Mangroves are woody plants that occur at the land and sea interface and inhabit the upper intertidal zones of saltwater areas, primarily in tropical and subtropical regions (Tomlinson 1986; Hutchings and Saenger 1987). Since mangrove forests are an ecotone between terrestrial and marine realms, they have both vegetative strata that are similar to many inland terrestrial ecosystems as well as an intertidal benthic component that is representative of the marine environment. As such, mangroves exhibit a very steep gradient from fresh to saline water (Mitsch and Gosselink 2000). The combination of terrestrial and marine environments occurring within mangrove forests can present many difficulties, in terms of abiotic stressors, such as salt water instead of fresh water and twice daily tidal inundations, for terrestrial species attempting to colonise this habitat (Mitsch and Gosselink 2000; Luther and Greenberg 2009). Salt water could present physiological challenges for some species and high tides could present challenges for ground-nesting and terrestrial species that prefer dry land to aquatic systems. Finally, competition or predation from estuarine and marine species has the potential to inhibit the colonisation of mangrove ecosystems by some terrestrial species. The aforementioned abiotic and biotic stressors could prevent some terrestrial species from invading mangrove habitats, which makes mangroves an excellent system in which to investigate the biological invasion of species into novel habitats.

Based on floral species composition, mangroves occur in two distinct biogeographical regions, the Indo-West Pacific and the Atlantic–Caribbean–East Pacific. These regions can be further subdivided into six distinct subregions: Eastern Pacific, West Atlantic–Caribbean, West African, East African, Indo-Malaysian and Australian (Duke 1992) (see Figure 6.1). The subregions of Australia and Indio-Malaysia have more than twice as many mangrove tree species as any other subregion (Duke 1992). Mangrove forests are one of the most productive ecosystems in the world with ample food resources for both marine and terrestrial animals (Mitsch and Gosselink 2000). However, mangroves are also cited as extremely depauparate of biodiversity when compared to other habitats (Tomlinson 1986; Hutchings and Saenger 1987; Hogarth 2007).

Foreword

Alas, the poor ecologist who is expected to follow the laws of scientific inference that have arisen from physics and chemistry. Erect hypotheses, make predictions, see if they are supported by evidence obtained by observations or manipulative experiments. Perhaps it would be easier if, instead of 30 million species, we had only 118 elements in the periodic table to study or only a few forces in physics to design hypotheses around. So how do we cope? We can deal with autecology or the ecology of individual organisms because we have a strong base in physiology and simple things like metabolic rates are constrained by how evolution has proceeded. We can deal with populations because they typically have a restricted nexus of interactions, as Andrewartha and Birch (1984) told us. But things are getting more complicated since the interactions can involve competition, predation, disease, food supplies, climate, and social effects. Perhaps we can cope with this amount of complexity, but it is certainly complex enough to allow many ecologists to argue extensively about the factors causing populations to rise or fall. In principle we can sort out these arguments at the population level by field or laboratory experiments, and this approach will often work to provide evidence-based explanations. But when we move up to community and ecosystem ecology problems multiply if only because experimental manipulations become more difficult and certainly more expensive. It is partly a reflection of why aquatic community ecology has progressed more than terrestrial community ecology – large-scale experiments in rivers and lakes are more prevalent than they are in terrestrial ecosystems. But it may also be partly a reflection of hypotheses that are not operational.

In an ideal universe we might be able to work out some of these problems but the arrival of human influences has added yet more complexity. Invasion biology is now one of the leading fields of community ecology both because of its intrinsic interest as a test case of how much we understand community interactions and even more because many species invasions have consequences written very large in dollars and cents.

Introduction

The honeyeaters (Meliphagidae) are one of the most speciose and distinctive elements of the Australo-Papuan bird fauna. Following a number of recent taxonomic revisions based on DNA analyses the family currently comprises 50 genera and 182 species (Cracraft and Feinstein 2000; Driskell and Christidis 2004; Ewen et al. 2006; Driskell et al. 2007; Norman et al. 2007; Fleischer et al. 2008; Nyari and Joseph 2011). The family reaches its highest diversity in Australia and New Guinea (125 species), but its distributional limits extend from Bali in the west, northward to Micronesia and eastward through New Zealand to islands of the southwest Pacific (Figure 5.1) (Mathew 2007; Higgins et al. 2008). The honeyeaters have a long evolutionary history in the region being a basal lineage of the oscine passerine radiation that arose in Gondwana prior to 34 Ma (Ericson et al. 2002). Unlike some elements of the Gondwanan biota, honeyeaters are absent from Africa and South America and, with the exception of a single species, do not extend west of Wallace’s Line.

The honeyeaters are an ecologically and morphologically diverse group. They range in size from 9–50 cm; the smallest species being the Myzomela honeyeaters and the largest the yellow wattlebird Anthochaera paradoxa (Daudin) (Higgins et al. 2008). Honeyeaters occur in nearly all habitats of the region and are often the most abundant species. They include both habitat specialists and generalists. The most distinctive morphological feature of the honeyeaters is the presence of a protrusible tongue with a brush-tip, an adaptation for nectar extraction (Paton and Collins 1989). Nectar is a major component of the diet of nearly all honeyeaters although some species are primarily insectivorous (e.g. green-backed honeyeater Glycichaera fallax (Salvadori) or frugivorous (e.g. painted honeyeater, Grantiella picta (Gould), a mistletoe specialist) (Higgins et al. 2008). Most commonly honeyeaters have a diet that consists of nectar supplemented with insects. Honeyeaters share a number of morphological, physiological and behavioural similarities with nectarivorous birds from other regions of the world (sunbirds and hummingbirds) (e.g. Pyke 1980) but they are unrelated (Sibley and Ahlquist 1990); the similarities are a consequence of convergent evolution.

Introduction

This book is primarily concerned with how species colonise and invade new areas. It examines the evidence relevant to a suite of hypotheses concerning processes that occur on present-day populations in present-day landscapes. This chapter takes an approach different to that of many other chapters. Before examining the main hypotheses, it begins by exploring how the deeper evolutionary history of any species will inform design of research into that species’ ecology, behaviour and present distribution. The ecological and behavioural hypotheses one might test to describe the evolution of species in their current geographical and ecological ranges will depend on tests of relevant phylogenetic hypotheses about a species – its systematics. For example, only from a phylogenetic analysis can it be understood whether a species is part of relatively species-diverse or species-poor clades. So, too, in this way can it be best understood whether it shares an ecological trait with other species because it is closely related to them, or because the trait has arisen multiple times, or has been retained from a relatively distance ancestor. In the context of this book, this might usefully be seen as an ecological and behavioural view of historical biogeography, or how clarifying the past is the first step to fully understanding the present.

Introduction

The Eurasian skylark (Alauda arvensis Linnaeus) is a small passerine that breeds across most of Eurasia from Western Europe and northwest Africa to eastern Russia, northern China, Japan and eastern Siberia (Figure 16.1). Its taxonomy is somewhat murky, A. arvensis is sometimes considered forming a superspecies with the Oriental skylark (A. gulgula Franklin) and some of its 13 subspecies are treated as species by some authors (del Hoyo et al. 2004). The subspecies are sometimes divided into ‘European Group’ and ‘Asian Group’, which I refer to as western and eastern skylarks in the remainder of this chapter. The western skylark (especially the subspecies A. a. arvensis) was the subject of most published studies as well as providing stock for introductions. Skylarks are mainly resident in the west of their range, but eastern populations are more migratory, moving further south in winter (Figure 16.1).

In the past, especially through the nineteenth century, the skylark’s range had presumably expanded, as large-scale habitat change, including increased deforestation and expansion of crops and pastures made it possible for the species to spread from diminishing natural steppe grasslands (Cramp 1988). Today it is one of the most common farmland birds in countries with extensive farmland, such as Germany (Toepfer and Stubbe 2001), France (Eraud and Boutin 2002), the Netherlands (Kragten et al. 2008), Sweden (Wretenberg et al. 2006), Finland (Suhonen et al. 1994), Poland (Sanderson et al. 2009) and the United Kingdom, where about 70% of the total skylark population occurs on farmland and 50% on arable land (Donald and Vickery 2000).

Introduction

We live on a continually shifting planet as the continents on which we have built our lives drift across the Earth’s surface, a process which has changed the configuration of the continents throughout Earth’s history and will continue to do so. A world map 250 Ma from now will reveal a very different arrangement of continents in the northern hemisphere than that observed today. Earth science studies have revealed the processes associated with continental movement known as plate tectonics; we have learnt that the Earth is made up of lithospheric plates carrying the continents and oceans, which driven by mantle processes are in continuous motion (Skinner and Porter 1987).

A map of the Earth’s plates (Figure 11.1) reveals the relative movement of each plate and their interactions along their boundaries such as subduction around the Pacific Ring of Fire, spreading along the Mid-Atlantic Ridge and transform faulting at the San Andreas Fault (Skinner and Porter 1987; Kearey and Vine 1996). Our world has been, and will continue to be, shaped by the movement of these plates (Skinner and Porter 1987).

Introduction

Prior to European settlement the Australian continental landmass (i.e. Australia plus the main island of New Guinea; Figure 10.1) was successfully invaded by only four groups of placental mammals: bats (Chapter 8), rodents, primates (our own species) and the dog (which almost certainly came with people; Chapter 19). Not surprisingly, bats were the earliest and in many respects, the most successful invaders (Hand 2006). Bats colonised Australia at least nine times, commencing sometime prior to the Early Eocene (Hand et al. 1994) when Australia was still connected to other Gondwanan landmasses. At the other end of the geological timescale, humans entered the region only during the late Pleistocene (c. 50 thousand years ago; kya) despite a much longer occupancy of islands to the immediate north (Morwood et al. 1999). Later still, the dog was transported to Australia around 4000 years ago (see Chapter 19). Rodents represent the middle ground in the history of placental invasion of Australia. They first entered the Australasian region during the late Miocene, after northward drift had brought the Australian continental plate into collisional contact with the Asian Plate (Lee et al. 1981). Despite the proximity of landmasses, the journey from Asia to Australasia still involved multiple water crossings, even during periods of low sea levels. Ultimately, only one of the many different kinds of rodents found in Asia proved fit for the challenge – the true rats and mice of the family Muridae – but members of this group succeeded on multiple occasions through natural dispersal and, more recently, with human assistance.

At the time of European settlement Australia supported around 66 species of native rats and mice, more species than in any family of marsupials or bats. Murine diversity is even more pronounced on the island of New Guinea and its major satellites to the north, with 114 species of native rodents already known and more being discovered (Helgen 2005a, b; Musser et al. 2008; Helgen and Helgen 2009; Musser and Lunde 2010). Native rodents thus comprised around 29% of the native terrestrial mammal fauna of Australia, and around 59% of that of the Melanesian islands. These figures do not rest comfortably with the common notion of Australasia as a continent of marsupials – but they do point to a fascinating history of invasion by what is clearly a highly successful evolutionary lineage.

Some of today’s most pressing issues deal with invasions by alien species into natural or man-made ecosystems such as agricultural landscapes. Invasions are not a new phenomenon having been a part of the relationship between man and the environment ever since humans moved out into the savannas; however, they became part of the ecological agenda in the middle of the last century. The foundations of invasion ecology stem from Charles Elton, who, in his book, The Ecology of Invasions by Animals and Plants (published in 1958) attempted to draw together three stands of ecology – faunal history, ecology, particularly population ecology, and conservation. Elton’s book had some traction at the time (e.g. Baker and Stebbins 1965), however, few ecologists paid much attention to invasions during the 1960s even though island biogeography theory (MacArthur and Wilson 1967) did provide theoretical frameworks for how new species fitted into the resident species communities on islands. It was not until the 1970s that invasion ecology began to gain traction in the literature (e.g. Baker 1974; Embree 1979) and continues to this day (Richardson 2011). There have been recent attempts to create unified theoretical frameworks for understanding the invasion process (Blackburn et al. 2011) and the traits that determine the degree to which a species can invade a new ecosystem or the degree to which an ecosystem can be invaded by a new species (Richardson and Pysek 2006). These developments provide a foundation upon which to assess the degree to which hypotheses concerning biological invasions relate to real-world case studies that are proliferating in the literature.