* Department of Zoology, University of Heidelberg, Im Neuenheimer Feld 230, D-69120 Heidelberg, Germany;
ESWE-Institute for Water Research and Water Technology, Söhnleinstrasse 158, D-65201 Wiesbaden, Germany;
Department of Food Toxicology, Center of Food Science, School of Veterinary Medicine Hannover, Bischofsholer Damm 15, D-30173 Hannover, Germany;
Bio Detection Systems b.v., Kanaal Laboratorium, Badhuisweg 3, 1031 CM Amsterdam, The Netherlands; and
¶ Department of Hygiene, University of Heidelberg, Im Neuenheimer Feld 354, D-69120 Heidelberg, Germany
Received April 16, 2003; accepted May 27, 2003
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ABSTRACT |
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Key Words: estrogenicity; STP effluent; surface water; Rhine; rainbow trout; primary hepatocytes; HEK 293 cells; yeast estrogen screen; solid phase extracted water samples; chemical analysis.
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INTRODUCTION |
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In Germany, data on estrogenic effects on aquatic organisms including fish due to STP effluents are scarce. A previous study using extracted water samples from two municipal STPs and river Rhine surface water sampled at the Rhine-Neckar-triangle (southern Germany) indicated estrogenic activities in both STP effluents and the surface water (Pawlowski et al., unpublished data). In the same study, steroidal estrogens (SE) and phytoestrogens (PE) could be identified in the STP effluents using chemical analysis at values up to 26 ng/l (SE) and 1.2 µg/l (PE), and in Rhine water of about 4 ng/l for steroidal estrogens and 0.25 µg/l for phytoestrogens. In order to evaluate the impact of domestic discharge on aquatic organisms, e.g., fish, the present study was designed to compare vitellogenin-mRNA induction in caged male or one-year-old immature rainbow trout (Oncorhynchus mykiss) with in vitro data obtained from (1) primary rainbow trout hepatocyte cultures (Islinger et al., 1999; Pawlowski et al., 2000
), the (2) yeast estrogen screen (YES; Routledge and Sumpter, 1996
), and (3) transfected human embryonic kidney 293 cells (HEK 293 cellER-luciferase assay; Kuiper et al., 1998
; Meerts et al., 2001
) exposed to extracts from water obtained from the sites where fish were exposed over nine months of sampling (from November 2000 to September 2001). Biological effects were related to data from corresponding chemical analyses.
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MATERIAL AND METHODS |
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Fish exposure at the STPs.
To prevent oxygen depletion in STP effluents, bypass systems consisting of 600 l glass fiber tanks were connected to the effluent via water pumps (4000 L/h; Eheim, Deizisau, FRG). The tank was covered with stainless steel grids to prevent loss of experimental fish. The water was aerated using a commercially available air pump (Rebie, Bielefeld, FRG). Measured oxygen concentrations in the tanks varied between 3.9 and 5.4 mg/l, and temperature ranged from 10 to 19°C, depending on the season (for details, see Table 1). In each experiment, five male rainbow trout were exposed to the effluents for a period of two weeks in November 2000 (site A), March (site B), May and September 2001 (sites A and B; Table 1
). Control fish were kept under similar conditions, receiving tap water only.
Fish exposure at the river Rhine.
Due to space limitations at the Rheingütestation Worms, one-year-old rainbow trout (between 15 and 25 cm) were used to measure estrogenic effects in Rhine water. For each experiment, 15 one-year-old rainbow trout were transferred to 100 l glass aquaria, receiving water from lane 4 of the river Rhine (right side), which is under the influence of, among others, the sewage effluent from site B. Exposure of fish was carried out over periods of 2 weeks in March, May, and September 2001 (Table 1). Control fish were kept under similar conditions, receiving tap water only.
Water sampling and solid phase extraction.
While fish were exposed to sewage effluents and river Rhine water, respectively, water samples (0.5 l) were taken concomitantly on a daily basis from every site and extracted using C18 solid phase-extraction columns (1 g, Bakerbond, J. T. Baker, Deventer, The Netherlands) within 24 h upon arrival in the laboratory. All subsequent steps were carried out in duplicate: Water samples were filtered over 0.1 µm glass fibre filters (type C5, MembraPure, Bodenheim, FRG), acidified with concentrated H2SO4 to pH 2, divided into two 1-liter samples and extracted using reverse phase C18 solid phase extraction columns (RP-C18 SPE; 1 g; Bakerbond, J. T. Baker, Deventer, The Netherlands), which had been conditioned with 3 x 3 ml hexane, 3 x 3 ml acetone as well as 1 x 3 ml deionized water according to a protocol by Spengler et al. (2001). After extraction, columns were centrifuged for 10 min at 2000 x g and dried under a nitrogen stream. Elution was carried out with 2 x 5 ml acetone. Eluted samples were blown to dryness under a nitrogen stream, reconstituted in 2.5 ml ethanol and stored at 4°C until use in the yeast estrogen screen (up to four weeks). All extracted water samples from one exposure regime (lasting 14 days) were mixed to obtain an average estrogenic activity over the entire experimental period. Deionized water and tap water were treated and extracted according to the same protocol and were used as process and field controls, respectively.
The yeast estrogen screen.
All medium constituents were obtained from Sigma (Deisenhofen, FRG) except for chorophenol red-ß-galactopyranoside (CPRG; Roche Diagnostics, Mannheim, FRG). Recombinant yeast cells (Saccharomyces cerevisiae) were kindly provided by John P. Sumpter from Brunel University, Uxbridge, U.K. Yeast cells were cultured, and the assay was carried out as described by Routledge and Sumpter (1996) with slight modifications. Briefly, 200 µl of the respective water sample extract (concentrated by a factor of 2000 if compared to the initial water sample) were added to the first column of a 96 well plate (TTP, Renner, Dannstadt, FRG) and then diluted with 100 µl ethanol in a 1:2 series across the plate, resulting in a 1000x concentrated extract as the highest concentration. 17ß-Estradiol (E2) was used as a positive control with 10 nM as the highest concentration in the first well. Ethanol (100 µl) was used as negative control in the YES. Values were compared with E2 induction rates and expressed as estradiol equivalents (E2-EQs).
ER-luciferase assay with HEK 293 cells.
Human embryonic kidney (HEK) 293 cells were kindly provided by Paul van der Saag, Hubrecht Laboratory, Netherlands Institute for Developmental Biology at Utrecht, The Netherlands. Extracted water samples were tested in HEK 293 cells stably transfected with recombinant human estrogen receptor (HEK 293 ER or 293 ERßs-luciferase assays; Kuiper et al., 1998
; Lemmen et al., 2002
). Cells were cultured and exposed to extracted water samples as described by Meerts et al. (2001)
and Kuiper et al. (1998)
, however, with 5% CO2 instead of 7.5%. Briefly, cells were trypsinized and resuspended in steroid-free assay medium and seeded in 96-well plates at a density of 15,000 cells per well in 200 µl assay medium. After 48 h of incubation, when cell density had reached 5060% confluence, the assay medium was replaced by incubation medium containing a 1000-fold dilution of the extracted water samples. Solvent concentrations never exceeded 0.1% in the culture medium, resulting in 2x concentrated water sample extracts in the test. 17ß-Estradiol dissolved in ethanol was used as a positive control at concentrations of 0.1, 1, 10, 100, and 1000 pM. After 24 h at 37°C, the plates were transferred onto ice and the medium was removed by suction. Cells were lysed, deep-frozen, and analyzed for luciferase activity. Each sample was tested in five independent replicates for its estrogenicity in the ER-luciferase assay. Values were compared with E2 induction rates and expressed as estradiol equivalents (E2-EQs).
Primary rainbow trout hepatocyte cultures.
Two-year-old male rainbow trout weighing 250350 g were used as donors for primary hepatocyte cultures. Fish were anesthetized with benzocaine (Sigma), and hepatocytes were isolated as described by Islinger et al. (1999) with slight modifications by Pawlowski et al. (2000)
. For culture, 2 ml aliquots of the hepatocyte suspension (1 x 106 cells/ml) were seeded in 24-well cell culture plates (TTP, Renner, Dannstadt, FRG) and incubated at 18°C.
Prior to measuring estrogenicity of extracted water samples, the solvent ethanol was replaced by dimethyl sulfoxide (DMSO). After 24 h of cell regeneration, 1 ml of the supernatant medium was replaced by the appropriate amount of fresh medium containing 0.15 % (v/v) DMSO for negative control, as well as 17ß-estradiol (E2) as positive control (0.1 and 1 nM) or the respective extract (concentration factors of 12.5, 25, and 50x) dissolved in DMSO. After 48 h of exposure at 18°C, 1 ml of medium was replaced.
For vitellogenin-mRNA (VTG-mRNA) determination, primary rainbow trout hepatocyte cultures were harvested after 96 h of exposure. Total RNA was prepared, and VTG-mRNA was quantified as described previously (Islinger et al., 1999; Pawlowski et al., 2000
). Samples were tested in independent cell cultures from two different fish, VTG-mRNA detection was carried out in six replicates. The amount of VTG-mRNA was determined using VTG-mRNA standards ranging from 300 to 0.12 pg. Values were compared with E2 induction rates and expressed as 17ß-estradiol equivalents (E2-EQs).
Isolation and quantification of vitellogenin-mRNA in exposed rainbow trout.
After two weeks of exposure, fish were anesthetised with benzocaine and opened ventrally. Pieces of liver (0.3 mm3) were cut and stored immediately in RNAlater (Qiagen, Hilden, FRG) for up to 6 h at room temperature. Total RNA was prepared using QIA-Shredder and RNeasy Mini-Kits (Qiagen; Islinger et al., 1999; Pawlowski et al., 2000
). Concentration and purity of RNA were determined by photometry at 260/280 nm, and RNA was stored at -70°C until use. Intactness of the RNA samples was checked by denaturating agarose/formaldehyde gel electrophoresis and ethidium bromide staining for 28 S and 18 S rRNA. Vitellogenin-mRNA was quantified in a dot blot/RNase protection assay (Islinger et al., 1999
; Pawlowski et al., 2000
; Zhan et al., 1997
). Vitellogenin-mRNA detection was carried out in six replicates each from three independent fish individuals. Values were compared with E2 induction rates and expressed as estradiol equivalents (E2-EQs).
Chemical analysis of water samples.
For chemical analysis of estrogens and phytoestrogens, water samples (1 l from each site collected over 24 h) were taken on 10 September 2001. Water samples were kept at 4°C during transport and were concentrated with RP-C18 (see above) within the next 24 h. The analytical methods for estrogens, fecal- and phytosteroids, and flavanoids were described in detail by Ternes et al. (1999a,b
,c
).
Statistical analysis.
In the YES, values higher than the triple standard deviation of the negative controls were registered as statistically significant. Statistically significant differences between means of luciferase activity in the ER luciferase assay, and vitellogenin-mRNA induction in vivo were first analysed for normality/homogeneity of variance before subjected to one-way ANOVA followed by Dunns post-hoc test (Sigma Stat, SPSS-Jandel, Erkrath, F.R.G.). Given the limited number of replicates (n = 2), vitellogenin-mRNA induction in primary rainbow trout hepatocytes was not evaluated statistically.
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RESULTS |
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Extracted Water Samples in HEK 293 Cells
With respect to the induction of estrogen receptors (ER) and ß activity in HEK 293 cells, no similarities in the reaction scheme of both types could be observed at any site (Table 3
). E2-EQs in HEK 293 cells were within the pM (pg/l) range with highest values for both ERs at sites A and B. Thus, at site A statistically significant values ranged from 0.002 ± 0.0006 nM (0.7 ± 0.2 ng/l) to 0.008 ± 0.002 nM (2.1 ± 0.7 ng/l) E2-EQs for ER
in May 01 and Nov 00, respectively. In contrast, values of site A samples remained below 0.003 ± 0.001 nM (0.9 ± 0.4 ng/l) E2-EQs within the ERß system. Site B showed low levels of E2-EQs for ER
up to 0.001 ± 0.00001 nM (0.3 ± 0.03 ng/l), whereas ERß values increased from 0.003 ± 0.002 (0.85 ± 0.52 ng/l) in Mar 01 to 0.01 ± 0.001 nM (2.7 ± 0.4 ng/l) E2-EQs in Sep 01 (Table 3
). At site C, no statistically significant elevated E2-EQs could be observed in both ER
and ß tests at any investigated season, if compared to the negative controls (Table 3
).
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DISCUSSION |
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Overall, there was a good correlation between results obtained from the in situ (rainbow trout in vivo) and in vitro experiments (YES, HEK 293 cells, primary hepatocytes) as well as the chemical analysis of representative water samples. Although primary rainbow trout hepatocytes data were not statistically evaluated, a tendency in reaction similar to those observed for the other in vivo and in vitro experiments could be observed. Comparing the estrogenicity of the three investigated sites, site A (STP 1 effluent) clearly displayed highest activity, followed by site B (STP 2 effluent) and C (river Rhine; Table 6).
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There was an apparent decrease in the sensitivity to estrogenic activity across the in vitro biotests applied with primary rainbow trout hepatocytes > YES > HEK 293 cell line. It is well known that biotransformation as well as bioactivating processes of chemicals including potential estrogens take place in primary fish hepatocytes and thus enhance their ability to detect potential estrogens (Braunbeck, 1992; Scholz et al., 1998
). Thus, the higher sensitivity of primary hepatocytes might be explained by differential bioactivation of estrogens in the detection systems used.
The relatively low levels of E2-EQs in HEK 293 cells might be due to the relatively high dilution rates of the extracts tested. Given the spectrum of substances analyzed chemically, the differences in reactivities of the two subtypes investigated ( and ß; see Kuiper et al., 1998
) could not be attributed to the specific chemicals. At present, it cannot be resolved whether these discrepancies were caused by variations within the test systems or whether they were caused by unknown (unanalyzed) endocrine active chemicals, for example, potential anti-estrogens or androgens with different modes of actions.
The estrogenic responses in the biotests used (particularly the in situ methods) may have been influenced by water temperature and mean flow rates at the different sampling periods. While E2 and EE2 concentrations were rather similar in STP effluents from September 97, January and October 1998 (Huang and Sedlak, 2001), a more in-depth study by Routledge et al. (1998)
into the time course of steroid levels over several months clearly documented much higher levels of steroidal estrogens during winter months from November 1997 to March 1998, if compared to summer months. This phenomenon might explain the decrease in VTG-mRNA expression in exposed rainbow trout at site A from November 00 to September 01. With respect to site B, dilution effects are likely to be responsible for the rather low levels of estrogenic activity in exposed fish during the winter time at higher dilution rates. In fact, several previous studies have indicated a reduction in estrogenic activity in parallel to an increase in the dilution factor of STP effluent in both laboratory and field experiments using Japanese medaka (Oryzias latipes; Metcalfe et al., 2001
) and roach (Rutilus rutilus; Rodgers-Gray et al., 2000
, 2001
).
The failure to detect any induced VTG-mRNA in 1-year-old male and female rainbow trout exposed to Rhine river water might be attributed to the age since fish were sexually inactive, although sexual development had been completed at this age. However, it seems more likely that either the low estrogenic activity as could be measured in the in vitro biotests or other chemicals acting as anti-estrogens, androgens, etc. interfere with VTG-mRNA induction in exposed fish at site C. Moreover, dilution effects most likely reduced the estrogenic activity in the river water below detection limits (mean flow rate: about 15002000 m3/s, Rhine at Worms versus 5000 m3/day at STP 2). Therefore, further contributors to the overall endocrine activity upstream the river must be assumed; in fact, other municipal STPs as potential sources for steroidal estrogens and a paper mill as a potential source for phytoestrogens) are located 10 and 15 km upstream. Thus, androgenic effects such as those observed in a previous study using guppy (Poecilia reticulata) and eelput (Zoarces viviparous) exposed to paper mill effluents (Larsson et al., 2000, 2002
) cannot be excluded. This clearly documents an advantage of in vivo systems, since all in vitro assays would inevitably fail to detect this. Unfortunately, no fish were caged upstream of site C at this time to further corroborate conclusions.
Taking into account previous studies, which indicated the presence of steroidal estrogens at biologically active lower ng/l-ranges in the river Rhine (E1 up to 2.9; E2 up to 2.9 and EE2 up to 4.3 ng/L; Belfroid et al., 1999; Ternes et al., 1999c
), estrogenic effects in fish could not be excluded at this time. Such effects might be even more pronounced by extended exposure of local fish in the river, as could be demonstrated for roach exposed to different STP effluent concentrations for various exposure durations (Rodgers-Gray et al., 2000
). According to Panter et al. (2000)
, fluctuations in the concentrations of estrogenically active chemicals within relative short periods of time do not necessarily reduce estrogenic impact on fish.
Given the results from the chemical analysis of steroidal estrogens from sites B and C, it may be assumed that the measured fecal- and phytosteroids, resorcyclic lactones, and flavonoids also contributed to the overall estrogenic activity of the water samples. Several studies indicated that phytoestrogens, with estrogenic activities 103104 times lower than that of the natural steroid E2, are biologically active at the µg/l range (Dubé and MacLatchy, 2001; Tremblay and Van der Kraak, 1999
). Thus, the concentrations of phytoestrogens measured in the present study could well be biologically relevant. Moreover, it cannot be excluded that other chemicals such as alkylphenols and bisphenols might influence the estrogenicity of water samples. Nevertheless, previous studies showed that most estrogenic activity, for example, of STP effluents, could be attributed (up to 90%) to steroidal estrogens (Desbrow et al., 1998
). However, these studies used the YES assay to determine this, which exclusively identifies ER agonists. Estrogenic activity may occur through a whole variety of indirect mechanisms such as disruption of E2 degradation/synthesis and biotransformation. All of these potential compounds were not considered, but will still cause estrogenic activity in vivo. This is why in vivo and in vitro assays should always be applied in parallel to determine the identity of non-ER agonist chemicals that are "estrogenic." Especially androgenic, antiestrogenic, and antiandrogenic effects that might be present in surface waters (Nimrod and Benson, 1996
; Thomas et al., 2001
) will normally not be identified in in vitro systems when used alone.
The correlation between results from the chemical analyses in the present study with from other studies at other locations (0.4220 ng/l for E1, 0.519 ng/l for E2, 0.57.5 ng/l for EE2 in municipal effluents; Belfroid et al., 1999; Rodgers-Gray et al., 2000
; Ternes et al., 1999b
,c
) suggests that steroidal estrogens regularly occur in STP effluents and rivers (e.g., small rivers with lower dilution effects) and cause effects in biological systems. An impact of estrogenically active chemicals on fish population might thus be a more general problem in European freshwater systems. As a consequence, the presence of a broad spectrum of hormonally active chemicals in STP effluents and surface waters (Bätscher et al., 1999
; Gülden et al., 1997
; Mellanen et al., 1996
; Tavera-Mendoza et al., 2002
; Tremblay and Van der Kraak, 1999
; Tyler et al., 2000
; Willingham and Crews, 1999
) requires a combination of both biological and chemical detection systems.
The present study clearly documents that in situ exposure of rainbow trout is capable of identifying potential estrogenic effects compounds in STP effluents; this in vivo result could be replicated using primary rainbow trout hepatocytes. Both YES and HEK 293 cells are less potent in their ability to detect estrogenic effects showing only about 50 and 10% of E2-EQs, if compared to primary rainbow trout hepatocytes. Furthermore, the presence of estrogenic activity in both STP effluents and river Rhine water indicates the risk of endocrine disruption in native fish populations. Further investigations are needed to quantify this risk of hormonal disruption of feral fish species.
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ACKNOWLEDGMENTS |
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NOTES |
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