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Thyroid Hormone Disruption by Water-Accommodated Fractions of
Crude Oil and Sediments Affected by the Hebei Spirit Oil Spill in
Zebrafish and GH3 Cells
Sujin Kim,† Ju Hae Sohn,† Sung Yong Ha,‡ Habyeong Kang,† Un Hyuk Yim,‡ Won Joon Shim,‡
Jong Seong Khim,§ Dawoon Jung,† and Kyungho Choi*,†
†
School of Public Health and §School of Earth and Environmental Sciences & Research Institute of Oceanography, Seoul National
University, Seoul 08826, Republic of Korea
‡
Oil and POPs Research Group, Korea Institute of Ocean Science and Technology (KIOST), Geoje 53201, Republic of Korea
S Supporting Information
*
ABSTRACT: A crude oil and the coastal sediments that were
affected by the Hebei Spirit Oil Spill (HSOS) of Taean, Korea
were investigated for thyroid hormone disruption potentials.
Water-accommodated fractions (WAFs) of Iranian Heavy
crude oil, the major oil type of HSOS, and the porewater or
leachate of sediment samples collected along the coast line of
Taean were tested for thyroid disruption using developing
zebrafish and/or rat pituitary GH3 cells. Major polycyclic
aromatic hydrocarbons (PAHs) and their alkylated forms were
also measured from the test samples. In zebrafish larvae,
significant decreases in whole-body thyroxine (T4) and
triiodothyronine (T3) levels, along with transcriptional
changes of thyroid regulating genes, were observed following
5 day exposure to WAFs. In GH3 cells, transcriptions of thyroid regulating genes were influenced following the exposure to the
sediment samples, but the pattern of the regulatory change was different from those observed from the WAFs. Composition of
PAHs and their alkylated homologues in the WAFs could partly explain this difference. Our results clearly demonstrate that
WAFs of crude oil can disrupt thyroid function of larval zebrafish. Sediment samples also showed thyroid disrupting potentials in
the GH3 cell, even several years after the oil spill. Long-term ecosystem consequences of thyroid hormone disruption due to oil
spill deserve further investigation.
1. INTRODUCTION
Oil spills along the coastal shoreline can cause serious
ecological and human health problems. The Hebei Spirit Oil
Spill (HSOS) accident which occurred in December 2007, near
Taean, Korea, is one of the largest oil spills in Korean history.
Many studies have assessed ecological and human health
damages of the HSOS accident along the Taean coastline.1−5
Petroleum products, e.g., crude oil, are extremely complex
mixtures and therefore may contain numerous toxic components. Polycyclic aromatic hydrocarbons (PAHs) or their
alkylated forms are considered to be among the primary
contributors to oil-related adverse effects.2,6 Most available
toxicological information on HSOS to date is limited to PAHassociated responses such as aryl hydrocarbon receptor (AhR)
binding affinity2 or cytochrome P4501A (CYP1A) induction.5,7
In addition, sex endocrine disruption and DNA damage
potentials of the oil spill have been documented from the
environmental samples of HSOS.3
Growing evidence indicates that crude oil and PAHs can
cause developmental toxicity in fish, e.g., cardiac malformation.8−12 Exposure to weathered Alaska North Slope crude oil
© 2016 American Chemical Society
led to changes in cardiac morphology in developing zebrafish.9,10 Similarly, cardiotoxicity syndromes were reported in
zebrafish embryos following exposure to Iranian Heavy crude
oil.13 Underlying mechanisms of such developmental changes
are not yet clear. However, thyroid hormone disruption could
be in part responsible for these adverse effects, as thyroid
hormones are important regulators of normal differentiation
and development of organs including heart and vascular
system.14−16 Indeed, in teleost fish, thyroid hormones such as
thyroxine (T4) and triiodothyronine (T3) have known to play
crucial roles in growth and development of an individual.17,18
Several studies have indicated that water-soluble components
of crude oil could disrupt thyroid hormone balances in fish,
including turbot (Scophthalmus maximus L.) and flounders
(Pleuronectes f lesus).19−22 Moreover, similar observations were
reported from avian species. In seabirds, e.g., black guillmot
Received:
Revised:
Accepted:
Published:
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February 11, 2016
April 27, 2016
May 4, 2016
May 4, 2016
DOI: 10.1021/acs.est.6b00751
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Environmental Science & Technology
Figure 1. Location of two sampling sites near Taean coast, Korea. Site A is located in Sogeunri mudflat and Site B in Sinduri beach. (Satellite photo
provided by Google Earth version 7.1.5. 2015. Image providers are shown in the bottom of each map: the upper left map was provided by Data SIO,
NOAA, U.S. Navy, NGA, GEBCO; Data Japan Hydrographic Association; Image Landsat. The other map was provided by Image Copyright 2016
TerraMetrics; Image Copyright 2016 CNES/Astrium.)
2. MATERIALS AND METHODS
2.1. Sample Collection and Preparation. WAFs were
prepared using Iranian Heavy crude oil following a standardized
method published elsewhere.25,26 Iranian Heavy crude oil is a
dominant type of the oils released from the Hebei Spirit.1,30
Briefly, 87.5 g Iranian Heavy crude oil was carefully added onto
the surface of a 5 L carboy filled with 3.5 L of dechlorinated
water (25 g oil/L water). The mixture was stirred gently for 24
h to avoid formation of oil droplets, after which the water phase
was siphoned out and was utilized. Prepared WAFs were sent
for chemical analysis and were immediately used for the
zebrafish exposure study.
Two types of environmental samples were collected in
September 2014 from two locations, i.e., Sogeunri mudflat
(latitude: 36° 48′46.47″, longitude: 126° 10′57.56″, henceforth
“Site A”) and Sinduri beach (latitude: 36° 50′20.95″, longitude:
126° 11′10.24″, henceforth “Site B”) along the coast of Taean
(Figure 1). The sampling locations were chosen based on
visible signs of oil residuals. From Site A, sediment samples
were collected, pooled, and transported to the laboratory under
cold condition. Upon delivery to the laboratory, the sample was
added with dechlorinated water (1 g/mL) and was vigorously
shaken for >12 h at 4 °C. Then the sediment pore water was
obtained by centrifugation at 1000g for 1 h at 4 °C. From Site
B, leachate from the sand was collected on site and moved to
the laboratory under cold condition. Both samples were stored
at 4 °C until the toxicological studies and chemical measurement.
2.2. Zebrafish Embryo/Larval Exposure. Fertilized
embryos of zebrafish (Danio rerio) were collected from 35 L
culture tanks with dechlorinated water, within 4 h postfertilization (hpf). Dechlorinated water has been used for maintaining
(Cepphus grille), herring gulls (Larus argentatus), and Leach’s
petrels (Oceanodroma leucorhoa), a single oral dosing to
petroleum products caused changes in plasma T4 level.23
Daily oral ingestion of crude oil resulted in increased plasma T3
and hepatic CYP levels in ducks (Anas platyrhynchos) as well.24
However, mechanisms underlying thyroid disruption of oil or
oil related compounds have not been fully understood yet.
In the present study, we investigated thyroid disrupting
potential of a crude oil and environmental samples that were
contaminated by the same crude oil spill. For this purpose, the
water-accommodated fractions (WAFs) of the crude oil, and
the pore water or leachate of the sediment samples which were
affected by HSOS accident, were employed. WAFs have been
recommended for studying short-term aquatic effects of poorly
soluble mixtures such as crude oil.25,26 These samples were first
tested for thyroid hormone disruption potentials using a rat
pituitary gland cell line (GH3). GH3 cells have been used for
screening thyroid related activities of chemicals.27−29 For in vivo
responses, zebrafish (Danio rerio) larvae were employed but
only for the WAFs of the oil, because of high salinity associated
with the near-coastal sediment samples. Through a combined
approach of zebrafish larvae and GH3 cells, possible effects of
the oil spill on thyroid hormone balance and its regulation via
hypothalamus-pituitary-thyroid (HPT) axis were investigated.
In addition, the test samples were analyzed for major PAHs and
their alkylated analogues, in order to link to the observed
thyroid hormone disruption. The result of this study will help
better understand the impacts of oil components on thyroid
system and development of aquatic organisms like fish.
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Table 1. Effects of WAF Exposure on Survival, Hatchability, Malformation Rate, and Wet Body Weight in Zebrafish at 120 hpfa
control
embryo survival (%)
hatchability (%)
larval survival (%)
malformation rate (%)
body weight (mg)
89.2
98.2
97.8
6.99
167
±
±
±
±
±
2.59
0.536
0.429
0.876
14.8
20% WAF
93.7
98.2
97.5
6.32
149
±
±
±
±
±
30% WAF
1.81
0.418
0.442
1.69
12.1
91.8
97.5
97.2
6.07
149
±
±
±
±
±
3.09
0.617
0.732
0.806
8.36
40% WAF
92.8
97.4
96.9
7.71
157
±
±
±
±
±
0.524
0.569
0.295
0.892
9.54
50% WAF
90.5
98.6
94.1
10.4
161
±
±
±
±
±
3.03
0.579
0.921
1.87
4.67
a
Embryo survival (%), the percentage of surviving embryos among total fertilized eggs; hatchability (%), percentage of hatchling among the live
embryos; larval survival (%), percentage of surviving larvae among the hatched; malformation rate (%), percentage of malformed individuals,
including dead fish; body weight (mg), wet weight of 150 zebrafish larvae per replicate. Results are shown as mean ± SEM of four replicates.
dissolved in a medium with dimethyl sulfoxide (DMSO, 0.1%
v/v). GH3 cell exposure was conducted following Kim et al.29
with some modifications. In brief, trypsinized GH3 cells from
culture plates were seeded into 24-well plates (Costar 3526) at
a density of 2.0 × 105 cells per well and incubated for 16 h.
After the growth medium was replaced with serum-free
medium, cells were incubated for additional 8 h. Then the
cells were dosed with serially diluted WAFs (0.0125, 0.025, and
0.05%), Site A pore water (0.00625, 0.0125, and 0.25%), and
Site B leachate (0.0125, 0.025, and 0.05%). For dilution,
dechlorinated water was employed. The 0.1% v/v dechlorinated
water was used as a control. For positive control, 0.023, 0.23, or
2.3 nM of T3 and 0.1% v/v DMSO were also dosed. The cells
were incubated in triplicates for each treatment for 48 h. At the
end of the exposure, RNA isolation, synthesis of cDNA, and
qRT-PCR for the GH3 cell experiment were conducted as
described elsewhere.29 Sequences and efficiencies of primers are
listed in Table S2. Cyclophilin was used as a housekeeping gene.
The exposure concentrations were determined at noncytotoxic
range (cell proliferation >80% of control), following a WST-1
cell proliferation assay (Roche Applied Science, Mannheim,
Germany; data not shown).
2.4. Analysis of PAHs by Gas Chromatography.
Concentrations of PAHs and alkylated PAHs were measured
in the 50% WAF, Site A pore water, and Site B leachate, using
an HP 5890 GC equipped with an HP 5972 MS (Agilent, Palo
Alto, CA, USA). Detection limits for target PAHs in the
samples ranged from 0.09 to 5.67 ng/L depending on their
physicochemical properties. The recovery efficiencies of PAHs
were generally comparable to those reported in our previous
works.34,35 List of target PAHs and detailed information are
shown in the Supporting Information.
2.5. Statistical Analysis. Statistics were carried out using
IBM SPSS Statistics (version 23.0; SPSS Inc., Chicago, IL,
USA). For zebrafish exposure study, normality and homogeneity of variances were evaluated by Shapiro-Wilk’s test and
Levene’s test, respectively. If data met the normality and
homogeneity assumptions, the difference among treatments
and control was analyzed by one-way analysis of variance
(ANOVA) followed by Dunnett’s test. Otherwise, a nonparametric Kruskal−Wallis test combined with Mann−Whitney
U test was used. Dixon’s Q test was performed for identification
of outliers, which were considered for exclusion before
statistical analysis. For GH3 cell exposure results, ANOVA
test followed by Dunnett’s test was performed. Logarithmic or
exponential data transformations were conducted when
necessary. In all statistical analyses, p values less than 0.05
were considered to be statistically significant. All data are shown
as mean ± standard error of mean (SEM).
fry and adult zebrafish in our laboratory. The eggs were
randomly placed into glass beakers containing 500 mL of 0, 20,
30, 40, or 50% WAF solutions with four replicates per group.
For each replicate, 240 eggs were placed. Dechlorinated water
was used for control and for dilution of WAFs. The exposure
concentrations of WAFs were chosen below a lethal level which
was determined based on a preliminary range finding test.
During 120 h of the exposure period, embryo and larval
survival, hatchability, and malformation rate were recorded
daily. Water quality parameters were measured at the beginning
and the conclusion of the exposure (Table S1 of the Supporting
Information). At 120 h, wet body weight (mg) of 150 larvae
was measured for each replicate and stored at −80 °C until
thyroid hormone measurement. For gene transcription analysis,
20 larvae were randomly chosen for each replicate and were
stored at −80 °C.
For thyroid hormone measurement, the zebrafish larvae (n =
150) per replicate were homogenized in 150 μL of sample
diluent, and supernatants were collected following Yu et al. with
some modifications (for details refer to the Supporting
Information).31 Whole body T4 and T3 were measured using
commercial enzyme-linked immunosorbent assays (ELISA) kits
(Uscnlif, Wuhan, China; Cat no. E0453Ge for T3; Cat no.
E0452Ge for T4) following the manufacturer’s instructions. All
the hormones were measured above the reported detection
limits (0.1 ng/mL for T3 and 1.2 ng/mL for T4).
For measurement of gene transcripts in zebrafish, 20 whole
body larvae were homogenized and total RNA was isolated
using Maxwell16 LEV simply RNA purification Tissue Kits
(Cat.#. AS1280, Promega, Madison, WI, USA). Complementary DNAs (cDNAs) were synthesized using iScript cDNA
synthesis kits (BioRad Hercules, CA, USA), and quantitative
real-time polymerase chains reaction (qRT-PCR) were carried
out with LightCycler 480 SYBR Green I Mastermix (Roche
Diagnostics Ltd., Lewes, UK) and LightCycler 480 instrument
(Roche Applied Science, Indianapolis, IN, USA). The primer
sequences and PCR efficiencies are listed in Table S2. The
transcription level of each target gene was normalized to that of
a housekeeping gene, 18s rRNA. The housekeeping gene was
chosen based on our preliminary study where this gene showed
the greatest stability compared to other candidate genes such as
β-actin, and elfa, in terms of the variability of Ct values. The 18s
rRNA gene has been successfully used in various zebrafish
exposure studies elsewhere.31,32 Target genes were quantified
by using the 2−ΔΔCT method.33
2.3. GH3 Cell Exposure. Transcriptional changes of
thyroid regulating genes were assessed with GH3 cells. For
GH3 cell exposure, stock solutions of WAFs, the pore water of
Site A, and the leachate of Site B were prepared in
dechlorinated water with the final concentration of 0.1% v/v.
T3 was chosen as a positive control for GH3 cell assay and was
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3. RESULTS
3.1. WAF Exposure in Zebrafish Embryos/Larvae.
3.1.1. Survival, Growth, and Developmental Effects. In
zebrafish embryos/larvae exposed to 20, 30, 40, or 50%
WAFs, no significant effects on survival, hatchability, and body
weight were observed until the conclusion of exposure (120 h)
(Table 1). However, among the exposure groups with higher
WAF concentrations, malformation with severe morphological
changes such as pericardial edema or yolk sac edema increased
slightly (Figure S1).
3.1.2. Whole-Body Thyroid Hormones. Exposure to WAFs
for 120 h caused significant decrease of the whole body thyroid
hormones in zebrafish larvae (Figure 2). T4 concentrations
gene by 4.8-fold, compared to the control. In addition,
transcription of uridine diphosphate glucuronosyltransferase
(ugt) gene was significantly increased by exposure to 40%
and 50% WAFs (Figure 3C). The transcription of cyp1a gene
was markedly increased in a concentration-dependent manner,
and the ahr2 gene was significantly up-regulated at 40% and
50% WAFs (Figure 3D).
3.2. WAF and Environmental Sample Exposure in GH3
Cells. In rat pituitary GH3 cells, concentration-dependent
down-regulation of tshβ gene was observed following exposure
to WAFs, and this pattern is similar to that observed by
exposure to T3 (Figure 4A). In contrast, the pore water of Site
A sediment (Sogeunri mudflat) and the leachate from Site B
(Sinduri beach) significantly up-regulated tshβ gene of GH3
cells. Both Sites A and B samples also up-regulated dio1 gene in
GH3 cells, unlike the changes observed from WAFs (Figure
4B). The transcription of dio2, thrα, and thrβ genes was not
significantly changed with an exception of the lowest
concentration of the Site A pore water where dio2 gene was
up-regulated (Figure S2).
3.3. Chemical Characterization of the WAF and
Environmental Samples. PAHs and alkylated PAHs
measured in the WAF (50%), Site A pore water, and Site B
leachate were different in levels and composition (Figure 5 and
Table S3). Total PAHs, calculated as a sum of PAHs and
alkylated PAHs, were the highest in Site A pore water (16,955
ng/L), followed by the 50% WAF (8,565 ng/L), and site B
leachate (672 ng/L). The PAHs with two benzene rings, such
as naphthalene (Na) and dibenzothiophene (DBT), and their
alkylated homologues were generally dominant in all three test
samples, perhaps because of their greater water solubility. In
WAFs, nonalkylated and C1−Na were dominant, while
concentrations of DBT compounds were very low or below
the detection limit. On the contrary, the pore water from Site A
was dominated by the alkylated DBT, whereas parent and alkylNa concentrations were relatively low. In the pore water sample
(Site A), alkylated homologues of three to four ring PAHs such
as fluorine, phenanthrene, and chrysene were also detected. For
the Site B leachate, although total PAH concentration was low,
relative composition of PAHs was similar to those of Site A,
e.g., with C2- and C3-DBT as dominant constituents.
Figure 2. Concentration (ng/g wet weight) of (A) T4 and (B) T3
measured in whole body zebrafish larvae at 120 hpf following exposure
to 0, 20, 30, 40, or 50% WAFs of Iranian Heavy oil. The results are
shown as mean ± SEM of four replicates (n = 4), and each replicate
includes 150 larvae. Asterisks (p* < 0.05) indicate significant
difference compared to dechlorinated water control.
4. DISCUSSION
4.1. Thyroid Disruption by WAFs in Zebrafish and
GH3 Cells. Significant decrease of thyroid hormone levels in
developing zebrafish by WAF exposure (Figure 2) clearly
demonstrates the thyroid endocrine disruption potential of the
crude oil. The regulatory changes of thyroid regulating genes in
GH3 cells (Figure 4) also support the thyroid disrupting effects
of the WAFs.
The significant decreases of whole body T4 and T3 contents
(Figure 2) may have implications in fish development, as
thyroid hormones play pivotal roles in cellular differentiation
and neurodevelopment.18,36,37 It is in agreement with a
previous report that observed reduced plasma T4 concentrations in flounder by the water-soluble fractions (WSFs) of
Omani crude oil.21 On the contrary, exposure to WSFs of a
crude oil (BP, Wytch Farm, Dorset, UK) for 6 h elevated whole
body T4 without alteration in T3 in turbot (Scophthalmus
maximus L.) larvae.20 These reports along with our present
observation suggest that oil components can disrupt thyroid
hormones, but the pattern of disruption may vary by species
and exposure duration.
were significantly decreased in zebrafish exposed to 40% and
50% WAFs (Figure 2A). Treatment with 50% WAF resulted in
significant decrease of whole-body T3 in the larvae (Figure 2B).
3.1.3. Regulation of Hypothalamic-Pituitary-Thyroid Axis
Genes in Zebrafish Larvae. Following exposure to WAFs,
several regulating genes of the HPT axis were affected in
zebrafish larvae (Figure 3). The thyroid stimulating hormone beta
(tshβ) and thyroid hormone receptor alpha (thrα) genes were
slightly up-regulated although the results were not statistically
significant (Figure 3A). Regulation of corticotrophin-releasing
hormone (crh), thyroid stimulating hormone receptor (tshr), and
thyroid hormone receptor beta (thrβ) genes were not altered.
However, the thyroglobulin (tg), NK2 homeobox1 (nkx2.1),
paired box protein 8 (pax8), and hematopoietically expressed
homeobox (hhex) genes were up-regulated following exposure
to WAFs (Figure 3B). 50% WAFs led to up-regulation of
transthyretin (ttr) gene by 20-fold and deiodinase type 1 (dio1)
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Figure 3. Gene transcription of (A) crh, tshβ, tshr, thrα, thrβ, (B) tg, nkx2.1, pax8, hhex, (C) dio1, dio2, ugt, ttr, and (D) ahr2 and cyp1a in zebrafish
larvae at 120 hpf following exposure to 0, 20, 30, 40, or 50% WAFs of Iranian Heavy oil. The results are shown as mean ± SEM of three to four
replicates (n = 3−4). Asterisks (p* < 0.05) indicate significant difference from those of the dechlorinated water control.
Figure 4. Gene transcriptions of (A) tshβ and (B) dio1 in GH3 cells following exposure to different concentrations of WAF, Site A pore water
(Sogeunri mudflat), or Site B leachate (Sinduri beach). T3 was used as a positive control. The results are shown as mean ± SEM of three replicates
(n = 3). Asterisks (p* < 0.05) indicate significant difference from the control, e.g., 0.1% v/v dechlorinated water or DMSO.
zebrafish larvae, tshβ gene showed up-regulating pattern
following WAFs exposure (Figure 3A), suggesting compensatory efforts of the zebrafish pituitary gland in response to
decreased thyroid hormones (Figure 2). Several other genes
that stimulate synthesis of thyroid hormones or development of
thyroid follicle, such as tg, nkx2.1, pax8, or hhex, were also up-
Significant down-regulation of tshβ gene in GH3 cells by
WAFs and T3 exposure (Figure 4A) indicates that WAFs may
act in the same way as T3 in the pituitary gland, i.e., downregulation of tshβ gene expression in the pituitary cells. TSH
released from pituitary gland stimulates thyroid gland and
stimulates synthesis of thyroid hormones.17 Interestingly, in
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Figure 5. Concentrations of 18 PAHs and alkyl-PAHs in 50% WAF of Iranian Heavy crude oil, Site A (Sogeunri mudflat) sediment pore water, and
Site B (Sinduri beach) leachate. Number of benzene rings were represented on the upper side of graph.
euthyroid state.45,47 In our present study, dio1 gene transcription observed in GH3 cells (Figure 4B) was different from
its transcriptional changes observed in whole body zebrafish
larvae (Figure 3C). This discrepancy may be attributed to lack
of feedback in the rat pituitary cell line, which may not reflect
regulatory feedback efforts that could be seen in vivo. In
addition, differential expression of those genes by organs should
be considered. The dio2 gene is predominant in the pituitary
gland, while the dio1 gene is abundantly expressed in liver,
kidney, and thyroid.47−49
Significant up-regulation of cyp1a gene in the fish larvae
following the exposure to WAFs (Figure 3D) is in line with our
expectation, as CYP1A induction has been considered as an
indication of PAH exposure.50,51 The ahr2 gene was upregulated, although the extent of increase was relatively smaller.
Between two orthologs of AhR in fish, namely AhR1 and AhR2,
ahr2 transcripts are more abundant and seem to be functionally
important in xenobiotic metabolism.52−54 These transcriptional
changes of cyp1a and ahr2 can be supported by chemical
composition of the WAFs with low molecular PAHs such as Na
and fluorene being dominant (Figure 5).
Taken together, our observations suggest that WAFs of crude
oil can lower thyroid hormone levels in zebrafish larvae,
possibly through hepatic UGT induction. However, interpretation of the results from this short exposure duration (120 h)
warrants caution. Because WAFs are not generally recommended for studies involving longer term exposure, we could
not continue the exposure long enough to understand longer
term consequences of the oil pollution. Study design that
involves long-term exposure to oil spill, and also includes time
series observations on gene expression, hormones, and thyroid
histology, deserves further consideration.
regulated (Figure 3B), implying compensatory efforts of
stimulating thyroid hormone synthesis in zebrafish.
The reason for decreased whole body thyroid hormones in
fish (Figure 2) is not clear but could be found from upregulation of ugt gene in zebrafish which was observed after
exposure to 40% and 50% WAFs (Figure 3C). UGT enzyme
plays a critical role in inactivation and excretion of many
exogenous and endogenous compounds, including T4 .
Induction of UGT may enhance glucuronidation and facilitate
elimination of T4, finally leading to decreased whole body T4
contents in zebrafish. Several reports indicated that induction of
UGT may possibly lead to decreased T4 levels.38,39 In
thyroidectomized male rats, serum thyroid hormones could
be reduced through UGT induction mechanism.38 Also in
zebrafish, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), one of
the most potent inducers of CYP1A, was reported to increase
hepatic UGT and result in decreased T4.40,41
Up-regulation of ttr gene (Figure 3C) may reflect a
compensatory effort against low thyroid hormone levels (Figure
2). TTR is an important T3 binding protein in teleost fish and
amphibia42,43 and is responsible for transportation of the
hormones to the target peripheral tissues. At the same time,
binding to TTR can delay metabolic elimination of free thyroid
hormones from the circulation.
Up-regulation of dio1 gene in zebrafish following exposure to
WAFs (Figure 3C) may be also seen as compensation against
the decrease in the T3 level. Among two types of deiodinases,
i.e., DIO1 and DIO2,44 DIO1 especially in liver plays an
important role in conversion of T4 to T3.45 In tilapia, upregulation of hepatic dio1 gene transcription was reported when
the fish were under hypothyroidism.46 Meanwhile, DIO2 plays
a role in catalytic conversion of T4 to active T3, generally in
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4.2. Potential Ecotoxicological Impact of Oil Spill. Two
environmental samples collected from the locations affected by
HSOS, e.g., Site A sediment pore water (Sogeunri mudflat) and
Site B leachate (Sinduri beach), showed clear up-regulation
patterns of major thyroid genes such as tshβ and dio1 gene in
GH3 cells (Figure 4), suggesting thyroid disrupting effects of
environmental samples. However, the regulatory changes in
GH3 cells from these environmental samples were different
from those observed in WAFs. Why are the responses observed
from oil spill affected area different from those observed in
WAFs? First, as the sampling areas are located near a city and a
popular beach, we cannot rule out the possibility of
anthropogenic contamination in our environmental samples,
not directly originated from the oil spill. On the other hand,
effects of weathering which might lead to clear difference in
composition of PAHs between WAFs and the environmental
samples (Figure 5) can in part explain the observed difference
of thyroid related gene regulations. In our present study, the
WAFs of Iranian Heavy crude oil contained higher concentrations of parent and alkyl-Na, while the environmental
samples collected from Sites A and B showed greater
proportions of alkyl-DBT, and other heavier PAHs. The
weathering process such as evaporation, dissolution, microbial,
and photochemical degradation generally shifts the PAH
composition to the dominance of heavier and more alkylated
compounds.13,55 It should be noted that our knowledge on
thyroid disrupting effects of low molecular weight PAHs is
quite limited. Only a few studies reported thyroid disrupting
potentials of PAHs with more rings, such as anthrancene,
phenanthrene, pyrene, and benzopyrene56 or by some
hydroxylated homologues.57 Further investigations on thyroid
disrupting effects of PAHs and their alkylated analogs are
warranted.
The results of our study clearly show that an oil spill can
cause thyroid hormone disruption in fish. Developing zebrafish
and GH3 cells could be used to understand the mechanisms of
the adverse effects. Screening areas with thyroid disrupting
potential and following up long-term consequences of thyroid
hormone disruption along coastal ecosystem warrant further
studies.
■
of Korea. We would like to thank NeoEnBiz Co. (Bucheon,
Korea) for assistance in sediment sampling.
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ASSOCIATED CONTENT
S Supporting Information
*
The Supporting Information is available free of charge on the
ACS Publications website at DOI: 10.1021/acs.est.6b00751.
Additional information, including analytical methods for
hormones, genes, and PAHs, and major results such as
morphology of fish, gene regulation of the cell, and the
levels of PAHs in the samples (PDF)
■
REFERENCES
AUTHOR INFORMATION
Corresponding Author
*Phone: 82-2-880-2738. Fax: 82-2-745-9104. E-mail:
[email protected].
Notes
The authors declare no competing financial interest.
■
ACKNOWLEDGMENTS
This study was supported by the “Oil Spill Environmental
impact Assessment and Environmental Restoration
(PM56951)” funded by the Ministry of Oceans and Fisheries
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