a freshwater fish biomonitor for edcs in southern victoria - … · a freshwater fish biomonitor for edcs in southern victoria Ana F. Miranda1,2,*, Vincent Pettigrove2, Dayanthi Nugegoda1,2

89

AustrAlAsiAn JournAl of Ecotoxicology Vol. 16, pp. 89-101, 2010

miranda et alfreshwater fish biomonitor for edcs

a freshwater fish biomonitor for edcs in southern victoria

Ana F. Miranda1,2,*, Vincent Pettigrove2, Dayanthi Nugegoda1,2

1School of Applied Sciences, RMIT University, PO Box 71, Bundoora, Victoria 3083, Australia.2Victorian Centre for Aquatic Pollution Identification and Management, Department of Zoology, The University of Melbourne, Parkville, VIC, 3052, Australia.

Manuscript received, 26/5/2011; accepted, 28/3/2012.

abstractRecent biomonitoring programs throughout Australia have revealed detectable release of steroidal hormones into waterways through Waste Water Treatment Plants and feedlot effluents. However, despite recent advances, it is difficult to determine whether the concentrations of these hormones are sufficient to produce effects on freshwater fish populations, and their potential effects on freshwater indigenous fish are still largely unknown. Further laboratory and field studies are needed to assess the impact of these chemicals in Australian aquatic ecosystems and especially on native fish. The identification of appropriate species, which are likely to be useful indicators of endocrine disruption in field conditions, is one of the key stages in a research strategy for investigating the ecological effects of endocrine disruption. This review aims to evaluate the potential of indigenous freshwater fish species of Southern Victoria for biomonitoring endocrine disruptive effects in the laboratory and in the field. We reviewed the characteristics of indigenous freshwater fish species found in the Southern Victoria catchments, and their life history patterns, as selection criteria for a “biomonitor’. Nineteen fish species were considered, and three species were identified as potential biomonitors that may justify further investigation for the development of toxicity tests and as biomarkers of endocrine disruption. These three species are the Australian smelt (Retropinna semoni), flathead gudgeon (Philypnodon grandiceps) and the southern pygmy perch (Nannoperca australis).

Key words: Endocrine disruption; biomonitor; Southern Victoria; Australia; indigenous fish.

introductionOver the past two decades, there has been substantial and increasing evidence that endocrine disruption is impacting wildlife adversely on a global scale (Colbourn et al. 1993; Tyler et al. 1998; Taylor and Harrison 1999; Vos et al. 2000). According to Goodhead and Tyler (2009) endocrine disruption can be defined as a hormonal imbalance initiated by exposure to a pollutant and leading to alterations in development, growth, and/or reproduction in an organism or its progeny. Chemicals that are termed Endocrine Disrupting Chemicals (EDCs) are wide ranging and include natural and synthetic steroids, pesticides and some personal care products and pharmaceuticals. The induced effects range from disrupted embryonic development (Rasmussen et al. 2002), complete sex reversal (Örn et al. 2003), feminisation of males, impaired reproduction (Jobling et al. 2002; Jobling and Tyler 2006). In worst-case scenarios, endocrine disruption has led to near extinction, as demonstrated by a study where fathead minnows (Pimephales promelas) were exposed to 5-6 ng L-1 in a seven-year whole lake experiment in Canada (Kidd et al. 2007). As surface waters are the main sink for EDCs and their metabolites, aquatic species are especially endangered, mainly due their higher exposure rates (Kloas et al. 2009).

Numerous biomonitoring programs have been performed in Australia and EDCs have been detected at concentrations comparable to those found in other developed countries (Leusch et al. 2006; Williams et al. 2007; Chapman et al. 2008; Allinson et al. 2010). However, limited information exists about the relative contribution of different potential sources of EDCs to the aquatic environment (e.g. pesticide run-off, animal farming operations, urban stormwater, and

*Author for correspondence, email: [email protected]

industrial inputs). It is also unknown if Australian fish populations are affected by EDCs, and whether toxicity values derived from the northern hemisphere are adequate for determining protective guidelines, considering the unique variations in climate and sunlight intensity, as well as potential differences in the response of Australian organisms to EDCs. The lack of knowledge about the effects of EDCs on indigenous fish species and the unique proprieties of the Australian environment could lead to an underestimation or overestimation of the potential effects of these chemicals in the indigenous fish populations (Williams et al. 2007).

In temperate Australia, the freshwater alien fish species Gambusia holbrooki (mosquitofish) has been used across different states as a biomonitor of endocrine disruption (Batty and Lim 1999). However, there are few data regarding the effects of EDCs on indigenous fish species, although the indigenous Murray River rainbowfish, Melanotaenia fluviatilis, has been proposed as model test species to investigate the effects of EDCs and several biomarkers have been developed (Pollino et al. 2007; Woods and Kumar 2011). This species can be found across four states throughout the Murray-Darling basin. This species can be found across four states throughout the Murray-Darling basin. However, it cannot be found in any urban catchments near the biggest populated areas where treated domestic sewage, waste from industries using high quantities of surfactants, and wastewater from hospitals, can all contribute to high levels of EDCs (Williams et al. 2007). When considering an organism as an environmental indicator, several criteria are often defined. In the case of estrogenic and androgenic effects from endocrine disruption, effects on sexual development of both males and

90

AustrAlAsiAn JournAl of Ecotoxicology

females and altered sex ratios are expected (vom Saal et al. 2008). Therefore, sexually dimorphic species should be considered and sexually selected traits might provide useful biomarkers to assess the risk of environmental contamination from EDCs (Jašarevic et al. 2011).

The aim of this review is to propose a potential fish biomonitor to assess reproductive endocrine disruptive effects in Victorian freshwaters. The use of an indigenous fish species to biomonitor the possible effects of EDCs will help to overcome the current lack of information on the effects of EDCs on indigenous fish species as well as to contribute to our knowledge on the selected indigenous fish, helping thereby to protect them in the near future. In this review we only pre-selected the indigenous fish species that can be found in the intended area of biomonitoring: catchments in Southern Victoria.

edcs in australiaRecently, several investigators have attempted to address the lack of knowledge about the presence, effects and behaviour of EDCs in the Australian environment (Sarmah et al. 2006; Williams et al. 2007; Allinson et al. 2008; Mispagel et al. 2009). These studies provide increased knowledge on the concentrations of EDCs in Australian freshwaters, their effects on wildlife and their environmental fate, as well as lead to the development of model species and biological markers. Recent freshwater biomonitoring programs have been established nationwide to determine concentrations of EDCs that can be of concern. These studies targeted riverine areas that are near potential EDC inputs such as municipal wastewater treatment plants (WWTPs), paper mills, runoffs from pesticide- or manure-laden farmlands and intensive livestock raising (Williams et al. 2007; Allinson et al. 2008; Allinson et al. 2010).

Endocrine disruptive effects have been described in both freshwater and marine environments. In the marine environment, the release of organotin antifouling agents causing imposex in snails is a classic example of endocrine disruption (Nias et al. 1993; Reitsema and Spickett 1999). In a survey along the New South Wales (NSW) coast of Australia, Gibson and Wilson (2003) found imposex in Thais orbita despite the fact that TBT had been partially banned from antifouling pints for ten years. In the freshwater environment, endocrine disruptive effects on the freshwater alien fish species Gambusia holbrooki (mosquitofish) have been reported across different Australian states (Batty and Lim 1999; Leusch et al. 2006).

In 1998, the Australian Government through NICNAS, the National Industrial Chemicals Notification and Assessment Scheme, published an information paper on EDCs stating the Australian position on endocrine disrupters. This document highlighted the necessity to identify suitable sentinel species to assess possible hazards caused by EDCs (NICNAS, 1998). On 23 November 2007 a workshop attended by researchers, policy makers, regulators, water suppliers and research investors was organised to discuss the matter of Endocrine Disrupting Chemicals (EDCs) in Australian waters which led

to a statement referred as the “Black Mountain Declaration” (LWA 2007).

In Australia as well as in most developed countries, regulatory authorities have established requirements for chemical analysis and methods of testing EDCs (WHO 2001). However, in contrast to other classes of chemical pollutants such as pesticides, heavy metals or PCBs where often guidelines exist, there is a lack of regulation for EDCS such as hormones, drugs and personal care products. This is partly because the methods of analysis are less developed, but also because there are insufficient data on ecotoxicological effects available in the public domain (Falconer et al. 2006; Williams et al. 2007).

monitoring edcs in australian freshwatersMonitoring hormone levels in Australian waterways has become a recent priority (Williams et al. 2007). For this purpose, several sampling methodologies were developed and areas surrounding potential EDC inputs were monitored (Leusch 2004; Williams et al. 2007).

Freshwater waterways receiving treated sewage from wastewater treatment plants, cattle feedlots, dairies, piggeries, agricultural lands and industrial sites are being monitored by diverse groups across the different states (Williams et al. 2007; Allinson et al. 2010; Rose et al. 2009). In WWTP effluents, E2 (17β-estradiol) and EE2 (17α-ethinylestradiol) are generally considered the major contributors to the estrogenic activity often present (Allinson et al. 2010). Thus the levels of E2 and EE2 are in general surveyed and used for comparison among sites and possible effects.

According to a study performed by Williams et al. (2007) in the Australian Capital Territory and South Australia in WWTPs effluents, levels of E1 (estrone) ranged up to 39.3 ng L-1, whereas levels of E2 and EE2 were up to 6.3 ng L-1 and 1.3 ng L-1 respectively. Similarly, in Queensland, E1 levels up to 32 ng L-1, E2 up to 6.4 ng L-1 and EE2 up to 1.2 ng L-1 were detected (Ying et al. 2009). In southern Victoria a survey conducted in 2004 on 12 Victorian WWTPs found that estradiol concentrations ranged from 1.3 to 18 ng L-1 (Mispagel et al. 2009). Recently Allinson et al. (2010) showed that in Victorian effluents, estrogen concentrations varied slightly with season, with E1 and E2 concentrations ranging from non detectable up to 16.8 and 12.4 ng L-1 respectively in winter, and 18.4 and 18.5 ng L-1 respectively in summer. EE2 concentrations ranged from non-detectable up to 0.6 ng L-1 and 0.4 ng L-1 in winter and summer respectively. Currently, the Water Quality Guidelines in British Columbia for 17β-ethinylestradiol (EE2) recommend that, for the protection of freshwater aquatic life, the 30-day average concentration of EE2 in water should not exceed 0.5 ng L-1 (Nagpal and Meays 2009). The recommended guideline is derived from a LOEC of 1.0 ng L-1 of EE2 for reproduction and egg production (Parrot and Blunt 2005; Thorpe et al. 2003).

Allinson et al. (2010) reported estrogenic activity in Victorian WWTP effluents to be up to 73.0 and 20.0 ng L-1 estrogenic equivalents (EEQ) in winter and summer respectively. The

Vol. 16, pp. 89-101, 2010


91


reported estrogenic activity in Victorian WWTPs effluents is similar to that measured by Tan et al. (2007) in Queensland, where activity ranged up to 67.8 ng L-1 EEQ. Victorian WWTPs effluents registered the highest estrogenic activity (EEQ) effluents compared to other states (Allinson et al. 2010).

The highest recorded estrogenic activity in Australian WWTPs is comparable to those in New Zealand, Canada, Germany and Japan, where estrogenic activity could be as high as 50 ng L-1 EEQ in New Zealand and 96 ng L-1 EEQ in Canada, up to 66 ng L-1 in Germany and 49 ng L-1 in Japan (Bandelj et al. 2006; Servos et al. 2005; Pawlowski et al. 2004; Hashimoto et al. 2007). Given that estrogenic levels within Australian freshwaters are comparable to other countries levels where effects of EDCs on fish have been recorded, Australian fish species may already be affected (Jobling 2007). According to Seki et al. (2004), concentrations as low as 8.66 ng L-1 of 17β-estradiol can have significant effects on the reproductive potential of Japanese medaka. Moreover, Woods and Kumar (2011) reported that 5 ng L-1 of EE2 is enough to significantly induce the levels of vitellogenin mRNA in male Australian indigenous fish Murray River rainbowfish. Additionally, according to Jobling (2007), the EDC levels reported by Williams et al. (2007) are high enough to have already affected species living in Australian waterways Australian indigenous fish species may be even more vulnerable, in particular due to the characteristics of Australian freshwater systems where low stream flows and declining freshwater resources are common due to persistent droughts. In addition, in Victoria, discharges from Victoria’s WWTPs can represent most of, and sometimes all of, the flows discharged into enclosed water bodies such as lakes and terminal wetlands (Allinson et al. 2010).

identification of a biomonitor for edcsThe identification of a suitable species that is likely to be a useful indicator of endocrine disruption in field conditions should be the first stage of any research strategy for identifying possible endocrine disruptive effects on wildlife populations (Taylor et al. 1999). A biomonitor species or biomonitor is commonly defined as a species that is sensitive and shows measurable responses to changes in the environment, such as changes in pollution levels (Barbour et al. 1999). Such a biomonitor is particularly important since humans and wildlife are linked by a common food chain, common environments and common physiological and molecular responses to toxicants. Any alteration in the morphology, physiology or behaviour in wildlife should be considered as an indication that there may be a threat to humans and wildlife (Hamling and Guillette 2010). Although data obtained from sentinel species are not expected to be the sole determinative factor in assessing human health concerns, such data can be particularly useful as additional weight-of-evidence in risk assessment, and may highlight suspected environmental contaminants (Hamling and Guillette 2010).

fish as ideal biomonitorsThere has been substantial research attention devoted to the use of fish to assess biological and biochemical responses to environmental contaminants. The main reasons are: firstly that fish live within aquatic ecosystems throughout their lifespan and are sensitive to many types of pollutants; and secondly, they are practical to work with as they are abundant, ever-present in the aquatic environment (freshwater, brackish and marine), and are easy to identify and collect in the field (Barbour et al. 1999). In addition the higher ventilation rate of fish compared with humans may increase their exposure to waterborne contaminants via the surfaces of the gills (Van Der Kraak et al. 2001). Moreover, other features of fish gills (countercurrent system of blood and water flow, thin epithelial membranes, and a high surface area) may increase the uptake of contaminants from the environment to the bloodstream (Damstra et al. 2002).

Compared to other groups of organisms such as invertebrates, fish provide the advantages of being less affected by natural microhabitat differences due to their larger ranges and longer life spans (2-10+ years) reflecting both long-term and current water resource quality (Barbour et al.1999).

Finally, fish are a vital link in the food chain, functioning as a carrier of energy from lower to higher trophic levels (Beyer et al. 1996). Fish, therefore, may be continually exposed to chemicals by diverse routes for extensive periods of time, and may also bioaccumulate and biomagnify toxins from their environment to extremely high levels (Geyer et al. 2001). Despite their limitations, such as their relatively high mobility, fish are considered the most feasible organism as biomonitors.

an indigenous fish biomonitor of endocrine disruPtion in australian freshwatersThe necessity to identify the current impacts of EDCs on indigenous Australian wildlife was defined as one of the priorities in the Black Mountain Declaration on Endocrine Disrupting Chemicals in Australian Waters 2007 (LWA 2007). Moreover, in the freshwater environment, the lack of knowledge about the effects of EDCs on indigenous fish species and the need for indigenous fish biomonitors has been highlighted over the last few years (Williams et al. 2007). There are few data relating to the effects of EDCs in native fish species, although the native Murray River rainbowfish (Melanotaenia fluviatilis) has been proposed as a model test species to investigate the effects of EDCs. This species can be found across four states, throughout the inland Murray-Darling system, where they are potentially affected by several run-offs (Pollino et al. 2007). However, to date in temperate Australian freshwaters, Gambusia holbrooki (mosquitofish) has been used as a routine biomonitor of endocrine disruption (Game et al. 2006). The strong sexual dimorphism of this species, evidenced by the presence of an elongated anal fin- the gonopodium - in males, in parallel with its wide distribution and abundance, has made this species Australia’s favourite fish biomonitor, throughout mainland Australia.

Vol. 16, pp. 89-101, 2010


92


Common name Scientific name Family Status Life cycle/

age to maturity

1. Pouched lampreys Geotria australis Geotriidae Common Long 2. Short headed

lampreys Mordacia mordax Mordaciidae Common Long

3. Short finned eel Anguilla australis Anguillidae Common Long

4. Marbled eel/ Long finned eel

Anguilla reinhardtii Anguillidae Common Long

5. Climbing galaxias Galaxias brevipinnis Galaxiidae Common Short

6. Mountain galaxias Galaxias olidus Galaxiidae Common Short

7. Eastern little galaxias/ dwarf galaxias

Galaxias pusilla Galaxiidae Potentially threatened according to the Australian Society for Fish

Biology

Short

8. Common jollytail Galaxias maculatus Galaxiidae Common 1 year

9. Spotted galaxias Galaxias truttaceus Galaxiidae Rare Short

10. Australian mudfish

Neochanna cleaveri Galaxiidae Vulnerable under the International Union for Conservation of Nature

(IUCN), FFG listed

Short

11. Trout cod Maccullochella macquariensis

Percichthyidae IUCN: Endangered Long

12. Murray cod Maccullochella peelii Percichthyidae Vulnerable, listed under the Victorian Flora & Fauna Guarantee

Act Threatened, Species of National

Significance, listed under the Commonwealth Environmental

Protection & Biodiversity Conservation (EPBC) Act

4-6 years till sexual maturity

13. River blackfish Gadopsis marmoratus Percichthyidae DSE advisory list: critically endangered

Short

14. Southern pygmy perch

Nannoperca australis Nannopercidae Common Short

15. Yarra pygmy perch

Nannoperca obscura Nannopercidae Red List of the World Conservation Union (UCN)

1 year

16. Cox’s gudgeon Gobiomorphus coxii Eleotridae DSE Advisory List: Endangered. Short

17. Flathead gudgeon Philypnodon grandiceps

Eleotridae Common 1 year

18. Australian smelt Retropinna semoni Retropinnidae Common 1 year

19. Freshwater herring

Potamalosa richmondia

Clupeidae Endangered Short

IUCN: International Union for Conservation of Nature; FFG: Victorian Flora and Fauna Guarantee Act 1988; DSE: Department of Sustainability and Environment (Victoria, Australia); EPBC: Environmental Protection & Biodiversity Conservation; UCN: World Conservation Union.

Table 1. List of candidate species and summary of two of the relevant attributes for use to determine their suitability as biomonitors and for development of toxicity tests and biomarkers of endocrine disruption. Species status and life cycle/age to maturity of each species is presented. Species in grey rows are endangered species and/or species with protected status.

Vol. 16, pp. 89-101, 2010


93


Interest in the Murray River rainbowfish as a biomonitor of endocrine disruption has recently been stimulated by the development of new sets of biological markers (Woods 2007; Woods and Kumar 2011, Shanthanagouda et al. 2012). Unfortunately, the distribution of the Murray River rainbowfish is limited to the Murray-Darling basin. It is not found in southern Victorian freshwaters where an alternative species is required due the lack of knowledge of EDCs effects on indigenous fish species inhabiting these freshwaters.

selection of an indigenous fish in southern victoriaIn order to select one or more fish species which are likely to be useful indicators of endocrine disruption in field conditions in southern Victoria, a set of indicative criteria was compiled, based on criteria listed in the literature (e.g., Taylor et al. 1999; Katsiadaki et al. 2007; Segner 2009). Nineteen Australian indigenous fish species from Southern Victoria were considered (Table 1), and their characteristics were evaluated against these criteria, as discussed below.

Many of the criteria are applicable to the selection of any general biomonitor, but since this species was to be used to detect reproductive endocrine disruption, particular emphasis was placed on sexual dimorphism, because secondary sexual characters can be used as endocrine biomarkers. The list of criteria is presented below, in order of importance. (Note that not all the criteria will have to be fulfilled in order for the species to be a potential biomonitor).

• Indigenousfishfromtheintendedareatobemonitored are preferred• Mustnotbelistedasthreatened• Common• Widespread• Likelytoreceivesignificantexposure• Sedentaryorterritorialorhavealocalhomerange• Smallsizeatmaturity• Ideally,rapidgenerationtimes• Sexualreproductionandpreferablysexualdimorphism• Ecologicallyimportant,preferablyatoppredator• Chosenspeciestoincludearangeoflifestylesand feeding habits• Biologywellunderstood• Sensitivetothecontamination• Experimentallyamenableandreadilycultured

It is preferable that the species is widespread within a large as it can be monitored throughout that range. Another important point is the species’ “home range” since the presence of migratory behaviour in some species makes it difficult to determine where and when the exposure to a chemical occurred. Consequently it is advisable to select a biomonitor that has a home range, is non-migratory and if possible territorial (Frame and Dickerson 2006). The time to sexual maturity should be short, reducing the time needed for multigenerational tests. Generally, species with short life cycles are also fish of small size. Small size at maturity will reduce maintenance costs and makes culturing and breeding easy (Dietrish et al. 2008). Short-lived species can be used for assessment of acute and sub-chronic effects of

contaminant exposure allowing a whole life cycle exposure and rapid effect observation. Asynchronous spawning, that is, spawning taking place repeatedly over an extended period, and the relatively large size of the brood are advantages in the selection process. A species being sexually dimorphic (i.e. male and female express secondary sexual characteristics that allow their differentiation) is one of the most important criteria when considering a possible biomonitor to assess endocrine disruptive effects. As male and female are different, the effects of EDC on sex ratios and the expression of male and/or female secondary sexual characteristics can be easily assessed and quantified as biomarkers of endocrine disruption. The selected species should be sensitive to the contaminant or contaminants of interest, but not necessary exquisitely sensitive. Moreover it would be an additional advantage if the selected species is at the top of the trophic chain as it could bio-concentrate and/or biomagnify the concentration of the contaminants. Finally, fish already cultured in captivity are an additional advantage to researchers, since more details of the fish biology are known and acclimatisation periods will be shorter. In order to analyse each species as potential candidates the details of the life cycle patterns of each species were categorised and analysed. Advantages and disadvantages of using one of the listed specific species are described below.

area intended to be monitored using a indigenous fish in southern victoriaSouthern Victorian freshwaters include five catchments: (1) East Gippsland (EG), (2) West Gippsland (WG), (3) Port Phillip and Western Port catchment (PP), (4) Corangamite(C) and (5) Glenelg Hopkins (GH). Each catchment includes three or more river basins as shown in Table 2.

species selectionGenerally, assessing endocrine disruptive effects involves sampling of the species. As sampling is often destructive, using an endangered species will not provide statistically-valid results. Applying this criterion, all the species endangered and/or with protected status were removed from the initial list as potential biomonitors as per Figure 1, criterion ‘B’. After removing all the species that are endangered and/or with protected status, the list of possible species to choose from was reduced to 9 species out of 19 (Figure 1).

Next, all species with a long life cycle and/or large size, (Figure 1, criterion C), were also excluded from the list, not only because time limitations are paramount in biomonitoring but also because it is difficult to maintain large fish under laboratory. Thus, pouched lampreys, short headed lampreys, freshwater eels and marble eels were also excluded (Figure 1).

The list was thus reduced to six species that includes three fish from the family Galaxiidae and one fish each from the Nannopercidae, Eleotridae and Retropinnidae (Figure 1). The family Galaxiidae is the most abundant family occurring in the fresh waters of southern Australia, and G. maculatus has possibly the widest natural distribution of any freshwater fish in the world. However, the three Galaxiids are migratory and not sexually dimorphic. Considering that it is advisable to select a biomonitor that has a home range is non-migratory and if possible territorial, these species were removed from the list (Figure 1).

Vol. 16, pp. 89-101, 2010


94


1. G. australis 10. N. cleaveri 2. M. mordax 11. M. macquariensis 3. A. australis 12. M. peelii 4. A. reinhardtii 13. G. marmoratus 5. G. brevipinnis 14. N. australis 6. G. olidus 15. N. obscura 7. G. pusilla 16. G. coxii 8. G. maculatus 17. P. grandiceps 9. G. truttaceus 18. R. semoni 19. P. richmondia

1. G. australis 6. G. olidus 2. M. mordax 8. G. maculatus 3. A. australis 14. N. australis 4. A. reinhardtii 17. P. grandiceps 5. G. brevipinnis 18. R. semoni

5. G. brevipinnis 6. G. olidus 8. G. maculatus 14. N. australis 17. P. grandiceps 18. R. semoni

14. N. australis 17. P. grandiceps 18. R. semoni

A. Native freshwater species in south eastern Victoria

B. Endangered species or those with protective status removed

C. Long-lived species/those that take many years to mature removed

D. Migratory species and those without sexual dimorphism removed

Figure 1. Scheme of the selection process used to reduce the number of candidate species from the 19 initially-considered species. Each species is represented by a specific number. Criteria A, B, C and D (on the left) were used to discriminate and select the final candidate species numbered 14, 17 and 18.

Vol. 16, pp. 89-101, 2010


95


This leaves three freshwater fish species: flathead gudgeon, Australian smelt and southern pygmy perch that are territorial and sexual dimorphic and thus potential ideal candidates for biomonitoring EDCs with reproductive effects. Their detailed characteristics are described below and summarised in the Table 3.

flathead gudgeonThis species is variable in colour and measures about 8 cm. It can be found in several habitats, from lakes and reservoirs to brackish waters from Queensland to the Murray River, South Australia occurring also in northern Tasmania (Pusey et al. 2003; McDowall 1996).

They are a territorial and show sexual dimorphism making it easy to distinguish males from females Male are generally darker, have a larger mouth, a more bulbous head, a broader interorbital space as well as larger pelvic fins than female fish. In male fish, the mouth generally reaches below the pupil, while in female it ends under the front of the eye. This species display an elaborate courtship display prior to spawning, and also shows parental care. Immediately after spawning, the males take care of the elongate eggs, fanning them with the pectoral fins until they hatch and showing aggression against possible intruders (Pusey et al. 2003; McDowall 1996).

southern pygmy perchThe southern pygmy perches belong to the family Nannopercidae and can be found in lakes, shallow wetlands, small creeks and large rivers such as the Murray throughout southern Australia, including Tasmania (Humphries 1995). This species is still common in Victoria, however it is rare in other Australian states where it has also been recorded (Pusey et al. 2003). This small fish is commonly found at sizes that range from 3 to 5 cm, and are territorial and sexually dimorphic. According to Humphries (1995) the sexes can be distinguished on the basis of the macroscopic appearance of the gonads when fish are more than 25 mm in length. Males and females differ in the region of the vent: males are heavily pigmented and possess a minute urinogenital papilla, while females have only a slightly pigmented vent and possess quite large urinogenital papilla (Llewellyn 1974). An additional advantage of using this species would be that it has multiple spawning events despite their fecundity being considered very low (Pusey et al. 2003). Sex ratio can be used as an indicator of health in wildlife populations and according to Humphries (1995) in wild caught fish from the Macquarie River, Tasmania, the sex ratios in each month did not differ significantly from 1 : 1 (P > 0.05).

australian smeltThe Australian smelt is a small fish (commonly from four to six cm in length), and is common, widespread and territorial. The species is sexually dimorphic: nuptial tubercles are present on the body and head of both sexes, but in males they are larger than in females, and also occur on pectoral and pelvic fins. During spawning events, Australian smelt males often present a bright orange red colour (McDowall 1996).

conclusionsThe three listed species offer great potential for the assessment of reproductive disturbances caused by estrogenic and androgenic xenobiotics. Nonetheless there are also disadvantages, such as the lack of baseline studies and the lack of commercial availability of the fish.

Among the selected species the southern pygmy perch (Nannoperca australis), a laterally compressed fish with length less than 6.5 cm, and the Australian smelt (Retropinna semoni) measuring commonly from 5 to 6 cm, are slightly bigger than the common laboratory species the zebrafish (Danio rerio) that usually measures from 4-5 cm (Allen et al. 2002; Segner 2009). Both southern pygmy perch and Australian smelt present all the advantages of using small fishes under laboratory conditions. However, it is difficult to collect more than a couple of microliters of blood from individual fish for analysis of sex steroids or vitellogenin. This will limit the information that can be obtained from a single specimen for small fishes like southern pygmy perch (N. australis) and Australian smelt (R. semoni), like other routinely used small fish models such as zebrafish (Danio rerio), Japanese medaka (Oryzias latipes) fathead minnow (P. promelas) or even the Australian Murray River rainbowfish (Melanotaenia fluviatilis). Therefore, in these cases, it is more practical to use whole body and liver homogenisation for vitellogenin analysis.

On the other hand, the flathead gudgeon is of body size 8 to 10 cm. This species could yield more blood, even though no attempts to collect blood have been made to date and the volume of blood that can be obtained from this species still unknown. Non destructive analysis of mucus has been employed to perform vitellogenin analysis in some species (Moncaut et al. 2003) and it is an alternative to the destructive methods that are generally used. This method may be used in the flathead gudgeon since it produces significant amounts of mucus.

Table 2. River basins in south eastern Victorian catchments: Glen Hopkins (GH), Corangamite(C), Port Philip (PP), West Gippsland (WG) and East Gippsland (EG).

Catchment GH C PP WG EG

River

Hopkins Moorabool Bunyip Thompson East Gippsland Portland Barwon Yarra La Trobe Snowy Glenelg Corangamite Maribyrnong South Gippsland Tambo Otway Werribee Mitchell

Vol. 16, pp. 89-101, 2010


96


Table 3. Comparison of the biology and life parameters of the three final selected species: Philypnodon grandiceps, Nannoperca australis and Retropinna semoni, adapted and compiled from Llewellyn (1974), Pusey et al. (2004), McDowell (1996).

Vol. 16, pp. 89-101, 2010


Flathead gudgeonP. grandiceps

Southern pygmy perchN. australis

Australian smeltR. semoni

Identification Small fish with variable colour (from black to brown), overall markings on the sides and dark spots at caudal fin base.

Small, laterally compressed with a deeply notched single dorsal fin. Body colour varies: white on the belly with series of dark botches on the side.

Small, elongate laterally compressed body. Large eyes and a slightly rounded mouth. Small dorsal fin set back on the body. Colour usually olive on the back to bright silver on the flanks, with a silver white belly (McDowall 1996).

Habitat Lakes, reservoirs and brackish, estuaries. Bottom dwelling species preferring mud, rock substrates and weedy patches.

Slow-flowing streams with abundant aquatic vegetation or in weedy lakes and billabongs.

From slow moving or still water in a variety of habitats.

Distribution Qld to the Murray River, SA. Also inland waters of the Murray-Darling drainage from Lachlan River, NSW to the Murray River, SA. Also northern Tasmania.

Vic (common), NSW (rare) and SA (a few tributaries of the Murray River in the Mount Lofty Ranges).

Coastal streams from central Qld through to the Murray Mouth in SA.

Diet Aquatic insects, molluscs and small fishes.

Small crustaceans and invertebrates.

Planktonic organisms, micro-crustaceans and small aquatic insects.

Age at sexual maturity

? 6 months 12 months

Length Maximum 12 cm, commonly 8.0 cm

Maximum 8.5 cm; commonly3-5 cm

Maximum 10 cm; commonly 4-6 cm.

Longevity ? Possibly 3 years 2+ years, possibly 3+

Sex ratio (female:male)

? 1:1 2:1-4:1

Sexual dimorphism

Male: generally darker colours, larger mouth, more bulbous head, broader interorbital and larger pelvic fins than female. In male the mouth generally reaches below pupil, while in female it ends under the front of the eye. Breeding females develop several flaps around the opening of the elongate urinogenital papilla, while males have only small bumps around the opening.

Region of the vent: males heavily pigmented and possessing a minute urinogenital papilla. Females only slightly pigmented in this region and possess quite a large urinogenital papilla.

Both sexes present nuptial tubercules on body and head but larger in male and also occur in fins. During spawning, males often present a bright orange red colour.

Critical temperature to spawn

18°C in field; 21-27°C in aquaria (Pusey et al. 2003).

21.6-22.1°C in field (Llewellyn, 1974).

15°C in field; 21-24°C in aquaria.

Peak spawning activity

? August- October July–September winter and mainly spring

Fecundity (number ova)

Average 1400-2300; number of eggs increases with size.

Depends on fish size: 560- 4217 eggs.

Total fecundity= 106-1023

Egg size (diameter)

Elongate, pointed at the basal sticky end and blunt at the other. 1.5-2.2 mm long; 0.7-0.9 mm wide.

Demersal, transparent, spherical, non-adhesive. 1.16-1.35 mm wide.

Intra-ovarian eggs 0.73 mm, newly-laid eggs 0.80 mm, water-hardened eggs 0.95 mm

Frequency of spawning

? Multiple times ?

97


NSW: New South Wales; Vic: Victoria; Qld: Queensland; SA: South Australia.

Table 3. Continued. Comparison of the biology and life parameters of the three final selected species: Philypnodon grandiceps, Nannoperca australis and Retropinna semoni, adapted and compiled from Llewellyn (1974), Pusey et al. (2004), McDowell (1996).

The Australian smelt covers a wide area of freshwaters biomonitoring ranging from central QLD through to the Murray mouth in SA and has the widest distribution of all the species considered. However, the flathead gudgeon, may offer the advantage of multiple approaches, allowing researchers to assess both estrogenic and androgenic mimics since the female and males have distinct secondary sexual characteristics while the elaborate courtship behaviour of this species as well as the parental care of this species could be potential endpoints of the effects of estrogenic chemicals on courtship behaviour and parental care. Elaborate courtship displays prior to spawning have been recorded in captivity as well as the intensive parental care by male fish (Pusey et al. 2003). Parental care starts during spawning, where male P. grandiceps have been observed to hover in a stationary position adjacent to the eggs mass (Robertson 1968). Immediately after spawning the male takes care of the eggs, guarding and fanning the eggs with pectoral fins until they hatch. The aggressive behaviour of this species when guarding the eggs has also been recorded (Pusey et al. 2003). Behavioural endpoints have been little integrated in aquatic toxicology, since until recently, there was a poor understanding of how alterations in behaviour may be related to ecologically relevant issues such as predation avoidance, prey capture, growth, stress resistance, reproduction, and longevity (Kane et al. 2005).

In conclusion, the three selected species are sexually dimorphic, widely distributed, non- migratory and have relatively short life cycles reaching sexual maturity in their first year. Clear sexual dimorphism allows non-destructive

sampling and morphological assessment. However, the same disadvantages of the lack of baseline studies and limited numbers that can be obtained from the field apply to all three species.

It is recommended that the proposed species may be used in a multi species approach in order to obtain more reliable results and avoid any under/over estimation of EDC effects on indigenous species caused by using a single biomonitor.

acKnowledgmentsThe author A. Miranda is supported by PhD scholarship from the Victorian Centre of Aquatic Pollution Identification and Management, CAPIM, The University of Melbourne, Parkville, VIC, 3052, Australia.

Vol. 16, pp. 89-101, 2010


Flathead gudgeonP. grandiceps

Southern pygmy perchN. australis

Australian smeltR. semoni

Oviposition and spawning site

Eggs are demersal, adhesive. Eggs are attached to a substrate (wood, rock) in a single cluster.

Slow flowing waters heavily weeded. Eggs are randomly dispersed in weedy areas.

Adhesive, demersal eggs scattered over gravel or attached to aquatic vegetation.

Courtship display Elaborate courtship displays prior to spawning.

Parental care Male takes care of the eggs immediately after spawning, aggressive against possible intruders. Fans the eggs with pectoral fins until they hatch.

Not known None known

Time to hatching 4-6 days after fertilisation 66-79 hours 9-10 days Length at hatching

Newly-hatched larvae are from 3.7 mm to 3.9 mm.

3.45 (3.18-3.92) mm ?

Length at free swimming

? 6 hours ?

Age at first feeding

3 days 1-2 days

Juvenile stage ? 2.5 months measuring about 1.15 cm length-6 months

?

98


referencesAllen GR, Midgley SH and Allen M. 2002. Field Guide to the Freshwater Fishes of Australia. Western Australia Museum, Perth, Australia.

Allinson G, Allinson M, Shiraishi F, Salzman SA, Myers JH, Hermon KM and Theodoropoulos T. 2008. Androgenic activity of effluent from forty-five municipal water treatment plants in Victoria, Australia. WIT Transactions on Ecology and the Environment 110, 293-304.

Allinson M, Shiraishi F, Salzman SA and Allinson G. 2010. In vitro and immunological assessment of the estrogenic activity and concentrations of 17 β estradiol, estrone and ethinyl estradiol in treated effluent from forty-five waste water treatment plants in Victoria, Australia. Archives of Environmental Contamination and Toxicology 58, 576- 586.

Bandelj E, van den Heuvel MR, Leusch FDL, Shannon N, Taylor S and McCarthy LH. 2006. Determination of the androgenic potency of whole effluents using mosquitofish and trout bioassays. Aquatic Toxicology 80, 237–248.

Barbour MT, Gerritsen J, Snyder BD and Stribling JB. 1999. Chapter 3. Elements of biomonitoring. In Rapid Bioassessment Protocols for Use in Streams and Wadeable Rivers: Periphyton, Benthic Macroinvertebrates and Fish, Second Edition. EPA 841-B-99-002. U.S. Environmental Protection Agency; Office of Water; Washington, D.C.

Batty J and Lim R. 1999. Morphological and reproductive characteristics of male mosquitofish (Gambusia affinis holbrooki) inhabiting sewage-contaminated waters in New South Wales, Australia. Archives of Environmental Contamination and Toxicology 36, 301-307.

Beyer WN, Heinz GH and Redmon-Norwood AW. 1996. Environmental Contaminants in Wildlife: Interpreting Tissue Concentrations, Lewis Publishers, Boca Raton, FL, USA.

Chapman H, Leusch F and Tan B. 2008. Risk management of chemicals in recycled water. In Purified Recycled Water for Drinking: The Technical Issues. Chapter 5. Queensland Water Commission, Brisbane, QLD, Australia. 273pp.

Colborn T, vom Saal FS, Soto AM. 1993. Developmental effects of endocrine–disrupting chemicals in wildlife and humans. Environmental Health Perspectives 101, 378-384.

Damstra T, Barlow S, Bergman A, Kavlock R and Van Der Kraak G. 2002. Global Assessment of the State-of-the-science of Endocrine Disruptors. WHO/PCS/EDC/02/2. Accessed at <http://www.who.int/ipcs/publications/new_issues/endocrine_disruptors/en/>.

Dietrich DR, Krieger H, Rumpf S, Segner H, Wester P, Fournie JW, Wolf JJ, van der Ven L and Gimeno S. 2008. Histology and Histopathological Analysis of Endocrine Disruptive Effects on Small Laboratory Fish. Wiley-Interscience, New York.

Falconer IR, Chapman HF, Moore MR and Ranmuthugala G. 2006. Endocrine-disrupting compounds: A review of their challenge to sustainable and safe water supply and water reuse. Environmental Toxicology 21, 181-191.

Frame LT and Dickerson RL. 2006. Fish and wildlife as sentinels of environmental contamination. In Endocrine Disruption: Biological Basis for Health Effects in Wildlife and Humans, eds. Norris DO and Carr JA. Oxford University Press, NY, pp 202-222.

Game C, Gagnon MM, Webb D and Lim R. 2006. Endocrine disruption in male mosquitofish (Gambusia holbrooki) inhabiting wetlands in Western Australia. Ecotoxicology 15, 665-672.

Geyer HJ, Rimkus GG, Scheunert I, Kaune A, Schramm K-W, Kettrup A, Zeeman M, Muir DCG, Hansen LG, Mackay D. 2000. Bioaccumulation and occurrence of endocrine- disrupting chemicals (EDCs), persistent organic pollutants (POPs), and other organic compounds in fish and other organisms including humans. In The Handbook of Environmental Chemistry, Vol. 2 Part J. Bioaccumulation. Beek B (Ed), Springer-Verlag, Berlin & Heidelberg, Germany. pp 1-166.

Gibson CP and Wilson SP. 2003. Imposex still evident in eastern Australia 10 years after tributyltin restrictions. Marine Environmental Research 55, 101–112.

Goodhead, RM and Tyler CR. 2009. Endocrine-disrupting chemicals and their environmental impacts. In Organic Pollutants - An Ecotoxicological Perspective. Walker CH (Ed), CRC Press, Boca Raton, Fl, USA. pp. 265-292.

Hamling HJ and Guillette LJ Jr. 2010. Wildlife as sentinels of environmental impacts on reproductive health and fertility. Ch.8. In: Environmental Impacts on Reproductive Health and Fertility. Woodruff TJ, Janssen SJ, Guillette LJ and Giudice LC (Eds). Cambridge University Press, UK. ISBN:9780511674686. pp. 103-114

Harris JH and Gehrke PC. 1997. Fish and Rivers in Stress: the NSW Rivers Survey.CRC for Freshwater Ecology, Canberra, ACT, and New South Wales Fisheries, Cronulla, NSW Australia.

Hashimoto T, Onda K, Nakamura Y, Tada K, Miya A, Murakami T. 2007. Comparison of natural estrogen removal efficiency in the conventional activated sludge process and the oxidation ditch process. Water Research 41, 2117-2126.

Humphries P, King AJ, and Koehn JD. 1999. Fish, flows and floodplains: links between freshwater fishes and their environment in the Murray-Darling River system, Australia. Environmental Biology of Fishes 56, 129-151.

Humphries P. 1995. Life history, food and habitat of southern pygmy perch, Nannoperca australis in the Macquarie River, Tasmania. Marine and Freshwater Research 46, 1159-1169.

Jašarevic E, Sieli PT, Twellman EE, Welsh Jr TH, Schachtman TR, Michael RR, Geary DC and Rosenfeld CS. 2011. Disruption of adult expression of sexually selected traits by developmental exposure to bisphenol A. Proceedings of the National Academy of Sciences 108, 11715–11720

Jobling S, Coey S, Whitmore JG, Kime DE, Van Look KJW, McAllister BG, Beresford N, Henshaw AC, Brighty G, Tyler CR, and Sumpter JP. 2002. Wild intersex roach (Rutilus

Vol. 16, pp. 89-101, 2010


99


rutilus) have reduced fertility. Biology of Reproduction 67, 515-524.

Jobling S and Tyler CR. 2006. Introduction: The ecological relevance of chemically induced endocrine disruption in wildlife. Environmental Health Perspectives 114, 7-8.

Jobling S. 2007. Drugs are in the water - does it matter? Keynote Address at What’s in Our Water: The Significance of Trace Organic Compounds. 2nd Australian Symposium on Ecological Risk Assessment and Management of Endocrine Disrupting Chemicals (EDCs), Pharmaceuticals and Personal Care Products (PPCPs) in the Australasian Environment, November 21-22 2007, CSIRO Discovery Centre, Black Mountain, Canberra, Australia. http://www.clw.csiro.au/video_html/2007/susanJobling/

LWA, 2007. The Black Mountain Declaration on Endocrine Disrupting Chemicals in Australian Waters 2007. Consensus statement prepared at What’s in Our Water: The Significance of Trace Organic Compounds (2007), 2nd Australian Symposium on Ecological Risk Assessment and Management of Endocrine Disrupting Chemicals (EDCs), Pharmaceuticals and Personal Care Products (PPCPs) in the Australasian Environment. November 21-22 2007, CSIRO Discovery Centre, Black Mountain, Canberra. Land and Water Australia, Canberra. http://www.clw.csiro.au/conferences/ourwater/EDC-conference-declaration.pdf

Kane AS, Salierno JD, Brewer SK. 2005. Fish models in behavioural toxicology: automated techniques, updates and perspectives. In Methods in Aquatic Toxicology, Vol. 2. Ostrander GK (Ed), Lewis Publishers, Boca Raton, FL, USA. pp 559–590

Katsiadaki I, Sanders M, Sebire M, Nagae M, Soyano K and Scott AP. 2007. Three-spined stickleback: An emerging model in environmental endocrine disruption. Environmental Sciences 14, 263-283.

Kidd KA, Blanchfield PJ, Mills KH, Palace VP, Evans RE, Lazorchak JM and Flick R W. 2007. Collapse of a fish population after exposure to a synthetic estrogen. Proceedings of the National Academy of Sciences of the United States of America 104, 8897-8901.

Kloas W. Urbatzka R, Opitz R, Würtz S, Behrends R, Hermelink B, Hofmann F, Jagnytsch O, Kroupova H, Lorenz C, Neumann N, Pietsch C, Trubiroha A, Van Ballegooy C, Wiedemann C and Lutz I. 2009. Endocrine disruption in aquatic vertebrates. Annals of the New York Academy of Science 1163, 187-200.

Legget R and Merrick JR. 1987. Australian Native Fishes for Aquariums. J.R. Merrick Publications, Artarmon, NSW, Australia. 241 pp.

Leusch FDL, Chapman HF, Gooneratne SR and Tremblay LA. 2004. Use of bioassays in the study of endocrine disruption: a case study with sewage in Australia and New Zealand. In Proceedings of the APEC Workshop MRC02/2004 - Modern Approaches to Linking Exposure to Toxic Compounds and Biological Responses, 12-16 July 2004, Gold Coast, Qld, Australia. p 14.

Leusch FDF, Chapman HF, Kay GW, Gooneratne SR and Tremblay LA. 2006. Anal fin morphology and gonadal histopathology in mosquitofish (Gambusia holbrooki) exposed to treated municipal sewage effluent. Archives of Environmental Contamination and Toxicology 50, 562-574.

Leusch FDL, Chapman HF, Korner W, Gooneratne SR and Tremblay LA. 2005. Efficacy of an advanced sewage treatment plant in southeast Queensland, Australia, to remove estrogenic chemicals. Environmental Science and Technology 39, 5781-5786.

Llewellyn LC. 1974. Spawning, development and distribution of the southern pigmy perch Nannoperca australis australis Gunther from inland waters in eastern Australia. Australian Journal of Marine and Freshwater Research 25, 121-149.

McDowall RM. 1996. Freshwater fishes of South Eastern Australia. Reed books, Chatswood, NSW, Australia.

Mispagel C, Allinson G, Allinson M, Shiraishi F, Nishikawa M and Moore M. 2009. Observations on the estrogenic activity and concentration of 17β-estradiol in the discharges of 12 wastewater treatment plants in southern Australia. Archives of Environment Contamination and Toxicology 56, 631–637.

Moncaut N, Lo Nostro F and Maggese MC. 2003. Vitellogenin detection in surface mucus of the South American cichlid fish Cichlasoma dimerus (Heckel, 1840) induced by estradiol-17beta. Effects on liver and gonads. Aquatic Toxicology 63, 127-137.

Nagpal NK and Meays CL. 2009. Water Quality Guidelines for Pharmaceutically-Active Compounds (PhACs): 17α-Ethinylestradiol (EE2). Overview Report. Science and Information Branch, Water Stewardship Division, Ministry of Environment, Ministry of Environment Province of British Columbia, Canada.

Nias DJ, McKillup SC and Edyvane KS. 1993. Imposex in Lepsiella vinosa from southern Australia. Marine Pollution Bulletin 26, 380-384.

NICNAS. 1998. Endocrine Disrupting Chemicals. An Information Paper. National Industrial Chemicals Notification and Assessment Scheme. Accessed 3rd June 2012 <http://www.nicnas.gov.au/international/Endocrine_Disputing_Chemicals/EDC_Policy_Paper_PDF.pdf>

Örn S, Holbech, H, Madsen TH, Norrgren L and Petersen GI. 2003. Gonad development and vitellogenin production in zebrafish (Danio rerio) exposed to ethinylestradiol and methyltestosterone. Aquatic Toxicology 65, 397–411.

Parrott JL and Blunt BR. 2005. Life-cycle exposure of fathead minnows (Pimephales promelas) to an ethinylestradiol concentration below 1 ng/L reduces egg fertilization success and demasculinizes males. Environmental Toxicology 20, 131-141.

Pawlowski S, Ternes TA, Bonerz M, Rastall AC, Erdinger L and Braunbeck T. 2004. Estrogenicity of solid phase-extracted water samples from two municipal sewage treatment plant effluents and river Rhine water using the yeast estrogen screen. Toxicology In Vitro 18, 129–138.

Vol. 16, pp. 89-101, 2010


100


Pollino CA, Georgiades E and Holdway DA. 2007. Use of the Australian crimson-spotted rainbowfish (Melanotaenia fluviatilis) as a model test species for investigating the effects of endocrine disruptors. Environmental Toxicology and Chemistry 26, 2171-2178.

Pusey B, Kennard M and Arthington A. 2004. Freshwater Fishes of North-Eastern Australia. CSIRO Publishing, Australia.

Rasmussen TH, Andreassen TK, Pedersen SN, Van der Ven LTM, Bjerregaard P and Korsgaard BK. 2002. Effects of waterborne exposure of octylphenol and oestrogen on pregnant viviparous eelpout (Zoarces viviparus) and her embryos in ovario. Journal of Experimental. Biology 205, 3857-3876.

Reitsema TJ and Spickett JT. 1999. Imposex in Morula granulata as bioindicator of tributyltin (TBT) contamination in the Dampier Archipelago, Western Australia. Marine Pollution Bulletin 39, 280-284.

Robertson DR. 1968. A Comparative Study of the Reproductive Behaviour of Australian Gudgeons (Pisces: Eleotridae). B.Sc. Honours Thesis, University of Queensland, Brisbane, Australia.

Rose G, Allen, D, Allinson G, Allinson M, AnhDuyen B, Wightwick A and ,Zhang P. 2009. Melbourne Water and DPI Agrochemicals in Port Philip Catchment Streams – Project Summary Report on 2008-09. Department of Primary Industries (DPI), Victoria, Australia.

Sarmah AK, Northcott GL, Leusch FDL and Tremblay LA. 2006. A survey of endocrine disrupting chemicals (EDCs) in municipal sewage and animal waste effluents in the Waikato region of New Zealand. Science of the Total Environment 355, 135–144.

Segner H (2007). Assessing the hazard of endocrine-disrupting compounds in fish: zebrafish (Danio rerio) as a model species. In Progress in Environmental Science and Technology Wang Y, Li S, Huang P, Yang Y, An Y and Sun X (Eds). Proceedings of the 2007 International Symposium on Environmental Science and Technology, Beijing, November 13-16, 2007. Science Press USA, Beijing. pp. 99-106.

Servos MR, Bennie DT, Burnison BK, Jurkovic A, McInnis R, Neheli T, Schnell A, Seto P, Smyth SA and Ternes TA. 2005. Distribution of estrogens, 17β-estradiol and estrone, in Canadian municipal wastewater treatment plants. Science of the Total Environment 336¸ 155–170.

Seki M, Yokota H, Maeda M and Kobayashi K. 2004. Fish full life-cycle testing for 17β-estradiol on medaka (Orizias latipes). Environmental Toxicology and Chemistry 24, 1259–1266.

Shanthanagouda AH, Patil JG and Nugegoda D. 2012.Ontogenic and sexually dimorphic expression of cyp19 isoforms in the rainbowfish, Melanotaenia fluviatilis (Castelnau 1878). Comparative Biochemistry and Physiology, Part A 161, 250-258.

Stahl RG Jr. 1997. Can mammalian and non-mammalian “sentinel species” data be used to evaluate the human health implications of environmental contaminants? Human and Ecological Risk Assessment 3, 329-335.

Sumida K, Ooe N, Saito K, Kaneko H. 2003. Limited species differences in estrogen receptor alpha-medicated reporter gene transactivation by xenoestrogens. Journal of Steroid Biochemistry and Molecular Biology 84, 33-40.

Sumpter JP and Jobling S. 1995. Vitellogenesis as a biomarker for estrogenic contamination of the aquatic environment. Environmental Health Perspectives 103 (Suppl. 7), 173–178.

Sumpter JR. 2003. Endocrine disruption in wildlife: The future? Pure and Applied Chemistry 75, 2355-2360.

Tan BLL, Hawker DW, Müller JF, Leusch FDL, Tremblay LA and Chapman HF. 2007. Comprehensive study of endocrine disrupting compounds using grab and passive sampling at selected wastewater treatment plants in South East Queensland, Australia. Environment International 33, 654–669.

Taylor MR and Harrison PTC. 1999. Ecological effects of endocrine disruption: current evidence and research priorities. Chemosphere 39, 1237–1248.

Taylor MR, Holmes P, Duarte-Davidson R, Humfrey CDN and Harrison PTC. 1999. A research strategy for investigating the ecological significance of endocrine disruption: report of a UK workshop. Science of the Total Environment 233, 181–191.

Tyler CR and Goodhead RM. 2010. Impact of hormone-disrupting chemicals on wildlife. In Silent Summer: The State of Wildlife in Britain and Ireland, Maclean N (Ed), Cambridge University Press, UK. pp. 125-140.

Tyler CR, Jobling S and Sumpter JP. 1998. Endocrine disruption in wildlife: A critical review of the evidence. Critical Reviews in Toxicology 28, 319–361.

Thorpe KL, Cummings RI, Hutchinson TH, Scholze M, Brighty G, Sumpter JP and Tyler CR. 2003. Relative potencies and combination effects of steroidal estrogens in fish. Environmental Science and Technology 37, 1142-1149.

Van Der Kraak G, Hewitt M, Lister A, McMaster ME and Munkittrick KR. 2001. Endocrine toxicants and reproductive success in fish. Human and Ecological Risk Assessment 7, 1017-1025.

VanVeld PA, Rutan BJ, Sullivan CA, Johnston LD, Rice CD, Fisher DF and Yonkos LT. 2005. A universal assay for vitellogenin in fish mucus and plasma. Environmental Toxicology and Chemistry 24, 3048–3052.

vom Saal, FS, Guillette JR Jr, Myers JP and Swan S. 2008. Endocrine disruptors: Effects in wildlife and laboratory animals. In Encyclopaedia of Ecology, Jorgensen SE, Fath BD (Eds). pp 1261-1264.

Vos JG, Dybing E, Greim HA, Ladefoged O, Lambré C, Tarazona JV, Brandt I and Vethaaak AD. 2000. Health effects of endocrine-disrupting chemicals on wildlife, with special

Vol. 16, pp. 89-101, 2010


101


reference to the European situation. Critical Reviews in Toxicology 30, 71-133.

WHO (World Health Organization). 2001. Integrated Risk Assessment. Report prepared for the WHO/UNEP/ILO International Programme on Chemical Safety. WHO/IPCS/IRA/01/12. Available online at: <http://www.who.int/ipcs/publications/new_issues/ira/en/>

Williams M, Woods M, Kumar A, Ying GG, Shareef A, Karkkainen M and Kookana R. 2007. Endocrine Disrupting Chemicals in the Australian Riverine Environment, A Pilot Study on Estrogenic Compounds. Land & Water Australia, Canberra, ACT Australia. 89 pp.

Williams W and Allen GR .1987. Origins and adaptations of the fauna of Australian inland waters. In Fauna of Australia. General Articles, Vol. 1A. Dyne GR and Walton DW (Eds) pp. 184-201. Australian Govt. Publishing Service, Canberra, Australia.

Woods M and Kumar A. 2011. Vitellogenin induction by 17β-estradiol and 17α-ethynylestradiol in male Murray rainbowfish (Melanotaenia fluviatilis). Environmental Toxicology and Chemistry 30, 2620-2627.

Woods M. 2007. Assessment of Xenoestrogens in the Australian Freshwater Environment: Use, Development and Validation of In-Vitro and In-Vivo Models. PhD Thesis. Sansom Institute, Division of Health Sciences. University of South Australia, Adelaide, Australia.

Ying GG and Kookana RS. 2002. Endocrine disruption: an Australian perspective. Water 29, 42-45.

Ying GG, Kookana RS, Kumar A and Mortimer M. 2009. Occurrence and implications of estrogens and xenoestrogens in sewage effluents and receiving waters from South East Queensland. Science of the Total Environment 407, 5147–5155.

Zelikoff JT. 1994. Chapter 5: Fish immunotoxicology. In Immunotoxicology and Immunopharmacology. Dean JH, Luster MI and Munson AE (Eds), Raven Press, New York, USA. pp. 71-95.

Vol. 16, pp. 89-101, 2010


Documents

a freshwater fish biomonitor for edcs in southern victoria - … · a freshwater fish biomonitor for edcs in southern victoria Ana F. Miranda1,2,*, Vincent Pettigrove2, Dayanthi Nugegoda1,2