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Keywords:

  • estrone;
  • endocrine disruption;
  • Japanese medaka;
  • fish multi-generation test;
  • fish short-term reproduction assay

ABSTRACT

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

The most potent chemicals potentially causing adverse effects on fish species are estrogens in human waste. Sewage is a source of these estrogens and it is difficult to reduce. In particular, although the bioactivity of estrone is estimated to be about half of that of estradiol, multiple studies report that more than 100 ng l–1 of estrone can be detected in urban rivers, including discharges from sewage treatment works; approximately two times as high as estradiol. Few studies have been conducted to investigate the long-term effects of estrone on wildlife; therefore, we conducted fish multigeneration test using Japanese medaka (Oryzias latipes). Medaka were exposed to estrone for 27 weeks across three generations in environmentally relevant concentrations, being 5.74, 11.4, 24.0, 47.1 and 91.4 ng l–1. No effects on reproduction were observed in the first generation; however, a decline in egg production and fertility was observed in the second generation exposed to 91.4 ng l–1 estrone, which is lower than some known environmental concentrations in urban environments. Furthermore, histopathological abnormalities were observed in the third generation exposed to both 47.1 and 91.4 ng l–1, suggesting that estrone possibly exerts severe effects on the third or later generations. However, appearances of testis–ova were observed in the second and third generation they were not consistent with actual effects on reproduction, notwithstanding the testis-ova is regarded as the key evidence for endocrine disruption. Accordingly, we consider that qualitative measurement of abnormalities using histopathological observations is required for appropriate evaluation of endocrine disruption. Copyright © 2014 John Wiley & Sons, Ltd.


Introduction

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

Although modernization of industry and a large number of human-made chemicals have made our lives richer and more convenient, at times they have also caused unforeseen adverse effects on both humans and wildlife. Some chemicals, called endocrine disrupting chemicals (EDCs), may covertly threaten the sustainability of animals by interfering with the body's endocrine system in even small amounts or concentrations (Colborn et al., 1996; McLachlan et al., 2002). From the 1980s, many studies have suggested that anthropogenic estrogens or estrogenic compounds exert endocrine disruption on wildlife (Tyler et al., 1998). In fish, induction of vitellogenin (VTG) in males, abnormal testes, and hermaphroditism have been reported in common roach (Rutilus rutilus), rainbow trout (Oncorhynchus mykiss) and common carp (Cyprinus carpio) in rivers primarily downstream from sewage treatment works in the UK (Harries et al., 1996; Jobling and Sumpter, 1993; Jobling et al., 1998; Purdom et al., 1994), USA (Folmar et al., 1996; Goodbred et al., 1997) and Japan (Hashimoto et al., 2000; Wanami et al., 2002). Other studies have been concerned with endocrine disruption in reptiles, birds and mammals, for example, abnormalities of the gonads and genitals in the American alligator (Alligator mississippiensis) in Florida (Guillette et al., 1994, 1996), thinning of the eggshell in glaucous-winged gulls (Larus glaucescens) in Seattle (Fry and Toone, 1981) and in white-tailed sea eagles (Haliaeetus albicilla) in the Baltic Sea (Helander et al., 1982), and impaired reproduction in the male Florida panther (Felis concolor coryi) in Florida (Facemire et al., 1995).

When considering EDCs, we tend to focus on industry-manufactured compounds such as polychlorinated biphenyls, dichloro-diphenyl-trichloroethane and other pesticides or pharmaceuticals. However, estrogens in human waste have the most potent effects on fish and quality, and significant prevention and reduction of emissions are difficult because they are derived from human physiological activities. Thus, investigation of long-term effects of estrogens on wildlife using environmentally relevant concentrations is needed, especially for urban rivers, where sewage treatment work discharges are situated.

In this case, exposure experiments in the laboratory using fish are useful for precise verification because ideal conditions that exclude unknown factors of field studies are possible. Accordingly, various testing methods have been developed and standardized by international organizations such as the Organization for Economic Cooperation and Development (OECD) (Harvey et al., 1999; McLachlan et al., 2002; OECD, 1992, 2009, 2013, 2011, 2012a; United States Environmental Protection Agency, 1982). To investigate population levels of adverse effects of chemicals, long-term tests that can evaluate full life-cycle or multigenerational effects using environmentally relevant concentrations are essential (OECD, 2010). Some studies have reported the full life-cycle effects of estrogen or estrogenic substances on various species, such as the effects of 17α-ethinylestradiol (EE2) (Lange et al., 2001; Parrott and Wood, 2002), and bisphenol A (Sohoni et al., 2001) on fathead minnow (Pimephales promelas); 17β-estradiol (E2) (Seki et al., 2005), 4-nonylphenol (Yokota et al., 2001) and 4-tert-pentylphenol (Seki et al., 2003) on Japanese medaka (Oryzias latipes); and EE2 on zebrafish (Danio rerio) (Fenske et al., 2005).

On the other hand, no studies have been found that evaluate the long-term effects of estrone (E1) on fish species. Most of the existing studies are short-term assessments of reproduction or short-term indirect reproductive tests using fathead minnow (Dammann et al., 2011; Panter et al., 1998; Thorpe et al., 2003), sexual development tests using zebrafish (Holbech et al., 2006), early life stage tests using common carp (Cyprinus carpio) and fathead minnow (Tyler et al., 1999), or sexual behavior tests using guppy (Poecilia reticulata). The effects observed on these studies are weaker than E2 and consistent with the theoretical study that estimates the bioactivity of E1 to be about half of that of E2 (Larson, 1999). These are the factors that the environmental risks of E1 have not been considered sufficiently. However, multiple studies have reported that more than 100 ng l–1 of E1 have been detected in urban environments, being approximately two times as high as E2 (Baronti et al., 2000; Kolpin et al., 2002; Rodgers-Gray et al., 2000; Shore et al., 1993; Van den Belt et al., 2004; Wang et al., 2008; Writer et al., 2010; Ying et al., 2002). Furthermore, Bradley et al. (2009) suggested that E1 has a longer half-time than the other estrogens due to its difficulty of biodegradation. Consequently, E1 is not in the least inferior to the other EDCs in its concern of adverse effects to wildlife.

We, therefore, conducted a multigeneration test on fish with a preliminary short-term reproduction assay for E1 using Japanese medaka to investigate long-term adverse effects.

Materials and Methods

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

Experimental Design

Japanese medaka (Oryzias latipes) were exposed to E1 under a scheme consisting of two tiers (Fig. 1), based on the frameworks of EDCs and other chemicals of Japan (Ministry of the Environment Government of Japan, 2010), USA (United States Environmental Protection Agency, 1998) and OECD (OECD, 2012b). Tier 1 is a 4-week short-term reproduction assay based on OECD TG229 (OECD, 2012a), as potential verification of estrogenic activity and range finding for tier 2. Tier 2 is a 27-week multigeneration test based on the draft protocol, being developed through joint research between Japan and the USA, as a definitive test (OECD, 2013).

image

Figure 1. Scheme of the experiments. Tier 1: Short-term reproduction assay. Tier 2: Multi-generation test.

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Animals

We used a strain of the orange–red variety of Japanese medaka (Temminck et al., 1846), which have been maintained for over 10 years at the National Institute for Environmental Studies (Tsukuba, Ibaraki, Japan) under a management scheme that excluded unintentional contamination with chemicals. All fish were maintained in whole stages in dechlorinated tap water at controlled temperatures (25 ± 1 °C) and pH (7.5 ± 0.5). The photoperiod was 16:8 h (light/dark) under visible light generated by a fluorescent lamp generating approximately 800 lux. Fertilized eggs were collected from stocks of mature fish via inbreeding. Embryos (approximately 300 drops per dish after removing sticky threads) were put into a Petri dish (15 cm in diameter) and subjected to dewatering once for 2 days. Hatched larvae were moved to cubic glass tanks (capacity: 5 l) at a density of 15 fish per tank, then cultured using a non-cycling overflow system (renewal rate, one-fifth per day) until start of each test. After absorption of the yolk sac, they were fed freshly hatched desalinated live brine shrimp (Artemia spp.) nauplii (resting eggs were purchased from Aquafauna Bio-Marine Inc., Hawthorne, CA, USA) at approximately 10% of the fish body weight per day, divided into two feedings.

Test Solutions

To prepare 100 mg l–1 stock solution A, 50 mg of estrone (E1, CAS No., 53-16-7; purity, > 98%; purchased from Wako Pure Chemical Industries Inc., Osaka, Japan) was dissolved in 500 ml acetone (CAS no., 67-64-1; purity, > 99.5%; Wako) by hand stirring. To prepare 100 µg l–1 stock solution B, 1 mg l–1 of stock solution A was placed into a 1 l glass medium bottle (body diameter, 101 mm) and air dried; the residue was then dissolved in 1 l of Milli-Q® water via sonication for 30 min in an ultrasonic bath, then stirring for 12 h using a magnetic stirrer with a PTFE stir bar. The stock solution B was diluted with Milli-Q water in hundred-fold concentrations of each nominal treatment, as stock solution C of each treatment. Then each stock solution C was continuously diluted to hundred-fold with dechlorinated tap water to the appropriate nominal concentrations for each treatment using a flow-through exposure system (SIS-1F; manufactured by Shibata Scientific Technology Ltd., Saitama, Japan). Stock solutions B and C were renewed once every 3 days using the same stock solution A, which was kept in a glass reagent bottle in a cold dark place.

Exposure and Observations: Short-Term Reproduction Assay

A short-term reproduction assay was conducted based on the OECD TG229. In brief, three males and three females at 84 days post-fertilization (dpf) were put into a 5 l glass cubssse tank and exposed to each test solution for 28 days via a flow-through exposure system (renewal rate, one-fifth per day). Four replicate tanks were used for each concentration (1000, 320, 100 and 32 ng l–1) as well as four clean-water control tanks.

Eggs produced during a test period were collected from each tank, counted and examined for embryogenesis using a stereomicroscope (MZ16; Leica Microsystems GmbH, Wetzlar, Germany). We calculated the total number of eggs, number of fertilized eggs and fertility (number of fertilized eggs/total number of eggs). Fish in the test tanks were visually examined daily to assess survival, appearance and behavior. In the present study, dead fish were not observed.

At the end of the exposure period, all surviving fish were dissected after ice anesthetization that conformed to our ethical treatment policy, and the following elements were observed or measured. Sex phenotype was identified from the shapes of the dorsal and anal fins. Total and standard lengths (from the tip of the snout to the posterior end of the last vertebra) were measured using a set of digital calipers (CD-S10C; Mitsutoyo Co. Ltd., Kanagawa, Japan). Wet body weight was measured using an electronic balance (AG204; Mettler-Toledo GmbH, Schwerzenbach, Switzerland). The liver was dissected and weighed using an electronic balance, then the hepatosomatic index (HSI; liver mass/body mass × 100) (Heidinger and Crawford, 1977; Htun-Han, 1978) was calculated. The VTG concentration in the liver was analyzed using an enzyme-linked immunosorbent assay (EnBioMedaka VTG ELISA system; EnBioTec Laboratories Co., Ltd, Tokyo, Japan) according to its operations manual (Ohkubo et al., 2003; Sumpter, 1997; Yamanaka et al., 1998). The gonad dissected and sex type (testis/ova/unknown) was identified from its appearance, weighed using an electronic balance and the gonadsomatic index (GSI; gonad mass/body mass × 100) (Nikolsky, 1963) was calculated. The gonads were fixed in modified Davidson's fluid (Creasy and Jonassen, 1999). The anal fin was dissected and fixed in 4% paraformaldehyde phosphate buffer solution (Wako). Thereafter, we counted the number of fin rays joint plates that had papillary processes under a stereomicroscope. For reference, the papillary process of the Japanese medaka appears on a fin ray of the anal fin and is normally characteristic of an adult male (Oka, 1931). The appearance and disappearance of it are controlled by androgens (Nakamura et al., 2013; Oka, 1931; Okada and Yamashita, 1944; Yamamoto 1958), and recently, genes (Lef1, Bmp7) related to development of this process in medaka have been elucidated (Ogino et al., 2013). Thus, OECD TG229 and TG230 specify that the quantity of this process as an optional indicator of secondary sexual characteristics.

Exposure and Observations: Multigeneration Test

Parent generation (P generation)

A pair of male and female medaka at 84 dpf were put into a 2 l glass rectangular solid tank and exposed to each test solution for 28 days via a flow-through exposure system (renewal rate, one-fifth per day). Four replicate tanks were used for each E1 concentration (100, 50.0, 25.0, 12.5 and 6.25 ng l–1) as well as eight clean-water control tanks. The concentrations were set based on the results of the short-term reproduction assay and detected concentrations in the environment.

Eggs produced during a test period were collected from each tank, counted and examined for embryogenesis using a stereomicroscope (MZ16, Leica Microsystems GmbH). We calculated the total number of eggs, number of fertilized eggs and fertility. Fish in the test tanks were visually examined daily to assess survival, appearance and behavior. In the present study, dead fish were not observed.

On the first day of the fourth week, 15 fertilized eggs per tank were used for the F1 generation after counting and examination. At that time, they were selected randomly from all the fertilized eggs from each tank; if fertilized eggs were fewer than 15, shortages were supplemented with eggs produced the next day, which were randomly selected.

At the end of the exposure period, all fish were dissected after ice anesthetization that conformed to our ethical treatment policy, and the following elements were observed or measured according to the same methods used in the short-term reproduction assay; sex phenotype, total and standard lengths, wet body weight, HSI, VTG concentration in liver, gonad type and number of fin ray joint plates that had papillary processes. Gonadal tissue sections were prepared for histopathological observation in control and groups of exposure to 50 and 100 ng l–1 by the following procedure: trunks were fixed by dipping into modified Davidson's fluid without detaching gonads. They were dehydrated and paraffin (Wako) penetrated via vacuum infiltration processor (Tissue-Tek VIP Jr.; Sakura Finetek Japan Co., Ltd., Tokyo, Japan), then embedded via a tissue embedding console (Tissue-Tek TEC Plus). After that, they were sliced into a coronal direction to a thickness of approximately 5 µm via a microtome (Tissue-Tek Feather Trustome), then stained by hematoxylin and eosin (both pigments were purchased from Wako), extended on the glass slide and included as a permanent prepared slide using coverglass and synthetic resin (Softmount; Wako) manually. Gonadal tissues were observed and photographed via a microscope. As gonadal tissue sections were prepared without detaching from the trunk, GSI could therefore not be calculated.

First filial generation

Fifteen fertilized eggs collected from the P generation were returned to each tank for further exposure to each treatment under the same condition of the P generation; at this time, eggs of each tank were separated in a glass cylinder (diameter: 5 cm, height: 10 cm) with a stainless wire mesh covering its bottom (0.25 ϕ, 32 mesh). The cylinder was shaken vertically (amplitude: approximately 5 cm) slowly (cycle: approximately once for 4 s) via the exposure system. Eggs were visually examined daily to assess survival and hatching. The determination of death was based on the absence of a heartbeat. Hatched larvae were moved out from the cylinder.

After the absorption of the yolk sac, larvae were fed freshly hatched desalinated live brine shrimp nauplii at approximately 10% of the fish body weight per day, divided into two feedings. Fish in the test tanks were visually examined daily to assess survival, appearance and behavior. Dead fish were removed as soon as they were observed.

At 10 weeks post-fertilization (wpf), the genotype sex of fish were determined by an identified and sequenced gene (Dmy), which is located on the Y chromosome (Matsuda et al., 2002). All surviving fish were ice anesthetized then tips of the caudal fin of each fish were clipped with a razor blade. DNA from each fin clip is extracted using DNeasy Blood & Tissue Kit (Qiagen Inc., Venlo, the Netherlands) and the presence or absence of Dmy is determined by polymerase chain reaction methods (Matsuda et al., 2002). From the determination, two pairs of genotype males and females per tank were selected randomly for breeding. Each pair was put in a 2 l tank and allowed to continue exposure to each test solution. In summary, eight replicate tanks were used for each concentration (100, 50.0, 25.0, 12.5 and 6.25 ng l–1) as well as 16 clean-water control tanks. At the end of the 10th wpf, all of the other fish (not selected for breeding) were dissected after ice anesthetization that conformed to our ethical treatment policy, and the following elements were observed or measured according to the same methods as for the P generation: sex phenotype, total and standard lengths, wet body weight, HSI, VTG concentration in liver, gonad type and number of fin ray joint plates that had papillary processes. Gonadal tissue sections were also prepared for histopathological observation on control and groups of exposure to 50 and 100 ng l–1 as well as P generation.

From the 12th to 15th wpf, eggs produced by breeding pairs were collected from each tank, counted and examined for embryogenesis using a stereomicroscope to calculate the total number of eggs, number of fertilized eggs and fertility. Fish in the test tanks were visually examined daily to assess survival, appearance and behavior. In the present study, dead fish were not observed.

On the first day of the fourth week, 15 fertilized eggs per tank were used for the F2 generation after counting. At that time, they were selected randomly from all of fertilized eggs of each tank; if fertilized eggs were fewer than 15, shortfalls were supplemented by eggs produced during the next day, which were selected randomly.

At the end of the 15th wpf, all of the other fish (not selected for breeding) were dissected using ice anesthetization that conformed to our ethical treatment policy, and the following elements were observed or measured the same as in methods for the P generation and 10th wpf of F1 generation: sex phenotype, total and standard lengths, wet body weight, HSI, VTG concentration in liver, gonad type and number of fin ray joint plates that had papillary processes. Gonadal tissue sections were also prepared for histopathological observation on controls and groups exposed to 50 and 100 ng l–1 as well as the P and the 10th wpf of the F1 generation.

Second filial generation

Fifteen fertilized eggs collected from the P generation were returned to each tank for continued exposure for each concentration under identical conditions to the P generation; at this time, eggs of each tank were separated in a glass cylinder (diameter: 5 cm, height: 10 cm) with a stainless wire mesh covering its bottom (0.25 ϕ, 32 mesh). The cylinder was shaken vertically (amplitude: approximately 5 cm) and slowly (cycle: approximately once for 4 seconds) by the flow-through exposure system. Eggs were visually examined daily to assess survival and hatching. Determination of death was based on the absence of a heartbeat. Hatched larvae were moved out of the cylinder.

After absorption of the yolk sac, they were fed freshly hatched desalinated live brine shrimp nauplii that were approximately 10% of the fish body weight per day, divided into two feedings. Fish in the test tanks were visually examined daily to assess survival, appearance and behavior. Dead fish were removed as soon as they were observed.

At the end of the 10th wpf, all surviving fish were dissected after ice anesthetization that conformed to our ethical treatment policy, and the following elements were observed or measured according to the same methods used for the P and F1 generation; sex phenotype and genotype, total and standard lengths, wet body weight, HSI, VTG concentration in liver, gonad type and number of fin ray joint plates that had papillary processes. Gonadal tissue sections were also prepared for histopathological observation on controls and groups exposed to 50 and 100 ng l–1 as well as the P and F1 generation.

Water Quality Measurements and Chemical Analysis

During all exposure tests, we measured the water quality for each test solution once a week to confirm that it met each of the test specifications (temperature, 25 ± 2 °C; pH 7.5 ± 0.5; DO, ≥ 60% oxygen saturation) using a thermometer (CT-430WP; manufactured by Custom Ltd., Tokyo, Japan), a pH meter (D-55; manufactured by Horiba Ltd., Kyoto, Japan) and a DO meter (HQ30d; manufactured by HACH Company Inc., CO, USA).

During all exposure tests, each test solution was sampled to verify the concentrations of E1. Each sample was extracted using solid-phase cartridges (Oasis HLB Plus LP Extraction Cartridge; purchased from Waters Co., MA, USA), dissolved in methanol and then analyzed by liquid chromatography/mass spectrometry (LCMS-2010EV; manufactured by Shimadzu Corporation, Kyoto, Japan).

Statistical Analysis

First, homogeneity of variance data were assessed with Bartlet's test (significance level, 5%) (Snedecor and Cochran, 1989). Then, if homogeneity of variances were not rejected, differences between each treatment and a control were assessed by Dunnett's test (Dunnett, 1964), otherwise they were assessed by Steel's test (Steel, 1959). When necessary, data were transformed for normalization and/or to reduce variance heterogeneity. All analyses were performed using R software (version 2.14) (Ihaka and Gentleman, 1996; R Development Core Team, 2004) and its packages mvtnorm (Genz et al., 2009) and EZR (Kanda, 2013). Results were considered significant at *P < 0.05 or **P < 0.01. All data are presented as means (± SD) of the four replicate exposure tanks unless otherwise mentioned.

Results

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

Short-Term Reproduction Assay

Water analyses

The concentration of each test solution is shown in Table 1. Measured concentrations are stable and close to the nominal concentrations for each test. E1 was not detected in the control solution in any of the tests. Results were shown as mean measured values.

Table 1. Nominal and measured concentrations of estrone (E1) for the short-term reproduction assay
Nominal concentrations (ng/l)Measured concentrations (ng/l)
 Week of exposure
 1234Mean (SD)
ControlNDNDNDND-
32.029.421.131.934.629.3 (5.05)
10013797.4113100112 (15.7)
320310202340235272 (55.5)
1000909840111011561000 (133)
Biological endpoints

There were no dead fish or abnormalities in appearance and behavior during E1 exposure in any of the tests. There were no significant concentration-related differences in either male or female in total length, standard length or wet body weight between E1-treated fish and controls at any concentrations (data not shown).

There were no dead fish during the exposure nor any abnormalities in the appearance or behavior. There were also no concentration-related differences in neither male nor female in total length, standard length and wet body weight.

Significant decreases in the number of total and fertilized eggs were observed at 1000 ng l–1; however, no concentration-related effects were observed on both total and fertilized eggs. No effect on fertility was observed in any of the concentrations (Fig. 2).

image

Figure 2. Total number of eggs, number of fertilized eggs and fertility (number of fertilized eggs/total number of eggs) per female fish in the short-term reproduction assay. Bar: ± SD, **P < 0.01 vs. control.

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A significant decrease in male HSI was observed at 1000 ng l–1; however, given the same conditions, no significant effects on female HSI were observed (Supplementary Table S1). Concentration-related increase in male VTG concentration in the liver was observed, significant increase was observed at 112, 272 and 1000 ng l–1. No change was observed in female VTG concentration in the liver at any of the concentrations (Fig. 3).

image

Figure 3. VTG concentrations in livers in the short-term reproduction assay. (A) Male, (B) Female. Bar: ± SD, *P < 0.05, **P < 0.01 vs. control. VTG, vitellogenin.

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No abnormal gonads were observed in any of the concentrations, nor were there any discrepancies between phenotypic sex and gonad type discrimination. A significant decrease in male GSI was observed at 1000 ng l–1; however, no effect on female GSI was observed for any of the concentrations (Supplementary Table S1).

No effect of E1 on the number of male and female papillary processes was observed in any of the concentrations (Supplementary Table S1).

Multigeneration Test

Water analysis

The concentration of each test solution is shown in Table 2. Measured concentrations are stable and close to the nominal concentrations in each test. E1 was not detected in the control solution in any of the tests. Results were shown as mean measured values.

Table 2. Nominal and measured concentrations of E1 for the multigeneration test
Nominal concentrations (ng/l)nAverage of measured concentration (ng/l)Measured value / nominal value (%)Coefficient of variation (%)
Control27N.D.--
6.25275.7491.87.42
12.52711.491.07.78
25.02724.096.27.02
50.02747.194.26.73
1002791.491.45.41
Biological endpoints

A summary of the biological endpoints of the multigeneration test is shown in Supplementary Table S2.

Reproductive adult period of P and F1 generations

There were no dead fish during the reproductive adult period, nor was there any abnormality in appearance and behavior in the P generation. In the F1 generation, there was no dead fish during the reproductive adult period and no abnormalities in appearance and behavior; however, decline or loss of courtship activities were observed in two males among eight pairs at 91.4 ng l–1 E1. No differences in total length, standard length, and wet body weight in the P generation in both sexes (data not shown). On the other hand, in the F1 generation, a significant increase in male total and standard lengths, and wet body weight were observed at 91.4 ng l–1; a significant decrease in female total length was observed at 91.4 ng l–1 (data not shown).

No effects on numbers of total eggs or fertilized eggs were observed in the P generation; in contrast, significant decreases in numbers of total and fertilized eggs were observed at 91.4 ng l–1 in the F1 generation (Fig. 4). No effect on fertility was observed in both the P and F1 generations (Fig. 4).

image

Figure 4. Total number of eggs, number of fertilized eggs and fertility (number of fertilized eggs/total number of eggs) per female fish during the reproductive adult period in the multigenerational test. (A) P generation, (B) F1 generation. Bar: ± SD, *P < 0.05 vs. control.

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No effects on either male or female HSI were observed in both P and F1 generations (Supplementary Table S3). A significant increase in male VTG concentration in the liver was observed at 47.1 and 91.4 ng l–1 in both the P and F1 generations (Fig. 5). Besides, a significant increase in female VTG concentration in the liver was also observed at 91.4 ng l–1 in the F1 generation (Fig. 5).

image

Figure 5. VTG concentrations in livers of reproductive adults (15th week per fertilization) in the multigenerational test. (A) P generation/male. (B) F1 generation/male. (C) P generation/female. (D) F1 generation/female. Bar: ± SD, *P < 0.05, **P < 0.01 vs. control. VTG, vitellogenin.

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No effect on the number of male papillary processes was observed in P and F1 generations, and no female papillary processes were observed in any concentrations in both P and F1 generations (Supplementary Table S3).

No discrepancy was encountered among three sexual indicators, phenotypic sex, gonad type and genotype sex in both P and F1 generations (Table 3). No abnormal gonads were observed in the P generation; however, in the F1 generation, six males with testis–ova were observed among eight males at 91.4 ng l–1 (Fig. 6). No ovarian abnormality was observed at any concentrations in both P and F1 generations.

Table 3. Phenotype sex, gonad type, and genotype sex in reproductive adult of the P and F1 generations in the multigeneration test
GenerationsConcentrations (ng/l)nPhenotypeGonadGenotype
muftuomf
  1. m; male; f, female; t; testis; o, ovary; u, unknown.

PControl16808808--
5.748404404--
11.48404404--
24.08404404--
47.18404404--
91.48404404--
F1Control3216016160161616
5.741680880888
11.41680880888
24.01680880888
47.11680880888
91.41680880888
image

Figure 6. Photomicrographs of typical male and female gonads of reproductive adults in the multigeneration test. (P-A) Control testis of the P generation. (P-B) Male exposed to 91.4 ng l–1 estrone of the P generation. (P-C) Male exposed to 91.4 ng l–1 estrone of the P generation showing myriads of stage I oocytes. (P-D) Control ovary of the P generation. (F1-A) Control testis of the F1 generation. (F1-B) Male exposed to 47.1 ng l–1 estrone of the F1 generation. (F1-C) Male exposed to 91.4 ng l–1 estrone of the F1 generation showing myriads of stage I oocytes (o). (F1-D) Control ovary of the F1 generation. Scale bars: 200 µm.

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From embryo to adult period (10 weeks per fertilization) of F1 and F2 generations

No effects on either hatchability or hatching day of embryos were observed in both F1 and F2 generations (Fig. 7). No effects on survival rate after hatching until 10 wpf were also observed in both F1 and F2 generations (Fig. 7).

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Figure 7. Hatchability, hatching day of embryos and survival rate after hatching until the 10th week per fertilization in the multigenerational test. (A) F1 generation. (B) F2 generation. Bar: ± SD.

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No differences in total length, standard length, nor wet body weight at 10th wpf in both sexes were observed in the F1 generation. On the other hand, in the F2 generation, a significant increase in male total and standard lengths at 10 wpf was observed at 91.4 ng l–1 (data not shown), and no differences in female total length, standard length and wet body weight at 10 wpf were observed.

No effects on male HSI were observed in neither 10 wpf F1 and F2 generations. A significant decrease in female HSI was observed at 47.1 and 91.4 ng l–1 in the 10 wpf F1 generation; however, no effects on female HSI were observed at the 10 wpf F2 generation (Supplementary Table S4). A significant increase in male VTG concentration in the liver was observed at 47.1 and 91.4 ng l–1 in both F1 and F2 generations (Fig. 8). No significant difference in female VTG concentration in the liver was observed in both 10 wpf F1 and F2 generations (Fig. 8).

image

Figure 8. VTG concentrations in livers at the 10th week per fertilization in the multigenerational test. (A) F1 generation/male. (B) F2 generation/male. (C) F1 generation/Female. (D) F2 generation/female. Bar: ± SD, *P < 0.05 vs. control. VTG, vitellogenin.

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A significant decrease in the number of male papillary processes was observed at 91.4 ng l–1 in the 10 wpf F1 generation. On the other hand, no significant effects on the number of male papillary processes were observed on the 10 wpf F2 generation. No female papillary processes were observed in any treatments among both the F1 and F2 generation (Supplementary Table S4).

There are no discrepancies observed among three sexual indicators, phenotype sex, gonad type and genotype sex on both the 10 wpf F1 and F2 generation (Table 4). Two males with testis-ova were observed among 19 males exposed to 91.4 ng l–1 in the 10th wpf F1 generation (Fig. 9). On the other hand, in the 10 wpf F2 generations, no testes–ova were observed with any treatments; however, some abnormalities of the testes such as cavities in the organ or dysfunctions of the seminiferous tubule were observed on exposure to 47.1 and 91.4 ng l–1 (Fig. 9). No abnormal ovaries were observed for any treatments on both the 10 wpf F1 and F2 generation.

Table 4. Phenotype sex, gonad type and genotype sex in the 10th week per fertilization of the F1 and F2 generations in the multigeneration test
GenerationsConcentrations (ng/l)nPhenotypeGonadGenotype
muftuomf
  1. m, male; f, female; t, testis; o, ovary; u, unknown.

P (10 w pf)Control5818634185352236
5.743723113231132314
11.43315117151171518
24.03017211172111812
47.13019110191101911
91.42817381738199
F1 (10 w pf)Control9546247462474649
5.744321121211212122
11.44011121181211822
24.04324415244152617
47.14125016250162516
91.44921523242232524
image

Figure 9. Photomicrographs of typical gonadal tissue preparations of typical males and females at the 10th week per fertilization in the multigeneration test. (F1-A) Control male testis of the F1 generation. (F1-B) Male exposed to 47.1 ng l–1 estrone of the F1 generation. (F1-C) Male exposed to 91.4 ng l–1 estrone of the F1 generation showing multiple stage I and II oocytes (o). (F1-D) Control ovary of the F1 generation. (F2-A) Control testis of the F2 generation. (F2-B) Male exposed to 47.1 ng l–1 estrone of the F2 generation showing cavities in the organ (c). (F2-C) Male exposed to 91.4 ng l–1 estrone of the F2 generation showing cavities in the organ (c). (F2-D) Control ovary of the F2 generation. Scale bars: 200 µm.

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Discussion

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

Panter et al. (1998) and Dammann et al. (2011) reported significant increases in male fathead minnow VTG concentrations after approximately 3 weeks’ exposure to 31.8 ng l–1 (Panter et al., 1998) or 50 ng l–1 (Dammann et al., 2011) E1. Thorpe et al. (2003) reported a significant decrease in the number of spawned eggs of fathead minnow after approximately 3 weeks’ exposure to 781 ng l–1 E1. The results of the short-term assay in the present study were in good agreement with these previous studies. We also confirmed consistency between results of the short-term assay and the P generation of a multigenerational test. As a result of long-term investigations, Seki et al. (2005) reported a significant decrease in Japanese medaka reproductivity after full life exposure to 27.9 ng l–1 of 17β-estradiol. This was approximately three times as high as the lowest observed effect concentration (91.4 ng l–1) of the F1 generation of the multigenerational test in this study.

When compared to environmental concentrations, the lowest observed effect concentration (91.4 ng l–1) of reproductive effects of E1 in the current multigeneration test is lower or nearly equal to the reported values, such as 134 ng l–1 (Tokyo Metropolitan Government Bureau of Sewerage, 2005), 112 ng l–1 (Kolpin et al., 2002), or 82.1 ng l–1 (Baronti et al., 2000). Based on these values, the predicted environmental concentration/predicted no effect concentration ratio in E1 is 2.85, even if an assessment factor is excluded (Fig. 10). This indicates there is a possibility that E1 exerts actual harm to fish species if these concentrations continue in the long term.

image

Figure 10. Comparison of environmental estrone concentrations and results of the multigeneration test. Reported highest value in the UK (Baronti et al., 2000). Reported highest value in the USA (Kolpin et al., 2002). Reported highest value in Japan (Tokyo Metropolitan Government, 2005).

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Figures 11 and 12 show detailed results for the reproductive F1 generation exposed to 91.4 ng l–1. Abnormalities in histopathology of the testes and decline or loss of courtship activities coincided with a decline in reproduction. The number of fin ray joint plates that have papillary processes also tended to be consistent with this. However, testes–ova were observed in six males, and no significant decline was observed in reproductivity for the five of them. Despite the fact that the testis–ova have been regarded as the key evidence for endocrine disruption (Gray et al., 1999; Gray and Metcalfe, 1997), this result shows that the appearance of testis–ova is not closely related to actual effects on reproduction. From the above results, we should consider the appearance of a testis–ova as an indication of endocrine disruption; however, qualitative consideration of abnormalities in histopathological observations is needed for appropriate evaluation of actual harm on fish reproduction.

image

Figure 11. Comparison of eight pairs of the F1 generation exposed to 91.4 ng l–1. Bar: ± SD. VTG, vitellogenin.

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image

Figure 12. Testes of eight males exposed to 91.4 ng l–1 estrone in the F1 generation. (No.1) Showing the dysfunction of the seminiferous tubule (d) with myriads of stage I oocytes (o), (No.2-5) Showing stage I oocytes (o), (No.6) Showing the process of normal spermatogenesis is observed with no oocytes and the other abnormalities, (No.7) Showing the dysfunction of the seminiferous tubule accompanied by significant dilatation (d), however no oocytes were observed, (No.8) Showing the dysfunction of the seminiferous tubule accompanied by dilatation (d) with stage I oocytes (o). Scale bars: 200 µm.

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Comparison of E1 effects between generations indicates that stronger effects were clearly observed in the F1 generation than the P generation, although we cannot determine the responsible elements, whether difference in exposure time, generational transitions or other reasons. Comparing the data in the 10th wpf of F1 and F2 generations, some effects were observed in only either one generation, when no testis–ova was observed, unlike F1, in which some abnormalities of the testis such as cavities in the organ or dysfunctions of the seminiferous tubule were observed at 47.1 and 91.4 ng l–1. From the viewpoint of consistency with the histopathological abnormalities and the decline in reproduction in the reproductive F1 generation, there is a possibility that adverse effects were observed even on exposure to 47.1 ng l–1.

Conclusion

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

In conclusion, E1 exposure at environmentally relevant concentrations did not exert adverse effects on Japanese medaka in the short-term test (one generation); however, adverse effects, such as decline in reproduction, were observed in the second generation. Furthermore, there is a possibility that there were severe effects in the third or later generations. Despite testis–ova being regarded as the key evidence for endocrine disruption, this study showed that the appearance of testis–ova is not always consistent with actual adverse effects on reproduction. Accordingly, we believe that qualitative consideration of abnormalities and histopathological observations are needed for appropriate evaluation of endocrine disruption. This method should be utilized to investigate multigenerational effects of chemicals, not only EDCs but also any other substances, on wildlife.

Acknowledgments

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information

We wish to thank Ms. Kazuyo Saito for her immense contribution throughout the experiment, Ms. Hiroko Takahashi for her help in breeding and culturing of medaka, and Mr. Masaaki Koshio for his technical assistance in the measurement of viellogenin concentration and determination of genotypic sex, at the National Institute for Environmental Studies, Japan. We would like to thank Ms. Eri Kamahara (Nagasaki University, Japan) for her help in the experiment, Mr. Teruhiko Kusano, Dr. Yoshihiro Kagami, and Ms. Kimiko Obuchi (Ecogenomics, Inc., Japan) for their technical assistance of histopathological observation. We would also like to thank Mr. Gregory King (The University of Tokyo, Japan) for the English language review. This study was supported by Ministry of the Environment, Government of Japan.

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  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information
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Supporting Information

  1. Top of page
  2. ABSTRACT
  3. Introduction
  4. Materials and Methods
  5. Results
  6. Discussion
  7. Conclusion
  8. Acknowledgments
  9. Conflict of Interest
  10. References
  11. Supporting Information
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