transport of perfluorooctanoic acid (PFOA) and perfluorooctanoate (PFO)
emitted from direct sources using a multispecies mass balance model.
Environ. Sci. Technol. 43 (4), 1134–1140.
T. Wang et al. / Chemosphere 129 (2015) 87–99
Bao, J., Liu, W., Liu, L., Jin, Y.H., Ran, X.R., Zhang, Z.X., 2010. Perfluorinated
compounds in urban river sediments from Guangzhou and Shanghai of China.
Chemosphere 80 (2), 123–130.
Bao, J., Liu, W., Liu, L., Jin, Y.H., Dai, J.Y., Ran, X.R., Zhang, Z.X., Tsuda, S., 2011.
Perfluorinated compounds in the environment and the blood of residents living
near fluorochemical plants in Fuxin, China. Environ. Sci. Technol. 45 (19), 8075–
8080.
Becker, A.M., Gerstmann, S., Frank, H., 2008. Perfluorooctanoic acid and
perfluorooctane sulfonate in the sediment of the Roter Main River, Bayreuth,
Germany. Environ. Pollut. 156 (3), 818–820.
Blaine, A., Rich, C., Hundal, L., Lau, C.S., Mills, M., Harris, K., Higgins, C.P., 2013.
Uptake of perfluoroalkyl acids into edible crops via land applied biosolids: field
and greenhouse studies. Environ. Sci. Technol. 47 (24), 14062–14069.
Boulanger, B., Vargo, J., Schnoor, J.L., Hornbuckle, K.C., 2004. Detection of
perfluorooctane surfactants in Great Lakes water. Environ. Sci. Technol. 38
(15), 4064–4070.
Brook, D., Footitt, A., Nwaogu, T.A., 2004. Environmental risk evaluation report:
perfluorooctane sulphonate (PFOS). In: Environment Agency (Ed.). Building
Research Establishment Ltd. & Risk and Policy Analysts Ltd.
Cai, F., 2009. Current situation of fluorochemicals and development proposal in our
country. Chem. Eng. 7, 59–61 (in Chinese).
Cai, M.H., Zhao, Z., Yang, H.Z., Yin, Z.G., Hong, Q.Q., Sturm, R., Ebinghaus, R., Ahrens,
L., Cai, M.G., He, J.F., 2012. Spatial distribution of per- and polyfluoroalkyl
compounds in coastal waters from the East to South China Sea. Environ. Pollut.
161, 162–169.
Chen, C.L., Lu, Y.L., Zhang, X., Geng, J., Wang, T.Y., Shi, Y.J., Hu, W.Y., Li, J., 2009. A
review of spatial and temporal assessment of PFOS and PFOA contamination in
China. Chem. Ecol. 25 (3), 163–177.
Chen, C.L., Wang, T.Y., Khim, J.S., Luo, W., Jiao, W.T., Lu, Y.L., Naile, J.E., Hu, W.Y.,
Zhang, X., Geng, J., Bi, C.C., Li, J., Giesy, J.P., 2011a. Perfluorinated compounds in
water and sediment from coastal regions of the northern Bohai Sea, China.
Chem. Ecol. 27 (2), 165–176.
Chen, C.L., Wang, T.Y., Lu, Y.L., Luo, W., Geng, J., 2011b. Estimation of perfluorinated
compounds emissions from major rivers and wastewater treatment plants in
China. Chin. J. Environ. Sci. 32 (4), 1073–1080 (in Chinese).
Chen, C.L., Wang, T.Y., Naile, J.E., Li, J., Geng, J., Bi, C.C., Hu, W.Y., Zhang, X., Khim, J.S.,
Feng, Y., Giesy, J.P., Lu, Y.L., 2011c. Perfluorinated compounds in aquatic
products from Bohai Bay, Tianjin, China. Hum. Ecol. Risk Assess. 17 (6), 1279–
1291.
Chen, H., Zhang, C., Han, J.B., Yu, Y.X., Zhang, P., 2012a. PFOS and PFOA in influents,
effluents, and biosolids of Chinese wastewater treatment plants and effluentreceiving marine environments. Environ. Pollut. 170, 26–31.
Chen, J., Wang, L.L., Zhu, H.D., Wang, B.B., Liu, H.C., Cao, M.H., Miao, Z., Hu, L., Lu,
X.H., Liu, G.H., 2012b. Spatial distribution of perfluorooctanoic acids and
perfluorinate sulphonates in surface water of East Lake. Environ. Sci. 33 (8),
2586–2591 (in Chinese).
Conder, J.M., Hoke, R.A., De Wolf, W., Russell, M.H., Buck, R.C., 2008. Are PFCAs
bioaccumulative? A critical review and comparison with regulatory lipophilic
compounds. Environ. Sci. Technol. 42 (4), 995–1003.
D’eon, J.C., Mabury, S.A., 2011. Is indirect exposure a significant contributor to the
burden of perfluorinated acids observed in humans? Environ. Sci. Technol. 45
(19), 7974–7984.
Ehresman, D.J., Froehlich, J.W., Olsen, G.W., Chang, S.C., Butenhoff, J.L., 2007.
Comparison of human whole blood, plasma, and serum matrices for the
determination of perfluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA),
and other fluorochemicals. Environ. Res. 103 (2), 176–184.
Fang, X., 2004. Current state and trend of fluoropolymer. Silicone Fluorine Inf. 3, 15–
19 (in Chinese).
Footitt, A., Nwaogu, T.A., Brook, D., 2004. Defra Risk Reduction Strategy and Analysis
of Advantages and Drawbacks for PFOS. J454/PFOS RRS.
Fromme, H., Tittlemier, S.A., Volkel, W., Wilhelm, M., Twardella, D., 2009.
Perfluorinated compounds – exposure assessment for the general population
in western countries. Int. J. Hyg. Environ. Health 212 (3), 239–270.
Fu, Y.N., Wang, T.Y., Wang, P., Fu, Q.L., Lu, Y.L., 2014. Effects of age, gender and
region on serum concentrations of perfluorinated compounds in general
population of Henan, China. Chemosphere 129, 104–110.
Fujii, S., Polprasert, C., Tanaka, S., Lien, N.P.H., Qiu, Y., 2007. New POPs in the water
environment: distribution, bioaccumulation and treatment of perfluorinated
compounds – a review paper. J. Water Supply Res. Technol.-Aqua. 56 (5), 313–
326.
Giesy, J.P., Kannan, K., 2001. Global distribution of perfluorooctane sulfonate in
wildlife. Environ. Sci. Technol. 35 (7), 1339–1342.
Giesy, J.P., Kannan, K., 2002. Perfluorochemical surfactants in the environment.
Environ. Sci. Technol. 36 (7), 146A–152A.
Giesy, J.P., Naile, J.E., Khim, J.S., Jones, P.D., Newsted, J.L., 2010. Aquatic
toxicology of perfluorinated chemicals. In: Whitacre, D.M. (Ed.), Reviews of
Environmental Contamination and Toxicology, vol. 202. Springer, New York,
pp. 1–52.
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 1996–2010.
Environ. Sci. Technol. 46 (16), 9071–9079.
Guo, F.F., Zhong, Y.X., Wang, Y.X., Li, J.G., Zhang, J.L., Liu, J.Y., Zhao, Y.F., Wu, Y.N.,
2011. Perfluorinated compounds in human blood around Bohai Sea, China.
Chemosphere 85 (2), 156–162.
97
Higgins, C.P., Field, J.A., Criddle, C.S., Luthy, R.G., 2005. Quantitative determination of
perfluorochemicals in sediments and domestic sludge. Environ. Sci. Technol. 39
(11), 3946–3956.
Holt, R., 2011. The Influence of Global Regulatory Changes and Customer
Preferences on the Development of Alternatives to Long Chain Fluorinated
Chemicals. DuPont for FluoroCouncil.
Houde, M., Balmer, B.C., Brandsma, S., Wells, R.S., Rowles, T.K., Solomon, K.R., Muir,
D.C.G., 2006a. Perfluoroalkyl compounds in relation to life-history and
reproductive parameters in bottlenose dolphins (Tursiops truncatus) from
Sarasota Bay, Florida, USA. Environ. Toxicol. Chem. 25 (9), 2405–2412.
Houde, M., Bujas, T.A.D., Small, J., Wells, R.S., Fair, P.A., Bossart, G.D., Solomon, K.R.,
Muir, D.C.G., 2006b. Biomagnification of perfluoroalkyl compounds in the
bottlenose dolphin (Tursiops truncatus) food web. Environ. Sci. Technol. 40
(13), 4138–4144.
Houde, M., De Silva, A.O., Muir, D.C., Letcher, R.J., 2011. Monitoring of perfluorinated
compounds in aquatic biota: an updated review. Environ. Sci. Technol. 45 (19),
7962–7973.
Jin, Y.H., Liu, X., Zhang, X., Li, T., Zhang, Y.H., Saitou, N., Sasaki, K., Koizumi, A., 2003.
Perfluorooctance sulfonate situation in serum of common population. Chin. J.
Public Health 19 (10), 1200–1201 (in Chinese).
Jin, Y.H., Liu, W., Sato, I., Nakayama, S.F., Sasaki, K., Saito, N., Tsuda, S., 2009. PFOS
and PFOA in environmental and tap water in China. Chemosphere 77 (5), 605–
611.
Ju, X.D., Jin, Y.H., Sasaki, K., Saito, N., 2008. Perfluorinated surfactants in surface,
subsurface water and microlayer from Dalian coastal waters in China. Environ.
Sci. Technol. 42 (10), 3538–3542.
Kälin, T., Will, R., Yamaguchi, Y., 2012. CEH Marketing Research Report:
fluoropolymers. IHS Chemical. <http://www.ihs.com/products/chemical/
planning/ceh/fluoropolymers>.
Kannan, K., Corsolini, S., Falandysz, J., Oehme, G., Focardi, S., Giesy, J.P., 2002.
Perfluorooctanesulfonate and related fluorinated hydrocarbons in marine
mammals, fishes, and birds from coasts of the Baltic and the Mediterranean
Seas. Environ. Sci. Technol. 36 (15), 3210–3216.
Kim, S.K., 2012. Watershed-based riverine discharge loads and emission factor of
perfluorinated surfactants in Korean peninsula. Chemosphere 89, 995–1002.
Kissa, E., 2001. Fluorinated Surfactants and Repellents, 2nd ed. Marcel Dekker Inc.
Labadie, P., Chevreuil, M., 2011. Biogeochemical dynamics of perfluorinated alkyl
acids and sulfonates in the River Seine (Paris, France) under contrasting
hydrological conditions. Environ. Pollut. 159 (12), 3634–3639.
Lau, C., Anitole, K., Hodes, C., Lai, D., Pfahles-Hutchens, A., Seed, J., 2007.
Perfluoroalkyl acids: a review of monitoring and toxicological findings.
Toxicol. Sci. 99 (2), 366–394.
Li, X.M., Yeung, L.W.Y., Xu, M.Q., Taniyasu, S., Lam, P.K.S., Yamashita, N., Dai, J.Y.,
2008. Perfluorooctane sulfonate (PFOS) and other fluorochemicals in fish blood
collected near the outfall of wastewater treatment plant (WWTP) in Beijing.
Environ. Pollut. 156 (3), 1298–1303.
Li, F., Zhang, C.J., Qu, Y., Chen, J., Chen, L., Liu, Y., Zhou, Q., 2010. Quantitative
characterization of short- and long-chain perfluorinated acids in solid matrices
in Shanghai, China. Sci. Total Environ. 408 (3), 617–623.
Li, F.S., Sun, H.W., Hao, Z.N., He, N., Zhao, L.J., Zhang, T., Sun, T.H., 2011.
Perfluorinated compounds in Haihe River and Dagu Drainage Canal in Tianjin,
China. Chemosphere 84 (2), 265–271.
Li, J.G., Guo, F.F., Wang, Y.X., Zhang, J.L., Zhong, Y.X., Zhao, Y.F., Wu, Y.N., 2013. Can
nail, hair and urine be used for biomonitoring of human exposure
to perfluorooctane sulfonate and perfluorooctanoic acid? Environ. Int. 53,
47–52.
Liu, J., Avendano, S.M., 2013. Microbial degradation of polyfluoroalkyl chemicals in
the environment: a review. Environ. Int. 61, 98–114.
Liu, W., Jin, Y.H., Quan, X., Dong, G.H., Liu, B., Wang, J., Wang, K., Yu, Q.L., Saito, N.,
2007. Investigation of PFOS and PFOA pollution in snow in Shenyang, China.
Environ. Sci. 28, 2068–2073 (in Chinese).
Liu, C., Hu, J.X., Liu, J.G., 2008a. Preliminary risk assessment of semiconductor
manufacturing PFOS emissions and near-site environmental concentrations. Sci.
Technol. Eng. 8 (11), 2898–2902 (in Chinese).
Liu, C., Hu, J.X., Liu, J.G., Tong, X.C., 2008b. Environmental risk assessment of
perfluorooctane sulphonate near chromium plating site. China Environ. Sci. 28
(10), 950–954 (in Chinese).
Liu, C., Hu, J.X., Liu, J.G., Wan, D., 2008c. Pollution status and release of
perfluorooctane sulfonate (PFOS) and risk analysis for PFOS in China. Environ.
Pollut. Control 7, 1–9 (in Chinese).
Liu, J.Y., Li, J.G., Luan, Y., Zhao, Y.F., Wu, Y.N., 2009a. Geographical distribution of
perfluorinated compounds in human blood from Liaoning Province, China.
Environ. Sci. Technol. 43 (11), 4044–4048.
Liu, W., Dong, G.H., Jin, Y.H., Sasaki, K., Saito, N., Sato, I., Tsuda, S., Nakayama, S.F.,
2009b. Occurrence of perfluoroalkyl acids in precipitation from Shenyang,
China. Chin. Sci. Bull. 54 (14), 2440–2445.
Liu, W., Jin, Y.H., Quan, X., Sasaki, K., Saito, N., Nakayama, S.F., Sato, I., Tsuda, S.,
2009c. Perfluorosulfonates and perfluorocarboxylates in snow and rain in
Dalian, China. Environ. Int. 35 (4), 737–742.
Liu, J.Y., Li, J.G., Zhao, Y.F., Wang, Y.X., Zhang, L., Wu, Y.N., 2010. The occurrence of
perfluorinated alkyl compounds in human milk from different regions of China.
Environ. Int. 36 (5), 433–438.
Liu, J.Y., Li, J.G., Liu, Y., Chan, H.M., Zhao, Y.F., Cai, Z.W., Wu, Y.N., 2011a. Comparison
on gestation and lactation exposure of perfluorinated compounds for newborns.
Environ. Int. 37 (7), 1206–1212.
98
T. Wang et al. / Chemosphere 129 (2015) 87–99
Liu, W., Xu, L., Li, X., Jin, Y.H., Sasaki, K., Saito, N., Sato, I., Tsuda, S., 2011b. Human
nails analysis as biomarker of exposure to perfluoroalkyl compounds. Environ.
Sci. Technol. 45 (19), 8144–8150.
Liu, W., Dong, G.H., Luo, Y.W., Liu, L., Cao, Z.W., Li, X.N., Jin, Y.H., 2012. Estimation of
reference values for PFOS and PFOA in human biomonitoring and relevance of
exposure among family members in China. J. Environ. Prot. 03 (04), 353–361.
Loi, E.I.H., Yeung, L.W.Y., Taniyasu, S., Lam, P.K.S., Kannan, K., Yamashita, N., 2011.
Trophic magnification of poly- and perfluorinated compounds in a subtropical
food web. Environ. Sci. Technol. 45 (13), 5506–5513.
Lu, G.H., Yang, Y.L., Taniyasu, S., Yeung, L.W.Y., Pan, J., Zhou, B.S., Lam, P.K.S.,
Yamashita, N., 2011. Potential exposure of perfluorinated compounds to
Chinese in Shenyang and Yangtze River Delta areas. Environ. Chem. 8 (4),
407–418.
Lu, H.X., Cai, Q.Y., Jones, K.C., Zeng, Q.Y., Katsoyiannis, A., 2013. Levels of organic
pollutants in vegetables and human exposure through diet: a review. Crit. Rev.
Environ. Sci. Technol. 44 (1), 1–33.
Martin, J.W., Asher, B.J., Beesoon, S., Benskin, J.P., Ross, M.S., 2010. PFOS or PreFOS?
Are perfluorooctane sulfonate precursors (PreFOS) important determinants of
human and environmental perfluorooctane sulfonate (PFOS) exposure? J.
Environ. Monit. 12 (11), 1979–2004.
Mcmurdo, C.J., Ellis, D.A., Webster, E., Butler, J., Christensen, R.D., Reid, L.K., 2008.
Aerosol enrichment of the surfactant PFO and mediation of the water–air
transport of gaseous PFOA. Environ. Sci. Technol. 42 (11), 3969–3974.
Meng, J., Wang, T.Y., Wang, P., Giesy, J.P., Lu, Y.L., 2013. Perfluorinated compounds
and organochlorine pesticides in soils around Huaihe River: a heavily
contaminated watershed in Central China. Environ. Sci. Pollut. Res. 20 (6),
3965–3974.
Moeller, A., Ahrens, L., Surm, R., Westerveld, J., van der Wielen, F., Ebinghaus, R., de
Voogt, P., 2010. Distribution and sources of polyfluoroalkyl substances (PFAS) in
the River Rhine watershed. Environ. Pollut. 158 (10), 3243–3250.
Murakami, M., Imamura, E., Shinohara, H., Kiri, K., Muramatsu, Y., Harada, A.,
Takada, H., 2008. Occurrence and sources of perfluorinated surfactants in rivers
in Japan. Environ. Sci. Technol. 42 (17), 6566–6572.
Myers, A.L., Crozier, P.W., Helm, P.A., Brimacombe, C., Furdui, V.I., Reiner, E.J.,
Burniston, D., Marvin, C.H., 2012. Fate, distribution, and contrasting temporal
trends of perfluoroalkyl substances (PFAAs) in Lake Ontario, Canada. Environ.
Int. 44, 92–99.
Naile, J.E., Khim, J.S., Wang, T.Y., Chen, C.L., Luo, W., Kwon, B.O., Park, J., Koh, C.H.,
Jones, P.D., Lu, Y.L., 2010. Perfluorinated compounds in water, sediment, soil and
biota from estuarine and coastal areas of Korea. Environ. Pollut. 158 (5), 1237–
1244.
Nøst, T.H., Vestergren, R., Berg, V., Nieboer, E., Odland, J.Ø., Sandanger, T.M., 2014.
Repeated measurements of per- and polyfluoroalkyl substances (PFASs) from
1979 to 2007 in males from Northern Norway: assessing time trends, compound
correlations and relations to age/birth cohort. Environ. Int. 67, 43–53.
Oliaei, F., Kriens, D., Weber, R., Watson, A., 2013. PFOS and PFC releases and
associated pollution from a PFC production plant in Minnesota (USA). Environ.
Sci. Pollut. Res. 20 (4), 1977–1992.
Olsen, G.W., Lange, C.C., Ellefson, M.E., Mair, D.C., Church, T.R., Goldberg, C.L.,
Herron, R.M., Medhdizadekashi, Z., Nobiletti, J.B., Rios, J.A., Reagen, W.K., Zobel,
L.R., 2012. Temporal trends of perfluoroalkyl concentrations in american red
cross adult blood donors, 2000–2010. Environ. Sci. Technol. 46 (11), 6330–6338.
Pan, Y.Y., Shi, Y.L., Wang, J.M., Cai, Y.Q., Wu, Y.N., 2010. Concentrations of
perfluorinated compounds in human blood from twelve cities in China.
Environ. Toxicol. Chem. 29 (12), 2695–2701.
Pan, Y.Y., Shi, Y.L., Wang, J.M., Jin, X.L., Cai, Y.Q., 2011. Pilot Investigation of
perfluorinated compounds in river water, sediment, soil and fish in Tianjin,
China. Bull. Environ. Contam. Toxicol. 87 (2), 152–157.
Paul, A.G., Jones, K.C., Sweetman, A.J., 2009. A first global production, emission, and
environmental inventory for perfluorooctane sulfonate. Environ. Sci. Technol.
43 (2), 386–392.
Peng, H., Wei, Q.W., Wan, Y., Giesy, J.P., Li, L.X., Hu, J.Y., 2010. Tissue distribution and
maternal transfer of poly- and perfluorinated compounds in Chinese Sturgeon
(Acipenser sinensis): implications for reproductive risk. Environ. Sci. Technol. 44
(5), 1868–1874.
Pico, Y., Farre, M., Llorca, M., Barcelo, D., 2011. Perfluorinated compounds in food: a
global perspective. Crit. Rev. Food Sci. Nutr. 51 (7), 605–625.
Pistocchi, A., Loos, R., 2009. A map of European emissions and concentrations of
PFOS and PFOA. Environ. Sci. Technol. 43 (24), 9237–9244.
Prevedouros, K., Cousins, I.T., Buck, R.C., Korzeniowski, S.H., 2006. Sources, fate and
transport of perfluorocarboxylates. Environ. Sci. Technol. 40 (1), 32–44.
Rayne, S., Forest, K., 2009. Perfluoroalkyl sulfonic and carboxylic acids: a critical
review of physicochemical properties, levels and patterns in waters and
wastewaters, and treatment methods. J. Environ. Sci. Health, Part A: Toxic/
Hazard. Subst. Environ. Eng. 44 (12), 1145–1199.
Roos, P.H., Angerer, J., Dieter, H., Wilhelm, M., Wolfle, D., Hengstler, J.G., 2008.
Perfluorinated compounds (PFC) hit the headlines: meeting report on a satellite
symposium of the annual meeting of the German society of toxicology. Arch.
Toxicol. 82 (1), 57–59.
Saint-Louis, R., Pelletier, E., 2004. Sea-to-air flux of contaminants via bubbles
bursting. An experimental approach for tributyltin. Mar. Chem. 84 (3–4), 211–
224.
Senthilkumar, K., Ohi, E., Sajwan, K., Takasuga, T., Kannan, K., 2007. Perfluorinated
compounds in river water, river sediment, market fish, and wildlife samples
from Japan. Bull. Environ. Contam. Toxicol. 79 (4), 427–431.
Shi, Y.L., Pan, Y.Y., Yang, R.Q., Wang, Y.W., Cai, Y.Q., 2010. Occurrence of
perfluorinated compounds in fish from Qinghai-Tibetan Plateau. Environ. Int.
36 (1), 46–50.
Shi, Y.L., Pan, Y.Y., Wang, J.M., Cai, Y.Q., 2012. Distribution of perfluorinated
compounds in water, sediment, biota and floating plants in Baiyangdian Lake,
China. J. Environ. Monit. 14 (2), 636–642.
So, M.K., Taniyasu, S., Lam, P.K.S., Zheng, G.J., Giesy, J.P., Yamashita, N., 2006a.
Alkaline digestion and solid phase extraction method for perfluorinated
compounds in mussels and oysters from south China and Japan. Arch.
Environ. Contam. Toxicol. 50 (2), 240–248.
So, M.K., Yamashita, N., Taniyasu, S., Jiang, Q., Giesy, J.P., Chen, K., Lam, P.K., 2006b.
Health risks in infants associated with exposure to perfluorinated compounds
in human breast milk from Zhoushan, China. Environ. Sci. Technol. 40 (9),
2924–2929.
So, M.K., Miyake, Y., Yeung, W.Y., Ho, Y.M., Taniyasu, S., Rostkowski, P., Yamashita,
N., Zhou, B.S., Shi, X.J., Wang, J.X., Giesy, J.P., Yu, H., Lam, P.K.S., 2007.
Perfluorinated compounds in the Pearl River and Yangtze River of China.
Chemosphere 68 (11), 2085–2095.
Squadrone, S., Ciccotelli, V., Favaro, L., Scanzio, T., Prearo, M., Abete, M.C., 2014. Fish
consumption as a source of human exposure to perfluorinated alkyl substances
in Italy: analysis of two edible fish from Lake Maggiore. Chemosphere 114, 181–
186.
Stock, N.L., Furdui, V.I., Muir, D.C.G., Mabury, S.A., 2007. Perfluoroalkyl
contaminants in the Canadian Arctic: evidence of atmospheric transport and
local contamination. Environ. Sci. Technol. 41 (10), 3529–3536.
Strynar, M.J., Lindstrom, A.B., Nakayama, S.F., Egeghy, P.P., Helfant, L.J., 2012. Pilot
scale application of a method for the analysis of perfluorinated compounds in
surface soils. Chemosphere 86 (3), 252–257.
Suja, F., Pramanik, B.K., Zain, S.M., 2009. Contamination, bioaccumulation and toxic
effects of perfluorinated chemicals (PFCs) in the water environment: a review
paper. Water Sci. Technol. 60 (6), 1533–1544.
Sun, H.W., Li, F.S., Zhang, T., Zhang, X.Z., He, N., Song, Q., Zhao, L.J., Sun, L.N., Sun,
T.H., 2011. Perfluorinated compounds in surface waters and WWTPs in
Shenyang, China: mass flows and source analysis. Water Res. 45 (15), 4483–
4490.
Tao, L., Ma, J., Kunisue, T., Libelo, E.L., Tanabe, S., Kannan, K., 2008. Perfluorinated
compounds in human breast milk from several Asian countries, and in infant
formula and dairy milk from the United States. Environ. Sci. Technol. 42 (22),
8597–8602.
The World Bank Group, 2014. Classification of Country and Lending Group. <http://
data.worldbank.org/about/country-and-lending-groups>.
Tong, H., Walton, A., Sang, J.G., Chan, J.C.L., 2005. Numerical simulation of the urban
boundary layer over the complex terrain of Hong Kong. Atmos. Environ. 39 (19),
3549–3563.
Wang, J., 2006. Current state and trend of fluoropolymer in China. Organo-Fluorine
Ind. 4, 28–31 (in Chinese).
Wang, Y., Yeung, L.W.Y., Taniyasu, S., Yamashita, N., Lam, J.C.W., Lam, P.K.S., 2008.
Perfluorooctane sulfonate and other fluorochemicals in waterbird eggs from
South China. Environ. Sci. Technol. 42 (21), 8146–8151.
Wang, T., Wang, Y.W., Liao, C.Y., Cai, Y.Q., Jiang, G.B., 2009. Perspectives on the
inclusion of perfluorooctane sulfonate into the Stockholm convention on
persistent organic pollutants. Environ. Sci. Technol. 43 (14), 5171–5175.
Wang, J.M., Shi, Y.L., Pan, Y.Y., Cai, Y.Q., 2010a. Perfluorooctane sulfonate (PFOS) and
other fluorochemicals in viscera and muscle of farmed pigs and chickens in
Beijing, China. Chin. Sci. Bull. 55 (31), 3550–3555.
Wang, Y.W., Fu, J.J., Wang, T., Liang, Y., Pan, Y.Y., Cai, Y.Q., Jiang, G.B., 2010b.
Distribution of perfluorooctane sulfonate and other perfluorochemicals in the
ambient environment around a manufacturing facility in China. Environ. Sci.
Technol. 44 (21), 8062–8067.
Wang, T.Y., Chen, C.L., Naile, J.E., Khim, J.S., Giesy, J.P., Lu, Y.L., 2011a. Perfluorinated
compounds in water, sediment and soil from Guanting Reservoir, China. Bull.
Environ. Contam. Toxicol. 87 (1), 74–79.
Wang, T.Y., Lu, Y.L., Chen, C.L., Naile, J.E., Khim, J.S., Giesy, J.P., 2011b. Perfluorinated
compounds in a coastal industrial area of Tianjin, China. Environ. Geochem.
Health 34 (3), 301–311.
Wang, T.Y., Lu, Y.L., Chen, C.L., Naile, J.E., Khim, J.S., Park, J., Luo, W., Jiao, W.T., Hu,
W.Y., Giesy, J.P., 2011c. Perfluorinated compounds in estuarine and coastal
areas of north Bohai Sea, China. Mar. Pollut. Bull. 62 (8), 1905–1914.
Wang, T.Y., Khim, J.S., Chen, C.L., Naile, J.E., Lu, Y.L., Kannan, K., Park, J., Luo, W., Jiao,
W.T., Hu, W.Y., Giesy, J.P., 2012. Perfluorinated compounds in surface waters
from Northern China: comparison to level of industrialization. Environ. Int. 42,
37–46.
Wang, B.B., Cao, M.H., Zhu, H.D., Chen, J., Wang, L.L., Liu, G.H., Gu, X.M., Lu, X.H.,
2013a. Distribution of perfluorinated compounds in surface water from
Hanjiang River in Wuhan, China. Chemosphere 93 (3), 468–473.
Wang, J.S., Zhang, Y.T., Zhang, F., Yeung, L.W.Y., Taniyasu, S., Yamazaki, E., Wang,
R.P., Lam, P.K.S., Yamashita, N., Dai, J.Y., 2013b. Age- and gender-related
accumulation of perfluoroalkyl substances in captive Chinese alligators
(Alligator sinensis). Environ. Pollut. 179, 61–67.
Wang, P., Wang, T.Y., Giesy, J.P., Lu, Y.L., 2013c. Perfluorinated compounds in soils
from Liaodong Bay with concentrated fluorine industry parks in China.
Chemosphere 91 (6), 751–757.
Wang, J.M., Pan, Y.Y., Shi, Y.L., Cai, Y.Q., 2014a. Perfluorinated compounds in snow
from downtown of Beijing, China. Scientia Sinica: Chimica. 41 (5), 900–906 (in
Chinese).
T. Wang et al. / Chemosphere 129 (2015) 87–99
Wang, P., Lu, Y.L., Wang, T.Y., Fu, Y.N., Zhu, Z.Y., Liu, S.J., Xie, S.W., Xiao, Y., Giesy, J.P.,
2014b. Occurrence and transport of 17 perfluoroalkyl acids in 12 coastal rivers
in south Bohai coastal region of China with concentrated fluoropolymer
facilities. Environ. Pollut. 190, 115–122.
Wang, Z.Y., Cousins, I.T., Scheringer, M., Buck, R.C., Hungerbühler, K., 2014c. Global
emission inventories for C4–C14 perfluoroalkyl carboxylic acid (PFCA)
homologues from 1951 to 2030, Part I: production and emissions from
quantifiable sources. Environ. Int. 70, 62–75.
Wu, Y.N., Wang, Y.X., Li, J.G., Zhao, Y.F., Guo, F.F., Liu, J.Y., Cai, Z.W., 2012.
Perfluorinated compounds in seafood from coastal areas in China. Environ. Int.
42, 67–71.
Xie, S.W., Lu, Y.L., Wang, T.Y., Liu, S.J., Jones, K.C., Sweetman, A.J., 2013a. Estimation
of PFOS emission from domestic sources in the eastern coastal region of China.
Environ. Int. 59, 336–343.
Xie, S.W., Wang, T.Y., Liu, S.J., Jones, K.C., Sweetman, A.J., Lu, Y.L., 2013b. Industrial
source identification and emission estimation of perfluorooctane sulfonate in
China. Environ. Int. 52, 1–8.
Yan, H., Zhang, C.J., Zhou, Q., Chen, L., Meng, X.Z., 2012. Short- and long-chain
perfluorinated acids in sewage sludge from Shanghai, China. Chemosphere 88
(11), 1300–1305.
Yang, L.P., Zhu, L.Y., Liu, Z.T., 2011. Occurrence and partition of perfluorinated
compounds in water and sediment from Liao River and Taihu Lake, China.
Chemosphere 83 (6), 806–814.
Yang, S.W., Xu, F.F., Wu, F.C., Wang, S.R., Zheng, B.H., 2014. Development of PFOS
and PFOA criteria for the protection of freshwater aquatic life in China. Sci. Total
Environ. 470, 677–683.
Yeung, L.W.Y., Miyake, Y., Taniyasu, S., Wang, Y., Yu, H.X., So, M.K., Jiang, G.B., Wu,
Y.N., Li, J.G., Giesy, J.P., Yamashita, N., Lam, P.K.S., 2008. Perfluorinated
compounds and total and extractable organic fluorine in human blood
samples from China. Environ. Sci. Technol. 42 (21), 8140–8145.
Yeung, L.W.Y., Miyake, Y., Wang, Y., Taniyasu, S., Yamashita, N., Lam, P.K.S., 2009.
Total fluorine, extractable organic fluorine, perfluorooctane sulfonate and other
related fluorochemicals in liver of Indo-Pacific humpback dolphins (Sousa
Chinensis) and finless porpoises (Neophocaena phocaenoides) from South China.
Environ. Pollut. 157 (1), 17–23.
Yu, N.Y., Shi, W., Zhang, B.B., Su, G.Y., Feng, J.F., Zhang, X.W., Wei, S., Yu, H.X., 2013.
Occurrence of perfluoroalkyl acids including perfluorooctane sulfonate isomers
in Huai River Basin and Taihu Lake in Jiangsu Province, China. Environ. Sci.
Technol. 47 (2), 710–717.
Zhang, T., Sun, H.W., Wu, Q., Zhang, X.Z., Yun, S.H., Kannan, K., 2010a.
Perfluorochemicals in meat, eggs and indoor dust in China: assessment of
sources and pathways of human exposure to perfluorochemicals. Environ. Sci.
Technol. 44 (9), 3572–3579.
99
Zhang, T., Wu, Q., Sun, H.W., Zhang, X.Z., Yun, S.H., Kannan, K., 2010b. Perfluorinated
compounds in whole blood samples from infants, children, and adults in China.
Environ. Sci. Technol. 44 (11), 4341–4347.
Zhang, T., Sun, H.W., Lin, Y., Wang, L., Zhang, X.Z., Liu, Y., Geng, X., Zhao, L.J., Li, F.S.,
Kannan, K., 2011. Perfluorinated compounds in human blood, water, edible
freshwater fish, and seafood in China: daily intake and regional differences in
human exposures. J. Agric. Food Chem. 59 (20), 11168–11176.
Zhang, D.W., Zhang, L., Wei, Y.H., Wang, D.G., Luo, L.G., Chen, Y.W., 2012a.
Investigation of perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate
(PFOS) pollution in surface water of Poyang Lake. Resour. Environ. Yangtze
Basin 21, 885–890 (in Chinese).
Zhang, R., Wang, M.X., Tang, J., Hua, R.M., 2012b. Investigation of PFOS in Chaohu Lake
water and risk assessment of drinking. J. Anhui Agric. Univ. 39, 92–96 (in Chinese).
Zhang, Y., Meng, W., Guo, C.S., Xu, J., Yu, T., Fan, W.H., Li, L., 2012c. Determination
and partitioning behavior of perfluoroalkyl carboxylic acids and
perfluorooctanesulfonate in water and sediment from Dianchi Lake, China.
Chemosphere 88 (11), 1292–1299.
Zhang, T., Sun, H.W., Lin, Y., Qin, X.L., Zhang, Y.F., Geng, X., Kannan, K., 2013a.
Distribution of poly- and perfluoroalkyl substances in matched samples from
pregnant women and carbon chain length related maternal transfer. Environ.
Sci. Technol. 47 (14), 7974–7981.
Zhang, W., Zhang, Y.T., Taniyasu, S., Yeung, L.W.Y., Lam, P.K.S., Wang, J.S., Li, X.H.,
Yamashita, N., Dai, J.Y., 2013b. Distribution and fate of perfluoroalkyl
substances in municipal wastewater treatment plants in economically
developed areas of China. Environ. Pollut. 176, 10–17.
Zhang, Y.F., Beesoon, S., Zhu, L.Y., Martin, J.W., 2013c. Biomonitoring of
perfluoroalkyl acids in human urine and estimates of biological half-life.
Environ. Sci. Technol. 47 (18), 10619–10627.
Zhang, Y.F., Beesoon, S., Zhu, L.Y., Martin, J.W., 2013d. Isomers of
perfluorooctanesulfonate and perfluorooctanoate and total perfluoroalkyl
acids in human serum from two cities in North China. Environ. Int. 53, 9–17.
Zhang, Y.Y., Lai, S.C., Zhao, Z., Liu, F.B., Chen, H.W., Zou, S.C., Xie, Z.Y., Ebinghaus, R.,
2013e. Spatial distribution of perfluoroalkyl acids in the Pearl River of Southern
China. Chemosphere 93 (8), 1519–1525.
Zhao, L.J., Zhou, M., Zhang, T., Sun, H.W., 2012a. Polyfluorinated and perfluorinated
chemicals in precipitation and runoff from cities across eastern and central
China. Arch. Environ. Contam. Toxicol. 64 (2), 198–207.
Zhao, Y.G., Wong, C.K.C., Wong, M.H., 2012b. Environmental contamination, human
exposure and body loadings of perfluorooctane sulfonate (PFOS), focusing on
Asian countries. Chemosphere 89 (4), 355–368.
Zhou, Z., Liang, Y., Shi, Y.L., Xu, L., Cai, Y.Q., 2013. Occurrence and transport of
perfluoroalkyl acids (PFAAs), including short-Chain PFAAs in Tangxun Lake,
China. Environ. Sci. Technol. 47 (16), 9249–9257.
<Supplemental Materials>
A review of sources, multimedia distribution and health risks of
perfluoroalkyl acids (PFAAs) in China
Tieyu Wanga, Pei Wanga,b, Jing Menga,b, Shijie Liua,b, Yonglong Lua,*, Jong Seong
Khimc, John Giesyd
* Corresponding author: Tel: +86 10 62849466; Fax: +86 10 62918177
E-mail address: yllu@rcees.ac.cn (Y. LU)
Table S1 Emission sources of PFOS and PFOA in China and other countries
Sources in China
PFOS
Sources in other countries
PFOS production, textiles, pesticides,
PFOS production, metal plating,
firefighting foams, semiconductors (IC
photographic industry, semiconductor
industry), metal plating, petroleum
industry, aviation (hydraulic fluids),
industry, cleaning products (solvents),
firefighting foams, fabric treatment,
rubber and plastics, leather, dope,
paper treatment, coatings, apparels, home
photography, aircraft hydraulic fluids,
furnishing and upholstery, carpet and
photoelectron, nanomaterials, medical
leather products, printing inks, cleaning
equipment, printing inks, papermaking,
products, paints and varnishes, glass
etc.
cleaning waxes and floor polishes,
pesticides, etc.
PFOA
PFOA manufacture, fluoropolymer
PFOA manufacture, fluoropolymer
manufacture, finishing agents, metal
manufacture, firefighting foams, floor
plating, firefighting foams, cleaning
polishes, cleaning formulations, hair care
products, glass cleaning waxes, paints
products, inks, medical inhalers, fuel
and varnishes, paper treatment, etc.
additives, paper, air fresheners, textile
treatments, etc.
Table S2 Concentrations of PFOS and PFOA from water (ng/L) and sediment (ng/g dw) from major water
systems in China
Water system
Location
Media
PFOS
PFOA
Reference
Songhua River
Heilongjiang
water
1.21
0.17
Liu et al. (2007)
Liaohe River
Liaoning
water
0.33
10.9
Yang et al. (2011)
sediment
0.15
0.08
water
0.42
169.04
sediment
nd
12.57
Daling River
Liaoning
Bao et al. (2011)
Hunhe River
Liaoning
water
7.32
0.51
Sun et al. (2011)
Haihe River
Tianjin
water
2.49
6.86
Wang et al. (2011a)
sediment
1.12
0.41
Hohhot
water
0.32
1.2
Shanxi
water
2.7
0.93
Huaihe River
Jiangsu
water
4.7
18
Yu et al. (2013)
Yangtze River
Chongqing-Wuhan
water
6.66
16.21
Jin et al. (2009)
Huangpu River
Shanghai
sediment
0.11
0.45
Bao et al. (2010)
Hanjiang River
Wuhan
water
51.8
81
Wang et al. (2013)
Pearl River
Guangzhou
water
3.3
3.7
Zhang et al. (2013)
sediment
0.86
0.2
Bao et al. (2010)
water
0.16
1.2
Wang et al. (2011b)
sediment
nd
0.28
water
0.55
18.4
sediment
0.24
nd
water
26.5
21.7
sediment
0.15
0.16
Yellow River
Guanting Reservoir
Baiyangdian Lake
Taihu Lake
Beijing
Hebei
Jiangsu
Wang et al. (2012)
Shi et al. (2012)
Yang et al. (2011)
Poyang Lake
Jiangxi
water
0.35
1.1
Zhang et al. (2012)
East Lake
Wuhan
water
60.4
55
Chen et al. (2012)
Tangxun Lake
Wuhan
water
357
372
Zhou et al. (2013)
sediment
74.4
2.35
Table S3 The criteria values and measured values of Total Daily Intake (TDI) for PFOS and PFOA.
TDI ng/kg bw/d
PFOS
PFOA
Objective
Reference
population
Nation
/Region
Criteria values
25
333
General
America
So et al. (2006)
100
100
General
Germany
Roos et al. (2008)
150
1500
General
European Union
Fromme et al. (2009)
300
3000
General
United Kingdom
Alexander et al. (2008)
So et al. (2006)
Measured values*
10.0 (4.0-30.0)
8.6 (4.0-17.0)
Infant
Zhoushan, China
(1.4-15.9)
(nd-88.4)
Infant
12 provinces, China
Liu et al. (2010a)
10.4
9.7
Infant
17 cities, China
Bao et al. (2010)
1.2
9.8
Adult
17 cities, China
Bao et al. (2010)
nd
0.2-0.9
Adult
Fuxin, China
Bao et al. (2010)
28.7
9.6
Infant
Japan
Tao et al. (2008b)
15
<4.7
Infant
Malaysia
Tao et al. (2008b)
12.1
<7.7
Infant
Philippines
Tao et al. (2008b)
10.3
<2.9
Infant
Indonesia
Tao et al. (2008b)
9.4
<3.7
Infant
Vietnam
Tao et al. (2008b)
8.3
<3.4
Infant
Cambodia
Tao et al. (2008b)
5.7
<4.4
Infant
India
Tao et al. (2008b)
1.6
1.0
Adult
Canada
Liu et al. (2010a)
1.4
2.9
Adult
Germany
Liu et al. (2010a)
10
10
Adult
British
Liu et al. (2010a)
1.1
nd
Adult
Spain
Tao et al. (2008b)
14.7
1.7
Infant
Massachusetts, USA
Tao et al. (2008a)
* indicated the mean value with the range in brackets
nd: no avaiable data
Reference:
Alexander, J., Auðunsson, G. A., Benford, D., Cockburn, A., Cravedi, J. P., Dogliotti, E., Domenico, A. D.,
Fernández-Cruz, M. L., Fink-Gremmels, J., Fürst, P., Galli, C., Grandjean, P., Gzyl, J.,
Heinemeyer, G., Johansson, N., Mutti, A., Schlatter, J., Leeuwen, R. V., Peteghem, C. V., Verger,
P. 2008. Perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA) and their salts. The
EFSA Journal. 653: 1-131.
Bao, J., Liu, W., Liu, L., Jin, Y., Dai, J., Ran, X., Zhang, Z., Tsuda, S., 2011. Perfluorinated Compounds in
the Environment and the Blood of Residents Living near Fluorochemical Plants in Fuxin, China.
Environ. Sci. Technol. 45, 8075-8080.
Bao, J., Liu, W., Liu, L., Jin, Y., Ran, X., Zhang, Z., 2010. Perfluorinated compounds in urban river
sediments from Guangzhou and Shanghai of China. Chemosphere. 80, 123-130.
Chen, J., Wang, L.L., Zhu, H.D., Wang, B.B., Liu, H.C., Cao, M.H., Miao, Z., Hu, L., Lu, X.H., Liu, G.H.,
2012. Spatial distribution of perfluorooctanoic acids and perfluorinate sulphonates in surface water
of East Lake. Environ. Sci. 33, 2586-2591.
Fromme, H., Tittlemier, S. A., Volkel, W., Wilhelm, M., Twardella, D. 2009. Perfluorinated compounds Exposure assessment for the general population in western countries. Int J Hyg Envir Heal. 212(3):
239-270.
Jin, Y.H., Liu, W., Sato, I., Nakayama, S.F., Sasaki, K., Saito, N., Tsuda, S., 2009. PFOS and PFOA in
environmental and tap water in China. Chemosphere. 77, 605-611.
Liu, B., Jin, Y.H., Yu, Q.L., Wang, K., Dong, G.H., Li, H.Y., Saitou, N., Sasaki, K., 2007. Investigation of
perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) pollution in the surface
water of the Songhua River. Acta Sci. Circumstantiae. 27, 480-486.
Liu, J., Li, J., Zhao, Y., Wang, Y., Zhang, L., Wu, Y. 2010a. The occurrence of perfluorinated alkyl
compounds in human milk from different regions of China. Environ Int. 36(5): 433-438.
Roos, P. H., Angerer, J., Dieter, H., Wilhelm, M., Wolfle, D., Hengstler, J. G. 2008. Perfluorinated
compounds (PFC) hit the headlines : meeting report on a satellite symposium of the annual
meeting of the German Society of Toxicology. Arch Toxicol. 82(1): 57-59.
Shi, Y., Pan, Y., Wang, J., Cai, Y., 2012. Distribution of perfluorinated compounds in water, sediment,
biota and floating plants in Baiyangdian Lake, China. J. Environ. Monit. 14, 636.
So, M. K., Yamashita, N., Taniyasu, S., Jiang, Q., Giesy, J. P., Chen, K., Lam, P. K. 2006. Health risks in
infants associated with exposure to perfluorinated compounds in human breast milk from
Zhoushan, China. Environ Sci Technol. 40(9): 2924-2929.
Sun, H., Li, F., Zhang, T., Zhang, X., He, N., Song, Q., Zhao, L., Sun, L., Sun, T., 2011. Perfluorinated
compounds in surface waters and WWTPs in Shenyang, China: Mass flows and source analysis.
Water Res. 45, 4483-4490.
Tao, L., Kannan, K., Wong, C. M., Arcaro, K. F., Butenhoff, J. L. 2008a. Perfluorinated compounds in
human milk from Massachusetts, USA. Environmental Science & Technology. 42(8): 3096-3101.
Tao, L., Ma, J., Kunisue, T., Libelo, E. L., Tanabe, S., Kannan, K. 2008b. Perfluorinated compounds in
human breast milk from several Asian countries, and in infant formula and dairy milk from the
United States. Environ Sci Technol. 42(22): 8597-8602.
Wang, B., Cao, M., Zhu, H., Chen, J., Wang, L., Liu, G., Gu, X., Lu, X., 2013. Distribution of
perfluorinated compounds in surface water from Hanjiang River in Wuhan, China. Chemosphere.
93, 468-473.
Wang, T.Y., Chen, C.L., Naile, J.E., Khim, J.S., Giesy, J.P., Lu, Y.L, 2011b. Perfluorinated Compounds in
Water, Sediment and Soil from Guanting Reservoir, China. Bull. Environ. Contam. Toxicol. 87,
74-79.
Wang, T.Y., Lu, Y.L., Chen, C.L., Naile, J.E., Khim, J.S., Giesy, J.P., 2011a. Perfluorinated compounds in
a coastal industrial area of Tianjin, China. Environ. Geochem. Health. 34, 301-311.
Wang, T.Y., Khim, J.S., Chen, C.L., Naile, J.E., Lu, Y.L., Kannan, K., Park, J., Luo, W., Jiao, W.T., Hu,
W.Y., Giesy, J.P., 2012. Perfluorinated compounds in surface waters from Northern China:
Comparison to level of industrialization. Environ. Int. 42, 37-46.
Yang, L., Zhu, L., Liu, Z., 2011. Occurrence and partition of perfluorinated compounds in water and
sediment from Liao River and Taihu Lake, China. Chemosphere. 83, 806-814.
Yu, N., Shi, W., Zhang, B., Su, G., Feng, J., Zhang, X., Wei, S., Yu, H., 2013. Occurrence of
Perfluoroalkyl Acids Including Perfluorooctane Sulfonate Isomers in Huai River Basin and Taihu
Lake in Jiangsu Province, China. Environ. Sci. Technol. 47, 710-717.
Zhang, D.W., Zhang, L., Wei, Y.H., Wang, D.G., Luo, L.G., Chen, Y.W., 2012. Investigation of
perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) pollution in surface water of
Poyang Lake. Resources and Environment in the Yangtze Basin. 21, 885-890.
Zhang, Y., Lai, S., Zhao, Z., Liu, F., Chen, H., Zou, S., Xie, Z., Ebinghaus, R., 2013. Spatial distribution of
perfluoroalkyl acids in the Pearl River of Southern China. Chemosphere. 93, 1519-1525.
Zhou, Z., Liang, Y., Shi, Y., Xu, L., Cai, Y., 2013. Occurrence and Transport of Perfluoroalkyl Acids
(PFAAs), Including Short-Chain PFAAs in Tangxun Lake, China. Environ. Sci. Technol. 47,
9249-9257.