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Lionfish invasion in nearshore waters of the Bahamas : an examination of the effects of artificial structures… Smith, Nicola Simone 2010

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    LIONFISH INVASION IN NEARSHORE WATERS OF THE BAHAMAS: AN EXAMINATION OF THE EFFECTS OF ARTIFICIAL STRUCTURES AND INVADER VERSUS NATIVE SPECIES COLONIZATION RATES  by  NICOLA SIMONE SMITH Hons. B.Sc. (with high distinction), The University of Toronto, 2006     A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF  MASTER OF SCIENCE  in  THE FACULTY OF GRADUATE STUDIES  (Zoology)      THE UNIVERSITY OF BRITISH COLUMBIA (Vancouver)  July 2010     © Nicola Simone Smith, 2010   ii Abstract  Artificial structures can facilitate invasion of non-native marine epibiota by providing unoccupied habitat for colonization.  Few studies have examined whether similar effects occur in mobile taxa like reef fishes, despite the widespread occurrence of anthropogenic structure in the world’s coastlines and the critical importance of structured habitat to many fishes.  I assessed the distribution and colonization of invasive Indo-Pacific lionfish Pterois volitans/miles in nearshore waters of the Bahamas where artificial structures are prevalent in human-modified seascapes.  I hypothesized that artificial structures may promote lionfish range expansion by providing sites for colonization, particularly if lionfish are superior colonizers relative to Atlantic taxa.  Using an observational survey, I examined how the type of human modification and habitat may influence invader abundance.  I also used a manipulative experiment to determine: (1) if lionfish were faster colonizers than native species, and (2) if local patterns of colonization were consistent at a regional (island-wide) scale.  I found that lionfish were evenly distributed between habitats dominated by sand and seagrass meadows, hard bottoms and coral patch reefs.  However, nearly 100% of lionfish were associated with artificial structures in sand-seagrass habitat, 25% in hard bottoms, and zero in patch reefs.  Lionfish were poor colonizers of experimental reefs relative to Atlantic taxa, although their colonization rate was not different from the most  iii ecologically similar native species in my study (the small-bodied grouper, Epinephelus guttatus).  Families, but not species, that were good colonizers at one site tended to arrive quickly at other sites as well, implying some predictability in colonization rate among fish families at an island-wide scale.  Artificial structures are a consequence of human coastal development and illegal dumping, but are also often added intentionally to create reef fish habitat.  My results suggest that lionfish are capable of invading natural patch reefs in the absence of such structures, but that their presence facilitates colonization of marginal habitats like sand- seagrass and to a lesser extent, hard bottoms.  Removing or preventing the dumping of debris may therefore slow the spread of lionfish, but is unlikely to prevent their expansion.   iv Table of contents  Abstract ........................................................................................................................... ii Table of contents ............................................................................................................ iv List of tables ....................................................................................................................vi List of figures ................................................................................................................. vii Acknowledgements ...................................................................................................... viii Co-authorship statement ................................................................................................ ix Dedication ……………………………………………………………………………………… x 1 Introduction .................................................................................................................. 1  1.1  Key definitions and concepts ........................................................................ 1  1.2  A theoretical approach: invasion biology as “unplanned experiments in the      natural world” ....................................................................................................... 4  1.3  A conservation approach: invasion biology as a value-based field of inquiry 6  1.4  How do historical invasions differ from contemporary ones? ……………… 11  1.5  Thesis statement and research objectives …………………………………... 13  1.6  References ……………………………………………………………………… 15 2  Artificial structures facilitate Indo-Pacific lionfish invasion into marginal Atlantic habitat  ………………………………………………………………………………………… 21  2.1  Introduction ……………………………………………………………………… 21  2.2  Materials and methods ………………………………………………………… 25   2.2.1  Lionfish distribution ………………………………………………....... 25   2.2.2  Lionfish versus native species colonization rates ………………… 28  v  2.3  Results …………………………………………………………………………… 33 2.3.1  Lionfish distribution ………………………………………………....... 33 2.3.2  Lionfish versus native species colonization rates ………………… 35 2.4  Discussion …………………………………………………………………........  40 2.5  References ……………………………………………………………………...  47 3  Conclusion ………………………………………………………………………………… 53  3.1  Contributions to invasion biology ……………………………………………..  53  3.2  Study strengths and limitations ……………………………………………….. 55  3.3  Management implications ……………………………………………………… 58  3.4  Future research …………………………………………………………………. 60  3.5  References ……………………………………………………………………… 62 4  Appendices ……………………………………………………………………………….. 66  4.1  Appendix A ……………………………………………………………………… 66  4.2  Appendix B ……………………………………………………………………… 71  4.3  Appendix C ……………………………………………………………………… 75  4.4  Appendix D ……………………………………………………………………… 77        vi List of tables  1.  Results of Kruskal-Wallis test of species colonization probabilities ………………… 36 2.  Results of Pearson correlation test for association in colonization probabilities ….. 39  vii List of figures  1.1  Model of the invasion process ……………………………………………………….... 3 2.1  Mosaic plot of habitat type and coastal modification ……………………………….. 27 2.2  Species accumulation curves for experimental reefs ………………………………. 30 2.3  Lionfish distribution in nearshore waters …………………………………………….. 34 2.4  Species colonization probabilities of experimental reefs …………………………… 37 2.5  Scatterplot of log transformed species colonization probabilities among sites ….. 40  viii Acknowledgements  I would like to thank my supervisor, Jon Shurin, for his patience, guidance, encouragement and humour.  I am also grateful for the support of Isabelle Côté and Daniel Pauly, who always made themselves available for discussion.  I would also like to thank Mike Whitlock for statistical help.  Many thanks to Kathleen Sullivan Sealey, Bahamas Department of Marine Resources, Bahamas Reef Environment Educational Foundation, College of the Bahamas Marine and Environmental Studies Institute and Kris Lemkuhl for logistical support with field work.  I am grateful to Elton Josephs, Everton Josephs, Javano Smith, Marcian Tucker, Meredith Turner, Ryan Wynder and countless volunteers (including numerous family members and high school friends) for assistance in the field.  I would also like to thank Lad Akins for technical advice.  Thank you to my family, Shurin lab mates and friends for emotional support throughout this process.  Funding for this project was provided by the Disney Worldwide Conservation Fund and a partial University of British Columbia graduate fellowship.  ix Co-Authorship statement  My supervisor, Dr. Jonanthan Shurin, is a co-author on chapter 2 of this thesis along with Dr. Kathleen Sullivan Sealey.  I was primarily responsible for all aspects of this thesis including the development of research questions and methodologies, data collection and analyses, and the writing of this manuscript.  Dr. Shurin provided guidance throughout this process and revised this manuscript.  Dr. Kathleen Sullivan Sealey helped to secure funding and logistical support for this research.                 x Dedication  This thesis is dedicated to my mother, Portia Maria Smith (1951 – 1998).  1 Chapter 1  Introduction  1.1  Key definitions and concepts  Biological invasions include both human-mediated and natural (i.e. without human agency) forms of dispersal in which a species arrives, establishes a self-sustaining population and subsequently spreads throughout a region in which it did not historically occur (Carlton 1989).  Human-assisted dispersal is referred to as introduction and can involve movement across geographic barriers and great distances.  In contrast, range expansion pertains to the natural diffusion of species into novel locations (Carlton 1989).  Gaps in our knowledge of systematics and biogeography, coupled with the fact that many species introductions occurred prior to the start of biological surveys, make it difficult to distinguish whether some species are native or introduced (Carlton 1989). Carlton (1996a) coined the term “cryptogenic species” to refer to taxa with unclear origins.  He argues that the concept of cryptogenic species is largely overlooked despite emerging evidence that such species may be common in some systems (Carlton 1996a).  Studies that quantitatively assess the prevalence and ecological impacts of invasions may be subject to substantial errors by ignoring cryptogenic taxa, particularly if they equal or exceed estimates of recognized exotics (Carlton 1996a).  2  Despite the challenges in determining whether a species is introduced or native, few species that are transported outside of their native range become widespread and abundant (i.e. invasive), and even fewer are likely to strongly impact invaded systems (Lockwood et al. 2007, Veltman et al. 1996, Williamson and Fitter 1996, Wonham 2006).  The former also holds true for species range expansions throughout geological time in which only a small proportion of taxa from the donor region invaded recipient biotas (Vermeij 1991a, 2005, see section 1.4).  Dispersal events may rarely produce biological invasions because non-indigenous species must overcome a series of barriers in order to progress from one invasion stage to the next (Fig. 1.1) (Lockwood et al. 2007).  For example, in order to establish in a new location, non-native species face demographic and environmental stochasticity, spatial variation in environmental suitability and reverse density dependence (Allee effects), all of which are exacerbated by initially small population sizes (Sax and Brown 2000).  By examining studies on a variety of imported plants and animals from a statistical perspective, Williamson and Fitter (1996) showed that on average, only about 10% of species successfully transition across an invasion stage, with the pool of potential invaders becoming increasingly smaller at each subsequent step.  3  Fig. 1.1  A model of the invasion process for introduced taxa.  Few species that are transported outside of their native range overcome the ecological and physical barriers necessary to transition across stages.  Each step in the invasion process is represented by a box.  Diagram modified from Lockwood et al. 2007.  The final stage of invasion is impact.  All invaders affect the recipient community as they become integrated into local food webs and interspecific interactions as predators, competitors, parasites, pathogens, hosts or mutualists (Elton 1958, Strauss et al. 2006, Wonham 2006).  Parker et al. (1999) proposed a framework for assessing the ecological impact of an invader as the product of its range, abundance and per capita effect on native biota.  Species that score highly in all three dimensions are considered to have the strongest impact.  However, whether or not the effect of an invader is  4 regarded as ‘harmful’ or ‘beneficial’ often depends on the value system and environmental goals of the evaluator (Wonham 2006, see section 1.3).  1.2  A theoretical approach: invasion biology as “unplanned experiments in the natural world”  As early as the mid-1800s naturalists recognized (but did not explicitly articulate) that invasions could be used to test fundamental ideas in ecology and evolution (Sax et al. 2007).  Joseph Grinnell was likely the first to refer to invasions as “unplanned experiments in the natural world” (Sax et al. 2007).  But others, including Charles Darwin, also employed this approach. For example, Darwin (1859) used observations from species introductions to support the notion that dispersal limitation (due to geographic barriers) was a primary constraint on species distributions (Sax et al. 2007). In The Origin of Species he states, “in the case of an island, or of a country partly surrounded by barriers, ... had the area been open to immigration, these same places would have been seized on by intruders” (p. 68).  Invasions also generated some of Darwin’s ideas on local adaptation and phylogenetic constraint on adaptive evolution, suggesting that native species may be competitively subordinate to invaders because they lack traits that have arisen elsewhere (Sax et al. 2007).  Darwin (1859) explains, “as foreigners have thus everywhere beaten some of the natives, we may safely conclude that the natives might have been modified with advantage, so as to have better resisted such intruders” (p. 69).  5  A little over a century after Darwin, Baker and Stebbins’ (1965) classic edited volume, The Genetics of Colonizing Species, was published in which scientists broadly considered colonizations (including species introductions and range expansions) in order to formulate generalizations about evolutionary processes.  But Baker and Stebbins’ (1965) theoretical approach to invasions is considered to be an exception for its time (Davis 2006, Sax et al. 2007).  It is only recently that the study of invasions to address basic questions in biology has become an explicit objective in mainstream science.  Entire books on the subject now exist (e.g. Cadotte et al. 2006, Sax et al. 2005).  Sax et al. (2007) propose that exotic species can serve as “model organisms” to further our understanding of fundamental processes in ecology.  For example, drawing on recent comparative studies (e.g. Sax et al. 2002, Sax and Gaines 2003), they argue that there is little evidence of species saturation in natural communities, as the number of established exotics generally exceeds the number of extinctions at regional scales for most taxa (Sax et al. 2007).  Similarly Strauss et al. (2006) assert that the “accidental experiments created by invasions” present novel opportunities to test hypotheses in evolution.  They point out that invasions provide some of the best examples of contemporary evolution operating in the natural world (e.g. Carroll et al. 2005) and may elucidate the role of evolutionary processes such as character displacement in community assembly.   6 Invasions are also beginning to inform broad issues in applied biology such as climate change and the environmental limits to species ranges.  Duncan et al. (2009) used the introduction of several species of dung beetles to test the fundamental assumption of climate envelope models (CEMs) that climate limits species distributions.  But non- climatic factors such as biotic interactions (e.g. predation and competition), dispersal and/or resource limitation, and local abiotic conditions (e.g. soil quality) can also prohibit the occurrence of species in otherwise climatically suitable areas (Duncan et al. 2009). Duncan et al. (2009) showed that CEMs derived from current native distributions generally transferred poorly to observed beetle distributions in the introduced range.  By comparing predictions of CEMs based on native distributions (where it was uncertain if climate is the primary constraint on species ranges) with those derived from distributions in the introduced range (where models performed well and climate was found to limit species ranges due to the removal of all other confounding non-climatic variables), they attributed the inability of the former to predict the beetles’ introduced range to the possibility that, unlike in the introduced range, climate is not the primary determinant of beetle distributions in the native range.  1.3  A conservation approach: invasion biology as a value-based field of inquiry  ...invasive species present a complex social and ethical quandary rather than solely a biological one.           (Larson 2007)   7 In the foreword to Charles Elton’s (1958) seminal text, The Ecology of Invasions by Animals and Plants, Daniel Simberloff states that Elton’s work “founded a whole field of research” (p. vii).  I illustrated above that the study of invasions (at least with the aim to inform basic ideas in biology) began in the mid-1800s (Sax et al. 2007).  Simberloff underscores that Elton (1958) was the first, however, to unite the three fundamental themes of invasion biology: (1) natural history, (2) ecology, and (3) conservation (although I would argue that invasion biology has four components, the final one being evolution).  Nonetheless, much of Elton’s (1958) work focused on the third theme of conservation. His extensive compendium of examples and insights into the pervasiveness and detrimental impacts of invaders on both the ecology of natural systems and human well- being undoubtedly aided in bringing the study of invasions in their own right to the forefront of many research agendas.  Elton (1958) highlighted the consequences of invasions for biodiversity (although it was not called this at the time) and suggested ways to manage them.  However, as Davis (2006) points out, it was likely Elton’s (1958) work in tandem with the emergence of the environmental movement in the 1960s-1970s and the subsequent development of conservation biology as an independent discipline that precipitated the current era in which many biologists study invasions from a value- based perspective.  Although few invaders strongly impact recipient communities (Williamson and Fitter 1996), the small proportion that do has made biological invasions one of the leading  8 threats to biodiversity worldwide (Sala et al. 2000, Wilcove et al. 1998).  Invasions have resulted in native species extinctions, population declines and displacements.  These effects are most often due to predation (e.g. Blackburn 2005, Savidge 1987), but may also occur as a consequence of competition (e.g. Holway 1999), the introduction of novel pathogens (e.g. Anagnostakis 1987, Van Riper et al. 1986), or hybridization of rare native species with exotics (e.g. Rhymer and Simberloff 1996).  Some invasive species have extensively altered habitat structure or ecosystem processes such as nutrient cycling that in turn, have transformed the biological composition of entire communities (e.g. Crooks 2002, Vitousek 1990).  Moreover, invasions can contribute to biotic homogenization at both a regional and global scale (Lodge 1993, McKinney and Lockwood 1999, Vitousek et al. 1997, Vander Zanden 2005).  The extinction of endemic species is occurring worldwide (largely due to habitat loss and/or antagonistic interactions with invaders), resulting in a decline in global species diversity (McKinney and Lockwood 1999, Sax and Gaines 2003). However, at the same time, the establishment and spread of non-indigenous taxa is generally increasing diversity at sub-global levels because the number of established exotics typically exceeds the number of native species extinctions (Sax et al. 2002, Sax and Gaines 2003).  Although invasions do not generally decrease local diversity, the combined effect of the loss of rare species and the spread of cosmopolitan (usually introduced) taxa is resulting in a decrease in biological distinctiveness among regions (i.e. declining beta diversity) and ultimately, a gradual homogenization of the world’s  9 biota  (Lodge 1993, McKinney and Lockwood 1999, Vander Zanden 2005, Vitousek et al. 1997).  Invasions are not only costly to the environment but also to the economy.  Pimentel et al. (2000) estimated the economic costs associated with damages caused by non- indigenous species (e.g. clogging of waterways, losses in agricultural production, fouling of ships) and control efforts (e.g. physical or chemical removal, biological control programs) at billions of dollars per year in the US alone.  For example, the state of Texas spends an estimated $300 million per year in damages to livestock, wildlife and public health caused by the introduced red fire ant, Solenopsis invicta, while an additional $200 million per annum is invested in control efforts (Pimentel et al. 2000).  It is important to note, however, that some non-native species are not invasive and benefit the economy like intentionally introduced food crops (e.g. corn and rice) and livestock (e.g. cattle).  Larson’s (2007) opening quotation to this section underscores the diversity of challenges presented by invasions.  Biological invasions are interesting from a theoretical perspective because they provide unique opportunities to explore fundamental concepts in biology (Sax et al. 2007, Strauss et al. 2006).  However as illustrated above, some invasive species are facilitated by humans (e.g. intentional or accidental introductions), pose a significant threat to the variety of life on this planet (Sala et al. 2000, Wilcove et al. 1998), and negatively impact the economy (Pimentel et  10 al. 2000).  As such, invasions also present biologists with a moral responsibility to create and apply scientific knowledge in ways that effectively address these issues.  It is important to note that for the small fraction of non-indigenous species that strongly affect invaded systems, a critical but forthright assessment of invader impacts that clearly articulates the investigator’s use of a valuing framework is needed.  Whether or not the impact of an exotic is regarded as ‘harmful’ or ‘beneficial’ depends on the goals of the evaluator (Wonham 2006).  The controversy surrounding the control of tamarisk, Tamarix spp., is an excellent case in point.  On one hand, tamarisk is a highly invasive shrub that displaces native vegetation, salinizes soil and degrades wildlife habitat for a variety of species in the western United States (Zavaleta 2000).  On the other, tamarisk provides substantial (and in some areas, preferred) nesting habitat for the endangered, native southwestern willow flycatcher, Empidonax trailii extimus (McKernan and Braden 1999, USFW 1997). Because of conflicting conservation goals (i.e. maximizing native species richness in riparian zones vs. protection of an endangered, native songbird), this situation has led to a contentious dispute between wildlife scientists (Malakoff 1999).  Subjective judgments about invasive species (e.g. this invader is ‘bad’ for biodiversity) are necessary in order to elicit a public response.  However, such appeals should be accompanied by an explicit statement indicating the use of a valuing system and its associated goals.   11  1.4  How do historical invasions differ from contemporary ones?  The rate and magnitude of biological invasions (relative to the 1800s) has been increasing worldwide at the close of the 20th century (Carlton 1996b, Cohen and Carlton 1998, Pimentel et al. 2000, Ricciardi and Atkinson 2004).  Most point to the expansion of global trade and travel in addition to widespread human-induced habitat alterations (which potentially favor invasive species) as likely explanations for this trend (e.g. Cohen and Carlton 1998, Cartlon and Ruiz 2005, Pimentel et al. 2000).  However, biological invasions are not a novel feature of the current age of human-domination of the biosphere.  Invasions have occurred throughout the history of life itself, as some taxa gradually evolved the ability to breach physical or ecological barriers while large- scale elimination of geographic barriers (e.g. due to plate tectonic activity) allowed for species from neighboring biotas to extend their ranges (also referred to as biotic interchange, Elton 1958, Vermeij 1991a, 2005).  One important way in which the Earth’s earliest invasions differ from contemporary ones, however, is that humans are now moving large numbers of organisms around either directly (by species introductions) or indirectly (by removal of former barriers such as in the case of marine invasions via the Suez canal, see Por 1971) at an unprecedented rate (Elton 1958, Lodge 1999, Vitousek et al. 1997). The predominant view of biologists and paleontologists is that invasions during geological times were a slow process occurring over thousands to millions of years while those today (due to  12 human-mediated dispersal) are comparably rapid, operating over the time scale of years to decades (Elton 1958, Vermeij 2005).  For example, when comparing the Pliocene invaders during the Great American interchange with the early 1900s invasion of sea lamprey into the inner Great Lakes of North America, Elton (1958) notes, “Of course the scale of time is totally different - one in millions and the other in decades” (pg. 40).  In terms of the number of species that invade new areas, studies from both contemporary and geological times reveal that only a minority of taxa from donor regions establish in recipient biotas (Williamson and Fitter 1996, Vermeij 1991a).  As previously stated, only about 10% of taxa transition across invasion stages during species introductions (Williamson and Fitter 1996).  Similarly, for cases of invasions during geological times like the Great American interchange, only 2% to 11% of North American mammalian taxa and 2% to 7% of their South American counterparts expanded their ranges at any given time (Marshall et al. 1982, Vermeij 1991a).  There is, however, considerable variation among habitats in the prevalence of invasive species (Vermeij 2005).  At the close of the 20th century, some habitats (especially oceanic islands, lakes and estuaries) are dominated by non-indigenous taxa in terms of number of species or individuals, or total biomass (e.g. Cohen and Carlton 1998, Sax et al. 2002) while others are not (e.g. Wasson et al. 2005).  The same can be said for patterns of invasion during geological times (Vermeij 2005).  For instance, during the Trans-Arctic interchange, North Atlantic rocky shores were heavily invaded by  13 molluscan species of Pacific origin while muddy and sandy bottoms were not (Vermeij 1991b).  A final point of comparison involves the types of species that come into contact during invasions.  Due to international commerce, humans are bringing together taxa from geographically distant biotas in which natural dispersal would be exceedingly rare (Lockwood et al. 2007, Vermeij 2005).  In contrast, biotic interchange during geological times was restricted to taxa from adjacent regions (Vermeij 2005).  The geographic scope of the kinds of species involved in present-day invasions is therefore substantially greater than in the distant past.  1.5  Thesis statement and research objectives  In response to the current magnitude and myriad of human impacts on the biosphere, Jane Lubchenco (1998) proposed a “new social contract for science” in which scientists create and apply knowledge in ways that address the most critical problems facing the planet.  Biological invasions are one of these problems.  They are an important element of human-driven global change (Vitousek et al. 1997).  In 1983 the Scientific Committee on Problems of the Environment (SCOPE) founded an international program on biological invasions that brought attention to this issue and resulted in the publication of several volumes (Davis 2006).  This initiative was the first of its kind and focused investigations into three main areas that have obvious conservation implications: (1) species traits that convey invasiveness, (2) characteristics of the recipient environment  14 that determine susceptibility to invasion, and (3) ways in which knowledge of the above can inform management decisions (Davis 2006).  My thesis research is in keeping with the guidelines of the SCOPE invasion program. Specifically, I used the ongoing range expansion of Indo-Pacific lionfish, Pterois volitans/miles, in the western Atlantic Ocean as a case study to determine whether there is a match between an organismal attribute of the invader and a physical property of the recipient environment such that it increases invasion success.  I also suggest how my findings may inform potential management strategies for lionfish in the Atlantic. Focusing on colonization (a prerequisite for both the establishment and spread phases of invasion), I hypothesized that artificial structures may facilitate lionfish expansion into nearshore waters by providing sites for colonization, particularly if lionfish are better colonizers than Atlantic taxa.  Research objectives included: (1) to determine whether coastlines with artificial structures contained more lionfish than less human-modified areas, (2) to determine whether lionfish are superior to native reef fishes as colonizers of artificial structures, and (3) to determine whether local patterns of colonization are consistent at a regional (island-wide) scale.         15 1.6  References  Anagnostakis, S.L.  (1987)  Chesnut blight: the classical problem of an introduced  pathogen.  Mycologia, 79, 23-37. Baker, H.G. & Stebbins, G.L., eds.  (1965)  The Genetics of Colonizing Species.  Academic Press. Blackburn, T.M., Petchey, O.L., Cassey, P. & Gaston, K.J.  (2005)  Functional diversity of mammalian predators and extinction in island birds.  Ecology, 86, 2916-2923. Cadotte, M.W., McMahon, S.M. & Fukami, T., eds.  (2006)  Conceptual Ecology and  Invasion Biology: Reciprocal Approaches to Nature.  Springer. Carlton, J.T.  (1989) Man’s role in changing the face of the ocean: biological invasions  and implications for conservation of near-shore environments.  Conservation  Biology, 3, 265-273. Carlton, J.T. (1996a)  Biological invasions and cryptogenic species.  Ecology, 77,  1653-1655. Carlton, J.T. (1996b)  Pattern, process, and prediction in marine invasion ecology.  Biological Conservation, 78, 97-106. Carlton, J.T. & Ruiz, G.M.  (2005)  The magnitude and consequences of bioinvasions in  marine ecosystems.  Marine Conservation Biology: The Science of Maintaining  the Sea’s Biodiversity.  (eds. Norse, E.A. & Crowder, L.B.), pp. 123-148.  Island  Press.  16 Carroll, S.P., Loye, J.E., Dingle, H., Mathieson, M., Famula, T.R. & Zalucki, M.P.  (2005)  And the beak shall inherit - evolution in response to invasion.  Ecology Letters, 8,  944-951. Cohen, A.N. & Carlton, J.T.  (1998)  Accelerating invasion rate in a highly invaded  estuary.  Science, 279, 555-557. Crooks, J.A.  (2002) Characterizing ecosystem-level consequences of biological  invasions: the role of ecosystem engineers.  Oikos, 97, 153-166. Darwin, C. (1859)  The Origin of Species.  Oxford University Press. Davis, M.A.  (2006)  Invasion biology 1958-2005: the pursuit of science and  conservation.  Conceptual Ecology and Invasion Biology: Reciprocal  Approaches to Nature.  (eds. Cadotte, M.W., McMahon, S.M. & Fukami, T.), pp.  35-64. Springer. Duncan, R.P., Cassey, P. & Blackburn, T.M.  Do climate envelope models transfer?  A  manipulative test using dung beetle introductions.  Proceedings of the Royal  Society B: Biological Sciences, 276, 1449-1457. Elton, C. (1958)  The Ecology of Invasions by Animals and Plants.  University of  Chicago Press. Holway, D.A.  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Mooney, H.A. & Hobbs,  R.J.), pp. 261-300. Island Press.  21 Chapter 2  Artificial structures facilitate Indo-Pacific lionfish invasion into marginal Atlantic habitats1  2.1  Introduction  All habitats can be invaded under some conditions (Davis et al. 2000; Davis & Pelsor 2001; Moyle and Light 1996a, b); however, human-modified habitats are generally more susceptible (Elton 1958).  Hypotheses proposed to explain this pattern include: (1) increased vectors and propagule pressure in human-disturbed habitats, and (2) changes to the abiotic and/or biotic environment as a result of human activities which favor invasive species (Cohen & Carlton 1998; Lockwood et al. 2007).  This chapter examines the latter hypothesis as it relates to increased invasion vulnerability of coastal areas due to adding artificial structures.  An estimated 1.2 billion people live within 100 km of a shoreline and 100 m of sea level where the population density is almost three times greater than the global average (Small & Nicholls 2003).  These areas are centers of large human settlements with concomitant demands for coastal development and socioeconomic activities (Small & Nicholls 2003).  Hence, the prevalence of artificial marine structures is increasing worldwide (Bulleri 2005; Glasby & Connell 1999).  This is cause for concern because  1  A version of this chapter has been submitted for publication.  Smith, N.S., Shurin, J.B. & Sullivan Sealey, K.  Artificial structures facilitate Indo-Pacific lionfish invasion into marginal Atlantic habitats.  22 artificial structures are often foci for the establishment and spread of non-native taxa (Ruiz et al. 2009).  For example, Bulleri and Airoldi (2005) found that breakwaters (coastal defense structures) along the northeast Adriatic coast of Italy facilitated the range expansion of the exotic alga, Codium fragile ssp. tomentosoides, by serving as a dispersal corridor across seascapes of unsuitable habitat.  Similarly, Glasby et al. (2007) showed that although non-indigenous epibiota occurred on both natural and artificial substrates in Sydney Harbour, Australia, exotic richness was up to 2.5 times greater on pontoons and pilings than on nearby natural rocky reefs.  Artificial structures may increase colonization success of non-native taxa by providing unoccupied or novel habitat for establishment (Tyrrell & Byers 2007).  Most studies of the role of artificial structures in facilitating marine invasions involve taxa whose life history is obligately dependent on hard substrate such as sessile fouling organisms (e.g., algae, sponges and bivalves) or invertebrates with relatively limited dispersal ability (e.g., polychaetes).  Few have addressed the consequences of artificial structure creation for marine invasions in mobile taxa like reef fishes, despite evident associations of some species with artificial substrates in invaded regions (e.g., Sapota & Skóra 2005).  The fluctuating resource availability hypothesis predicts that an environment is most susceptible to invasion when there is an increase in a resource that limits both resident and invading species (Davis et al. 2000; Davis & Pelsor 2001).  Space is frequently limiting in fouling communities (e.g., Stachowicz et al. 1999).  Increased availability of  23 hard substrate could thus enhance the likelihood of establishment of non-indigenous marine epibiota.  Likewise, the availability of structural habitat is an important resource for reef fishes.  Structures provide refuge from predation, storms and wave surge as well as nesting and foraging sites (Steele 1999).  Suitable habitat is therefore often limiting in reef fishes (e.g., Hixon & Beets 1989; Munday 2004; Sale 1978).  I propose that artificial structures may facilitate reef fish invasions into coastal areas by providing sites for colonization, as they do for sessile benthic organisms.  In addition to attributes of the recipient environment, attempts have been made to identify biological characteristics that predict successful invaders (Kolar & Lodge 2001, 2002; Moyle & Marchetti 2006).  Whether or not a trait increases invasiveness likely depends on the environmental context and the stage of the invasion process (Kolar & Lodge 2001, 2002).  I examined whether invasive Indo-Pacific lionfish Pterois volitans/miles are superior to native Atlantic taxa as colonizers of artificial substrate. Given the ubiquity of human-made structures as potential colonization sites for reef fishes, an exotic could become widespread and abundant (i.e., invasive) if it has superior colonizing ability relative to natives, allowing it to monopolize a regionally abundant resource in human-modified environments.  I used the ongoing range expansion of Indo-Pacific lionfish in the western Atlantic Ocean to test the above hypotheses.  P. volitans and P. miles are carnivorous, reef- associated fishes native to the Indian and Pacific Oceans (Schultz 1986).  Likely introduced off the east coast of Florida in the early 1990s via aquaria releases  24 (Semmens et al. 2004), lionfish populations are expanding along the southeast US seaboard and the Caribbean (Schofield 2009).  More recently, P. volitans has been reported in the Gulf of Mexico and off the Atlantic coast of both Central and South America (Schofield 2009).  P. volitans and P. miles are sister species that are difficult to differentiate visually due to close morphological resemblance and some overlap in meristics (Hamner et al. 2007).  Although both P. volitans and P. miles occur in US waters (Hamner et al. 2007), only P. volitans has been confirmed elsewhere in the Atlantic.  It is likely that most lionfish in the Bahamas are P. volitans (Freshwater et al. 2009), however I do not distinguish the two species in this study.  The lionfish invasion of the Atlantic was identified as one of fifteen emerging global conservation issues by Sutherland et al. (2010). Considerable concern arises due to the potential negative effects of lionfish on Atlantic species via predation and competition for both food and space on reefs (Albins & Hixon 2008; Whitfield et al. 2007).   The first empirical investigation into the impacts of P. volitans in the Atlantic by Albins and Hixon (2008) showed that lionfish reduced recruitment of native fishes to experimental reefs in the Bahamas by an average of 79% over a five-week period.  My study investigates the distribution and colonization of lionfish in nearshore waters of New Providence Island, Bahamas, where current lionfish densities are estimated at nearly five times higher than reported in its native range (Green & Côté 2009).  I predicted that artificial structures may facilitate lionfish invasion into coastal waters by providing sites for colonization, particularly if lionfish are superior colonizers relative to  25 natives.  The aims of my study were threefold: (1) to determine whether coastlines with artificial structures contained more lionfish than less human-modified areas, (2) to determine whether lionfish are superior to native reef taxa as colonizers of artificial structure, and (3) to determine whether local patterns of colonization are consistent at a regional (island-wide) scale.  2.2  Materials and methods  2.2.1 Lionfish distribution  I conducted an observational survey to determine whether areas with artificial structures contained more lionfish than less altered coastlines.  A total of 16 nearshore sites were selected opportunistically along the northern coast of New Providence based on ease of access from shore and level of boat traffic.  Sites surveyed ranged in size from 5,000 - 10,000 m2, were less than 5 m in depth and no more than 200 m offshore.  I classified sites into three types of modification based on the presence of human-made substrate: (1) none (no artificial structure), (2) debris (presence of one or more items of hard land- based debris that could serve as reef fish habitat such as wrecks and sunken pipes), and (3) infrastructure (e.g., docks, jetties, and seawalls).  Natural habitat may also influence lionfish abundance, hence sites were identified as being dominated by: (1) coral patch reefs (hereafter referred to as patch reefs), (2) hard bottom (limestone), or (3) sand-seagrass (sandy substrate with varying densities of seagrass).   26 All surveys were conducted during July - August 2007.  Due to the relatively low abundance and patchy distribution of lionfish in nearshore waters, all surveys were conducted on snorkel using the roving diver technique (RDT; Schmitt et al. 2002). During RDT surveys, the observer snorkeled within a pre-defined area for 40 to 80 minutes (depending on site size) and recorded the total number of lionfish encountered as well as the substrate that lionfish were found directly on.  The abundance of non- cryptic native species was also recorded.  In order for comparison among sites, I standardized lionfish abundance by time to calculate lionfish sighting rate.  The types of habitat altered by humans were biased such that less modified sites were always dominated by patch reefs (Fig. 2.1).  I was therefore unable to separate the effects of artificial structure and natural habitat on lionfish abundance, or determine whether there was a significant interaction between the two. Hence, I performed a one- factor ANOVA with each of the two explanatory variables (presence/absence of artificial structures and habitat type).  Visual inspection of normal quantile plots and the Shapiro- Wilk test indicated that all data conformed to the assumption of normality while the Bartlett test revealed that the assumption of homogeneity of variance had been met. Fisher’s exact test was used to determine if the presence of artificial structures varied between habitat types (i.e., human bias in the types of habitat modified).  Although I could not test directly for an interaction between artificial structure and habitat on lionfish abundance, I examined whether the proportion of lionfish occupying artificial structures differed between habitats. No lionfish were associated with artificial substrate around patch reefs.  Hence for sites where lionfish were present, I used the non-  27 parametric Wilcoxon rank sum test to compare the proportion of lionfish associated with artificial substrate in hard bottom versus sand-seagrass.  Figure 2.1  Mosaic plot of the proportion of sites within each habitat that were modified by adding artificial structures (grey) and those that were not (black).  The area of each box is proportional to the number of sites surveyed: Sand-seagrass = 4, Hard bottom = 7, Patch reef = 5.  There was a strong association between habitat and whether or not artificial structures were present (Fisher’s exact test, P = 0.003).     28 2.2.2  Lionfish versus native species colonization rates  I conducted a manipulative experiment to determine whether lionfish are faster to colonize artificial structures than native taxa.  I used a block design with four nearshore sites along the northern (two sites: Manor Beach and Triplets) and southern (two sites: Adelaide and Coral Harbour) coastlines of New Providence.  Sites covered 20,000 m2 of predominantly sandy bottom, ranged from 100 - 600 m offshore, were at least 100 m from the nearest substantial reef structure, and were all less than 5 m deep.  During June - August 2008, 15 identical 0.5 m2 artificial concrete block reefs were laid out in a grid pattern at each site with 50 m between each reef.  Each reef was treated as an independent replicate.  I recorded natural colonization by both native taxa and lionfish of all size classes, including recruits less than 5 cm total length, to experimental reefs over five to six months until December 2008 (except for at one site, Manor Beach, where colonization was followed for only one month due to logistical difficulties).  The rate of colonization is expected to be proportional to species abundance, as more numerous taxa are more likely, by chance, to be in close proximity to experimental reefs.  I therefore determined per capita colonization probability (hereafter referred to as PC) for each taxon on experimental reefs:  PC = Per capita Pr [colonization] = 1 - (1 - 1 / t) 1/abundance   29 where t is time (recorded in days, and then converted to hours) before the first individual of a species appeared on a reef while abundance represents the estimated abundance of the species in the study area at the start of the experiment, determined by visual surveys (described below). PC therefore measures the likelihood of arrival of an individual in the study area given its species abundance.  PC assumes that: (1) the probability that a species colonizes a reef remains constant over time, and (2) interspecific processes do not influence whether or not an individual colonizes.  I measured PC at the level of the reef for each taxon (n = 15 per site).  PC for each species was then averaged across reefs to compare colonization rate among all taxa at a site.  In cases in which a species did not colonize a reef during the study period, we assumed that PC was zero so that the sample size was always the same for all species.  Because reefs were sampled at discrete intervals, we used the mid-point between the day when a species was initially observed on a reef and the previous census day to estimate colonization time.  The census schedule consisted of four to six surveys per site during the study period (except for Manor Beach, which was sampled twice over a one-month period). The precision of estimates of PC is likely sensitive to how frequently reefs were surveyed and the time interval between surveys.  The census schedule was most frequent during the first month of colonization when the rate of species accumulation was likely to be greatest and became progressively less frequent later in the study (Fig. 2.2).  30  Figure 2.2  Cumulative number of species observed on experimental reefs from June - December 2008.  Census frequency was highest during the first month of colonization when the rate of species accumulation was greatest and became progressively less frequent later in the study.  Data are means (+/- SE) typically from 14 to 15 independent reef replicates per site.  (a) Adelaide, (b) Coral Harbour, (c) Triplets.  Data from Manor Beach are not shown due to a prematurely terminated sampling period.  I used snorkel and SCUBA to survey reefs (depending on site depth) and followed the protocol of Sale and Douglas (1981).  Fishes were identified to species except for grunts (Haemulidae), which often arrived as young recruits that were classified only to  31 genus.  Transient predators to experimental reefs such as jacks (Carangidae) were also noted but not included in PC calculations since they are not considered to be reef residents (Appendix A).  Local abundance of all non-cryptic species was determined via modified RDT surveys of the entire 20,000 m2 in which the study area was divided into 50 x 50 m squares.  For each square, I swam in an S-pattern for 15 minutes and recorded the total number of individuals of each species encountered within 2.5 m on either side.  Because of uncertainty in detection likelihood, all cryptic species, taxa with diurnal crevice use and those that burrow in sand were excluded from abundance estimates and PC calculations despite being observed on experimental reefs (Appendix A).  All of the study sites, except for one (Manor Beach), were surveyed between two to four times within the week of reef deployment in order to estimate local abundance early in the experiment.  At each site, I conducted repeated surveys on separate days, at different starting times and used a different starting point within the grid each time in order to obtain the most comprehensive list of the local species pool given the degree of effort.  A few species that colonized experimental reefs were not detected in local abundance surveys (Appendix B).  In these instances, it was assumed that the species either: (1) emigrated from an area larger than that defined by the study, or (2) was so rare that it was undetected in surveys.  These species were arbitrarily assigned a local abundance of half of the least abundant species estimated in site-specific surveys so that a PC could still be calculated.  Conversely, a small number of reef species were detected in  32 local abundance surveys but were not observed on experimental units, presumably due to: (1) predation, or (2) failure to colonize during the study period.  These taxa were not included in PC calculations (Appendix C).  To determine if lionfish are faster colonizers than native species, I compared PC among species within each site by examining the position of lionfish in the distribution of PC among taxa.  I also compared lionfish colonization rate to the closest ecologically similar native species in each assemblage (selected based on weight and diet) by a planned comparison using a Wilcoxon rank sum test.  Finally, I tested whether species that were good or poor colonizers at one site display a similar pattern at other sites.  If this was the case, I expected that species PC would be positively correlated among sites.  I therefore calculated Pearson’s correlation for species-specific PC among sites using only species that were common to each pair of sites considered.  In all instances, data were natural log transformed in order to meet the assumption of bivariate normality.  The above analyses, including transformations, were repeated to determine if there was any consistency in fish family colonization rate among sites.       33 2.3  Results  2.3.1  Lionfish distribution  There was no difference in lionfish abundance between sites that were modified by adding artificial structures and those that were not (single-factor ANOVA, F2,13 = 0.158, P = 0.856; Fig. 2.3a).  Although a greater proportion of lionfish occurred on artificial structures in areas containing infrastructure (mean = 72%) than debris (mean = 25%), this trend was non-significant (Wilcoxon rank sum, P = 0.178; Fig. 2.3b).  Despite similar abundances in modified versus unmodified habitats, lionfish never occurred on artificial structures in the latter.  There was no difference in lionfish abundance among sand-seagrass, hard bottom or patch reefs (single-factor ANOVA, F2,13 = 0.952, P = 0.411; Fig. 2.3c).  There was a strong association between habitat and type of modification.  Artificial structures were always found in hard bottom and sand-seagrass but rarely in areas containing patch reefs (Fisher’s exact, P = 0.003; Fig. 2.1).  For sites in which lionfish were present (i.e., 11 out of the 16 sites), the proportion of lionfish on artificial structures was significantly greater in sand-seagrass, where nearly all lionfish occupied artificial structures (mean = 96%), than in hard bottom (mean = 25%) (Wilcoxon rank sum, P = 0.046; Fig. 2.3d).  No lionfish were observed on artificial structures in habitats dominated by patch reefs, although lionfish abundance was similar to the other habitats (Figs. 2.3c & 2.3d).  34  Figure 2.3  Lionfish sighting rate (proxy for abundance, top panels) and proportion of lionfish in invaded sites that occupied artificial structures (bottom panels) at 16 nearshore sites grouped according to type of modification (grey) and habitat (white). Different letters above boxplots denote statistical significance, ns signifies non- significance.  (a) There was no difference in lionfish abundance between sites that were modified by adding artificial structures and those that were not (single-factor ANOVA, F2,13 = 0.158, P = 0.856).  (b) Although a greater proportion of lionfish occupied artificial structures at sites containing infrastructure (mean = 72%) than debris (mean = 25%), this trend was non-significant (Wilcoxon rank sum, P = 0.178).  Statistical tests  35 comparing modification categories did not include sites without artificial structures since every lionfish observed occupied natural substrate.  (c) There was no difference in lionfish abundance between habitats (single-factor ANOVA, F2,13 = 0.952, P = 0.411). (d) A significantly greater proportion of lionfish occupied artificial structures in sand- seagrass (mean = 96%) than in hard bottom (mean = 25%) (Wilcoxon rank sum, P = 0.046).  No lionfish were observed on artificial structures near patch reefs.  2.3.2  Lionfish versus native species colonization rates  There was significant among-species variation in PC at all sites (Table 2.1 & Figs. 2.4a- d).   Lionfish colonized experimental reefs at only two sites (Coral Harbour and Manor Beach).  Although no lionfish were detected in initial local abundance surveys at Coral Harbour (where reefs were surveyed for five months), I observed five juveniles in the study area at Manor Beach during reef deployment (where reefs were surveyed for one month).  Despite these differences, lionfish occurred in the tail of the PC distribution at both sites (Figs. 2.4b & 2.4c).         36 Table 2.1.  Kruskal-Wallis test showing significant among-species variation in colonization probabilities.  AD = Adelaide, CH = Coral Harbour, MB = Manor Beach, TR = Triplets. **P<0.01, ***P<0.001. Site !2 df P AD 116.193 21 <0.001*** CH 156.887 29 <0.001*** MB 61.274 16 0.003** TR 157.905 22 <0.001***   37  Figure 2.4  Species-specific per capita probability of colonization (PC) of experimental reefs estimated during June - December 2008 (except at Manor Beach, where reefs were surveyed twice during August 2008).  There was significant among-species variation in PC at all sites (see Table 2.1).  Data are means (+/- SE) from 14 to 15 independent reef replicates per site.  PC for Pterois volitans/miles and the closet ecologically similar native species in the study, Epinephelus guttatus, are indicated by arrows.  (a) Adelaide, (b) Coral Harbour,  (c) Manor Beach, (d) Triplets.  Numbers below bars represent fish species (see Appendix D).   38 Lionfish showed PC values comparable to the only similarly sized native carnivore at the study sites.  The redhind Epinephelus guttatus is a small-bodied grouper that colonized experimental reefs at all sites.  E. guttatus was also similar to lionfish in that it was locally rare.  E. guttatus was initially detected at only one site (Triplets), where its mean abundance was less than one.  Like lionfish, E. guttatus consistently occurred in or near the tail of the PC distribution (Figs. 2.4).  A planned comparison revealed that lionfish PC was not significantly different from E. guttatus (Wilcoxon rank sum, Coral Harbour: P = 1, Manor Beach: P = 1).  Interestingly, at sites that lionfish failed to invade (Adelaide and Triplets), E. guttatus appeared further in the tail of the PC distribution relative to sites in which both E.guttatus and lionfish had colonized (Figs. 2.4).  There was a positive but non-significant correlation in species-specific PC values among sites (Table 2.2).  In contrast, PC became consistent among sites once species were grouped according to families.  One test revealed a highly significant, positive correlation in fish family PC among sites (Table 2.2 & Fig. 2.5c).  The remaining site comparisons were also positively correlated but marginally non-significant at " = 0.05 (Table 2.2 & Figs. 2.5a and 2.5b).  Manor Beach was excluded from site comparisons because estimates were based on one month of reef surveys while those from all other sites were based on five to six months.      39 Table 2.2.  Results of Pearson correlation tests for association in colonization rate at the species and family levels between pairs of sites.  AD = Adelaide, CH = Coral Harbour, TR = Triplets.  Manor Beach was excluded from analyses because colonization rates were based on one month of census data while all other sites were based on five to six months of observations.  ***P<0.001; †0.05 < P < 0.15, marginally non-significant. Site comparison t df r P Species-level AD vs. CH 1.196 17 0.279 0.248 AD vs. TR 1.571 13 0.399 0.140 CH vs. TR 1.265 17 0.293 0.223 Family-level AD vs. CH 2.22 7 0.643 0.062† AD vs. TR 1.655 7 0.530 0.142† CH vs. TR 5.718 8 0.896 <0.001***   40  Figure 2.5 There was a consistent, positive trend in association between log transformed per capita probability of colonization (PC) of fish families among sites, which ranged from marginally non-significant (0.05 < P < 0.15) to highly significant (Pearson correlation; Table 2.2).  Data for only those families common to both sites being compared were analyzed. (a) Adelaide vs. Coral Harbour (r = 0.643, P = 0.062), (b) Adelaide vs. Triplets (r = 0.530, P = 0.142), (c) Triplets vs. Coral Harbour (r =  0.896, P <0.001). Manor Beach was excluded from comparisons because estimates were based on one month of surveys while those from all other sites were based on five to six months.  2.4  Discussion  Patterns of distribution indicate that patch reefs are invasible by lionfish, but that artificial structures facilitate colonization into marginal sand-seagrass and to a lesser extent, into hard bottom.  Contrary to my expectation, areas that had been modified by adding artificial structures did not contain more lionfish than less modified sites.  Also  41 unexpected was the finding that lionfish abundances were similar among different habitats.  In its native range of the Indo-Pacifc, lionfish have been documented to occur in all three habitats (Schultz 1986). However, P. volitans is a reef-associated fish that frequently occupies sheltered sites (Morris & Akins 2009; Schultz 1986).  Structured habitats are also generally more likely to concentrate higher densities of forage fishes than sandy substrates.  I therefore expected patch reefs and hard bottoms to be better- quality habitats than sand-seagrass and thus, harbor more individuals.  Although abundances were comparable among habitats, lionfish tended to occupy artificial structures in sand-seagrass while they most often occurred on natural patch reefs and hard bottoms when available.  Estimates of lionfish abundances and densities are reported in the literature.  However, I cannot infer whether lionfish are more likely to occur in one habitat over another based on these studies because they all occurred at much greater depths than the present study (12 - 100 m vs. < 5 m) and focused on only one habitat type (coral reefs or rocky bottoms combined with artificial substrate).  Fishelson (1997) and Green and Côté (2009) estimated lionfish densities exclusively on coral reefs in the Red Sea and the Bahamas, respectively, while Morris and Whitfield (2009) and Whitfield et al. (2007) grouped abundances on rocky bottom and artificial structures into a single category along the US Atlantic coast.  Lionfish likely arrived in the Bahamas via natural dispersal of pelagic eggs and larvae on ocean currents from neighboring US waters (Freshwater et al. 2009).  Systematic  42 differences in propagule supply between habitats are therefore unlikely to account for the high lionfish abundance in suboptimal sand-seagrass habitat reported here. However, the pattern of human activity along coastlines and the proportion of lionfish found on artificial structures among invaded habitats allowed me to propose a possible explanation for the lack of a detectable difference in lionfish abundance between types of modification and habitat.  New Providence encompasses the capital city of the Bahamas and is the most densely populated in the archipelago, where it is home to approximately 69% of the country’s human population (Bahamas Dept. of Statistics 2000).  Given the demands of human settlements, it is not surprising that all hard bottom and sand-seagrass sites in this study were modified.  In contrast, only one of the patch reef sites contained artificial structures.  Coral reefs provide natural shoreline protection (Pendleton 1995), therefore lessening the need for coastal defense structures like jetties and seawalls.  Similarly, the economic benefits of coral reefs (fisheries and dive tourism) likely reduce the propensity to build docks and illegally dump in these areas.  Patch reefs therefore tended to occur in areas with relatively little human modification.  In invaded hard bottom sites, 25% of lionfish occurred on artificial structures.  By comparison, nearly all lionfish occurred directly on artificial substrate (mean = 96%) in sand-seagrass and none in patch reefs.  This pattern suggests that artificial structures represent a resource subsidy to lionfish and that sand-seagrass is marginal habitat but artificial substrates allow it to support similar lionfish abundances as those found on  43 coral reefs and hard bottoms.  The fluctuating resource availability hypothesis predicts that an environment is most susceptible to invasion when there is an increase in a resource that limits both resident and invading species (Davis et al. 2000; Davis & Pelsor 2001).  Structural habitat may be more limiting for lionfish in sand-seagrass than on hard bottoms or patch reefs.  Hence, adding artificial structures in marginal reef fish habitats could facilitate invasion.  There was no difference in lionfish abundance between sites that had artificial substrates and those that did not.  However, sites contained a variety of natural habitats.  This could not be avoided due to the bias in the types of habitat that humans alter.  I have provided evidence supporting the view that sand-seagrass is marginal habitat for lionfish but adding artificial structures enables it to support similar lionfish abundances as those found in more structurally complex, higher-quality natural habitats. The similarity in lionfish abundance among sites of varying human impact is therefore likely due to the presence of the most natural structure (coral reefs) in sites of least human impact.  My artificial reef experiment indicates that lionfish are not fugitive species with higher capacity to colonize empty habitat relative to natives.  Various traits are likely to be advantageous at some, but rarely all, stages of the invasion process (Kolar & Lodge 2001, 2002).  Nevertheless, species characteristics suggested to be important determinants of invader success include: diet or habitat generalists (Moyle & Light 1996a, b; Moyle & Marchetti 2006), high fecundity (Lodge 1993), and rapid dispersal  44 under some conditions (Moyle & Marchetti 2006; Shea & Chesson 2002).  I predicted that lionfish would be superior colonizers relative to native taxa.  If this is the case, lionfish could gain a competitive advantage during the spread phase of invasion.  My data show that lionfish are slow to colonize artificial reefs, with a comparable PC as the most ecologically similar native species in this study.  On one hand, in open systems like reef fish communities, larval recruitment from the plankton onto reefs tends to vary unpredictably in space and time (Sale et al. 1984).  On the other hand, I suggest that species traits such as home range size, swimming ability, pelagic larval duration and fecundity may constrain colonization rates in relatively predictable ways.  I examined the combined effect of stochastic and deterministic mechanisms on colonization rate by comparing the PC of lionfish and native species to experimental reefs.  I observed significant among-species variation in PC at all sites.  However, lionfish colonized experimental reefs at only two sites (Manor Beach and Coral Harbour).  I observed lionfish in the study area at Manor Beach prior to the deployment of artificial reefs, where experimental units were surveyed for one month.  By comparison, I did not detect any lionfish in the study area at Coral Harbour during the week of reef deployment, where units were surveyed for five months.  Despite these differences, the pattern of invasion was the same: lionfish occurred in the tail of the PC distribution. Because lionfish did not colonize reefs at Triplets or Adelaide, I assumed that lionfish PC for the study period was effectively zero at these sites. Taken together, this provides strong evidence that the rapid range expansion of lionfish in the Atlantic is not due to superior colonizing ability.  45  Although there was a trend toward positive correlation between species PC in all combinations of site comparisons, none was significant. This suggests that species colonization rate largely depends on environmental context or stochastic processes.  In contrast, PC was more consistent among sites once species were grouped according to families.  In this instance, all site comparisons were positively correlated and ranged from highly significant to marginally non-significant.  Families that were good colonizers at one site also tended to arrive quickly at all other examined sites.  This implies that there is some degree of predictability in fish family colonization rates at an island-wide scale.  My data suggest that superior colonizing ability does not account for the ongoing natural range expansion of lionfish in the Atlantic, “one of the most rapid marine finfish invasions in history” (Morris et al. 2009).  Lionfish have specialized bilateral swim bladder muscles that enable fine-scale control of position in the water column, potentially enhancing access to prey in structurally complex habitats (Hornstra et al. 2004).  Furthermore, lionfish employ a novel predation strategy in the Atlantic by corralling prey with enlarged, fan-like pectoral fins (Albins & Hixon 2008).  An alternative explanation for lionfish success is that they are relatively slow at colonizing, but once they arrive, they are better able to defend or displace natives from prime microhabitat and that they benefit from a superior predation tactic once established.  Evidence of a similar situation has been documented in Hawaii by Schumacher and Parrish (2005) who found that due to asymmetrical competition for shelter on reefs, the introduced  46 blueline snapper Lutjanus kasmira displaced the native but competitively subordinate yellowtail goatfish Mulloidichthys vanicolensis higher into the water column where they may be more susceptible to predation.  Two important management implications emerge from my study.  First, lionfish can invade patch reefs and hard bottoms; however, artificial structures are the main habitat in marginal sand-seagrass areas.  Due to the pattern of human activity along coastlines, lesser-quality natural habitat supported similar lionfish abundances as more structurally complex, higher-quality habitats.  Second, despite lionfish being relatively poor colonizers, artificial structures likely facilitated their invasion into sand-seagrass and to a lesser extent, into hard bottom, where the availability of suitable substrate may be limiting.  This has implications for lionfish local persistence and rates of regional spread. Sand-seagrass may naturally function as sink habitats for lionfish in which local populations are maintained by continued migration from more productive sources such as coral reefs.  Adding artificial structures may promote a transition from a sink to a source habitat in which the formation of self-sustaining populations allows for the net export of surplus individuals to new areas.  Future investigations into lionfish demographic rates between different habitats are therefore urgently needed to explore this possibility.  Nonetheless, given the current pattern of lionfish distribution, removing or preventing the dumping of debris may slow the rate of spread, but is unlikely to prevent their expansion.    47 2.5  References  Albins, M.A. & Hixon, M.A.  (2008)  Invasive Indo-Pacific lionfish Pterois volitans reduce  recruitment of Atlantic coral-reef fishes.  Marine Ecology Progress Series, 367,  233-238. Bahamas Dept. of Statistics (2000)  Bahamas Department of Statistics.   "http://statistics.bahamas.gov.bs" Cited 20 April 2010. Bulleri, F.  (2005)  The introduction of artificial structures on marine soft- and hard-  bottoms: ecological implications of epibiota.  Environmental Conservation, 32,  101-102. Bulleri, F. & Airoldi, L.  (2005)  Artificial marine structures facilitate the spread of a  nonindigenous green alga, Codium fragile spp. tomentosoides, in the north  Adriatic Sea.  Journal of Applied Ecology, 42, 1063-1072. Cohen, A.N. & Carlton, J.T. (1998)  Accelerating invasion rate in a highly invaded estuary.  Science, 279, 555-558. Davis, M.A., Grime, J.P. & Thompson, K.  (2000)  Fluctuating resources in plant  communities: a general theory of invasibility.  Journal of Ecology, 88, 528-534. Davis, M.A. & Pelsor, M.  (2001)  Experimental support for a resource-based  mechanistic model of invasibility.  Ecology Letters, 4, 421-428. Elton, C. (1958)  The Ecology of Invasions by Animals and Plants.  University of  Chicago Press, Chicago.  48 Fishelson, L.  (1997)  Experiments and observations on food consumption, growth and  starvation in Dendrochirus brachypterus and Pterois volitans (Pteroinae,  Scorpaenidae).  Environmental Biology of Fishes, 50, 391-403. Freshwater, D.W., Hines, A., Parham, S., Wilbur, A., Sabaoun, M., Woodhead, J., Akins, L., Purdy, B., Whitfield, P.E. & Paris, C.B.  (2009)  Mitochondrial control region sequence analyses indicate dispersal from the US East Coast as the source of the invasive Indo-Pacific lionfish Pterois volitans in the Bahamas. Marine Biology, 156, 1213-1221. Glasby, T.M. & Connell, S.D.  (1999)  Urban structures as marine habitats.  Ambio, 28,  595-598. Glasby, T.M., Connell, S.D., Holloway, M.G. & Hewitt, C.L.  (2007)  Nonindigenous biota  on artificial structures: could habitat creation facilitate biological invasions?  Marine Biology, 151, 887-895. Green, S.J. & Côté, I.M.  (2009)  Record densities of Indo-Pacific lionfish on Bahamian  coral reefs.  Coral Reefs, 28, 107. Hamner, R.M., Freshwater, D.W. & Whitfield, P.E.  (2007)  Mitochondrial cytochrome b  analysis reveals two invasive lionfish species with strong founder effects in the  western Atlantic.  Journal of Fish Biology, 71, 214-222. Hixon, M.A. & Beets, J.P.  (1989)  Shelter characteristics and Caribbean fish  assemblages: experiments with artificial reefs.  Bulletin of Marine Science, 44,  666-680. Hornstra, H.M., Herrel, A. & Montgomery (2004)  Gas bladder movement in lionfishes: a  novel mechanism for control of pitch.  Journal of Morphology, 260, 299-300.  49 Kolar, C.S. & Lodge, D.M. (2001)  Progress in invasion biology: predicting invaders.  Trends in Ecology and Evolution, 16, 199-204. Kolar, C.S. & Lodge, D.M. (2002)  Ecological predictions and risk assessment for alien  fishes in North America.  Science, 298, 1233-1236. Lockwood, J.L., Hoopes, M.F. & Marchetti, M.P.  (2007)  Invasion Ecology.  Blackwell  Publishing, Oxford. Lodge, D.M. (1993)  Biological invasions: lessons for ecology.  Trends in Ecology and  Evolution, 8, 133-136. Morris, J.A. & Akins, J.L. (2009)  Feeding ecology of invasive lionfish (Pterois volitans)  in the Bahamian archipelago.  Environmental Biology of Fishes, 86, 389-398. Morris, J.A., Akins, J.L., Barse, A., Cerino, D., Freshwater, D.W., Green, S.J., Muñoz,  R.C., Paris, C. & Whitfield, P.E.  (2009)  Biology and ecology of the invasive  lionfishes, Pterois miles and Pterois volitans.  Proceeding of the 61st Gulf and  Caribbean Fisheries Institute Conference, 61,1-6. Morris, J.A. & Whitfield, P.E. (2009)  Biology, ecology, control and management of the  invasive Indo-Pacifc lionfish: an updated integrated assessment.  NOAA  Technical Memorandum NOS NCCOS, 99, 1-57. Moyle, P.B. & Light, T.  (1996a)  Fish invasions in California: do abiotic factors determine success?  Ecology, 77, 1666-1670. Moyle, P.B. & Light, T.  (1996b)  Biological invasions of freshwater: empirical rules and  assembly theory.  Biological Conservation, 78, 149-161. Moyle, P.B. & Marchetti, M.P.  (2006)  Predicting invasion success: freshwater fishes in  California as a model.  Bioscience, 56, 515-523.  50 Munday, P.L.  (2004)  Competitive coexistence of coral-dwelling fishes: the lottery hypothesis revisited.  Ecology, 85, 623-628. Pendleton, L.H.  (1995)  Valuing coral reef protection.  Ocean and Coastal Management, 26, 119-131. Ruiz, G.M., Freestone, A.L., Fofonoff, P.W. & Simkanin, C.  (2009)  Habitat distribution and heterogeneity in marine invasion dynamics: the importance of hard substrate and artificial structure.  Marine Hard Bottom Communities: Patterns, Dynamics, Diversity and Change (ed.  Wahl, M.),  pp. 321-332. Springer, Heidelberg. Sale, P.F.  (1978)  Coexistence of coral reef fishes - a lottery for living space. Environmental Biology of Fishes, 3, 85-102. Sale, P.F., Doherty, P.J., Eckert, G.J., Douglas, W.A. & Ferrell, D.J.  (1984)  Large scale spatial and temporal variation in recruitment to fish populations on coral reefs. Oecologia, 64, 191-198. Sale, P.F. & Douglas, W.A. (1981)  Precision and accuracy of visual census technique  for fish assemblages on coral patch reefs.  Environmental Biology of Fishes, 6, 333-339. Sapota, M.R. & Skóra, K.E.  (2005)  Spread of alien (non-indigenous) fish species  Neogobius melanostomus in the Gulf of Gdansk (south Baltic).  Biological Invasions, 7, 157-164. Schmitt, E.F., Sluka, R.D. & Sullivan-Sealey, K.M.  (2002)  Evaluating the use of roving  diver and transect surveys to assess the coral reef fish assemblage off southeastern Hispaniola.  Coral Reefs, 21, 216-223.  51 Schofield, P.J.  (2009)  Geographic extent and chronology of the invasion of non-native  lionfish (Pterois volitans [Linnaeus 1758] and P. miles [Bennett 1828]) in the  Western North Atlantic and Caribbean Sea.  Aquatic Invasions, 4, 473-479. Schultz, E.T. (1986)  Pterois volitans and Pterois miles: two valid species.  Copeia,  1986, 686-690. Schumacher, B.D. & Parrish, J.D.  (2005)  Spatial relationships between an introduced  snapper and native goatfishes on Hawaiian reefs.  Biological Invasions, 7,  925- 933. Semmens, B.X., Buhle, E.R., Salomon, A.K. & Pattengill-Semmens, C.V.  (2004)  A  hotspot of non-native marine fishes: evidence for the aquarium trade as an  invasion pathway.  Marine Ecology Progress Series, 266, 239-244. Shea, K. & Chesson, P.  (2002)  Community ecology theory as a framework for  biological invasions.  Trends in Ecology and Evolution, 17, 170-176. Small, C. & Nicholls, R.J.  (2003)  A global analysis of human settlement in coastal  zones.  Journal of Coastal Research, 19, 584-599. Stachowicz, J.J., Whitlatch, R.B. & Osman, R.W.  (1999)  Species diversity and invasion  resistance in a marine ecosystem.  Science, 286, 1577-1579. Steele, M.A.  (1999)  Effects of shelter and predators on reef fishes.  Journal of  Experimental Marine Biology and Ecology, 233, 65-79. Sutherland, W.J., Clout, M., Côté, I.M., Daszak, P., Depledge, M.H., Fellmen, L., Fleishman, E., Garthwaite, R., Gibbons, D.W., De Lurio, J., Impey, A.J., Lickorish, F., Lindenmayer, D., Madgwick, J., Margerison, C., Maynard, T., Peck, L.S., Pretty, J., Prior, S., Redford, K.H., Scharlemann, J.P.W., Spalding, M. &  52 Watkinson, A.R.  (2010)  A horizon scan of global conservation issues for 2010. Trends in Ecology and Evolution, 25, 1-7. Tyrrell, M.C. & Byers, J.E.  (2007)  Do artificial substrates favor nonindigenous fouling  species over native species?  Journal of Experimental Marine Biology and  Ecology, 342, 54-60. Whitfield, P.E., Gardner, T., Vives, S.P., Gilligan, M.R., Courtenay, W.R., Ray, G.C. &  Hare, J.A.  (2002)  Biological invasion of the Indo-Pacific lionfish Pterois volitans  along the Atlantic coast of North America.  Marine Ecology Progress Series, 235,  289-297. Whitfield, P.E., Hare, J.A., David, A.W., Harter, S.L., Muñoz, R.C. & Addison, C.M.  (2007)  Abundance estimates of the Indo-Pacific lionfish Pterois volitans/miles  complex in the Western North Atlantic, Biological Invasions, 9, 53-64.            53 Chapter 3  Conclusion  3.1  Contributions to invasion biology  Attempts to identify traits that predict which species will become invasive have produced mixed results, depending on the taxonomic level of analysis.  For instance, Kolar and Lodge (2002) developed a quantitative model of risk assessment based on a suite of organismal attributes that correctly predicted invasion success (with 87 to 94% accuracy) for attempted fish introductions into the North American Great Lakes.  By comparison, Jeschke and Strayer (2006) examined 20 factors in several vertebrate groups (freshwater fishes, birds and mammals) but found that only two (propagule pressure and human affiliation) were strong determinants of invasion for all taxa.  Identification of attributes that confer invasiveness and are applicable across many broadly defined taxonomic groups remains elusive, despite more than a century of effort (e.g. Gray 1879, Henslow 1879, Moyle and Marchetti 2006, Kolar and Lodge 2001). Indeed, the search has been referred to as the “Holy Grail of invasion biology” (Enserink 1999) and some argue that “such overarching characteristics do not exist” (Kolar and Lodge 2002).   54 Part of the challenge in pinpointing biological characteristics that predict invasion (even within narrowly defined taxonomic groups) is that: (1) different properties of the invader are important at different stages of invasion (Kolar and Lodge 2001, 2002, Moyle and Marchetti 2006), and (2) it is often difficult to obtain data on species that failed to establish (Kolar and Lodge 2001).  It was not the aim of my research to determine species traits that predict invasiveness.  However, I was interested in a closely related issue.  I wanted to know if the success of an exotic in the recipient community could be explained by its superior performance relative to native species (as opposed to relative to introduced taxa that failed to establish) in a trait that is important at multiple stages of invasion.  My research tested a common assumption in invasion biology that invaders (typically from more species-rich donor biotas) are superior to natives in traits related to fitness in the invaded region (Daehler 2003, Moyle and Marchetti 2006, Sax and Brown 2000, Shea and Chesson 2002, Vermeij 2005).  More specifically, I examined whether Indo-Pacific lionfish are superior colonizers relative to Atlantic reef fishes.  Successful colonization is a prerequisite for both the establishment and spread of a species, whether native or exotic.  A fish must first arrive at a suitable location before it can establish a self-sustaining population (particularly in cases where priority effects are important, e.g. Almany 2003, Sale 1978), and must be an effective disperser to spread throughout a region.  Contrary to my expectations, I found that lionfish were slower to colonize experimental reefs than most native Atlantic fishes.  This suggests that superior colonization ability does not explain the rapid establishment and expansion of lionfish throughout the western Atlantic.  The assumption of superior performance of  55 invaders relative to natives was therefore not supported by my research (at least in regard to the examined trait in shallow, nearshore waters).  This finding is consistent with a comparative study by Daehler (2003), which showed that invasive plants rarely out-performed natives in fitness-related traits such as growth, dispersal and fecundity. My results highlight the difficulty of identifying traits associated with invasiveness, and indicate that complex interactions between the characteristics of species and those of recipient communities determine which species are likely to flourish in novel environments.  A second contribution of my research is that it provided observational evidence in support of the widespread view that humans can promote invasions via landscape/seascape-scale habitat alterations (which favor non-indigenous species). Although lionfish were slower to colonize artificial structures than most native taxa, my research suggests that the presence of such structures can nonetheless facilitate invasion into marginal habitats by acting as a resource subsidy to lionfish.  Artificial structures can serve as foci for the establishment and spread of non-indigenous, sessile, marine epibiota (Ruiz et al. 2009).  My research showed that similar effects can occur in mobile taxa like reef fishes.  3.2  Study strengths and limitations  In order to test the hypothesis that lionfish are superior colonizers relative to Atlantic taxa, I first quantified their colonization ability.  In my study, the rate of colonization is  56 expected to be proportional to local species abundances because more numerous taxa are more likely, by chance, to be in close proximity to experimental reefs.  I therefore derived a formula from first principles of probability theory, which allowed me to calculate the likelihood that an individual present at the study site colonized an experimental reef per day, given its species local abundance (i.e. per capita probability of colonization).  The development of a quantitative measure of colonization that accounts for natural variation in species abundances in the field was certainly a novel aspect of my research.  However, the real strength of this new development is that it can be broadly applied to dispersal studies involving a wide range of taxa, not just reef fishes.  For example, seasonal migration times for bird populations are commonly measured as the day that the first individual of a species is observed at a site, but do not incorporate migratory cohort size, which can also influence first arrival date (Miller-Rushing et al. 2008, Tryjanowski et al. 2005).  In a study of bird migratory responses to climate change, Miller-Rushing et al. (2008) found that declining bird population sizes over time largely accounted for some of the documented shifts in species migration times.  They concluded that previous studies on the subject need to be reinterpreted in light of the effect of temporal variation in population sizes on estimates of migration times, as some of the documented migratory changes may be real while others may be due to a methodological artifact (Miller-Rushing et al. 2008).  The per capita probability of  57 colonization formula developed in my thesis can potentially be applied to studies such as these.  A second strength of my research was the considerable replication involved (both within and among sites) in the manipulative field experiment, which compared colonization ability of lionfish relative to native taxa.  This not only increased the power of statistical analyses but more importantly, it allowed me to make generalizations about my findings that transcended the particularities of an individual site.  For example, lionfish did not colonize experimental reefs at two sites while they appeared in the tail of the distribution of species per capita colonization probabilities at invaded sites, despite differences in sampling effort and local lionfish abundances at the start of the experiment.  Taken together, this provides strong evidence that lionfish are not superior to Atlantic reef fishes as colonizers of artificial substrate in nearshore waters at a regional (island-wide) scale.  There was also a cost to the high level of replication in my study.  Substantial amounts of time and labour were associated with the field component of my research.  The large number of sites (four in total) and experimental reefs (60 in total) severely constrained: (1) the number of visual surveys that I could conduct in order to estimate local species abundances prior to reef deployment, and (2) the frequency with which I could sample reefs post deployment.  However, it was not the goal of my research to determine precise estimates of local species abundances.  Furthermore, although sampling effort varied somewhat between sites, it was fairly consistent within a site.  Because  58 colonization comparisons (both qualitative and quantitative) were generally restricted to within a site, there is no reason to suspect that these limitations would bias my results or change the general conclusions made here.  3.3  Management implications  The Indo-Pacific lionfish invasion of the Atlantic was identified as one of fifteen emerging global conservation issues this year (Sutherland et al. 2010).  Lionfish pose a potential threat to public health (due to its venomous spines, Vetrano et al. 2002) as well as to native biodiversity and commercial fisheries (due to predation on and competition with resident taxa, Albins and Hixon 2008, Whitfield et al. 2007).  Once an invader has established, as is the case of lionfish throughout much of the western Atlantic (Schofield 2009), it is very difficult if not impossible to eradicate (Elton 1958, Lockwood et al. 2007).  Hence, invasion management at this stage is typically devoted to control efforts (Lockwood et al. 2007).  My study has implications for the control of lionfish in particular, as well as for the management of coastlines, in general.  My findings suggest that lionfish are capable of invading natural, coral patch reefs in the absence of artificial structures, but that the presence of such structures facilitates colonization of marginal nearshore habitats like sand-seagrass and to a lesser extent, hard bottoms.  There is little that one can do to limit lionfish expansion into coral reef systems and otherwise marginal, natural habitats with existing coastal infrastructure (e.g. docks and jetties), except for long-term  59 monitoring programs and subsequent culling of lionfish (which in many instances, may not be feasible).  One can, however, potentially slow the spread of lionfish into marginal habitats littered with hard land-based debris.  In this instance, removal of artificial substrates (and stricter enforcement that prevents coastal dumping) may reduce the carrying capacity (and thus local abundances) of all reef fishes that rely on structural habitat, including lionfish.  However even then, efforts are unlikely to prevent lionfish expansion into these areas.  In more general terms, the proliferation of anthropogenic substrates along coastlines is occurring worldwide as the human population continues to increase and landscape/seascape-scale modifications to natural habitat intensify (Bulleri 2005, Bulleri and Airoldi 2005, Glasby and Connell 1999).  Artificial structures are intentionally added to coastal waters for a variety of reasons: (1) to provide shoreline protection (a demand that is on the rise due to global climate change, Bulleri and Airoldi 2005), (2) to enhance dive tourism or fisheries production, (3) to “mitigate” habitat loss, (4) to prevent trawling, and (5) for recreational boating and industrial purposes (e.g. construction of marinas and ports, see Bain 2001 for a review of the multiple uses of artificial marine structures). Given the plethora of emerging evidence that human-made marine substrates can facilitate invasions (see chapter 2 of this thesis, Bulleri and Airoldi 2005, Glasby et al. 2007, Ruiz et al. 2009, Tyrrell and Byers 2007), there is a need now more than ever for society to critically re-evaluate the unintended environmental (and economic) costs of such activities.   60 3.4  Future research  My research revealed that although lionfish abundances were similar among nearshore habitats dominated by coral patch reefs, hard bottoms, and sand-seagrass, lionfish tended to occupy artificial structures in sand-seagrass while they most often occurred on natural patch reefs and hard bottoms when available.  I also provided a potential explanation for this pattern: the availability of natural structural habitat may be more limiting in sand-seagrass than in the other two examined habitat types.  Artificial substrates may thus represent a resource subsidy to lionfish in marginal natural, nearshore habitats, thereby allowing these areas to support similar lionfish abundances as more structurally complex, higher-quality coral patch reefs.  What is lacking from my study is information on lionfish density, age and size distributions as well as other important demographic features (e.g. birth and death rates) between different nearshore habitat types (both with and without artificial structures).  Knowledge of the above would allow one to more confidently determine whether coral patch reefs are indeed high-quality natural habitat for lionfish and the extent to which the addition of artificial structures affects population parameters other than abundance.  My research also showed that although lionfish were consistently poor colonizers of experimental reefs relative to Atlantic taxa in nearshore waters, their colonization rate was not different from the most ecologically similar native species in my study (the  61 small-bodied grouper, Epinephelus guttatus).  Lionfish are most likely to compete with functionally similar native species for food and space resources in the Atlantic. Investigations into the performance of lionfish relative to several native species of small- bodied groupers in other fitness-related traits such as fecundity and predation rate and efficiency would allow scientists to better predict the likely impacts of lionfish on native biota.                   62 3.5  References  Albins, M.A. & Hixon, M.A.  (2008)  Invasive Indo-Pacific lionfish Pterois volitans reduce  recruitment of Atlantic coral-reef fishes.  Marine Ecology Progress Series, 367,  233-238. Almany, G.  (2003)  Priority effects in coral reef fish communities.  Ecology, 84, 1920- 1935. Baine, M.  (2001)  Artificial reefs: a review of their design, application, management and  performance.  Ocean and Coastal Management, 44, 241-259. Bulleri, F.  (2005)  The introduction of artificial structures on marine soft- and hard-  bottoms: ecological implications of epibiota.  Environmental Conservation, 32,  101-102. Bulleri, F. & Airoldi, L.  (2005)  Artificial marine structures facilitate the spread of a  nonindigenous green alga, Codium fragile spp. tomentosoides, in the north  Adriatic Sea.  Journal of Applied Ecology, 42, 1063-1072. Daehler, C.C.  (2003)  Performance comparisons of co-occuring native and alien  invasive plants: implications for conservation and restoration.  Annual Review of  Ecology, Evolution and Systematics, 34, 183-211. Elton, C. (1958)  The Ecology of Invasions by Animals and Plants.  University of  Chicago Press. Enserink, M.  (1999)  Biological invaders sweep in.  Science, 285, 1834-1836. Glasby, T.M. & Connell, S.D.  (1999)  Urban structures as marine habitats.  Ambio, 28,  595-598.  63 Glasby, T.M., Connell, S.D., Holloway, M.G. & Hewitt, C.L.  (2007)  Nonindigenous biota  on artificial structures: could habitat creation facilitate biological invasions?  Marine Biology, 151, 887-895. Gray, A.  (1879)  The pertinacity and predominance of weeds.  American Journal of  Science and Arts, 18, 161-167. Henslow, G.  (1879)  On the self-fertilization plants.  Linnaen Society Transaction Series  II, 1, 317-389. Jeschke, J.M. & Strayer, D.L.  (2006)  Determinants of vertebrate invasion success in  Europe and North America.  Global Change Biology, 12, 1608-1619. Kolar, C.S. & Lodge, D.M. (2001)  Progress in invasion biology: predicting invaders.  Trends in Ecology and Evolution, 16, 199-204. Kolar, C.S. & Lodge, D.M. (2002)  Ecological predictions and risk assessment for alien  fishes in North America.  Science, 298, 1233-1236. 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All transient predators were excluded because they are not considered to be reef residents. Because of uncertainty in detection likelihood, all cryptic species and taxa that use crevices during the day or burrow in sand were also excluded.  AD = Adelaide, CH = Coral Harbour, MB = Manor Beach, TR = Triplets.  Reason for exclusion of taxon: crevice = diurnal use of crevices; cryptic = cryptic taxa; sand = burrows in sand; transient = roving predator. Site Family Species Reason AD Apogonidae Apogon maculatus crevice   A. sp. crevice  Carangidae Caranx bartholomaei transient   C. ruber transient  Holocentridae Holocentrus adscensionis crevice  67 Site Family Species Reason   Sargocentron coruscum crevice   Labrisomidae Malacoctenus macropus cryptic CH Apogonidae Apogon maculatus crevice   A. sp. crevice  Carangidae Caranx bartholomaei transient   C. ruber transient  Holocentridae Holocentrus adscensionis crevice    Sargocentron coruscum crevice  Labrisomidae Malacoctenus macropus cryptic  Ophichthidae Myrichthys ocellatus sand  68 Site Family Species Reason  Sphyraenidae Sphyraena barracuda transient MB Apogonidae Apogon sp. crevice   Bothidae Bothus lunatus cryptic   B. ocellatus cryptic  Carangidae Caranx bartholomaei transient   C. ruber transient  Gobiidae Coryphopterus glaucofraenum cryptic   Ctenogobius saepepallens cryptic  Holocentridae Holocentrus adscensionis crevice   Holocentrus rufus crevice   Myripristis jacobus crevice  69 Site Family Species Reason   Sargocentron coruscum crevice  Labrisomidae Malacoctenus macropus cryptic  Scorpaenidae Scorpaena plumieri cryptic  Serranidae Diplectrum formosum crevice  Synodontidae Synodus intermedius sand TR Apogonidae Apogon maculatus crevice   Apogon sp. crevice  Bothidae Bothus ocellatus cryptic  Carangidae Caranx bartholomaei transient   C. ruber transient  Gobiidae Coryphopterus glaucofraenum cryptic  70 Site Family Species Reason  Holocentridae Holocentrus adscensionis crevice   Holocentrus rufus crevice   Sargocentron coruscum crevice  Labrisomidae Malacoctenus macropus cryptic  Synodontidae Synodus intermedius sand   Synodus sp. sand  71 4.2  Appendix B  Taxa that colonized experimental reefs but were not detected in local abundance surveys during the week of reef deployment.  It was assumed that these species either: (1) migrated from an area greater than that defined by the study; or, (2) were so rare that they were undetected in local surveys. These taxa were arbitrarily assigned a local abundance of half of the least abundant species observed during site specific surveys so that PC could be estimated.  AD = Adelaide, CH = Coral Harbour, MB = Manor Beach, TR = Triplets. Site Family Species AD Acanthuridae Acanthurus coeruleus  Chaetodontidae Chaetodon capistratus   C. ocellatus  Haemulidae Haemulon sp.  Lutjanidae Lutjanus griseus   L. synagris   Ocyurus chrysurus  Mullidae Mulloidichthys martinicus  Pomacentridae Abudefduf saxatilis  72 Site Family Species   Stegastes diencaeus  Sciaenidae Pareques acuminatus  Serranidae Epinephelus adscensionis   E. guttatus CH Chaetodontidae Chaetodon capistratus  Lutjanidae Lutjanus griseus  Mullidae Mulloidichthys martinicus  Pomacanthidae Holacanthus ciliaris   Pomacanthus paru  Scaridae Scarus taeniopterus  Scorpaenidae Pterois volitans / miles  Serranidae Epinephelus guttatus  Tetraodontidae Canthigaster rostrata MB Acanthuridae Acanthurus chirurgus   A. coeruleus  73 Site Family Species  Serranidae Epinephelus guttatus  Sciaenidae Pareques acuminatus TR Acanthuridae Acanthurus bahianus   A. chirurgus   A. coeruleus  Chaetodontidae Chaetodon ocellatus   C. striatus  Labridae Thalassoma bifasciatum  Lutjanidae Lutjanus synagris   Ocyurus chrysurus  Pomacanthidae Pomacanthus arcuatus  Pomacentridae Stegastes diencaeus  Priacanthidae Heteropriacanthus cruentatus  Scaridae Sparisoma aurofrenatum  Sciaenidae Pareques acuminatus  74 Site Family Species  Serranidae Epinephelus striatus  Tetraodontidae Canthigaster rostrata  75 4.3  Appendix C  Taxa detected in local abundance surveys that were not observed on experimental reefs during the study period.  Species absences were presumably due to: (1) predation or other forms of mortality; or, (2) failure to colonize reefs during the study. AD = Adelaide, CH = Coral Harbour, MB = Manor Beach, TR = Triplets. Site Family Species AD Monacanthidae Monacanthus tuckeri  Mullidae Pseudupeneus maculatus  Ostraciidae Acanthostracion quadricornis  Pomacentridae Microspathodon chrysurus CH Labridae Halichoeres poeyi MB Balistidae Balistes vetula  Pomacentridae Stegastes partitus  Tetraodontidae Canthigaster rostrata TR Ostraciidae Acanthostracion quadricornis  76 Site Family Species  Scaridae Scarus taeniopterus  77 4.4  Appendix D  Species for which PC was calculated.  AD = Adelaide, CH = Coral Harbour, MB = Manor Beach, TR = Triplets.  Values in the graphic column correspond to the colonization rank assigned to taxa at each site in Figures 2.5a-d. Site Family Species Graphic AD Acanthuridae Acanthurus bahianus 8   A. chirurgus 2   A. coeruleus 6  Chaetodontidae Chaetodon capistratus 14   C. ocellatus 20  Haemulidae Haemulon spp. 22  Labridae Halichoeres bivittatus 15   H. maculipinna 13   H. poeyi 9   Thalassoma bifasciatum 16  Lutjanidae Lutjanus griseus 17   L. synagris 12  78 Site Family Species Graphic   Ocyurus chrysurus 3  Mullidae Mulloidichthys martinicus 19  Pomacentridae Abudefduf saxatilis 1   Stegastes diencaeus 5   S. leucostictus 11  Scaridae Sparisoma aurofrenatum 4   S. radians 10  Sciaenidae Pareques acuminatus 21  Serranidae Epinephelus adscensionis 7   E. guttatus 18 CH Acanthuridae Acanthurus bahianus 21   A. chirurgus 1   A. coeruleus 4  Chaetodontidae Chaetodon capistratus 10   C. ocellatus 5  79 Site Family Species Graphic   C. striatus 9  Haemulidae Haemulon spp. 20  Labridae Halichoeres bivittatus 29   H. maculipinna 13   H. radiatus 15   Thalassoma bifasciatum 28  Lutjanidae Ocyurus chrysurus 18   Lutjanus griseus 2   L. synagris 7  Mullidae Mulloidichthys martinicus 22   Pseudupeneus maculatus 17  Pomacanthidae Holacanthus ciliaris 27   Pomacanthus paru 11  Pomacentridae Abudefduf saxatilis 3  80 Site Family Species Graphic   Stegastes adustus 30   S. diencaeus 8   S. leucostictus 23  Scaridae Scarus iserti 14   S. taeniopterus 25   Sparisoma aurofrenatum 16   S. radians 19   S. viride 6  Scorpanidae Pterois volitans / miles 24  Serranidae Epinephelus guttatus 12  Tetraodontidae Canthigaster rostrata 26 MB Acanthuridae Acanthurus bahianus 9   A. chirurgus 3   A. coeruleus 6  Chaetodontidae Chaetodon capistratus 10  81 Site Family Species Graphic  Haemulidae Haemulon spp. 15  Labridae Halichoeres bivittatus 11   H. poeyi 8   Thalassoma bifasciatum 2  Lutjanidae Lutjanus synagris 4   Ocyurus chrysurus 5  Pomacanthidae Pomacanthus paru 14  Pomacentridae Stegastes leucostictus 17  Scaridae Scarus iserti 16   Sparisoma radians 13  Sciaenidae Pareques acuminatus 1  Scorpanidae Pterois volitans / miles 12  Serranidae Epinephelus guttatus 7 TR Acanthuridae Acanthurus bahianus 10   A. chirurgus 5  82 Site Family Species Graphic   A. coeruleus 3  Chaetodontidae Chaetodon capistratus 23   C. ocellatus 4   C. striatus 9  Haemulidae Haemulon spp. 19  Labridae Halichoeres bivittatus 16   Thalassoma bifasciatum 13  Lutjanidae Ocyurus chrysurus 2   Lutjanus synagris 1  Pomacanthidae Pomacanthus arcuatus 11   P. paru 14  Pomacentridae Stegastes diencaeus 20   S. leucostictus 7  Priacanthidae Heteropriacanthus cruentatus 6  83 Site Family Species Graphic  Scaridae Sparisoma aurofrenatum 12   S. radians 15   S. viride 22  Sciaenidae Pareques acuminatus 18  Serranidae Epinephelus guttatus 21   E. striatus 8  Tetraodontidae Canthigaster rostrata 17     

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