Land–water interface


Sea level rise

Climate change is driving global and regional sea level rise, and more intense and frequent extreme sea levels. Global averaged sea level has been rising at a significantly higher rate during the 20th and 21st centuries (1.7 ± 0.2 millimetres per year; mm/y) (Church et al. 2013a) than in pre-industrial times, and at an accelerating rate since the late 19th century (Church et al. 2013a, Kopp et al. 2016). The widely accepted estimate of global average sea level rise since the start of the record in 1993 is 3.2 ± 0.4 mm/y (Church et al. 2013a). Rising sea level is a result of the expansion of ocean waters and the loss of land-based ice, because of the increasing temperatures associated with climate change. The depletion of groundwater also plays a role, as most extracted groundwater ends up in the ocean through run-off, evaporation and precipitation (Church et al. 2013b, Slangen et al. 2014).

Sea level rise is felt primarily at the coast, exacerbating erosion, inundation, and loss of coastal ecosystems (Figure COA6). Currently, the impact of sea level rise is slight, but impacts will increase as trends become more pronounced during the background of storm surges and longer-term weather-related variability. Background variability makes it necessary to consider climate-related sea level rise during periods longer than 5 years, particularly at the regional level. For 1966–2009 and 1993–2009, the average trends in relative sea level around the coastline are 2.1 ± 0.2 mm/y and 3.1 ± 0.6 mm/y, respectively (White et al. 2014). The increases in sea level on the east and west coasts of Australia have resulted in a significant increase in extreme high sea levels during the 20th and 21st centuries (Church et al. 2006).

Trends and projections of Australian sea levels for the first decades of the 21st century are related to greenhouse gas emissions and begin to diverge depending on emission trajectory from about 2050. The Intergovernmental Panel on Climate Change (Church et al. 2013a) developed 4 greenhouse gas concentration trajectories, known as Representative Concentration Pathways (RCPs), based on possible mitigation scenarios. For the business-as-usual scenario (RCP8.5), the rates steadily increase through the 21st century, reaching 11.2 mm/y by 2100. For the intermediate scenarios of RCP6.0 and RCP4.5, the rates stabilise in about 2090 and 2060 at 7.4 and 6.1 mm/y, respectively. For the strong mitigation scenario (RCP2.6) where significant and urgent action occurs, the rate of rise stabilises much sooner and declines to 4.4 mm/y. Projections for 2090 relative to 1986-2005 are shown in Figure COA7 (based on McInnes et al. 2015).

The largest uncertainty in future sea level trends is the behaviour of ice sheets, particularly the Antarctic ice sheet. There is also uncertainty about the sensitivity of the climate system greenhouse gas concentrations (e.g. whether feedback loops in the system may accelerate or decelerate a warming trend), and the regional distribution of sea level change. Sea level rise is expected to continue for centuries after greenhouse gas emissions reach near zero because of the increased baseline concentrations of greenhouse gases in the atmosphere.

Erosion and inundation

Shorelines naturally erode and accrete sediments, but these processes can be exacerbated by human activities and can present major issues for human settlements. Parts of Australia’s coastline are at risk of periodic erosion and inundation because of the impact of extreme storms. Sea level rise is considered to have low impact on these processes at present, but there is the potential for erosion events to worsen in the future and lead to long-term shoreline recession.

Beach erosion, and consequent shoreline recession, increase in response to sea level rise. Originally described by the simplistic Bruun’s rule (Bruun 1962), the future response of coasts to sea level rise is now understood in terms of contemporary frameworks that involve sediment compartment and budget concepts, and response models (e.g. Cowell et al. 2003, Woodroffe et al. 2012, Kinsela et al. 2016). However, observations along many sections of the open coast during the past 40 years highlight the capacity of beaches to recover after storm events with no net shift in shoreline position. This highlights the role of local and regional sediment supplies in determining changes in shoreline position over time.

Periodic erosion is of most concern in developed areas, such as urban centres or ports, where property is at risk (see Coastal development and land use). It is most acute during extreme weather events accompanied by sustained periods of high wave action and storms, and is exacerbated by the loss of dune vegetation, shoreline modification, spring high tides, rising sea levels, and changes to hydrodynamics and sediment supply. Inundation is particularly concerning around estuaries with low-lying properties, as documented nationally by the Department of Climate Change report (DCC 2009).

Susceptibility of the Australian coastline to inundation and erosion or recession has recently been mapped at a national scale as part of CoastAdapt. This used the sediment compartment approach, and assessed all 354 Australian coastal compartments for their susceptibility to climate change (Thom 2016). Most states have a system of identifying locations that are sensitive to periodic erosion, recession and inundation in association with monitoring programs undertaken by local councils and community groups. In New South Wales, there are approximately 30 individual coastal zone management plans, and 15 coastal erosion hotspots. In South Australia, the Coastal Protection Board established under the Coast Protection Act 1972 (SA) works to manage erosion and other coastal issues, and Western Australia has a state planning policy that explicitly maps predicted erosion under various sea level scenarios.

Although we have risk assessments and a framework for managing erosion and inundation, monitoring of beach erosion and shoreline change is a major gap in most states, and a national erosion monitoring plan is needed. The best large-scale data could be obtained from repeated surveying using a combination of LIDAR (light detection and sensing; dry land), LADS (laser airborne depth sounder; underwater) and high-resolution satellite images. Intertidal and subtidal imagery is necessary because sediment deposits can provide an important source of beach nourishment and a buffer for the coastal zone. There is value in maintaining field observations because of the broader range of variables that can be measured and for the calibration of remotely sensed data.

Rising sea levels, saline intrusion and coastal erosion are likely to affect coastal ecosystems. Frontal dune systems are likely to be affected by coastal recession, whereas wading birds that breed on low-lying sandy spits are likely to be affected by increased overwash. Landward migration of seagrass beds, mangroves and saltmarshes is likely, but only where terrain and coastal structures allow, although patterns of response are likely to vary depending on sediment supply and local processes. Some wetland systems may be able to adapt to modest rates of sea level rise through vertical accretion.

The short-term outlook is continued impact of extreme storm events, leading to periodic erosion at hotspots where property and infrastructure are exposed, and inundation increasing in line with sea level. In the long term, climate-driven sea level rise is expected to gradually increase exposure of property, infrastructure and coastal ecosystems to the effects of inundation and erosion. Erosion will eventually lead to shoreline recession, and increase the area of inundation as a result of both storm overwash and higher tides in estuaries. Long-term changes in shoreline behaviour will also be dependent on changes in wave climate, cyclone frequency and intensity, tide range and sediment supply, so caution is needed in making predictions based on sea level change alone.

There is, however, potential for long-term improvement if management appropriately considers the dynamic and ambulatory nature of the coasts. Adaptive planning in the context of sea level rise is necessary to manage risks to existing infrastructure and avoid future risk. Some erosion hotspots may require managed realignment, whereby coastal property and infrastructure are moved inland. An important consideration in coastal armouring is the ‘pinch effect’, which refers to the squeeze on freshwater wetlands from both land and sea directions. Sea level rise and the movement of saline water inland create pressure from the seaward side, whereas barriers erected to protect urban and agricultural lands from rising water and salinisation create pressures on the landward side (Sheaves et al. 2007).

Sediment transport

Sediment input to coastal waterways varies greatly around Australia, and is often site specific. Transport of sediments by wind and water is a natural process that shapes the geomorphology of beaches and estuaries, but anthropogenic factors can change the spatial and temporal patterns of movement, as well as the quantity and type of sediment transported.

Catchment modification is one of the major drivers of change, and generally increases sediment input. High proportions of annual sediment budgets can come from the catchment and be transported during flood flows (Hossain & Eyre 2002). Sedimentation of navigation channels is also a major management issue for ports and harbours. Coastal lagoons and shallow coastal waters are among the most vulnerable habitats, because of their limited volumes and rates of water exchange with the ocean. Sediment input and nutrient input are closely linked processes, and many coastal waterways with elevated sediment input also suffer from nutrient enrichment (see Nutrient pollution).

Sedimentation influences a range of taxa, including mangroves, saltmarshes, corals and other sessile invertebrates, seagrasses, phytoplankton, microphytobenthos, and infauna (Todd et al. 2015). Impacts of excess sediments on pelagic (open-ocean) species include smothering, gill irritation and reduced light penetration, whereas impacts on benthic (ocean-floor) species arise from changes in grain size distribution, light penetration and seabed depth. Increased density of fine suspended sediments can also influence the behaviour and functioning of vertebrates (e.g. fish, turtles and dugongs), or have indirect impacts through changes in prey abundance.

In addition to sediment input, increased resuspension of bottom sediments can have significant ecological impacts (Knott et al. 2009), change water quality parameters and alter patterns of sediment transport. Sediment resuspension is exacerbated by vessel movements in shallow waters, dredging operations, and changes in hydrodynamics because of coastal engineering. Resuspension of contaminated sediments is particularly concerning because it releases contaminants into the water column, increasing their bioavailability and bioaccumulation (Hedge et al. 2009).

For the Great Barrier Reef coast, the Reef Plan report cards (2014) present sediment, nutrient and pesticide loads to the coast, and include modelled estimates of how they have changed because of catchment management works. Modelling of river plumes has been done for the Great Barrier Reef (e.g. Kroon et al. 2012; Álvarez-Romero et al. 2013; Fabricius et al. 2014, 2016). Although some dredge plume modelling has been done, it is generally not published in peer-reviewed literature (but see review by McCook et al. 2014). Sediment and nutrient transport to coastal waters is understood to be one of the major ongoing threats to the health of the Great Barrier Reef. Measures to reduce sediment inputs in Queensland are currently under way, although it is likely to take many years to achieve any outcomes (Reef Water Quality Protection Plan Secretariat 2014).

Sediment inputs from several north coast rivers in New South Wales have been studied by Eyre and Ferguson (2006), among others, but there is no ongoing comprehensive monitoring of sediment inputs to waterways of New South Wales. Research on storm water in Sydney Harbour (Birch & Rochford 2010) and plumes in selected locations near Sydney (Birch & O’Hea 2007) provides some insight into the role of urbanisation in sediment transport. A study of 184 New South Wales estuaries found that total suspended sediments were 16 to 10,594 per cent higher in areas of significant human land use, compared with relatively undisturbed catchments (Roper et al. 2011).

Assessment of the impacts of catchment management decisions on sediment loads requires either modelling of source inputs or monitoring across decades. Such long-term monitoring programs are economically costly, and have only been implemented at a small number of locations (e.g. Great Barrier Reef, Port Phillip Bay, Moreton Bay, Nepean–Hawkesbury River, Swan River and Blackwood River). More information on the effect of sediment transport on bedded sediments is crucial, as they are often the dominant habitat type and are good indicators of integrated environmental condition over time (Burton & Johnston 2010).

Coastal sediments around Australia have recently been mapped in the Coastal Sediment Compartments Project, which aims to improve coastal risk assessment by classifying coast based on landforms and patterns of sediment movement. It uses 3 levels to capture processes at different scales, and each level is suitable for different types of decision-making:

  • primary—based on the influence of large landforms and offshore processes; suitable for regional planning or positioning of large-scale engineering such as ports
  • secondary—based on medium landforms and regional sediment processes; useful for smaller engineering or local planning decisions
  • tertiary—based on individual beaches; suitable for very small projects that are unlikely to restrict sediment movement, such as deciding the exact location of a groyne or sea wall within a broader management plan.

The outlook for sediment transport varies around Australia. Sediment quality is improving in some urban areas, particularly where organic inputs have declined because of changed waste management and/or the restoration of riparian vegetation, but coastal sprawl into previously undeveloped areas is increasing sediment and contaminant transport. Continued land clearing and poor agricultural practices result in increased transfer of terrestrial sediments to aquatic systems. Management should focus on combined sediment and nutrient input reductions, since the 2 inputs often co-occur, and can have compounding or interactive effects on sediment and water quality.

Artificial structures

The addition of artificial structures is one of the first modifications made to estuaries and coastal foreshores. Structures include marinas, ports, seawalls, groynes, wharves, aquaculture facilities, stormwater pipes and breakwaters. An estimated 20 per cent of New South Wales estuaries have between 4 and 20 per cent of their perimeter modified by a foreshore structure (Roper et al. 2011), and, in some estuaries, more than 50 per cent of the foreshore is artificial (e.g. Sydney Harbour; Chapman 2003). Approximately 10 per cent of the Great Barrier Reef World Heritage Area coastline has been modifiedan increase of 70 per cent in 3 years in some areas because of port developments (Waltham & Sheaves 2015).

Pressures associated with artificial structures (reviewed by Dafforn et al. 2015) include:

  • removal of soft-sediment, rocky reef and mangrove habitats
  • stuctures acting as stepping stones and havens for invasive species
  • increased shading underneath structures during the day
  • artificial light surrounding the structures at night
  • introduction of contaminants
  • altered hydrodynamics and erosion.

Artificial seawalls often support lower biodiversity than natural rocky shores, which is potentially related to their vertical orientation, lower surface area and fewer microhabitats (Chapman 2003, Ferrario et al. 2016, Bishop et al. in press). A recent meta-analysis found that seawalls supported 23 per cent less biodiversity and 45 per cent fewer organisms than natural shorelines (Gittman et al. 2016). In some instances, built structures can contain higher densities of introduced species than adjacent natural rocky reefs as a result of the novel habitat they provide and their proximity to vessels that may carry introduced species. It remains unclear whether populations on these structures function as well as they would in natural habitats. In addition to the direct effects of structures themselves, there are also pressures from associated activities such as shipping and fishing.

Eco-engineering, as has been implemented in parts of Sydney Harbour, aims to incorporate ecological principles into the design of artificial structures to reduce their impact and provide useful functions such as increased biodiversity (Browne & Chapman 2011). Until eco-engineering principles are adopted more widely, however, artificial structures at a national scale will continue to remain poor substitute habitat for marine communities. Best-practice guidelines that are supported by comprehensive research are a priority, to improve the design, deployment and monitoring of artificial structures. This research priority was outlined in the National Marine Science Plan 2015 (Treloar et al. 2016), and the Hawkesbury Shelf marine bioregion assessment identified eco-engineering as a key solution-based research need.

Looking forward, as artificial structures continue to spread as coastal features, there is a need to balance engineering needs with environmental protection and conservation. In many cases, implementing eco-engineering designs early in the development and construction phase can be less expensive than hard engineering solutions. The benefits of eco-engineering can also extend to include social and economic stability in coastal areas across Australia.


Desalination plants are increasingly common in major Australian cities. Most are designed to serve as secondary water supplies for use during drought. Desalination plants draw in large volumes of sea water from which fresh water is extracted, creating a byproduct of highly saline brine. Impacts on the coastal environment include the footprints of desalination plants on coastal land, the intake of large volumes of sea water with associated plankton, and the release of the brine effluent into coastal waters. The brine can affect marine biota through higher salinity (often double that of ambient sea water), changed flow conditions and, in some cases, contaminants (Roberts et al. 2010). Environmental impacts are somewhat proportional to the size of the plant, which varies from small community potable water supplies to large plants for major cities. However, impacts can be ameliorated if brines are discharged into high-energy or highly dispersive waters.

Pressures associated with desalination are marginally increasing because more large-capacity desalination plants are being commissioned, although several of these are currently nonoperational. Of the plants that are operating, there is little evidence of impact beyond the vicinity (hundreds of metres) of the outfalls, and affected areas usually represent a small proportion of that habitat type. Technological advances, such as high-velocity diffusers that promote mixing of brine with surrounding sea water, are also helping to minimise the impacts of hypersalinity.

Ecological impacts are generally monitored by changes to water quality parameters, such as salinity and dissolved oxygen. Impacts are usually attributed to the stress of hypersalinity, although there have not been manipulative tests at the appropriate scales to separate salinity stress from other potential stressors. Uncertainty still surrounds potential impacts of increased flow and shear stress, chemicals used to protect the membranes of desalination cells when the plants are not operating, and antiscalants used during the commissioning and maintenance of pipework (Roberts et al. 2010).

Substantial environmental impact assessments are performed for large desalination plants, although the amount of monitoring following construction varies. The Sydney Desalination Plant implemented a robust monitoring program (Clark & Johnston 2014), and the Victorian, Western Australian and South Australian major desalination plants were well assessed at the environmental impact statement stage. The Victorian plant is yet to become fully operational. The surrounding waters of the Western Australian plants are monitored for water quality; however, monitoring data are not available for ecological impacts apart from effects on seagrass. In South Australia, the Adelaide (Gulf St Vincent) and proposed Point Lowly (Spencer Gulf) plants were well assessed initially, but there is ongoing debate about the potential impacts of the Point Lowly plant. There is concern that the calm waters of the South Australian gulf increase the risk of negative impacts (Kämpf et al. 2009), but only a minor salinity rise (less than 0.5 parts per thousand) was recorded within 100 metres of the Adelaide plants. This, together with cuttlefish toxicological studies, suggests that the risk posed by the hypersaline effluent is low (Dupavillon & Gillanders 2009). The Perth and Adelaide desalination plants have ecological monitoring (before–after control–impact), although data are generally not publicly available.

The short-term outlook for desalination as a pressure is stable, because many existing plants are not operating at capacity and the addition of new plants is unlikely. In the longer term, impacts from desalination are expected to rise, given increasing human population and climate-driven aridity (particularly in south-eastern and western Australia). This will likely be a low-risk pressure, however, because desalination plants are economically costly to build and maintain, and their impacts are highly localised. It is likely that ongoing global research will improve the management and design of desalination plants to further minimise direct environmental impacts.

Clark GF, Johnston EL (2016). Coasts: Land–water interface. In: Australia state of the environment 2016, Australian Government Department of the Environment and Energy, Canberra,, DOI 10.4226/94/58b659bdc758b