Connectivity among species, populations and habitats is important for their long-term persistence, contributing to gene flow and supporting adaptation to environmental change. Connectivity has been altered in most habitats to some degree, because of habitat loss or fragmentation, anthropogenic change to mechanisms of dispersal, or impacts on species’ biology and lifecycles. There is high consensus that changes in connectivity are occurring, but direct evidence (e.g. from genetic studies) is limited because of the inherent difficulties in studying connectivity in nature, and the relatively recent availability of technologies to measure it. Understanding connectivity is important for conservation planning, particularly in the design and placement of marine reserves (Palumbi 2003).
On land and in nearshore waters, coastal urbanisation presents a barrier to species that are restricted to the coastal zone. For example, in New South Wales, there are known distributional gaps for species such as Phyllospora comosa (Coleman et al. 2008). This is particularly problematic for species that migrate or are reproductive seasonally or as part of their lifecycle, or for those attempting range extensions in response to changing climate. For example, disruption of connectivity through habitat loss on international flyways is considered the major threat to most shorebirds (see Nursing, roosting and nesting sites).
Of the jurisdictions, New South Wales, Western Australia and Queensland are the most likely to have altered marine connectivity because of their strong and rapidly changing boundary current systems (Cetina-Heredia et al. 2015). The ranges of tropical species (Vergés et al. 2014) are expanding, and temperate species are retreating polewards (Wernberg et al. 2011). A strengthening East Australian Current is facilitating expansion of the long-spined sea urchin (Centrostephanus rodgersii) into Tasmania. The importance of increased connectivity relative to other pressures for larval outbreaks of crown-of-thorns starfish (Acanthaster planci) along the Queensland coast is still uncertain (see the Marine environment report for further discussion of crown-of-thorns starfish). In Western Australia, connectivity of lobster populations is altered by climate change–driven heatwaves, and the loss of macroalgae from the north may have decreased connectivity. In contrast, connectivity may be most stable in the Northern Territory, because of the largely unmodified coastline and lack of a rapidly changing boundary current.
Connectivity has changed, and will continue to change, in differing directions, particularly on larger temporal scales, but quantifying this requires long-term measurements using genetic tools at the least. On short temporal scales, changes to connectivity will likely be related more to short-term ‘pulse’ events that alter species distributions (e.g. pollution events, development) or episodic climatic events such as the 2012 marine heatwave in Western Australia (Wernberg et al. 2013).
Conservation strategies such as spatial closures and fisheries restrictions should aim to maintain connectivity of native species, and reduce connectivity for non-native species. How to integrate connectivity into conservation planning across species and habitats is uncertain, as is anticipating changes to vectors of dispersal. More temporal studies are needed, and generalities need to be explored (but see Durrant et al. 2013). One tool to assist is the Australian Connectivity Interface (currently known as ), which estimates the probability that any 2 regions of upper water column are connected through passive dispersal (Condie et al. 2005).
Nursing, roosting and nesting sites are generally protective and productive habitats, such as mangroves, beaches, islands and wetlands. Several groups of animals, including shorebirds, seabirds and turtles, use these habitats to breed, and for juvenile development and rest.
The pressures on these sites vary between locations and species. Pressures on important beach sites include nourishment, armouring, grooming and seaweed harvesting (Schlacher et al. 2014). Seaweed harvesting is a growing industry in south-eastern South Australia, where a key ecosystem component is removed from coastal food webs. Other more general pressures include urbanisation, port developments, recreational disturbance, water abstraction and sea level rise. Rising sea levels threaten to move turtle nesting inland in the short term, and to remove suitable habitat in the longer term. Many turtle nesting sites are threatened by human settlement and associated light pollution (Kamrowski et al. 2012). Non-native predators such as dogs, foxes and rats also increase mortality at nursing, roosting and nesting sites. Predation by corvids, dogs and foxes threatens penguins and other ground-dwelling birds (Ekanayake et al. 2015), and foxes and pigs dig up the turtle nests and eat the eggs. The Queensland and Australian governments are contributing $7 million across 4 years to combat the threat of feral predators to turtle nests.
The outlook for coastal nursing, roosting and nesting sites is poor. Increasing sea levels, combined with coastal armouring, will modify habitats to the point that they will be unsuitable for many species, and nursing, roosting and nesting sites need to be explicitly considered when developing coastal management plans to accommodate these changes.
Trophic structures and relationships
The nature of trophic (feeding and predation) interactions among ecosystem members has implications for food-web diversity, stability and function (Rooney & McCann 2012). Human activities in the coastal environment disrupt trophic structure and relationships by removing larger species and predators, predominantly through overharvesting. Trophic structures may also be affected through bottom-up processes such as increased nutrient addition to local waters.
For estuaries, there are many uncertainties regarding the condition of trophic dynamics at the national scale, particularly for turbid estuaries in northern Australia. Good but patchy information on trophic relationships comes from stomach content analysis and stable isotope studies, and data on trophic structure come from fisheries data or diver-based surveys of outer estuaries and embayments (Edgar & Stuart-Smith 2014). In general, surveys have suggested that food webs are lacking in higher trophic level representatives that should be present and, in some cases, are extremely lacking in expected species and therefore particular functional groups.
Evidence for human impacts on trophic dynamics can be seen when management actions remove or reduce the activities that disrupt relationships. For example, data from green zones in the Great Barrier Reef show major changes in trophic structure in response to closure from fishing, demonstrating the effects that had previously been imposed by fishing. Other evidence has been found in Tasmania, where, outside the MPAs, removing lobsters decreases kelp resilience to invasive long-spined sea urchins (Centrostephanus rodgersii), increasing the potential for the creation of urchin barrens (Ling et al. 2009). Higher-level predators are also expected to increase in locations such as Jervis Bay, Port Stephens, Port Davey and Moreton Bay, where MPAs offer opportunities to restore more natural trophic structures.
The outlook for trophic structures in the future is uncertain. Trophic relationships will always exist, loss of species and trophic levels will reduce the range of relationships, and there will be an overabundance of biota that have been freed from their predators. In the short term, trends are likely to remain stable, with the potential to improve in the long term where human activities are well managed.
Filter feeding (where animals such as sponges, bivalves and barnacles feed by straining suspended food particles from water) is a crucial ecological process that maintains water quality and biogeochemical cycling in coastal waters. Filter feeders can be particularly important in nutrient-poor environments. For example, because of their filter feeding and ability to recycle dissolved organic matter, sponges were recently found to facilitate coral reefs in nutrient-poor waters (de Goeij et al. 2013).
The state of coastal filter-feeding communities varies tremendously based on exposure to both local anthropogenic pressures (e.g. nutrients and point-source pollutants) and natural pressures (e.g. rainfall causing reduced salinity and increased run-off). Consequently, their state varies on small spatial scales (kilometres) and across relatively short time periods. For example, a filter-feeding community near a metropolitan river is likely to be in a different condition from a filter-feeding community in a remote area, although both will respond to recent changes in rainfall. Clark et al. (2015) found increased abundances of active filter feeders such as barnacles and sea squirts in highly developed estuaries, and this is a potential response to increased pelagic food availability in nutrient-enriched areas (Lawes et al. 2016).
There is no general source of information for filter feeding nationally. Online resources such as present a range of environmental variables and predictions relevant to Queensland coastal filter-feeding communities (e.g. catchment flow, water quality forecasting, sediment modelling); however, there is limited inclusion of fauna other than reef-forming corals.
To appropriately manage coastal filter-feeding processes, a focus is needed on measuring and modelling water quality parameters (e.g. nutrient and pollutant run-off, turbidity) and other pressures (e.g. temperature, salinity) (Fisher et al. 2015), and then linking these to the abundance of filter feeders and rates of feeding. A recent study compiling data from 3 dredging programs in tropical Australia summarised water quality parameters so that thresholds may be developed for filter-feeding organisms (Fisher et al. 2015). This form of research will facilitate best environmental practice and monitoring of filter-feeding communities in response to potential pressures.
The outlook for filter feeding as an ecological process is dependent on the spatial scale of interest. The likelihood of high rainfall because of extreme storm events associated with ENSO and climate change will increase the likelihood of extreme freshwater stress to estuarine populations of filter feeders. Other ecological communities (particularly those in intertidal and mudflat areas) may also be exposed to thermal stress because of warming waters and hotter days.
Microbial processes and nutrient cycling
Natural populations of endemic marine microorganisms form the base of the marine food web and provide a suite of valuable ecosystem services (Falkowski et al. 2008). Marine microbes transform elements in a series of processes that control biogeochemical cycling in coastal waters (Worden et al. 2015). These processes and elemental cycles in turn affect the availability of nutrients such as carbon and nitrogen to the food web, and the ocean’s influence on global climate (Falkowski et al. 1998, Worden et al. 2015). Shifts in the productivity and composition of microbial populations can strongly influence food-web structure and even drive yields of commercial fisheries.
Microbial processes in the coastal zone are strongly influenced by the input of allochthonous nutrients (nutrients coming from outside the ecosystem), other contaminants (Doney 2010, Nogales et al. 2011) and organic matter. These inputs can lead to dramatic shifts in microbial community composition (Sun et al. 2012, 2013) that result in changes in function (e.g. elevated respiration leading to anoxia) or bloom events, which can be toxic. Eutrophication can also disrupt the balance of microbial interactions with marine plants and animals, leading to disease outbreaks. In the short term, increased human population densities near the coast have the potential to increase the input of nutrients and contaminants (Kennish 2002) and exacerbate these impacts, although stricter environmental regulations and better awareness may buffer these effects. In the longer term, microbial diversity and functionality may be influenced by climate change processes (Doney et al. 2012). Microbial activity is strongly influenced by temperature, and there is evidence that rising seawater temperatures may cause significant shifts in the composition of marine microbial assemblages (Fan et al. 2013, Kroeker et al. 2013). This can ultimately lead to changes in the biogeochemical function of microbial assemblages, or increases in the occurrence and virulence of diseases (Campbell et al. 2011, Case et al. 2011), including potentially dangerous pathogens such as Vibrio spp. (Kelly 1982).
Microbial pathogens cause disease in marine animals and plants, including species of economic interest. Other marine microbes form transient noxious blooms, which can disrupt the function of entire coastal ecosystems. Microbe-related blooms are being observed more frequently, across larger areas and with more severity (Hallegraeff 1993). Introduced microbes, including enteric microbes (bacteria and viruses), enter marine environments through various pollution point sources (e.g. sewage overflows, storm water) and can threaten the health of human populations that use coastal environments for recreation and food supply (Nogales et al. 2011). Whether increased reporting of microbial-mediated disease reflects more incidences or greater awareness is unclear.
Because marine microbial populations are extremely sensitive to environmental change, the balance of these positive and negative effects can shift rapidly. This can have implications for ecosystem function (e.g. primary production) and, in some cases, can produce large-scale ecological impacts (e.g. noxious bloom events) (Hambright et al. 2014). Understanding the factors that control and alter the structure and function of marine microbial assemblages is critical from an ecosystem management perspective (Bodelier 2011), because even subtle shifts in the composition of microbial assemblages can affect ecosystem and human health.
Lack of data on microbial processes within marine systems makes it difficult to judge the state of microbial processes within the Australian coastal marine environment. Additionally, the complexity and spatiotemporal heterogeneity of microbial processes (Stocker 2012, Ladau et al. 2013) complicate our ability to understand microbial dynamics within Australian waters, or predict how microbial processes will change under different environmental conditions. Little is known about patterns of diversity (Ladau et al. 2013), specific responses of taxonomic units to stress and their contribution to ecosystem function (Evans-Illidge et al. 2013). We lack baseline information such as the identity of natural communities of microbes living within specific habitats at a given time, and how the composition and function of natural marine microbial assemblages change temporally or spatially (Ottesen et al. 2012; see Box COA9). At a higher level, we also lack anything beyond a simple understanding of how microbial community structure and function respond to shifting environmental conditions and hydrodynamic processes.
To remedy this, we need sustained measurement of important microbial processes (e.g. nitrogen and carbon dioxide fixation, carbon turnover, host–microbe interactions), coupled with information on microbial diversity (Worden et al. 2015). This would enable us to link function to specific microorganisms, and hence develop a full model of the microbial process in the coastal environment. This information, combined with experimentation and ecosystem models (Coelho et al. 2013), would help to build a more complete picture of coastal ecosystem functioning in the face of environmental pressures.
Knowledge of coastal microbial processes is improving on the back of technological advances and greater research interest (Stocker 2012). Modern approaches have enabled researchers to gather valuable information in the past 5 years, with a limited body of recent work demonstrating human impacts on microbial communities. Currently, our knowledge of microbial processes within Australian marine ecosystems is derived from 2 main sources:
- Research and monitoring conducted by universities, CSIRO and the Australian Institute for Marine Science. Australia has a strong research community in marine microbial ecology, and research is conducted in a range of habitats, spanning the coasts to the open ocean, and from the water column to sediments, and in animal and plant hosts (Marzinelli et al. 2015). A recent development that will advance this research is the Bioplatforms Australia Marine Microbiology Project, which involves members of several universities, as well as CSIRO, the Australian Institute for Marine Science and the Integrated Marine Observing System (IMOS). The project will sequence microbial assemblages from 7 oceanographic IMOS National Reference Stations and 3 coastal monitoring sites around Australia, and will describe the monthly, seasonal and interannual dynamics of microbial taxonomic and metabolic capacity in waters ranging from the tropics to the Southern Ocean.
- Monitoring of coastal environments for microbial contaminants from pollution (sewage and storm water). Monitoring by various state and Australian government agencies provides data on the occurrence of enteric indicator microbes (e.g. enterococci counts), to gain an insight into the occurrence of potentially dangerous human pathogens within coastal habitats. One example of this basic data source in New South Wales is provided by Beachwatch, which uses Escherichia coli counts to consider the water quality of beaches along the New South Wales coastline. Given the advanced technology available to detect a huge range of more sensitive indicators of pathogenic potential, basic cell counts of cultivated microbes such as this are likely to be rapidly replaced.