Project Kimberley

Global Environmental Change (GEC) includes a range of processes that will affect ecosystems over the coming decades; climate change, invasive species, pathogens, and habitat loss. Most if not all of these processes have a strong anthropogenic drivers and most will lead to loss of biological diversity.

The United National Educational, Scientific and Cultural Organisation state that “Global environmental change includes processes such as biodiversity loss, freshwater scarcity and climate change. Understanding of the complexity, interconnectedness and sensitivity of social systems, ecosystems and their interfaces is needed to address and plan for the environmental change” (ISSC/UNESCO 2013).

The Knowledge gaps relevant to biodiversity and ecosystem management affect all countries worldwide, but the sheer size of Australia's natural heritage, the significant logistical issues involved with studying it, and the relatively low numbers of people available to collect information mean that Australia is particularly affected by knowledge gaps. Especially but not only in remote areas, many fundamental questions remain concerning natural history, ecological function, and even basic taxonomy of many species.

These knowledge gaps restrict efficient the ability to implement efficient management strategies, leading to poor outcomes including loss of biodiversity, genetic diversity, and (via ecological cascades) ecosystem function. Worse, the processes feeds back on themselves; loss of biodiversity has been identified as a cause, not just a consequence, of GEC (Sutherland et al. 2012; Hooper et al. 2012), with compounding effects (Reich et al. 2012).

Of particular importance to loss of biodiversity is inaccurate identification of management units; incorrect identification of species, often resulting from a lack of relevant population genetic data, means that management strategies are focused at the wrong level and the outcomes are a major and permanent loss of biological and genetic diversity, both within and between populations.

The dangers of this lack of understanding and knowledge is now widely acknowledged by management agencies and a major goal in many situations is to collect the baseline data required to describe population structures and connectivity patterns (Potter et al. 2012).

For the Kimberley region of north-west Australia, addressing these knowledge gaps is an urgent task for biologists. The Kimberley is one of the last remaining wilderness areas in Australia and has long been recognised for its high levels of unique biodiversity; recent work suggests that those levels are actually far higher than previously thought and that large amounts of undiscovered diversity remain.

The cane toad Rhinella (Bufo) marinus is an invasive species that has spread across tropical Australia since its introduction near Townsville in 1935 (Lever, 2001; Urban et al., 2007). The most significant ecological impact of the cane toad on native Australian biodiversity is the lethal toxic ingestion by large native predators (reviewed in Shine, 2010).

In the Northern Territory, the arrival of toads has caused significant population-level declines of varanid lizards (71-96% declines in three spp.; Doody et al., 2009), and regional/local extinctions of northern quolls (Dasyurus hallcatus; Shine 2010; Woinarski et al. 2010, 2011).

A lack of ecological data predating the spread of toads through Queensland hampers qualitative analysis of their impact upon native predators (Shine 2010), but more recent monitoring of their spread through the NT indicates a range of effects. Varanid lizards (goannas) in riparian systems (Varanus panoptes, V. mitchelli, V. mertensi) show significant impacts following the arrival of toads, with declines of over 90% in local populations (Doody et al. 2009).

Individual freshwater crocodiles are killed by lethal toxic ingestion of toads, although whether these individual level impacts result in population level declines appears to depend on additional factors; Crocodylus johnstoni populations were affected in the Victoria River (Letnic et al 2008) but were not in the Daly River (Doody et al. 2009, Letnic et al. 2008), possibly due to differences in aridity or productivity between these two river systems (Doody et al. 2009; Somaweera et al 2012).

In the medium term, toads are likely to spread south through the Chamberlain and Durham systems, and once into the headwaters of these systems they will be close to the headwaters of the Fitzroy River ; once into the Fitzroy their spread westwards will be downstream and is likely to be rapid.

In contrast, their spread through the northern Kimberley may take some time; in order to reach the Drysdale and Mitchell catchments toads need to move up the Durack River from the Cambridge Gulf, against the eastwards flow of the river.

Toads spread rapidly across the slopes and plains of the Top End of the NT, but the landscape of the Kimberley is very different; rivers are enclosed by steep cliffs and gorges for large sections, with dry ridge country separating river valleys.

Toad dispersal in the Kimberley may be confined to the major waterways to a larger extent than in the Top End; this may present novel management issues and opportunities.

Knowledge of population structure can have a profound effect on understanding of a species' ecology and conservation biology. For example, the magnificent tree frog (Litoria splendida) is a large native frog that is endemic to the Kimberley. It inhabit the wet gorges that are common in the sandstone landscape.

The level of movement and dispersal between gorges is unknown; if dispersal is low, then genetic exchange between adjacent gorges may be low and the population will be highly structured, and management strategies that treat the species as a single homogeneous population will result in loss of genetic diversity and evolutionarily important structure. If, on the other hand, gene exchange between populations is high then different management approaches are suitable.

Long term effects of rapid, severe reduction in population numbers are evident as a loss of genetic diversity within populations, a process termed genetic bottlenecks (England et al, 2003); bottle-necks result in a high loss of genetic diversity measured as loss of allelic diversity and heterozygosity.

Once gone, this genetic diversity is impossible to recover within the timeframe of most current conservation management programmes.

This loss of genetic diversity has important long term effects, including declines in individual health resulting from the loss of heterozygosity associated with inbreeding, and a loss of adaptive or evolutionary potential (a consequence of low allelic diversity following genetic drift) required for adaptation to environmental changes such as climate change, invasive species, new pathogens, or just small shifts in community and ecosystem dynamics (e.g. Tassie Devils; also Frankham, 2005; Reed &Frankham, 2003; Reid et al, 2003; Webb et al, 2011).

This is a serious issue (perhaps the most serious issue) for many conservation programs since the resilience and persistence of populations and species is dependent on the capacity to adapt to dynamic and changing environments.

There is now strong evidence of extinctions occurring because of inbreeding and loss of genetic diversity (Frankham, 2005). In many cases, the extent and nature of the original diversity is simply unknown, so that management lacks the means to assess the loss of adaptive functional genes and the capacity to set a relevant target outcome for restoration of genetic diversity (and no capacity to restore it anyway, since alleles have been permanently lost).

Gene-banking, where genetic diversity is stored in advance of a bottleneck taking effect, is a strategy that can provide long-term insurance against loss of genetic diversity, and has been applied to a range of plant and animal species.

Genebanking programs can be implemented for a fraction of the cost of other ex-situ management approaches (Wildt et al, 1997; Pukazhenthi et al, 2006), and are preferable to maintaining captive populations that lose genetic fitness due to adaptation to captivity (Frankham, 2008).

Working ecosystems are essential for sustainable human societies, but ecosystem function is threatened by loss of biodiversity, and biodiversity is itself under serious threat from the processes that are collectively termed 'Global Environmental Change'; overexploitation and habitat loss, the effects of invasive species and disease, and climate change.

The relationship between ecosystem function and biodiversity is complex; many ecological communities can withstand reductions in species diversity and abundance and yet still function (for example, in terms of trophic interactions) much as they did before.

This trait - the resilience of an ecosystem (Orians 1975) - is critical to environmental managers seeking to maximise ecosystem function. When the community's biodiversity is reduced beyond a threshold level, trophic interactions are significantly altered and the resilience of the community is exceeded, with the result that is severely reduced or even lost. For managers, changes that exceed an ecosystem's capacity for resilience pose the greatest threat and thus need to be managed with the highest priority.

Ecologists have documented many instances where resilience has been exceeded and ecosystem function lost, with the most serious examples - in terms of the scale of the collapse, and the impact upon human society - coming from fisheries.

The reduction in large predatory fishes worldwide (Myers & Worm 2003) has led to both targeting of lower trophic levels (Pauly et al. 1998) and the decline of other fisheries - for example, the reduction in bivalve catches following the removal of large predatory sharks from Western Atlantic coastal waters (Myers et al 2007).

This effect of predator removal, where unintended consequences are observed throughout the food web, is termed a ecological cascade; where such as cascade involves a keystone species the consequences on the ecosystem c an be pronounced.