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3.2Overview of Critical Services, Components and Processes

A summary of the critical and supporting wetland components, processes and services/benefits for the Kakadu National Park Ramsar site as determined in the present study is shown in Table 3 -10. In summary, the following have been identified:

  • eleven critical components and six supporting components

  • four critical processes and five supporting processes, and

  • three critical services/benefits and five supporting services/benefits.

The broad interaction of wetland components, processes and services/benefits (both critical and supporting) at a whole-of-site level is shown in Figure 3 -22. The figure shows three broad processes (climate, geomorphology and regional-scale hydrodynamic and hydrological processes) that together have shaped the topography, marine and freshwater flow regime and other important aspects of the site. At the local habitat scale, there is a mix of physical and chemical processes as well as biological processes that control the wetland habitats and associated biota. The interaction of the wetland components with the wetland processes yields a range of wetland services/benefits that are characterised as biodiversity (ecosystem services) and cultural services (relevant to providing a social or economic benefit to humans) using the terminology in the National Framework and Millennium Ecosystem Assessment.

The following sections provide a more detailed description of critical components, processes and services for Kakadu National Park that form the basis of this ECD. Where possible, information on the different original listing and extension dates (1980,1989,1995 and 2010) has been identified to aid in the description of natural variability for the components, processes and services at the time of listing.

Table 3 10 Summary of critical and supporting components, processes and services/benefits





C1 – Mangroves

C2 – Melaleuca Forests

C3 – Palustrine Wetlands and Billabongs

C4 – Waterfalls, Seeps and Waterholes

C5 – Populations of Migratory and Resident Waterbirds

C6 – Populations of Freshwater Fish

C7 – Populations of Freshwater and Saltwater Crocodiles

C8 – Populations of Threatened Sharks

C9 – Yellow Chat Populations

C10 – Pig-nosed Turtle Populations

C11 – Locally Endemic Invertebrate Species

P1 – Fluvial Hydrology

P2 - Fire Regimes

P3 – Breeding of Waterbirds

P4 – Flatback Turtle Nesting

S1 – Maintenance of Global Biodiversity

S2 – Fisheries Resource Values

S3 – Contemporary Living Culture



Monsoon Rainforests and Riparian Vegetation

Other Wetland Habitats

Terrestrial Habitats

Aquatic Invertebrates

Regionally Endemic Species



Tidal Hydraulics


Water Quality

Ecosystem Processes

Recreation and Tourism

Scientific Research and Education

Historical Cultural Heritage

Biological Products

Sites/Items of Cultural Significance

Figure 3 22 Conceptual model showing interactions between critical and supporting components, processes and services/benefits within the Ramsar site

3.3Critical Components

As outlined in Section 2.4, a range of wetland habitat types are known to be present within the Ramsar site boundary, including a diversity of those designated within the coastal/marine and inland wetland categories under the Ramsar classification scheme. Within these habitat types, a rich diversity of wildlife exists from all the major groups of organisms (from planktonic organisms to vertebrates) which make up the ecosystem components of the wetland. Critical components of the site have been selected on the basis of those habitats, key species and wildlife populations that are fundamental in determining the site’s ecological character and underpin the critical services/benefits, and are described below.

3.3.1C1 – Mangroves

Reasons for Selection as ‘Critical’

Mangroves were selected as a critical component due to the essential role that these communities have in terms of provisioning for species of fisheries value (Service 2). Furthermore, mangrove communities are important with respect to habitat provisioning for waterbird breeding colonies, and therefore, in turn, mangrove communities are fundamental to determining the site’s ecological character, contribute to support of Ramsar Nomination Criteria 4, 5 and 6, and provide opportunities for recreation and tourism.


Refer to Wetland Type I in Section 2.4.1.

Patterns in Variability

There are two data sources describing patterns in mangrove extent: Mitchell et al. (2007) and Cobb et al. (2007).

Mitchell et al. (2007) provides the most detailed mangrove mapping within Kakadu National Park (refer Table 3 -11). The assessment was undertaken based on 1990 aerial photography, which is close to baseline pre-listing conditions in Stage II (1989), and includes the estuarine portions of Wildman, West and most of South Alligator Rivers, as well as Field and Barron Islands. The estuarine portions of East Alligator River are almost entirely within Stage I, which was declared in 1980, and therefore these data describe mangrove extent post-Ramsar site listing.

Table 3 -11 shows that West and South Alligator Rivers had the largest mangrove areas. Tree height, as a relative measure of tree age, varied inconsistently between locations and broad geomorphological zones within locations (inland, coast and creek). Mitchell et al. (2007) concluded that the patterns in tree height are related to contemporary changes in erosion and sedimentation patterns.

Cobb et al. (2007) examined changes in mangrove area over time within Kakadu National Park. Note that this assessment was based on a review of remote imagery that did not involve ground-truthing, and was based on different mapping methods to those used by Mitchell et al. (2007). Mangroves extent was mapped for 1950 and 1975 (pre-listing for the site), 1985 (pre-listing for Stage II, post-listing Stage I) and 1992 (post-listing for both sites).
Table 3 11 Mangrove area and canopy height mapped from 1990 aerial photography (source: Mitchell et al. 2007)


Listing date

Mangrove area (ha)

Tree height range (metre)




West Alligator

Stage II - 1989


6 (edge) 20 (central)



South Alligator

Stage II - 1989






Stage II - 1989



7-13 (19 mouth)


East Alligator

Stage I - 1980

688 (coast)

10-14 (10 inland)



Field Island

Stage II - 1989



5 (fringe); 18 (shore)


Barron Island

Stage II - 1989





Blue shading – data collected post listing and unlikely to reflect conditions at time of listing
Cobb et al. (2007) demonstrated that there was a long-term trend of mangrove expansion within the site between at least 1950 and 1991 (refer Figure 3-2). Figure 3-2 shows that the rate of change varies between catchments, as well as between time periods (Cobb et al. 2007).

The highest rate of change occurred at West and South Alligator Rivers (Stage II – 1989 declaration), particularly after 1975. Between 1984 and 1991 mangroves increased by 13 and nine square kilometres at South and West Alligator, respectively, whereas a three to four square kilometres increase was recorded at Wildman and West Alligator Rivers. While recognising that the 1991 data-set describes conditions just after listing, for the purposes of this study, these patterns in variability between 1950-1991 are assumed to represent baseline, pre-listing conditions for Stage II estuarine waterways.

In terms of Stage I estuarine waterways (East Alligator River), pre-listing mangrove extents were mapped on two occasions: 1950 and 1975. Between these two periods, mangrove extent increased from nine to 9.5 square kilometres. Note that by 1991 (11 years post-listing), mangrove extent had increased to 15 square kilometres.

Figure 3-2 also shows that the rates of change in mangrove extent varied among catchments. Between 1950 and 1991 the annual rate of increase in mangrove extent ranged from 0.12 square kilometres per year (Wildman), to 0.43 square kilometres per year (West Alligator), and 0.49 square kilometres per year (South and West Alligator Rivers). The highest rate of mangrove increase was between 1984 and 1991 at South Alligator River (2.1 km2 per year), West Alligator River (1.5 km2 per year), East Alligator River (0.6 km2 per year) and Wildman (0.42 km2 per year).

With respect to mangroves, it is evident that there is an inherent difficulty with quantifying ‘natural variability’ when the overall trend or trajectory is operating over a longer timescale than that assessed during the measurement period.

Figure 3 23 Area of mangroves mapped from 1950 to 1991 aerial photographs (source: Cobb et al. 2007)

Note: ‘funnel’ refers to the estuarine funnel at the river mouth, ‘cuspate’ refers to cuspate meanders, ‘sinuous’ refers to sinuous meanders and ‘fluvial’ refers to the main channel of the river.

3.3.2C2 - Melaleuca Forests

Reason for Selection as ‘Critical’

In a study that identified qualities of the Alligator Rivers Region that favoured National Park establishment, one of the recognised qualities was vegetation communities associated with various landscape features and the concentration of aquatic birds that occur in the swamp lands during the dry season (refer Fox et al. 1977). Melaleuca forests fit the description of these vegetation communities, and as such were selected as a critical component.


Refer Wetland Type Xf in Section 2.4.2.

Patterns in Variability

Melaleuca forest can display natural variation in the extent of areas covered as well as the density of the forest. Due to the longevity of Melaleuca individuals, natural variability of Melaleuca forest communities is typically not detected over short timescales, with the exception of sudden dieback events that result from certain unexpected changes in environmental conditions. Specifically, saltwater intrusion is one of the principal factors contributing to the loss of Melaleuca communities within the Ramsar site. In contrast to this, concern has also been expressed regarding possible encroachment of Melaleuca forests over long timescales into areas of the Ramsar site that were previously not occupied by Melaleuca (for example, see Riley and Lowry 2002).

As discussed in section 2.4.2, Brocklehurst and van Kerckhof (1994) mapped Melaleuca forest extent in Kakadu National Park in 1992, which describes post-listing conditions for Stage I and II, and pre-listing condition for Stage III. This was a one-off study, and therefore, there are no empirical data describing natural variability in Melaleuca forest over time at the whole-of-site scale. However, a number of studies have examined changes in Melaleuca extent on localised scales, as described below.

Winn et al. (2006) used aerial photography interpretation to examine changes in Melaleuca extent (and morphological change – see section 3.2.3) at the mouth of the East Alligator River (Stage I) between 1950 and 19971. Between 1950 and 1975 (pre-listing), dramatic increases in tidal creek expansion (and associated saltwater intrusion) occurred, resulting in a 45 percent reduction in Melaleuca forest extent. Around the time of listing, a further 26 percent reduction in Melaleuca extent occurred between 1975 and 1984, and a further seven percent reduction occurred between 1985 and 1997. Similar to temporal trends in mangrove extent (and intertidal flat extent; see Winn et al. 2006), it is apparent that there has been a long-term trend of increased salt water intrusion into freshwater wetlands (and associated changes in vegetation communities), which has been evident for at least 30 years prior to Ramsar site listing. These changes represent part of the ecological character of the Ramsar site.

There has been a large body of research describing changes in Melaleuca density and/or extent in the Magela floodplain, which is predominantly within Stage II (1989 listing). Two studies examined changes in Melaleuca forests prior to Stage II listing: Williams (1984) examined Melaleuca densities, and Staben (2008) examined Melaleuca extent. Williams (1984) reported an overall decline in the density of Melaleuca on the Magela floodplain between 1950 and 1975.

Staben (2008) mapped Melaleuca extent on the Magela floodplains during four time periods: 1950, 1975, 1996 and 2004 (Figure 3-3). In contrast to the observed reduction in Melaleuca densities documented by Williams (1984), Staben (2008) observed an increase in Melaleuca extent between 1950 (118.9 hectares) and 1975 (133.4 hectares) time periods. Between 1975 and 1995 (i.e. six years after listing of Stage II), Melaleuca extent remained almost static (three hectare increase), but declined in 120.4 hectares by 2004. Staben (2008) suggested that differences in the results of these two studies could relate to differences in mapping methodologies and/or sampling (classification) errors in either of the studies.

Two other studies undertaken in the Magela floodplain describe changes in Melaleuca forests between a pre-listing sampling episode (either 1975 or 1983) and one post-listing sampling episode (1996, 2006). Riley and Lowry (2002) observed a net decline in the number of trees between 1975 and 1996 in the study area as a whole, as well as within four out of five sub-areas that together comprised the study area (refer Table 3 -12). Similarly, Boyden et al. (2008) examined changes in the percentage cover of Melaleuca on the Magela floodplain between 1983 and 2003, and found a ten percent decrease in cover (refer Figure 3 -25). These two studies are consistent with the findings of all other studies undertaken on the Magela floodplain except Staben (2008).

A number of processes could ultimately control these temporal patterns. Williams (1984) suggested that factors such as late dry-season fires, strong winds, buffalo and in some instances saltwater intrusion may be responsible for the decline in Melaleuca trees prior to listing. Other authors also highlight the potential impacts of fire, feral pigs, changes in rainfall patterns and successional changes due to sediment accumulation. This suggests that both natural and anthropogenic factors could have resulted in the observed changes to wetland communities in the period prior to listing.
Table 3 12 Number of Melaleuca trees on the Magela Floodplain (source: Riley and Lowry 2002)


    Sub-area 2

    Sub-area 3

    Sub-area 4

    Sub-area 6

    Sub-area 7





    11 238


    10 317

    31 433






    13 695

    24 704








Figure 3 24 Spatial extent of Melaleuca cover (green shading) on the Magela floodplain for four time periods (source: Staben 2008)

3.3.3C3 – Palustrine Wetlands and Billabongs

Reasons for Selection as ‘Critical’

Fox et al. (1977) considered these habitats as important land type features that are well-represented within the Alligator Rivers Region, and therefore these habitats contributed significantly to proposals for declaration as a National Park. Furthermore, palustrine wetlands and billabongs were selected as a critical component as they are fundamental to determining the site’s ecological character.

Specifically, billabongs provide areas of deep water habitat for aquatic flora and fauna, as well as dry season refuge for many of the aquatic fauna species that inhabit the floodplains. These fauna species include a diversity of freshwater fish (Criterion 7), a large number of waterbirds (Criteria 5 and 6), certain threatened species (for example, pig-nosed turtles – Criterion 2 and 9) and a number of traditional food species (for example, file snakes, freshwater turtles – see Section 3.8.4). Additionally, many traditional dietary staple plant species are associated with billabongs (for example, water lilies – see Section 3.8.4). Billabongs such as Yellow Water are also of value due to their tourism and recreational significance (section 3.8.1).


Refer Wetland Types O, Tp and Ts in Section 2.4.2.

Patterns in Variability

The spatial arrangement of palustrine wetlands within the landscape emulates some constancy due to geology and topo-climatic patterns, however the distribution and abundance of wetland plant species exhibits substantial natural variation over time (Finlayson 2005). Parameters contributing to the dynamic nature of palustrine wetlands may include rainfall, fire and magpie goose Anseranas semipalmata foraging.

While broad-scale mapping of billabongs are available (refer Section 2.4), there are no empirical data describing natural variability of palustrine wetlands over time at a whole-of-site scale. Furthermore, there area no data describing changes in vegetation communities at more localised spatial scales prior to listing.

The only study to date describing temporal changes in vegetation distribution was undertaken on the Magela floodplain (located in Stage II) between 1983 (six years prior to listing) and 2003 (14 years after listing) (Boyden et al. 2008). Boyden et al. (2008) mapped the relative change in the percentage cover and distribution of eight native vegetation classes in the Magela floodplain (refer Figure 3 -25). Most plant classes exhibited little change, except for Eleocharis that decreased by 57 percent and Nelumbo that decreased by 85 percent. While this study considered changes occurring both before and after site listing, the study provides a basis for demonstrating the nature of temporal variability in these communities.

Empirical data for billabongs are limited to the mapping layer for the catchment of the South Alligator River (refer Section 2.4.2). As such, patterns in natural variability over time can not be quantitatively described. However, it is known that a high degree of variability is exhibited by billabongs in terms of their size and permanency, predominantly due to seasonal and annual variations in climatic parameters (particularly rainfall) (for example, see Finlayson et al. 2006).

Figure 3 25 Generalised vegetation class changes on the Magela floodplain between 1983 and 2003 (source: Boyden et al. 2008)

3.3.4C4 – Waterfalls, Seeps and Waterholes

Reasons for Selection as ‘Critical’

Key features that were identified in proposals for the National Park declaration included the plateau and its gorges and the wilderness values of these areas, as well as the main escarpment, scenic landforms and associate flora and fauna (refer Fox et al. 1977). Furthermore, these habitats were selected as a critical component as they support endemic aquatic invertebrates (Criterion 3 and Critical Service 1) and provide opportunities for recreation, tourism and scientific research (for example, Wilson et al. 2009).


As briefly mentioned in Section 2.4.2, the sandstone cliffs of the Arnhem Land Escarpment in Kakadu National Park contain a number of waterfalls. Water levels of the creeks forming the waterfalls typically drop very quickly at the end of the wet season, and the waterfalls are either greatly reduced in volume or may totally dry up, leaving a number of deep pools at the base of the escarpment.

The top of the plateau is generally a harsh and dry environment. However, water seeping from rock walls, together with deep alluvial soils, has allowed development of tall monsoon forests within gorges. These habitats are important for fauna refuge during the drier months, and also support a number of threatened flora species2. In places the dominant plant species is Allosyncarpia ternata, a tree species that is endemic to the Kakadu and Arnhem Land region.

Patterns in Variability

There are no empirical data describing variability over time in extent of permanent waterholes and seeps in the stone country.

3.3.5C5 - Populations of Migratory and Resident Waterbirds

Reasons for Selection as ‘Critical’

Populations of migratory and resident waterbirds are fundamental to determining the site’s ecological character, and were noted as a feature of importance with respected to proposals for declaration of the National Park (refer Fox et al. 1977). Furthermore, populations of migratory and resident waterbirds were selected as a critical component due to the support for Ramsar Nomination Criteria 4 and 5, the opportunities presented for recreation and tourism, the provisioning of traditional foods (Section 3.8.4) and opportunities presented for scientific research (for example, Traill et al. 2010a,b).


The Ramsar site supports significant overall numbers of waterbirds, as well as significant numbers of individual bird species. Specific details have been described for justification of the Ramsar Nomination Criteria in Sections 2.5.7 and 2.5.8 (also see Appendix C for waterbird species list).

Waterbirds feed on aquatic invertebrates, vertebrates such as fish and frogs, and plant material. As such, these populations of migratory and resident waterbirds are important to ecosystem functioning, particularly with respect to wetland nutrient cycling processes. Herbivorous birds (for example, magpie geese Anseranas semipalmata, wandering whistling duck Dendrocygna arcuata) are most abundant and are supported by the vast expanse of vegetation.

Waterbirds are also important with respect to plant recruitment processes. Specifically, waterbirds may disperse seeds through endozoochory (ingestion of seeds) or epizoochory (for example, transportation of seeds in mud stuck to feet), and intensive turning over of floodplain soils occurs as a result of magpie geese foraging for water chestnut Eleocharis dulcis tubers.

Species that are listed under international migratory agreements are listed in Appendix C. Important sites for migratory shorebirds include:

  • intertidal feeding habitats of Field Island (roost and intertidal feeding sites); the coast between Finke Bay and East Alligator and between the South Alligator River and Minimini Creek (and nearby open saline wetlands)

  • roost sites of Field Island and beaches of West Alligator Head area (especially Middle Beach – an important mainland site)

  • inland wetlands associated with all the major rivers, including: Gulungal, Narramoor, Kapalga South, Kapalga Ruins, Fischer’s Hole Billabongs; Magela Creek; Mirangie Spring, Obiworbby Spring, Gaden’s Spring, Boggy Plain, Causeway Point, Jarrahwingkoombarngy Swamp, and Billyangardee Spring, and

  • for individual species – the mouth of East Alligator River (black-tailed godwit Limosa limosa); South and East Alligator River wetlands (lesser sand plover Charadrius mongolus), and Boggy Plain, Jarrahwingkoombarngy Swamp and Billyangardee Spring (little curlew Numenius minutus).

There is a range of biological processes that, together with physical (abiotic) processes described elsewhere, are critical to the maintenance of wetland ecosystem functioning and waterbird values. The availability of food sources will affect the frequency and intensity of use of the site as a feeding habitat by waterbirds, noting that a broad range of feeding techniques are used by the array of waterbirds that use the site. These feeding adaptations range from shorebirds feeding on macroinvertebrates within intertidal habitats to herbivorous waterbirds of the freshwater floodplain wetlands. The following is a summary of some of the key processes which are required to maintain feeding habitat values for waterbirds:

  • freshwater flow regimes to support freshwater wetland characteristics and buffers to increasing salinity levels

  • primary and secondary productivity of aquatic flora, algae and micro- and macro-invertebrates within shallow wetland habitats

  • water quality to a level required to support high primary and secondary productivity, and

  • maintenance of natural patterns of tidal inundation. Tidal inundation influences intertidal feeding habitat characteristics, that is, overall extent, productivity and daily availability to shorebirds.

Patterns in Variability

For the majority of waterbird species, birds appear to migrate into the Alligator Rivers Region from shallower floodplains to the west, from sub-tropical areas to the south, from distant locations in southern Australia, and from the northern hemisphere (Morton et al. 1993a,b, Bamford 1990, Bayliss and Yeomans 1990, Chatto 2000-2006). For waterbirds (excluding migratory shorebirds), a consistent pattern in changes in waterbird abundance is primarily linked to the seasonality of rainfall and the annual wetting and drying cycle of tropical wetlands (Morton et al. 1990a, Bayliss and Yeomans 1990, Chatto 2006). Rainfall is highly seasonal so that wetlands are flooded annually (wet season December to March) and then followed by an extended dry season (April to November). For waterbirds (excluding migratory shorebirds), this pattern of abundance is generally characterised by lower numbers during the wet season (typically dispersed throughout floodplain habitats to nest), then increasing to dramatic peaks in the late dry season and concentrating in higher densities at remnant wetlands as the only remaining sources of permanent freshwater and food (Morton et al. 1990a, 1993c). The influx is typically dominated numerically by magpie geese, but all species become more abundant in the dry season (Morton et al. 1990a; Bayliss and Yeomans 1990; Chatto 2006). Whilst variations in abundance and distribution between years within the same season have been recorded, the most likely explanation is reflected in annual and local variations in rainfall (Morton et al. 1990a).

In regards to the migratory shorebird component of the waterbird assemblage, there is an increase in numbers between July and September (linked to the arrival of birds migrating from northern breeding grounds), then followed by a decline over the wet season as some birds continue migration or move into freshly inundated wetlands, then an exodus of those remaining birds from May to July (Chatto 2006). There is some evidence which indicates that there is variation in the timing of peak abundance of certain species (though not significantly altering the overall pattern of shorebird abundance mentioned above) and this is probably linked to differences in migratory paths used to enter and depart from Australia (Geering et al. 2007). Examples include higher abundances coinciding with northern migration (marsh sandpiper Tringa stagnatilis, common sandpiper Actitis hypoleucos, sharp-tailed sandpiper Calidris acuminata) and southern migration (little curlew, oriental pratincole Glareola maldivarum) (Chatto 2003b, 2006). Differences in a species’ use of site on southward and northward migration elsewhere within the flyway have been widely reported (for example, Minton 2004, Gosbell and Clemens 2006). Interpretation of changes in abundance for migratory species needs to also consider potential external factors (potential variability in breeding success) and in particular, anthropogenic impacts to key stopover sites within other parts of the flyway.

Of the three general functional groupings of migratory shorebirds recorded in Kakadu National Park (species of marine shorelines; species of freshwater shorelines; and species of grasslands), Bamford (1990) noted that most species of freshwater shorelines and grasslands were scarce or absent from early wet season (December) until towards the end of the following dry season (September-November). Of these, the little curlew provides a dramatic example, where numbers may increase dramatically within a short period (30 000 to 50 000 birds recorded), and then plummet, all within the period September through to October with a large proportion of the population moving to the sub-coastal floodplains some distance to the east and west of the Alligator Rivers Region (Bamford 1990, Morton et al. 1991). Those species of marine shorelines, not unexpectedly, do not exhibit the concomitant pattern of change in abundance as their preferred wetland habitat is not affected to the same extent by the seasonal effects of flooding and drying.

3.3.6C6 – Populations of Freshwater Fish

Reasons for Selection as ‘Critical’

Freshwater fish are a key component of aquatic ecosystems throughout the Ramsar site, and were noted as an important biological feature in proposals for National Park declaration (refer Fox et al. 1977). The importance of freshwater fish populations within the site is also reflected through the support of Ramsar Nomination Criterion 7, the presence of endemic species (Service 1) and the fisheries resource values provided by the site (Critical Service 2).


Freshwater fish abundance and diversity in Kakadu National Park are high in an Australian context (Allen et al. 2002). Freshwater fish can be found in all aquatic habitats, including floodplain wetlands, billabongs, creeks, rivers and permanent pools at creek headwaters in the escarpment. Due to differences in habitat characteristics and food resources, the greatest fish diversity is supported by channel, backflow and floodplain billabongs, whilst escarpment habitats and sandy creek beds typically contain the lowest fish diversity (Cowie et al. 2000).

Freshwater fish are a key dietary component for the top aquatic predators in the site (for example, crocodiles, fishing eagles) and therefore contribute to controlling ecosystem processes and biological interactions. Barramundi are also opportunistic predators, primarily feeding on aquatic invertebrates and fish. Typically the diet of larger barramundi consists of 60 percent fish and 40 percent crustaceans (predominantly prawns/shrimp), whilst smaller barramundi primarily feed on crustaceans (Allsop et al. 2006).

Densities of freshwater fish are highly seasonal and are related to flooding and water depth. Most fish migrate seasonally, moving out of dry season refuges at the start of the wet season to colonise wetted creeks, floodplains and billabongs (Cowie et al. 2000). The proliferation of freshwater fish during the wet season and their progressive concentration in shrinking water bodies from mid to late dry season presents ideal feeding conditions for fish-eating birds (Cowie et al. 2000). Exclusively fish-eating species include darters Anhinga novaehollandiae, little black cormorants Phalacrocorax sulcirostris, Australian pelicans Pelecanus conspicillatus, ospreys Pandion haliaetus and great egrets Ardea alba.

Freshwater fish communities in the upper Arnhem Land escarpment are distinctly different to those in the lowland floodplain habitats as a result of a gradient in environmental conditions occurring along the creek systems of the site (Gardner et al. 2002). Escarpment communities inhabit aquatic areas with low temperatures, high dissolved oxygen and low turbidity over a rocky substrate. In contrast, habitats in the lower freshwater reaches are typically warmer, with lower dissolved oxygen and soft sediment substrates (Gardner et al. 2002).

Most freshwater fish migrate seasonally, moving out of dry season refuges (for example, permanent billabongs) at the beginning of the wet season to colonise floodplains. The main purpose of undertaking these migrations is to benefit from increased food availability during the wet season and to breed (Griffin 1995, Cowie et al. 2000). Most of the fish species in Magela Creek, a seasonally flowing tributary of the East Alligator River, show a peak in breeding activity soon after the start of the wet season (Bishop and Forbes 1991).

Approximately 20 percent of species within the Alligator Rivers Region are catadromous, migrating from freshwater areas to breed in marine or estuarine environments (Bayliss et al. 1997). The remaining fish species are typically potadromous and do not have an obligate estuarine phase (i.e. primarily migrate from rivers to floodplain areas to use increased habitat and food availability during the wet season). Perhaps the most important catadromous fish species within the site is barramundi Lates calcarifer (refer Section 3.7.2). Mature barramundi (and also possibly the Ord River mullet and tarpon) migrate to coastal areas early in the wet season to breed. Barramundi eggs and larvae require saltwater so spawning occurs from September to February in coastal swamps, river mouths and marine embayments (Davis 1985; Allsop et al. 2006).

Towards the end of the wet season there is a large migration of adult and juvenile fish upstream from coastal areas to permanent freshwater bodies (Bishop and Forbes 1991). These upstream migrations are thought to be crucial to transferring assimilated aquatic productivity from floodplain areas to the less productive rivers and streams. For example, in Magela Creek, the upstream migration of black-striped rainbowfish Malanotaenia nigrans has been estimated to be up to one tonne wet weight per day, which is almost an order of magnitude greater than the downstream migration (Pidgeon and Boyden 1993).

Patterns in Variability

There are few data describing temporal patterns in fish abundance in the periods prior to listing of the Ramsar site. Data on abundance and size distribution of fish species in eight pools within Magela Creek (Stages I and II) during a sampling season in 1981 is presented in Woodland and Ward (1992). Spangled perch Leiopotherapon unicolour and bony bream Nematalosa erebi were the most dominant species according to biomass, while Magela hardyhead Craterocephalus marianae was numerically the most dominant species at the start of the study, prior to suffering high levels of predation by larger fish (Woodland and Ward 1992). Surveys were not systematic so do not constitute a baseline for identifying changes in communities.

As noted, many of the freshwater species within the site are migratory, therefore diversity and abundance of fish assemblages vary greatly at different times of year, even in permanent waterbodies. During major migration times, fish numbers can also vary greatly from day to day (Humphrey et al. 2005).

Based on long-term data from Magela Creek, Gardner et al. (2002) note a number of patterns in freshwater fish community dynamics as follows:

  • Annual variation in community structure showed species richness was highest in the late-dry and early-wet season, and was also greatest in shallow muddy lagoons.

  • Muddy channel lagoons (backflow billabongs) had the lowest annual variability.

  • Many fish species demonstrated a preference for particular structural habitat features (for example, substrate type) and, to a lesser extent, physicochemical parameters such as water depth.

Long-term fish monitoring studies by the Australian Government Environmental Research Institute of the Supervising Scientist (eriss) provide a basis for assessing long-term patterns in fish species richness at Mudginberri Billabong (Magela Creek – Stage II) and Sandy Billabong (Nourlangie Creek – Stage I) (for example, Woodland and Ward 1992, Gardner et al. 2002, Humphrey et al. 2005). These surveys were undertaken using systematic methods and therefore constitute a reliable baseline (that is standardised fish counts along 50 metre transects). Note that the fundamental processes that control fish communities in these billabongs are not known to have been fundamentally altered since the time of listing, therefore these data are expected to represent a reliable baseline.
Table 3 13 Mean abundance (number fish per 50 metres) of fish species from Mudginberri and Sandy billabongs for the period 1994 to 2005 (source: Humphrey et al. 2005)

The results of the long-term fish monitoring program undertaken by eriss are documented by Humphrey et al. (2005). At Mudginberri and Sandy Billabongs, the species with the highest abundances were typically fly-specked hardyhead Craterocephalus stercusmuscarum, followed by chequered rainbowfish Melanotaenia splendida inornata and banded grunter Amniataba percoides (refer Table 3-4, Humphrey et al. 2005). At Mudginberri Billabong, a total of 30 species was recorded, with mean species density of 12.9 species per 50 metre transect, and at Sandy Billabong a total of 29 species was recorded and the mean species density was 13.2 species per 50 metre transect (Humphrey et al. 2005). The mean species density values recorded by Humphrey et al. (2005) were comparable to baseline species richness values recorded by Bishop et al. (1990), despite differences in sampling methods and effort.

3.3.7C7 – Populations of Freshwater and Saltwater Crocodiles

Reasons for Selection as ‘Critical’

Saltwater and freshwater crocodiles represent critical components not only in terms of their ecological roles within the site, but also in terms of their iconic and cultural values.


Saltwater Crocodiles

The Australian distribution of saltwater crocodile Crocodylus porosus extends across the north from Broome and down the east coast to Gladstone. They inhabit both salt and fresh waters, including tidal rivers, estuaries, nearby freshwater billabongs, lagoons and wetlands. Saltwater crocodiles also bask on riverbanks and sandbars.

Saltwater crocodiles are opportunistic predators with a wide ranging diet, feeding in channel, billabong and floodplain habitats. In Kakadu National Park, the diet of juveniles largely consists of crabs and prawns, as well as fish, amphibians and small reptiles, and larger adults also consume fauna such as birds, kangaroos, wild pigs and sharks (NTG undated). The relatively high number of saltwater crocodiles in Kakadu National Park means that their predatory activities can have a significant effect on the population dynamics of their prey species. For example, in populations of waterbird species that are not typically present in high numbers, their reproductive success and consequent abundances may be controlled in part by saltwater crocodile predation on eggs, chicks and adult birds, particularly during the wet/breeding season.

Compared with other parts of northern Australia, saltwater crocodile densities are greatest in the Northern Territory, with Kakadu National Park containing a significant proportion of the Australian population (S. Ward pers. comm. 2009). Fukuda et al. (2007) suggest that a number of environmental influences are linked to the greater abundances in this area, namely:

  • the proportion of a catchment area that consists of favourable wetland vegetation types (Melaleuca, grass and sedge)

  • rainfall seasonality, and

  • temperature (mean temperature in the coolest quarter of the year).

Saltwater crocodiles nest over the wet season between October and May, with an increase in temperature triggering reproductive activities (Webb 1991). The extent and timing of nesting is related to rainfall and water levels in the late dry season: years with high rainfall and cool conditions between August and November are associated with high nesting effort, while years with poor rainfall and hot conditions between August and November are associated with low nesting effort (Webb 1991).

Nest mounds are constructed out of live or dead vegetation and mud. These nests are typically located among dense aquatic grasses or on floating mats, close to a permanent water source (billabong margins, riverbanks etc.) (Grigg and Taylor 1980). The nest mounds are approximately 1.8 metres high, and a hole is excavated in the mound into which approximately 50 eggs are laid (Leach et al. 2009) and incubated for 75 to 90 days. The mounds serve a number of functions, including insulation of the eggs from temperature extremes, prevention of dehydration, prevention of predation and minimisation of flood damage to the embryos. There is a high mortality of saltwater crocodile eggs, predominantly due to flooding that may kill over half of the eggs laid each year (Webb and Manolis 1989).

Freshwater Crocodiles

Freshwater crocodiles Crocodylus johnstoni are endemic to northern Australia, occurring in Western Australia, Northern Territory and Queensland. They inhabit rivers wetlands, billabongs and creeks, remaining in permanent waters during the dry season. Basking also occurs on riverbanks.

Freshwater crocodiles are ambush predators, with fish and crustaceans (for example, crayfish and shrimp) comprising the majority of their diet, although a significant proportion is derived from other fauna such as amphibians, small reptiles, birds and insects (Cogger 2000). Approximately 40 percent of the diet of freshwater crocodiles is of terrestrial origin (for example, birds, reptiles) (Webb et al. 1983).

Freshwater crocodiles typically inhabit the floodplains during the wet season, and move to river channels late in the wet season to stay in close proximity to permanent water during the dry season. Female freshwater crocodiles dig holes in sand embankments as nests. Nesting occurs during the dry season, after the water levels fall and riverbanks are exposed. Early wet season flooding can be detrimental to nesting success as embryos will drown if eggs are inundated. The temperature at which eggs are incubated determines the sex-ratio of hatchlings (Whitehead et al. 1990).

Pattern in Variability

Saltwater Crocodiles

Since protection in 1971, the Northern Territory population increased from approximately 3000 post-hatchlings (juveniles to adults age classes) to approximately 70 000 – 75 000 by 1994 (NTG undated) and now represents a large proportion of Australia’s saltwater crocodile population.

The most recent information on saltwater crocodile trends in the Kakadu National Park Ramsar site is provided by Britton (2009), who analysed population trends in the four major tidal rivers of Kakadu National Park from 1977 to 2007. For the purposes of the present study, the tidal reaches of Wildman, South Alligator and West Alligator Rivers are considered to occur in Stage II (1989 listing) whereas the tidal reaches of East Alligator are considered to occur in Stage I (1980 listing).

Table 3-5 shows saltwater crocodile densities prior to Ramsar site declaration. Since this time, there has been a general increase in saltwater crocodile densities (refer Figure 3 -26). However, this trend was not consistent across all four river systems (Figure 3 -27). Saltwater crocodile densities (number counted per kilometre) were generally greatest in the Wildman, East and South Alligator Rivers (Figure 3-6). Densities in the West Alligator Rivers were lower than the other rivers, perhaps due to issues with survey effort and other biases (Britton 2009). There is also a trend of increasing numbers of saltwater crocodiles recorded in freshwater areas, possibly in response to increasing population densities in other more optimal habitats.

Table 3 14 Saltwater crocodile densities prior to site listing in 1980 (Stage I) or 1989 (Stage II) (source: Britton 2009)



No. per km (Min. and Max.)

East Alligator (Stage I)


2.3 – 2.8

West Alligator (Stage II)


1.4 – 3.8

South Alligator (Stage II)


1.1 – 2.7

Wildman (Stage II)


2 – 6.2

Another trend noted by Britton (2009) is that the site’s saltwater crocodile population appears to be gradually shifting towards a greater proportion of larger crocodiles (greater than 1.8 metres in length), accompanied by a decline or stabilisation in the proportions of smaller size classes. Britton (2009) suggests this is due the increased densities of large crocodiles, which are known to prey on small crocodiles and drive them out of territorial areas.

Figure 3 26 Changes in non-hatchling densities of saltwater crocodiles for all four major rivers surveyed in Kakadu National Park – combined data (source: Britton 2009)

Figure 3 27 Changes in non-hatchling densities of saltwater crocodiles for Wildman River (WLDM), West Alligator River (WAR), South Alligator River (SAR) and East Alligator River (EAR) (source: Britton 2009)

Freshwater Crocodiles

Kakadu National Park has undertaken annual surveys of freshwater crocodiles within the Park since 1994. A summary of the data available from these surveys is presented in Figure 3-7 from 1994 to 2007. The data presented here include data collected from six key freshwater crocodile locations: Twin Falls (Stage I), Maguk (Stage I), Koolpin (Stage I), Jim Jim (Stage I), East Alligator (Stage I and II) and Coirwong Gorge (Stage III). With the exception of Coirwong Gorge which occurs in Stage III (1996), all data were collected following Ramsar site listing.

The greatest numbers of freshwater crocodiles were most consistently recorded at the Jim Jim and Twin Falls areas and Coirwong Gorge. In general, freshwater crocodile sightings during the surveys increased between 1994 and 2007. However, the most recent surveys conducted in 2006 and 2007 show a marked decline in densities, returning to total sighting numbers not recorded since 1998.

In the absence of systematic survey data, it is difficult to assess the likely causal factors for the apparent changes in abundance over time. Given the long life-span and low reproductive success of this species, it is unlikely that the dramatic changes in numbers from one year to the next represent actual changes in population densities. It is far more likely that such dramatic inter-annual changes reflect survey error due to inconsistent levels of sampling effort. For example, the available data suggest that not all locations were surveyed in every year, with some years containing data for only one or two locations (Figure 3-7).

However, it is notable however that the observed decline in sightings is coincident with the timing of cane toads arriving at the site. Letnic et al. (2008) recorded mass mortality of freshwater crocodiles in the Victoria River (Northern Territory), with population densities of crocodiles plummeting by as much as 77 percent following arrival of cane toads. The lack of information on the population status of freshwater crocodiles in the Park and the impacts of cane toads on local populations represent key information gaps.

Note: available data does not cover all locations for every year
Figure 3 28 Total freshwater crocodile sightings over time at each location (source: Kakadu National Park, unpublished data)

3.3.8C8 – Populations of Threatened Sharks

Reasons for Selection as ‘Critical’

Kakadu National Park supports two threatened shark species: speartooth shark Glyphis glyphis (formerly Glyphis sp. A) and northern river shark Glyphis garricki sp. nov. Maintenance of populations of threatened species is an important factor contributing to the maintenance of global biodiversity values (see Critical Service 1).


Speartooth shark

Speartooth shark Glyphis glyphis (formerly Glyphis sp. A) is listed as critically endangered under the EPBC Act and endangered under the IUCN Red List. This species has a restricted and highly fragmented population (refer Section 2.5.11). It has only been captured in tidal rivers and estuaries indicating that large tropical river systems appear to be the primary habitat for this shark, although it has been suggested that this species may move offshore to feed (Stevens et al. 2005).

Speartooth sharks hunt close to and among the soft substrate, and feed on fish and large crustaceans. The large number of sensory ampullae and the small eye of the speartooth shark indicate that it may have adapted to feeding on benthic and demersal species in turbid waters (Peverell et al. 2006).

There are no data describing natural variability in abundance of this species. In the context of this service, it would appear that the most notable life-history function provided by the site is a feeding area for neonate, juvenile and adult sharks. Note that this species has not been captured outside rivers and estuaries which may suggest that it complete its life-cycle in rivers. Based on tracking studies in the Wenlock River, neonates and juveniles of this species are thought to have highly restricted home ranges that are confined to the upper Wenlock estuary (less than 10 square kilometres stream reach), whereas adults may have wider home ranges (Pillans et al. 2008).

Due to the lack of data on specific habitat requirements of this species, it is not possible at this stage to determine critical components and processes supporting this species. It is likely however that the following are important to maintenance of this species (and the northern river shark) within the site:

  • presence of suitable prey, including fish and large demersal invertebrates (crabs, prawns etc.) (Critical Component 6)

  • physical (tides, fluvial) processes controlling stream morphology and habitat suitability, as well as movement patterns of sharks (Pillans et al. 2005; 2008) (Critical Process 1), and

  • tidal and fluvial processes controlling important water quality variables (salinity and turbidity) that are thought to determine habitat suitability (see Pillans et al. 2008).

Northern river shark

Northern river shark Glyphis garricki (formerly Glyphis sp. C) is listed as endangered under the EPBC Act and IUCN Red List. This species has a restricted and highly fragmented population (refer 2.5.11), and is thought to be restricted to the relatively shallow, upper freshwater to brackish (0-26 ppt) reaches of the Fitzroy, Adelaide, Mary and Alligator (East, West, South) River systems (TSSC 2001a; Morgan et al. 2004; Field et al. 2008). Despite considerable fishing and collecting activity in the Northern Territory, no specimens have ever been found in coastal marine habitats (Thorburn et al. 2003, Larson et al. 2004).

Very little is known of its life-history or ecology. It is likely that the key service offered by the site for this species is a feeding area for juveniles and adults. It is unknown whether the site supports mating or breeding habitat for this species. Refer to speartooth shark for possible controls on abundance.

Patterns in Variability

Insufficient data to assess pre-listing or present day distribution and abundance patterns of either species.

3.3.9C9 - Yellow Chat Populations

Reasons for Selection as ‘Critical’

The site supports one wetland-dependent threatened bird species: yellow chat (Alligator Rivers) Epthianura crocea tunneyi. Maintenance of populations of threatened species is an important factor contributing to the maintenance of global biodiversity values (see Critical Service 1).


Yellow chat (Alligator Rivers) is listed as endangered under both the EPBC Act and Territory Parks and Wildlife Act 2000. Yellow chat (Alligator Rivers) is endemic to the Northern Territory and is restricted to a small geographic area encompassing the floodplains from the Adelaide River to the East Alligator River, between Oenpelli and Darwin (Armstrong 2004, Woinarski and Armstrong 2006, TSSC 2006).

Yellow chat is typically associated with low vegetation (for example, saltmarsh, samphire, chenopod shrublands or grasslands) bordering wetlands (especially ephemeral wetlands on floodplains) (Higgins et al. 2001). Within the Northern Territory, the Alligator Rivers subspecies is only known from a small number of sites. It is mainly found on seasonally-inundated alluvial floodplains where low lying areas support cover of grasses, herbs and sedges, but is also known from vegetated margins of channels, including mangrove stands (Keast 1958; Armstrong 2004). Within the Ramsar site, the majority of recorded sightings are derived from floodplain wetlands associated with the South Alligator River (and north of the Arnhem Highway) (see Figure 3 -29).

The subspecies is thought to be relatively sedentary (Keast 1958), though known to undertake local movements where they concentrate around wetter areas of floodplain habitat at the end of the dry season (Armstrong 2004). Yellow chat (Alligator Rivers) is mainly insectivorous and typically forages on the ground in dense grass or in low shrubs (TSSC 2006, Woinarski et al. 2007).

The subspecies has been suspected to occur in a single contiguous population (Garnett and Crowley 2000) though may actually comprise multiple subpopulations (Woinarski and Armstrong 2006). The extent of occupancy is estimated to be less than 500 square kilometres, based on the extent of the floodplain habitats that yellow chat (Alligator Rivers) has been recorded (Armstrong 2004, TSSC 2006).

There is no accurate information on the total population size for this species, though Garnett and Crowley (2000) conditionally estimated that the number of breeding birds was approximately 500 individuals. The most recent targeted survey undertaken in Kakadu National Park in 2004, estimated that the Kakadu National Park population is probably fewer than 300 individuals (Armstrong 2004). Results from the 2004 survey were regarded as largely comparable to earlier survey records within Kakadu National Park, and providing no evidence of recent decline in numbers within the site (TSSC 2006).

Patterns in Variability

There are no data describing patterns in variability of this species.

I:\B17399_I_GML Kakadu Ramsar GWF\DRG\ ECO_007_091124_ Yellow_Chat.wor
Figure 3 29 Yellow chat records for Kakadu National Park (source: Parks Australia unpublished and Armstrong 2004)

3.3.10C10 - Pig-nosed Turtle Populations

Reasons for Selection as ‘Critical’

The site supports the threatened (IUCN-Vulnerable) pig-nosed turtle Carettochelys insculpta. Maintenance of populations of threatened species is an important factor contributing to the maintenance of global biodiversity values (see Critical Service 1).


Pig-nosed turtle Carettochelys insculpta is listed under as vulnerable under the IUCN Red List, and as a near threatened species under the Territory Parks and Wildlife Conservation Act 2000. Pig-nosed turtles have been recorded in South Alligator River (Schodde et al. 1972, Legler, 1980, 1982, Press 1986) and East Alligator River (Georges et al. 1989).

Pig-nosed turtles are a freshwater species, favouring still waters with an approximate depth of two metres (Legler 1980, 1982, Georges and Kennett 1989). Billabongs along the Alligator River systems are known to represent a significant refuge for this species (Press et al. 1995a). Cover for pig-nosed turtles within billabongs is provided by characteristics such as fallen branches, exposed roots and undercut banks.

Pig-nosed turtles are omnivorous, with a diet including leaves, flowers, fruit, invertebrates and fish (Schodde et al. 1972). This diversity of food sources enables opportunism, allowing varying exploitation of resources dependent on availability. While males are almost entirely aquatic, females leave the water to nest on sandy banks and lay eggs during the dry season.

The following are important to maintenance of these species within the site:

  • presence of suitable habitat in terms of vegetation communities (Components 3 and 4), fluvial hydrology (Process 1) and water quality,

  • presence of food resources (Component 6), and

  • biological processes including breeding and migration.

Patterns in Variability

Georges and Kennett (1989) found pig-nosed turtles to be widespread between the tidal reaches and the head-waters of the South Alligator River, and that high densities were present in the upper reaches during the dry season (33.8 ± 11.3 turtles per hectare, or 67 turtles per kilometre of channel, in small discrete ponds on the main channel).

There are few quantitative data describing temporal trends in the number of pig-nosed turtles within the site. It is believed that feral animals and other stock caused a decline in the South Alligator River pig-nosed turtle population prior to the declaration of Stage III of Kakadu National Park (1987-1991) (A. Carr pers. comm. in Pritchard 1979), but this decline was not quantified, and it is not clear whether the population has subsequently recovered (TSSC 2005). However, it is known that declines in yellow-spotted monitor lizard Varanus panoptes population numbers associated with the arrival of toxic cane toads Rhinella marina have reduced predation on pig-nosed turtle eggs by monitor lizards (Doody et al. 2006).

Figure 3 30 Map of know dry season distribution of pig-nosed turtle in the Ramsar site (source: Georges and Kennett 1989)

3.3.11C11 – Locally Endemic Invertebrate Species

Reasons for Selection as ‘Critical’

Support of locally endemic fauna species was selected as a critical component on the basis of support for Ramsar Nomination Criterion 3, as well as the importance of endemic species in terms of justification for National Park nomination (refer Fox et al. 1977) and the fundamental importance of threatened species to determining the site’s ecological character.


The following aquatic invertebrate taxa are considered to be locally endemic species (occur exclusively within the catchments of the Ramsar site and the Arnhem Plateau – see Appendix D):

  • An endemic family of shrimps (Kakaducarididae), which contains two mono-specific genera, namely Leptopalaemon gagadjui and Kakaducaris glabra. Leptopalaemon gagadjui typically occurs in upland permanent streams in the north-western area of the Arnhem Land plateau and is widely distributed (South Alligator River, Nourlangie and Magela Creeks, Namarrgon Gorge), while K. glabra is restricted to a single location (the type location in Lightning Dreaming Creek, Namarrgon Gorge) (Bruce 1993, Page et al. 2008, refer Appendix D).

  • A genus of isopod (Eophreatoicus; Family Phreatoicidea) that reportedly has exceptional species-level diversity (approximately 30 species lineages, Wilson et al. 2009). Specimens have been collected from the King River region of western Arnhem Land, and various aquatic habitats associated with the west Arnhem Land plateau and escarpments, including sites in the East Alligator (for example, Magela, Ngarradji and Catfish Creeks), South Alligator, Katherine and Liverpool River catchments (Wilson et al. 2009, refer Appendix D). The species within this genus are extremely narrow range endemics, with juveniles migrating very small distances downstream (approximately two to six kilometres) from their headwater refuges (Wilson et al. 2009).

  • At least one species of mayfly from the family Leptophlebiidae (see section 2.5.3).

These species occur exclusively in upland areas, specifically the ancient stone country. Humphrey (1999) identified two key controls on endemism in the stone country:

  • the antiquity and persistence of the escarpment and associated perennial streams, springs and seeps, and

  • isolating mechanisms, including processes leading to fragmentation of habitat operating over geological time-scales (climate change, erosion etc.), and the generally poor dispersal characteristics of the crustaceans.

Note that regional endemic species (including fish, aquatic plants and invertebrates) are considered as supporting components in section 3.4.6.

Patterns in Variability

There are few data describing patterns in variability of these endemic aquatic invertebrate species. This is because it is only relatively recently that many of the endemic invertebrates have been observed, with most remaining undescribed (for example, Eophreatoicus; Wilson et al. 2009).

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