Ecological processes in the marine environment are highly spatially and temporally variable. They are often specific to habitats, communities and biogeographic regions, and incorporate highly complex interactions between biophysical parameters and species groups. Measuring many processes can be challenging because of their complexity and highly dynamic nature (Griffith & Fulton 2014, Kool & Nichol 2015).
State and trends of indicators of marine ecosystem health: Ecologial processes
State and trends of indicators of marine ecosystem health: Ecologial processes
Connectivity and trophic structure
Defining connectivity of marine populations is highly complex, and can vary depending on the systems and processes being focused on. Definitions of connectivity include source-sink relationships7 between populations; key pathways for movement (migration or general movement) between habitats; mixing of multiple populations or stocks of a particular species within a defined region; or fidelity and lack of mixing of breeding populations, resulting in genetic structuring of populations (Cowen et al. 2000, 2006; Figure MAR33). Ultimately, the concept of connectivity can be regarded as relating to the rates, scale and spatial structure of exchange between populations (Cowen et al. 2006).
A large proportion of connectivity within the marine environment is driven by the dispersal of larval stages of species by physical processes such as winds, waves, tides and currents, modified by relatively small larval movements and behaviours (Kool & Nichol 2015). High connectivity can have positive or negative influences on ecosystem components, depending on the circumstances, as can low connectivity (i.e. high retention). For example, transport of eggs and larvae from spawning grounds to nursery areas may be critical to successful breeding, but may also contribute to the spread of harmful species and diseases. An increase or decrease in connectivity may not be directly indicative of a better or worse state, and the consequences are dependent on the system and processes within the relevant ecosystem.
A composite image that includes 3 maps of Australia and surrounding waters, a series of coloured boxes and arrows, and a line graph. Together, the figure demonstrates 4 examples of connectivity in marine habitats. The top 2 maps show connectivity of habitats for marine larvae via source–sink pathways, using a coloured scale to show the distribution of particle densities moving westward from the west and north of Australia further into the Indian Ocean. The third map shows the connectivity of habitats via migration pathways of southern bluefin tuna, indicated by a purple oval (spawning grounds) to the north-west of Australia; orange patches (juvenile summer feeding grounds) in the Great Australian Bight and off the coast of South Africa; and purple arrows (winter migration routes) between Australia, South Africa and New Zealand. The coloured boxes and arrows represent the mixing of populations within and between particular habitats for a marine fish, indicated by 5 purple ovals (spawning areas) pointing to 2 yellow rectangles (larval retention areas), pointing to 1 green rectangle (juvenile nursery area), pointing to 3 blue rectangles (adult feeding areas), pointing again to 5 purple ovals (spawning areas). The bottom bar graph shows the genetic structure of yellowfin tuna populations between 2 sites in the Pacific Ocean (the Coral Sea and Tokelau), represented by membership probability; the values are much higher for the Tokelau population.
If we assume that changes in physical processes such as boundary currents provide an indication of the state and trends of connectivity, observations of a weakening Leeuwin Current (Feng et al. 2012) suggest that connectivity of populations in waters off Western Australia has decreased. Conversely, with strengthening of the polewards extension of EAC eddies (Cetina-Heredia et al. 2014), connectivity of populations in waters off eastern Australia could be assumed to have increased. Across habitats such as coral reefs, connectivity based on physical processes occurs on much finer scales (e.g. Hock et al. 2014). Investigation of such connectivity requires fine-scale modelling of ocean processes across large-scale regions. Changes in physical processes associated with climate change are assumed to be affecting connectivity, with varying effects on marine populations, including both limiting and expanding habitat (Hartog et al. 2011, Johnson et al. 2011), and altering larval dispersion (Cetina-Heredia et al. 2015). Progress on modelling approaches through initiatives such as eReefs is allowing theconnectivity to be modelled (Cetina-Heredia et al. 2015, Schiller et al. 2015), and impacts of changes to physical processes on connectivity within and across ecosystems to be determined.
Connectivity and trophic processes are linked on evolutionary scales. Colonisation history influences food-web structure and food-chain length. As new colonisers move into habitats, they alter the food web, modifying trophic structures within the habitat (Post 2002), and can alter habitat structure and associated biological communities (e.g. the southerly extension of the black-spined sea urchin from the mainland to Tasmania; Ling et al. 2009). Selective mortality on specific ecosystem components-for example, the removal of large predators by fishing (see Box MAR5)-will also affect trophic structure. The removal of larger predatory lobsters by commercial and recreational fishers off eastern Tasmania is implicated in the successful climate-mediated colonisation by the black-spined sea urchin, which has, in turn, reduced theresilience of kelp forests, contributing to the listing of giant kelp forests in south-eastern Australia as Australia's first threatened marine community (F).
Much of our understanding of trophic processes is based on dietary studies, whereas shifts in trophic structure are typically inferred from observations of changing species distributions (e.g. Hobday et al. 2008, Last et al. 2011a), or from modelling studies that also incorporate changes in species abundance (e.g. Fulton et al. 2005, Bulman et al. 2012, Dichmont et al. 2013). Modelling studies (Fulton et al. 2005, Klaer 2005) suggest that some food webs in south-eastern Australia have been restructured during the past century, particularly as a result of the intensification of commercial fisheries up to the 1990s. Eastern Australian ecosystems, including the Great Barrier Reef, are also known to be highly modified (Butler & Jernakoff 1999, GBRMPA 2014a). Reduced fishing pressure, particularly in the past 5-10 years (Flood et al. 2014, Patterson et al. 2015a), should support the recovery of trophic structures across these ecosystems. However, complete recovery is unlikely given ongoing pressures (e.g. remaining recreational and commercial fishing, habitat modification, pollution), because some highly depleted species (e.g. eastern gemfish) may not recover fully from past overexploitation, and climate-related changes in connectivity are changing the underlying structure of the physical environment. Trends of the state of trophic processes are currently unclear.
The ecosystems in the north, west, south-west and south of Australia see less direct, and spatially more variable, pressures than those in the east and south-east, and as a result are likely to have trophic structures that are not as highly modified. However, there are little or no data for deeper-water habitats and more remote locations, and changes to trophic structure are unlikely to be recognised if these were to occur. Areas closed to local pressures, including marine protected areas and reserves (IUCN categories I and II), and fishery closures, provide an opportunity to recognise what (locally) undisturbed trophic structures could look like.
Diseases, outbreaks and blooms
Connectivity influences the spread of diseases and parasitic infestations. It can transport and sustain harmful algal blooms that can suffocate organisms (particularly in semi-enclosed areas; see also the Coasts report) by reducing dissolved oxygen concentrations, or poison organisms directly.
Australia has a reporting system for aquatic animal diseases of national significance. All the diseases currently reportable to the World Organisation for Animal Health and any other aquatic animal diseases of national significance are included on Australia's National List of Reportable Diseases of Aquatic Animals.
Based on reporting of events, there are currently no regionally or nationally significant changes to marine ecosystems because of diseases, parasitic infestations or mass die-offs. Recent trends appear to be stable, with only 2 major fish die-offs reported in 2011-15: 1 in South Australia in 2014 and 1 in Western Australia in 2015. This suggests that conditions have not changed since SoE 2011.
During 2011-16, blooms of the alga Karenia mikimotoi resulted in major fish kills in South Australia in 2014, and the alga Chaetoceros sp. bloomed in 2015 in Western Australia with similar results. A prolonged and extensive bloom of Alexandrium tamarense along the east coast of Tasmania in 2015-16 contaminated mussels, oysters, scallops and, ultimately, rock lobsters. Shellfish harvest areas were closed from late July to late November 2015, and some wild fisheries blocks remain closed. This followed a harmful algal bloom in Tasmanian waters in 2012, which cost $23 million in lost fishery and aquaculture production (Campbell et al. 2013; see also the Coasts report).
Based on observations of concentrations of harmful algal species from the IMOS National Reference Stations, there is no evidence of an increase in the frequency of harmful algal blooms since 2011 (Figure MAR34). Although not specifically designed to monitor harmful algal species, the network of stations provides one of a limited number of datasets from areas outside estuaries, embayments and near-coastal regions that can currently be used to investigate algal blooms. Large-scale algal blooms can also be monitored via satellite imagery; however, these datasets are limited by cloud cover, and algal blooms have to be large enough to be identifiable.
Similar to harmful algal species, many marine animals associated with 'outbreaks' occur naturally within the marine environment. It is only when spawning coincides with favourable conditions that these species reach outbreak densities and can affect marine ecosystems, marine industries and, in some cases, humans. Examples of these are the crown-of-thorns starfish7 and jellyfish blooms.
The high fecundity of crown-of-thorns starfish means that, when spawning coincides with favourable conditions, resulting recruitment can lead to outbreak densities of large starfish that can deplete local coral cover within 3-5 years (Kayal et al. 2012). Populations then collapse through starvation and disease, but not before they spawn abundant planktonic offspring, which can form secondary outbreaks on downstream reefs. There have been 4 synchronised eruptions of secondary outbreaks of crown-of-thorns starfish in the highly connected Great Barrier Reef since 1960. From 1990 to 2012, crown-of-thorns starfish have contributed to approximately 40 per cent of the overall decline in coral cover on the Great Barrier Reef (GBRMPA 2014a). Outbreaks have also been recorded in all other parts of the starfish's Australian range (far northern Great Barrier Reef, Torres Strait, north-western Western Australia), but their significance is uncertain because systematic monitoring was only recently implemented or is non-existent (see Box MAR5). Trends in crown-of-thorns starfish numbers in other regions of northern Australia are less certain because of lack of observations, but numbers appear to have increased since 2011 on both coasts. The state of the environment in relation to crown-of-thorns starfish is regarded as poor and deteriorating. Although outbreaks are difficult to predict, both the dynamics of previous outbreaks and population models suggest that the central Great Barrier Reef may experience another wave of crown-of-thorns starfish outbreaks in the next decade, with negative impacts on the abundance of coral within the reef (Mumby & Anthony 2015).
Jellyfish blooms affect marine industries by stinging humans (potentially leading to fatalities), clogging intake pipes and fishing nets, killing fish, and reducing the abundance of commercial fish through competition and predation (Richardson et al. 2009). No systematic monitoring of jellyfish currently exists for areas outside estuaries, enclosed embayments and near-coastal regions. Given a lack of reporting of large blooms of jellyfish in the marine environment, it is assumed that the current state is good, with trends currently unclear.
Outbreaks of crown-of-thorns starfish and jellyfish blooms have been associated with high run-off events that increase nutrients in the marine environment and thereby increase phytoplankton abundance-a food source for crown-of-thorns starfish and jellyfish larvae (Richardson et al. 2009, Fabricius et al. 2010). Direct drivers and processes for these outbreaks are still uncertain, and, as a result, projections of outbreaks are also highly uncertain.
More than 250 introduced marine plants and animals are established in Australian waters (NSPMMPI 2014). Some have hitchhiked to Australian waters on the hulls of vessels of all types, from yachts to commercial cargo vessels, or in their ballast water. Others have been introduced to support local aquaculture, or through the aquarium industry. Some have displaced our native species from their habitats, modifying ecosystems and affecting marine industries (Bax et al. 2002, Ross et al. 2003, Hayes & Sliwa 2003; see also the Coasts report).
Many of the species introduced to Australia, however, do not become established (i.e. survive long enough to reproduce, complete a full lifecycle and establish a population), and most established species do not become widespread or invasive in terms of their distribution and numbers. This is largely because environmental conditions at introduction sites are not suitable, and/or native species outcompete species before they can become established and invasive. Many species will remain restricted to areas in ports or other semi-enclosed areas close to their point(s) of introduction. Introduced species in these environments are discussed in the Coasts report.
Several species are more widespread in the marine environment. The New Zealand screw shell (Maoricoplus roseus) was introduced to southern Tasmania in the 1920s, probably as part of rock ballast discharged by ships taking timber to New Zealand, and is now widespread across the continental shelf (Gunasekera et al. 2005). The northern Pacific seastar (Asterias amurensis), originally from Japan, was first established in the Derwent River estuary in the early 1980s, through either hull fouling or ballast water discharge, although it was not detected for 10 years. It has now spread along the eastern Tasmanian coastline to Banks Strait and was detected in Port Phillip Bay in 1995, where it increased to approximately 30 million individuals within 2 years (Ross et al. 2003). Japanese seaweed, or wakame (Undaria pinnatifidia), was found in 1988 near Triabunna on Tasmania's east coast, probably introduced in ballast water or as biofouling, and was subsequently spread along the coast by fishing and recreational boats, where its establishment was aided by dieback of the native macroalgae canopy (Valentine & Johnson 2004).
Once introduced species have become widespread in the marine environment, control can be prohibitively expensive or unfeasible. An exception occurred in 1999 when Australia became the first country to eradicate an established introduced species, the black-striped mussel (Mytilopsis sallei), from 3 marinas in Darwin (Bax et al. 2002).
Under the EPBC Act, introduced species that threaten, or may pose threats to, native species or ecological communities can be listed as key threatening processes, and a subsequent threat abatement plan developed. To date, no introduced species has been listed.
Fortunately, the establishment and spread of introduced species remain rare events, with no change in the state and recent trends of invasive species in the marine environment since 2011. Many species have been in Australia for a significant period, and many have potentially reached an equilibrium state. If this is the case, and given that the major vectors for most introduced species (hull fouling, ballast water, the aquarium trade) have existed for some time, it could be assumed that, if translocation of additional introduced species could occur, it would have done so already. However, changes in international trade routes and a changing receiving environment can provide new opportunities.
Biosecurity measures are in place to manage several major vectors for introduction of species (ballast water and the international aquarium trade), and are under development to manage biofouling on international vessels. These biosecurity measures are designed to minimise new incursions, which are inherently difficult to predict. The history of introductions is replete with unexpected events that establish a species at one location, with the spread of the species then occurring decades later when changed environmental conditions facilitate it. This highlights the need for early detection; some states have developed programs that encourage the public to report unknown species that may potentially be introduced and invasive. Ongoing monitoring for introduced species is cost and labour intensive, and monitoring effort has varied considerably between jurisdictions and is mostly limited, despite concentrated efforts to develop monitoring systems (DAWR 2015). As a result, any limitations to Australia's national and local prevention arrangements are likely to be identified through the establishment of an introduced species. The development of new technologies may provide more viable monitoring options in the future, which are likely to not only allow early detection but also inform regulation development.