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Chemosphere 85 (2011) 122–128
Contents lists available at ScienceDirect
Chemosphere
journal homepage: www.elsevier.com/locate/chemosphere
Assessment of risk to humans of bisphenol A in marine and freshwater fish from
Pearl River Delta, China
Xi Wei a, Yeqing Huang a, Ming H. Wong a, John P. Giesy b, Chris K.C. Wong a,⇑
a
b
Department of Biology, Croucher Institute of Environmental Sciences, Hong Kong Baptist University, Hong Kong, China
Department of Veterinary Biomedical Sciences & Toxicological Center, University of Saskatchewan, Canada
a r t i c l e
i n f o
Article history:
Received 26 November 2010
Received in revised form 9 March 2011
Accepted 21 May 2011
Available online 22 June 2011
Keywords:
Fish pollution
Bisphenol A
LC/MS/MS
Health
a b s t r a c t
Bisphenol A (BPA) is a high production-volume chemical used in the manufacture of a wide variety of consumer products. However it is also a ubiquitous contaminant that can interfere with endocrine systems of
wildlife and humans. China is the ‘‘world factory’’ and the Pearl River Delta is the major manufacturing center and is consequently polluted. Concentrations of BPA in meats of marketable fish had not been previously
reported for this region. In the study upon which we report here concentrations of BPA were determined in
20 common species of freshwater and marine fish, collected from markets in Hong Kong, SAR, China. A comprehensive analytical method based on SPE extraction and liquid chromatography electrospray ionization
tandem mass spectrometry (LC–ESI–MS/MS) was developed, validated and applied. The method limit of
detection (LOD) and limit of quantification (LOQ) were 0.5 and 1.25 ng g1 dw, respectively. BPA was
detected in 19 species of fish at concentrations, ranging from 0.5 to 2.0 ng g1 ww. Average daily BPA intake
per person ranged from 1.1 102 ng d1 for marine fish and 2.2 102 ng d1 for freshwater fish. Concentrations of BPA in fish from Hong Kong markets unlikely would be causing adverse population-level effects
in humans.
Ó 2011 Elsevier Ltd. All rights reserved.
1. Introduction
Bisphenol A (BPA) is a chemical widely used in production of
epoxy resins and polycarbonates, and is especially abundant in
PVC plastics (Vandenberg et al., 2009). The market for BPA has
been growing with the increasing demand for polycarbonates
and epoxy resins. Global demand for BPA is predicted to grow from
3.9 million tons in 2006 to about 5 million tons in 2010 (Tsai,
2006). Many countries throughout the world have large production
capacities for BPA, especially Germany, the Netherlands, the USA
and Japan (Vandenberg et al., 2009). In recent years, the annual
consumption of BPA in China has been approximately 206 000
ton (Peng et al., 2007). The extensive manufacture and use of
BPA has led to its ubiquitous in the environment. BPA concentrations in the range of 5–320 ng L1 in river waters (Fromme et al.,
2002; Ballesteros-Gomez et al., 2007), 20–700 ng L1 in sewage
effluents (Ballesteros-Gomez et al., 2007; Ruiz et al., 2007), 2–
208 ng m3 in air, 0.2–199 ng g1 in dust (Rudel et al., 2001; Wilson et al., 2007) and 0.1–384 ng g1 in food-stuffs (Goodson
et al., 2002; Thomson and Grounds, 2005) have been reported.
Concentrations of BPA in blood and urine of 0.3–4.4 lg L1 and
0.47–9.5 lg L1 have been reported in healthy adults from some
countries such as Japan, Korea, Belgian and China (Vandenberg
⇑ Corresponding author. Tel.: +852 3411 7053; fax: +852 3411 5995.
E-mail address: ckcwong@hkbu.edu.hk (C.K.C. Wong).
0045-6535/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.chemosphere.2011.05.038
et al., 2007, 2010). The presence of BPA in food is of special concern
since it constitutes the primary route of human exposure. Detection of BPA in human fluids and tissues has led to questions of
the safety of BPA and what effects it could cause to humans and
wildlife (Vandenberg et al., 2007).
BPA is classified as an endocrine disruptor, on the basis of its
detectable estrogenic (Wozniak and Murias, 2008) and/or antiandrogenic potency (Lee et al., 2005). In animal studies, prenatal
and neonatal exposures to BPA have been linked to early onset of
sexual maturation (Howdeshell et al., 1999), altered development
and tissue organization of mammary glands (Markey et al.,
2001), reproductive tract lesions (Newbold et al., 2007), increase
of prostate size (Timms et al., 2005) and decrease of sperm production (vom Saal et al., 1998) in offspring. Exposure to BPA has also
been associated with chronic effects in humans, including cardiovascular disease and diabetes (Lang et al., 2008). Because of the
large volume of production, wide dispersive use and endocrine disrupting properties, BPA is a candidate to be included in the list of
substances subjected to authorization in the new policy on chemicals approval in Canada, Europe, and several key states of
United States (i.e. Chicago, Connecticut, Minnisota and New York)
(EFSA, 2006; Health Canada, 2008). Thus, there are pressing needs
of more research studies for risk assessment and the control of human exposure to BPA.
Over the past 20 years, the Pearl River Delta (PRD) in South
China has rapidly transformed from a traditional agriculture-based
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X. Wei et al. / Chemosphere 85 (2011) 122–128
to an industry-based economy (Ouyang et al., 2006). This has resulted in a substantial increase in industrial discharges, and a
widespread distribution of various pollutants in this region (Peng
et al., 2007). At present the PRD is classified as a highly polluted region (Richardson et al., 2005; Zheng et al., 2008). Various types of
industrial chemicals, pesticides and dioxins have been detected in
the PRD (So et al., 2004; Zheng et al., 2004, 2009). Recently BPA
pollution has been reported in surface water and sedimentary
cores from the Pearl River (Peng et al., 2007; Gong et al., 2009;
Zhao et al., 2009). With the benefit of hindsight, the wildlife and
human health in PRD could potentially be affected by pollution
in the environment and various foods. Since a data gap exists on
BPA contamination in commercially food fish species from PRD
market, therefore the monitoring of its contamination in fish may
be useful to reveal the environmental occurrence and the exposure
risk. The objectives of the present study were to: (1) measure concentrations of BPA in 20 species of freshwater and marine fishes
available in local markets of Hong Kong; (2) compare concentrations of BPA in fishes cultivated in different regions of South China;
and (3) estimate the local daily intake of BPA through fish consumption by people in Hong Kong.
2. Materials and methods
2.1. Chemicals and equipment
Methanol and ethanol (HPLC/MS grade), acetonitrile (HPLC
grade) were purchased from Sigma–Aldrich, USA and Tedia, USA.
Pesticide grade petroleum ether and ethyl acetate were purchased
from LAB SCAN, UK. Authentic standards of bisphenol A and
bisphenol A-d16 were obtained from AccuStandard, Connecticut,
USA and Chiron, Trondheim, Norway. Stock solutions
(1000 ng mL1) of BPA and BPA-d16 were prepared in methanol.
Milli-Q water (Millipore, Milli-Q system) was used in sample preparation. All equipment, glassware and polypropylene (PP) centrifuge tubes (IWAKI, Japan) were pre-washed three times by
acetone followed by methanol.
An Agilent 1200 liquid chromatography (Waldbronn, Germany),
equipped with a quaternary high-pressure gradient pump and an
automatic sample injector was used for LC–MS/MS analysis. Chromatographic separation was performed by using an Agilent C8
(2.1 mm 12.5 mm, 5 lm) guard column (ZORBAX Eclipse XDBC8, Narrow-Bore) and a C18 ODS column (Agilent Zorbax XDBC18, 3.5 lm 2.1 mm 50 mm, 3. 5 lm). Tandem mass detection
was conducted by an Agilent 6410B Triple Quadrupole mass spectrometer system equipped with an Agilent Masshunter Workstation (version B.02.01) and an electrospray ionization source. In
order to achieve greater sensitivity, analytes were detected in a
Multiple Reaction Monitoring (MRM) mode with a dwell time of
10 min. The ionization source parameters were as follow: ion spray
voltage, 5000 V; source temperature, 350 °C; nebulizer pressure,
40 psi; dry gas flow, 10 L min1; delta EMV, 900 V for negative. Collision energy (CE) of BPA-d16 and BPA were also optimized to obtain maximum sensitivity (Table 1).
2.2. Sampling and preparation
Ten species of marine fishes and ten species of freshwater fishes
were purchased from local markets in Hong Kong. The sources of
different fish species are given (Fig. 1A) while species names and
numbers of samples per species are shown (Table 2). Briefly, samples were wrapped in aluminum foil and stored on ice 0–4 °C during transportation. On arrival at the laboratory, individual fish
were dissected immediately. Filleted muscle was freeze-dried
and water content was determined. Dried samples were then
Table 1
BPA and BPA-d16 MRM conditions.
Compound
Retention
time
Precursor
Product
Dwell
(ms)
Fragmentor
(V)
Collision
energy
(V)
BPA-d16
6.72
241.1
BPA
6.74
227.1
223.1
142.1
212.1
133.1
80
80
80
80
95
95
95
95
12
20
10
20
ground into fine powder, homogenized and stored in desiccators
until extraction followed by BPA quantification.
2.3. Extraction and clean-up
An aliquot of 0.2 g dry weight (accurate to three significant figures) was extracted. Ten ng of deuterated BPA (BPA-d16) was
added to each sample as an internal recovery standard. Each sample was extracted with 10 mL acetonitrile in a pre-washed 50 mL
PP centrifuge tube. The sample was extracted in an ultrasonic bath
(Models 3510, Branson, USA) for 30 min and was then mixed in a
digital reciprocating shaker (HS501, IKA, Germany) for 30 min at
300 mot min1 at room temperature. The solution was centrifuged
(Allegra 6R, BechMan, USA) at 1500 rcf for 15 min. The supernatant
(the acetonitrile phase) was saved. The extraction was repeated
twice and all the extracts pooled. An aliquot of 0.5 mL of the
extract was saved for lipid content determination. The remaining
extract was mixed with 15 mL n-hexane and was shaken vigorously for 30 min to remove lipids (Grumetto et al., 2008). The acetonitrile layer was then mixed with 150 mL Milli-Q water in a
250 mL glass conical flask and ultrasonic for 5 min. The extract
was cleaned-up according to previously reported methods (Sajiki
et al., 2007), with minor modification. HLB cartridges (Oasis HLB
6 cc, 200 mg, Waters, UK) were conditioned and equilibrated with
5 mL ethanol, followed by 5 mL MilliQ water. The extract was
loaded into the cartridge at a flow rate of 1 mL min1, washed with
5 mL 15% ethanol, 5 mL MilliQ water and 20 mL petroleum ether.
Each sample was eluted with 14 mL of ethyl acetate at a flow rate
of 1 mL min1. The eluent was dried under N2 and was redissolved
in 1 mL of methanol/water (50:50) prior to LC/MS/MS analysis.
2.4. Determination of lipid content in fish
Twenty grams of freeze-dried sample was soxhlet extracted
with 400 mL of 50% v/v hexane:dichloromethane for at least
40 h. Sample extracts were heated in a furnace set at 80 °C for at
least 5 h. The extracts were kept in a desiccator for at least half
hour and weighted. The heating and weighing procedures were repeated until the difference between two consecutive readings is
less than 0.05 g. The final reading is the lipid content in the 20 g
freeze-dried fish sample.
2.5. HPLC–MS/MS analysis
Gradient LC conditions were adopted for separation and quantification of BPA in fish samples. In brief, mobile phases were A: MilliQ
water and B: methanol. A six-step gradient was used as follows:
0 min, 20% B (flow rate: 0.25 mL min1); 0–2.5 min, linear gradient
from 20%B to 95%B (flow rate: 0.25 mL min1); 2.5–5.5 min, 95%B
(flow rate: 0.25 mL min1); 5.5–7 min, linear gradient from 95%B
to 20%B (flow rate: 0.25 mL min1); 7–8 min, 20%B (flow rate:
0.8 mL min1); 8–10 min, 20%B (flow rate: 0.25 mL min1). The
injection volume was 20 lL and the column temperature was maintained at 25 °C throughout the chromatography. The MS/MS was
operated in the negative ESI mode and the MRM detection mode.
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X. Wei et al. / Chemosphere 85 (2011) 122–128
(A)
Shunde
New Territories
Hong Kong
Hong Kong
island
South China Sea
Hainan
(B)
geye
Big
eeker'ss Group
Ble
per
ellow Se
Ye
eafin
ongue S
To
Sole
Orrange-s
spotted
per
d group
artail Fla
Ba
athead
olden T
Go
Threadffin Brea
am
oldspottted rab
Go
h
bbitifish
nubnos e Pom pano
Sn
ellow Croaker
Ye
Grrey Mullet
potted S
Sp
head
Snakeh
Marine Fish
apia
Tila
andarin
Ma
n Fish
nakehead
Sn
Ricce field
d eel
atfish
Ca
Grrass Ca
arp
g head Carp
Big
ud Fish
Mu
h
Freshwater Fish
BPA
A Concc (ng/g lw)
30
25
20
15
10
5
0
New
Territories
Shunde
Hong Kong
Island
South China
Sea
Hainan
Fig. 1. (A) Locations of cultured fish and (B) the concentrations of BPA in the fish samples, in respective to the origins of production. Bars with the same letter are not
significantly different according to the results of one-way ANOVA followed by Duncan’s multiple range test (p < 0.05).
2.6. Quality control
Instrumental Limits of Detection (LOD) (S/N > 3) and Limits of
Quantification (LOQ) (S/N > 10) for BPA were 0.15 ng g1 dw and
0.25 ng g1 dw respectively. The method LOD and LOQ for BPA in
fish samples were 0.5 ng g1 dw and 1.25 ng g1 dw respectively.
Data were reported as ‘‘not detected’’ (ND) when the level was
lower than LOD. One procedural blank was analyzed after each
batch of eight samples to check for any interference and contamination from solvent and glassware. Rates of recovery of BPA-d16
standard with 3 ng and 10 ng spiking levels ranged between 76–
78% and 78–79%, respectively (n = 3).
2.7. Statistical analysis
Statistical evaluations were conducted by SPSS16. All data were
tested to be normally distributed and independent by using the
Normal Plots in SPSS and Shapiro–Wilk significance were 0.05. Differences among groups were tested for statistical significance by
analysis of variance (ANOVA) followed by Duncan’s Multiple Range
test (significance at p < 0.05) SPSS16. Results are presented as the
mean ± SEM. Groups were considered significantly different if
p < 0.05.
3. Results and discussion
3.1. Concentrations of BPA in fishes from Hong Kong market
Concentrations of BPA in meats of fishes collected from Hong
Kong markets are tabulated (Table 2). BPA was detected in all
fishes except in the Bleeker’s groupers. In freshwater fish samples,
concentrations of BPA ranged from 0.5 to 2.0 ng g1 ww (wet
weight), in which rice field eels exhibited the greatest concentration. In marine fishes, concentrations of BPA ranged from 0.0 to
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X. Wei et al. / Chemosphere 85 (2011) 122–128
Table 2
Mean concentrations of BPA in 20 species of fish from HK market.
Sample
Common name
Species name
N
Weight (g)
Lipid (%)
Original locations
BPA (ng g1 ww)
BPA (ng g1 lw)
Freshwater
Big head carp
Catfish
Grass carp
Grey mullet
Mandarin fish
Mud fish
Rice field eel
Snakehead
Spotted snakehead
Tilapia
Aristichthys nobilis
Claris fuscus
Ctenopharyngodon idellus
Mulgil cephalus
Siniperca kneri
Cirrhina molitorella
Monopterus albus
Channa asiatiea
Channa maculate
Oreochromis ossambicus
6
21
6
18
3
15
14
10
10
10
1000 ± 45.8
315.7 ± 36.7
916.7 ± 16.7
378.8 ± 27.7
1518.6 ± 118.6
413.6 ± 132.1
287.6 ± 22.6
450.0 ± 68.3
253.8 ± 18.2
430.6 ± 28.8
8.55
21.40
19.02
17.60
17.30
10.43
7.40
19.53
23.43
21.80
New Territories
New Territories
New Territories
Shunde
Shunde
New Territories
Shunde
Shunde
Shunde
Shunde
1.9 ± 1.9 10
2.0 ± 3.3 101
1.3 ± 1.0 102
0.6 ± 1.4 101
1.9 ± 5.9 101
2.0 ± 5.2 101
0.5 ± 2.3 101
0.6 ± 1.0 101
1.3 ± 1.4 101
1.4 ± 2.4 101
4.7 ± 1.1 100
3.3 ± 5.1 101
7.0 ± 1.6 100
3.4 ± 1.4 100
10.7 ± 4.2 101
18.5 ± 6.0 102
22.4 ± 3.9 100
9.1 ± 1.5 100
5.5 ± 6.0 101
6.3 ± 1.1 100
Marine
Bartail flathead
Bigeye
Bleeker’s grouper
Goldspotted rabbitfish
Golden threadfin bream
Orange-spotted grouper
Snubnose pompano
Tongue sole
Yellow croaker
Yellow seafin
Platycephalus indicus
Priacanthus acracanthus
Epinephelus bleekeri
Siganus punctatus
Nemipterus virgatus
Epinephelus coioides
Trachinotus blochii
Cynoglossus robustus
Pseudosciaena crocea
Acanthopagrus latus
6
33
10
15
36
9
9
27
15
9
489.1 ± 407.0
201.7 ± 55.8
358.0 ± 150.7
174.6 ± 62.6
121.9 ± 33.3
432.3 ± 16.7
409.4 ± 36.7
184.6 ± 32.4
352.2 ± 65.4
416.6 ± 32.8
18.28
4.31
16.8
25.66
8.44
24.38
44.53
16.52
44.61
27.74
South China Sea
Hainan
Hainan
Hong Kong Island
South China Sea
South China Sea
Hong Kong Island
South China Sea
Hong Kong Island
South China Sea
0.6 ± 8.0 102
0.7 ± 2.6 101
ND
0.7 ± 8.0 102
0.8 ± 2.2 101
0.7 ± 2.0 101
1.0 ± 3.5 101
1.1 ± 1.0 101
0.9 ± 0.8 101
0.5 ± 1.7 101
2.8 ± 3.0 101
3.3 ± 4.4 101
ND
9.6 ± 2.6 100
3.0 ± 8.3 101
2.2 ± 7.9 101
6.7 ± 1.0 100
1.7 ± 6.5 101
16.6 ± 3.0 100
1.9 ± 2.0 101
1
ND = not detected, and assumed as 0.0 for the BPA concentration values.
1.1 ng g1 ww, while yellow croaker exhibited the greatest concentration. Freshwater fishes (1.38 ± 0.2 ng g1 ww) collected from the
Hong Kong markets contained significantly (p < 0.05) greater concentrations of BPA than did marine fishes (0.69 ± 0.5 ng g1 ww).
Although it has been reported that BPA is biodegradable, the leaching of BPA from plastic products, effluents from wastewater treatment plants and landfills contribute significant amounts of BPA in
the freshwater systems (Oehlmann et al., 2009). According to the
studies conducted in the United States, Germany, Japan, Spain,
and Netherlands, concentrations of BPA in surface water from rivers and fish ponds were ranged from 2 to 21 000 ng L1 (Belfroid
et al., 2002; Crain et al., 2007). Although BPA concentrations in
freshwater environments can be greater, concentrations in marine
environments might be less due to dilution and microbial degradation (Miceli et al., 2009).
In both freshwater and seawater fishes, relatively great concentrations of BPA were observed in carnivorous fish (rice field eel and
yellow croaker). The likely explanation for this result is that carnivorous fish species are the top consumers in food chains (UNEP,
2002) and they commonly accumulate more pollutants (Nie
et al., 2005). In aquaculture of PRD, carnivorous fishes is generally
fed with fish offal, those might contain relatively high levels of pollutants (Cheung et al., 2007; Arvanitoyannis and Kassaveti, 2008).
In addition to carnivorous fishes, bottom-feeding fish, such as
mud fish was also contaminated with a greater concentration of
BPA. This may be due to the greater concentrations of BPA exposure in sediments (Heemken et al., 2001). This assumption is supported by a dated monitoring study, which reported that BPA was
well preserved in river sediments collected from the PRD estuary
(Peng et al., 2007).
3.2. Comparison of fish from Hong Kong Markets to those from other
locations
The origins of fishes sold in the Hong Kong markets were shown
(Fig. 1A). The greatest BPA contaminations were found in the fish
from Hong Kong (New Territories and Island) and Shunde at
Guangdong province (Fig. 1B). Comparatively BPA concentrations
in fishes from Hainan and the South China Sea were less. This result
was expected since urban areas usually display greater concentrations of BPA than suburb areas (WSP, 2007). In the PRD, the aquatic
environment has been considerably polluted with the drastic increase in population density and industrial activities. It is
estimated that domestic and industrial wastewater amounted to
3.0 109 m3 annually in the PRD, of which more than 60% of the
sewage was not treated before discharge. The detectable BPA contamination in the Pearl River ranged from 43.5 to 639.1 ng L1
(Gong et al., 2009).
In Hong Kong, the rapid economic development has pushed
manufacturing into the New Territories, where the most intensive
freshwater aquaculture and three strategic landfills are located
(Chung and Poon, 2001). In addition to the industrial discharges,
landfill leachates also contribute significant source of pollutants
(Eggen et al., 2010). A report from Japan highlighted that concentrations as great as 17 200 lg BPA L1 can be detected in leachates
from landfills (Yamamoto et al., 2001). In the other location Shunde, this region has a network of cities that form the aquacultural
and industrial base of the PRD, where over 400 fishponds are located. Downstream tributaries of the Zhujiang River in Guangdong
Province, are the major water source used for aquaculture in Shunde. Relatively greater concentrations of BPA, ranging from 43.5 to
639.1 ng L1 were detected in surface waters collected from the
Zhujiang River (Gong et al., 2009).
3.3. Comparison of BPA concentrations
Previous studies have reported concentrations of BPA in fresh
marketable fish, either on body lipid weight (lp) or wet weight
(ww) basis. To facilitate data comparison, in the present study
the data were presented in both lp and ww (Table 3). On the lipid
weight basis, BPA concentrations in fishes collected from Hong
Kong is comparable to the carp samples from Japan but lesser than
the black seabream from Yundang Lagoon in Xiamen, China (Zhang
et al., 2010). On the wet weight basis, concentrations of BPA observed in this study (0.5–2.0 ng g1 ww) were comparable to the
concentrations detected in marketable fishes in Netherland,
Sweden and Italy but were lesser than those in fishes collected
from markets in Beijing, China (Shao et al., 2005), and coasts in
Norway (Fjeld et al., 2004).
3.4. Estimated Intake of BPA via consumption of fish
Consumption of fish often contributes a significant proportion
of total intake of persistent organic pollutants (POPs) in human
diets (Cheung et al., 2008). The dietary habits of Chinese people
in South China consume more fish than other meats or dairy
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Table 3
BPA found in fresh marketable fish from other studies.
Sampling date
Sample size (species)
Country
2001
2004
2005
2006
2009
2007
2010
2009
6(3)
3(5)
6(ND)
1–10(4)
10(5)
30(1)
6(3)
6–30(20)
Netherland
Norway
Beijing, China
Sweden
Italy
Japan
Xiamen, China
Hong Kong
BPA
products. A person of 60 kg body weight in Hong Kong consumes
fish or shellfish four or more times a week, averaging about
60 kg of fish per year, which is equal to 164.4 g of fish per person
per day (Dickman and Leung, 1998). Using this value, the calculated average daily intakes (ADI) ranged from 1.86 ng BPA per kg
body weight per day for marine fish to 3.69 ng BPA per kg body
weight per day for freshwater fish. Accordingly the estimated
ADI of BPA intake via consumption of fish in Hong Kong market
for an individual with 60 kg bw would be as much as 169.2 ng
BPA per person per day. According to the USEPA (2010), the RfD
for BPA is 50 lg BPA kg1 d1. Based on the average concentration
of BPA detected in fishes in this study, the HR (Eq. (1)) for BPA
(2.81 ng kg1 d1/50 lg kg1 d1) in edible portions of fishes sold
in Hong Kong is 0.000056. This value is much less than 1.0 and
the calculated margin of exposure (MoE) is as high as 17 793 (Eq.
(2)). Several toxicokinetic studies on BPA metabolism in healthy
adult have indicated that BPA can be rapidly cleared from blood
(Volkel et al., 2005, 2008), hence the exposure of BPA through fish
consumption seems to present no remarkable risk to humans in
Hong Kong.
Authors
Wet weight
Lipid weight
0.24–2.6 ng g1 ww
1–14 ng g1 ww
0.33–7.8 ng g1 ww
<0.24–4.7 ng g1 ww
0.1–4.9 ng g1 ww
–
–
0.5–2.0 ng g1 ww
–
–
–
–
–
0–30 ng g1 lw
64.8–177.4 ng g1 lw
2.2–22.4 ng g1 lw
Belfroid et al. (2002)
Fjeld et al. (2004)
Shao et al. (2005)
WSP (2007)
Mita et al. (2011)
NITE (2007)
Zhang et al. (2010)
Present study
4. Conclusions
HR ¼ ADI=RfD
ð1Þ
Fish is a recommended bio-indicator for monitoring POPs
(UNEP, 2004) and consumption of fish has many health benefits.
In many areas of the world fish provides the major source of protein and essential amino acids in the diet of humans (Usydus
et al., 2009). The nutritional benefits of fish consumption are due
to the presence of essential omega-3, unsaturated fatty acids and
minerals (Sidhu, 2003). Consumption of omega-3 fatty acids in fish
or fish oil reduces the risk of coronary heart disease, and lessens
hypertension and plasma triglycerides, and prevents cardiac
arrhythmias and sudden death (Berry, 1997; Albert et al., 2002).
Thus, the potential adverse effects of contaminants in fish cannot
be made without the analysis of these risks balanced against the
benefits of consuming fish. This study provided the first evaluation
of concentrations of the endocrine-disrupting chemical BPA in fish
and its potential to cause adverse effects in humans in Hong Kong.
Moreover concentrations of other pollutions, such as DDTs and
PAHs may cause a greater risk (Cheung et al., 2007). While better
control of contaminants entering the aquatic environments of
southern China is warranted, focusing on minimizing releases of
BPA is or relatively low priority.
MoE ¼ 1=HR
ð2Þ
Acknowledgements
However, this HR does not take into account of other food sources
and non-dietary routes to BPA exposure (Stahlhut et al., 2009; Zalko
et al., 2011). Furthermore the RfD value of 50 lg BPA kg1 d1 is
somewhat controversial since in animal studies, endocrine disrupting effects have been reported by exposure to concentrations of BPA
at 102–103 fold less than the RfD (vom Saal and Hughes, 2005; vom
Saal et al., 2005). Adverse effects like meiotic abnormalities in fetal
oocytes and defects in the male and female reproductive tracts were
found at exposure levels of less than 20 lg kg1 d1 during prenatal
and neonatal development in mice (Hunt et al., 2003, 2009). If these
effects are found to be relevant to humans, using the lowest RfD of
50 ng BPA kg1 d1, in the present study the calculated HR (0.056)
is still less than 1. Therefore it is unlikely that BPA derived from
the fish in Hong Kong markets would be causing adverse population-level effects in humans. According to the existing migration
data, the daily intake of BPA from the food contact sources, like
polycarbonate plastic and canned foods, is estimated in the range
of 0.5–1.6 lg kg1 bw (USFDA, 2002). The data indicate that the
fresh fish samples from Hong Kong are probably a minor dietary
source of BPA. Moreover the additive effects of BPA and other pollutants cannot be ignored (Schmidt et al., 2006; Xiao et al., 2010).
It has been estimated that a person can be exposed to more than
10 types of chemical pollutants per day through dietary route (Pompa et al., 2003). In marketable fish from Hong Kong, contamination
with other pollutants like DDTs, PAHs and PBDEs were reported
(Cheung et al. 2007, 2008). Accordingly these additive effects may
make the HR value of BPA itself more significant even if it is less
than 1.0.
This work was supported by the Super Faculty Research Grant,
Hong Kong Baptist University and Collaborative Research Fund
(HKBU 1/CRF/08 to Prof. C.K.C. Wong) University Grants Committee. Prof. Giesy was supported by the Canada Research Chair program and an at large Chair Professorship at the Department of
Biology and Chemistry and State Key Laboratory in Marine Pollution, City University of Hong Kong.
References
Albert, C.M., Campos, H., Stampfer, M.J., Ridker, P.M., Manson, J.E., Willett, W.C., Ma,
J., 2002. Blood levels of long-chain n-3 fatty acids and the risk of sudden death.
N. Engl. J. Med. 346, 1113–1118.
Arvanitoyannis, I.S., Kassaveti, A., 2008. Fish industry waste: treatments,
environmental impacts, current and potential uses. Int. J. Food Sci. Technol.
43, 726–745.
Ballesteros-Gomez, A., Ruiz, F.J., Rubio, S., Perez-Bendito, D., 2007. Determination of
bisphenols A and F and their diglycidyl ethers in wastewater and river water by
coacervative extraction and liquid chromatography–fluorimetry. Anal. Chim.
Acta 603, 51–59.
Belfroid, A., van Velzen, M., van der Horst, B., Vethaak, D., 2002. Occurrence of
bisphenol A in surface water and uptake in fish: evaluation of field
measurements. Chemosphere 49, 97–103.
Berry, E.M., 1997. Dietary fatty acids in the management of diabetes mellitus. Am. J.
Clin. Nutr. 66, 991S–997S.
Cheung, K.C., Leung, H.M., Kong, K.Y., Wong, M.H., 2007. Residual levels of DDTs and
PAHs in freshwater and marine fish from Hong Kong markets and their health
risk assessment. Chemosphere 66, 460–468.
Cheung, K.C., Zheng, J.S., Leung, H.M., Wong, M.H., 2008. Exposure to
polybrominated diphenyl ethers associated with consumption of marine and
freshwater fish in Hong Kong. Chemosphere 70, 1707–1720.
Chung, S.S., Poon, C.S., 2001. A comparison of waste-reduction practices and new
environmental paradigm of rural and urban Chinese citizens. J. Environ.
Manage. 62, 3–19.
Author's personal copy
X. Wei et al. / Chemosphere 85 (2011) 122–128
Crain, D.A., Eriksen, M., Iguchi, T., Jobling, S., Laufer, H., LeBlanc, G.A., Guillette Jr.,
L.J., 2007. An ecological assessment of bisphenol-A: evidence from comparative
biology. Reprod. Toxicol. 24, 225–239.
Dickman, M.D., Leung, K.M.C., 1998. Mercury and organochlorine exposure from
fish consumption in Hong Kong. Chemosphere 37, 991–1015.
European Food Safety Authority (EFSA), 2006. Opinion of the scientific panel on food
additives, flavourings, processing aids and materials in contact with food on a
request from the commission related to 2,2-bis (4-hydroxyphenyl)propane
(bisphenol A). EFSA J. 428, 1–75. <http://www.efsa.europa.eu/cs/BlobServer/
Scientific_Opinion/afc_op_ej428_bpa_op_en,1.pdf>.
Eggen, T., Moeder, M., Arukwe, A., 2010. Municipal landfill leachates: a significant
source for new and emerging pollutants. Sci. Total Environ. 408, 5147–5157.
Fjeld, E., Schlabach, M., Berge, J.A., Eggen, T., Snilsberg, P., Kjellberg, G., Rognerud, S.,
Enge, E.K., Borgen, A., Gundersen, H., 2004. Screening of Selected New Organic
Contaminants – Brominated Flame Retardants, Chlorinated Paraffins,
Bisphenol-A and Triclosan. Norwegian Institute for Water Research, Report
4809, p. 115.
Fromme, H., Kuchler, T., Otto, T., Pilz, K., Muller, J., Wenzel, A., 2002. Occurrence of
phthalates and bisphenol A and F in the environment. Water Res. 36, 1429–
1438.
Gong, J., Ran, Y., Chen, D., Yang, Y., Ma, X., 2009. Occurrence and environmental risk
of endocrine-disrupting chemicals in surface waters of the Pearl River, South
China. Environ. Monit. Assess. 156, 199–210.
Goodson, A., Summerfield, W., Cooper, I., 2002. Survey of bisphenol A and bisphenol
F in canned foods. Food Addit. Contam. 19, 796–802.
Grumetto, L., Montesano, D., Seccia, S., Albrizio, S., Barbato, F., 2008. Determination
of bisphenol a and bisphenol B residues in canned peeled tomatoes by reversedphase liquid chromatography. J. Agric. Food Chem. 56, 10633–10637.
Health Canada, 2008. Notice: Identification of Medical Devices Containing DEHP or
BPA.
<http://www.hc-sc.gc.ca/dhp-mps/md-im/activit/announce-annonce/
md_notice_im_avis_dehp_bpa-eng.php>.
Heemken, O.P., Reincke, H., Stachel, B., Theobald, N., 2001. The occurrence of
xenoestrogens in the Elbe river and the North Sea. Chemosphere 45, 245–259.
Howdeshell, K.L., Hotchkiss, A.K., Thayer, K.A., Vandenbergh, J.G., vom Saal, F.S.,
1999. Exposure to bisphenol A advances puberty. Nature 401, 763–764.
Hunt, P.A., Koehler, K.E., Susiarjo, M., Hodges, C.A., Ilagan, A., Voigt, R.C., Thomas, S.,
Thomas, B.F., Hassold, T.J., 2003. Bisphenol a exposure causes meiotic
aneuploidy in the female mouse. Curr. Biol. 13, 546–553.
Hunt, P.A., Susiarjo, M., Rubio, C., Hassold, T.J., 2009. The bisphenol A experience: a
primer for the analysis of environmental effects on mammalian reproduction.
Biol. Reprod. 81, 807–813.
Lang, I.A., Galloway, T.S., Scarlett, A., Henley, W.E., Depledge, M., Wallace, R.B.,
Melzer, D., 2008. Association of urinary bisphenol A concentration with
medical disorders and laboratory abnormalities in adults. JAMA 300, 1303–
1310.
Lee, C.K., Kim, S.H., Moon, D.H., Kim, J.H., Son, B.C., Kim, D.H., Lee, C.H., Kim,
H.D., Kim, J.W., Kim, J.E., Lee, C.U., 2005. Effects of bisphenol A on the
placental function and reproduction in rats. J. Prev. Med. Public Health 38,
330–336.
Markey, C.M., Luque, E.H., Munoz De Toro, M., Sonnenschein, C., Soto, A.M., 2001. In
utero exposure to bisphenol A alters the development and tissue organization
of the mouse mammary gland. Biol. Reprod. 65, 1215–1223.
Miceli, A., Capoun, R., Fino, C., Narayan, P., Bryan, A.J., Angelini, G.D., Caputo, M.,
2009. Effects of angiotensin-converting enzyme inhibitor therapy on clinical
outcome in patients undergoing coronary artery bypass grafting. J. Am. Coll.
Cardiol. 54, 1778–1784.
Mita, L., Bianco, M., Viggiano, E., Zollo, F., Bencivenga, U., Sica, V., Monaco, G.,
Portaccio, M., Diano, N., Colonna, A., Lepore, M., Canciglia, P., Mita, D.G., 2011.
Bisphenol A content in fish caught in two different sites of the Tyrrhenian Sea
(Italy). Chemosphere 82, 405–410.
Newbold, R.R., Jefferson, W.N., Padilla-Banks, E., 2007. Long-term adverse effects of
neonatal exposure to bisphenol A on the murine female reproductive tract.
Reprod. Toxicol. 24, 253–258.
Nie, X., Lan, C., Wei, T., Yang, Y., 2005. Distribution of polychlorinated biphenyls in
the water, sediment and fish from the Pearl River estuary, China. Mar. Pollut.
Bull. 50, 537–546.
National Institute of Technology and Evaluation (NITE), 2007. Summary of the
Interim Report – Bisphenol A. <http://www.safe.nite.go.jp/risk/pdf/
interimreport_summary_bpa.pdf>.
Oehlmann, J., Schulte-Oehlmann, U., Kloas, W., Jagnytsch, O., Lutz, I., Kusk, K.O.,
Wollenberger, L., Santos, E.M., Paull, G.C., Van Look, K.J., Tyler, C.R., 2009. A
critical analysis of the biological impacts of plasticizers on wildlife. Philos.
Trans. R. Soc. London, B 364, 2047–2062.
Ouyang, T.P., Zhu, Z.Y., Kuang, Y.Q., 2006. Assessing impact of urbanization on river
water quality in the Pearl River Delta Economic Zone, China. Environ. Monit.
Assess. 120, 313–325.
Peng, X., Wang, Z., Mai, B., Chen, F., Chen, S., Tan, J., Yu, Y., Tang, C., Li, K., Zhang, G.,
Yang, C., 2007. Temporal trends of nonylphenol and bisphenol A contamination
in the Pearl River Estuary and the adjacent South China Sea recorded by dated
sedimentary cores. Sci. Total Environ. 384, 393–400.
Pompa, G., Caloni, F., Fracchiolla, M.L., 2003. Dioxin and PCB contamination of fish
and shellfish: assessment of human exposure. Review of the international
situation. Vet. Res. Commun. 27 (Suppl. 1), 159–167.
Richardson, B.J., Lam, P.K., Martin, M., 2005. Emerging chemicals of concern:
pharmaceuticals and personal care products (PPCPs) in Asia, with particular
reference to Southern China. Mar. Pollut. Bull. 50, 913–920.
127
Rudel, R.A., Brody, J.G., Spengler, J.C., Vallarino, J., Geno, P.W., Sun, G., Yau, A., 2001.
Identification of selected hormonally active agents and animal mammary
carcinogens in commercial and residential air and dust samples. J. Air Waste
Manage. 51, 499–513.
Ruiz, F.J., Rubio, S., Perez-Bendito, D., 2007. Vesicular coacervative extraction of
bisphenols and their diglycidyl ethers from sewage and river water. J.
Chromatogr., A 1163, 269–276.
Sajiki, J., Miyamoto, F., Fukata, H., Mori, C., Yonekubo, J., Hayakawa, K., 2007.
Bisphenol A (BPA) and its source in foods in Japanese markets. Food. Addit.
Contam. 24, 103–112.
Schmidt, S., Degen, G.H., Seibel, J., Hertrampf, T., Vollmer, G., Diel, P., 2006.
Hormonal activity of combinations of genistein, bisphenol A and 17betaestradiol in the female Wistar rat. Arch. Toxicol. 80, 839–845.
Shao, B., Han, H., Li, D., Zhao, R., Meng, J., Ma, Y., 2005. Analysis of nonylphenol,
octylphenol and bisphenol A in animal tissues by liquid chromatographytandem mass spectrometry with accelerated solvent extraction. Se Pu 23, 362–
365.
Sidhu, K.S., 2003. Health benefits and potential risks related to consumption of fish
or fish oil. Regul. Toxicol. Pharmacol. 38, 336–344.
So, M.K., Taniyasu, S., Yamashita, N., Giesy, J.P., Zheng, J., Fang, Z., Im, S.H., Lam, P.K.,
2004. Perfluorinated compounds in coastal waters of Hong Kong, South China,
and Korea. Environ. Sci. Technol. 38, 4056–4063.
Stahlhut, R.W., Welshons, W.V., Swan, S.H., 2009. Bisphenol A data in NHANES
suggest longer than expected half-life, substantial nonfood exposure, or both.
Environ. Health Perspect. 117, 784–789.
Thomson, B.M., Grounds, P.R., 2005. Bisphenol A in canned foods in New Zealand: an
exposure assessment. Food. Addit. Contam. 22, 65–72.
Timms, B.G., Howdeshell, K.L., Barton, L., Bradley, S., Richter, C.A., vom Saal, F.S.,
2005. Estrogenic chemicals in plastic and oral contraceptives disrupt
development of the fetal mouse prostate and urethra. Proc. Natl. Acad. Sci. U.
S. A. 102, 7014–7019.
Tsai, W.T., 2006. Human health risk on environmental exposure to Bisphenol-A: a
review. J. Environ. Sci. Health, Part C: Environ. Carcinog. Ecotoxicol. Rev. 24,
225–255.
United Nations Environment Programme (UNEP), 2002. UNEP Fish Supply Chains
Report.
<http://www.unep.ch/etb/publications/Fish%20Supply%20Chains/
UNEP%20fish%20supply%20chains%20report.pdf>.
United Nations Environment Programme (UNEP), 2004. Guidance for a Global
Monitoring Programme for Persistent Organic Pollutants. <http://
www.chem.unep.ch/gmn/guidancegpm.pdf>.
US Environmental Protection Agency (USEPA), 2010. Bisphenol A. (CASRN 80-05-7).
<http://www.epa.gov/iris/subst/0356.htm>.
US Food and Drug Administration, 2002. Draft Assessment of Bisphenol A for Use in
Food Contact Applications. <http://www.fda.gov/ohrms/dockets/ac/08/briefing/
2008-0038b1_01_02_FDA%20BPA%20Draft%20Assessment.pdf>.
Usydus, Z., Szlinder-Richert, J., Polak-Juszczak, L., Komar, K., Adamczyk, M., MalesaCiecwierz, M., Ruczynska, W., 2009. Fish products available in Polish market –
assessment of the nutritive value and human exposure to dioxins and other
contaminants. Chemosphere 74, 1420–1428.
Vandenberg, L.N., Hauser, R., Marcus, M., Olea, N., Welshons, W.V., 2007. Human
exposure to bisphenol A (BPA). Reprod. Toxicol. 24, 139–177.
Vandenberg, L.N., Maffini, M.V., Sonnenschein, C., Rubin, B.S., Soto, A.M., 2009.
Bisphenol-A and the great divide: a review of controversies in the field of
endocrine disruption. Endocr. Rev. 30, 75–95.
Vandenberg, L.N., Chahoud, I., Heindel, J.J., Padmanabhan, V., Paumgartten, F.J.,
Schoenfelder, G., 2010. Urinary, circulating, and tissue biomonitoring studies
indicate widespread exposure to bisphenol A. Environ. Health Perspect. 118,
1055–1070.
Volkel, W., Bittner, N., Dekant, W., 2005. Quantitation of bisphenol A and bisphenol
A glucuronide in biological samples by high performance liquid
chromatography–tandem mass spectrometry. Drug Metab. Dispos. 33, 1748–
1757.
Volkel, W., Kiranoglu, M., Fromme, H., 2008. Determination of free and total
bisphenol A in human urine to assess daily uptake as a basis for a valid risk
assessment. Toxicol. Lett. 179, 155–162.
vom Saal, F.S., Hughes, C., 2005. An extensive new literature concerning low-dose
effects of bisphenol A shows the need for a new risk assessment. Environ.
Health Perspect. 113, 926–933.
vom Saal, F.S., Cooke, P.S., Buchanan, D.L., Palanza, P., Thayer, K.A., Nagel, S.C.,
Parmigiani, S., Welshons, W.V., 1998. A physiologically based approach to the
study of bisphenol A and other estrogenic chemicals on the size of
reproductive organs, daily sperm production, and behavior. Toxicol. Ind.
Health 14, 239–260.
vom Saal, F.S., Nagel, S.C., Timms, B.G., Welshons, W.V., 2005. Implications for
human health of the extensive bisphenol A literature showing adverse effects at
low doses: a response to attempts to mislead the public. Toxicology 212, 244–
252 (author reply 253–244).
Wilson, N.K., Chuang, J.C., Morgan, M.K., Lordo, R.A., Sheldon, L.S., 2007. An
observational study of the potential exposures of preschool children to
pentachlorophenol, bisphenol-A, and nonylphenol at home and daycare.
Environ. Res. 103, 9–20.
Wozniak, M., Murias, M., 2008. Xenoestrogens: endocrine disrupting compounds.
Ginekol. Pol. 79, 785–790.
WSP, 2007. Screening of Bisphenol A in Fish from Swedish Waters. <http://
www.naturvardsverket.se/upload/02_tillstandet_i_miljon/Miljoovervakning/
rapporter/miljogift/bpa_fisk.pdf>.
Author's personal copy
128
X. Wei et al. / Chemosphere 85 (2011) 122–128
Xiao, Y., Liu, R., Xing, L., Xu, Y., Shang, L., Hao, W., 2010. Combined developmental
toxicity of bisphenol A and genistein in micromass cultures of rat embryonic
limb bud and midbrain cells. Toxicol. In Vitro.
Yamamoto, T., Yasuhara, A., Shiraishi, H., Nakasugi, O., 2001. Bisphenol A in
hazardous waste landfill leachates. Chemosphere 42, 415–418.
Zalko, D., Jacques, C., Duplan, H., Bruel, S., Perdu, E., 2011. Viable skin efficiently
absorbs and metabolizes bisphenol A. Chemosphere 82, 424–430.
Zhang, S., Peng, P., Huang, W., Li, X., Zhang, G., 2009. PCDD/PCDF pollution in soils and
sediments from the Pearl River Delta of China. Chemosphere 75, 1186–1195.
Zhang, X., Gao, Y., Li, Q., Li, G., Guo, Q., Yan, C., 2010. Estrogenic compounds and
estrogenicity in surface water, sediments, and organisms from Yundang Lagoon
in Xiamen, China. Arch. Environ. Contam. Toxicol.
Zhao, J.L., Ying, G.G., Wang, L., Yang, J.F., Yang, X.B., Yang, L.H., Li, X., 2009.
Determination of phenolic endocrine disrupting chemicals and acidic
pharmaceuticals in surface water of the Pearl Rivers in South China by gas
chromatography–negative chemical ionization–mass spectrometry. Sci. Total
Environ. 407, 962–974.
Zheng, G.J., Martin, M., Richardson, B.J., Yu, H., Liu, Y., Zhou, C., Li, J., Hu, G.,
Lam, M.H., Lam, P.K., 2004. Concentrations of polybrominated diphenyl
ethers (PBDEs) in Pearl River Delta sediments. Mar. Pollut. Bull. 49, 520–
524.
Zheng, G.J., Leung, A.O., Jiao, L.P., Wong, M.H., 2008. Polychlorinated dibenzo-pdioxins and dibenzofurans pollution in China: sources, environmental levels
and potential human health impacts. Environ. Int. 34, 1050–1061.