Environment International 42 (2012) 117–123
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Environment International
j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / e n v i n t
Thyroid hormone disrupting activities associated with phthalate esters in water
sources from Yangtze River Delta
Wei Shi a, Feng-Xian Zhang a, Guan-Jiu Hu b, Ying-Qun Hao b, Xiao-Wei Zhang a, Hong-Ling Liu a,⁎, Si Wei a,
Xin-Ru Wang c, John P. Giesy a, d, e, f, g, h, Hong-Xia Yu a,⁎
a
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210093, PR China
State Environmental Protection Key Laboratory of Monitoring and Analysis for Organic Pollutants in Surface Water, Jiangsu Provincial Environmental Monitoring Center, Nanjing, 210036,
PR China
c
Key Laboratory of Reproductive Medicine & Institute of Toxicology, Nanjing Medical University, Nanjing, 210029, PR China
d
Department of Veterinary Biomedical Sciences and Toxicology Centre, University of Saskatchewan, Saskatoon, SK, Canada
e
Department of Zoology, and Center for Integrative Toxicology, Michigan State University, East Lansing, MI, USA
f
Zoology Department, College of Science, King Saud University, P. O. Box 2455, Riyadh 11451, Saudi Arabia
g
Department of Biology & Chemistry, City University of Hong Kong, Kowloon, Hong Kong, SAR, China
h
School of Biological Sciences, University of Hong Kong, Hong Kong, SAR, China
b
a r t i c l e
i n f o
Available online 17 June 2011
Keywords:
Reporter gene assay
Agonist equivalents
Antagonist equivalents
Dibutyl phthalate
ATR-EQ20–80 range
Human risk
a b s t r a c t
Thyroid hormone disrupting compounds in water sources is a concern. Thyroid hormone (TH) agonist and
antagonist activities of water sources from the Yangtze River, Huaihe River, Taihu Lake and ground water in
the Yangtze River Delta region were evaluated by use of a TH reporter gene assay based on the green monkey
kidney fibroblast (CV-1). While weak TH receptor (TR) agonist potency was observed in only one of 15 water
sources, antagonist potency was present in most of the water sources. TR antagonist equivalents could be
explained by the presence of dibutyl phthalate (DBP), with concentrations ranging from 2.8 × 10 1 to
1.6 × 10 3 μg DBP /L (ATR-EQ50s). None of the ground waters exhibited TH agonist potencies while all of the
samples from Taihu Lake displayed notable TR antagonist potencies. To identify the responsible thyroid active
compounds, instrumental analysis was conducted to measure a list of potential thyroid-disrupting chemicals,
including organochlorine (OC) pesticides and phthalate esters. Combining the results of the instrumental
analysis with those of the bioassay, DBP was determined to account for 17% to 144% of ATR-EQ50s in water
sources. Furthermore, ATR-EQ20–80 ranges for TR antagonist activities indicated that samples from locations
WX-1 and WX-2 posed the greatest health concern and the associated uncertainty may warrant further
investigation.
© 2011 Elsevier Ltd. All rights reserved.
1. Introduction
Increasing attention has been given to the environmental contaminants which can disrupt the endocrine system in human and wildlife
(Colborn et al., 1993). Most research has focused on chemicals that can
modulate androgen and estrogen homeostasis (Kuster et al., 2010; Gracia
et al., 2008; Hill et al., 2010). In contrast, less information is available
regarding the compounds with thyroid disrupting activities (Jugan et al.,
2009). However, several contaminants from agriculture and industry,
such as pesticides and plasticizers, have been shown to exert toxic effects
on thyroid gland function, which also lead to adverse effects on growth
and development (Brucker-Davis, 1998; Darnerud et al., 2010) .
Thyroid hormone (TH) is a key molecule involved in regulating
growth, tissue differentiation, energy metabolism, reproduction, and
⁎ Corresponding authors. Tel.: + 86 25 8359 3649; fax: + 86 25 8368 6761.
E-mail addresses: hlliu@nju.edu.cn (H.-L. Liu), yuhx@nju.edu.cn (H.-X. Yu).
0160-4120/$ – see front matter © 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.envint.2011.05.013
formation of the central nervous system (Jugan et al., 2009). Normal
thyroid hormone levels are essential for mammals during development
of the central nervous system and disruption of normal hormone levels
can impair brain maturation, that results in permanent mental
retardation (Flamant and Samarut, 2003). TH is also important for
several aspects of reproduction for fish, including ovary maturation
(Blanton and Specker, 2007).
Synthetic chemicals that occur in the environment, such as
organochlorine (OC) pesticides, polybrominated diphenyl ethers
(PBDEs) and phthalate esters are potential endocrine disruptors by
modulating thyroidal system (Hofmann et al., 2009; Jugan et al., 2007; Li
et al., 2010). Thyroid hormone disrupting chemicals could interfere with
the thyroid receptor (TR) by agonism or antagonism. A transactivation
screening method for chemicals with TR ant/agonistic properties based
on CV-1 cells was used (Li et al., 2010; Shi et al., 2009). The anti/thyroid
hormone effects of bisphenol A (BPA), tetrachlorobisphenol A, carbaryl,
1-naphthol, 2-naphthol, DBP, mono-n-butyl phthalate (MBP) and DEHP
have been demonstrated (Shen et al., 2009; Sun et al., 2008; Sun et al.,
118
W. Shi et al. / Environment International 42 (2012) 117–123
2009). The present study used the CV-1 cell based TR gene reporter
assay to measure thyroid disrupting potentials of source water samples
in the Yangtze River Delta.
Chemical pollution has recently caused public concerns about
drinking water safety in the Yangtze River Delta region. The Yangtze
and Huaihe River, Taihu Lake and groundwater are sources of drinking
water in this region. The Yangtze River is the primary source of drinking,
serving more than 50 million residents in 8 large cities in Yangtze River
Delta. More than 40 chemical industrial complexes had been set up
along the river during the past decade (Shen et al., 2006). The Huaihe
River provides drinking water to more than 50 million people, but it is
one of the most densely populated rivers in China. Due to the rapid
development of industry and economy in the area, the Huaihe River has
been moderately polluted since the 1980s. Thirdly, Taihu Lake is an
indispensable water resource for drinking water, agriculture, aquaculture and industrial plants in the Yangtze River Delta region (Song et al.,
2007). In the past decades, the extraordinary economic growth,
industrialization, and urbanization, coupled with inadequate investment in basic water supply and treatment infrastructure, have resulted
in more contamination of the whole basin (Wu et al., 2004). More than
20% of the Yangtze River Delta population depends on ground water for
drinking water supply from either a public source or private wells.
Recent studies have suggested that groundwater quality could be
threatened by chemical pollution in this region (Chen et al., 2010). These
concerns on source water quality in the Yangtze River Delta warrant
comprehensive monitoring studies on chemical pollutants in source
water.
Previous studies have reported TH-active compounds in extracts of
environmental samples (Ishihara et al., 2009), however, limited
information is available for the contaminants that cause the effects.
PBDEs are well-known TH disrupters, but concentrations in surface
water are generally small. Although no concentrations of PBDEs in
Yangtze River Delta have been reported, studies have been employed in
the Zhujiang River Estuary, which is much more polluted by e-waste.
Concentrations ranged from 9.0 to 1.3 × 102 pg/L (Luo et al., 2008),
which are insufficient to cause the observed TR antagonist potency (Li
et al., 2010). We have previously reported that concentrations of
phthalate esters were 100- to 1000-fold greater than those of OC
pesticides, polychlorinated biphenyls and some phenols in source water
in East China (Shi et al., 2011). There is a growing concern to utilize the
mass balance analysis to these chemicals with high concentrations for
the identification of the responsible compounds (Li et al., 2010).
Deviations from parallelism between the dose–response curves of
reference chemical and samples cause uncertainty in the analysis of
toxic equivalency by bioassays (Villeneuve et al., 2000). In the present
study, ATR-EQ20–80 ranges for TR antagonist activities were employed in
the potency balance analysis to estimate the uncertainty associated with
the bioassay approach.
The objectives of the present study were to: 1) examine the
agonist and/or antagonist effects in primary water sources at the
Yangtze River Delta region by the transient reporter gene assays based
on the CV-1 cell line; 2) identify the responsible thyroid-active
compounds by combining instrumental analysis with bioassays. 3)
evaluate the uncertainty of the REP estimation by the employment of
ATR-EQ20–80 ranges.
2. Materials and methods
2.1. Chemicals and reagents
List, purity, abbreviation and source of analytical chemicals are given
in Table 1. 3-(4,5-dimethylthiazol-2-ol)-2,5-diphenyltetrasodium
bromide tetrazolium (MTT) and L-3,5,3′-triiodothyronine (T3) with
the purity of over 99% were purchased from Sigma Chemical Co. (St.
Louis, MO, USA).
Table 1
Sources and purities of tested chemicals.
Classes (providers)
Chemicals
Purity (%)
99.5%
γ-chlordane, α- chlordane, α-HCH,
β-HCH, γ-HCH, δ-HCH, p,p'-DDT,
o,p'-DDT, p,p'-DDD, p,p'-DDE
Phthalate esters (Labor Dr. Dibutyl phthalate (DBP), di-2-ethylhexyl
N 99%
phthalate (DEHP), dimethyl phthalate
Ehrenstorfer-Schafers,
(DMP), diethyl phthalate (DEP), benzyl
Germany)
butyl phthalate (BBP), diisodecyl phthalate
(DIDP), bis(2-ethylhexyl) adipate (DEHA),
di-n-octyl phthalate (DnOP), diisononyl
phthalate (DINP)
Organochlorine pesticides
(Sigma–Aldrich)
2.2. Sample collection and preparation
Untreated underground and surface water as sources of drinking
water in the Yangtze River Delta were studied. Sites were chosen in
areas that were known or suspected to have industrial, human and (or)
animal wastewater sources upstream or in the vicinity. Fifteen
waterworks were selected in the Yangtze River Delta, whose drinking
water daily outputs were more than 300,000 m 3. Water samples were
collected in March 2009 from Yangtze River at locations LYG-1, YC-1,
YC-2, XZ-1, YZ-4, Huaihe River at NT-1, NT-2, TZ-2, NJ-3, Taihu Lake at
SZ-4, WX-1, WX-2 and groundwater from locations XZ-12, XZ-3, XZ-7
(Fig. 1). Samples of water (15 L) were collected in a glass vessels precleaned and rinsed with methanol at each location (10 L for bioassay
and 5 L for chemical analysis). The water samples were transported and
stored at 4 °C pending extraction and analysis within 24 h.
Water samples were passed through Oasis cartridges (200 mg Oasis
HLB glass cartridge; Waters, Milford, MA, USA) under vacuum at a flow
rate of 6–8 mL/min. 2 L sample was passed through each column to
avoid over filtration. A series of 5 columns were used for bioassay and 2
columns were used for instrumental analysis for each water sample.
Cartridges were sequentially activated and conditioned with highpurity hexane (Merck, Darmstadt, Germany), dichloromethane (Tedia
Co. Ltd, Fairfield, OH, USA), acetone (Tedia Co. Ltd, Fairfield, OH, USA)
and methanol (Tedia Co. Ltd, Fairfield, OH, USA). Each cartridge was
eluted stepwise as follows: 10 mL hexane, 10 mL hexane: dichloromethane (4:1), followed by 10 mL acetone: methanol (1:1, v/v). All
eluates were evaporated by rotary evaporation (type TVE-1000, EYELA,
Tokyo, Japan) in a thermostatic bath. Then the dehydrated extracts were
blown to dryness under gentle nitrogen flow and reconstituted in
0.2 mL dichloromethane for chemical analysis. For the bioassays,
extracts were blown to dryness under a gentle nitrogen flow and
reconstituted in 0.2 mL of dimethyl sulfoxide (DMSO, BDH Laboratory
Supplies, UK). Extracts in DMSO were diluted with appropriate culture
medium to be equivalent to 12.5, 25, 50, 100 and 200 times greater the
original concentration in source water before bioassays with a final
solvent less than 0.5% (v/v). Blanks prepared with purified water were
used to exclude endocrine disrupting toxicity during the working
procedure, using the same procedure as for the environmental samples.
Extracts were stored at −20 °C.
2.3. Bioassay
All media used for the assay were prepared according to the original
protocol (Shi et al., 2009). Green monkey kidney fibroblast (CV-1) cells
which contain no endogenous receptors were obtained from the
Institute of Biochemistry and Cell Biology in Shanghai, Chinese Academy
of Science. CV-1 cells were routinely cultured in Dulbecco's modified
Eagle's medium (DMEM) (Sigma, St. Louis, MO, USA) supplemented
with 10% fetal bovine serum (FBS; Gibco, Invitrogen Corporation,
Carlsbad, CA, USA), 100 U/mL penicillin (Sigma) and 100 μg/mL
streptomycin (Sigma, St. Louis, MO, USA) in an atmosphere containing
5% CO2 at 37 °C. Cells were seeded into 48-well microplates and were
W. Shi et al. / Environment International 42 (2012) 117–123
119
Fig. 1. Map of the chosen water sources (LYG-1, YC-1, YC-2, XZ-1, YZ-4, NT-1, NT-2, TZ-2, NJ-3, SZ-4, WX-1, WX-2, XZ-12, XZ-3, XZ-7) from Yangtze River, Huaihe River, Taihu Lake
and groundwater.
transfected 12 h later with 0.25 μg Gal4 responsive luciferase reporter
pUAS-tkluc, 0.1 μg pGal4-L-TR using 2.5 μg Sofast TM transfection
reagent per well. After further 12 h incubation, the cells were exposed to
various concentrations of standard compounds and sample extracts
dissolved in medium for 24 h.
The 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide (MTT) assay was performed in parallel with the luciferase
induction assays to examine cytotoxicity caused by sample extracts
(Shen et al., 2009). CV-1 cells were plated in 96-well plates using
DMEM with 10% dextran-coated charcoal (DCC) serum. After 24 h
incubation, CV-1 cells were treated with vehicle or chemicals for 24 h.
Then MTT (5 mg/mL in PBS, sigma-Aldrich, St. Louis, MO, USA) was
added to each well. After an additional 4 h incubation at 37 °C,
absorbance was measured by a microplate reader (EL808, Bio-Tek,
Winooski, VT, USA) at 570 nM (Shen et al., 2009).
2.4. TR ant/agonist equivalents
Determination of thyroid receptor agonist potency equivalents
(TR-EQ) and antagonist potency equivalent (ATR-EQ) was based on
previous reports with modifications (Conroy et al., 2007; Urbatzka et al.,
2007; Villeneuve et al., 2000). The TR-EQ of the sample was calculated by
dividing the EC50 values for the TH by sample concentration factors that
produced the same bioassay response. ATR-EQi was reported as an
equivalent DBP concentration, defined as the concentration of DBP
divided by the sample concentration factor that produced an equivalent
(i%) depression in the bioassay response to 5 nM T3.
ATR-EQ max is the maximum response observed for the sample
expressed as % DBP max (Conroy et al., 2007; Urbatzka et al., 2007).
The first step is to fit an appropriate regression model to the dose–
response relationship. When maximal inhibition is achieved, the
assumption of equal efficacy can be evaluated. If maximal inhibition is
not achieved and cannot be tested at greater concentrations, the
extrapolation should be employed and potential uncertainty due to
unknown efficacy must be identified and discussed (Villeneuve et al.,
2000).
2.6. Fractionation and instrumental analyses
Samples were fractionated and analyzed with the methods previously described (Koh et al., 2005). Briefly, concentrated extracts for the
instrumental analysis were passed through activated Florisil (60–100
mesh size; Sigma Chemical Co., St. Louis, MO, USA) glass column
separately. OC pesticides were collected in the first column eluted with
100 mL of high-purity hexane and 80 mL hexane/dichloromethane
(4:1). Plasticizers were collected in the second column eluted with
150 mL acetone/ dichloromethane (1:1). OC pesticides and plasticizers
were quantified using a Thermo TSQ Quantum Discovery triplequadrupole mass spectrometer (San Jose, CA, USA) in multiple-reaction
monitoring (MRM) mode. The detection limits for OC pesticides and
plasticizers were 0.1 × 10 −1 ng/L. Recoveries of plasticizers were
between 90% and 120%. Recoveries of OC pesticides were between 85%
and 105%. The reported results were corrected with the recoveries of the
surrogate standards. None of the procedural blanks contained detectable
concentrations of target compounds.
2.5. Antagonism equivalent range
2.7. Potency balance
The antagonism equivalent range was determined by a previous
protocol with improvements (Villeneuve et al., 2000). Thyroid
receptor antagonist equivalent range in the bioassay (ATR-EQ20–80
range) was derived by 20%, 50%, and 80% DBP maximal inhibition
(ATR-EQ20, ATR-EQ50, and ATR-EQ80) to calculate antagonist equivalents for each response value.
ATR−EQ 20–80 range = ATR−EQ 20 to ATR−EQ 80 :
The relative potencies (REPs) of the standard compounds were
calculated by dividing the T3 (or DBP) EC50 by the test compound
EC50, which were indicated in this study and previously published
results (Table 2). Instrumentally derived T3 equivalents (T3-EQs) were
calculated by multiplying the concentrations of known T3 agonists,
such as DnOP, DEHP, and DBP by specific relative potencies (REPs-T3)
and summing the products for each agonist present in the sample (not
done on a per fraction basis) of interest. Instrumentally derived TR
120
W. Shi et al. / Environment International 42 (2012) 117–123
Table 2
TR agonistic potency refer to T3 (REPs-T3) and antagonistic potency refer to DBP (REPs-DBP) of the well-known thyroid hormone disrupting compounds.
TR agonistic potency
TR antagonistic potency
REPs-T3a
−4
DBP
DEHP
γ-HCH
DnOP
DiNP
a
1.3 × 10
1.8 × 10− 4
/
1.2 × 10− 4
/
References
REPs-DBPa
References
(Ghisari and Bonefeld-Jorgensen, 2009)
(Ghisari and Bonefeld-Jorgensen, 2009)
/
(Ghisari and Bonefeld-Jorgensen, 2009)
/
1.0
0.05
0.15
0.9
0.7
This study
This study
(Li et al., 2010)
(Ghisari and Bonefeld-Jorgensen, 2009)
(Ghisari and Bonefeld-Jorgensen, 2009)
REPs: Relative potencies, calculated by dividing the T3 (or DBP) EC50 by the test compound EC50.
antagonist activity equivalents (DBP-EQs) were calculated by summing the product of concentrations of individual congeners by their
respective relative potencies (REPs-DBP).
2.8. Statistical analyses
Data were reported as mean ± SD (n = 3). Triplicate wells were
dosed for each treatment in the bioassays. SPSS statistical software
(version 11, SPSS Inc., Chicago, Illinois) was employed for calculations.
Curve-fitting analyses were carried out with Microsoft Excel (USA,
Seattle, WA, USA). The normality of each sample set was assessed with
the Kolomogrov–Smirnov one-sample test before parametric analysis,
followed by Duncan's multiple comparisons when appropriate. The
level of significance was set at *p b 0.05 and ** p b 0.01.
3. Results
3.1. Cell viability and system creditability
None of the tested groups showed cytotoxicity alone or in the presence of 5 nM T3. T3
induced luciferase expression in a concentration-dependent manner (Fig. 2). Luciferase
activity was induced by T3 in the range of 10− 10 M/L to 10− 6 M/L, with maximal induction
of 89.38-fold relative to that of the vehicle control achieved at 10− 6 M/L T3. The typical
dose–response curves were obtained by treating CV-1 cells with increasing concentrations
of DBP and DEHP in 5 nM T3 (Fig. 3). No significant induction of luciferase was observed in
any of the solvent controls (data not shown). The assumption of equal efficacy was
evaluated based on the dose–response curve. The DBP concentration leading to halfmaximum activity with 5 nM/L T3 was 33 μM DBP /L. For DEHP, a maximal response was
not achieved and the minimal response of the sample is less than 80% maximum response.
The estimated 20, 50, and 80% maximum response of DEHP were 1.2× 10− 1, 6.9 and
4.0× 101 mM/L.
3.2. TR agonist activity
The sample extract from Taihu Lake WX-1 site was the only water of the fifteen
locations that exhibited weak TR agonist activity. The activity was observed in extract at
the maximal tested concentration (200 times the original concentration in the source
water), resulting in increase expression as 1.41 fold of control. The corresponding T3
equivalent for extract was 2.9 × 102 ng/L (Table 3).
YC-2, XZ-1, NT-1, NT-2, TZ-2, SZ-4, WX-1 and WX-2 with the maximal concentration
factors (200 times the original concentration in the source water) decreased luciferase
expressions to as low as 24.3%, 21.8%, 27.1%, 14.4%, 2.9%, 54.6%, 37.6%, 17.5% and 3.3% of
5 nM/L T3 activity, respectively. The corresponding TR antagonistic equivalent ranges for
the extracts were referred to DBP (Table 3). WX-1 and WX-2 in Taihu Lake caused
significant TR antagonist activities with the highest ATR-EQ50 referring to DBP as 1.6 × 103
and 8.8 × 102 μ M/L, respectively (Table 3).
3.4. Concentrations of contaminants
OC pesticides were detectable in waters from all the locations at concentrations
ranging from 1.5 × 101 to 1.8 × 102 ng/L. Generally, concentrations of OC pesticides in
source water from Yangtze River and Taihu Lake were greater than in other locations.
Concentrations of HCHs ranged from 3.6 to 8.5 × 101 ng/L. Concentrations of HCHs
from water sources in Yangtze River and Taihu Lake were generally greater than those
at other locations. The ratios of α-HCH to γ-HCH in water sources were generally more
than 1.0. However, the ratios at some locations from the Huaihe River and Taihu Lake,
such as XZ-1, YZ-4, SZ-4 and WX-1, were less than 1.0, which indicated the use of
Lindane rather than technical mixtures. It should be noted that Lindane has been
banned in China in 2009 and can still be detected in various locations (Gao et al., 2008;
Li et al., 2001; Shen et al., 2009). Concentrations observed in this study were
comparable with the concentration ranges found in previous detected the Yangtze
River source of drinking water (ND-4.8 × 101 ng/g), but greater than those detected in
water sources from Beijing (Li et al., 2010; Wu et al., 2009).
Concentrations of DDTs varied among sampling locations, ranging from 3.1 to
7.8 × 101 ng/L. The greatest concentrations of DDTs were observed in water sources from
NT-1 and NT-2 in Yangtze River. P, p'-DDT were the predominant DDT contaminants in
almost all the samples, accounting for over 30%. Ratios of (DDDs + DDEs)/total DDTs in the
water sources collected from different regions were generally close to or greater than 1.0,
and indicated that more than half of the DDTs were transformation products (Wong et al.,
2006). Although the use of DDTs has been banned in China since 1983, the residue in the
soil may enter the water sources via run-off or surface water. Concentrations of DDTs in
water sources were greater than those detected in Beijing water sources which were not
detectable (Li et al., 2010).
Γ-chlordane and α- chlordane were detectable in all source water. Γ-chlordane was
dominant in all the samples. Concentrations of Γ-chlordane in water sources at location
NJ-3 from Yangtze River were greater than those at other locations. The ratio of αchlordane to γ -chlordane in the source water from XZ-1, NT-1 and NT-2 was similar to
the ratio in the technical mixture (0.8) (Singh et al., 2007). This indicated industrial
input of chlordane in XuZo city and NaTo city. Production of chlordane has been banned
in China since 2009, and the detectable concentrations suggested the possible release of
these chemicals into the water sources via surface run-off.
Relative luciferase activity
(n-fold of control)
All water extracts from Taihu Lake exhibited TR antagonist potency that inhibited
luciferase activity in the presence of 5 nM/L T3, but none of the groundwater extracts
displayed the activity (Fig. 4). Nine of the extracts of fifteen water sources including YC-1,
100
T3
80
60
40
20
0
-11
Relative luciferase activity
(n-fold of control)
3.3. TR antagonist potency
25
DEHP+5nM T3
DBP+5nM T3
20
**
**
**
15
**
**
10
**
5
0
-8
-7
-6
-5
-4
**
-3
log concentration (M)
-10
-9
-8
-7
-6
-5
log concentration (M)
Fig. 2. Concentration-dependent luciferase activities in CV-1 cell line thyroid receptor
(TR) reporter gene assay treated with T3 (n = 3).
Fig. 3. Anti-thyroid hormone activities of dibutyl phthalate (DBP) and di-2-ethylhexyl
phthalate (DEHP) measured by the CV-1 cell line TR reporter gene assay. Dibutyl
phthalate and di-2-ethylhexyl phthalate were diluted as indicated. Results are
expressed as mean ± SD (n = 3). Significant differences were indicated by asterisks
(* p b 0.05 and ** p b 0.01).
W. Shi et al. / Environment International 42 (2012) 117–123
Table 3
Thyroid receptor agonist and antagonist equivalents derived from bioassays and
instrumental analysis.
LYG-1
YC-1
YC-2
XZ-1
YZ-4
NT-1
NT-2
TZ-2
NJ-3
SZ-4
WX-1
WX-2
XZ-12
XZ-3
XZ-7
Equivalents from bioassays
Equivalents from
instrumental analysis
TR-EQa
(ng T3/L)
ATR-EQ50sb
(ng DBP /L)
T3-EQsc
(ng T3/L)
DBP-EQsd
(ng DBP /L)
–
–
–
–
–
–
–
–
–
–
291
–
–
–
–
–
883.34
693.06
409.27
–
824.19
717.50
39.60
–
28.35
1625.57
883.34
–
–
–
b0.01
0.02
0.3
0.32
b0.01
0.36
0.38
0.01
b0.01
0.01
1.19
0.43
b0.01
b0.01
b0.01
0.39
154.14
439.19
395.9
0.54
1291.64
771.16
45.22
0.56
25.48
1712.9
916.1
0.75
0.4
0.9
– Activity was undetectable.
a
TR-EQ: Thyroid receptor agonist equivalents derived from reporter gene assays.
b
ATR-EQ50s: Thyroid receptor antagonist equivalents derived from reporter gene
assays.
c
T3-EQs: Instrumentally derived T3 equivalents, calculated by summing the product of
concentrations of Dibutyl phthalate (DBP), di-2-ethylhexyl phthalate (DEHP), di-n-octyl
phthalate (DnOP), diisononyl phthalate (DINP) and γ-HCH by their respective relative
potencies (REPs-T3).
d
DBP-EQs: Instrumentally derived TR antagonist activity equivalents, calculated by
summing the product of concentrations of DBP, DEHP, DnOP, DINP andγ-HCH, by their
respective relative potencies (REPs-DBP).
All of the phthalate esters were detected in water samples. DEHP, DEP and DBP were
the major phthalate esters accounting for more than 60% of total phthalate esters studied.
Concentrations of DBP and DEHP in water samples ranged from 8 × 10− 2 to 1.4 × 103 and
5 × 10− 2 to 5.6 × 103 μg/L, respectively. Elevated concentrations of DBP and DEHP were
observed in source water at YC-2, XZ-1, NT-1, NT-2, XZ-1 and XZ-2, which were almost
1000-fold higher than the concentrations of OC pesticides. Measurable concentrations of
DMP, DEP, BBP, DEHA, DnOP and DiNP were found in some samples from the chosen water
sources. Concentrations of DBP and DEHP observed in this study were comparable with the
concentration ranges found in raw drinking water in Southern California (up to 8.3 and
5.9 μg/L) and Yangtze River (up to 6.2 and 6.2 μg/L) (Loraine and Pettigrove, 2006; Wu et
al., 2009).
Concentrations of total T3-EQs for the water sources ranged from b 0.01 to 1.2 ng/L
(Table 3). Concentrations of DBP-EQs, instrumentally derived TR antagonist activity
equivalents, were from 3.9 × 10− 1 to 1.7 × 103 μg/L.
Relative luciferase activity
(n-fold of control)
30
12.5
121
4. Discussion
This study examined for the first time the possible occurrence of
TR agonists and antagonists in the primary water sources in the
Yangtze River Delta by means of the reporter gene assay. The chemical
pollutants in water source had less TR agonist-like effect but greater
TR antagonist-like effect. This is compatible with the results of
previous reports, which also indicated the common occurrence of TH
antagonist activities in China (Li et al., 2010).
Overall, a weak T3 response was detectable in water sources of one
out of fifteen locations when extracts were measured at concentrations
200-fold greater than the original concentration in the source water.
During the present study, extracts could not be tested at greater
concentrations to reach a maximal response (Villeneuve et al., 2000).
Previous studies indicated that, T3 can induce the transcription of the
TRβ gene in tadpole tail at the concentration of 3.1 × 10 2 ng/L (Hogan
et al., 2007; Zhang et al., 2006). This was at the same concentration as T3
equivalents at location WX-1 (2.9 × 102 ng/L). Although the experimental conditions were different, the results still indicated the thyroid
hormone disrupting potential of water sources in Taihu Lake. It is well
known that the lake is polluted by the industrial effluents, domestic
wastewater and agricultural runoff. WX-1 is located at the western part
of Wangyu River, which is the primary outlet for Taihu Lake. The
presence of TR agonist potency in this region was hypothesized based
on concentrations of contaminants from the whole lake. Concentrations
of the main contaminants at this location were reported (Table 4).
Concentrations of detected chemicals at location WX-1 were not
sufficient to cause the observed TR agonist activity measured by the
bioassay. Less than 1% of the total concentrations of TR-EQ in water
sources were contributed by the detected thyroid hormone disrupting
chemicals.
TR antagonist potencies were detected in most of the surface water
samples with the equivalents ranging from 2.8 × 10 1 to 1.6 × 10 3 μg/L
(refer to DBP). Previous work has studied the effect of DBP on T3dependent activation of TRβ gene in T3-induced metamorphosing
tadpoles. The TR antagonist response was detected at 1.1 × 10 3 μg/L
DBP (Sugiyama et al., 2005). The concentration of TR antagonistic
equivalents at location WX-1 was greater than this level, and thyroid
hormone disrupting potentials were indicated. For location WX-1 in
Taihu Lake, both TR agonist and antagonist potencies exceed the least
observed effect concentration, and thus this is not an appropriate
location for drinking water use. None of the ground water extracts
contained TR antagonists and more attention should be paid to the
25
50
100
200
25
*
20
*
**
15
**
**
**
**
**
**
**
**
10
5
3
X
Z7
XZ
-
12
X
Z-
SZ
-4
W
X
-1
W
X
-2
J3
N
TZ
-2
1
NT
-2
N
T-
Z4
Y
5
nM
T3
LY
G
-1
YC
-1
YC
-2
X
Z1
0
Fig. 4. Concentration-dependent TR antagonist activities in the water extracts measured by the CV-1 cell line TR reporter gene assay. Water extracts were tested at 12.5, 25, 50, 100
and 200 times the original concentration. Cells were exposed to extracts in parallel with 5 nM T3 as indicated by the dashed line. The TR antagonist activity was expressed as relative
expression versus the untreated cells (control) (mean ± SD). Significant differences between the extracts and the T3 treatment were tested using ANOVA, Dunnett's test. Significant
differences were indicated by asterisks (* p b 0.05 and ** p b 0.01). The results of statistical analysis at higher concentrations which also exhibited significant differences (**p b 0.01)
were not shown.
122
W. Shi et al. / Environment International 42 (2012) 117–123
in these locations. ATR-EQ20–80 ranges may be suitable for comparative purposes. However, they may limit the utility of the estimate for
risk assessment or potency balances.
In summary, reporter gene assays were utilized for the first time to
identify TR agonist and antagonist activities of the water sources from
the Huaihe River, Yangtze River, Taihu Lake and ground water in the
Yangtze River Delta. Our results suggest that TH antagonist potencies
were present in most of the detected water sources. None of the
ground water samples exhibited TR disrupting activities while all of
the detected samples in Taihu Lake contained measureable TR
antagonist potencies. It can be speculated that DBP is the primary
TR antagonist in water sources in the Yangtze River Delta, while DEHP,
DnOP and DiNP also contributed. For water from Yangtze River, the
equivalents calculated from instrumental analysis equivalents were
greater than the bioassay equivalents, which indicated that combined
toxicities or contaminants with proliferation effects were suspected
to be responsible for the observed phenomena. Furthermore, for the
very first time, ATR-EQ20–80 ranges for TR antagonist activities were
employed for comparative purposes. Water sources from TZ-2 and
SZ-4 posed the least TH antagonistic activities, while WX-1 and WX-2
posed the greatest potential risk of TH disruption. The results also
indicated the uncertainty in the REP estimate due to deviations from
parallelism between the standard and sample dose–response curves.
Integrated management practices and further treatment techniques
should be emphasized to prevent the entering of thyroidal disruptive
chemicals to water sources in the Yangtze River Delta.
water quality of the water sources in Huaihe River, Yangtze River and
Taihu Lake.
To identify the causality for TR antagonistic activities, TR
antagonist potencies of DBP and DEHP were compared to the results
of the reporter gene assay. DBP-EQs accounted for 89.9% to 105.4% of
the ATR-EQ50s derived from bioassays. More than 67% of the total
ATR-EQ50s concentrations in Taihu Lake were contributed by DBP. The
DBP-EQs derived from instrumental analysis accounted for 17.5% to
96.7% of the bioassay equivalents in water sources from Huaihe River.
Thus, other chemicals which were not measured in the present study
or the combined effects might be responsible for the agonistic
activities in Huaihe River. The DBP-EQs accounted for 107.5% to
156.7% of the ATR-EQ50s derived from bioassays in Yangtze River.
Previous studies indicated that BBP, DEHA could exhibit proliferation
effect in the presence of T3 by providing additional ligands that may
bind to TR (Ghisari and Bonefeld-Jorgensen, 2009). The presence of
these contaminants in Yangtze River may account for the observed
proliferation activities.
In the present study, ATR-EQ20–80 ranges for TR antagonist
activities were calculated from the dose–response relationships for
the very first time (Fig. 5). Because of the limited sample volumes,
extracts from water sources could not be tested at greater concentrations. In some cases they exhibited significant activity but did not
reach a maximal/minimal response, and significant extrapolation
should be used at this time. Furthermore, previous studies indicated
that dose–response curves for complex mixtures analyzed by the
same in vitro bioassay will not be parallel or exhibit equal efficacy to
the standard (Villeneuve et al., 2000). Because the extracts did not
result in equal efficacy with the standard curve used, the ATR-EQ50s
derived only from EC50 would tend to underestimate or overestimate
for the antagonist equivalents. ATR-EQ20–80 ranges indicates the
uncertainty in the REP estimate due to deviations from parallelism
between the standard and sample dose–response curves. Moreover
the ATR-EQ20–80 ranges could be utilized for comparison of risk
potential posed by different mixture. For example, the results
indicated that the TH antagonistic activities posed in water sources
at locations TZ-2 and SZ-4 has the least risk potential (Fig. 5). Water
sources at locations WX-1 and WX-2 posed the greatest toxicity
potential. Although the ATR-EQ50s of the samples were different from
each other, the ATR-EQ20–80 ranges of YC-1, YC-2, XZ-1, NT-1, and
NT-2 were similar, which indicates the comparable toxicity potentials
Acknowledgements
This work was funded by Major State Basic Research Development
Program (No. 2008CB418102), the National Program in Control and
Management of the Polluted Water Bodies (No. 2008ZX07101-003006), the Program for Postgraduates Research and Innovation in
Jiangsu Province (CX10B_022Z) and the Program for Postgraduates
Research and Innovation in Jiangsu Province (CX10B_022Z). Prof.
Giesy was supported by the Canada Research Chair program, an at
large Chair Professorship at the Department of Biology and Chemistry
and State Key Laboratory in Marine Pollution, City University of Hong
Kong, The Einstein Professor Program of the Chinese Academy of
Sciences and the Visiting Professor Program of King Saud University.
Table 4
Concentrations of thyroid hormone disrupting compounds (ng/L).
Chemicals
α-HCH
β-HCH
γ-HCH
δ-HCH
ΣHCH
α- chlordane
γ-chlordane
ΣChlordane
p,p'-DDT
o,p'-DDT
p,p'-DDD
p,p'-DDE
ΣDDT
DMP
DEP
DBP
BBP
DEHA
DEHP
DnOP
DiNP
Locations
LYG-1
YC-1
YC-2
XZ-1
YZ-4
NT-1
NT-2
TZ-2
NJ-3
SZ-4
WX-1
WX-2
XZ-12
XZ-3
XZ-7
3.81
3.34
1.21
5.42
13.77
3.17
5.70
8.88
3.30
3.58
11.70
2.41
20.99
0.06
1.87
0.16
0.44
0.13
0.11
0.02
0.03
3.38
16.12
0.91
0.06
20.46
2.94
6.50
9.43
1.77
0.80
12.84
4.53
19.94
2.34
2.63
152.75
22.53
26.11
3.67
0.45
0.95
5.23
3.91
2.09
2.77
14.01
1.19
2.54
3.73
1.68
2.71
6.64
2.21
13.24
213.70
81.56
315.93
196.81
28.71
1412.87
41.06
21.92
0.68
2.37
1.03
2.57
6.65
3.14
3.68
6.82
1.71
1.76
9.24
2.77
15.48
239.32
84.12
259.50
183.90
48.33
1571.78
49.10
19.24
0.82
3.31
2.83
4.69
11.65
1.61
4.96
6.57
1.72
1.96
7.16
1.42
12.26
0.09
0.48
0.08
0.16
0.11
0.05
0.01
0.02
9.94
30.19
2.43
41.95
84.51
6.09
8.21
14.31
4.22
17.72
47.02
9.05
78.01
189.34
106.80
1190.78
160.48
11.72
1097.47
39.14
14.84
10.80
16.85
2.79
6.58
37.02
12.91
13.60
26.51
5.12
12.76
19.65
7.77
45.30
688.85
242.22
541.76
861.60
48.49
1631.28
118.11
58.74
0.58
1.07
0.36
1.58
3.59
0.61
1.49
2.09
0.35
1.01
1.31
0.39
3.06
3.57
3.96
44.76
13.07
13.97
4.64
0.10
0.12
6.68
18.44
2.93
15.32
43.37
7.83
62.72
70.55
8.66
12.16
15.69
14.10
50.61
0.03
1.12
0.09
0.13
0.08
0.08
0.02
0.02
4.50
10.29
9.92
45.31
70.02
2.91
2.58
5.49
1.10
1.74
19.03
3.15
25.02
32.67
29.28
19.05
9.80
8.14
23.17
1.51
3.46
1.87
12.44
5.13
7.77
27.21
2.26
6.43
8.69
3.03
10.92
10.95
9.40
34.31
1451.12
706.31
1351.60
473.70
61.51
5578.12
60.52
38.50
15.98
5.62
6.45
4.32
32.37
4.23
11.98
16.21
3.07
5.96
21.95
9.49
40.47
293.17
144.00
753.01
374.16
26.83
1778.48
60.87
26.58
0.35
2.77
0.27
1.10
4.49
1.05
3.44
4.49
0.52
1.38
1.80
1.13
4.83
1.71
4.31
0.44
1.05
0.62
1.70
0.05
0.19
0.96
2.54
0.47
1.62
5.59
1.23
1.15
2.38
0.53
13.66
3.16
1.08
18.44
0.14
0.83
0.26
0.35
0.84
0.36
0.02
0.05
1.54
1.88
0.59
1.60
5.61
0.70
3.20
3.90
0.54
1.06
1.92
0.96
4.48
0.28
2.00
0.56
1.05
0.34
0.54
0.12
0.18
W. Shi et al. / Environment International 42 (2012) 117–123
100000
REP-20
REP-80
REP-50
REP-max
nM/L
10000
1000
100
10
1
YC-1
YC-2
XZ-1
NT-1
NT-2
TZ-2
SZ-4
WX-1 WX-2
Locations
Fig. 5. Thyroid receptor antagonist equivalent ranges (ATR-EQ20–80 ranges) for water
sources. ATR-EQ20, ATR-EQ50, and ATR-EQ80 refer to relative potencies calculated as a
ratio of potency estimates where the defined level of response was 20, 50, and 80%.
ATR-EQmax means the maximum magnitude of response the maximum magnitude of
response observed for the sample expressed as % DBP max. The ATR-EQ20–80 range
indicates the uncertainty in the ATR-EQ estimate due to deviations from parallelism
between the standard and sample dose–response curves.
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