Thesis Submitted in Fulfilment of the Requirements for the Degree of Masters of Science by Research

Factors influencing hybridisation and introduction success of the critically endangered Running River rainbowfish, Melanotaenia sp.

Karl Moy B. Zool. (Hons)

Institute for Applied Ecology

University of Canberra

A thesis submitted in fulfilment of the requirements for the degree of Masters of Science by Research

7th of August 2019

Abstract Given that conservation introductions are essentially biological invasions, researching the main factors which influence them will provide insight for both conservation and management. The factors affecting invasion success in small-bodied Australian freshwater fishes are largely unstudied. From a conservation-oriented perspective this is worrying as small-bodied freshwater species are more likely to become threatened than large-bodied species. It is equally concerning from an invasive species management perspective as many species have the potential to negatively impact native species and ecosystems. This thesis consists of two data chapters preceded by a general introduction and followed by a synthesis.

The first data chapter examines potential pre and post zygotic barriers to hybridisation between the Running River rainbowfish (RRR) and eastern rainbowfish (Melanotaenia splendida). Eastern rainbowfish is a widespread native fish of northern Australia with an alien population in Running River (a tributary of the Burdekin River, Queensland). Hybridisation between RRR and eastern rainbowfish has been detected and in the absence of barriers to further hybridisation and introgression will likely lead to the loss of pure RRR from the wild. Dichotomous mate choice experiments and egg survival experiments were used to determine the presences of pre and post-zygotic barriers to hybridisation between RRR and eastern rainbowfish. The findings of this study do not support the presence of barriers to hybridisation between the two species. Mate choice experiments suggested that males and females of both species examined did not exhibit preferential mate choice, which is most unusual. Similarly, egg survival experiments found no differences in fertility or reproductive success between hybrid and pure crosses. The chapter is presented as a stand-alone article and has been accepted for publication by the journal Ethology.

The second data chapter focuses on conservation introductions conducted for RRR into refuge habitats free of eastern rainbowfish. Releases made into one of these release sites (Deception Creek) were used to examine the impact of predator avoidance training on the survival of captive bred RRR after release. Deception Creek received approximately 2500 fish while Puzzle Creek received approximately 1500 fish. RRR bred for release into Deception Creek exposed to a novel predator found within the release sites (spangled perch, Leipotherapon unicolor) using repeated controlled exposures prior to release. Experiments regarding predator training were ended prematurely due to unexpected weather conditions. This reduced the number of replicates and prevented any reliable statistical analysis of collected data. However, analysis of collected data suggested there was no difference between

iii the survival of trained and untrained fish. Despite this, anecdotal observations suggest that pre-exposure to predators may have some beneficial effects on the survival of captive fish. This chapter also discusses movement data gained from these releases, something which is lacking for most native small-bodied fish species. This chapter is presented as a stand-alone article that will form part of a larger publication covering the discovery, description and conservation of RRR.

Acknowledgments I would like to thank my supervisory panel (Peter Unmack, Mark Lintermans, Richard Duncan, and Culum Brown) for their advice, support, and knowledge. I hold each of them in the highest regard and aspire to the examples of a scientist that they have set. A special thanks to Jason Schaffer for his help with predator training, to Michael Hammer for providing gill nets, advice and mentorship, and to Brendan Ebner for his advice on sampling techniques. I must also thank the Australian Wildlife Conservancy and in particular Tim and Bree White for their support of conservation actions for the Running River rainbowfish and for providing me with lodgings while conducting field work. Steve Hume has been instrumental in communicating the research and conservation of Running River rainbowfish to the wider community. This, alongside his contributions to field work earn him a great deal of my gratitude. A large thanks also goes out to many other students and peers who have helped me along the way to complete my masters and supported my passion for small silvery fish.

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Table of Contents Abstract ...... iii Form B Certificate of Authorship of Thesis ...... v Acknowledgments ...... vii Table of Contents ...... ix Chapter 1: Introduction ...... 1 The specific problem of hybridisation with invasive/translocated species ...... 2 The potential for conservation introductions to alleviate these problems...... 3 The case of rainbowfishes ...... 5 The aims of this project...... 7 References ...... 9 Chapter 2: Barriers to hybridisation and their conservation implications for a highly threatened Australian fish species ...... 23 Declaration for Thesis Chapter 2 ...... 23 Abstract ...... 25 Introduction ...... 26 Methods ...... 30 Maintenance ...... 30 Mate preference ...... 30 Treatments ...... 30 Apparatus ...... 31 Protocol ...... 31 Analysis ...... 32 Embryonic survival ...... 33 Treatments ...... 33 Protocol ...... 33 Analysis ...... 33 Results ...... 34 Discussion ...... 34 References ...... 38 Chapter 3: A test of the introduction of a predator-naïve and predator trained threatened small- bodied freshwater fish to a novel environment...... 51 Abstract ...... 51 Introduction ...... 51 Methods ...... 54 Study species...... 54 Breeding and Rearing ...... 55

Training ...... 56 Habitat Assessment at release sites ...... 56 Deception Creek releases ...... 57 Puzzle Creek releases ...... 59 Analysis ...... 59 Results ...... 60 Habitat ...... 60 Predator effects ...... 60 Dispersal ...... 61 Discussion ...... 61 References ...... 64 Chapter 4 Synthesis ...... 73 Thesis overview ...... 73 Hybridisation and introgression risk ...... 73 Embryonic survival experiments ...... 74 Conservation releases ...... 75 Predator Training ...... 75 Rainbowfish movement ...... 76 Implications for management ...... 77 The importance of monitoring ...... 77 An argument for moving more than the species ...... 78 Contrasts between small- and large-bodied fish translocations ...... 78 Knowledge gaps and directions for future research ...... 79 Conclusion ...... 80 References ...... 81 Appendix A – Weight and body depth presented as a function of length in RRR and eastern rainbowfish (Melanotaenia splendida) ...... 87 Appendix B – Correlation between spangled perch density and adult rainbowfish density at 11 months from release ...... 89 Appendix C – Summary statistics of body size measurements of fish used in mate preference experiments ...... 91 Appendix D – Statistical output from analysis of length, weight and depth relationships ...... 93

List of Figures and Tables

Figure 1 Freshly fixed morphological specimens of male Running River rainbowfish (sp. nov., top) and eastern rainbowfish (Melanotaenia splendida, bottom). Photograph credit to Michael Hammer from the Northern Territory Museum...... 20 Figure 2 Map of the study area showing the location of Puzzle and Deception creeks and their positions relative to Running River and its gorges. Purple arrows indicate the range of Running River rainbowfish while orange arrows indicate the range of eastern rainbowfish (Melanotaenia splendida). Credit goes to Tani Cooper with permission provided by the Australian Wildlife Conservancy...... 21 Figure 3 Images of male Running River rainbowfish (a*, c) and Melanotaenia splendida introduced into Running River (b*, d) displaying their natural colours in an aquarium (top) and the colours they displayed during mate preference trials (bottom).*Credit Gunther Schmida...... 46 Figure 4 A sketch of the apparatus showing the layout. The dotted line represents the clear acrylic barrier, which restricts movement but not water or sight, while the solid lines represent solid barriers that restrict movement, water and sight. The dashed and diagonal lines represent a 10 cm strip of white corflute laid on the substrate which was used as the preference zone. The dashed circle represents the plastic bottle use to acclimate the test fish, the circle at the far left shows the air stone and the six- pointed stars represent spawning mops...... 47 Figure 5 Distribution plot showing the proportion of time each test group spent associating with a potential mate of the same species during interspecific mate preference trials involving Melanotaenia splendida (M. splendida) and Running River rainbowfish (RRR). Mean and standard error from left to right: 0.586± 0.0622, 0.519 ± 0.0691, 0.393 ± 0.0840, and 0.438 ± 0.0662...... 48 Figure 6 The proportion of time that a subject spent with an option fish as a function of its size relative to the other option in intraspecies trials separated by sex and species: (a) male Running River rainbowfish (RRR), (b) male Melanotaenia splendida, (c) female RRR, (d) female M. splendida...... 49 Figure 7 The mean proportion of collected eggs which were fertile, mean proportion of fertilised eggs which survived to hatching, and mean proportion of collected eggs which resulted in hatching with 95% confidence intervals for all tested combinations involving Melanotaenia splendida (ER) and the Running River rainbowfish (RRR). The first individual in a cross represents the male...... 50 Figure 8 Abundance of released fish over time during the first field season in Deception Creek for trained fish and untrained fish. Different colours represent different release sites. Note, the number of released fish cannot increase as fish were only released once into each site...... 72

Table 1 Output from regression analysis and one-way analysis of variance from intraspecies and interspecies dichotomous mate choice experiments respectively, involving males and females from both Running River rainbowfish (RRR) and Melanotaenia splendida. Regression analysis examined the connection between proportional size of a subject and the proportion of time it received, while the one-way analysis of variance compared the time subjects spent with their own species with the time they spent with members of the other species...... 45 Table 2 Statistical output comparing trained and untrained releases of fish. Statistical output comparing trained and untrained releases of fish. A Welch’s t-test (W) compared total observed abundance at 2-3 weeks from release, while a paired t-tests (P) compared the density of adults at 6 and 11 months from release. Standard error (SE) was calculated from 11 abundance observations between two sites which all fell within 4 days of one another converted to a percentage and then applied to all samples...... 71 Table 3 Statistical output from linear regression analysis testing predator density as a predictor of rainbowfish abundance in the first month, and density 6 and 11 months from release...... 71

Table 4 Upstream and downstream movement distances of Running River rainbowfish from their release sites in Deception and Puzzle creeks over time. The distance upstream is shown first in kilometres, while the changes in elevation are shown in brackets given in metres...... 71

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Chapter 1: Introduction

Translocation is a worldwide phenomenon, with alien species currently present on every continent (Prins and Gordon 2014). Translocation of organisms is a major threat to many ecosystems worldwide (Clavero and Garcia-Berthou 2005; Gallardo et al. 2016; Vitousek et al. 1997), but can also be a useful tool for the conservation of biodiversity (IUCN 2013; Minckley 1995; Tuberville et al. 2005). The International Union for Conservation of Nature defines translocation as the human-mediated movement of living organisms from one area, with release in another (IUCN 2013). When a species is translocated to an area outside its natural range, this is known as an introduction (IUCN 2013). When a species is introduced it is considered alien within the new range, even if its natural range is close by, or the species is considered native to the country. Translocations can be intentional or unintentional, although both kinds often have unintended consequences (Arbačiauskas et al. 2010; Cunningham 1996; Pusey et al. 2003). Alien species, are often considered invasive, meaning that once established in a new environment they have the potential to disperse and cause significant harm to ecosystems, species, or habitats (McNeely 2001). Although often used in conjunction, the terms invasive and alien are not interchangeable and a species may fit the definition of one but not the other (McNeely 2001).

Invasive alien species are one of the greatest causes of species declines in the world (Bax et al. 2003; Clavero and Garcia-Berthou 2005; Gallardo et al. 2016; McNeely 2001) and can cause significant ecological and economic damage (Francis 2012; McLeod and Norris 2004; McNeely 2001). Alien species can negatively impact native species through habitat degradation (Costin and Moore 1960; Nimmo and Miller 2007), competition (Holway 1999; McIntosh et al. 1992), predation (Barlow et al. 1987; Pen and Potter 1992), hybridisation (Huxel 1999; Wolf et al. 2001), spreading disease (Cunningham 1996), and aggressive interactions (Howe et al. 1997; Keller and Brown 2008). These effects can lead to the extinction of native species (Burbidge and Manly 2002; Clavero and Garcia-Berthou 2005; Vitousek et al. 1997). In addition to causing significant environmental and ecological damage, alien species also have a considerable economic impact. In Australia alone, the estimated combined cost of 11 alien vertebrate species was over $720 million per year (McLeod and Norris 2004), and these costs have undoubtedly continued to rise.

For an invasion to succeed, a species must enter a novel environment, establish a self- sustaining population, and then disperse to occupy areas outside of the initial release site (Blackburn et al. 2011). However, environmental (e.g. drought, habitat availability/suitability) and biotic pressures (propagule size, predation, competitors) will often reduce the chance of establishment (Blackburn et al. 2011; Duncan 2016; Prins and Gordon 2014). Many invasive species tend to have generalist dietary and habitat requirements (Harris 2013; Marchetti et al. 2004) and fill ecological niches underutilised by native species (Harris 2013; Olden et al. 2006; Vila-Gispert et al. 2005). In freshwater fishes, habitat suitability seems to play a larger role in determining the success of an introduction/invasion than biotic resistance, e.g. competitors and predators (Harris 2013; Moyle and Light 1996b).

The specific problem of hybridisation with invasive/translocated species Hybridisation occurs when two individuals from separate species reproduce to make offspring (hybrids). Introgression occurs when hybrids backcross with one or both of the parental species, resulting in the movement of genetic material from one species to another. Hybridisation is influenced by both biotic (organism density, behaviour, range expansions) and abiotic (turbidity, environmental chemical levels) factors (Fisher et al. 2006; Seehausen et al. 1997). Typically, more distantly-related species are less likely to hybridise and introgress, however, barriers to hybridisation and introgression are rarely completely effective between closely related species (Seehausen et al. 1997). These barriers can be prezygotic (behavioural, physical) and postzygotic (hybrid inviability and inferiority) (Boughman 2001; Jones et al. 2006; Rentz 1972; Seehausen et al. 1997). Examples of behavioural barriers include differences in courtship or migratory behaviour (Boughman 2001; Stratton and Uetz 1986). Physical barriers are usually phenotypic differences that physically prevent individuals from mating, such as differences in genital morphology (Rentz 1972). Hybrid inviability is caused by genetic factors, which prevent hybrid pairings from producing viable offspring. They may prevent the fusion of gametes, cause the embryo to die during development, or produce a healthy but sterile individual (Ålund et al. 2013; Jones et al. 2006; Rhymer and Simberloff 1996). Hybrid inferiority refers to maladaptions of hybrid individuals to selection pressures such as competition or predation when compared to one or both parent species (Gilk et al. 2004; Jones et al. 2006; Rhymer and Simberloff 1996).

Hybridisation and subsequent introgression threaten many species world-wide through the loss of genetic and species diversity (Huxel 1999; Riley et al. 2003; Soorae 2018; Unmack et

2 al. 2016). In spite of this, it is also a natural phenomenon which may lead to the transfer of adaptive alleles from ones species to another and increased species diversity (Dowling and Secor 1997). The phenomenon of hybridisation is widespread, but is more commonplace in fishes than any other vertebrate group (Hubbs 1955; Scribner et al. 2000; Verspoor and Hammart 1991). Hybrid offspring may possess competitive advantages over individuals from one or both of their parental species (Hammer et al. 2013a). This can result in the extirpation of one or both parental species through genetic swamping. Alternatively, hybrids may possess deleterious traits via outbreeding depression, leaving them at a disadvantage. Reduced survivorship and reproductive viability are typical symptoms of outbreeding depression and may not occur until the second generation or later (Frankham et al. 2011; Gharrett et al. 1999; Jones et al. 2006; Rhymer and Simberloff 1996). As a result, outbreeding depression may maintain divergence between closely related species in areas of sympatry (Jones et al. 2006). However, outbreeding depression caused by hybridization between native and alien species may lead to the decline of one or both parent species through wasted reproductive effort or reduced survival of offspring (Rhymer and Simberloff 1996).

Throughout Europe and North America, extinctions and declines of native fishes caused by hybridisation with alien species are well documented (Hitt et al. 2003; Ludwig et al. 2009; Meldgaard et al. 2007; Rosenfield and Kodric-Brown 2003). In Australia, introgressive hybridisation between native and alien species has not typically been considered a threat to native biodiversity compared to overseas (Hitt et al. 2003; Ludwig et al. 2009; Meldgaard et al. 2007), because most of the alien species have come from other continents whose biota is taxonomically distant (Lintermans 2013). However, recent broad-scale genetic studies of Australian freshwater fishes have revealed high levels of genetic structuring between populations, as well as many new cryptic species (Hammer et al. 2013a; Hammer et al. 2007; Raadik 2014; Shelley et al. 2018; Unmack 2013; Unmack et al. 2013; Unmack and Dowling 2010). As a result, introgressive hybridisation caused by translocations of ‘native’ species outside their natural range, or from one part of a species’ range to another, has become a more recently recognised threat to conservation for Australian freshwater fishes (Couch et al. 2016; Harris 2013; Lintermans et al. 2005).

The potential for conservation introductions to alleviate these problems While there has been considerable research examining the negative effects of alien species (McNeely 2001; Prins and Gordon 2014), translocations can also be an effective tool for

3 conservation and management (Chilcott et al. 2013; Field et al. 2007; IUCN 2013; Pusey et al. 2003). Translocation has become a key tool for conserving freshwater fishes, utilising both wild and captive-bred fishes (Lintermans 2013; Lintermans et al. 2015; Minckley 1995). In Australia most early conservation translocations of fish have been conducted with larger threatened species that are potential angling targets (Lintermans et al. 2015), such as trout cod (Maccullochela macquariensis, Cadwallader and Gooley 1984), eastern cod (Maccullochella ikei, Nock et al. 2011), Mary River cod (Maccullochella mariensis, Simpson and Jackson 2000) and Macquarie perch (Macquaria australasica, Cadwallader 1981). The practice has recently been applied to smaller threatened native fishes in Australia (Attard et al. 2016; Hammer et al. 2013b; Lintermans et al. 2015). The continued existence of Pedder galaxias (Galaxias pedderensis) is solely due to conservation introductions (Chilcott et al. 2013), while the conservation status of several critically-endangered species such as red-finned blue- eye (Scaturiginichthys vermeilipinnis, Kerezsy and Fensham 2013), and several other galaxiid species (Ayres et al. 2009; Hardie et al. 2006; Koster 2003) have benefited substantially from translocations. A review of factors influencing the success of freshwater fish reintroductions reported that, second to addressing the cause of initial decline, habitat related factors were the greatest predictors of reintroduction success (Cochran-Biederman et al. 2015). The importance of suitable habitat in determining the success or failure of conservation introductions is echoed by studies of invasive fish species, which found that if the habitat characteristics of the receiving environment were suitable, then an invasion was likely to succeed regardless of other factors (Harris 2013; Moyle and Light 1996a; Moyle and Light 1996b). That an introduction will likely fail in the absence of suitable habitat seems rather straightforward, as habitat loss and degradation are one of the greatest threats to freshwater environments (Walker and Steffen 1997); however, some reintroductions may fail even in the presence of apparently adequate habitat (Barlow et al. 1987; Leggett and Merrick 1997).

A review of the factors influencing the success of conservation releases of freshwater fish found that 71% of all failed releases utilised hatchery-reared fish (Cochran-Biederman et al. 2015). Captive-reared fish are often raised under conditions which are vastly different from the environment into which they are released (Brown et al. 2003). Consequently, hatchery- reared fish often exhibit behaviours that are detrimental to their survival in the wild and as a result often suffer high mortality rates once released (Brown and Day 2002; Ebner et al. 2007; Sparrevohn and Støttrup 2007). The behavioural impacts of hatchery rearing have been known for some time (Brown and Day 2002), with hatchery-reared fish showing deficiencies

4 in key behaviours such as predator recognition and avoidance (Ebner et al. 2007; Hutchison et al. 2012) and foraging skills (Brown et al. 2003). Brown et al. (2003) showed that environmental enrichment such as plants, and exposure to live foods resulted in fish being better able to handle novel prey items. Meanwhile, several studies have shown that repeated exposure to predators, or their stimulus, will improve the predator-avoidance behaviours of captive-bred fish (Abudayah and Mathis 2016; Brown 2003a; Vilhunen 2006). As a result, research and implementation of environmental enrichment and predator training of captive- reared fish is becoming more commonplace in Australia (Hammer et al. 2012; Hutchison et al. 2012; Lintermans et al. 2015).

Most research investigating methods to improve the survival of hatchery-reared fishes has taken place overseas, although some recent research has been conducted in Australia (Hutchison et al. 2012). In both cases, research investigating introduction success has focused almost entirely around large-bodied, predatory, recreationally-important species, such as brown and rainbow trouts (Alvarez and Nicieza 2003; Brockmark et al. 2010; Brown and Smith 1998) or percichthyids (Ebner and Thiem 2009; Ebner et al. 2007; Hutchison et al. 2012). However, of the 17 Australian species used in conservation introductions documented by Lintermans et al. (2015), 10 were small-bodied (< 20 cm). Small-bodied species usually have vastly different requirements to large-bodied species (Allen et al. 2002), and a technique that works well for large species may not be as effective for small-bodied species. As small- bodied freshwater fish have a higher risk of extinction (Kopf et al. 2017; Olden et al. 2007; Reynolds et al. 2005), there is a need for a better understanding of factors influencing, and methods for improving, the survival of captive-bred small-bodied freshwater fish once released.

The case of rainbowfishes Rainbowfishes are colourful, pelagic, small-bodied fishes belonging to the family Melanotaeniidae, with more than 100 described species (Eschmeyer and Fong 2017) plus several undescribed taxa (Unmack et al. 2013). They are distributed throughout much of New Guinea and Australia, inhabiting a wide range of habitats, such as ephemeral desert streams, upland perennial creeks, swamps, rivers and lakes (Allen et al. 2002). Hybridisation and introgression have been common between some rainbowfish species in the wild and in most cases, species within different rainbowfish lineages are allopatrically distributed, suggesting an inability to co-occur (Unmack et al. 2013). As a result, translocation of rainbowfish

5 outside their natural range can have serious negative implications for the conservation of other rainbowfish species within the same lineage. Rainbowfishes are common where they occur, often being the most abundant species present (Pusey et al. 2004). Most species reproduce throughout the year, with males competing amongst each other for access to females and spawning substrate (usually aquatic vegetation) while actively courting females. They are often easily caught and thrive in both aquaria and outdoor ponds, which has made them quite popular within the aquarium trade. It has also led to some species being introduced outside their natural range (Brown et al. 2012; Martin 2016; O'Donnell 1935).

Currently, there are four Melanotaenia species listed by the Australian Society for Fish Biology (ASFB) as either vulnerable, endangered, or critically endangered (Lintermans 2016b). In at least three cases, the main threat is hybridisation with eastern rainbowfish (Melanotaenia splendida, Lintermans 2016a). In these cases, hybridisation has occurred because M. splendida has been introduced outside of its natural range. Conservation action for one of the critically endangered species; Running River rainbowfish (RRR, Melanotaenia sp.) provides an excellent opportunity to study the factors influencing the success of both invasive translocations and conservation focused translocations.

Although currently included within eastern rainbowfish, RRR differs in size and appearance, being smaller and differently patterned and coloured to eastern rainbowfish (Figure 1, Martin and Barclay 2016). Recent genetic work has determined that it is an undescribed species (Unmack 2016). RRR is naturally confined to a 13 km reach of Running River within the Burdekin River system (Figure 2, Martin and Barclay 2016; Unmack and Hammer 2015). Running River is divided into three reaches by two gorges, each with distinct fish faunas. Eastern rainbowfish is naturally found within the lowest reach, unable to spread further upstream due to barriers in the lower gorge. RRR is naturally found in the stretch between the two gorges, with no rainbowfish naturally found further upstream (Unmack and Hammer 2015).

In August 2015, eastern rainbowfish were detected upstream and within the habitat of RRR (eastern rainbowfish were not present in 2012 [Martin and Barclay 2016; Unmack and Hammer 2015]). Several wild fish were genotyped, confirming the two species were hybridising. By 2017, pure RRR was declining in the upstream portions of its range in Running River as the frequency of hybrids increased (Unmack, unpub. data). The underlying

6 cause(s) of hybridisation between the two species are currently unknown. Research on the closely-related western rainbowfish (M. australis) demonstrated that male size had a positive correlation with reproductive success (Evans et al. 2010; Young et al. 2010); eastern rainbowfish grows to a larger size than RRR and it’s possible that this may have driven hybridisation between the two.

Removing the main threat of hybridisation with eastern rainbowfish from the natural range of RRR was not feasible, and distinguishing hybrids from pure individuals was difficult without genetic analysis. As a result, the introduction of captive-bred fish from genotyped broodstock into areas free of other rainbowfish was the best option for conserving RRR. Fish stocked to establish secure populations needed to be genetically pure, so a breeding program at the University of Canberra was initiated using genotyped wild fish. It was important to release fish as close to their original distribution as possible, to reduce the potential of them becoming invasive, and to ensure that sites supported the habitat and hydrological conditions required by the species. Deception and Puzzle creeks, which flow into Running River were identified as the best potential release sites. Both these creeks already had a resident fish fauna, meaning that the potential impacts on invertebrates and frogs of introducing a new fish species was minimal. Purple spotted gudgeon (Mogurnda sp.) was found in both creeks, while spangled perch (Leiopotherapon unicolor) was found throughout Deception Creek and reaches below the release sites in Puzzle Creek. While both species have the potential to prey upon small fish, spangled perch grow to a larger size than purple spotted gudgeon and are more active hunters (Pusey et al. 2004). Therefore, predation pressure on small fish in Deception Creek was likely to be higher than that in Puzzle Creek and had the most potential to interfere with the introduction success.

The aims of this project Knowledge regarding the best methods to maximise the survival of captive-reared fish is particularly lacking for small-bodied Australian species. Conservation introductions of captive bred RRR into Deception Creek had the potential to be hampered by predation pressure from resident predatory fish species and thus provided an opportunity to examine the effect of predation on released fish. Previous laboratory trials and field experiments have shown that predator training; pre exposure to predators and conditioning to certain predator associated cues, may help increase the post-release survival of captive bred fish (Abudayah

7 and Mathis 2016; Brown 2003b; Hutchison et al. 2012; Vilhunen 2006). This thesis had two aims to: 1. investigate the presence of pre or postzygotic barriers to hybridisation between eastern rainbowfish and RRR 2. determine the effect of pre-exposure to predators on introduction success for RRR. The first aim of this study was investigated by pre-exposing half of the intended Deception Creek releases to spangled perch; the dominant predatory fish at the release site. It was hypothesised that pre-exposed fish would be better prepared for survival in the wild, resulting in increased introduction success. This was determined by assessing the abundance of populations immediately after release. Snorkel survey were used to monitor the relative abundance of spangled perch and RRR after release as it is an effective method for gathering relative abundance data within clear water habitats (Ebner et al. 2015; Fulton et al. 2012; O’Neal 2007).

Given that hybridisation with eastern rainbowfish is currently threatening at least four other Melanotaenia species (Lintermans 2016a), the mechanisms underlying it are worth investigation. The second aim of this study was investigated using dichotomous mate choice and breeding experiments. Prezygotic barriers to hybridisation often occur as preferences for specific visual or chemical cues. Dichotomous choice experiments, whereby an individual is given two options to choose from can be used to examine the mate preferences of fish and the factors that influence their choices (Wong and Jennions 2003; Wong et al. 2004a; Wong et al. 2004b; Young et al. 2010). As increasing body size has been linked to reproductive success in both sexes in other rainbowfish species (Pusey et al. 2004; Young et al. 2010) it was hypothesised that both sexes from both species would show a preference for larger fish, giving eastern rainbowfish a competitive advantage over RRR. To investigate the presence of postzygotic barriers, the reproductive success of hybrid pairings (eggs laid and fertilised, plus % survival to hatching) was examined using methods similar to Doupé et al. (2009). As hybrids have been detected in the wild, it was predicted that hybrid pairings would have equal or greater reproductive success than non-hybrid pairings.

This thesis is written as two stand-alone data chapters intended for publication, hence there is some repetition between chapters; preceded by a general introduction and followed by a synthesis. Chapter 2 investigates the presence of barriers to hybridisation between RRR and eastern rainbowfish. Specifically, it examines the influence of body size and species on mate

8 preference and the reproductive success of hybrid crosses compared to pure crosses. It has now been published in the journal Ethology. Chapter 3 reports on the results of conservation translocations of RRR to establish two separate refuge populations. Chapter 3 is intended to form part of a larger publication which contains the full account of the discovery and conservation of RRR.

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Figure 2 Map of the study area showing the location of Puzzle and Deception creeks and their positions relative to Running River and its gorges. Purple arrows indicate the range of Running River rainbowfish while orange arrows indicate the range of eastern rainbowfish (Melanotaenia splendida). Credit goes to Tani Cooper with permission provided by the Australian Wildlife Conservancy.

Chapter 2: Barriers to hybridisation and their conservation implications for a highly threatened Australian fish species

Declaration for Thesis Chapter 2 Declaration by candidate In the case of Chapter 2, the nature and extent of my contribution to the work was the following: Nature of Contribution Extent of Contributions (5)

I, Karl Moy, performed most of the lab 80 work, analysed the data, wrote the paper, prepared figures and tables, reviewed drafts of the paper, submitted it to the journal, made corrections, corresponded with the editor and wrote responses to the reviewers

The following co-authors contributed to the work: Name Nature of Contribution Contributor is also a student at UC Y/N Peter Unmack Assisted with collection of wild N specimens and reviewed drafts of the paper for submission Richard Duncan Assisted with statistical analysis N and figure production and reviewed drafts of the paper for submission Mark Lintermans Assisted with collection of wild Y specimens and reviewed drafts of the paper for submission Culum Brown Assisted apparatus and N experimental design and reviewed drafts of the paper for submission

Candidate’s Signature Date 20/12/2018 Declaration by co-authors The undersigned hereby certify that: (1) the above declaration correctly reflects the nature and extent of the candidate’s contribution to this work, and the nature of the contribution of each of the co-authors. (2) they meet the criteria for authorship in that they have participated in the conception, execution, or interpretation, of at least that part of the publication in their field of expertise; (3) they take public responsibility for their part of the publication, except for the responsible author who accepts overall responsibility for the publication; (4) there are no other authors of the publication according to these criteria; (5) potential conflicts of interest have been disclosed to (a) granting bodies, (b) the editor or publisher of journals or other publications, and (c) the head of the responsible academic unit; and (6) the original data are stored at the following location(s) and will be held for at least five years from the date indicated below:

23 Location(s): Institute for Applied Ecology, University of Canberra

Signatures Date

20/12/2018 20/12/2018

20/12/2018

[This is the peer reviewed version of the following article: Moy KG, Unmack PJ, Lintermans M, Duncan RP, Brown C. Barriers to hybridisation and their conservation implications for a highly threatened Australian fish species. Ethology. 2019;125:142–152. https://doi.org/10.1111/eth.12837, which has been published in final form at the listed DOI. This article may be used for non- commercial purposes in accordance with Wiley Terms and Conditions for Use of Self-Archived Versions.]

24 Running Title: Hybridisation barriers in a threatened fish

Karl G. Moy1*, Peter J. Unmack1, Mark Lintermans1, Richard P. Duncan1 and Culum Brown2

1 Institute for Applied Ecology, University of Canberra ACT 2617, Australia

2 Department of Biological Sciences, Macquarie University, Sydney NSW 2109, Australia

*Correspondence: [email protected]

Acknowledgements

The primary author would like to thank L. Tousetto for her help with animal husbandry.

Abstract Hybridisation and introgression are natural phenomena that may lead to the transfer of adaptive alleles from one species to another and increased species diversity. At the same time hybridisation and subsequent introgression threaten many species world-wide through the loss of genetic and species diversity. In Australia, introgressive hybridisation between native and alien species has not typically been considered a significant threat to native biodiversity because of the taxonomic distance between native and alien biota. However, many native fish have been introduced outside their natural range. Recently, four taxa in the genus Melanotaenia have been nationally listed as threatened due to introgressive hybridisation with introduced Melanotaenia splendida. We examined pre and post zygotic barriers to hybridisation between M. splendida and one of these threatened taxa – Running River rainbowfish (RRR) – to assess the potential for hybridisation to occur. We used dichotomous mate choice experiments to examine pre-zygotic barriers and mating experiments to examine post-zygotic barriers. Size was not a significant predictor of the proportion of time subjects spent with a potential mate, nor was there any significant difference in the amount of time subjects spent with potential mates of their own or the opposite species. Eggs from hybrid pairings with female RRR had a slightly higher hatching rate than those from hybrid pairings with female M. splendida, but neither were significantly different from intraspecies crosses. We could not identify any definite barriers to hybridisation, demonstrating that the

25 introduction of ‘native’ fish species outside their natural range poses a higher risk of hybridisation than previously thought. We call for better education around the consequences of moving ‘native’ fish and the development of rapid response plans to deal with recently established alien populations of Australian fish species in order to prevent future extinctions due to introgressive hybridisation.

Keywords: Invasive species, mate choice, introgression, behaviour, prezygotic, postzygotic.

Introduction Hybridisation and introgression are natural phenomena that may lead to the transfer of adaptive alleles from ones species to another and increased species diversity (Dowling and Secor 1997). However, hybridisation and subsequent introgression also threatens many species world-wide through the loss of genetic and species diversity (Huxel 1999; Seehausen et al. 1997). Hybrid offspring may possess competitive advantages over individuals from one or both parental species (Hammer et al. 2013) or they may possess deleterious traits via outbreeding depression. Hybridisation occurs when two individuals from separate species reproduce to make offspring (hybrids). Introgression occurs when hybrids backcross with one or both of the parental species resulting in the movement of genetic material from one species to another. Hybridisation and introgression are widespread phenomena, but are more common in fishes than other vertebrate groups (Hubbs 1955; Scribner et al. 2000; Verspoor and Hammart 1991). Typically, more distantly-related species are less likely to hybridise and introgress but barriers to hybridisation and introgression are rarely completely effective between closely- related species (Seehausen et al. 1997). Barriers can be prezygotic (behavioural, physical) (Boughman 2001; Rentz 1972) and postzygotic (hybrid inviability and inferiority) (Gilk et al. 2004; Rhymer and Simberloff 1996). These barriers are influenced by both biotic (organism density, behaviour, habitat modification) (Atsumi et al. 2018; Boughman 2001; Jones et al. 2006) and abiotic (turbidity, environmental chemical levels) (Fisher et al. 2006; Jones et al. 2008; Seehausen et al. 1997) factors. In certain situations, however, processes which can act as barriers, such as mate preference, may actually work in favour of hybridisation, whereby one or both species have a stronger preference to mate with the opposite species (Rosenfield and Kodric-Brown 2003). Reduced survivorship and reproductive viability are typical symptoms of outbreeding depression and may not occur until the second generation or later (Frankham et al. 2011; Gharrett et al. 1999; Jones et al. 2006; Rhymer and Simberloff 1996).

As a result, hybridization between native and alien species may lead to the decline or extinction of one or both parent species through genetic swamping, wasted reproductive effort or reduced survival of offspring (Rhymer and Simberloff 1996). Throughout Europe and North America, extinctions and declines of native fishes caused by hybridisation with alien species are well documented (Hitt et al. 2003; Ludwig et al. 2009; Meldgaard et al. 2007; Rosenfield and Kodric-Brown 2003). In Australia, introgressive hybridisation between native and alien species has not typically been considered a significant threat to native biodiversity because of the taxonomic distance between native and alien fishes (Harris 2013). However, recent broad-scale genetic studies of Australian freshwater fishes have revealed high levels of genetic structuring between populations as well as many new cryptic species (Hammer et al. 2013; Hammer et al. 2007; Unmack 2013; Unmack et al. 2013; Unmack and Dowling 2010). As a result, introgressive hybridisation caused by introductions of ‘native’ species outside their natural range or from one part of a species range to another, is a more recently recognised threat to the conservation for Australian freshwater fishes (Couch et al. 2016; Harris 2013; Lintermans et al. 2005). Keeping Australian fish in aquariums, ponds and stock troughs is becoming increasingly popular (e.g. Australia New Guinea Fishes Association 2018). Unfortunately, releases or escapes of alien ornamental fish has resulted in increasing numbers of these species becoming established in Australian waterways (García-Díaz et al. 2018). Once established, these species are often spread further by secondary human movement such as bait buckets or stocking (Lintermans 2004). Several native species have also been introduced within Australia due to stocking, stocking contamination and aquarium/pond escapes, though the latter is harder to identify (see Lintermans, 2004 & Martin, 2016) . As many fish species from Australia and New Guinea are kept in aquariums (e.g. Australia New Guinea Fishes Association, 2018), an investigation into the potential of aquarium species to hybridise with native fauna seems timely. Rainbowfishes are colourful, pelagic, small-bodied fishes in the family Melanotaeniidae with 110 species described (Eschmeyer and Fong 2017) plus several undescribed species (Unmack et al. 2013). They are distributed throughout much of New Guinea and northern and eastern Australia, inhabiting a wide range of habitats such as ephemeral desert streams, upland perennial creeks, swamps, rivers and lakes (Allen et al. 2002). Phylogenetic work on the group identified seven lineages, several of which can be sympatric (Unmack et al. 2013). Hybridisation and introgression were commonly found between species in these sympatric lineages. However, within a lineage, species were almost

27 never sympatric and, in some cases such as in the ‘Australis’ lineage, the different species are contiguously distributed but never overlapping (Unmack et al. 2013). This suggests that when these species come together they are unable to co-occur, likely as a result of a lack of mating barriers, although this hypothesis remains untested. As a result, introduction of rainbowfish outside their natural range can have serious negative implications for the conservation of other rainbowfish species within the same lineage. Rainbowfishes are frequently common where they occur, often being the most abundant species present. Most species reproduce throughout the year, with males competing amongst each other for access to females and spawning substrate (usually aquatic vegetation) while actively courting females. Because rainbowfishes are very colourful, easy to breed and keep in captivity they have become popular aquarium fishes such that they are now available in pet stores globally (Tappin 2011). Unfortunately, their popularity has also led to some species being introduced outside their natural range (Brown et al. 2012; Martin 2017; O'Donnell 1935). Despite the great potential for addressing questions around mate preference and sexual selection due to their bright and variable colouration and ease of captive care, only two studies have examined mate preference in rainbowfishes. Arnold (2000) examined the role of kinship in female mating preference, while Young et al. (2010) examined the role that male colour pattern and size played in female mate preference and male competitive performance of Melanotaenia australis. Although these studies were unable to find any relationship between female preference and colour pattern characteristics, they did detect a significant relationship between male size and female preference and competitive ability. Adding to this, Evans et al. (2010) found that females spawned more eggs when reproducing with larger males. These findings suggest that larger males have higher reproductive success than smaller males, though this finding has not been tested in other species. The Running River rainbowfish (Melanotaenia sp.; RRR) was until recently, included within the Melanotaenia splendida (eastern rainbowfish) complex, but recent genetic work determined that it is an undescribed species (Unmack 2016). Running River rainbowfish is naturally confined to a 13 km reach between two gorges of Running River within the Burdekin River system in Queensland, Australia (Martin and Barclay 2016; Unmack and Hammer 2015). Running River rainbowfish is one of four Melanotaenia species listed by the Australian Society for Fish Biology (ASFB) as either vulnerable, endangered, or critically endangered (Lintermans 2016b). Hybridisation with M. splendida is one of the key threats to the conservation of these species (Lintermans 2016a). In the case of RRR, hybridisation has

28 already occurred due to the introduction of M. splendida outside of its natural range (Unmack 2016; Unmack and Hammer 2015). Melanotaenia splendida is a generalist species, found in nearly all lowland freshwater habitats from the Adelaide River near Darwin in the Northern Territory to the Deepwater River near Gladstone in Queensland, as well as throughout the Lake Eyre Basin which encompasses a large portion of Australia’s dry interior (see G. Allen et al., 2002 or Tappin, 2011). One key ecologically significant difference is that M. splendida grows larger (up to 100 mm SL) relative to other narrow-range rainbowfishes (typically up to 65 mm SL, Pusey et al. 2004). Colouration, body depth and, to a lesser extent, fin size, varies across the range of M. splendida. In addition to M. splendida being generally larger than RRR, there are also obvious colour differences between the species (Martin and Barclay 2016). Running River rainbowfish differs in coloration and patterning, with three dark zig zagging lines that run along the bottom rear half of the body, a bright green mid-lateral stripe, solid red fins with blue rays and an overall golden hue compared to M. splendida’s lack of stripes, occasional dark mid lateral stripe, speckled fins and overall blue hue (in the alien population in upper Running River; Figure 3), although colouration and pattern varies considerably between localities. Melanotaenia splendida is naturally found below the lowermost gorge in Running River, with no rainbowfish naturally found further upstream above the upper gorge (Martin 2017). An alien population of M. splendida upstream of the habitat of RRR has spread downstream and has begun to hybridise with RRR (Unmack 2016). The alien population was first detected in 2015 (it was not present in 2012 [Martin and Barclay 2016]), and was thus fortuitously detected shortly after its establishment. By 2017 pure RRR was declining in the upstream portions of its range in Running River as the frequency of hybrids increased (Unmack, unpub. data). Given that hybridisation with M. splendida is listed as a threatening process for at least three Melanotaenia species (Lintermans 2016a), the mechanisms underlying this process are important to understand. Rarely are species introductions detected close to the time of their establishment; thus the introduction of M. splendida provides an ideal opportunity to examine the mechanisms around invasion success. While we expect that RRR will decline due to invasion by M. splendida and subsequent hybridisation, the mechanisms and likely impacts on RRR are unknown. A greater understanding of these processes should help determine the speed of impact and the possibility that RRR may be able to resist invasion in the longer term. Therefore, I aim to 1: determine whether mate preference (a prezygotic barrier) affects hybridisation between these two species and, 2: investigate potential differences in fertility

29 and egg survival (a postzygotic barrier) between these species and their hybrids. Based on the findings of Young et al. (2010) we hypothesised that both species would prefer larger mates. As hybrids have been detected in the wild, we also hypothesised that the proportion of fertile eggs produced, and the proportion that survive to hatching may be similar to pure parentage combinations.

Methods Maintenance Sixty-three Running River rainbowfish were collected from Running River in February 2016 and then maintained in a 150x60x60 cm aquarium. In February 2017, 35 females and 28 males were moved to two 120x45x45 cm aquaria and segregated and visually isolated by sex. Fish were sexed by eye as adults of both species are sexually dimorphic, with males possessing brighter colours and longer, pointed dorsal and anal fins. Seventy-one introduced M. splendida were collected from the upstream reach of Running River in August 2016 and housed in five 60x30x30 cm aquaria separated by sex. In February 2017, 31 female and 40 male M. splendida were moved to two 120x45x45 cm aquaria and segregated by sex. Before and after trials, fish were physically and visually isolated from the opposite sex. All fish were given a weekly 25% water change and fed a diet of flake food, frozen brine shrimp and frozen bloodworm. Under these conditions fish were in spawning condition and eager to mate.

Mate preference Treatments Dichotomous choice experiments were used to examine the factors influencing mate preference of both species. The first treatment allowed test subjects to choose between two individuals of their own species (intraspecies trials) allowing us to examine intraspecific mate-choice selection. Thirty male and 30 female M. splendida and 28 males and 32 female RRR were tested in intraspecies trials. The second treatment allowed test subjects to choose between one member of their own species and one member of the other species (interspecies trials), thus examining interspecific mate-choice selection. Thirty-one male and 30 female M. splendida and 28 male and 31 female RRR were tested in interspecies trials.

Apparatus The test apparatus consisted of a 120x45x45 cm glass aquarium which was divided into three compartments (S, A, B [Figure 4]). Compartments A and B were separated from each other by a white sheet of corflute, while both were separated from compartment S by a sheet of clear Perspex with 2 mm holes drilled into it. This prevented fish in compartments A and B from interacting with one another physically and visually but allowed them to visually and chemically interact with fish in compartment S. A spawning mop was placed in compartments A and B to provide cover and to encourage courtship behaviour, and an air stone was placed at the far end of compartment S. A 10 cm strip of corflute was placed on the substrate in front of compartments A and B to allow us to measure the time each test subject spent in association with fish in each compartment. The sides of the aquarium were covered in brown cardboard to minimise any outside influences on behaviour.

Protocol Mate selection trials took place from the 23rd of February 2017 to the 22nd of April 2017 between 0700 and 1600 hours. Trials were recorded in HD1080p using a Gopro camera positioned 30 cm above the centre of the fish tank. The proportion of time subjects spent associating with either of the option fish was later determined using BORIS (Friard and Gamba 2016) a program for behavioural research.

Each fish was used as a test subject for both treatments, and as a preference option for both treatments. To prevent re-testing of individuals within a treatment, fish that had been used in trials were housed separately from those that had not until all trials from the active treatment had been run. A minimum period of four days was allowed between each treatment to minimise the impacts of handling stress. Fish were physically and visually isolated from the opposite sex for one month before the first trials started to maximise sexual motivation and remained isolated outside of use in a trial. As there were fewer RRR males, four were used twice when testing female preferences.

One fish was randomly selected from the species and sex required for the treatment and placed into each compartment (A, B). The test fish - termed the ‘subject’ - was placed in the main compartment (S) and held within a clear, plastic container with the bottom cut off. The bottle had holes drilled in it allowing the subject to assess its surroundings using visual and chemical cues. The subject was left to acclimate in the container for 25 minutes, after which the container was raised using a pulley system allowing the subject to roam the main

31 compartment. Observations started immediately and lasted for at least fifteen minutes; in the event that the subject did not approach the option fish within the first seven minutes, filming was extended to thirty minutes. We recorded the first compartment that was approached, and the amount of time the subject spent within 10 cm (approx. 2 body lengths) of each compartment during the filming interval.

After each trial the subject was transferred to a bucket containing clove oil (0.04 mL/L). Fish remained supervised in the bucket until sedated (evident from loss of equilibrium). They were then placed on laminated 1 mm graph paper in a shallow tray and photographed before being transferred to an electronic scale and weighed. Summary statistics of this data can be found in Appendix C. The fish was then transferred to a bucket of clean water to recover before being placed in an aquarium housing fish that had been tested.

During intraspecific trials, this process was repeated for the fish from compartment A, with the fish in compartment B transferred to compartment A shortly after with a new fish added to compartment B. During double species trials, the fish from compartments A and B were used for two trials (one test subject of each species) before both being measured and replaced. A 15% water change was performed after each trial to dilute any impacts from pheromones and the apparatus was reset with a new subject added. ImageJ was used to measure the body length and depth of the photographed fish (Abràmoff et al. 2004), which then allowed for accurate size comparisons to be made.

Analysis As time spent with one individual is not independent of time spent with another, the association time for either option was calculated as the time spent with either option divided by the total time spent with both options (A/A+B). The presence of species based preferences was tested with a one-way analysis of variance (ANOVA) fitted in R 3.4 (R Development Core Team 2010), mapping the proportion of time each subject of each treatment spent with the option of its own species. There was a strong correlation between the measured body traits for all groups (Appendix A and D), so we used body length in all further analyses.

To quantify the difference in size between the two option fish, option A was treated as a reference fish, with size difference expressed as option A length minus option B length. To investigate the influence of body size on mate preference, we fitted a linear regression with

32 size difference (A–B) as the predictor and the proportion of time spent with option A (A/(A+B)) as the response variable.

Embryonic survival Treatments Four different crosses were compared based on all possible parentage combinations between RRR and M. splendida. Between 22–28 replicates were collected for each cross. Each replicate consisted of at least 20 eggs from a pair of fish. Three measures of reproductive success were examined: the proportion of eggs fertilised (fertility), the proportion of fertilised eggs that survived to hatching (hatching success), and the overall proportion of eggs that survived to hatching (total success).

Protocol Fish were housed in pairs of the desired cross with a synthetic yarn spawning mop which was checked daily for eggs. Eggs were collected from spawning mops and placed into a petri dish. Once a spawning mop had been harvested it was placed in boiling water for ten minutes. Any remaining eggs were removed and counted before the mop was re-used for spawning. A minimum of 20 eggs were collected for each replicate; and multiple spawnings were often conducted to obtain sufficient eggs. If a subsequent spawning produced 20 eggs they were all collected, raising the total number of eggs examined for that replicate.

Collected eggs were checked daily for development until all had hatched or died. Any eggs that did not show signs of development or ‘died’ within the first two days were counted as infertile. Death of an egg was evident when it became opaque and all dead/infertile eggs were removed upon identification to prevent contamination.

Analysis We modelled each measure of reproductive success as coming from a binomial distribution with a different mean for each parental cross and included a term to allow for over-dispersion among pairs within each parental cross. We tested for a significant effect of parentage on each measure of reproductive success using a likelihood ratio test (LRT), comparing models with parentage included to models without. These analyses were conducted using R (R Development Core Team 2010) and models were fitted in the lme4 package (Bates et al. 2015).

Results There was no evidence that subject fish showed either a size or species preference (Figure 5 & Figure 6). The relative size of an option fish did not successfully predict the proportion of time a subject spent with it for males and females of both M. splendida and RRR (Table 1). Similarly, there was no significant difference in the proportion of time a subject spent with an option of the same or different species for males and females of both species (Table 1).

Intraspecies pairings of both species had similar rates of fertility (78% and 79% RRR and M. splendida respectively, Figure 7). While male RRR – female M. splendida crosses had slightly lower fertility rates (66%), and male M. splendida – female RRR crosses slightly higher fertility rates (87%), the 95% confidence intervals overlapped (Figure 7), meaning observed differences were not significant. The survival of fertilised eggs was very similar for all tested combinations, meaning total success differed in the same way as fertility. Overall there were no significant differences between parentage crosses for fertility (df = 3, LRT = 7.35), survival of fertile eggs (df = 3, LRT = 0.27) or total success (df = 3, LRT = 7.39) with a chi-square test indicating p> 0.05 in all cases.

Discussion This is the first study to look at mate preference between two closely related fish species within Australia and the first to look at it between melanotaeniids. Mate recognition and preferences can play a major role in shaping species boundaries (Boughman 2001). In many cases individuals will base mate choices on predictors of fitness such as body size (Katvala and Kaitala 2001; Wong and Jennions 2003; Young et al. 2010). This can encourage hybridisation if preference for a certain trait overrides preference for a member of the same species (MacGregor et al. 2017; Wymann and Whiting 2003). Members of many closely related species are able to distinguish between members of their own species and those of a closely related species although this is not always the case. Individuals may prefer members of their own species (Boughman 2001), members of a related species (Rosenfield and Kodric- Brown 2003), or they may mate randomly (Becher and Gumm 2018; Garrett 1979). When members of a species show no preference toward mating with members of their own species it often leads to introgression between the two (Childs et al. 1996; Echelle and Echelle 1997; Garrett 1979; Rosenfield and Kodric-Brown 2003).

Contrary to expectation, body size and species identity did not play an important role in determining mate preference in either of the rainbowfish species examined here. As both species are sexually dimorphic, with males growing larger, having larger anal and dorsal fins and exhibiting more colourful patterns on their body and fins, it is reasonable to assume that females should exhibit sexual preference for larger, more colourful males. Males of both species have been observed sparring with one another and courting females whilst in an aquarium, which also suggests the presence of intrasexual selection. In many animal groups there is an established link between size and mate preference (Bateman et al. 2001; Blanckenhorn et al. 2000; Charlton et al. 2007; Cooper Jr and Vitt 1993). Additionally, the findings of Young et al. (2010) supported these assumptions for the closely related M. australis, as contrary to the present findings, they found a positive relationship between male size and female preference as well as male-male competitive ability.

The evidence suggests that the two species examined here, despite displaying sexual dimorphism, do not select mates based on physical characteristics. This is the case for Cyprinodon variegatus, in which females spawn randomly through the territories of males, giving males with a larger territory an advantage (Draud and Itzkowitz 2004; Itzkowitz 1978). Unlike Young et al. (2010) and Draud and Itzkowitz (2004) we did not examine the effect of intrasexual interactions between males on mate preference and male success. As there is a demonstrated link between body size and intrasexual competitive success across many animal groups (McElligott et al. 2001; Partridge et al. 1987; Schuett 1997) the effect of intrasexual interactions between males on male success should be examined in any future studies of rainbowfish mate preferences within the Running River hybrid zone.

Studies on hybridisation between other small-bodied fishes such as Cyprinidon rubrofluviatilis (Becher and Gumm 2018), Cyrpinodon pecoensis (Rosenfield and Kodric- Brown 2003), Cyprinodon variegatus (Becher and Gumm 2018; Rosenfield and Kodric- Brown 2003), and Gambusia nobilis (Swenton 2011) found results similar to those here, where one or both species showed no preference for or against conspecific mates. Unlike Rosenfield and Kodric-Brown (2003), we did not examine the preferences or reproductive success of F1 hybrid individuals; they found that parent species showed no preference for or against mating with F1 hybrids. If F1 hybrids possess hybrid vigour, this would accelerate the formation of a hybrid swarm (Rosenfield et al. 2004).

Similar to our behavioural experiment, which suggested an apparent lack of prezygotic reproductive barriers, our fertility and egg survival experiment did not reveal any clear post-zygotic barriers to hybridisation. Male RRR – female M. splendida crosses had slightly lower fertility and overall reproductive success than male M. splendida – female RRR crosses, but these differences were not significant and neither hybrid crosses were significantly different from pure crosses. However, examining the reproductive success of F1 and F2 hybrids may yield different results as the deleterious effects of outbreeding may not emerge within the first generation (Gharrett et al. 1999; Gilk et al. 2004).

Studies of mate preference between native and introduced species that have formed hybrid swarms such as Cyprinodon pecoensis x Cyprinodon variegatus (Childs et al. 1996) and Cyprinodon bovinus x Cyprinodon variegatus (Echelle and Echelle 1997) have found results similar to those here (Garrett 1979; Rosenfield and Kodric-Brown 2003). Introgressive hybridisation with alien species has not previously been identified as a threat to Australian freshwater fish species due to the distant phylogenetic relationship of most alien species (Harris 2013; Lintermans 2013). However, it is clear that introductions of native fishes, which can result in introgressive hybridisation, is a major threat to some species. This is the first study from Australia to confirm the vulnerability of a native fish species to hybridisation with an alien species using aquaria trials. If the long-term survival of RRR is to be predicted from what has been observed overseas in similar situations, then this is the first identified and documented case of an Australian fish facing extinction solely due to hybridisation. However, this may have already happened in other situations as a result of widespread fish stocking, river diversions and potential aquarium escapes that have occurred in parts of Australia, but remain largely undocumented. Recent broad-scale phylogenetic work may have identified situations where natural populations have hybridised or introgressed with alien species as a result of human-assisted movement. Examples of this include Hephaestus fuliginosus, a naturally widespread species which has been stocked within and outside its natural range (Pusey et al. 2016; Pusey et al. 2004), Glossamia aprion, a widespread species complex where there is evidence that a population of a unique species within the lower reaches of the Brisbane River is becoming introgressed as a result of introduced fish (Cook et al. 2017) and Macquaria ambigua in the Dawson River which is admixed with fish from the Murray- Darling Basin likely due to fish stocking (Beheregaray et al. 2017).

The two species examined here are not unlike M. australis in form or behaviour. It therefore seems unusual that the relationship identified by Young et al. (2010) was not

36 identified here. Our methods were similar to those of Young et al. (2010) and others (Arnold 2000; Wong et al. 2004) investigating dichotomous preferences of association, and contained a similar number of replicates. Young et al. (2010) compared male fish ranging in size from 40.3–62.8 mm; we compared a similar range of fish sizes from 37.8–62.8 mm with mean size differences of 3.85–7.15 mm between the option fish in each treatment (Appendix C) though Young et al. (2010) did not present the mean size difference between males used in their study. It remains unclear, therefore, as to why a relationship similar to that observed by Young et al. (2010) was not found.

We can only speculate as to why our results differed, but under typical experimental conditions males and females are unlikely to display their full natural nuptial colours. This would be the same for all studies involving rainbowfish as colouration is influenced by many factors including stress, turbidity, predators, lighting and background colouration. Replicating natural conditions in the laboratory is difficult and lighting is known to affect mate choice in birds (Evans et al. 2006) as well as fish (Milinski and Bakker 1990). Due to these conditions, subjects may have been unable to differentiate between species if they relied on visual cues. RRR and M. splendida are closely related – from the same lineage and although they have quite different colouration in nature, they looked relatively similar during experimental trials as compared to their typical full display colouration (Figure 3)

The extinction risk of range-restricted species such as RRR and other nationally listed Melanotaenia species is amplified simply because they are range-restricted. The risk of extinction for range-restricted species through introgressive hybridisation is further increased if they are closely related to a widespread, dispersal-oriented species; as is the case for the aforementioned species, with introgressive hybridisation with M. splendida listed as a key threating process for each. These widespread generalist species often become popular within the aquarium trade, often used as angling bait (Lintermans 2004), thus increasing their chances of being spread through bait bucket introductions and the release or escape from aquariums and outside ponds/small farm dams (Drake and Mandrak 2014; Lintermans 2004).

We have demonstrated a lack of significant barriers to introgressive hybridisation between two closely related species of Australian fish. This implies that introductions of native fishes, particularly those belonging to widespread genera are likely to result in the loss of genetic integrity of native populations. The number of alien species released into Australian waterways continues to increase with aquarium species contributing to the

37 increasing numbers (García-Díaz et al. 2018; Lintermans 2004; Whittington and Chong 2007). When combined with the increased availability and popularity of Australian/New Guinean species the potential for native species to introgress with alien species increases. Without improvements to community education and management capabilities such as those suggested in Lintermans (2004) the extinction of other fish species due to introgressive hybridisation with alien species will be inevitable.

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Intraspecies trials (regression) Test group df t value p value Male RRR 27 0.221 0.827 Female RRR 31 -0.330 0.744 Male M. splendida 29 -0.022 0.983 Female M. splendida 29 1.689 0.103 Interspecies trials (ANOVA) Test group df F statistic p value Male RRR 1, 25 0.074 0.788 Female RRR 1, 25 0.044 0.836 Male M. splendida 1, 25 0.415 0.525 Female M. splendida 1, 28 2.017 0.167

Figure 3 Images of male Running River rainbowfish (a*, c) and Melanotaenia splendida introduced into Running River (b*, d) displaying their natural colours in an aquarium (top) and the colours they displayed during mate preference trials (bottom).*Credit Gunther Schmida.

Figure 4 A sketch of the apparatus showing the layout. The dotted line represents the clear acrylic barrier, which restricts movement but not water or sight, while the solid lines represent solid barriers that restrict movement, water and sight. The dashed and diagonal lines represent a 10 cm strip of white corflute laid on the substrate which was used as the preference zone. The dashed circle represents the plastic bottle use to acclimate the test fish, the circle at the far left shows the air stone and the six-pointed stars represent spawning mops.

Figure 5 Distribution plot showing the proportion of time each test group spent associating with a potential mate of the same species during interspecific mate preference trials involving Melanotaenia splendida (M. splendida) and Running River rainbowfish (RRR). Mean and standard error from left to right: 0.586± 0.0622, 0.519 ± 0.0691, 0.393 ± 0.0840, and 0.438 ± 0.0662.

Figure 6 The proportion of time that a subject spent with an option fish as a function of its size relative to the other option in intraspecies trials separated by sex and species: (a) male Running River rainbowfish (RRR), (b) male Melanotaenia splendida, (c) female RRR, (d) female M. splendida..

Figure 7 The mean proportion of collected eggs which were fertile, mean proportion of fertilised eggs which survived to hatching, and mean proportion of collected eggs which resulted in hatching with 95% confidence intervals for all tested combinations involving Melanotaenia splendida (ER) and the Running River rainbowfish (RRR). The first individual in a cross represents the male.

Chapter 3: A test of the introduction of a predator-naïve and predator trained threatened small-bodied freshwater fish to a novel environment Abstract Conservation translocations utilising captive bred individuals for release into the wild have been widely adopted by conservation managers around the world. This strategy is commonly used for recovery programs focusing on fishes, due to the ease by which large numbers of individuals may be produced. However, captive reared fish are often poorly equipped for survival in the wild, suffering high mortality rates upon release. Previous research on improving post-release survival of captive reared fish has focused mainly on large-bodied species. Here, we examined the effect of pre-release predator exposure on the success of conservation introductions using hatchery-reared individuals for a critically endangered small-bodied freshwater fish, the Running River rainbowfish (Melanotaenia sp.). Before release, fish received repeated exposures to spangled perch (Leipotherapon unicolor) a novel predator present within the receiving habitat which is also present throughout the native range of RRR. The abundance of released fish was then monitored using visual surveys in the weeks after release and then several months later. The experiment was cut short by a large rainfall event which caused flooding and therefore connection of experimental pools, confounding any chance of detecting statistically significant differences. Observational evidence suggested behavioural differences between exposed and unexposed fish. Regardless, all but one release succeeded prior to flooding, suggesting that for rainbowfish, predation pressure may not be an overriding deciding factor of establishment success.

Introduction Human-mediated movement of a species outside its natural range for conservation purposes is referred to as ‘assisted colonisation’ or a ‘conservation introduction’ (IUCN 2013; Seddon et al. 2014). Conservation introductions differ from reintroductions because they do not take place within the historic range of the species (IUCN 2013; Seddon et al. 2014). They can be an effective tool for conservation and management, especially when cause of decline is unknown or cannot be removed from the species’ natural range (Chilcott et al. 2013; Field et al. 2007; IUCN 2013; Pusey et al. 2003). Because they require a species to colonise a new area, conservation introductions are essentially biological invasions (e.g. alien species). Like other invasions, for conservation introductions to succeed a species must enter a novel environment, establish a self-sustaining population, and then disperse to occupy areas outside

51 the initial release site (Blackburn et al. 2011; Seddon et al. 2014). However, environmental pressures (e.g. droughts, habitat availability) and biotic pressures (propagule size, predation and competitors) may reduce the chance of establishment (Blackburn et al. 2011; Hayes and Barry 2008; Prins and Gordon 2014). Conservation introductions may differ from other invasions in the traits possessed by the invading species. For example, many invasive freshwater fish species tend to have generalist dietary requirements (Harris 2013; Marchetti et al. 2004) and fill ecological niches underutilised by native species (Harris 2013; Olden et al. 2006; Vila-Gispert et al. 2005). On the other hand, threatened freshwater fish species are more likely to be small- or very large- bodied with specific habitat requirements (Kopf et al. 2017; Olden et al. 2007). As a result, species used in conservation introductions may not possess traits that would make them successful invaders of the receiving habitat. Additionally, many conservation introductions utilise captive-reared individuals which may be behaviourally deficient (Araki et al. 2007; Philippart 1995; Snyder et al. 1996), thereby reducing the likelihood of a successful invasion. Freshwater fishes are one of the most threatened vertebrate groups worldwide (IUCN 2018). Certain species are also among the most invasive aquatic species, with some achieving a near global distribution, while being implicated in the extinction and decline of other aquatic species (Cambray 2000; Francis 2012). In freshwater fishes, habitat suitability seems to play the largest role in determining the success of an introduction/invasion, while biotic resistance (resident competitors/predators) seems to have very little effect (Harris 2013; Moyle and Light 1996b). Translocation has become a key tool for conserving freshwater fishes, utilising both wild and captive-bred fishes (Lintermans 2013; Lintermans et al. 2015; Minckley 1995). In Australia most early conservation translocations were conducted with larger threatened species that were potential angling targets (Lintermans et al. 2015) such as trout cod (Maccullochela macquariensis, Cadwallader and Gooley 1984), eastern cod (Maccullochella ikei, Nock et al. 2011), Mary River cod (Maccullochella mariensis, Simpson and Jackson 2000) and Macquarie perch (Macquaria australasica, Cadwallader 1981). The practice has recently been applied to small-bodied (< 20cm) threatened native fishes in Australia (Attard et al. 2016; Hammer et al. 2013; Lintermans et al. 2015). The continued existence of Pedder galaxias (Galaxias pedderensis) is solely due to conservation introductions (Chilcott et al. 2013), while the conservation status of several critically endangered species such as red-fin blue-eye (Scaturiginichthys vermeilipinnis Kerezsy and Fensham 2013) and several threatened galaxiid species (Ayres et al. 2009; 2006) have massively benefited from translocations. A review of factors influencing the success of

52 freshwater fish reintroductions reported that second to addressing the cause of initial decline, habitat related factors were the greatest predictors of reintroduction success (Cochran- Biederman et al. 2015). The importance of suitable habitat in determining the success or failure of an introduction is also echoed by studies of invasive fish species (Harris 2013; Moyle and Light 1996a; Moyle and Light 1996b) which found that if the habitat characteristics of the receiving environment were suitable then an invasion was likely to succeed regardless of other factors. That an introduction will fail in the absence of suitable habitat seems rather straightforward, as habitat loss and degradation is one of the greatest threats to freshwater environments (Walker and Steffen 1997); however, some reintroductions may fail even in the presence of adequate habitat (Barlow et al. 1987; Leggett and Merrick 1997). Cochran-Biederman et al. (2015) also noted that 71% of all failed releases utilised hatchery-reared fish. Captive-rearing, usually in the form of hatcheries, has been used to conserve fish populations around the world, because it is capable of producing large numbers of fish at a low cost, in a controlled environment. However, captive-reared fish are often raised under conditions which are vastly different from the environment they are released into (Brown et al. 2003). Consequently, hatchery-reared fish often exhibit behaviours that are detrimental to their survival in the wild and as a result often suffer high mortality rates once released (Brown and Day 2002; Ebner et al. 2007; Sparrevohn and Støttrup 2007). The behavioural impacts of hatchery rearing have been known for some time (Brown and Day 2002), with hatchery- reared fish showing deficiencies in key behaviours such as predator recognition and avoidance (Ebner et al. 2007; Hutchison et al. 2012) and foraging skills (Brown et al. 2003). Brown et al. (2003a) showed that environmental enrichment (e.g. adding artificial plants) and exposure to live foods resulted in fish being better able to utilise novel prey items. Meanwhile, several studies have shown that repeated exposure to predators, or their stimuli, will improve the predator avoidance behaviours of captive-bred fish (Abudayah and Mathis 2016; Brown 2003a; Vilhunen 2006). As a result, research and implementation of environmental enrichment and predator training of captive-reared fish is becoming more commonplace in Australia (Hammer et al. 2012; Hutchison et al. 2012; Lintermans et al. 2015). Most research investigating methods to improve the survival of hatchery-reared fishes has taken place overseas, although some recent research has been conducted in Australia (Hutchison et al. 2012). In all cases, research investigating introduction success has focused almost entirely on large-bodied, predatory, recreationally important species such as

53 brown and rainbow trouts (Alvarez and Nicieza 2003; Brockmark et al. 2010; Brown and Smith 1998) or percichthyids (Ebner and Thiem 2009; Ebner et al. 2007; Hutchison et al. 2012). However, of the 17 Australian species used in conservation introductions or reintroductions documented by Lintermans et al. (2015), 10 were small-bodied. Small-bodied species usually have vastly different requirements to large-bodied species (Allen et al. 2002), and a technique that works well for large species may not be as effective for small-bodied species. As small-bodied freshwater fish have a higher risk of extinction (Kopf et al. 2017; Olden et al. 2007; Reynolds et al. 2005), there is a need for better understanding factors influencing, and methods for improving, the survival of captive-bred small-bodied freshwater fish once released. Unlike studies investigating the role of biotic resistance on freshwater fish invasions (Moyle and Light 1996a; Moyle and Light 1996b), studies on conservation introduction success of freshwater fishes within Australia (Brown et al. 2012; Ebner et al. 2009; Ebner et al. 2007) and abroad (Alvarez and Nicieza 2003) suggest that biotic resistance is likely to play a role in translocation success. Several laboratory experiments have shown that pre-exposure to predators may have the potential to increase the survival of released fish (Abudayah and Mathis 2016; Brown 2003b; Hutchison et al. 2012; Vilhunen 2006). Here we aim to determine the effect of predator training on translocation success of captive bred Running River rainbowfish (RRR, Melanotaenia sp.) – a small bodied freshwater species threatened by extinction.

Methods Study species RRR is a recently identified and undescribed species which was naturally confined to a 13 km reach between two gorges of Running River within the Burdekin River system (Martin and Barclay 2016; Unmack and Hammer 2015) and currently listed by the Australian Society for Fish Biology as critically endangered (Lintermans 2016b). Recent unpublished work using SNP markers has determined that it is an undescribed species (Unmack unpub. data). RRR was previously classified as eastern rainbowfish, however, RRR differs by being smaller and is differently patterned and coloured (Martin and Barclay 2016). The key threatening process is hybridisation with an alien population of eastern rainbowfish (Melanotaenia splendida) which has established above the upstream limit of RRR’s natural range (Lintermans 2016a). The eradication of eastern rainbowfish and hybrids with RRR was not a viable conservation

54 option, therefore translocations into isolated reaches nearby were considered the best option for long term conservation of the species. Study area Running River is a tributary of the Burdekin River catchment in far north Queensland Australia which flows off the western slope of the Great Dividing Range (Figure 2). Deception Creek, which flows into Running River just below the lowermost gorge, and Puzzle Creek which flows into Running River just above the uppermost gorge were identified as the best potential translocation sites (Figure 2). Both creeks contained barriers to upstream dispersal by rainbowfish and already had resident fish fauna, meaning that the potential impacts on invertebrates and frogs of introducing a new fish species was less. Throughout most of the year, Deception Creek consists of several disconnected pools without flow, whereas Puzzle Creek flows for most of the year reducing to disconnected pools during periods of low rainfall. Purple spotted gudgeon (Mogurnda sp.) was found in both creeks, and spangled perch (Leiopotherapon unicolor) was abundant throughout Deception Creek and reaches below the release sites in Puzzle Creek. Although both species have the potential to prey upon small fish, spangled perch grows to a much larger size than purple spotted gudgeon and are more active hunters (Pusey et al. 2004). Therefore, as predation pressure on small fish in Deception Creek was likely to be higher than that in Puzzle Creek, releases into Deception Creek were used to assess the effect of predator training on translocation success.

Breeding and Rearing In 2015, 52 RRR with an equal sex ratio were collected from Running River and transported to a glasshouse at the University of Canberra which had heating and cooling systems to avoid extreme fluctuations. Broodstock were genotyped and compared to wild fish that had been collected in 1997 (18 years earlier) to ensure genetic purity. These fish were set up as 26 breeding pairs and used as brood stock for Deception Creek releases. Subsequently, an additional 32 wild fish were genotyped and added as brood stock for Puzzle Creek releases in February 2016, and fish were spawned in 17 groups of 2 males and 2 females. Some breeding groups had extra individuals added such that half the breeding groups consisted of 5, 6 or 7 individuals. Approximately 6500 fish were produced, 2500 in the first round for Deception Creek, and 4000 in the second round for Puzzle Creek within a temperature regulated glasshouse at the University of Canberra.

Eggs were collected on synthetic wool mops placed into the breeding tanks. After two days of spawning, mops were transferred to small fish tanks (40x20x20 cm) and juvenile fish were raised for ~2 months, then transferred to larger tanks (91x35x45 cm). Breeding and rearing tanks had painted sides and bottom and sponge filters, but were completely unfurnished in any other way. Larvae were started on a diet of live vinegar eels (Turbatrix aceti) and as they grew larger, moved onto a diet of baby brine shrimp (Artemia) in about a week, plus commercial flake food.

Once large enough for transport, fish were air freighted to James Cook University (JCU) Townsville and distributed evenly into ten outdoor ponds (130 cm in diameter and 30 cm deep) to grow out. At JCU, fish were fed commercially available flake food three times a day and a mixture of frozen brine shrimp and blood worms (Chironomidae) once a day. Rearing ponds all contained several large river stones and plastic mesh 50x100 cm with 2.5 cm holes, which was contorted into different shapes and added to provide cover. This was to encourage natural behaviours, such as using cover to escape threats, establishing and holding territories, and foraging, which have previously been found to result in improved survival rates (Brown et al. 2003; Roberts et al. 2014). Although there were differences in the shape and size of the rocks, all the ponds were arranged in a similar pattern.

Training A live adult spangled perch approximately 15 cm in length was placed in a 25x25 cm “box” made from plastic 2.5 cm mesh within the rearing ponds of trained fish. RRR were able to swim freely in and out of the mesh box. In addition to providing the predator, we also provided a cutaneous alarm cue, which is often released when a fish’s skin is damaged and can be used in associative learning (Abudayah and Mathis 2016; Brown 2003b; Brown and Chivers 2007). We euthanised one RRR (with an overdose of clove oil) per week of training, crushed it up, mixed it with water and sieved out larger fragments. This solution was then frozen in an ice cube tray and one cube was added at the same time as the spangled perch, in the hope that juvenile RRR would associate the olfactory cue of dead/injured conspecifics with the stimulus of a spangled perch. The spangled perch was left in the rearing pond for 15 minutes a day for 7 days immediately prior to fish being released into the wild.

Habitat Assessment at release sites Release sites (pools) within Deception Creek were paired based on similarities between habitat variables in an attempt to isolate the effects of predator training, with one site randomly selected to receive trained fish while the other received untrained fish. Puzzle creek

56 release sites were also assessed, but due to the number of accessible pools, habitat assessments were only used to identify suitable release sites. The variables examined were pool length, average pool width, substrate composition, average depth, deepest point, and riparian cover. Pool length was measured from the uppermost water’s edge to the most downstream water’s edge. Average pool width was calculated by taking three measurements at 25%, 50% and 75% of the total length of the pool. A transect consisting of 5 sample points was taken along each width measurement at 0% (+25 cm), 25%, 50%, 75% and 100% (-25 cm) of channel width. At each sample point, depth, substrate composition, macrophyte cover, and leaf litter were measured. Riparian cover was estimated by eye to the nearest 5% while depth was measured using a 120 cm metal ruler; where this was not enough depth was measured using a rock tied to a rope, which was then measured against the tape measure. All other variables were measured using a 50x50 cm quadrat.

Deception Creek releases Ten releases, of 250 fish per site were made into Deception Creek between the 2nd of November 2016 and 13th of January 2017. This time of year was chosen because, based on what was known about other Melanotaenia species, it was thought that conditions would be best for growth and reproduction (Beumer 1979; Milton and Arthington 1984). Releases were made in groups of 250 so that there would be five replicates of each treatment (trained, untrained), as grow-out facilities could only facilitate 2500 fish. At release, fish were 3 cm long on average but varied from 2 – 5 cm in length. Pools were paired based on similar habitat characteristics and were matched as best as possible based on size, riparian cover and substrate, in that order; with one pool randomly assigned to trained or untrained fish before releases took place. A release was made into Deception Creek once every week or so, however, there was no assigned order due to the logistical constraints regarding predator avoidance training. Fish were transported from rearing ponds at JCU to their release sites in 20 L plastic buckets. Buckets were filled to 1/3 and water was dosed with sea salt at 2.6 g/L and API stress coat dosed at 4 mL/5 L with 25 fish per bucket. Fish were delivered to their release site on the same day as collection in all but one case, which was hampered by heavy rainfall. In this instance, fish were held in buckets for two days with a daily water change, before being delivered to their release site. Fish were given a soft release; that is they were held at the release site overnight in a holding net made from shade cloth and pvc pipe with dimensions of 1x1x1 m. This allowed fish to acclimate to the water conditions without

57 predation pressure. The following day the holding net was up-ended and the fish were released into the pool.

After release, snorkel surveys were used to estimate the abundance of spangled perch and RRR in each pool; a suitable method in clear water (Ebner et al. 2015; Fulton et al. 2012; O’Neal 2007). Due to logistical constraints; such as the remoteness of the site, weather, and the need to carry out other releases, surveys occurred somewhat opportunistically. However, at least one survey was undertaken in the first week following release and was often followed by others up to fifty-six days after release. A total of 41 surveys across five untrained and two trained release sites were made between 2nd November 2016 and 5th January 2017. Snorkel surveys consisted of three passes: down the left bank, up the right bank, and a final pass down the pool centre. During passes along the edge of the pool, the researcher swam as far from the bank as possible while still being able to clearly identify fish (approximately 2.5 metres) only recording what was between them and the bank, whereas during the final pass they recorded what was present 3m either side of them, unless that area had already been covered by a previous pass of the bank. The researcher kept a steady pace; to prevent double counts of fish, and recorded on a waterproof notepad a tally of the total number seen as well as the maximum seen at one time, with a separate count for larvae. Spangled perch were also recorded in this way to estimate predator density. A large rainfall event (over 200 ml across four days at the nearest rainfall gauge) occurred in early January 2017, which restored flow to the channel and reconnected release pools, before the experiment in Deception Creek could be completed. This prevented any survey data being collected for the final three releases which were all of trained fish. Follow-up surveys were undertaken for all sites in Deception Creek in May and October 2017.

After the first field season, the extent of fish occurrence throughout each of the drainages was mapped by walking along the creek and stopping at each pool for five minutes to observe the presence or absence of rainbowfish. If no rainbowfish were observed within five minutes, the researcher moved to a different region of the pool and continued to observe for a further five minutes. If no rainbowfish were observed, the next pool downstream/upstream was also checked. This was repeated until three pools in a row were found without rainbowfish. This was carried out for Deception Creek in May and October 2017 and May 2018. An attempt was made to map the extent of RRR in Deception Creek after the large rainfall event in early January 2017, following the same protocol above but was hampered by low visibility as a result of increased turbidity.

Puzzle Creek releases Four releases, each consisting of 375 untrained fish were made into Puzzle Creek in May 2017 in the same manner as those made into Deception Creek. Although fish released into Puzzle Creek were the same size as those released into Deception Creek, only 1500 of the originally intended 4000 fish were released as a result of attrition in the growing out ponds. Due to funding and weather constraints on field work, no monitoring was undertaken in the weeks immediately after release for the Puzzle Creek releases. Planned monitoring of Puzzle Creek in October 2017 was prevented by a large rainfall event, but a survey of all release sites and distribution mapping following the same protocol above took place in May 2018.

Analysis Measures of abundance were used for analysing data collected in the first field season (up to a month from release), whereas measures of density (fish per metre of stream length) were used for analysing data collected in subsequent field seasons. Measures of density were calculated by dividing abundance by the length of the pool. This was done because even though pools differed in size, all pools received the same number of fish, meaning that in the weeks after release, density would represent differences in survival appropriately. Density was used for analysis of data collected later, because after that time the population had grown enough that density was a more accurate representation. Welch’s t-tests were used to test for differences in RRR abundance between trained and untrained fish in the 2-3 weeks following release, whereas paired t-tests were then used to compare the density of trained and untrained fish at release sites in Deception Creek. Linear regression analysis was used to investigate the effect of predator density on the abundance of RRR in the first month after release, and the density of RRR 6 and 11 months after release. Linear regression analysis was also used to examine trends in abundance within the first month after release. All analysis was performed in R (R Development Core Team 2010) using the R commander package (Fox and Bouchet- Valat 2019).

Results Habitat In October 2016, release sites in Deception Creek varied between 100 and 280 m long and 8 to 14 m wide. The average depth varied between 42 and 113 cm, while the deepest points ranged from 1.65 – 3 m. Riparian cover ranged from 60 – 99%. Substrate was dominated by sand (68%), followed by boulder (13%), bedrock (11%) and cobble (7%). On average, aquatic plants (macrophytes and charophytes) covered approximately 40% of the substrate, and leaf litter covered approximately 25%.

Release sites within Puzzle Creek were between 150 and 265 m long and between 9.9 and 22.4 m wide, with an average depth between 84 and 125 cm. Riparian cover varied between 95 and 80%, and substrate was dominated by sand (51%) followed by bedrock (25%), cobble (20%), and boulder (4%). On average, aquatic macrophytes and charophytes covered 20% of the substrate, and leaf litter covered 20%.

Predator effects There was no significant difference in abundance or density of adult fish between trained and untrained release sites at any point after release (Table 2). Of the seven releases in Deception Creek before flooding, fish failed to establish at only one site following the release of untrained fish. This site was surveyed five times from 2–31 days after release without a single RRR observed, and was otherwise unremarkable compared to other sites. At the remaining sites, the abundance of released fish appeared to decline continuously over time for both treatments at sites where samples were collected over longer time scales (Figure 8), however, linear regression analysis did not provide statistical support for this (t=0.27, p=0.788); though this could have been due to low detection power caused by small sample sizes and variation in detectability. The abundance of RRR appeared to increase at some sites in the weeks immediately after release (Figure 8), however, this is not possible as enough time for reproduction of the next generation had not elapsed and therefore these perceived increases were due to variation in detectability.

Linear regression analysis found no significant link between predator density and rainbowfish abundance or density for any survey season (Table 3). This was the case even when the analysis was broken up into different size classes for both rainbowfish and spangled perch. While these results were not statistically significant, there was a positive correlation between adult rainbowfish density and the density of all spangled perch (Appendix B).

Rainbowfish larvae were detected within the first field season at four sites (two trained, two untrained) 30–40 days after release. In May 2017, juveniles and adults too small to have been released fish were detected at all sites. When the total density of rainbowfish – including fry and juveniles was compared, untrained release sites were significantly higher than trained release sites at six months, but at no other time (Table 2). No significant difference in rainbowfish density was found between releases that took place before or after the flooding occurred during May (t = -1.91, P = 0.09) and October (t = 0.557, P = 0.59) surveys.

Dispersal When flooding occurred in Deception Creek, fish moved between release sites, which made comparisons between treatment pools invalid. Ten days after flooding in Deception Creek, one individual was recorded in an ephemeral gully stream 660 m upstream and approximately 24 m higher in elevation than the nearest release site. The movements of fish from their upper and lower-most release sites in both systems are summarised in Table 4. The population in Puzzle Creek spread downstream at approximately half the rate of the population in Deception Creek, but spread upstream at approximately one fifth of the rate of the Deception Creek population (Table 4). In one year, RRR from Puzzle Creek dispersed a total of 460 m upstream, 200 m less than the distance covered by a fish from Deception Creek in 10 days. In Deception Creek there was a dramatic increase in the distance spread downstream between October 2017 and April 2018 in Deception Creek (Table 4). The actual downstream limit in Deception Creek in April 2018 could not be determined, due to time constraints and limited access into that portion of the creek.

Discussion Although impacted by flood conditions, the data presented here do not support the hypothesis that pre-exposure to predators or predation pressure increased introduction success in RRR. While the only unsuccessful release was of untrained fish, all other releases of untrained fish were successful, suggesting that predator naïve-fish are still capable of establishing. However, sample sizes for this experiment were small enough that only strong differences would have been detected. It is unclear whether the fish that were released after flow was restored survived to reproduction. There was no significant differences in rainbowfish density between sites that had been released before or after flow was restored. This was the same even for the failed release site, meaning that dispersal and colonisation were enough to result in RRR densities resembling that of release sites in areas that had failed or not been stocked. Given that rainbowfish are known to utilise social learning (Brown and

Warburton 1999), and that predator experienced fish from other releases dispersed to post flood release sites, it is likely that post flood releases were less impacted by predation encounters than before flood releases. Introductions into Puzzle Creek were made during a high flow event and still established a sustaining population, so it is likely that post flood releases in Deception Creek survived to reproduction. Few released fish, if any, were present at release sites six months later as most fish were not large enough (i.e. released fish were 2- 5cm and were expected to be around 7-10 cm following another 6 months of growth). Thus it was likely that most fish observed were spawned in the wild. Therefore, we would not expect any differences in rainbowfish density at 6 or 11 months. As the rainfall, flow regime, habitat, vegetation, and resident fish biota of Puzzle Creek were different to that of Deception Creek and it was only surveyed once, we are limited in the information we can glean from this introduction. We can, however, state that predation and competition with purple spotted gudgeon and flooding during introduction did not prevent RRR establishing.

While there is a paucity of information regarding the movements of Australian small- bodied freshwater fishes, studies on ephemeral waterholes (Kerezsy et al. 2013) and genetics (Unmack et al. 2013) suggest that some of these species are capable of dispersing great distances. The study of dispersal in small-bodied fishes has often been hampered by their size, however, these releases provided a unique opportunity regarding the rate at which rainbowfishes may spread throughout a previously unoccupied waterway. Puzzle Creek flows more frequently than Deception Creek, so it is reasonable to assume that RRR would spread through Puzzle Creek at a faster rate. Contrary to what might have been expected, RRR dispersed throughout Deception Creek faster than Puzzle Creek. One possible explanation is that because Puzzle Creek was experiencing high flow at the time of release, many individuals may have dispersed but perished in the process. It’s almost certain that many fish died during dispersal in Deception Creek as well, with many individuals observed occupying habitats that would not persist long term in the absence of rain (e.g. the individuals observed within the ephemeral gully). In Deception Creek, however, opportunities to disperse were less frequent and initially limited, restricting released fish to their release sites where they increased in population size, thereby increasing the success of subsequent dispersal. The site fidelity of translocated individuals is consistently lower than that of wild individuals across most faunal groups (Clarke and Schedvin 1997; Tuberville et al. 2005) including fish (Ebner and Thiem 2009). Immediate dispersal from the point of release may increase the likelihood of translocation failure, for example as individuals may disperse to sub optimum habitats or

62 encounter predators in unfamiliar environments. In some instances, ‘penning’, whereby translocated organisms are kept in pens at the release site for several days or weeks before being allowed to roam free, has been an effective method of increasing site fidelity and overall success of establishment (Tuberville et al. 2005). It is possible that during periods of low flow disconnected pools acted in a similar fashion, forcing fish to develop some site fidelity with their new habitat and allowing them to increase in number, thereby increasing the number of fish that dispersed when it became possible to do so.

Observations in Deception Creek made in the hour and days immediately after release suggest there may have been some behavioural differences between trained and untrained fish. In both pre-flood releases, trained fish shoaled and found a shallow, sandy area out of the reach of larger spangled perch close to the point of release and remained there for around six days before venturing out. In contrast, untrained fish were often observed swimming near the surface in open water and swimming toward the spangled perch, which were trying to eat them, before eventually finding shallow areas to hide. One reason predation may not have been found to have a significant impact, may have been that neither spangled perch or purple spotted gudgeon are primarily piscivorous (Pusey et al. 2004). If a more specialised piscivore such as mouth almighty (Glossamia aprion) was found within the study reach, we may have found different results. Mouth almighty has been implicated in the extirpation of Lake Eacham rainbowfish (M. eachamensis) from Lake Eacham (Barlow et al. 1987), and it is not unreasonable that a similarly proficient piscivore could negatively impact an introduction of small bodied fish if they did not possess the ability to recognise or escape predators (Brown and Warburton 1997).

These releases were an unusual success for Australian freshwater fish conservation and may be attributable to traits of the species itself, as rainbowfishes are mostly generalists (Allen et al. 2002). Eggs were detected within the holding pen at some sites before fish were released, indicating that conditions were good for reproduction. The ability to reproduce immediately after release is likely an important distinction between small and large-bodied species to consider when performing conservation releases. Larger-bodied species, such as Macquarie perch (Macquaria australasica) take four years to reach maturity (Appleford et al. 1998). In contrast, many small-bodied species can reproduce within their first year of life, meaning that released fish could begin breeding immediately even if they are unable to avoid predation or forage effectively, etc. As their offspring develop in the wild, they are unlikely to

63 suffer the same deficiencies as captive-reared fish such as predator naivety (Alvarez and Nicieza 2003).

Our sample sizes were too small to conclusively say whether predator training had an effect on post release survival. Based on the data at hand it appeared to make no difference, although anecdotal evidence from observations immediately after release suggest there may have been some behavioural effects of training. Future research on the effect of predation on releases of small-bodied species should focus on both laboratory and field trials while also utilising multiple kinds of predators. Given its low cost and support from laboratory-based experiments (Hutchison et al. 2012; Vilhunen 2006) we recommend the continued implementation of predator training in release programs.

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Tables and Figures

Table 2 Statistical output comparing trained and untrained releases of fish. Statistical output comparing trained and untrained releases of fish. A Welch’s t-test (W) compared total observed abundance at 2-3 weeks from release, while a paired t-test (P) compared the density of adults at 6 and 11 months from release. Standard error (SE) was calculated from 11 abundance observations between two sites which all fell within 4 days of one another converted to a percentage and then applied to all samples.

Adults t-test Trained ± SE Untrained ± SE T P df 2-3 weeks W 85.7 ± 18.85 38.25 ± 8.42 -1.60 0.260 1.85 6 months P 1.3 ± 0.30 1.86 ± 0.41 -2.14 0.100 4 11 months P 2.0 ± 0.45 1.58 ± 0.35 -0.65 0.554 4 Juveniles 6 months P 0.2 ± 0.04 0.36 ± 0.07 -1.22 0.291 4 11 months P 1.3 ± 0.28 1.06 ± 0.21 -0.67 0.542 4 All Rainbowfish 6 months P 1.5 ± 0.44 2.22 ± 0.49 -2.88 0.045 4 11 months P 3.1 ± 0.62 2.97 ± 0.59 0.11 0.920 4

Table 3 Statistical output from linear regression analysis testing predator density as a predictor of rainbowfish abundance in the first month, and density 6 and 11 months from release.

T P R df 2-3 weeks 0.527 0.621 -0.137 5 6 months -0.014 0.989 -0.125 8 11 months 1.545 0.161 0.133 8

Distance (km) (elevation [m]) Months since release Upstream Downstream Deception Creek May 2017 6 1.9 (31) 1.3 (46) Deception Creek October 2017 11 2.4 (39) 2.7 (62) Deception Creek April 2018 17 2.5 (41) >6.3 (>171) Puzzle Creek May 2018 12 0.46 (9) 1.33 (30)

First Field season Untrained 140

120

100

60 TotalAbundance 40

0 0 5 10 15 20 25 30 35 40 45 Days from release

First Field Season Trained 250

200

150

100 TotalAbundance

0 0 10 20 30 40 50 60 Days from release

Figure 8 Abundance of released fish over time during the first field season in Deception Creek for trained fish and untrained fish. Different colours represent different release sites. Note, the number of released fish cannot increase as fish were only released once into each site.

Chapter 4 Synthesis Thesis overview This thesis aimed to investigate the factors influencing invasion success in rainbowfishes through two avenues. Firstly, by investigating the presence of pre and post zygotic barriers to hybridisation between eastern rainbowfish (Melanotaenia splendida) and Running River rainbowfish (RRR, Melanotaenia sp.). Secondly it looked at the effect of pre- exposure to predators on the introduction success of captive bred RRR in situ. There was also a unique opportunity to gain in situ information on dispersal rates for RRR, providing a rare account of the dispersal abilities of a rainbowfish.

Hybridisation and introgression risk Introgressive hybridisation with eastern rainbowfish is currently threatening RRR and at least four other Melanotaenia species (Lintermans 2016). In an attempt to better understand the underlying processes surrounding introgressive hybridisation between RRR and eastern rainbowfish, this study examined mate choice between the two species using dichotomous mate choice experiments. Mate recognition and preferences can play a major role in shaping species boundaries (Boughman 2001). In many cases, individuals will base mate choices on predictors of fitness, such as body size (Katvala and Kaitala 2001; Wong and Jennions 2003; Young et al. 2010). This can encourage hybridisation if preference for a certain trait overrides preference for a member of the same species (MacGregor et al. 2017; Wymann and Whiting 2003). When members of a species show no preference toward mating with members of their own species it often leads to introgression between the two (Childs et al. 1996; Echelle and Echelle 1997; Garrett 1979; Rosenfield and Kodric-Brown 2003). As increasing body size has been linked to reproductive success in both sexes in other rainbow fish species (Evans et al. 2010; Pusey et al. 2004; Young et al. 2010), it was hypothesised that both sexes from both species would show a preference for larger fish, giving eastern rainbowfish a competitive advantage over RRR.

Contrary to expectation, body size and species identity did not play an important role in determining mate preference in either of the species examined here. As both species are sexually dimorphic, with males growing larger, having larger anal and dorsal fins and exhibiting more colourful patterns on their body and fins, it is reasonable to assume that females should exhibit sexual preference for larger, more colourful males. In many animal groups there is an established link between size and mate preference (Bateman et al. 2001; Blanckenhorn et al. 2000; Charlton et al. 2007; Cooper Jr and Vitt 1993). Additionally, the

73 findings of Young et al. (2010) supported these assumptions for the closely related western rainbowfish (M. australis), as contrary to the present findings, they found a positive relationship between male size and female preference, as well as male-male competitive ability.

However, the findings of my study may have been influenced by the study design, because under typical experimental conditions males and females are unlikely to display their full natural nuptial colours. This would be the same for all studies involving rainbowfish, as colouration is influenced by many factors, including stress, turbidity, predators, lighting and background colouration (Morrongiello et al. 2010; Sumner 1935). Replicating natural conditions in the laboratory is difficult and lighting is known to affect mate choice in birds (Evans et al. 2006) as well as fish (Milinski and Bakker 1990). Due to these conditions, subjects may have been unable to differentiate between species if they relied on visual cues. RRR and M. splendida are closely related – from the same lineage – and although they have quite different colouration in nature, they appeared relatively similar during experimental trials as compared to their typical full display colouration (Figure 3). However, fish also use other cues to select potential mates (e.g. chemical), and the experimental approach used in the current study has been used by previous studies, including those examining rainbowfish (Arnold 2000; Young et al. 2010). Therefore I stand by my findings, which suggest that RRR and eastern rainbowfish may mate randomly in the wild. Unless mate preferences of these two species differ under natural conditions, the natural population of RRR will likely become fully introgressed with the alien M. splendida, becoming extinct within its natural range. Initial results from ongoing genetic monitoring of the wild RRR population support this, as it has found an increase in the frequency of hybrid genotypes in Running River over time (Unmack, unpub. data).

Embryonic survival experiments Post-zygotic barriers to hybridisation generally manifest in the reduced fitness of hybrid offspring but may also result in lower fertility in hybrid pairings (Gilk et al. 2004; Rhymer and Simberloff 1996). I found no significant differences in fertility rates, or the survival rate of fertilised eggs between any of the parental or hybrid combinations. The apparent lack of behavioural barriers combined with the lack of post zygotic barriers suggests that RRR will likely be replaced by a hybrid swarm within its natural range unless hybrid individuals possess maladaptive traits, though this seems unlikely based on the observed trend of an increasing population of hybrids in Running River (Unmack, unpub. data).

Conservation releases RRR were introduced to Puzzle and Deception creeks, as removing the main threat to this species from the natural population (i.e. hybridisation with alien eastern rainbowfish) would be extremely difficult. Conservation introductions have become a useful tool for conservation managers, especially when the cause of decline is unknown or cannot be removed from the species original range (Chilcott et al. 2013; Field et al. 2007; IUCN 2013; Pusey et al. 2003). This is certainly the case for the Lake Pedder galaxias (Chilcott et al. 2013), but is also true across taxonomic groups (Soorae 2018). Conservation introductions into Deception and Puzzle creeks successfully established two self-sustaining populations of RRR. Both of these refuge populations are secured against future incursion of alien fish species by large waterfalls and by virtue of being situated on private land used for conservation purposes.

Predator Training Introgression with eastern rainbowfish was the main threat to RRR, and hybrids were not easily identified without genetic analysis. Releasing fish bred from genotyped brood stock was therefore the best available option. Although biotic resistance is not a determining factor of most freshwater fish invasions (Moyle and Light 1996a; Moyle and Light 1996b), it has been shown to play a role in many introductions utilising captive-bred or reared individuals (Alvarez and Nicieza 2003; Ebner et al. 2007). As a result, research has focused on reducing the effect of predation and better preparing released fish for life in natural settings (Brown and Laland 2002; Brown and Chivers 2007; Roberts et al. 2014). Pre-exposure to predators and their chemical and visual cues have been demonstrated to improve the post release survival of captive-bred fish (Abudayah and Mathis 2016; Brown and Smith 1998; Vilhunen 2006). We attempted to test the impact of repeated exposure to a generalist predator under controlled conditions before release on the survival of captive reared RRR in Deception Creek.

Limitations in the number of fish we had on hand for release and the number of similar sized release pools meant our sample sizes were small enough that only strong differences in survival would have been detected. However, high rainfall, resulting in the connection of release sites during the experiment resulted in the experiment ending before all releases and monitoring could take place, confounding statistical power of any findings. As sample sizes

75 were too small to conduct rigorous statistical analysis, the data from this study cannot be used to draw definitive conclusions.

While predator-trained fish appeared to have a higher abundance than untrained fish in the months immediately after release, there were only small differences in RRR density between trained and untrained sites six months after release. Additionally, differences were confounded by variation in the detectability of visual census between surveys (impacted by visibility; though this was consistent across all sites). Unfortunately there are no highly accurate / truly quantitative ways to measure fish numbers without substantially interfering with released fish and their habitat. Trained and untrained fish appeared to exhibit behavioural differences immediately after release. Untrained fish (5 releases) seemed more likely to endanger themselves by scattering upon release and swimming near the surface in open water. On the other hand, trained fish (2 releases) usually formed one large school which settled and remained in a shallow sandy area for several days before venturing out. Adding to this, the only failed release was one of untrained fish, the release site of which was colonised after flow was restored. While we cannot definitively say that predator training had an impact on release success, the anecdotal information obtained suggests that further research into these techniques is warranted. Future studies could also test behavioural differences between trained and untrained fish prior to release.

Rainbowfish movement Movement and migration is a major knowledge gap for most Australian small-bodied freshwater fishes (Koehn and Crook 2013). Small body size makes methods such as tagging and telemetry difficult. However, conservation introductions of RRR provided a unique opportunity to examine the extent and timing that RRR moved throughout a creek system. With no rainbowfish present in either Deception or Puzzle creeks prior to the release of RRR, the minimum distance traversed by any individual spotted outside release sites could be calculated using the nearest release site. While detailed information regarding flow and barriers traversed, precise timing, or the distances traversed by individual fish could not be obtained, a general estimate of the potential dispersal rate of RRR was obtained.

Rainbowfish moved throughout Deception Creek faster than Puzzle Creek and moved throughout Deception Creek at different speeds between sampling seasons. The reasons for these differences can only be speculated about, but it is likely that different flow regimes and

76 topography were responsible, though it may also have been caused by differences in release times resulting in differences in population growth. Fish in Deception Creek spread further and faster than those in Puzzle Creek. In Puzzle Creek, fish spread at least 460 m upstream with a rise of 9 m elevation and 1.3 km downstream with a 30 m decrease in elevation in 12 months. By contrast, in Deception Creek, fish spread at least 1.9 km upstream with a rise of 31 m elevation in six months, and then 2.4 km with a rise of 39 m in elevation from the closest release site in 11 months, then spread no further because the upper limit of navigable water had been reached. They also spread 1.3 km downstream in the first six months, then 2.7 km in the 12 months and then at least 6.3 km after 18 months though the downstream limit could not be identified during that sampling event time period. Such dispersal ability is impressive, considering the species has previously been confined to a 13 km stretch of river. The speed and timing of large-scale movements will likely differ between species, and widespread species may be inclined to move further or more frequently, though this requires further investigation. This information is directly applicable to the management of new incursions of rainbowfish. As more rainbowfish introductions are likely to take place in the future (given the increasingly imperilled nature of many taxa), understanding the dispersal ecology of rainbowfishes should help in planning for conservation releases as well as limiting rainbowfish invasions.

Implications for management The importance of monitoring Despite its importance in formulating effective conservation release plans, there are few studies incorporating follow-up monitoring on Australian native fish releases, fewer still looking at survival in the weeks immediately after release and even fewer still examining small-bodied species. Monitoring revealed that the release which failed, did so within two days of release. Observations from successful releases suggested that fish introduced to Deception Creek gradually decreased in abundance over time. Natural processes such as predation and finding suitable resources, combined with behavioural deficiencies of captive- reared fish made such declines likely. However, a month from release, surviving fish were able to reproduce during favourable conditions, allowing the population to grow quickly and overcome initial declines. This may suggest that the time of year a release takes place may play an important role in determining whether or not it is successful, both in the establishment of a self-sustaining population and in the genetic diversity of the resulting population, which can be an important determinant of long term success (Frankham 2003). Due to constraints on

77 funding and time, it was not possible to obtain detailed information on initial population growth for fish in Puzzle Creek in the first months after release. As a result, it will be important that future monitoring takes both population size and genetic diversity into account.

An argument for moving more than the species Commonly, conservation introductions of Australian small-bodied freshwater fishes take place in areas where there are no other fish species present (Ayres et al. 2009; Chilcott et al. 2013). One of the main reasons for this is that predation or competition from alien species is often seen as the main cause of a species’ decline (Cadwallader 1996; Lintermans 2000; Morgan et al. 2003) and therefore conservations introductions have to be undertaken in places free of alien predators or competitors. While negative interactions with alien species are the leading causes of decline for some species (Chilcott et al. 2013; Morgan et al. 2003), conservation introductions of RRR have shown that that a complete lack of other species is not required. Studies on hatchery-reared fish have shown rapid loss of behavioural traits, such as predator recognition (Kydd and Brown 2009) and reduced competitive ability (Rhodes and Quinn 1998), suggesting that the recovery or adequate conservation of a species will be detrimentally affected if the species is maintained away from all predators and competitors. I would suggest that when conservation introductions are required, and predation is not an overwhelming threat (e.g. suitable shelter from predators is available), effort should be made to include release sites which contain native competitors and/or predators, and the introduction of natural competitors and predators be considered in the absence of such sites where appropriate.

Contrasts between small- and large-bodied fish translocations Conservation releases for RRR contrast with those of larger-bodied Australian species. Unlike releases for trout cod (Maccullochella macquariensis Ebner et al. 2009; Ebner et al. 2007) and Macquarie perch (Macquaria australasica Lintermans 2013), we were able to determine whether or not our releases were successful over a much shorter time period. This can likely be explained by two factors – the first is that RRR was released into habitats free of the cause of decline, and secondly that RRR, like most small-bodied species, has a much shorter reproductive cycle than large-bodied species. This means that released fish can reach maturity and reproduce in a relatively short period of time, so even if released fish are unable to survive long-term due to behavioural deficiencies caused by captive rearing, wild spawned fish which will not suffer these deficiencies will be present almost immediately. However, it’s also worth noting that a shorter life cycle poses an extra risk. While it’s already been noted

78 that the conservation benefits of captive maintenance for a species may be limited (Araki et al. 2007; Attard et al. 2016; Philippart 1995; Snyder et al. 1996), the short lifespans and generation times of most small-bodied species mean it will take less time for the negative effects of captive maintenance to take effect. However, using fish sourced from more natural settings, such as farm dams or previously translocated populations may reduce these negative effects.

This work establishes two things important to the continued management and conservation of small-bodied fish species; that they may easily establish new populations when the dominant threat is removed, and secondly, that hybridisation between small-bodied species is a bigger risk than previously thought. Most small-bodied species are less likely to be intentionally translocated outside their natural range than large bodied species (Hunt and Jones 2017) and are more likely to enter a new area through other pathways, such as bait bucket translocations and stocking contamination (Lintermans 2004). Given the number of widespread species complexes of small-bodied species in Australia (Hammer et al. 2007; Page et al. 2004; Raadik 2014) the chance of an accidental translocation resulting in establishment, hybridisation, and subsequent introgression is quite high. However, the ease with which populations may be established is also beneficial for the establishment of refuge populations used for conservation, assuming suitable refuge habitat is available.

Knowledge gaps and directions for future research My mate choice experiments failed to find any inter or intra species preferences, differing from the findings of Young et al. (2010). Future research should utilise conditions that encourage the natural colouration of the subject fish. More importantly, unlike Young et al. (2010), this study did not examine the role of intrasexual competition between or within species, which was shown to play an important role in reproductive success for western rainbowfish. Other factors such as reproductive effort, should also be investigated as female reproductive effort has been shown to increase with male body size in western rainbowfish (Evans et al. 2010). Additionally, comparing the fecundity of narrow-range endemics, such as RRR or the Malanda rainbowfish (Melanotaenia sp.) to widespread invasive species such as eastern rainbowfish or their hybrids, may help to explain how widespread species are able to invade and replace them in habitats already occupied by closely related species. Although native species may share a recent common ancestor and may have adapted to the local

79 environment, human-induced environmental changes may shift conditions to the advantage of the invading species and/or hybrids. Therefore, it may be worth comparing environmental variables associated with fitness such as temperature, turbidity, and water velocity between narrow range and widespread species under current and predicted historical conditions.

I was unable to provide conclusive evidence regarding the effectiveness of pre-release predator training, due to small sample size and unplanned weather events. Anecdotal evidence suggested though that predator training may have had some positive effect. Currently, there is no widely agreed upon standardised method for training captive fish to recognise predators and so this is a key area for future research. The methods I used were based on those used by Hutchison et al. (2012), although a variety of other approaches have also been developed (Brown and Smith 1998; Vilhunen 2006). The ability of fish to learn to recognise predators (Brown and Chivers 2007; Brown and Smith 1998; Suboski and Templeton 1989) and the various mechanisms that they may use to do so (Abudayah and Mathis 2016; Vilhunen et al. 2005) have already been established; the next logical step is to determine which methods are best suited for preparing captive-reared fish for release into the wild. Ultimately, multiple standardised methods will be required depending on the traits of the species being trained and the threat(s) they are being trained for (e.g. birds vs fishes; Kydd 2014). An ideal protocol would be interchangeable between species and threat types, utilise easily obtained and constructed apparatus and be time efficient.

A successful conservation introduction not only requires the continuation of a species in the wild but also the maintenance of genetic diversity in the introduced population. Although two populations of RRR have been established, it remains untested as to how well these new populations represent the genetic diversity of the brood fish used to establish each population. Genetic samples have been collected over multiple field seasons for future comparison. Further ongoing retention and analysis of samples will allow the genetic diversity of these populations to be monitored over time allowing for a better understanding of any ongoing changes.

Conclusion As humans increasingly translocate native fishes, there will be more alien populations becoming established, and thus research into the species’ interactions and how to manage the consequences of their introduction is critical. Rainbowfish species with broad distributions

(e.g., eastern rainbowfish, western rainbowfish) possess many traits which allow them to establish new populations quickly, and as a result, the number of rainbowfish species threatened by translocation is likely to increase in the future. Many small-bodied species in Australia are likely to face the same challenges. To date, Australia boasts that it has not yet had any freshwater fish extinctions (at least for known described species), but this is unlikely to remain the case in the future unless we take appropriate management measures. To prevent future declines and extinctions, careful management and continual monitoring will be required. Changes to government policy allowing funding for the conservation of undescribed, but identified taxa, will also be required. The establishment of conservation populations for small-bodied species should be more easily achieved, due to being easier to breed or translocate and having early maturity, but this effort requires a small but important investment of funds towards conservation of smaller native fish.

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Appendix A – Weight and body depth presented as a function of length in RRR and eastern rainbowfish (Melanotaenia splendida)

Body depth Body

)

(

(mm)

Weight

Standard length (mm) Figure S1 Weight and body depth presented as a function of standard length for males and females of Running River rainbowfish (RRR) and M. splendida (Male RRR (a), Male ER (b) Female RRR (c) and Female ER (d) respectively). Body length is plotted along the horizontal axis in millimetres while weight is represented in grams on the primary vertical axis by ● and body depth is represented on the secondary vertical axis in millimetres by▲.

Appendix B – Correlation between spangled perch density and adult rainbowfish density at 11 months from release 3.50

3.00

2.50

2.00

1.50

1.00

0.50 Density Density adult of rainbowfish (fish/m) 0.00 0.00 0.10 0.20 0.30 0.40 0.50 0.60 0.70 0.80 0.90 Density of spangled perch (fish/m)

Figure S2 The density of adult Running River rainbowfish 11 months after release plotted as a function of the density of all spangled perch at release sites in Deception Creek. Density is measured as fish per metre of stream length.

Appendix C – Summary statistics of body size measurements of fish used in mate preference experiments Table S1 Summary statistics of body size measurements of fish used in mate preference experiment trials between Running River rainbowfish (RRR) and Melanotaenia splendida (M. splendida). ‘Average difference in length’ relates to the average difference in body length of the pairs of fish presented as options for each experimental treatment.

Standard length (mm) Body depth (mm) Weight (mg) Male RRR Range 37.8–62.8 9.9–20.4 1146–4845 Mean (± std. deviation) 49.7(6.06) 14.5 (2.26) 2385.0 (939.07) Female RRR Range 40.9–57.3 10.1–14.0 1184–2788 Mean (±std. deviation) 49.5 (4.36) 12.4 (1.03) 2023.1 (477.52) Male M. splendida Range 38.8–58.5 10.3–18.6 952–3970 Mean (±std. deviation) 47.5 (4.91) 13.8 (1.89) 2084.3 (698.94) Female M. splendida Range 39.9–63.6 10.7–34.9 1238–4435 Mean (±std. deviation) 49.1 (5.74) 13.6 (1.74) 2279.3 (804.76) Average difference in length between option fish (mm) Test group Intraspecies Interspecies Male RRR 5.33 4.16 Female M. splendida 7.15 5.27 Male ER 6.39 4.04 Female M. splendida 3.85 5.27

Appendix D – Statistical output from analysis of length, weight and depth relationships Table S2 p and t values from linear regression analysis examining the relationship between length, depth, and weight of male and female Running River rainbowfish (RRR) and M. splendida.

Test Group p-value t-value Length - Width Male RRR 6.41e-16 17.526 Male M. splendida 0.0000000268 7.617 Female M. splendida 2.91e-15 15.083 Female RRR 6.03e-11 10.040 Length - weight Male RRR 6.81e-16 8.25 Male M. splendida 0.0000000122 7.932 Female M. splendida 0.00000000312 8.376 Female RRR 1.73e-14 14.069