COMPARATIVE FORAGING DISTRIBUTION AND ECOLOGY SUGGESTS INTERSPECIFIC COMPETITION BETWEEN TWO SYMPATRIC SHEARWATERS FROM THE SEYCHELLES

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The “anthropocene”: an extinction era

At present, human beings are considered the dominant species on Earth (Vitousek et al. 1997). Anthropogenic actions can modify entire ecosystems and lead to the disappearance of the most vulnerable species and the less resilient communities (Peterson et al. 1998; Hoegh-Guldberg 2011). The negative impact of human beings is positively correlated to the humanity development (Steffen et al. 2011) and it started already in early history with the first appearance of Homo sapiens throughout the continents. In fact, first the change in climate and then the arrival of humans were most probably the cause of the late Quaternary megafaunal extinction (Prescott et al. 2012). However, anthropogenic actions started to deeply modify the planet with the XIX century industrial revolution and they grew exponentially after the World War II, a phase called by scientists “Great Acceleration” (Steffen et al. 2011). The first half of the XX century saw indeed at least 70 known species extinctions according to Cragg (1968a). However, to quantify the exact number of extinctions, information on the total number of species present on earth is necessary, a data which is at present still unknown (Wilson 1988). In any case, it is clear that human beings have been responsible for what is called the 6th mass extinction (Ceballos et al. 2015).
This long, and still ongoing, chain of extinctions was triggered and powered by the human perception of natural environments. In fact, since the human development begun, natural ecosystems and habitats were only seen like an unlimited resource, something to be fully exploited and that will never end. The idea of sustainable harvest lasted only in very isolated and rural community while the rest of the world started to produce and consume at a quick pace. Since the “Great Acceleration” started, the examples of vertebrate extinctions are many. The areas more threatened by extinctions are inevitably the most rich in biodiversity. It has been estimated that the number of species of a particular group of organisms in an island system (or habitat island) increases approximately as the fourth root of the land area (Wilson 1988). This calculation can help to estimate the biodiversity loss in many habitats. For example, the tropical world is heading towards an extreme reduction and fragmentation of tropical forests, a major hotspot for biodiversity.
Habitat fragmentation increases the insularity syndrome and some species – particularly the large ones that need greater extensions of territory – end up disappearing (Mc Arthur & Wilson 1967; Blondel 1986; Whittaker & Fernández-Palacios 2007). If the reduction of tropical forests continues at this pace it will be accompanied by a massive extinction of species (Wilson 1988; Brooks et al. 2002). The basic models of island biogeography are particularly suited to tropical forests as tropical species are more localized than the temperate ones (Helmus et al. 2014). Then, even if a portion of the species survives, these will probably have suffered significant reduction in genetic variation (Wilson 1998, (Helmus et al. 2014). For example three patches of subtropical forest in Brazil, varying from 0.2 to 14 square kilometers, were isolated by agricultural clearing for 100 years and their resident bird species suffered a 14 to 62% extinction rate (Willis 1979).
There are many other examples of extinctions, especially in islands hosting rare endemic or native species never exposed to predation pressure. These species are considered particularly vulnerable either because they are flightless birds, or/and because they have evolved in the absence of predators and have no defences against them (Milberg & Tyrberg 1993).
The biodiversity of the small islands in the Pacific and Indian Oceans started to be significantly affected by anthropogenic factors much earlier than its continental counterpart, after the beginning of the European maritime expansion in 1400. Island extinctions after 1600 are well documented and in many cases bird species have been lost from single islands or have become globally extinct as a result (Milberg and Tyrberg 1993). It is important to underline that ninety percent of the c.108 species thought to have become extinct since 1600 were restricted to islands (Johnson & Stattersfield 1990). The most famous example of islands extinctions is the case of the Dodo (Raphus cucullatus), a flightless bird living in the uninhabited island of Mauritius (Indian Ocean). When the firsts Dutch settlements started to widespread in 1638, the interference of introduced foreign animals (mainly cats and rats) together with the continued overuse of the birds for food led to their total extinction by the end of the century (Cheke & Hume 2008). But not only flightless land birds were affected by humans’ arrival in small islands. Procellariidae and Columbidae are the most frequent orders among the extirpations of island populations, based on fossil recoveries and historical evidences (Milberg & Tyrberg 1993). In fact, many seabird species that were breeding undisturbed in many oceanic islands have been extirpated or have become globally extinct after the arrival of humans. For some species,, the only elements documenting their presence are fossil evidences (Milberg and Tyrberg 1993). It is the case of two species of petrels (Pterodroma rupinarum and Bulweria bifax) breeding at St. Helena (to UK), which became extinct after humans colonized this island in 1502 (BirdLife international 2012). In general, the major anthropogenic factors affecting biodiversity are notoriously: habitat loss and fragmentation, introduced invasive species, pollution, overharvest, and lately the climate change (Mills 2012).

Conservation biology: definition and development

Although species declines and extinctions caused by humans started already in 1400 and intensified in 1800 (as documented above), awareness and willingness of the scientific community and the general public to protect natural ecosystems and species is a fairly recent feeling (Pfeffer et al. 2001).
As the cases of population reduction and species loss were increasing, around the years 1940s and 1950s conservation thoughts started to emerge. The first minds emerged as an “aesthetic concern for wilderness”, and when ecological studies and the concept of ecosystem started to develop, this “romantic era” became a more science-based movement known as conservation science (Cragg 1968). Since the beginning, conservation biology required a multidisciplinary approach, which involved social and cultural studies (Cragg, 1968b). The foundation of the International Union for the Conservation of Nature (IUCN) in 1948 and of its financial supporter the World Wild Fund for Nature (WWF) in 1961 can be considered as the starting point of the modern conservation awareness. At the beginning, the operations of the IUCN were mainly about exposing problems and persuading people, and the focus was on the species particularly closed to extinction (IUCN Red List). Nature was still very much seen as a benefit for human kind in an economic prospective (Cragg 1968a) but more protectionist thoughts started to emerge. Dasmann in 1968 and Erenfeld in 1970 introduced the concept of conservation biology. This new discipline was then described by Soulé (1985) as “a new synthetic discipline addressing the dynamics and problems of perturbed species, communities and ecosystems” and with time it became a multidisciplinary science developed to address the loss of biological diversity (Hunter 1996; Meffe 2006; van Dake 2008). After years of research the main goals of conservation biology have evolved and became: first to evaluate human impact on biological diversity and second to develop practical approaches to prevent the extinction of species (Soulé 1985; Wilson 1999). Conservation biology was described as a “mission-orientated crisis discipline”. In fact, in order to address pressing problems, conservationists have often to act fast without being completely comfortable with the theoretical and empirical bases of the analysis. Therefore tolerating uncertainty is often necessary although the principle of precaution needs to prevail.
According to Soulé (1985) conservation biology is supported by strong postulates both functional and normative. The functional postulates suggest the rules of action in order to maintain both, form and function of natural biological systems. They follow evolutionary, ecological, demographic and spatial rules regulating the ecosystem functioning and they need to be followed in order to achieve the subsistence of natural ecosystems. The normative postulates are value statements that create the ethic basis towards other forms of life. They cannot be tested or proven completely, but they provide the ethic upon which conservation decisions should be made.
After the milestone publication of Soulé (1985), that gives the basic definition and aim of conservation biology, other authors have enriched this new emerging discipline based on the experience accumulated during the firsts two decades of experiences in this field. It was in fact only after ten years from the definition of conservation biology that Caughley (1994) implemented this discipline with some important directions driven by the two major paradigms. The first is the small-population paradigm, it has a mainly theoretical approach and it is the center of all conservation actions carried out during the 1980s. It is centered on the notion that small populations are at major risk of extinction than big ones. In fact, a small number of individuals can lead to genetic and population dynamic problems, which in turn can bring the population into the extinction vortex (inbreeding depression combined with demographic stochasticity and genetic drift) (Lynch et al. 1995; Masel 2011).
The second paradigm addresses population decline and focuses on means for detecting, diagnosing and stopping it. This paradigm has a pragmatic origin and it is rooted in practical examples. It states that population declines have always one or more tangible causes that can be defeated with appropriate skills. It is therefore applicable mostly on a case-by-case basis and it lacks of a proper theory. Within this paradigm, four main agents of populations decline (called the “evil quartet”) can be identified: overkill, habitat destruction and degradation, impact of introduced species and chains of extinction (Pullin 2010).

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The ecological niche concept in a competition framework

The modern concept of ecological niche was proposed by George Evelyn Hutchinson (1957) and it is defined as a quantitative n-dimensions hypervolume constructed by the range of environmental features that enable a species to maintain a viable population indefinitely (Blonder et al. 2014). Independent axes, that have a biological meaning for the species, characterize the dimensions of the niche (Maire et al. 2012). If we assume that interspecific competition does truly occur among co-existing species, the ecological niche can be divided into two categories: the fundamental niche is defined in absence of competition while the realized niche is characterized by the presence of competition among species (Hutchinson 1957, Maire et al. 2012).
Based on the principle of competitive exclusion, niche theory implicitly assumes that, in order to have a stable co-existence, niches of co-occurring species must differ (Hutchinson 1957; Chesson 1991) although a certain degree of similarity is permissible (May & Mac Arthur 1972; Pianka 1974). By measuring competitors’ niche overlap it is possible to assess the effects of density-dependent competition on the tolerable upper limit of niche overlap (Young 2004). Moreover, the conceptual niche framework allows testing if competition between sympatric species is happening with the use of species assemblages modeling.
The ecological niche concept is strongly related to habitat selection (Maire et al. 2012). The latter is considered as a density dependent process. In fact, when populations are at low-density levels, individuals can freely occupy the habitat that maximizes their fitness. At the opposite, when population density levels increase, the individual fitness decreases within the most favourable habitat, making adjacent and less favourable habitats providing the same fitness. If habitat suitability can vary in function of population densities (Morris 1988), then habitat selection depends not only on resources abundance but also on the density of the same and/or different species sharing the same area. In the latter case, which is the most commonly represented, the organization of the community can be based on shared or distinct preferences (Morris 1988). In the case of shared preferences, competition for resources may occur. It is therefore possible to use habitat distribution patterns to evaluate the role of interspecific competition.

The necessary link between ecosystem knowledge and conservation

We have seen above how the loss of biodiversity is becoming an urgent issue that needs to be addressed by the conservation and scientific world before it becomes irreversible. Effective conservation measures are therefore required to guarantee the persistence of ecosystems that might provide invaluable services for human well-being (Braat & de Groot 2012), in most cases still undiscovered (Wilson 1992). However, successful conservation management cannot be implemented without well-developed knowledge on species biology and processes occurring among species sharing the same habitat and, eventually, the same preferences. The discipline of conservation biology has already contributed to mitigating anthropogenic actions on biodiversity at different organization levels and it has helped to reveal underlying mechanisms inducing variation in populations’ demographic parameters (Primack & Miller-Rushing 2012). Thanks to this information, policy makers were able to act, most of the time under high urgency, setting up appropriate conservation programs to save endangered species from extinction.
Most of the conservation measurements put into place to preserve particular species deal only with the management of physical or habitat features without accounting for species interactions, which, most of the time, are unknown before the management action has been undertaken (Soulé et al. 2005). This kind of procedure is a risk for the ecosystem as ecological cascades may bring a better or a worst result than the expected one. For example, the eradication of invasive rats on North Island (Seychelles) in 2005 led to an unexpected decrease in the number of invertebrates, both on ground and leaves, despite the fact that rats were known to be feeding on invertebrates. This unexpected result was probably due to the trophic cascading effects of the removal of rats, which triggered a significant increase in land birds and lizards, and also in large invertebrates, all of which are feeding on small invertebrates and limited by rats (Galman 2011; Rocamora & Henriette, in press). Therefore, when conservationists fail to understand the interactions occurring between species, conservation measurements might not achieve the desired results (Soulé et al. 2005). A solution to improve biodiversity maintenance can therefore be found in both practice and theory. Conservation evidence provide practical examples of what works in conservation (Sutherland et al. 2004a; Sutherland 2015) i.e. which actions have been already undertaken with successful results (see above). However, empirical studies on wild communities are strongly required to improve knowledge on mechanisms driving species coexistence (Morris 2003) and to track and predict changes on populations and communities when humans alter ecosystems properties.

Calculating the number of pairs per sampled plot

To obtain an unbiased number of pairs per species and per plot, we need to account for the content of invisible burrows, detection rate and breeding failure, i.e. treating i) the burrows that could not be visually inspected and for which no response was obtained after playback, and ii) the currently empty burrows, that may have contained a pair earlier in the season (e.g. premature failures) or not. We used two different methods: in the first method, census data were integrated within a Bayesian statistical framework (see CHAPTER 2). This method accounts for the imperfect detection based on response rate to playback in order to calculate the probability of occupancy. This probability is then used to estimate the number of burrows with unknown content (content not visible and no response to the playback) that are occupied by one or the other species. However, while accounting for imperfect detection, this method does not account for breeding failures. Therefore in our case, the resulting estimates correspond to a minimum population size for the 2014 census. To account for breeding failure in both species we used a second method that considers the data collected during the three breeding seasons. To estimate which proportion of empty burrows results from breeding failure, we took into account time (since total number of breeding failure will increase with season) and location in the island (as spatial variation in breeding density of the two species might occur). We assumed that the plots surveyed at the beginning of the breeding season (November) were at the maximum occupancy rate. Data were available in November (all years combined) for 73 plots and were used to map (using inverse distance weighting interpolation) the relative proportions of wedge-tailed, tropical shearwaters and empty burrows for the whole island, at a 1 ha resolution (see Appendix C). Then, we used the following procedure to calculate the number of pairs per plot accounting for imperfect detection, breeding failures and coexistence of the two species (i.e., an unknown burrow cannot be occupied simultaneously by the two species). The logic used in the calculation is described in details in Appendix C (which discusses also the assumptions used), a summary is provided below. The total number of burrows is obtained for tropical (TS) and wedge-tailed shearwater (WS) using the following formula (see Figure 3.2 for explanations on the various counts, N1 to N20):
TOT TS = N1+N4+ N14+N7+N16 (EQN 1)
TOT WS = N2+N5+ N13+N8+N15 (EQN 2)

Table of contents :

TABLE OF CONTENTS
ACKNOWLEDGEMENTS
RÉSUMÉ EN FRANÇAIS
ENGLISH SUMMARY
CHAPTER 1
1.1 GENERAL INTRODUCTION
1.2 GENERAL METHODS
CHAPTER 2 ANALYSIS OF PLAYBACK CENSUS TO ESTIMATE THE DENSITY OF CAVITY-DWELLING BIRDS
2.1 INTRODUCTION
2.2 MATERIALS AND METHODS
2.3 RESULTS
2.4 DISCUSSION
TABLES
FIGURES
EXTERNAL SUPPLEMENTARY MATERIALS (ESMS)
CHAPTER 3 ASSESSING POPULATION SIZE IN NOCTURNAL DWELLING SEABIRDS ACCOUNTING FOR DETECTION PROBABILITY AND HABITAT PREFERENCES: A SEQUENTIAL APPROACH 
3.1 INTRODUCTION
3.2 MATERIAL AND METHODS
3.3 RESULTS
3.4 DISCUSSION
TABLES
APPENDIXES
CHAPTER 4 COMPARATIVE FORAGING DISTRIBUTION AND ECOLOGY SUGGESTS INTERSPECIFIC COMPETITION BETWEEN TWO SYMPATRIC SHEARWATERS FROM THE SEYCHELLES
4.1 INTRODUCTION
4.2 MATERIALS AND METHODS
4.3 RESULTS
4.4 DISCUSSION
TABLES
FIGURES
EXTERNAL SUPPLEMENTARY MATERIALS (ESMS)
CHAPTER 5 MOVEMENT PATTERNS AND HABITAT SELECTION OF WEDGE-TAILED SHEARWATERS (PUFFINUS PACIFICUS) BREEDING AT ARIDE ISLAND, SEYCHELLES
5.1 INTRODUCTION
5.2 MATERIALS AND METHODS
5.3 RESULTS
5.4 DISCUSSION
TABLES
FIGURES
CHAPTER 6 GENERAL DISCUSSION
APPENDIX I
BREEDING SUCCESS ANALYSIS OF TROPICAL SHEARWATER AND WEDGE-TAILED
SHEARWATER POPULATIONS ON ARIDE ISLAND.
BIBLIOGRAPHY 

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