Arch. Environ. Contam. Toxicol. 40, 151–160 (2001)
DOI: 10.1007/s002440010158
A R C H I V E S O F
Environmental
Contamination
a n d Toxicology
© 2001 Springer-Verlag New York Inc.
In Vitro Bioassay Determination of Dioxin-Like and Estrogenic Activity in
Sediment and Water from Ulsan Bay and Its Vicinity, Korea
J. S. Khim,1,2 K. T. Lee,1 D. L. Villeneuve,2 K. Kannan,2 J. P. Giesy,2 C. H. Koh1
1
School of Earth and Environmental Sciences (Oceanography Program), College of Natural Sciences, Seoul National University, Seoul 151-742,
Korea
2
National Food Safety and Toxicology Center, Department of Zoology, and Institute for Environmental Toxicology, Michigan State University,
East Lansing, Michigan 48824-1311, USA
Received: 11 June 2000/Accepted: 28 August 2000
Abstract. Extracts of sediment and water samples collected
from Ulsan Bay, Korea, were screened for their ability to
induce dioxin-like and estrogenic gene expression in vitro.
Each sample was tested as raw extract (RE) and fractionated
extract (FE). Based on the initial screening of RE, 23 of 31
sediment samples showed significant dioxin-like activity in
H4IIE-luc bioassay, whereas most sediment samples did not
elicit estrogenic response in MVLN bioassay. Most of the
activities associated with FE samples revealed that mid-polar
(F2) and most polar (F3) fractions were responsible for the
significant reporter gene expression in H4IIE-luc bioassay. The
results suggest that complex interactions may have depressed
the activities of the known arylhydrocarbon receptor (AhR)
agonists present in F1 samples. The F2 samples were the most
active fraction. All F2 samples except one induced significant
dioxin-like activity, and over half of the F2 samples induced
significant estrogenic activity. Ten of the F2 samples produced
magnitudes of response in H4IIE-luc bioassay similar to those
induced by a 2,3,7,8-tetrachlorodibenzo-p-dioxin standard.
Sediment associated with F2 samples was estimated to contain
24.9 – 826 pg TCDD-EQ/g DW. Based on a qualitative mass
balance analysis, polycyclic aromatic hydrocarbons (PAHs)
appeared to account for both the estrogenic and dioxin-like
responses observed. Over half of the F3 samples were either
cytotoxic or caused morphological changes in both H4IIE-luc
and MVLN cells. Known concentrations of alkylphenols and
bisphenol A were not great enough to account for both the
estrogenic response and cytotoxicity observed for F3 samples.
Despite the apparent toxic or stressful effects, most of F3
samples induced significant dioxin-like activity in vitro, adding
to a growing body of evidence that suggests the presence of
unidentified, relatively polar, AhR agonists in sediment from
some areas.
Correspondence to: C. H. Koh
Contamination of aquatic environments is of particular concern
because food resources and drinking water come from surface
waters and coastal areas in many countries such as Korea,
where seafood is one of the major sources of diet. The area
adjacent to Ulsan Bay, Korea, is one of Korea’s most highly
industrialized regions. In 1985, for example, over 10,000 people suffered from diseases attributed to exposure to unknown
industrial pollutants (Kang et al. 1999).
Persistent halogenated organic contaminants, such as polychlorinated biphenyls (PCBs) and certain polycyclic aromatic
hydrocarbons (PAHs), have been shown to elicit a wide variety
of adverse effects in fish and other organisms, which are
mediated by an aryl hydrocarbon receptor (AhR)– dependent
mechanism of action (Safe 1990; Giesy and Kannan 1998; Van
den Berg et al. 1998; Willet et al. 1997; Clemons et al. 1998).
In addition to dioxin-like compounds, there has been increasing
concern in recent years over the compounds that mimic or
disrupt hormones in the body (Colborn et al. 1993; Dibb 1995;
McLachlan and Arnold 1996). Alkylphenols (APs), such as
nonylphenol (NP) and octylphenol (OP), are degradation products of alkylphenol ethoxylates (APEs) are a class of weakly
estrogenic compounds released into the environment since the
1940s (White et al. 1994; Nimrod and Benson 1996). Bisphenol A (BPA), a synthetic chemical widely used in polycarbonate manufacture, has also been shown to elicit estrogenic
effects in vitro and in vivo (Krishnan et al. 1993; Steinmetz et
al. 1997).
Instrumental analysis was used to evaluate the concentrations of many of the above compounds in samples from Ulsan
Bay and its vicinity, Korea (Khim et al. 2000b). Instrumental
analysis alone provides little information regarding the integrated biological potency of the samples. Furthermore, instrumental analysis alone cannot identify novel chemicals, which
may act independently or interact with known agonists to
influence the biological potency of the mixture of organic
contaminants present in the environment. Thus, this study
employed two mechanism-specific in vitro bioassays to support
and complement analytical characterization of samples from
the Ulsan Bay area. Recombinant rat hepatoma cells (H4IIEluc, Sanderson et al. 1996) with a stably transfected luciferase
152
reporter gene under control of dioxin-responsive elements
(DREs) were used to screen Ulsan Bay samples for their
capacity and potency to elicit dioxin-like responses in vitro.
Recombinant MCF-7 human breast cancer cells (MVLN cells;
Demirpence et. al. 1993) with a luciferase reporter gene under
control of estrogen-responsive elements (EREs) were used to
examine the overall estrogenic potency of samples.
Where appropriate, bioassay-directed fractionation and mass
balance analyses were utilized to identify the causative agents
responsible for in vitro bioassay responses observed. Bioassay
results were compared to predicted responses, based on published relative potencies for the target compounds and the
analytical results (Khim et al. 2000b). The use of bioassaybased toxicity identification and evaluation (TIE) and mass
balance approaches was important, as the sediment extracts
may contain a myriad of potentially bioactive compounds,
which were not analyzed for using instrumental methods
(Khim et al. 1999a, 1999b, 1999c, 2000a). Unlike previous
studies, which focused solely on sediments, we analyzed several environmental compartments, including sediment, pore
water (PW) , suspended particulate matter (PM), and dissolved
fraction (DF).
Materials and Methods
Sample Preparation
Detailed description of the sample collection, extraction, and fractionation procedures have been provided elsewhere (Khim et al. 2000b).
Each sample was tested as raw extracts after Soxhlet extraction and
fractionated samples after florisil column. All standards and samples
were prepared in high-purity hexane and/or acetonitrile (ACN) (Burdick & Jackson, Muskegon, MI) prior to dosing cells.
Cell Culture and Bioassay
Culturing conditions and assay procedure for H4IIE-luc and MVLN
cells have been described previously (Khim et al. 1999a). In brief,
cells for bioassay were plated into the 60 interior wells of 96-well
culture plates (250 ␮l per well) at a density of approximately 15,000
cells per well. Cells were incubated overnight prior to dosing. Test and
control wells were dosed with 2.5 ␮l of the appropriate extract or
solvent. Blank wells received no dose. At least three replicate wells
were analyzed for each sample dilution, control, and blank tested.
Luciferase and protein assays (Kennedy and Jones 1994; Khim et al.
1999c) were conducted after 72 h of exposure. Briefly, culture media
was removed by vacuum manifold, and the cells were rinsed with
phosphate-buffered saline. Cells were lysed, and luciferase assay reagent (containing a luciferin substrate) was added to the wells. Plates
were incubated for 5 or 10 min (depending on method) at 30°C, then
scanned with an ML3000 microplate reading luminometer (Dynatech
Laboratories, Chantilly, VA). Following the luminometer scan a 1.08
mM solution of fluorescamine (Sigma) in ACN was added to each well
and plates were assayed for protein, using a Cytofluor 2300 (excitation
400 nm, emission 460 nm), after a 15-min incubation at room temperature. Detailed methods for the H4IIE-luc and MVLN in vitro
bioassays have been described elsewhere (Khim et al. 1999a, 1999c).
J. S. Khim et al.
Bioassay Data Analysis
Protein content per well was calculated by regression against a bovine
serum albumin standard curve. Total protein per well was used as an
index of cell number to detect outliers that were not apparent by visual
inspection. Relative luminescence units (RLUs) were not adjusted for
protein. Sample responses, expressed as mean RLU over three replicate wells, were converted to relative response units, expressed as a
percentage of the maximum response observed for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD; %-TCDD-max) or 17-␤-estradiol (E2;
%-E2-max) standard curves generated on the same day (Khim et al.
1999, 1999c; Villeneuve et al. 2000). For screening purposes, significant responses were defined as those outside the range defined by
three times the standard deviation (expressed in %-standard-max) of
the mean solvent control response (0%-standard-max). Where appropriate, sample potency relative to the TCDD or E2 standard was
estimated. Relative potencies (REPs) were expressed as a range of
values calculated over multiple levels of response from 20 – 80%standard-max. (REP20 – 80-ranges) in order to account for potential
uncertainty in the estimated due to deviations from parallelism to the
standard curve (Villeneuve et al. 2000).
Mass Balance Analysis
Mass balance analysis, sometimes referred to as potency balance
analysis, is an approach that is commonly used to examine whether or
not the known composition of a sample, as identified by instrumental
analysis, can account for the magnitude or potency of biological
response observed (Sanderson and Giesy 1998). This study used two
types of mass balance analysis to aid in discussion and hypothesis
generation regarding the probable causes of estrogenic and dioxin-like
activity associated with samples from the Ulsan Bay.
The first approach employed was potency-based. Instrumentally
derived dioxin equivalents (TEQs) or estrogen equivalents (EEQs)
were calculated based by multiplying the concentrations of known
AhR or ER agonists (i.e., NP, OP, BPA, certain PAHs) by their
assay-specific REP values and summing the products for each agonist
present in the sample or fraction of interest (Villeneuve et al. 1998;
Khim et al. 2000c). Where it was possible to obtain sample doseresponse relationships by testing samples at multiple levels of dilution,
bioassay-derived dioxin equivalents (TCDD-EQs) or estrogen equivalents (E2-EQs) were derived. These were estimated directly from
sample dose-response curves using methods described elsewhere
(Khim et al. 1999c; Villeneuve et al. 2000). Instrumentally derived
values (TEQs or EEQs) were then compared to the bioassay-derived
relative potency estimates (TCDD-EQs or E2-EQs).
The second approach applied was magnitude-based mass balance
analysis. This approach was used in situations where a dose-response
relationship could not be generated for a sample, but there was sufficient analytical information to derive TEQs or EEQs. In this case,
TEQs or EEQs were assumed to behave exactly as if they were TCDD
or E2, respectively. Based on this assumption, regression of TEQs or
EEQs against the appropriate standard curve was used to predict the
magnitude of bioassay response that should have been elicited by the
known agonist/antagonist composition of the sample. Predicted magnitudes of response were then compared to magnitudes of response
observed for nondiluted sample in order to generate hypotheses regarding the potential contribution of the known agonists/antagonists to
the bioassay responses observed. Magnitude based mass balance analysis has been employed previously (Khim et al. 1999c). Significant
differences between observed and predicted response magnitudes were
defined as those, which differed by greater than three times the mean
standard deviation of the solvent control response (expressed as
%-TCDD-max, %-E2-max).
In Vitro Bioassay for Dioxin-Like and Estrogenic Activity
153
Table 1. Screening result of H4IIE-luc (dioxin responsive) and MVLN (estrogen responsive) in vitro bioassay
H4IIE-luc Bioassay
Sample Type
Sediment
Pore water
Particulate matter
Dissolved
fraction
MVLN Bioassay
No. of
Samples
No.
Tested
Significant
Inductiona
Cytotoxicity
or Stressedb
No.
Tested
Significant
Induction
Cytotoxicity
or Stressed
31
17
18
31
17
17
23
0
7
9
0
0
31
17
18
2
6
2
12
0
1
16
16
0
0
16
6
0
a
Significant induction means above the 3 standard deviations (expressed in %-TCDD-max) of the mean solvent control response (set to 0%TCDD-max)
b
Indicates cells exhibited an altered or “stressed” morphology
Fig. 1. Luciferase induction in the H4IIE-luc (dioxin-responsive) and MVLN (estrogen-responsive)
cell bioassay elicited by Ulsan Bay sediment, pore
water (PW), water samples (DF, dissolved fraction;
PM, particulate matter) Response magnitude presented as percentage of the maximum response
observed for a 2,000 pM 2,3,7,8 tetrachlorodibenzo-p-dioxin standard (%-TCDD-max)
Results
Dioxin-Like Activity
Dioxin-like activity was primarily associated with sediment
and particulate matter samples (Table 1, Figure 1). Raw
extracts (REs) for 23 of the 31 sediment samples tested
elicited a significant response in the H4IIE-luc bioassay
(Table 1). Seven of the 17 PM REs induced significant
luciferase activity (Table 1). None of the PW (n ⫽ 17) or DF
(n ⫽ 16) REs caused a dioxin-like response. The greatest
magnitude of activity observed for sediment REs was 42%TCDD-max. The mean activity (⫾ SD) was 22 ⫾ 8%TCDD-max. REs of PM were similarly active, with a mean
activity (⫾ SD) of 18 ⫾ 11%. Nine of the sediment RE
samples caused noticeable growth inhibition over the 72-h
exposure period (Table 1), but none of the other RE samples
were overtly toxic to the H4IIE-luc cells.
REs of sediment, which produced a significant response in
the H4IIE-luc bioassay, were separated into three fractions for
additional characterization. Fraction 1 (F1) samples, which
were known to contain PCBs and DDE, elicited relatively low
magnitudes of response (Figure 2). Fraction 2 (F2) samples
caused much greater magnitudes of induction relative to TCDD
(Figure 2). Whereas the maximal activity observed for sediment REs was 42%, sediment F2 samples elicited responses as
great as 109%-TCDD-max. Eighteen of the 22 F2 samples
tested yielded greater magnitudes of response, relative to
TCDD, than the corresponding RE (Figure 2). Eleven of the 18
yielded responses were at least double that observed for the
RE. Only two of the F2 samples caused noticeable growth
inhibition/toxicity to the H4IIE-luc cells. Fraction 3 (F3) sam-
154
J. S. Khim et al.
Fig. 2. Luciferase induction in
the H4IIE-luc (dioxin-responsive)
cell bioassay elicited by Ulsan
Bay sediment raw extracts (RE),
Florisil fraction 1, 2, 3 (F1, F2,
F3). Response magnitude presented as percentage of the maximum response observed for a
2,000 pM 2,3,7,8 tetrachlorodibenzo-p-dioxin standard (%TCDD-max). Πindicates cells
exhibited an altered or “stressed”
morphology; ● indicates the sample was toxic to the cells
ples also induced relatively high magnitudes of dioxin-like
activity (Figure 2). Sixteen of the 22 F3 samples tested yielded
greater responses than the corresponding RE, with 13 of the 16
at least doubling the response (Figure 2). Nine of the F3
samples were cytotoxic.
Eleven sediment F2 samples yielded responses greater than
65%-TCDD-max (Y4, Y5; J1, J2, J3, J4; U1, U6, U7, U9, U16),
and 12 F3 samples that yielded responses greater than 55%TCDD-max (Y1, Y2, Y4, Y5; J2, J3, J4; U6, U7, U9, U13, U16)
were selected for full dose-response characterization and mass
balance analysis. Six dilutions were tested and REPs, expressed as
bioassay-derived TCDD equivalents (TCDD-EQs), were estimated based on the resulting dose-response relationships (Villeneuve et al. 2000). TCDD-EQ estimates for the F2 samples ranged
from approximately 0.02– 0.83 ng TCDD-EQ/g DW (Table 2).
Deviations from parallelism to the TCDD standard curve yielded
some uncertainty in the TCDD-EQ estimates over the range of
response from 20 – 80%-TCDD-max (Table 2). Even considering
the potential ranges of uncertainty in the bioassay-derived
TCDD-EQ and instrumentally-derived TCDD equivalent (TEQ)
estimates calculated based on the known concentration of PAHs
(TEQPAH) and their H4IIE-luc-specific REPs (Khim et al. 2000c),
bioassay-derived TCDD-EQ were generally one to two orders of
magnitude greater than TEQPAH (Table 2). The range of
TCDD-EQ for F3 samples was approximately 0.007– 0.27 ng
TCDD-EQ/g DW. No known AhR agonists were quantified in F3
samples, thus TEQ estimates could not be derived for F3 samples.
Estrogenic Activity
Overall, relatively few of the Ulsan Bay RE samples induced
a significant response in the MVLN bioassay. Only 2 of the
31 sediment REs and 2 of the 18 PM REs samples caused
significant responses (Table 1, Figure 1). Twelve of the
sediment REs were overtly toxic to the cells, however. Six
of the 17 PW and 6 of the 16 DF REs elicited a significant
estrogenic response (Table 1). Estrogenic pore waters
and surface waters were associated with rivers in the Yeocheon, Jangsaengpo, and Taehwa areas near Ulsan Bay. No
water samples from the outer bay induced an estrogenic
response.
Twenty-two sediment REs were separated into three fractions for additional characterization with the MVLN cell bioassay. Nine of the 22 REs selected caused noticeable stress
and/or growth inhibition to the MVLN cells after 72 h of
exposure. Fractionation isolated the toxicity to F3 (Figure 3).
Only one F1 sample and one F2 sample showed signs of toxic
stress (Figure 3). Among the F3 samples, however, 15 of the 22
showed visual signs of stress and/or growth inhibition or cytotoxicity (Figure 3). F1 samples were generally not estrogenic
(Figure 3). Only the F1 sample associated with site J1 induced
a significant response in the MVLN bioassay. Fifty percent of
the mid-polarity F2 samples yielded a significant estrogenic
response. One F3 sample induced a significant MVLN response, but, for the most part, any estrogenic activity of the F3
samples was obscured by the toxic effects of compounds
present in F3 (Figure 3).
In addition to fractionation, 20 sediment REs were
analyzed at 6 dilutions. Complete dose-response relationships useful for REP estimation were not obtained. For
several of the samples, however, dilution alleviated
cytotoxicity, allowing a significant estrogenic response to
be detected. At concentrations less than the 100% and
33% RE concentrations (2.5 ␮l and 0.83 ␮l extract per
well) used for screening, 13 of the 20 samples
analyzed induced a significant response in the MVLN bioassay. Two samples, Y2 and Y5, induced responses over
50%-E2-max, but after dilution the magnitudes of MVLN
responses for the sediment REs were generally less than
20%-E2-max. It is unclear what magnitudes of response may
have been observed if the responses were not confounded by
cytotoxicity.
In Vitro Bioassay for Dioxin-Like and Estrogenic Activity
155
Table 2. Potency-based (TCDD-EQ) and magnitude-based (%-TCDD-max) mass balance analysis for PAH compounds associated with
sediment samples from Ulsan Bay and its inland areas
Sampling
Location
T1
T2
T3
Y0
Y1
Y2
Y3
Y4
Y5
J0
J1
J2
J3
J4
U1
U2
U3
U4
U5
U6
U7
U8
U9
U10
U11
U12
U13
U14
U15
U16
Ri
TEQPAH-rangea
TEQPAHb
pg TEQPAH/g
0.05–0.46
0.04–0.36
0.04–0.36
0.04–0.43
0.05–0.49
0.04–0.39
0.05–0.53
0.32–2.32
1.61–12.0
0.13–1.21
0.23–2.55
0.09–1.17
0.43–4.21
0.81–7.27
0.43–3.74
0.07–0.65
0.09–0.89
0.30–2.92
0.25–2.42
2.85–22.4
3.95–35.7
0.25–2.49
1.38–11.3
0.10–1.03
0.17–1.73
0.12–1.22
0.31–3.06
0.06–0.56
0.17–1.72
5.50–41.3
0.11–1.11
TCDD-EQ20–80c
TCDD-EQ50d
pg TCDD-EQ/g
0.14
0.11
0.11
0.13
0.15
0.12
0.16
0.76
4.29
0.39
0.72
0.32
1.31
2.39
1.24
0.20
0.27
0.91
0.75
7.80
11.6
0.76
3.84
0.30
0.52
0.37
0.94
0.17
0.52
14.5
0.33
62.0–447
165–386
293–826
78.5–165
44.2–134
37.4–55.8
24.9–110
38.7–225
42.8–391
26.5–72.4
34.4–160
NAg
NA
NA
NACh
NAC
NAC
NAC
166
252
NAC
492
114
76.9
45.7
52.2
NA
NAC
NAC
NA
93.2
129
NA
43.8
NA
NA
NA
NAC
NAC
NAC
74.1
NA
%-TCDD-max
Calculatede
Observedf
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
3.90
12.5
0.19
4.31
⬍ 0.00
7.14
9.40
5.50
⬍ 0.00
⬍ 0.00
4.23
3.16
16.6
19.8
2.57
11.6
⬍ 0.00
0.66
⬍ 0.00
4.29
⬍ 0.00
0.99
22.5
⬍ 0.00
NA
NA
NA
5.23
25.9
26.3
2.58
102
109
26.4
100
83.5
98.2
68.8
72.6
NA
30.8
43.1
NA
74.2
83.4
NA
67.0
NA
NA
NA
43.0
24.3
44.1
71.9
NA
a,b
Instrumentally derived 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) equivalents (TEQs) of PAHs associated with sediment samples For
TEQPAH calculation, concentrations of benzo(a)enthracene, chrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, indeno(1,2,3cd)pyrene, and dibenz(a,h)anthracene used
a
Refer to the range of TEQPAH derived using assay-specific REP-20 – 80 values generated from multiple point estimates and selected PAH
concentrations
b
Refer to the TEQPAH generated from assay specific REP-50 value and selected PAH concentrations
c,d
Bioassay-derived TCDD equivalents (TCDD-EQs) of PAHs sediment fraction (pg TCDD/g DW)
c
Refer to the range of dioxin equivalents generated from multiple point estimates made for responses ranging from 20 – 80%-TCDD-max
d
Refer to the dioxin equivalents generated from one-point estimates made for response of 50%-TCDD-max
e
Regression of TEQPAH against the TCDD standard curve was used to predict the magnitude of bioassay response
f
Observed bioassay response (%-TCDD-max) of sediment fraction F2 samples contained PAHs
g
NA: not analyzed
h
NAC: not available data for the calculation of TCDD-EQs, i.e., dose-response curve could not obtained in the full dose-response bioassay
i
Reference site
Discussion
Dioxin-Like Activity
All the dioxin-like activity detected in this study was associated
with sediments and PM samples. This was not unexpected, as
most known AhR agonists are relatively nonpolar compounds
with high log Koc values (McGroddy et al. 1996). Based on the
method detection limit for the H4IIE-luc bioassay and the
minimum pore water volume extracted, PW samples contained
less than 600 pg TCDD-EQ/L. This agrees with concentrations
of total TEQs calculated for pore water samples, which ranged
from 0.2 to 122 pg TEQs/L. TEQs of Three DF samples from
river and/or stream locations (Y3, Y4, and J0) exceeded the
detection limit of 50 pg TCDD-EQ/L for water samples. This
result suggests there would be antagonistic activities among
dissolved organic compounds in water samples. There is also a
possibility of chemical interaction, which may affect the luciferase enzyme activity in the cells. The mass of particulate
matter filtered from each 4 L surface water sample varied
among sample locations. There was a poor correlation (r2 ⫽
0.23) between H4IIE-luc responses to PM and the mass of
156
J. S. Khim et al.
Fig. 3. Luciferase induction in
the MVLN (estrogen-responsive)
cell bioassay elicited by Ulsan
Bay sediment raw extracts (RE),
Florisil fraction 1, 2, 3 (F1, F2,
F3). Response magnitude presented as percentage of the maximum response observed for a
1,000 pM 17␤estradiol standard
(%-E2-max). Πindicates cells
exhibited an altered or “stressed”
morphology; ● indicates the sample was toxic to the cells
Table 3. Instrumentally derived dioxin equivalents of PAHs (TEQPAH) and PCBs (TEQPCB) in water samples (PM: particulate matter, DF:
dissolved fraction) and predicted/observed dioxin-like responses of samples in H4IIE-luc bioassay
Particulate Matter
TEQPAH
a
TEQPCB
Sampling
Location
(ng TEQs/g PM)
T0
T1
T2
T3
T4
Y0
Y1
Y3
Y4
Y5
J0
J1
J2
U1
U2
U5
U7
U16
0.44
⬍ 0.01
0.08
0.09
0.04
0.03
0.05
0.02
0.03
0.02
0.30
0.05
0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
0.28
0.01
49.0
30.3
41.2
34.6
31.8
12.1
21.6
16.7
13.0
7.96
53.2
0.01
6.46
NA
19.1
NA
NA
15.9
Dissolved Fraction
b
c
TEQs
%-TCDD-max
d
49.4
30.3
41.2
34.7
31.9
12.1
21.6
16.7
13.0
7.98
53.5
0.06
6.47
NA
19.1
NA
0.28
15.9
TEQPAH
e
Calculated
Observed
(pg TEQs/L)
17.8
20.0
21.9
21.0
23.7
18.6
22.4
20.2
20.8
20.5
18.4
⬍ 0.00
18.3
NA
18.8
NA
NA
22.4
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
3.45
2.98
26.1
10.6
18.0
18.9
3.0
36.3
3.9
3.93
NA
0.60
10.1
⬍ 0.01
⬍ 0.01
0.55
⬍ 0.01
0.25
0.15
0.67
0.91
1.42
0.21
0.62
0.99
1.47
⬍ 0.01
⬍ 0.01
0.32
⬍ 0.01
⬍ 0.01
1.14
TEQPCB
27.6
1.11
1.29
0.19
19.9
22.3
23.2
238
127
1.68
83.8
2.85
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
0.15
0.06
TEQs
27.6
1.66
1.29
0.44
20.1
23.0
24.1
239
127
2.30
84.8
4.32
⬍ 0.01
⬍ 0.01
0.32
⬍ 0.01
0.15
1.20
%-TCDD-max
Calculated
Observed
3.65
⬍ 0.00
⬍ 0.00
⬍ 0.00
1.75
2.55
2.84
16.5
12.7
⬍ 0.00
10.3
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
⬍ 0.00
0.12
⬍ 0.01
⬍ 0.01
⬍ 0.01
0.12
NAf
0.36
1.07
1.31
1.07
0.12
2.50
⬍ 0.01
⬍ 0.01
NA
⬍ 0.01
⬍ 0.01
⬍ 0.01
a,b
Instrumentally derived 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) equivalents (TEQs) of PAHs and PCBs in water samples
TEQs are sum of TEQPAH and TEQPCB
d
Regression of TEQPAH against the TCDD standard curve was used to predict the magnitude of bioassay response
e
Observed bioassay response (%-TCDD-max) of sediment raw extracts contained PAHs and PCBs
f
NA: not analyzed
c
particulate matter extracted, which suggests that concentrations
of TCDD-EQ associated with PM varied among sampling
locations. This was in agreement with instrumental analysis,
which showed TEQ concentrations for PM samples ranging
from approximately 600 –53,500 pg TEQ/g DW. Generally,
calculated, %-TCDD-max was close and/or greater than observed %-TCDD-max in the bioassay (Table 3). Based on
magnitude-based mass balance analysis, PAHs and PCBs in
PM samples were great enough to induce the dioxin-like responses observed in H4IIE-luc bioassay.
Most of the sediment samples from the Ulsan Bay area
caused significant luciferase induction in the H4IIE-luc cell
bioassay (Table 1). Based on PCB concentrations detected in
F1 (Khim et al. 2000b), concentrations of TEQPCB in F1
samples ranged from 0.003 to 0.67 pg TEQ/g DW. Following
the regression against the TCDD standard curve, these concen-
In Vitro Bioassay for Dioxin-Like and Estrogenic Activity
trations would not have been expected to produce a significant
response. Thirteen of the F1 sediment samples induced a significant H4IIE-luc response, however. This suggests that there
may be nonpolar dioxin-like compounds in F1 or supradditive
interactions that were not identified by instrumental analysis.
Responses of F2 and F3 sediment samples relative to corresponding REs further complicated the assessment. Over 80%
of the F2 sediment samples and over 70% of the F3 samples
yielded activities that were greater than that of the corresponding RE. This strongly suggests the presence of antagonists in
F1. In some cases, high concentrations of noncoplanar PCBs
have been shown to reduce the potency of coplanar congeners,
presumably by restricting access to the AhR (Sanderson et al.
1996). Attributing the apparent antagonism of RE responses to
noncoplanar PCBs present in F1 does not explain why F1
responses were greater than the magnitudes predicted based on
instrumentally derived TEQs, however. Furthermore, such attribution would also imply the presence of other potent, nonpolar, non-PCB, AhR agonists in F1. Such agonists would be
needed to counteract the antagonistic compounds in F1 and still
yield a significant response. Alternatively, competitive antagonists may have been present in all three fractions but diluted
enough through the separation procedure to allow for greater
magnitudes of response to F2 and F3 samples. Overall, the
results suggest a complicated sample composition for F1 samples, which is not easily explained from the instrumental results. Further bioassay directed fractionation and identification
would be needed to characterize the antagonistic compounds
responsible for the bioassay activity observed in RE and fraction samples.
Mass balance assessment of F2 samples was a bit more
straightforward. PAHs were shown to partition to F2 (Khim et
al. 1999b). Benzo(a)anthracene, chrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, indeno(1,2,3-cd)pyrene, and dibenz(a,h)anthracene have all been shown to elicit
dioxin-like responses in the H4IIE-luc bioassay and similar
assays (Willet et al. 1997; Khim et al. 2000c). H4IIE-lucspecific REPs for these PAHs were used to calculate TEQPAH
for the F2 sediment samples (Khim et al. 2000c). On average,
there was approximately 10-fold uncertainty in the TEQPAH
estimates due to uncertainties in the REP estimates for PAHs.
Assuming the maximum concentration of TEQPAH over each
range of TEQ uncertainty, 8 of the 22 F2 samples tested were
predicted to yield a significant H4IIE-luc response (Table 2).
The greatest magnitude of response predicted for an F2 sample,
based on TEQPAH was 22.5%-TCDD-max for sample U16
(Table 2). The observed response for U16 was 72%-TCDDmax (Table 2). Only two of the F2 samples tested yielded
response magnitudes less than 22.5%-TCDD-max. In general,
magnitudes of response observed for sediment F2 samples
were much greater than magnitudes predicted based on the
maximum potential TEQPAH in the sample. This suggests the
presence of other, unidentified dioxin-like compounds in F2 of
the Ulsan Bay sediment samples.
To investigate this mass balance relationship further, doseresponse curves were generated for 11 of the 22 F2 samples by
testing 6 dilutions of each sample. The dose response relationships were used to estimate bioassay-derived dioxin equivalents (TCDD-EQ) for each sample. These were compared to
instrumentally derived TEQPAH estimates to provide a potencybased mass balance analysis. TCDD-EQ and TEQPAH were
157
both presented as a range to account for the degree of uncertainty in the estimates due to deviations from parallelism to the
TCDD standard curve or uncertainty in REP estimates for
individual PAHs (Table 2). Only 1 of the 11 samples examined,
U16, had overlapping TCDD-EQ and TEQPAH ranges (Table
2), which suggested that PAHs may account for a substantial
portion of the dioxin-like potency of the sample. Samples U6,
U7, and U9 from Ulsan Bay had TCDD-EQ ranges that were
relatively close to their corresponding TEQPAH range. This
implies that PAHs may have contributed some activity to the
samples, but other unidentified compounds were likely to be
more significant sources of dioxin-like activity. Samples Y4,
Y5, J1, J2, J3, and J4 were all from river sites. TEQPAH for all
these samples were much lower than the corresponding
TCDD-EQ ranges (Table 2), indicating that PAHs contributed
a relatively minor portion to the overall potency of the samples.
Sample U1 was collected from the Ulsan Bay, closest to the
mouth of Yeocheon Stream. Correspondingly, the degree of
difference between its TEQPAH range and TCDD-EQ range
was intermediate relative to the other Ulsan Bay and river sites.
As a whole, potency-based mass balance supports a hypothesis
that unidentified compounds associated with F2 of Ulsan Bay
area sediment samples were responsible for much of the dioxin-like activity observed. Furthermore, comparison of the magnitudes of difference between TEQPAH ranges and TCDD-EQ
ranges suggests the hypothesis that there is a gradient in the
relative contribution of PAHs to total dioxin-like activity as
one moves from inland streams out to Ulsan Bay. Both response-based and potency-based mass balance analysis of sediment F2 samples suggested the presence of unidentified dioxin-like compounds in F2. This suggests that the presence of
other dioxin-like compounds in F2, such as polychlorinated
dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans
(PCDFs), and polychlorinated naphthalenes (PCNs), which are
known to partition in F2 of the Florisil column chromatography. However a study from Tokyo Bay, Japan, showed that the
contribution of PCDDs/DFs and PCNs to TEQs were less than
those by PAHs (Kannan et al. 2000; Yamashita et al. 2000)
Extensive mass balance analysis was not possible for F3
samples because no target dioxin-like compounds have been
shown to partition into F3 (Khim et al. 1999b). Despite the lack
of known AhR agonists, however, F3 samples were found to be
nearly as active as F2. TCDD-EQ estimates generated for 12
selected F3 samples ranged from approximately 14 –1,050 pg/g
DW, which suggests that the overall potency of compounds
associated with the most polar fraction of Ulsan Bay area
sediment samples was relatively similar to that of the midpolarity dioxin-like compounds. It is unclear, however,
whether the potency is attributed to low concentrations of
highly potent agonists or high concentrations of less potent
agonists.
Estrogenic Activity
Approximately one-third of the water samples (PW and DF)
exhibited some estrogenic activity. Response magnitudes observed ranged considerably from barely significant to as great
as 91%-E2-max (Figure 1). Sites that yielded estrogenic pore
water and surface water samples also tended to have relatively
158
high concentrations of EEQ in sediment. Water concentrations
of estrogenic compounds, such as NP, OP, and BPA, were not
determined, however; thus the correlation between MVLN
responses and waterborne EEQ could not be determined. The
localization of estrogenic waters to inland lotic systems is
consistent with the hypothesis that municipal and industrial
waste discharges to Ulsan streams is the major source of APs
and BPA in the region (Khim et al. 2000b)
Sediment and PM REs generally did not induce an estrogenic
response in the MVLN bioassay. Sediment concentrations of
NP, OP, BPA, and two estrogenic PAHs (benzo(a)anthracene,
and dibenz(a,h)anthracene) were multiplied by their MVLN
assay–specific relative potencies (Villeneuve et al. 1998b;
Khim et al. 2000c) to estimate the total EEQs associated with
sediment REs. Total EEQ estimates ranged between 0.008 and
15.3 pg EEQ/g DW. Based on regression against an E2 standard curve, these concentrations of EEQ would not have been
sufficient to induce a significant estrogenic response in the
MVLN cell bioassay. Thus, the relative lack of response was
consistent with the known composition of the samples. Some
activity was detected, however, and RE responses for at least
12 of the sediment samples were confounded by cytotoxicity,
thus selected sediment samples were studied in greater detail.
Twenty sediment REs were tested in the MVLN bioassay at
six different dilutions. Dose-response relationships suitable for
E2-EQ estimation were not obtained for most samples, but
dilution was found to alleviate the cytotoxic effects of some
samples, allowing an estrogenic response to be observed. Over
50% of the samples showed a significant estrogenic response
when diluted to noncytotoxic concentrations. It was unclear
how high response magnitudes may have been observed if the
RE samples have not been shown the cytotoxic effects, however.
Florisil fractionation was employed to separate estrogenic
compounds from overtly toxic ones. Upon fractionation, cytotoxicity was only associated with F3. This indicated that relatively polar compounds were responsible for the overtly toxic
effects of the extracts. The general lack of estrogenic activity
associated with F1 samples was not surprising, as F1 was not
expected to contain any estrogenic compounds. PAHs, some of
which are able to induce estrogenic responses in vitro, partition
to F2 (Clemons et al. 1998; Khim et al. 2000c). Correspondingly, a number of F2 samples showed significant estrogenic
activity. As mentioned earlier, however, total EEQs calculated
for the samples included the contribution of estrogenic PAHs.
The total EEQ concentrations measured were not sufficient to
account for the responses observed for the F2 samples. As a
result, the estrogenic activity of F2 does not appear to be
attributable to estrogenic PAHs. OC pesticides can also partition to F2, but were not present in sufficient concentration to
yield an estrogenic response (Soto et al. 1994).
Over half of the F3 samples were toxic to the MVLN cells.
In general, MVLN cells have been found to be more sensitive
to cytotoxic agents than H4IIE-luc cells. Thus, the disparity
between the toxic effects of the extracts on the different cell
types does not imply that the toxicity was mediated through an
estrogenic mechanism of action. Based on previous studies, the
known concentrations of NP, OP, and BPA were not sufficient
to kill the MVLN cells (Khim et al. 1999a, 1999c). The
potential estrogenic activity of compounds in F3 could not be
J. S. Khim et al.
discerned without additional fractionation and/or treatment to
separate toxic compounds from estrogenic ones.
Recurring Trends
This is the third in a series of studies that employed a combination of instrumental analysis and in vitro bioassay to study
dioxin-like and estrogenic contaminants in sediments from
Korea. Over the course of these studies, several recurring
conclusions have emerged. First, dioxin-like activity has consistently been associated primarily with compounds present in
florisil F2 and F3 samples. The F2 responses may be explained
by the potential presence of PAHs and their derivatives, PCDDs, PCDFs, and/or PCNs in F2. Future studies should employ
high-resolution mass spectrometry and method for PCN identification and/or quantification to address the potential contribution of these known AhR agonists. However a recent study
has shown that PCDDs, PCDFs, and PCNs contribute less to
TEQs compared to those by PAHs (Kannan et al. 2000). Thus,
either unidentified compounds or interactions among chemicals
contribute to the activity. F3 responses are more perplexing.
Prototypical dioxin-like compounds would not be expected to
partition to F3. Current evidence suggests that the dioxin-like
compounds present in F3 are relatively polar and acid labile
(Khim et al. 1999a, 1999c). Additional fractionation could be
used to further separate and isolate active agents present in F3
samples. Due to the polarity of compounds in F3, however,
liquid chromatograph equipped with a mass selective detector
(LC/MS) would probably be required to identify and quantify
suspect agents.
Second, most estrogenic activities associated with F3 samples could not be explained by the concentrations of known ER
agonists, such as APs and BPA. Based on qualitative and
quantitative mass balance analysis, known concentrations of
prototypical xenoestrogens can account for only a portion of
estrogenic activities observed in F3 samples. These results
suggest that florisil F3 samples contained unidentified or nondetectable bioactive compounds that contributed to the MVLN
responses. Low concentrations of highly potent ER agonists,
such as E2, ethynyl estradiol (EE2), and/or estrone (E1), are
one possibility. High concentrations of natural or synthetic ER
agonists that were not detected or identified by instrumental
analysis may also explain the responses. Finally, it is also
possible that one or more compounds within the complex
mixture may act synergistically with the APs and BPA or with
some other agonists, such as phytoestrogens, to yield responses
of the magnitude observed in the bioassay.
Through the studies conducted, most of the cytotoxicity
and/or altered or “stressed” morphology in MVLN cells were
observed for sediment F3 samples. Based on previous and
current studies, the known concentrations of NP, OP, and BPA
in F3 samples were not related to the degree of cytoxicity
observed. Additionally known concentrations of NP, OP, and
BPA were not sufficient to kill the MVLN cells (Khim et al.
1999a, 1999c). This indicates that unidentified compounds in
F3 samples may be responsible for the majority of cytotoxicity
observed. Further fractionation and/or clean-up techniques are
needed to separate these toxic compounds from nontoxic and
estrogenic compounds to evaluate the estrogenic potency of
In Vitro Bioassay for Dioxin-Like and Estrogenic Activity
environmental samples without interference caused by the
toxic compounds.
Summary
Based on the initial screening of raw extracts, most sediment
and some particulate matter of water samples showed significant dioxin-like activity in H4IIE-luc bioassay, whereas no
pore water and dissolved fraction of water samples elicited
bioassay response. Most of the activities associated with florisil
column–fractionated samples showed that F2 and F3 fractions
were responsible for the significant reporter gene expression in
H4IIE-luc and MVLN bioassay. Based on a qualitative mass
balance analysis, PAHs appeared to account for a portion of
dioxin-like responses observed and xenoestrogens, such as NP,
OP, and BPA, were not responsible for the estrogenic activities
observed. This suggests the presence of other dioxin-like compounds in F2, such as PCDDs, PCDFs, and PCNs. Again most
F3 samples induced dioxin-like activity significantly. Over half
the F3 samples were either cytotoxic or caused morphological
changes in H4IIE-luc and MVLN cells, however. Known concentrations of both APs and BPA were not great enough to
account for cytotoxicity observed in F3 samples.
Acknowledgments. This work was supported by the National Institute of Environmental Research (NIER), Ministry of Environment,
Korea (Sediment Organic Compound Bioassay Study; SORGBIOS
98 –2000). We thank Dr. M. D. Pons, Institut National de la Sante de
la Recherche Medicale for the MVLN cells, and Dr. Jac Aarts, University of Wageningen, The Netherlands, for the H4IIE-luc cells. We
also acknowledge colleagues from the Benthos Lab at Seoul National
University, Korea, and Aquatic Toxicology Laboratory at Michigan
State University, MI, for their technical assistances.
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