Arch. Environ. Contam. Toxicol. 48, 575–586 (2005)
DOI: 10.1007/s00244-004-0166-1
Estrogenic and Dioxin-like Activities and Cytotoxicity of Sediments and Biota from
Hong Kong Mudflats
H. L. Wong,1 J. P. Giesy,1,2 W. H. L. Siu,1 P. K. S. Lam1
1
Department of Biology and Chemistry, City University of Hong Kong, Tat Chee Avenue, Kowloon, Hong Kong
Zoology Department, National Food Safety and Toxicology Center and Center for Integrative Toxicology, Michigan State University, East Lansing,
Michigan 48824-1311, USA
2
Received: 20 July 2004 /Accepted: 19 October 2004
Abstract. Persistent organic pollutants, such as organochlorine
insecticides and polychlorinated biphenyls (PCBs), were
measured in several environmental matrices including aerial
deposition, seawater, sediment, and biota in two important
coastal wetlands of Hong Kong, China. Specifically, samples
were collected from within the Mai Po Marshes Nature Reserve (Mai Po), an internationally acclaimed wetland situated
in the northwestern part of the New Territories of Hong Kong,
and A Chau in Starling Inlet, a relatively remote island on the
eastern side of Hong Kong. Hexachlorobenzene, dichlorodiphenyltrichloroethanes, and hexachlorocyclohexanes were
detected in all samples collected from Mai Po. Environmental
endocrine disruptors (including dioxin-like compounds and
estrogenic chemicals), measured by the use of cell-based
chemical activated luciferase expression assays, were found to
occur at concentrations that might pose a risk to the ecologic
systems in Mai Po. Dioxin-like PCBs were detected at small
concentrations in some of the samples. Concentrations of
2,3,7,8 tetrachlorodibenzo-p-dioxin equivalents (TEQs) were
primarily related to the relatively great concentrations (>100
ng/g dry weight) of high molecular–weight polycyclic aromatic hydrocarbons in sediments, whereas the relative proportion of TEQs contributed by nonortho-substituted PCBs
was small. Polar compounds primarily contributed estrogen
equivalents, which were measured in sediments. Significant
concentrations of cytotoxic compounds were detected in fish
samples collected from the Mai Po but not in fish collected
from A Chau.
A study was conducted to characterize the current magnitude
and extent of contamination in two important coastal marine
habitats, Mai Po Marshes Nature Reserve (Mai Po) and A
Chau, in Hong Kong. Mai Po is located at the northwest edge
of Hong Kong where mudflats and mangroves are important
Correspondence to: P. K. S. Lam; email: bhpksl@cityu.edu.hk
habitats for a wide variety of birds, particularly migratory
shore birds. Mai Po is an important midpoint of the Eastern
Asian–Australian Flyway. For this reason, the area has been
designated a wetland of international importance under the
Ramsar Convention. Persistent organic pollutants (POPs)—
such as polycyclic aromatic hydrocarbons (PAHs), organochlorine (OC) insecticides, polychlorinated biphenyls (PCBs),
polychlorinated dibenzo-p-dioxins (PCDDs), and polychlorinated dibenzofurans (PCDFs)—have been considered to be the
major contaminants of concern that might pose a threat to this
valuable and protected system. A Chau is a remote island in
the northeastern part of Hong Kong where several species of
waterbird breed. A Chau is expected to be less contaminated
than Mai Po (Connell et al. 2003), and was selected as a reference area. In this study, a survey of the primary contaminants of concern was conducted at Mai Po and A Chau.
Because POPs are not the only toxicants that could possibly
cause effects on wildlife, several bioanalytical techniques were
used to screen for larger classes of compounds that might affect wildlife. Previous studies demonstrated that several OC
pesticides and PCBs in the sediment and biota at Mai Po exceed the guideline values of regulatory authorities, e.g., the
United States Environmental Protection Agency (Zheng et al.
1999; Liang et al. 1999; MEller et al. 2002). Notwithstanding,
there has been no attempt to correlate body concentrations
with hormonal and other signal transduction interference effects. Therefore, baseline studies aimed at characterizing the
key environmental contaminants, especially those exerting
endocrine-disrupting effects, are needed to establish protective
schemes for these valuable ecosystems. Specifically, these
bioassays, based on in vitro cell systems, were used to screen
for estrogenic potency, total dioxin-like activity, and nonspecific cytotoxicity. These screening assays were used in conjunction with toxicity identification and evaluation techniques
to determine potential classes of compounds contributing to
each of the observed effects.
In vitro bioassays are useful for screening biorelevant
toxicological end points and providing information that can
complement instrumental analysis and allow a more complete
and cost-effective characterization of environmental samples
(Murk et al. 1996; Hilscherova et al. 2000). Bioassays can
576
H. L. Wong et al.
Fig. 1. Map showing the Mai Po and A Chau study sites. Sampling locations in the Mai Po area: 1 = fish pond (FP); 2 = gei wei (GW); 3 = the
gate of gei wei (NG); 4 = middle part of the mangrove (MM); 5 = mudflat near the bird-viewing hide (NB); and 6 = Inner deep bay (ID).
integrate the effects of a number of compounds in complex
mixtures and provide an estimate of the overall potency of the
mixture to cause a specific biologic effect. For instance, the
H4IIE-luc bioassay used in this study provides an estimate of
the overall potency of compounds that cause their effects
through interaction with the aryl hydrocarbon receptor (AhR).
The relative potency is determined by comparing the results of
the unknown mixture to that of a well-characterized reference
compound, e.g., 2,3,7,8- tetrachlorodibenzo-p-dioxin (TCDD),
for AhR-mediated toxicity. This method has been widely
adopted in various environmental studies around the world
(e.g., Hilscherova et al. 2000; Khim et al. 1999, 2000; Tillitt
et al. 1996). In this study, chemically activated luciferase
expression (CALUX) bioassays with genetically transfected
cells (H4IIE-luc and MVLN-luc) were used to evaluate the
dioxin-like and estrogenic effects of environmental samples,
respectively, and provide a baseline estimation of the potential
hazard of endocrine-disrupting chemicals in various environmental compartments to allow cost-effective monitoring and/
or remediation.
Materials and Methods
Sampling and Instrumental Analysis
Seawater, sediment, bulk atmospheric deposition, and biota (including
polychaetes, shrimps, and fish) were collected from Mai Po and A
Chau. Sediments were collected from six locations at the Mai Po
Marshes and one at A Chau (Fig. 1). Relative potencies of AhR-active
(TCDD-like) compounds with respect to 2,3,7,8-TCDD were determined by two methods: the H4IIE-luc bioassay and instrumental
analysis. Toxic equivalency quotients (TEQs) were obtained by
multiplying the concentration of each AhR-active PCB measured by
instrumental analysis with its relative toxic equivalency factor (TEF)
provided by the World Health Organization (WHO) based on the
responsiveness of mammals, birds, and fish. These were designated as
TEQWHO-mammal, TEQWHO-bird, and TEQWHO-fish, respectively. The
methods used follow those of Khim et al. (1999) and Snyder et al.
(2001).
In brief, sediment, suspended particulates in seawater and biota
were freeze dried and extracted with dichloromethane and hexane for
16 hours in a Soxhlet apparatus. Each seawater sample (20 L) was
filtered through a precleaned 0.5-lm glass fiber filter (TOYO GC-50)
Toyo Roshi, Tokyo, Japan. The suspended particles and the filtrate
were extracted by dichloromethane. Extracts were concentrated by
rotary evaporation and then fractionated by a Florisil column (10 mm
i.d. with 10 g Florisil as described in Khim et al. [1999]) for subsequent analysis. PCBs, dichlorodiphenyltrichloroethanes (DDTs),
hexachlorobenzene (HCB), and hexachlorocyclohexanes (HCHs)
were eluted in fraction 1 (F1) by 80 to 120 mL hexane. Most of the
PAHs were eluted in fraction 2 (F2) by 100 mL hexane/dichloromethane (4:1). Six-ring PAHs (in fish samples only) and polar compounds were eluted in fraction 3 (F3) by 100 mL dichloromethane/
methanol (1:1). The efficiency of the Florisil separation was confirmed by spiking samples with appropriate standards. Concentrated
sulfuric acid was applied to F1 to decrease interference from other
chlorinated compounds to improve the detection limit of PCBs for
instrumental analysis after testing for their potencies using bioassay.
PAHs in a sample were only determined if positive response of
H4IIE-luc bioassay was found. OCs and PCBs were separated and
quantified using a Hewlett Packard 5890 series II gas chromatograph
(GC) equipped with an electron-capture detector (ECD) (Palo Alto,
CA). A fused silica capillary column (60 m x 0.25 mm diameter)
coated with DB-5MS (5% phenyl-methyl-polysiloxane; J&W Scientific, Folsom, CA) at 0.25-lm film thickness was used. The temperature of the column oven was programmed from an initial
temperature of 100C with a 3-minute holding time to 210C at a rate
of 25C/min and was then ramped at a rate of 1C/min to 270C, with
a final hold time of 60 minutes. The injector and detector temperatures were maintained at 250C and 280C, respectively. PAHs were
quantified using a Hewlett Packard 5890 series II GC equipped with a
flame-ionization detector). A fused silica capillary column (30 m x
0.25 mm diameter) coated with DB-5MS at 0.25-lm film thickness
was used. The column oven was programmed from an initial temperature of 120C to 160C at a rate of 10C/min and was then
ramped at a rate of 9C/min to 280C, with a final hold time of 24
minutes. The injector and detector temperatures were maintained at
280C. The relative potency (REP) values (Tables 1 and 2) were
calculated by multiplying the concentration of PAHs with REP50
5.6 €
0.1 €
0.3 €
0.1 €
2.6 €
0.6 €
0.03 €
0.02 €
1.2 €
5.1 €
0.001 €
0.07 €
0.6 €
<0.001
0.4
0.01
0.05
0.01
1.0
0.6
0.01
0.02
0.8
0.2
0.001
0.01
0.7
TEQWHO-Mammal
1.7 €
0.01 €
0.004 €
<0.001
0.8 €
0.1 €
<0.001
<0.001
0.3 €
1.5 €
<0.001
0.4 €
0.1 €
<0.001
0.02
0.1
0.6
2.5
1.3
0.06
2.8
0.001
0.001
TEQWHO-Fish
10.4 €
2.2 €
1.2 €
0.01 €
2.6 €
6.2 €
<0.001
0.003 €
3.2 €
5.1 €
<0.001
1.0 · 101 €
0.01 €
<0.001
5.5 · 10)1
0.001
0.001
0.75
2.4 · 10)1
0.8
0.3
0.17
0.001
1.0
3.2
TEQWHO-Bird
19.2
33.1
12.8
4.3
9.8
16.7
1.4
15.6
30.4
0.3
3.2
28.6
25.7
2.5
€
€
€
€
€
€
€
€
€
€
€
€
€
€
9.0
3.9
0.7
0.2
0.8
1.2
0.06
5.1
0.9
0.08
0.3
0.9
0.6
0.3
Total DDTs
0.2 €
1.5 €
2.3 €
0.8 €
5.4 €
6.6 €
0.4 €
4.4 €
4.6 €
0.4 €
1.1 €
9.0 €
8.6 €
ND
0.02
0.06
0.14
0.07
0.4
0.5
0.1
0.2
0.3
0.02
0.09
4.4
0.6
Total HCHs
9.6
23.8
14.7
0.05
7.5
14.6
1.5
8.0
19.8
3.8
3.9
1.5
11.1
3.3
HCB
€
€
€
€
€
€
€
€
€
€
€
€
€
€
0.3
1.0
0.7
0.01
0.4
2.8
0.2
0.2
0.8
0.2
0.2
0.2
0.9
0.5
9.3
37.5
22.1
5.6
13.2
9.8
0.08
8.6
17.2
2.5
3.7
9.8
14.5
0.4
€
€
€
€
€
€
€
€
€
€
€
€
€
€
0.7
0.7
6.0
0.2
1.2
1.1
0.05
0.7
0.5
0.1
0.1
4.1
0.8
0.4
Cyclodienes
–
1200 €
350 €
810 €
680 €
820 €
110 €
–
633 €
622 €
501 €
220 €
503 €
510 €
160
290
21
44
650
66
160
43
300
69
130
27
Total PAHs
€
€
€
€
€
€
0.8 €
2.0 €
0.8 €
1.4 €
3.1 €
4.1 €
NA
7.8
8.5
6.7
110.4
141.2
0.6
0.3
2.4
1.4
1.0
1.5
3.6
1.2
2.7
3.9
2.3
3.6
0.8
REP Value
The concentrations of PAHs, DDTs, HCHs, HCB, and cyclodienes were expressed as ng/g dw, and PCBÕs were expressed as I-TEQ (pg TCDD-TEQ/g dw mean € SDs). REP was calculated by
the multiplication of the REP50 values from Villeneuve et al. (2000) with the concentration of PAHs.
– = No measurement was made.
DDTs = Dichlorodiphenyltrichloroethanes.
HCB = Hexachlorobenzene.
HCHs = Hexachlorocyclohexanes.
NA = Not applicable for calculations.
ND = Data below detection limit.
OCs = Organochlorine insecticides.
PAHs = Polycyclic aromatic hydrocarbons.
PCBs = Polychlorinated biphenyls.
REP = Relative potency.
a
A Chau
Mai Po
Wet
A Chau
Fish pond
Gei wai
Near the gate
Middle mangrove
Near bird-viewing hide
Inner deep bay
Mai Po
Dry
Fish pond
Gei wai
Near the gate
Middle mangrove
Near bird-viewing Hide
Inner deep bay
Location
Season
PCB
Table 1. Summary of OCs in the sediment samples from Mai Po and A Chaua
Sediments and Biota from Hong Kong Mudflats
577
Total DDTs
0.02
0.004
0.01
0.001
0.001
5.0 €
48.3 €
1.9 €
0.6 €
ND
ND
0.1 €
0.001 €
0.011 €
0.1 €
ND
2.6 €
ND
2.5
0.1
0.001
0.02
0.006
0.7
2.2
0.4
0.3
0.5 € 0.2
2.6 € 2.5
139.7
100.3
401.8
332.4
14.8
12.3
8.0
11.4
15.4
23.4
48.3
30.2
22.8
€
€
€
€
€
€
€
€
€
€
€
€
€
122.1
8.5
22.3
33.4
9.8
7.3
8.1
33.3
8.1
7.7
10.2
33.1
7.7
292.1 € 34.8
173.3 € 16.7
1.5 €
16.1 €
28.1 €
3.7 €
1.1 €
9.4 €
ND
4.7 €
3.2 €
3.1 €
7.1 €
3.6 €
5.4 €
0.4
0.4
0.3
2.7
3.1
2.2
0.3
1.5
1.8
0.4
1.3
3.5
7.7 € 0.5
9.1 € 0.9
13.2 € 1.9
8.2 € 0.8
–
–
Total
PAHs
NA
NA
Rep
Value
162.2 € 12.1 893 € 75 5.5 € 1.6
63.2 € 9.3
57.5 € 2.3
Cyclodienes
55.4
391.2
751.3
141.5
34.4
22.3
15.5
3.9
20.7
11.2
50.6
17.2
57.8
€
€
€
€
€
€
€
€
€
€
€
€
€
2.4
72.8 € 2.9
–
NA
20.5 131.2 € 18.2 134 € 45 1.5 € 0.3
50.8
138 € 7.3 362 € 42 0.3 € 1.5
9.3
77.2 € 12.3 311 € 25 0.5 € 0.9
9.8
ND
–
NA
34.4
ND
–
NA
5.7
ND
–
NA
0.5
ND
–
NA
1.3
ND
–
NA
1.3
3.4 € 1.5
–
NA
5.0
10.3 € 1.6
–
NA
2.0
0.7 € 0.1
–
NA
17.5 12.8 € 2.1
–
NA
251.2 € 22.1 123.1 € 16.3 801 € 14 4.3 € 0.4
131.3 € 6.0
71.8 € 11.8
58.8 € 3.9
Total HCHs HCB
The concentrations of PAHs, DDTs, HCHs, HCB, and cyclodienes were expressed as ng/g lipid wt, and PCB were expressed as TEQWHO (pg TCDD-TEQ/g lipid wt mean € SDs). REP was
calculated by the multiplication of the REP50 values from Villeneuve (2000) with the concentration of PAHs.
– = No measurement was made.
D = Dry season.
dw = Dry weight.
DDTs = Dichlorodiphenyltrichloroethanes.
HCB = Hexachlorobenzene.
HCHs = Hexachlorocyclohexanes.
NA = Not applicable for calculations.
ND = Data below detection limit.
OCs = Organochlorine insecticides.
PAHs = Polycyclic aromatic hydrocarbons.
PCBs = Polychlorinated biphenyls.
REP = Relative potency.
W = Wet season.
a
0.0001
0.002
0.004
0.01
0.01 €
0.02 €
0.004 €
0.001 €
ND
ND
ND
ND
ND
ND
ND
0.01 €
ND
0.02 € 0.01
0.2 € 0.2
0.1
0.3
0.001
0.002
0.01 € 0.013
0.2 €
0.3 €
0.005 €
0.02 €
ND
ND
0.004 €
0.01 €
ND
0.03 €
ND
0.01 €
ND
TEQWHO-Bird
0.2 € 0.001
0.6 € 0.29
100.2 € 18.3
0.001 € 0.001 0.002 € 0.0003 90.8 € 4.6
0.006 € 0.001
2.4 € 0.2
0.05 € 0.001
Lipid content
(% dw)
TEQWHO-Mammal TEQWHO-Fish
Mai Po
Polychaetes (D)
1.0 – 1.2
Near birdPolychaetes (W)
1.2–l2.6
viewing hide
Boleophthalmus
6.4–8.1
boddaerti (D)
Boleophthalmus
14–19
boddaerti (W)
Metapenaeus ensis
5.4–7.0
Gei wai
Mugil cephalus
18–19
Mylio marcocephalus
9.0–9.5
Tilapia zilla
9.3–10.3
A Chau
Polychaetes (D)
2.5–4.5
Polychaetes (W)
2.5–3.2
Acetes spp.
1.0–1.6
Cyclina orientalis (D)
2.0–2.4
Cyclina orientalis (W) 1.8–1.9
Gobiidae A (D)
8.6–10.2
Gobiidae A (W)
9.6–10.3
Gobiidae B (D)
12.9–13.7
Gobiidae B (W)
10–11
Species
PCB
Table 2. Summary of the OCs in the biota samples from Mai Po and A Chaua
578
H. L. Wong et al.
579
Sediments and Biota from Hong Kong Mudflats
values in Villeneuve et al. (2002). External standards of OC pesticides and PCBs were added to water, particulates, sediment, and biota
with various lipid content to determine recoveries (n = 3). Preliminary results indicated that lipids and matrix effects needed to be
decreased in samples with high lipid contents to improve recoveries
and precision. To achieve this, samples with high lipid contents were
divided up and fractionated in two or three Florisil columns until no
particles were observed in F1 and F2. Recoveries of all ortho- and
nonortho-substituted PCBs and OCs ranged from 90% to 102% in
sediment and biota samples. Fifteen PAHs were spiked into the
sediment and biota samples to test the recoveries of their extraction.
The average recoveries of individual PAHs ranged from
65.2 € 17.2% for Fluorene to 124 € 5.08% for Pyrene with an
average of 92.6 € 5.1 % for the other 13 PAHs. The data by instrumental analysis were not corrected for recoveries.
Bioassay Analysis
Bioassays were conducted with either raw extracts or extracts that had
been treated to remove specific compounds or fractionated into different classes based on polarity. Samples were divided into two
aliquots. The first aliquot was treated with concentrated sulfuric acid
to remove acid-labile chemicals such as PAHs, cyclodienes and lipids.
The second aliquot was not acid treated. Both aliquots were then
screened for the AhR and estrogen receptor (ER)–mediated activities
by the H4IIE-luc and MVLN-luc bioassays, respectively. The response measured was the production of the protein luciferase, which
was detected by the use of Luclite reagent (Packard Bioscience,
Merdian, CT). Light formed by the interaction of luciferase and the
Luclite reagent was quantified with a luminometer (BMG, PolarStar,
Durham, NC). The limit of quantification (LOQ) for all of the assays
was specified to be three times the blank response (obtained by dosing
with the same amount of solvent control). Those samples exhibiting
positive responses (light production) in the cell-based bioassay systems were fractionated on a Florisil column. H4IIE-luc and MVLNluc cells were dosed with the three fractions obtained (Khim 1999;
Hilscherova et al. 2000). Relative potencies of AhR-mediated (dioxinlike) compounds in samples exhibiting positive results were reported
in terms of their potencies relative to 2,3,7,8-TCDD. For AhR-activity
measured using the H4IIE-luc in vitro bioassay, the potency of the
entire mixture was designated as TCDD-EQH4IIE-luc. The selection
criteria for the methods used to calculate TCDD-EQH4IIE-luc in this
study were based on the decision making process given by Villeneuve
et al. (2000). In this study, H4IIE-luc cells were exposed to three
concentrations of 15 different individual PAHs (400 ng/mL, 132 ng/
mL, and 44 ng/mL), which were similar to the concentrations in the
environmental samples collected from Mai Po and A Chau, to
examine the structural specification effects from PAHs to the H4IIEluc induction. Because the sample maxima in the bioassay were <20%
maximum of the TCDD standard, point estimation of relative potency
based on EC20 or EC10 were applied for F2 and F3, respectively. The
same approach was applied for testing estrogenic compounds by
MVLN-luc, but REP-50 and the REP 20-80 ranges were calculated
because the sample maxima were shown to be >50% maximum of the
17b-estradiol (E2) standard.
Cytotoxicities of the environmental samples to the H4IIE-luc cells
were screened using a colorimetric cell viability test based on the
dehydrogenase-mediated conversion of a yellow-colored 3-(4,5-dimethylthiazol-2-yl)-5-(3-carboxymethoxyphenyl)-2-(4-sulfophenyl)-2Htetrazolium inner salt to a brownish colored formazan product (Promega, G5421, Madison, WI) within the metabolically active H4IIEluc cells. The relative viability of H4IIE-luc cells was calculated by
comparing the response of cells exposed to the extract with that of the
solvent only. In addition, composite samples of Mugil cephalus
(n = 3) were divided into three aliquots and stored at –20C, at 4C,
and at room temperature, respectively, for 3 months. Cell viabilities in
the three treatments were assayed using the previously mentioned
cytotoxicity test and compared with the cytotoxicities in their
respective samples at the start of the 3-month period. Finally, based on
results from instrumental analysis, mixtures containing amounts of
OC and PAH standards equivalent to those found in the environmental
samples were prepared for dosing to examine the relative contribution
of various chemicals in accounting for the observed cytotoxicity in the
positive cytotoxic samples.
In addition to the values determined by the bioassay, concentrations
of TCDD-equivalents were estimated by the use of relative potency
(REP) values for each congener derived in the H4IIE-luc bioassay
(REPH4IIE-luc). 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents were
designated as TEQH4IIE-luc (Hilscherova et al. 2000). The calculated
TEQs were the sum of the products of the concentration of each
congener and its associated REP value. These calculated TEQs were
used to conduct mass balance calculations to determine if all of the
compounds contributing to the TEQ had been identified by instrumental analysis.
Results and Discussion
Instrumental Analysis
In this study, o,pÕ and p,pÕ isomers of the following were
detected in most samples (Tables 1 and 2): DDDs, DDEs, and
DDTs; HCB; a-, b-, and c-HCHs; cyclodienes (aldrin, dieldrin, eldrin, endosulfans I and II, heptachlor, kepone, mirex);
and PCBs (28, 37, 52, 71, 81, 105, 114, 118, 126, 156, 157,
167, and 189). Concentrations of TEQWHO-bird (based on
TEFWHO-bird for birds) were greater than TEQWHO-mammal or
TEQWHO-fish (based on TEFWHO-mammal or TEFWHO-fish for
mammals or fish, respectively). The greater concentration of
TEQWHO-bird was related to the presence of PCB77 and
PCB81, both of which have greater TEQWHOs for birds than
for mammals or fish (van den Berg et al. 1998). OC pesticides were widely distributed in different biotic and abiotic
compartments. Cyclodienes were reported as the sum of the
concentrations of aldrin, chlordanes, dieldrin, endrin, endosulfans, heptachlor, kepone, and mirex. DDTs and HCB were
the dominant OCs in the samples. Concentrations of OCs in
all the Hong Kong samples were comparable with those
collected from other locations along the Pearl River (Mai et
al. 2001). In the samples with positive results in the H4IIEluc bioassay, the high molecular–weight PAHs were dominant (Fig. 2).
Bioassay Analysis
Concentrations of TCDD-EQH4IIE-luc and E2-EQMVLN-luc,
which were obtained by comparing, respectively, the luminescence in the bioassay with the standard curves in the
H4IIE-luc and MVLN-luc bioassays, were less than the LOQ
in atmospheric deposition and seawater and plankton samples.
These results indicated that these samples had insignificant
dioxin-like, estrogenic, and cytotoxic effects and that the
endocrine disrupting hazards were low.
580
Fig. 2. Composition of PAHs classified by ring structures in sediment
and biotic samples with positive response in H4IIE-luc bioassay.
D = dry season; FP = fish pond; GW = gei wei; ID = Inner deep bay;
MM = middle part of the mangrove; NB = near bird-viewing hide;
NG = the gate of gei wei; PAHs = polycyclic aromatic hydrocarbons;
W = wet season.
H. L. Wong et al.
Fig. 3. AhR activity expressed as percent of the maximum response
of H4IIE-luc cells to TCDD (max € SDs) caused by extracts of sediment collected from Mai Po or A Chau (initial screening of raw
extracts) in the dry season (a) and the wet season (b). The LOQ was
defined as being three times the response to the solvent blank.
AC = A Chau; FP = fish pond; GW = gei wai; ID = Inner deep bay;
LOQ = limit of quantification; MM = middle part of the mangrove;
NB = near bird-viewing hide; NG = near the gate TCDD = 2,3,7,8tetrachlorodibenzo-p-dioxin.
AhR (Dioxin-Like) Activities Based on H4IIE-Luc Bioassay
Most of the sediment samples from both Mai Po and A Chau
contained concentrations of TCDD-EQH4IIE-luc that were
greater than the LOQ (1.2 pg, dry weight (dw) TCDD with a
linear range of 2.0 to 600 pg dw TCDD), which indicated the
presence of compounds able to induce or depress AhR- and/or
DRE-mediated gene transcription. Most of the AhR-mediated
activities observed in the raw extracts were eliminated by
sulfuric acid treatment, and this suggested that the dioxin-like
compounds in the extract were acid labile and therefore unlikely to be PCBs, dioxins, or furans (Hilscherova et al. 2001)
(Fig. 3).
Concentrations of TCDD-EQH4IIE-luc measured in fraction 1
(F1) were similar to the results for the acid-treated samples
(Fig. 4). The absence of measurable concentrations of TCDDEQH4IIE-luc in F1 is consistent with the instrumental analyses
with GC-ECD, which showed that concentrations of AhR-active PCB congeners were small (<1 pg TEQWHO-Mammals/g dw
in most sediment samples with a maximum of 68 pg TEQWHOMammals/g dw). The observation that some samples expressed
no detectable activity in the H4IIE-luc assay, when a relatively
great concentration of TEQWHO-mammal was predicted from the
concentrations of AhR-active compounds present, may have
been related to the antagonistic interactions among other
PCBs. Previous studies have also suggested that the relatively
small values of CALUX-TEFs could be caused by coplanar
Fig. 4. TCDD-like activity in the H4IIE-luc cell bioassay (expressed
as percent of maximum) in three Florisil fractionated extracts of
sediments. The maximum response was observed for 480 pg 2,3,7,8TCDD standard/ml (% TCDD maximum € SD). The LOQ was defined as being three times the response to the solvent blank. FP = fish
pond; GW = gei wai; ID = Inner deep bay; LOQ = limit of quantification; MM = middle part of the mangrove; NB = near bird-viewing
hide; NG = near the gate; TCDD = 2,3,7,8- tetrachlorodibenzo-p-dioxin.
PCB congeners using TEFWHOs (0.01 to 0.7), except for
PCB77, for which the value of CALUX-TEFs is 10-fold
greater (Sakai and Takigami 2003). Significant concentrations
Sediments and Biota from Hong Kong Mudflats
Fig. 5. Luciferase induction in the H4IIE-luc cell bioassay elicited by
PAH standards mixtures (100, 33, and 11 ng/well for each PAH).
Response magnitude is presented as percentage of the maximum response observed for 480 pg/ml 2,3,7,8-TCDD standard (% TCDD
maximum € SD). The LOQ was defined as being three times the
response to the solvent blank. ‘‘Total’’ indicates the response of all the
PAHs with different ring structures combined at a specific concentration. LOQ = limit of quantification; PAH = polycyclic aromatic
hydrocarbons.
of TCDD-EQH4IIE-luc were determined in the fraction 2 (F2) of
sediment samples collected from several sampling locations in
Mai Po (Fig. 4). The magnitude of induction was approximately 12% to 40 % of the maximum response in the assay,
which was elicited by 480 pg TCDD standard/mL. This result
might be related to the binding affinity of AhR for PAHs of the
sediment as suggested in other studies (Khim et al. 1999;
Hilscherova, 2000). This result further explained the difference
in AhR-mediated luciferase expression in H4IIE-luc cells because most of the PAHs are acid-labile (Villeneuve et al.
2002). Marginal induction of Ah-R-mediated luciferase
expression was found in fraction 3 (F3) (Fig 4.).
In sediments, PAHs might have caused a major part of the
observed AhR-mediated response because most of the TEQs
were determined to be in F2, which contained mostly PAHs
and certain OCs. These effects were not caused by dioxins and
furans, which are acid labile as was shown in the first
screening (Fig. 3). PAHs were quantified so that a mass balance could be calculated by comparing the observed TCDDEQH4IIE-luc with the TEQH4IIE-luc calculated by using the
known REP from the H4IIE-luc assay (Villeneuve et al. 2000).
In most samples collected from Mai Po, the greatest proportion
of the TEQ could be attributed to high molecular–weight
PAHs. Some studies have suggested that PAHs, especially the
high molecular–weight congeners, could mediate AhR-induced expression (Araũjo et al. 2000; Koh et al. 2001). Indeed,
PAHs with five and six rings could induce luciferase expression at concentrations of 400 ng/mL and 132 ng/mL respectively (Fig. 5). Induction was, however, less than those
reported in other studies (Machala et al. 2002), and this might
be related to the difference in exposure durations because cells
could have various mechanisms to metabolize PAHs, especially for lower molecular–weight compounds, over prolonged
periods. Thus, the moderate induction of AhR-mediated
luciferase expression observed might not play a critical role
under realistic field-exposure situations (Villeneuve et al.
581
2000). Because the TCDD%max of F2 and F3 were 10% and
20% compared with the standard curve of the TCDD exposure
experiment, EC10 and EC20 were applied to estimate the
REP-10 and REP-20 of the samples, respectively, as suggested
in Villeneuve (2000). Concentrations of TEQH4IIE-luc in F2 of
the sediment extracts ranged from several pg to 68 pg/g dw in
Mai Po. The relative potencies of PAHs in the F2 extracts,
which had been normalized to a concentration of PAHs of
approximately 1 to 14 pg TCDD-EQH4IIE-luc/g (dw) by using
the REP values for H4IIE-luc reported by Villeneuve et al.
(2002), are listed in Table 3. There was little correlation
(R2 = 0.3 to 0.6) between TCDD-EQH4IIE and the total concentrations of PAHs containing two, three, four, five, and six
rings, and this might be because of interlaboratory differences
in responses of the H4IIE-luc bioassay induced by PAHs. The
results of the mass balance calculation, together with the acidlabile nature of the TCDD-like compounds, indicated that it
was likely that PAHs caused most of the activity. Concentrations of TEQ and TCDD-EQH4IIE-luc were similar but not
identical. It should be noted that the small differences might
have been related to the fact that the TEQs were calculated
from REP values reported in the literature (Villeneuve et al.
2002), and there could have been interlaboratory variations in
the bioassay results used for calculating the REP for PAHs and
the TCDD-EQ reported here. Similarly, REP-10 was used to
estimate TCDD-EQH4IIE-luc for F3. TCDD-like responses of
the H4IIE-luc bioassay were greater with the raw extracts than
fractionated samples, and similar results have been reported by
Khim et al. (1999). Concentrations of TCDD-EQH4IIE-luc observed in the sediments studied here were small compared with
sediments collected from the Detroit River (Michaellet-Ferrier
2004), Korea (Khim et al. 1999), and the Czech Republic
(Hilscherova et al. 2001). The sediment results were therefore
consistent with the undetectable responses from seawater and
plankton samples. The potential hazard caused by dioxin-like
compounds at the study sites was therefore small.
The initial screening with the H4IIE-luc assay indicated
that raw extracts of Mugil cephalus, Mylio macrocephalus,
Tilipa zillia, and Boleophthalmus boddaerti contained relatively small concentrations of TCDD-EQH4IIE-luc, but the
extracts exhibited significant cytotoxicity (Fig. 6). The cytotoxic effects adversely affected AhR-mediated expression,
which could invalidate the estimation of relative potency.
When the biota extracts were treated with concentrated sulfuric acid, all of the TCDD-EQH4IIE-luc were found to be acid
labile, and the acid-labile chemicals were eluted in F2 in
Florisil fractionation as in the sediment samples (Figs. 7 and
8). Concentrations of TCDD-EQH4IIE-luc measured in the raw
extracts of polychaetes were below detection after Florisil
fractionation (data not shown). It remains possible that some
of these active compounds could have been trapped in the
Florisil column.
Concentrations of TCDD-EQH4IIE-luc in F1 of all animal
extracts were less than the LOQ, which is consistent with the
finding of low concentrations of dioxin-like PCBs, dioxins,
and furans in this fraction extracted from the sediment
samples (Fig. 8). Indeed, compounds with known TCDD-like
characteristics were mainly found in F2 rather than F1. Although PAHs are readily metabolized by fish or other high
trophic-level animals (Neff 2002), some residual unmetabolized PAHs could still be found in their tissues. Some
H. L. Wong et al.
582
Table 3. Concentrations of TCDD-EQH4IIE-luc (2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents as determined in the H4IIE-luc bioassay) in
sediment samples (n = 4) (TCDD-EQs € SDs) determined with in vitro bioassays.
Season
Location
Raw (pg TCDD eqv/g dry wt)
F2
F3
Fish pond
Gie wai
Near the gate
Middle mangrove
Near bird-viewing hide
Inner deep bay
68
53
27
43
47
10
20 € 50
31 € 20
5€4
9€5
18 € 11
19 € 21
15 € 5
8 € 10
NA
2€5
1€7
12 € 11
Fish pond
Gei wai
Near the gate
Middle mangrove
Near bird-viewing hide
Inner deep bay
NA
19 € 3
16 € 3
3€5
9€1
21 € 53
NA
10 €
68 €
28 €
30 €
18 €
NA
4€1
3€2
33 € 1
3€1
51 € 23
Dry
€
€
€
€
€
€
17
8
25
5
60
50
Wet
1
1
2
12
4
a
TCDD-EQH4IIE-luc determined by the response equivalency approach at the level of response equivalent to 10% (F3) and 20% (F2) median
effective concentration (EC10 and EC20) of the maximal response produced by the standard TCDD max are presented.
NA = (Not applicable:), indicates that it is not appropriate to use bioassay to estimate the TCDD-EQ in the samples.
undesirable dioxin-like and/or carcinogenic effects in both fish
and humans (International Agency for Research on Cancer
1983), a further risk assessment to determine the risks posed
by PAHs in sediments and fish from the gei wai and mudflats
was undertaken. In this study, most five-ring PAHs were found
in F2, but an underestimation of TCDD-EQH4IIE-luc from high
molecular–weight PAHs might have been caused by metabolism of lighter PAHs by the cells during incubation at 37C
(Villeneuve et al. 2000).
Cytotoxicity Assay
Fig. 6. Cytotoxicity (mean € SD) of different fractions (F1, F2, F3)
of the biota samples from Florisil fractionation. The LOQ was defined
as being three times the response to the solvent blank. AC = Gobiidase spp. from A Chau; Blk = solvent blank; D = dry season;
LOQ = limit of quantification; MD = Boleophthalmus boddaerti;
MU = Mugil cephalus; MY = Mylio marcocephalus; TP = Tilapia
zilla from Mai Po W = wet season.
metabolites of PAHs, particularly the five- and six-ring
forms, are AhR-active. Concentrations of TCDD-EQH4IIE-luc
in F2 of fish tissues were less than those of F2 of the sediment samples. It is possible that matrix effects of this fraction
might also have interfered with the luciferase expression
because of cytotoxicity, which decreased the density of viable cells. This is evident in F3 in this study (Fig. 6) and from
a previous study (Khim et al. 1999). Concentrations of
TCDD-EQH4IIE-luc in most of F3 of the animal extracts were
less than the LOQ.
Fish collected from the gei wai contained high molecular–
weight PAHs (Fig. 2). This is consistent with the PAH contents in sediments collected from the gei wai and mudflats.
Because most fish collected from Mai Po were detritivores
feeding on sediment-dwelling organisms, these fish were directly exposed to the PAHs in the sediments. Because high
molecular–weight PAHs can interact with AhR and trigger
In addition to receptor-specific effects, some chemicals can
cause cytotoxicity through more general disruption to cellular
functions (Hollert et al. 2000). Consequently, possible cytotoxic effects of the extracts had to be monitored to allow
correct interpretation of the bioassay responses. Furthermore,
cytotoxicity in the in vitro assay could also be used as an
additional assessment endpoint. In this study, marked cytotoxicity was observed in extracts, especially those from fish
collected at Mai Po. Although it is difficult to extrapolate the
results of the Promega non-radioactive cell titer proliferation
and cell protein assay on H4IIE-luc cells to other toxic effects
in fish and wildlife, it could still serve as an indicator of
possible toxic effects. The degree of confluence was less in
cells exposed to greater doses of cytotoxic extracts. The cells
under stress had a low metabolic activity and, furthermore, the
absorbance values of the cell proliferation assay and the fluorescence of the protein assay were shown to be decreased
compared with the control groups.
The cytotoxic components were acid labile, polar compounds, which were probably not the ‘‘common’’ OCs and
PAHs reported in other investigations (Liang et al. 1999; Mai
et al. 2001) and analyzed in this study. No significant cytotoxicity was observed in sediment extracts from either location, except for sediments from the gei wai during the dry
season. Only raw extracts and F3 of some sediment samples or
fish collected from Mai Po exhibited significant cytotoxicity to
Sediments and Biota from Hong Kong Mudflats
583
Fig. 7. TCDD-like response in H4IIE bioassay for biota samples collected during the dry and wet seasons (first screening). Response magnitude
is presented as percentage of the maximum response observed for 480 pg 2,3,7,8-TCDD standard/ml (% TCDD maximum). The LOQ was defined
as being three times the response to the solvent blank. AC = Acetes spp. from A Chau; CO = Cyclina orientalis; GB(A) = Gobiidae sp. A;
GB(B) = Gobiidae sp. B; LOQ = limit of quantification; MD = B. boddaerti; MU = M. cephalus; MY = M. marcocephalus; PL = polychaetes;
ME = Metapenaeus ensis; TCDD = 2,3,7,8- tetrachlorodibenzo-p-dioxin; TP = T. zilla from Mai Po.
Fig. 8. TCDD-like response in H4IIE bioassay for biota (F1, F2, and
F3) following Florisil fractionation. Response magnitude is presented
as the percentage of the maximum response observed for 480 pg
2,3,7,8-TCDD standard/mL (% TCDD maximum). The LOQ was
defined as being three times the response to the solvent blank. D = dry
season; LOQ = limit of quantification; MD = B. boddaerti; MU =
M. cephalus; MY = M. marcocephalus; TCDD = 2,3,7,8- tetrachlorodibenzo-p-dioxin; TP = T. zilla (from Mai Po); W = wet season.
H4IIE-luc cells (Figs. 3 and 6). However, such cytotoxicity
was not evident in other biotic samples from Mai Po or A
Chau. The occurrence of cytotoxic chemicals in the raw ex-
tracts and F3 was similar to the results of Cerna et al. (1996)
and Hollert et al. (2000). The cytotoxicity observed was unlikely to be caused by the lipids in the fish extracts because no
significant cytotoxicity was triggered by an equivalent amount
of solvent-extracted and nonextracted peanut oil and cod-liver
oil under similar experimental conditions (data not shown).
Significant cytotoxicity to H4IIE-luc cells was observed in
undiluted (1x) raw extracts of the environmental samples, and
the cytotoxicity decreased with dilution. Synthetic cocktails of
appropriate concentrations of OCs and PAHs caused minimal
cytotoxicity, i.e., <8% to 13 % of the cytotoxicity caused by F3
of the fish extracts. Higher molecular–weight PAHs and acidlabile OC pesticides may therefore have contributed part of the
observed cytotoxicity in the environmental samples, although
none of them was present individually at concentrations sufficiently great to cause cell death. Moreover, only the undiluted mixture of OCs and PAHs was shown to be capable of
triggering cell death, whereas significant cytotoxicities were
recorded in three-fold diluted fish sample extracts. In the
present study, many OC pesticides (including dieldrin, endosulfan I, endrin, kepone, and methoxychlor) and two six-ring
PAHs [dibenzo(1,2,5,6)anthracene and benzo(g,h,i)perylene]
present in the raw extracts were acid labile. Because lM
concentrations of these chemicals would be expected to cause
significant cytotoxicity (Bayoumi et al. 2000; Raychoudhury
H. L. Wong et al.
584
ER-Active Compounds (MVLN Bioassay)
Fig. 9. Effect of storage temperature on the cytotoxicity (mean € SD)
of extracts of M. marocephalus from Mai Po after being stored for 3
months.
Fig. 10. Luciferase induction (mean € SD) expressed as percent of
maximum induction in MVLN cells by raw and fractionated sediment
samples collected from Mai Po and A Chau. Response magnitude is
presented as percentage of the maximum response observed for 272
pg E2/ml (% E2 maximum.). The LOQ was defined as being three
times the response to the solvent blank. LOQ = limit of quantification. AC = A Chau; FP = fish pond; GW = gei wei; ID = Inner Deep
Bay; MM = middle part of the mangrove; NB = mudflat near the
bird-viewing hide; and NG = the gate of gei wei.
and Kubinski 2003), it was unlikely that any of the measured
chemicals were the cause of the cytotoxicity. When samples
were stored at –20C, 4C, and room temperature for a period
of 3 months, only those extracts stored at 4C or –20C
exhibited cytotoxicity (Fig. 9). These results all suggested that
the cytotoxic components were polar and labile when the extracts were stored at room temperature. Further study would be
required to identify and characterize the potential toxicants or
biotoxins and specific part(s) of the fish containing these
cytotoxic compounds.
Because the toxicity of F3 to H4IIE-luc cells was significant
for fish extracts, bioassay-derived TEQ of dioxin might provide only a qualitative estimate. An underestimation of REPTCDD-EQ was found in all the samples with induction of
H4IIE-luc when the calculation of toxic equivalency was based
on a summation of PAHs only, probably because the cell
density and dose-related response of luciferase expression per
well were not standardized.
The standard curves used for calculation were obtained from
the mean of the dose response curves generated by exposing
MVLN cells to E2. The linear ranges of the curves were
from 0.1 to 500 pg E2/g dw. The LOQ was 0.01 pg/g dw.
Among the sediment extracts from all sampling sites, fractions F2 and/or F3 of the extracts of gei wai, fish pond, near
the gate, and near the bird-viewing hide from Mai Po were
found to elicit significant estrogenic activities as measured
by the MVLN bioassay (Fig. 10). Luciferase indication was
greater in the wet season samples than the dry season samples. Bioassay derived E2 equivalents (EEQMVLN-luc), based
on REP-50, for these samples ranged from 10 to 460 pg E2/g
dw (Table 4). Sediments from the gei wai contained significantly greater EEQMVLN-luc than those from the fish pond.
The observed difference in EEQMVLN-luc concentrations between the two locations might be related to both tidal
transport and atmospheric deposition (especially during the
wet season).
Among those potential estrogenic compounds, polar chemicals eluted in F3 by Florisil fractionation were the main
contributor for triggering the response in MVLN-luc cells.
PCDDs, PCDFs, and PCBs have been reported to elicit antiestrogenic responses in vitro (Chen et al. 2001). Although
small concentrations of mono-ortho–substituted PCBs and diortho–substituted PCBs were measured in the sediment samples (Table 1), they were at concentrations insufficient to elicit
a response in the MVLN assay. The estrogenic compounds
were likely to be acid-labile substances rather than the acidstable chemicals such as PCBs and dioxins. OC pesticides—
such as chlordane, heptachlor, endosulfan, DDTs, aldrin, and
dieldrin—have also been reported to cause weak estrogenic
responses in vitro. Such responses were, however, reported at
concentrations generally exceeding 10 lM, e.g., 1 to 5 lg/g for
OC pesticides (Andersen et al. 2002). Concentrations of OC
insecticides measured in the Mai Po sediment extracts were
approximately 1000-fold less than the concentrations reported
to elicit estrogenic responses. Induction of E2-like responses
was therefore unlikely to be caused by OC pesticides solely in
F2 and F3. Results of previous studies have indicated that
polar chemicals such as endogenous estrogens, synthetic
estrogens, xenoestrogens, and other unknown acid-labile
chemicals might be important in eliciting estrogenic responses
(Khim et al. 1999; Hilscherova et al. 2002). Concentrations of
these compounds were not measured in the current study.
Although there was no clear indication of estrogenic effects,
further investigations are needed to identify the major contributors of E2-like responses in the sediment samples. The
results of this study indicate that traditional chemical analysis
and risk assessments of OC pesticides, PCBs, and PAHs might
not be sufficient to estimate the risks caused by environmental
estrogen agonists.
Acknowledgments. This study was supported by a Central Allocation
Grant (No. 8730020) awarded by the Research Grants Council, Hong
Kong, and the Area of Excellence Scheme under the University
Grants Committee of the Hong Kong Special Administration Region,
China (Project No. AoE/P-04/2004).
585
Sediments and Biota from Hong Kong Mudflats
Table 4. EEQMVLN-luc (mean € SD) of E2-like response of sediment samples by MVLN-luc bioassay
EEQMVLN-luc (pg E2/g dw)
Location
Fraction
Wet Season
Dry Season
Deep bay
Near bird viewing hide
Middle mangrove
Near gei wai
Gei wai
Whole
Whole
Whole
Whole
F1
F2
F3
Whole
F1
F2
F3
Whole
Whole
ND
ND
ND
ND
ND
62 € 14a
464 € 112a (11 € 31 – 490 € 210)b
421 € 113a (115 € 54 – 440 € 110)b
ND
ND
43 € 21a
50 € 1.0a
ND
ND
ND
ND
ND
363 € 53a (82 € 41 – 332 € 18)b
320 € 78a (110 € 51 – 280 €120)b
ND
ND
54 € 17a
42 € 22a
ND
Fish pond
A Chau
a
REP-50 = Point estimate of relative potency (REP = EC-std/ECx) where 50% std max is the selected magnitude of response was applied to
estimate the EEQMVLN-luc.
b
REP20-80 range = Range of relative potency estimates generated from multiple point estimates for responses ranging from 20 to 80% std max.
dw = Dry weight.
N.D. Data below detection limit.
References
Andersen HR, Vinggaard AM, Rasmussen TH, Gjermandsen IR,
Bonefeld-Jørgensen EC (2002) Effects of currently used pesticides in assays for estrogenicity, androgenicity, and aromatase
activity in vitro. Toxicol Appl Pharmacol 179:1–12
Araũjo CSA, Marques SAF, Carrondo MJT, Goncalves LMD (2000)
Chemical-activated luciferase gene expression (CALUX): A novel in in vitro response of the brown bullhead catfish (BB) and
rainbow trout (RTG-2) cell lines to benzo[a]pyrene. Sci Total
Environ 247:127–135
Bayoumi AE, GarcTa-FernUndez AJ, Fett C, CubrTa JC, Reguera RM,
OrdVňez C, et al. (2000) Evaluation of basal cytotoxicity of organochlorines insecticides by in vitro alternative methods using
BF-2 and RTG-2 fish cell lines. Ecotoxicology and Enviromental
Restoration 3:81–86
Cerna M, Pastorkova A, Smid J, Bavorova H, Ocadlikova D, Rossner
P, et al. (1996) Genotoxicity of industrial effluents, river waters
and their fractions using the Ames test and in vitro cytogenetic
assay. Toxicol Lett 88:191–197
Chen I, Hsieh T, Thomas T, Safe S (2001) Identification of estrogeninduced genes downregulated by AHR agonists in MCF-7 breast
cancer cells using suppression subtractive hybridization. Gene
22:207–214
Connell DW, Fung CN, Minh TB, Tanabe S, Lam PKS, Wong BSF,
et al. (2003) Risk to breeding success of fish-eating Ardeids due
to persistent organic contaminants in Hong Kong: Evidence from
organochlorine compounds in eggs. Water Res 37:459–467
Hilscherova K, Machala M, Kannan K, Blankenship AL, Giesy JP
(2000) Cell bioassays for detection of aryl hydrocarbon (AhR)
and estrogen receptor (ER) mediated activity in environmental
samples. Environ Sci Pollut Res Int 7:159–171
Hischerova K, Kannan K, Kang Y-S, Holoubek I, Machala M,
Masunaga S, et al. (2001) Characterization of dioxin-like activity
of sediments from a Czech river basin. Environ Toxicol Chem
20:2768–2777
Hilscherova K, Kannan K, Holoubek I, Giesy JP (2002) Characterization of estrogenic activity of riverine sediments from the Czech
Republic. Arch Environ Contam Toxicol 43:175–185
Hollert H, DErr M, Erdinger L, Braunbeck T (2000) Cytotoxicity of
settling particulate matter and sediments of Necar River (Germany) during a winter flood. Environ Toxical Chem 19:528–534
International Agency for Research on Cancer (1983) Polycyclic aromatic compounds. Part I: Chemicals, environmental and experimental data. IARC Monographs on the Evaluation of the
Carcinogenic Risk of Chemicals to Humans. Volume 32. International Agency for Research on Cancer, Lyon, France
Khim JS, Villeneuve DL, Kannan K, Lee KT, Snyder SA, Koh CH,
et al. (1999) Alkylphenols, polycyclic aromatic hydrocarbons,
and organochlorines in sediment from lake Shihwa, Korea:
Instrumental and bioanalytical characterization. Environ Toxicol
Chem 18:2424–2432
Khim JS, Villeneuve DL, Kannan K, Hu WY, Giesy JP, Kang SG,
et al. (2000) Instrumental and bioanalytical measures of persistent
organochlorines in blue mussel (Mytilus edulis) from Korean
coastal waters. Arch Environ Contam Toxicol 39:360–368
Koh CH, Kim GB, Maruya KA, Anderson JW, Jones JM, Kang SG
(2001) Induction of the P450 reporter gene system bioassay by
polycyclic aromatic hydrocarbons in Ulsan Bay (South Korea)
sediments. Environ Pollut 111 3:437–445
Liang Y, Wong MH, Shutes RBE, Revitt DM (1999) Ecological risk
assessment of polychlorinated biphenyl contamination in the Mai
Po marshes nature reserve, Hong Kong. Water Res 33:1337–1346
Machala M, VondrUček J, BlUha L, Ciganek M, Neča J (2002) Aryl
hydrocarbon receptor-mediated activity of mutagenic polycyclic
aromatic hydrocarbons determined using in vitro reporter gene
assay. Mutat Res 497:49–62
Mai BX, Fu JM, Sheng GY, Kang YH, Lin Z, Zhang G, et al. (2001)
Chlorinated and polycyclic aromatic hydrocarbons in riverine and
estuarine sediment from Pearl River Delta. China Environ Pollut
117:457–474
Murk AJ, Legler J, Denison MS, Giesy JP, van de Guchte C, Brouwer A
(1996) Dioxin-like toxic potency in ForsterÕs tern eggs from Green
Bay, Lake Michigan, North America. Chemosphere 26:2079–2084
MEller JF, Gaus C, Prange JA, Olaf P, Poon KF, Lam MHW, et al.
(2002) Polychlorinated dibenzo-p-dioxins and polychlorinated
dibenzofurans in sediments from Hong Kong. Mar Pollut Bull
45:372–378
Neff JM (2002) Polycyclic aromatic hydrocarbons in the ocean. In: Neff
JM, (ed) Bioaccumulation in marine organisms—Effect of contaminants from oil well produced water. Elsevier Science, UK,
Raychoudhury SS, Kubinski D (2003) Polycyclic aromatic hydrocarbon-induced cytotoxicity in cultured rat sertoli cells involves differential apoptotic response. Environ Health Perspect 111:33–38
586
Sakai SI, Takigami H (2003) Integrated biomonitoring of dioxin-like
compounds for waste management and environment. Ind Health
41:205–214
Snyder SA, Villeneuve DL, Snyder EM, Giesy JP (2001) Identification and quantification of estrogen receptor agonists in wastewater effluents. Environ Sci Technol 35:3620–3625
Tillitt DE, Kubiak TJ, Ankley GT, Giesy JP (1996) Chemical-activated luciferase gene expression (CALUX): A novel in vitro
bioassay for ah receptor active compounds in sediments and pore
water. Fundam Appl Toxicol 33:149–160
van den Berg M, Birnbaum L, Bosveld ATC, Brunstrçm B, Cook P,
Feeley M (1998) Toxic equivalency factors (TEFs) for PCBs,
H. L. Wong et al.
PCDDs, PCDFs for humans and wildlife. Environ Health Perspect
106:775–792
Villeneuve D L, Blankenship AL, Giesy JP (2000) Derivation and
application of relative potency estimates based on in vitro bioassay results. Environ Toxicol Chem 19:2835–2843
Villeneuve DL, Khim JS, Kannan K, Giesy JP (2002) Relative
potencies of individual polycyclic aromatic hydrocarbons to induce dioxinlike and estrogenic responses in three cell lines.
Environ Toxicol 17:128–137
Zheng GJ, Richardson BJ (1999) Petroleum hydrocarbons and polycyclic aromatic hydrocarbons (PAHs) in Hong Kong marine
sediments. Chemosphere 38:2625–2632