1 literature review - Victoria University · chapter 1: literature review 5 1 literature review 1.1 Background to the study A primary driving force behind this project is the recognition

chapter 1: literature review 5

1 literature review

1.1 Backgroundtothestudy

A primary driving force behind this project is the recognition that Australian coastal saltmarsh has suffered from long-term neglect by researchers, natural resource managers and the wider community. The two main plant communities on the seaward side of coastal saltmarsh – seagrasses and mangroves – have been subject to intensive study over the past three decades, and there is now an extensive literature on the ecology and management of these wetland types in Australia (e.g. MacNae 1966; Lear & Turner 1977; Clough 1982; Hutchings & Recher 1982; Hutchings & Saenger 1987; Larkum et al. 1989; Robertson & Alongi 1992; Claridge & Burnett 1993; Harty 1997; Pollard et al. 2003; chapters in Green & Short 2003; Duke 2006; Carruthers et al. 2007; Alongi 2009; Paling et al. 2009).

Australian saltmarshes have not benefited from a comparable investment in research or management. Indeed, until the publication this year of Australian Saltmarsh Ecology (edited by Saintilan 2009), some chapters in Coastal Wetlands: An Integrated Ecosystem Approach (edited by Perillo et al. 2009) and one chapter in Human Impacts on Salt Marshes (edited by Silliman et al. 2009), the most recent text with substantive sections on Australian coastal saltmarsh is 20 years old: Adam’s (1990) Saltmarsh Ecology. The presence of separate entries for mangroves (Bridgewater 1999) and seagrasses (Walker 1999) – but not for saltmarsh – in the introductory volume of Flora of Australia (Orchard 1999) may say something about the overall neglect of coastal saltmarsh by even experienced ecologists and those charged with managing coastal vegetation.

From at least two perspectives the neglect of coastal saltmarsh is paradoxical. First, coastal saltmarsh in the Northern Hemisphere has been the subject of intensive research for decades (e.g. Teal & Teal 1969; Ranwell 1972; Chapman 1974; Long & Mason 1983; Kennish 1990; French 1997; Weinstein & Kreeger 2000; Doody 2008; Perillo et al. 2009; Silliman et al. 2009; Weis & Butler 2009). Second, a number of early botanical studies in Australia were undertaken on mangroves and coastal saltmarsh: Hamilton (1919) and Collins (1921), for example, studied saltmarsh and mangrove vegetation in the Sydney region; Pidgeon (1940) reported on spatial zonation (which she interpreted as successional patterns) in mangroves and saltmarshes along the central coast of New South Wales; Patton (1942) reported on Victorian saltmarshes; and Curtis & Somerville (1947) described coastal saltmarshes in Tasmania.

This impressive early start was not maintained in subsequent decades. The most detailed studies of any Australian saltmarsh were undertaken 40 years ago, by Clarke & Hannon (1967, 1969, 1970, 1971). The most detailed studies of Victorian saltmarshes – Bridgewater (1975, 1982) – are nearly as old. Rather than embracing the more useful ecological orientation taken by Clarke & Hannon, they were concerned mostly with phytosociological classification and, for many potential users, have produced not-entirely-satisfying results. The most detailed identification guide to Victoria coastal saltmarshes is the booklet by Bridgewater et al. (1981), and it is not only dated but suffers from restricted availability. In fact, many of the studies undertaken on Victorian mangroves and coastal saltmarsh are available only in the ‘grey’ literature, as government or consultants’ reports or as unpublished student theses. In this aspect the literature on mangroves and coastal saltmarsh differs little from that on other types of Australian wetlands, which also suffer from poor availability and a lack of peer review (Boon & Brock 1994). An illustration of the importance of unpublished reports is provided by the review by Ross (2000) of the mangroves and coastal saltmarsh of Western Port. Of the 88 reports cited, about one half were in the ‘grey’ literature.

mangroves and coastal saltmarsh of victoria: distribution, condition, threats and management6

Nevertheless, there are encouraging signs that the neglect of Australian coastal saltmarsh and other estuarine wetlands is coming to an end. The past few years have seen the publication of handbooks that allow the identification of saltmarsh plants in Queensland ( Johns 2006) and western Victoria (Allen 2007). A new identification guide to saltmarsh vegetation in south-eastern Australia, edited by Geoff Sainty et al. and published by Sainty & Associates, is expected to be released in 2011. A good start has been made into resolving the taxonomic issues associated with problematic plant species (e.g. Shepherd et al. 2005; Shepherd & Wilson 2007; Adams et al. 2008; Wilson 2008). Duke (2006) has recently produced a delightful field manual on mangroves (Australia’s Mangroves), which expands on the earlier field guide by Lovelock (1993). Australian Saltmarsh Ecology has just been published (edited by Saintilan 2009), and Laegdsgaard (2006) has published her review of the status and threats to coastal saltmarsh, mostly from the perspective of New South Wales systems. A wider perspective on anthropogenic threats to coastal saltmarsh in Australia and New Zealand was prepared by Thomsen et al. (2009). The biogeography of Australian saltmarsh vegetation has been re-examined by Saintilan (2009a,b). Two of the recent books on estuarine and coastal vegetation (Coastal Wetlands: An Integrated Ecosystem Approach: Perillo et al. 2009; and Human Impacts on Salt Marshes: Silliman et al. 2009) contain a number of chapters written by Australian authors or covering Australian systems (e.g. Adam 2009b; Alongi 2009; Paling et al. 2009; Thomsen et al. 2009).

Comprehensive statewide mapping and inventory studies have been completed recently, or are nearing completion, for coastal and estuarine systems in Queensland (Queensland EPA 2005, 2009c), South Australia (Canty et al. 2006) and New South Wales (Williams et al. 2006). A suite of regional studies has addressed aspects of particular mangrove or saltmarsh systems in south-eastern Australia (e.g. Clarke 1993, 2003; Kessler 2006; Kelleway et al. 2007; Jugovic 1985, 2008; Pacific Wetlands 2008; Horlock & Houtgraaf 2008). Condition-assessment protocols have been developed for aspects of Victorian estuaries (Arundel et al. 2009) and, in at least one case, a method specifically tailored to assess saltmarsh condition (Kessler 2006).

The first component of our project – the preparation of a comprehensive literature review – aims to bring together these disparate streams of information and, where appropriate, place them within the wider context of coastal wetlands in south-eastern Australia. In addition to the reference sources provided in monographs or edited books (e.g. Ranwell 1972; Long & Mason 1983; Adam 1990; Weinstein & Kreeger 2000; Doody 2008; Saintilan 2009; Perillo et al. 2009; Silliman et al. 2009; Weis & Butler 2009), we surveyed the available literature in recent scientific journals: over 300 articles published after 2000 were retrieved using the keywords ‘salt+marsh’ in Web of Science. We examined also the university archives for student theses, which were particularly helpful for studies on *Spartina in Victoria.

To complement these formal sources of information, the project was discussed with many saltmarsh researchers and those charged with managing coastal areas: Paul Adam (University of New South Wales, Sydney), Geoff Wescott (Deakin University, Melbourne), Janine McBurnie (Deakin University, Melbourne), Rae Moran (Department of Sustainability and Environment, Melbourne), Mike Ronan (Queensland Environment Protection Authority, Brisbane), Neil Saintilan (New South Wales Department of Environment and Climate Change, Sydney), Jeff Shimeta (RMIT University, Melbourne), Mike Vanderzee (Victoria Department of Sustainability and Environment, Melbourne), Rob Williams (Aquatic Ecosystems Research Unit, Cronulla Fisheries Centre, Sydney), Debra Canty (South Australian Department of Environment and Heritage, Adelaide), David Carew (Melbourne Water, Melbourne), Greg Hunt (Western Port Greenhouse Alliance) and Steffan Howe and Leslie Leunig (Parks Victoria, Melbourne and Foster, respectively). Glenn Ehmke (Birds Australia) kindly contributed text and images for the habitat-value of coastal saltmarsh for birds, especially for Orange-bellied Parrot Neophema chrysogaster.


1.2 Whataremangrovesandcoastalsaltmarsh?

mangroves – a definition

Woodroffe & Davies (2009, page 66) defined mangroves as ‘trees, shrubs, or palms, exceeding 0.5 m in height that occur in the upper intertidal zone’. Duke (2006, page 12) adopted a similar definition with regard vegetation structure but was more prescriptive about tidal position: he indicated that mangroves normally grew ‘…above mean sea level in the intertidal zone of marine coastal environments and estuarine margins’.

In Victoria, vegetation is classified and described in terms of Ecological Vegetation Classes, or EVCs. Ecological Vegetation Classes are defined as one or a number of floristic and structural types that appear to be associated with a recognisable environmental niche and which can be characterised by their adaptive responses to ecological processes that operate at the landscape scale. The EVC approach to vegetation classification therefore uses floristic and structural criteria, combined with geographic information on niches and distributions (Department of Natural Resources and Environment 2002b).

Only one mangrove species, Avicennia marina subsp. australasica is present in Victoria (Duke 2006) and the plant community it forms is classified under the Victorian typology as Ecological Vegetation Class (EVC) 140 Mangrove Shrubland (Department of Sustainability and Environment 2009e). Avicennia marina is globally the most widespread of all mangrove species and the only one that occurs south of Merimbula (36o50’S) on the Australian east coast. Three subspecies of Avicennia marina are recognised in Australia: Avicennia marina subsp. australasica along the south-eastern coast; Avicennia marina subsp. eucalyptifolia along the tropical northern coast; and Avicennia marina subsp. marina along the Western Australian and South Australian coasts (Everett 1993; Duke 2006; Barker & Orchard 2008). Avicennia marina subsp. australasica is commonly known as Grey Mangrove or White Mangrove.

coastal saltmarsh – differentiation from other estuarine wetlands

Coastal saltmarsh presents a more formidable problem with definition and spatial delimitation. Although almost everyone finds it easy to recognise a saltmarsh when they see it, it is surprisingly difficult to devise a definition that unequivocally delimits saltmarshes from the habitats on their seaward and landward sides and, even more problematically, from the other types of wetland that occur in estuaries. (Tapliapietra et al. 2009 recently completed a useful review of meanings of the term ‘estuaries’ and associated environments.) The difficulty is evident from Table 1.1, which shows the range of meanings devised by various authors to define and delimit coastal saltmarsh; some have used features of the vegetation; others, characteristics of the soils; some have used hydrological criteria; and others, various geomorphological attributes. Despite the differences, there is general agreement that coastal saltmarsh is a wetland type that occurs in the intertidal zone and supports the growth of halophytic plants.


Table1.1: Definitions of coastal saltmarsh provided by various authors.

Definition Author

‘…areas of land bordering on the sea, more or less covered with vegetation and subject to periodic inundation by the tide.’

Chapman (1960, cited in Dalby 1987, p. 38)

Coastal saltmarshes are characterised by three attributes: i) consist of alluvial sediments on the shore by the sea; ii) are subjected to tidal or weather-affected inundation by more or less diluted seawater, which results in soils that are continuously waterlogged or at least temporarily moist; and iii) have soils are of variable salinity and consequently support plant communities that are completely or mainly consisting of halophytic species.

Beeftink (1977, p. 109)

‘…areas of alluvial or peat deposits, colonized by herbaceous and small shrubby terrestrial vascular plants, almost permanently wet and frequently inundated with saline waters.’

Long & Mason (1983, p. 1)

‘…intertidal ecosystems developed in sheltered situations where silt and mud can accumulate.’ Knox (1986, p. 37)

‘…beds of intertidal rooted vegetation which are alternately inundated and drained by the tides’ Day et al. (1989, p. 189)

‘…areas, vegetated by herbs, grasses or low shrubs, bordering saline waterbodies…subjected to periodic flooding as a result of fluctuations (tidal or non-tidal) in the level of the adjacent water body.’

Adam (1990, p. 1)

‘The terms saltmarsh and mangrove describe communities dominated by vascular plants occurring in the intertidal zone. Mangrove communities are characterised by trees and shrubs, and saltmarshes by herbs and dwarf shrubs but an absolute distinction between the two vegetation types in terms of structure cannot be maintained.’

King et al. (1990, p. 226)

‘…open-shrubland to closed-herbfield community of salt-tolerant, often succulent species…on intertidal mudflats…’

Conn (1993, p. 128)

‘…intertidal ecosystems backed up against the land on one side while opening to the estuary and the sea on the other.’

Lefeuvre & Dame (1994, p. 170)

‘…areas of intertidal soft sediments occupied by grasses, reeds or sedges and small shrubs.’ Morrisey (1995, p. 205)

‘…communities of emergent herbs, grasses, or low shrubs rooted in soils alternately inundated and drained by tidal action.’

Nybakken (2001, p. 348)

‘High inter-tidal to supra-tidal salt-tolerant vegetation such as grasses, sedges, reeds and small shrubs, that occur in muddy sediment.’

Ryan et al. (2003, p. 7)

‘…complex mosaics of closed sedgelands, grasslands and open herbfields, and occasionally have emergent shrubs. They are restricted to estuarine mudflats that are exposed to intermittent tidal inundation and to small soaks on exposed headlands that receive abundant salt spray from onshore winds.’

Keith (2004, p. 238)

‘Saltmarshes tend to occupy the hyper-saline soils of the upper inter-tidal zone, where saltwater inundation occurs less frequently (usually only during high spring tides). These communities are generally found growing on the landward side of mangroves, and are made up of salt tolerant, flowering plants in the form of low growing shrubs, herbs and grasses.’

Goudkamp & Chin (2006, p. 3)

‘Saltmarshes occur in low-energy environments that allow for the accumulation of fine sediments…Saltmarsh vegetation consists of a limited number of halophytic (salt tolerant) species adapted to regular immersion by the tides.’

Best et al. (2007, p. 205)

‘…a community of plants and animals that grow along the upper-intertidal zone (above the mean spring-tide height) of coastal waterways…mainly in temperate regions.’

OzCoasts (2008, p. 1)

‘…intertidal communities dominated by flowering plants, principally herbs and low shrubs.’ Adam (2009a, p. 1)

‘…a mostly treeless plant community recognised by a low mosaic of succulent herbs, salt tolerant grasses and sedges, found in the tidal flats of estuaries and on the edges of intermittently opened coastal lagoons. They are characterised by vegetation interspersed with unvegetated patches or salt pans.’

Department of Environment and Climate Change (2009, p. 1)

‘Saltwater wetland occupied mainly by herbs and dwarf shrubs, characteristically able to tolerate extreme environmental conditions, notably waterlogging and salinity.’

Queensland EPA (2009b, p. 14)

‘…coastal wetlands that are transitional zones between the aquatic and terrestrial worlds.’ ‘Salt and brackish marshes are also called tidal marshes because they occur in the intertidal zone between low and high tides.’

Weis & Butler (2009, pp. 3 & 4)


In the typology used to classify vegetation in Victoria, coastal saltmarsh falls into EVC 9 Coastal Saltmarsh Aggregate (Department of Sustainability and Environment 2009e). The unit is termed an Aggregate because it is not ecologically homogeneous and, with future work, could profitably be more finely resolved into a number of smaller units. As it currently stands, EVC 9 describes coastal saltmarsh as a low shrubby (to herbaceous or grassy) type of vegetation that occurs on salinised coastal soils that are tidally influenced, and includes shrubby dicots such as Tecticornia (previously Sclerostegia) arbuscula, grasses such as Austrostipa stipoides, and dicot herbs such as Sarcocornia quinqueflora. Chapter 4 of the report outlines a new typology for Victorian coastal saltmarsh which resolves the severe limitations of the current classification.

The issue that confronts all studies that try to map or make an inventory of coastal saltmarsh is how to differentiate this particular type of wetland from the diverse range of other plant communities that occur in estuaries and along the coast in other intertidal locations. The recent publication on wetland EVCs in Victoria (Department of Sustainability and Environment 2009e) indicates the large number of different wetland vegetation classes that can occur in the state, and Table 1.2 shows the variety of estuarine wetlands that may be found near to, or in association with, EVC 9 Coastal Saltmarsh Aggregate in Victoria. Coastal saltmarsh therefore needs to be differentiated from the range of other wetland types that occur as discrete units or sometimes intermingled with, saltmarsh in the intertidal zone; moreover, it is critical for mapping and inventory studies, as well as for legislative and regulatory needs, to be able to set a clear landward boundary (Adam 1992; Adam et al. 1985; Winning 1991).

Table1.2: Range of estuarine wetlands and fringing communities, classified by Ecological Vegetation Classes (EVCs), that may be found in association with coastal saltmarsh in Victoria.

EVC EVC name Characterisation Indicator species

10 Estuarine Wetland Rushland/sedgeland vegetation, variously with component of small halophytic herbs, occurring in coastal areas where freshwater flows augment otherwise saline environments.

Juncus kraussii, occasionally with Phragmites australis or species of Cyperaceae.

13 Brackish Sedgeland Sedgeland dominated by salt-tolerant sedges in association with a low grassy/herbaceous ground-layer with a halophytic component.

Gahnia trifida (sometimes Gahnia filum), Baumea juncea, with a mixture of species as for Brackish Herbland and species which are not obligate halophytes.

14 Estuarine Flats Grassland

Tussock grassland or grassy sedgeland beyond zone of normal tidal inundation but sometimes subject to seasonal waterlogging or rarely brief intermittent inundation.

Poa poiformis with Ficinia nodosa, and including non-halophytic species such as Senecio spp., Clematis microphylla and Acaena novae-zelandiae.

53 Swamp Scrub Dense (and potentially tall shrubby vegetation of swampy flats), dominated by Myrtaceous shrubs (to small trees), ground-layer often sparse, aquatic species conspicuous, sphagnum and/or waterlogging tolerant ferns sometimes present.

Melaleuca ericifolia, Leptospermum lanigerum, with aquatic / semi-aquatic spp. (e.g. Isolepis inundata, Triglochin procera s.l., Villarsia spp., Sphagnum spp.).

842 Saline Aquatic Meadow

Submerged ephemeral or perennial herbland of slender monocots, occurring in brackish to saline water bodies subject or not to dry periods. The vegetation is characteristically extremely species-poor, consisting of one or more species of Lepilaena and/or Ruppia.

Variously Ruppia megacarpa, Ruppia polycarpa, Lepilaena spp. (e.g. L. preissii, L. bilocularis, L. cylindrocarpa).


EVC EVC name Characterisation Indicator species

140 Mangrove Shrubland Extremely species-poor shrubland vegetation of intertidal zone, dominated by mangroves.

Characteristically occurs as mono-specific stands of Avicennia marina. In some stands, species from adjacent Coastal Saltmarsh or Seagrass Meadow also present.

196 Seasonally Inundated Sub-saline Herbland

Very species-poor low herbland of seasonal saline wetland within relicts of former tidal lagoons, dominated by Wilsonia spp.

Wilsonia humilis sometimes with W. backhousei and/or W. rotundifolia.

538 Brackish Herbland Low herbland dominated by species tolerant of mildly saline conditions and rare intermittent inundation.

Lobelia irrigua, Sebaea spp., Ranunculus spp., Apium annuum, Lachnagrostis spp., Isolepis cernua, Schoenus nitens, Wilsonia rotundifolia; variously Selliera radicans, Distichlis distichophylla and/or Samolus repens.

656 Brackish Wetland Collective label for the various zones of sedgy-herbaceous vegetation associated with sub-saline wetlands. Components variously include wetter versions of Brackish Sedgeland, Brackish Herbland and Saline Aquatic Meadow.

Bolboschoenus caldwellii and/or Schoenoplectus pungens and aquatic semi-aquatic species tolerant of at least moderate salinity.

821 Tall Marsh Wetland dominated by tall emergent reeds, rushes or sedges, typically in dense, species-poor swards.

Typically Phragmites australis, Typha spp., Schoenoplectus tabernaemontani. Associated species are quite variable and can include Calystegia sepium and Urtica incisa and a range of aquatics.

845 Sea-grass Meadow Sward-forming aquatic herbland of sheltered marine shallows, intertidal flats and lower estuarine habitats.

Dominated by Zostera and/or Heterozostera spp. (or localised variant also including Lepilaena marina and Ruppia tuberosa).

934 Brackish Grassland Grassland on sub-saline heavy soils, including dominants of Plains Grassland (and a portion of associated herbaceous species) in association with herbaceous species indicative of saline soils.

Poa labillardierei/Themeda triandra, Austrodanthonia spp., Distichlis distichophylla, Calocephalus lacteus, Selliera radicans, Sebaea spp., Wilsonia rotundifolia, Lobelia irrigua; Poa poiformis in some coastal sites.

947 Brackish Lignum Swamp

Wetland dominated by Muehlenbeckia florulenta with a component or patches of salt-tolerant herbs (at least at low to moderate levels of salinity) and usually also with some species common to freshwater habitats. Can be very species-poor.

Muehlenbeckia florulenta, variously with Samolus repens, Isolepis cernua, Triglochin striata, Chenopodium glaucum, Myriophyllum verrucosum, Selliera radicans, Mimulus repens, Distichlis distichophylla, Lobelia irrigua, Wilsonia rotundifolia, Lachnagrostis spp. and/or Gahnia filum.

952 Estuarine Reedbed Vegetation dominated by tall reeds (usually 2–3 m or more in height), in association with a sparse ground-layer of salt tolerant herbs. Distinguished from Estuarine Wetland by the vigour and total dominance of reeds, and from Tall Marsh by the presence of halophytes.

Phragmites australis, with associated species variously including Samolus repens, Juncus kraussii, Triglochin striatum, Bolboschoenus caldwellii and Suaeda australis.

953 Estuarine Scrub Shrubland to scrub of Myrtaceous shrub species of sub-saline habitat, occurring in association with ground-layer including halophytic herbs.

Melaleuca ericifolia (in eastern Victoria), with other Melaleuca spp. (e.g. Melaleuca lanceolata, Melaleuca gibbosa) or Leptospermum lanigerum in marginal sites in western Victoria. Gound-layer includes Samolus repens, Triglochin striata and Selliera radicans, variously with Sarcocornia quinqueflora, Gahnia filum, Poa poiformis, Juncus kraussii, Disphyma crassifolium, Distichlis distichophylla.


Differentiationofsaltmarshesfromseawardplantcommunities

On their seaward side, saltmarshes are usually fringed by seagrasses and/or mangroves. Differentiation from seagrasses beds is quite unequivocal, as seagrasses are either fully submerged or experience regular inundation by neap tides and have a distinctive floristic composition, usually consisting of vascular species from five monocot plant families: Cymodoceaceae, Hydrocharitaceae, Posidoniaceae, Zannichelliaceae and Zosteraceae (Green & Short 2003).

Differentiation from mangroves, however, depends strongly upon location. Like coastal saltmarsh, mangroves are tidally flooded but are always dominated by shrubs or trees and reach their optimum development – in terms of floristic diversity and size of plants – in tropical regions (Turner et al. 2004; Saintilan et al. 2009b). Saltmarshes are usually dominated by plants of shorter stature, especially by grasses and dicot herbs, although low shrubs are not uncommon in coastal saltmarsh of southern Australia (e.g. Tecticornia arbuscula, Tecticornia halocnemoides and Tecticornia pergranulata, as well as several Atriplex species). Duke (2006) argued that saltmarshes were generally distinguishable from mangroves in northern Australia on the basis of the smaller size of the plants and their forming a shrubby ground layer that was usually < 0.5 m high. A further simple distinction, particularly useful in southern Australia, is that mangroves are usually broad-leaved trees, while woody plants in coastal saltmarsh are generally strongly succulent. (Note that many species of mangrove in northern Australia have succulent (cf. sclerophyllous) leaves: Duke 2006.)

The distinction between saltmarshes and mangroves is usually quite clear along the wetter subtropical and warm temperate coast of northern New South Wales and Queensland, where large mangrove trees form a dense woodland or forest on the seaward side and saltmarshes are usually poorly developed. It is less striking along the semi-arid tropical coastlines of Queensland, where stunted mangroves merge into fringing saltmarsh with much the same structure. In fact, Jacobs (1999) postulated that saltmarshes in the tropics could be harder to delimit than saltmarshes in the temperate zone.

The distinction between mangroves and saltmarshes can be unclear also along some parts of the Victorian coast, where Avicennia marina may form a stunted shrubland, sometimes with an equal or lower stature than the Tecticornia arbuscula shrubs of the adjacent saltmarshes and into which mangroves trees often recruit. In such cases, the distinction between mangroves and saltmarshes is subtle and based on floristics and convention (Adam 1990). In other parts of the Victorian coast, however, structural differences between mangroves and saltmarshes make differentiation simple (Figure 1.1).


Figure1.1: Mangroves Avicennia marina on the seaward side of a band of coastal saltmarsh, Stony Creek Backwash, Melbourne. Mangroves were planted here in the 1980s and have since formed an extensive mangrove shrubland.

Mangroves can be found also as sparse individuals within what would appear to be a large expanse of saltmarsh. In these cases an arbitrary differentiation between the two wetland types is usually made on the basis of distance between individual mangrove plants (Wilton & Saintilan 2000; Saintilan & Rogers 2001). Such intimate juxtapositions of mangroves and saltmarsh are not uncommon on a global scale: several parts of the world, including the Gulf Coast of the USA, central Florida, many parts of northern, eastern and southern Australia, as well as parts of New Zealand and southern Japan, have an obvious intertidal zone in which mangroves and saltmarshes intermingle (Chapman 1977).

Differentiationfromothertypesofestuarinewetland

In Victoria, perhaps the greatest difficulties are encountered when attempting to delimit coastal saltmarsh from other sorts of wetlands that occur next to or intermingled with them in estuaries (Table 1.2). Overseas research has shown repeatedly that such broad-scale zonation of vegetation is driven strongly by a complex mosaic of gradients in salinity and waterlogging (Pratolongo et al. 2009; Weis & Butler 2009), and the same factors probably hold sway in south-eastern Australia, as shown by the pioneering work of Clarke & Hannon on saltmarshes of the Sydney region in the late 1960s.

Although coastal saltmarshes are intertidal and hence saline, it is often difficult to differentiate them in the field from the suite of other estuarine wetland types that may be periodically inundated by fresh and/or saline water over highly variable spatial and temporal scales. Even nominally freshwater wetlands, for example, may


be influenced by seawater during extremely high tides or storm surges. Saltmarshes can be differentiated from tidal freshwater marshlands because the former are inundated by tidally influenced seawater rather than by tidally forced fresh water. Tidal freshwater marshes are very common along the eastern coast of continental USA and in Alaska (Kennish 1990; Connor et al. 2007; Whigham et al. 2009); they are not uncommon along the New South Wales coast (Adam et al. 1985).

Another context where coastal saltmarsh can be difficult to distinguish from other types of estuarine wetland is where freshwater flows emerge at the coast as estuaries or from groundwater seepage. In the brackish lagoons and the upper reaches of estuaries in New South Wales, for example, coastal saltmarsh can merge into tall reedswamps dominated by Phragmites australis and Schoenoplectus, Bolboschoenus and Typha spp. (Adam 1994). In Victoria, saltmarshes are similarly often found nearby or in mosaics with a range of different estuarine wetlands that experience a mixture of tidal and freshwater influence. These wetlands are frequently vegetated by dense stands of rhizomatous perennial monocots, including Juncus kraussii and Phragmites australis in the seaward reaches, and Bolboschoenus caldwellii and Schoenoplectus pungens in areas further from the influence of seawater (Sinclair & Sutter 2008). Halophytic and glycophytic species are often intermixed to form complex patterns and mosaics that reflect the interplay of site elevation, saline and freshwater inundation, and groundwater flows. The co-existence of such diverse plant taxa creates real difficulties in differentiating between saltmarsh and other types of estuarine wetlands, with obvious implications for mapping saltmarsh and the preparation of statewide inventories.

Settingalandwardboundarytocoastalsaltmarsh

Adam (1990, 1994) argued that the upper limit of coastal saltmarsh was often difficult to define, particularly on dry coastlines where intertidal saltmarsh vegetation can merge ‘imperceptibly with inland grassland or chenopod shrublands’ (Adam 1994, page 398). Three approaches have commonly been used to set the landward boundary: i) presence of anthropogenic structures such as sea walls; ii) floristic composition, especially the presence of halophytic indicator species; and iii) hydrological criteria.

The landward limit of saltmarshes sometimes can be defined by the presence of an artificial structure, such as a sea wall, which provides an unequivocal limit to tidal influence. Even in these cases, however, remnant and degraded saltmarsh can survive on the the alienated landward side, particularly if fed by shallow saline groundwater. Because of their long history of land-claim for agricultural use, northern European saltmarshes are often bounded by sea walls (French 1997) and such structures have been often used for delimiting the landward boundary of saltmarshes in the Northern Hemisphere (e.g. Dalby 1987). Most Australian systems lack such conveniently delimiting structures, although many saltmarshes in West and South Gippsland are confined by low earthen or stone sea walls (Figure 1.2). In the absence of engineering-set boundaries, it may be difficult delimit the landward side of saltmarsh. The difficulty becomes especially relevant when attempts are made to map saltmarsh distributions (Dalby 1987; Adam 1990, 1992; Winning 1991).


Some jurisdictions make use of halophytic indicator species to set the landward boundary (Dalby 1987), and it is not uncommon for coastal-zone managers to set the landward limit in terms either of the presence of vegetation generally thought to be halophytic or, more particularly, the presence of a limited number of plant taxa that are (sometimes arbitrarily) chosen as characteristic indicators of saltmarshes. In New South Wales, for example, coastal saltmarsh is identified on the basis of five criteria (Table 1.3), one of which is the presence of plant species listed as characteristic taxa (Table 1.4). Note that some of the species identified in Table 1.4 as characteristic saltmarsh plants in New South Wales are not considered indicative of saltmarshes in Victoria (e.g. Baumea juncea, Bolboschoenus spp., Ficinia nodosa, Phragmites australis, Schoenoplectus spp., Tetragonia tetragonioides and Typha spp.).

Table1.3: Criteria used to identify coastal saltmarshes in New South Wales. Source: Department of Environment and Climate Change (2009).

Criterion Description

1 Location on a coastal floodplain

2 Close to or fringing a saline or semi-saline coastal estuary or lagoon

3 Vegetation is treeless

4 Located on a flat surface behind a mangrove stand

5 Presence of characteristic plant species (see Table 1.4 below)

Figure1.2: Sea wall, Port Albert, South Gippsland. The photograph illustrates not only the role of sea walls in alienating coastal saltmarsh from hinterland areas, but also the growth of Avicennia marina into adjacent coastal saltmarsh.


Table1.4: Characteristic plant species used to identify coastal saltmarshes in New South Wales. Source: Department of Environment and Climate Change (2009).

Scientific name Common name

Overstorey – tree or shrub layer (> 1.5 m)

Aegiceras corniculatus River Mangrove

Avicennia marina Grey Mangrove

*Baccharis halimifolia Groundsel Bush

Melaleuca ericifolia Swamp Paperbark

Tecticornia (Sclerostegia) arbuscula Shrubby Glasswort

Groundcover (< 1.5 m )

Austrostipa stipoides Prickly Spear-grass

Baumea juncea Bare Twig-sedge

Bolboschoenus spp. Marsh Club-sedge

*Cortaderia spp. Pampas Grass

§Ficinia (Isolepis) nodosa Knobby Club-rush

Gahnia filum Coast Spiny Saw-sedge

*Juncus acutus Rush

§Juncus kraussii Sea Rush

Limonium australe Sea-lavender

Phragmites australis Common Reed

§Samolus repens Creeping Brookweed

Sarcocornia quinqueflora ssp. quinqueflora Beaded Glasswort

§Selliera radicans Selliera

§Suaeda australis Austral Seablite

§Sporobolus virginicus Salt Couch

Schoenoplectus spp. Club-sedge

Tetragonia tetragonioides New Zealand Spinach

§Triglochin striata Streaked Arrowgrass

Typha spp. Cumbungi

§Zoysia macrantha Prickly Couch

* Introduced species§ ‘Key indicator species’ in this classification

There are a number of problems with using floristic criteria to define the landward limit of saltmarsh or to delimit coastal saltmarsh from other types of estuarine wetland. The presence of halophytic indicator species, for example, is difficult to apply where saltmarshes front waters of relatively low salinity, such as along the Baltic Sea in western Europe (Dalby 1987). Moreover, the term ‘halophyte’ itself is open to different meanings (cf. problems with defining what is a hydrophyte: see Tiner 1991). Halophytes can be defined on the basis of physiological or ecological criteria and, for the present purpose, an ecological definition (plants that


can complete their life cycles in a saline environment: Flowers et al. (1986) is the most applicable. Such a definition, however, raises a further problem in that the term ‘saline’ requires definition and there is no broad agreement as to what this word means. Table 1.5 shows, as an example, how water managers and limnologists differ in their understanding of the term.

Table1.5: Comparison of meanings of various salinity terms by water managers and limnologists. Seawater is 35 g L–1.

Source: Williams (1987).

Salinity (g L–1)

Fresh Slightly saline Saline Highly saline

Water managers < 0.3 0.3–1 1–3 > 3

Limnologists < 0.3 3–10 10–100 > 100

Thus the differentiation of halophytic plants from glycophytic plants is by no means unequivocal. The identification, for example, of Typha spp. as characteristic saltmarsh plants in New South Wales (Table 1.4) would cause surprise in Victoria, where they are never seen as plants typical of coastal saltmarsh. Moreover, not only can different genotypes of glycophytic species vary widely in their salt sensitivity (e.g. see Hanganu et al. 1999 and Zhao et al. 1999 for Phragmites australis), not all genotypes of a supposedly halophytic species can survive in saline environments (e.g. Adam 1990 for Festuca rubra in Europe).

The final obstacle to using the presence of halophytic taxa to set a landward boundary is that it is likely to be tautological: saltmarshes are defined in terms of the presence of specific taxa of halophytic plants, and in turn halophytes are salt-tolerant plants that, by definition, grow in saline environments such as saltmarsh. The objection can be negated if independent studies can show the taxa of interest to be salt-tolerant, but must continue to hold if it is inferred by its presence in saltmarshes that a particular species of plant must be halophytic. In fact, the indigenous and exotic flora of Victorian coastal saltmarsh is generally categorisable into two broad groups: i) species (halophytes) that are always or characteristically found in saline environments; and ii) those species (glycophytes) that occur opportunistically in saline environments (see Appendix D). In the latter case, the plants can occupy spatial or temporal niches in saltmarsh where salinity is lowered, for example, by restriction of tidal inundation or after heavy rain has ponded fresh or low-salinity water into surface depressions. Such plants may be eliminated later, as evaporation increases the salinity of the water and/or soil. Many opportunistic glycophytes perform poorly in these moderately saline environments.

The third approach to saltmarsh delineation is to invoke the landward limit in hydrological terms. On the surface it appears an attractive method, but it may be complicated by the influence of rare events (e.g. storm surges, king tides, etc.) and the presence of salinised ground above tidal reach due to geomorphological changes. In Australia, the Queensland Parks and Wildlife Service describes saltmarshes as being situated in the intertidal zone, below the level of the highest astronomic tide but well above the low tide level. It recognises that the highest areas of saltmarsh are inundated with seawater only by the highest spring tides (Queensland EPA 2009b). Although not as explicit as that implemented in Queensland, the New South Wales approach to defining saltmarshes also implies a tidal delimitation, with mangrove swamps found on land inundated by ‘every tide’, saltmarshes in the zone inundated by ‘higher and spring tides’, and terrestrial vegetation limited to the zone above ‘flood, storm or king tides’ (Department of Environment and Climate Change 2009). The list of different definitions of the term ‘coastal saltmarsh’ indicates also the near ubiquity of tidal inundation as a distinguishing characteristic (Table 1.1).


Invoking a landward boundary in hydrological terms, specifically the extent of tidal inundation, seems at first to have a number of attractions. It has the benefit of allowing coastal saltmarsh to be clearly differentiated from inland saltmarsh. Many salt-affected areas of inland Australia are vegetated with halophytic plants that would be recognised as characteristically saltmarsh plants. For example, shrublands dominated by succulent chenopods are found across much of inland continental Australia, especially around periodically inundated salt lakes (Briggs 1981; Leigh 1981), as well as around the margins of salt lakes in the dry midlands of Tasmania (Kirkpatrick & Harris 1999). In Victoria, EVC 708 Hypersaline Inland Saltmarsh Aggregate contains genera (and, in some cases, also species) of halopytic plants that are common in coastal saltmarsh too. But being in an intertidal zone is not sufficient, as many intertidal areas (e.g. beaches and rock-platforms) do not support saltmarsh. Instead, saltmarshes are almost always found along low wave-energy parts of the coast, often in bays, inlets, estuaries, the coastal side of barrier islands and in the lee of sand spits (Barson 1977). Along such sheltered coastlines saltmarsh develops on land that is inundated to the extent of high spring tides. Tidal factors control also the seaward boundary of a saltmarsh, as most saltmarsh plants (with the exception of the introduced *Spartina: Bird & Boston 1968) seem unable to withstand permanent inundation with seawater, or intense tidal or wave action.

A final complication with the hydrological criterion is that groups of plants identical to those found in coastal saltmarsh may occasionally occur in coastal areas not subject to tidal inundation at all. Salt spray, sometimes combined with seepage, for example, can encourage the development of halophytic plant communities on cliffs, bluffs and rock shelves well above tidal reach (e.g. the cliffs of Lady Julia Percy Island, south-west Victoria; see Carr & Mathews 2004). In New South Wales, extensive swards of Zoysia, Sporobolus and Sarcocornia spp. can occur on local seepage zones, and may form an important linkage between estuaries or contribute propagules that could enter estuarine areas (Paul Adam, University of New South Wales, pers. comm.).

Conclusions

We conclude that the best approach to defining ‘coastal saltmarsh’, and especially the problematic issues of setting a practical landward boundary and separating saltmarsh from other types of estuarine wetland, is to adopt a definition that uses geomorphic, hydrological and structural vegetation attributes. First, it is recognised that saltmarshes are distinctive plant communities that are found mostly along low wave-energy sectors of the coast. Along such sheltered coastlines, saltmarsh develops on land that is subject to inundation by the tides, the periodicity of which can vary from regular to rare or episodic, according to elevation and meteorological conditions. The landward extent of coastal saltmarsh in Victoria is largely determined by the penetration of spring high tides, whereas the seaward extent is determined by the depth, duration and periodicity of routine tidal submergence, intensity and frequency of mechanical disturbance due to coastal processes (i.e. tidal or wave action), and type of substratum (e.g. rocky vs muddy). Although soil type seems not to be limiting to the development of saltmarsh vegetation, the drainage characteristics of coastal soils, along with tidal patterns and elevation, certainly influence strongly the arrangement of plant species and communities within saltmarshes (Clarke & Hannon 1967, 1969, 1970, 1971).


The definition we adopt for the purposes of this report and to meet its stipulated aims (Appendix A) is that coastal saltmarsh is:

land that experiences recurrent low-energy inundation by seawater and which is vegetated by low-growing vascular plants (generally < 1.5 m height), such as succulent chenopods and salt-tolerant monocots.

Note that this definition excludes vegetation in the salt-spray or splash zone in high-energy environments, such as along cliffs. While the focus of the current project is on mangroves and coastal saltmarsh, it necessarily also considers some other types of estuarine wetlands where they are spatially integrated with coastal saltmarsh (Table 1.2). Previous mapping studies on estuarine wetlands in Victoria (e.g. Sinclair & Sutter 2008) indicate that swards of Juncus kraussii, for example, frequently occupy the interface between coastal saltmarsh and other estuarine vegetation systems. Moreover, Juncus kraussii is often considered a component of ‘core’ intertidal saltmarsh in south-eastern Australia (see Table 1.4). These other wetland types are considered in the typology and inventory components of the project, as outlined in Chapters 4 and 5, respectively.

1.3 Inundationregimes

tidal inundation

As the relationship between the level of the land and the level of the sea is central to the formation and function of coastal wetlands (Woodroffe & Davies 2009), tidal regimes are critical to understanding the ecology of mangroves and coastal saltmarsh (Por 1984a,b; Perillo 2009). Tidal regimes not only delimit the upper boundaries of coastal saltmarsh and the commencement of a truly terrestrial flora, but they also set the limit to which saltmarsh can colonise into areas currently vegetated with mangroves or seagrasses (Figure 1.3).

Figure1.3: Mangroves Avicennia marina at high tide, Barwon River estuary.


Tides are generated by periodic changes in sea level caused by the gravitational effects of the sun and moon on the ocean. There are four astronomic sources of temporal variability in tidal amplitude. At the shortest temporal scale, high tides alternate with low tides with a periodicity of about 12.5 hours along coasts that experience a semi-diurnal tidal cycle (such as much of southern Australia). Longer-term variations occur on a roughly fortnightly time scale when the sun, moon and earth are aligned. That alignment results in tides with unusually large amplitude – spring tides – and takes place 1.5 days after new and full moons. Tides with much smaller amplitude – neap tides – occur 7 days after spring tides, when the three celestial bodies are not aligned. The amplitude of spring tides varies also over an annual time scale, and is greatest around the solstices and equinoxes. Much longer-term variations in tidal amplitude occur over a time frame of 18.6 years, and are produced by variations in solar and lunar declination. Such long-term changes in tidal amplitude have a profound effect in structuring coastal biotic communities (Bleakney 1972; Swinbanks 1982). Adam (1990), for example, noted that historical fluctuations in the dominance of Poaceae and Cyperaceae in a Long Island (USA) saltmarsh were related to the 18.6 year tidal cycle, with sedges dominating over grasses when the tidal range was smallest. Wolanski (2007) outlined the various components of estuarine tidal cycles in considerable detail.

Although tidal amplitude can be predicted on the basis of the superimposition of these four astronomic factors, variability in the actual tidal regime experienced by a given saltmarsh varies widely, according to site-specific geomorphology and more general meteorological conditions. Funnel-shaped estuaries, for example, usually give rise to an amplified tidal range, whereas the tidal range may be greatly attenuated in those estuaries with a very constricted opening to the sea. Prolonged onshore winds can raise tidal heights, offshore winds can depress tidal heights, and storm surges can greatly modify the amplitude of tidal inundation. In estuaries that experience large inputs of fresh water, river discharge can strongly influence the extent of saline intrusions and tidal extent. In the case of the Gippsland Lakes, for example, riverine discharge of over ~130 GL month–1 into Lake Wellington prevents the intrusion of saline water from McLennan Strait and the easterly lakes that are connected directly to the ocean (Robinson 1995).

Within individual saltmarshes, the periodicity and amplitude of tidal inundation varies also according to elevation and microtopographical relief. Ranwell (1972) differentiated between ‘intertidal’ and ‘high’ marshes; the former were posited to experience frequent tidal inundation and have relatively constant soil moisture and salinity regimes, whereas the latter experienced unpredictable inundation and had highly variable edaphic conditions, being influenced by evaporation and rainfall as well as seawater. A similar distinction seems to be valid for Victorian coastal saltmarsh, where ‘lower’ and ‘upper’ saltmarsh can be differentiated on the basis of elevation and hydrological conditions (see Chapter 1.5 for more detail).

A simple differentiation of saltmarshes along a putative elevational gradient from the ocean, however, does not fully reflect the microtopographic diversity in a given marsh. Saltmarshes and mangroves in Victoria are often criss-crossed with tidal channels, which modify the penetration of saline intrusions (Figure 1.4). These complications in extrapolating the tidal regime experienced by saltmarshes from tidal cycles that operate in the open ocean have major implications for determining the likely impact of climate change on coastal saltmarsh. Perillo (2000) provided a detailed overview of the importance of tidal channels in the ecology of coastal wetlands.


Figure1.4: Tidal channels in mangroves and coastal saltmarsh, Tooradin.

freshwater inundation

Although their connection, albeit sometimes rare or episodic, with the ocean is the most obvious aspect of the hydrology of coastal saltmarshes, freshwater inputs play an important role in structuring their vegetation, fauna and ecological interactions (Boorman 2009). This is because saltmarshes receive water not only via tidal inundation with seawater but also receive fresh water directly via rainfall, indirectly as runoff from the terrestrial hinterland, as groundwater flows, and as episodic inundation from flood-swollen rivers that discharge into estuaries. The relative ratio of inundation with tidally forced seawater water and with fresh water is thought to be factor that controls many aspects of saltmarsh ecology (Adam 1990, 2002; Boorman 2009). Some of the interactions have been well described for coastal saltmarsh in the Sydney region, following the classic studies by Clarke & Hannon (1969). The interactions among tidal inundation, rainfall, height of the water table, soil salinity and waterlogging, and their effects on plant performance are shown in Figure 1.5.

Inputs of fresh water affect not only the extent and duration of land that is inundated and thus waterlogged, but also the salinity of surface soils and water that ponds in low-lying areas within a saltmarsh (Beeftink 1977) and the importation of sediments and nutrients (Boorman 2009). Groundwater intrusions can influence soil salinities and the position of the freshwater-brackish water interface in coastal saltmarsh (Werner & Lockington 2006; Wolanksi 2007; Carter et al. 2008), as well as possibly having a role also in nutrient dynamics (Krest et al. 2000). As demonstrated with the Coorong in South Australia, a severe reduction in freshwater inflows can have devastating effects on coastal saltmarsh and other coastal wetlands (Adam 2002).


Many perennial saltmarsh plants require periods of reduced salinity to germinate (see Chapter 1.8) and seasonal relief of high salinities can prompt annuals to divert resources to reproductive structures and set seed. Ponds of stagnant rainwater have been shown to be critical factors in the occurrence of many taxa in coastal saltmarshes of western and northern Europe (Beeftink 1977). Many intertidal wetlands along the Victorian coast possess large depressions that can fill with rainwater and remain unvegetated or soon be colonised by submerged angiosperm taxa such as Ruppia and Lepilaena spp. (Figure 1.6). Such small depressions constitute an important component of the environmental heterogeneity of coastal saltmarshes, and are often valuable foraging sites for waterbirds. They are important also as breeding sites for mosquitoes and biting midges, a characteristic that is likely to become more significant to nearby human populations with climate change and increasing population densities around the coast (see Chapter 1.3). In Victoria, vegetation in these ponds is usually described as EVC 842 Saline Aquatic Meadow.

Figure1.5: Importance of tidal inundation and rainfall in controlling hydrological and salinity regimes and plant performance in coastal saltmarsh. Source: Nybakken (2001, Figure 8.18), in turn based on Clarke & Hannon (1969).


1.4 Victorianmangroves

Australia has the fourth-highest species diversity of mangrove taxa of any country after the Philippines, Indonesia and Papua New Guinea (Duke 2006). Only one taxon of mangrove, Avicennia marina subsp. australasica, occurs in Victoria. It has a discontinuous distribution, from the Barwon River in the west (38o17’S, 144o30’E) to McLoughlins Beach in the east (~38o40’S, 146o52’E) of the state. Mangroves are particularly well developed around Western Port (including on French Island) and in the Corner Inlet-Nooramunga area of South Gippsland. Interestingly, they are present also in Cunninghame Arm at Lakes Entrance, and are believed to be progeny of planted specimens (Figure 1.7).

Figure1.7: Avicennia marina, Lakes Entrance, Gippsland Lakes.

Figure1.6: Pools in saltmarsh backed by Tecticornia arbuscula and Avicennia marina zones, Hastings.


The most southerly occurrence of mangroves in Australia (and indeed the highest latitude occurrence of mangroves anywhere in the world) is in Victoria, in Corner Inlet, where Avicennia marina occurs to a latitude of 38o45’S (Duke 2006; Spalding et al. 2010). Figure 1.8 shows mangroves at Millers Landing, their most southerly limit in Australia.

In general, Avicennia marina occurs along the Victorian coast as a dense, monospecific shrubland, with individuals growing as shrubs or small trees from 0.3 to 4 m tall (Denis 1994; Harty 1997; Duke 2006). Their densities, however, can vary from individual plants growing sparsely on the shoreline to dense, near-continuous belts of vegetation (Figure 1.10). There is little or no angiosperm understorey, except in the better-drained zones or in landward areas where mangroves become more scattered and where Sarcocornia quinqueflora and Triglochin striata may occur as a sparse ground layer. Nevertheless, much of the sediment may be covered by brown algae, including Hormosira banksii. Algal mats of Bostrychia and Catanella species are usually associated with pneumatophores and lower tree stems (Morgan 1986).

Mangroves extend considerably further down the eastern Australian coast than they do along the Western Australian coast, where Avicennia marina is found only as far south as Bunbury (33o45’) (Lear & Turner 1977). Frost and/or low winter temperatures are believed to be the environmental factor that limits the distribution and vigour of mangroves in southern Victoria (Ashton 1971; Lear & Turner 1977; Oliver 1982). Presumably because of these environmental factors, mangroves in southern Victoria are often quite stunted and can take on what has been called a ‘bonsai’ form (Figure 1.9).

Figure1.8: Avicennia marina, Millers Landing.

Figure1.9: ‘Bonsai’ Avicennia marina, Toora.

Figure1.10: Comparison of densities of Avicennia marina mangrove shrubland in Victoria: the right-hand image shows a ‘bonsai’ mangrove, Jam Jerrup Red Bluff; the left-hand image shows a continuous belt of mangroves, Tooradin.


Victorian mangroves are usually found on wholly muddy substrata. In a few locations, however, they can be found growing on a rocky shoreline. The rocky substrata can be quite varied: basalt in the case of mangroves at The Jawbone; granite at Millers Landing; and Silurian sedimentary rocks south of Hastings in Western Port.

There are few estimates of the extent of mangroves in south-eastern Australia, and the estimates that do exist are in conflict. The variation is due mainly to three factors. First, the delineation of mangroves is sometimes difficult, especially where they merge into other types of estuarine wetland (Wilton & Saintilan 2000). Second, mangrove zones are often intricately shaped and their area measurement varies greatly depending on how closely and accurately the boundaries are delimited. Accordingly, it is well known that the resolution of any mapping greatly affects area calculations (Turner et al. 1989). Third, in some areas mangroves have been cleared or are expanding (see Chapter 1.11), which makes areal estimates quickly outdated.

Galloway (1982) estimated the total area of mangroves in Australia, on the basis of a 1 km2 dot grid on 1:85,000 aerial photographs, as 11,500 km2, of which 95% occurred in tropical Australia (see also Galloway 1981). Victoria was estimated as having 12 km2 of mangroves; New South Wales and South Australia as 992 and 201 km2, respectively. On the basis of the mapping undertaken by Bucher & Saenger (1991), Harty (1997) indicated there were 41 km2 of mangroves in Victoria. The area of mangroves in New South Wales and South Australia, again drawn from Bucher & Saenger (1991), was cited by Harty (1997, 2003, 2005a) as 107 km2 and 111 km2, respectively. The estimate for New South Wales is similar to that of Galloway (1982), but that for South Australia is different by a factor of nearly two. The OzCoasts database (OzCoasts 2009) reported that there were 109.83 km2 of mangrove in New South Wales, out of an Australia-wide area of 3,976 km2. The discrepancy with earlier estimates of nationwide extent is apparent. No values are available in the database for Victorian mangroves. The most recent and reliable mangrove mapping in New South Wales is based on interpretation of 1:25,000 aerial photographs and extensive field reconnaissance by Williams et al. (2006), who updated earlier work by West et al. (1985). They estimated the total area of mangroves in New South Wales to be approximately 124 km2.

Currently available vegetation mapping undertaken by the Department of Sustainability and Environment suggests that Victoria supports about 66 km2 of Mangrove Shrubland (EVC 140). On the basis of the mapping completed for this project, we present in Chapter 5 an updated estimate of the area of this EVC of 51.77 km2 for Victoria. As our mapping is by far the most detailed and consistent statewide, that figure is probably the most accurate.

1.5 Victoriancoastalsaltmarsh

distribution and area

Saltmarsh is found along many parts of the Victorian coast, although it is most extensive along the western coast of Port Phillip Bay, northern parts of Western Port, in the Corner Inlet-Nooramunga complex, and behind the sand dunes that line the Ninety Mile Beach in Gippsland, especially in Lake Reeve. In their analysis of vegetation of the Victorian coast, Barson & Calder (1981) concluded that coastal saltmarsh was best developed in the region between Barwon Heads and Corner Inlet. To the west of Barwon Heads, notable patches occur at Breamlea, the mouth of the Anglesea River, Painkalac Creek, Aireys Inlet, at Port Fairy, and


in the estuary of the Glenelg River. To the east of Corner Inlet, saltmarsh fringes the shoreline of parts of the Gippsland Lakes (especially in Lake Reeve), and extends as far east as the mouth of the Snowy River, to Wingan Inlet, and to Mallacoota.

There are few estimates of the area of coastal saltmarsh in Victoria and, as with those for mangroves, the data show little agreement. Drawing from the work of Bucher & Saenger (1991), Kelleway et al. (2009) stated that there were 125 km2 of saltmarsh in Victoria. (Note that this estimate may include all saltmarsh, not only coastal saltmarsh.) That is about twice that of the area of saltmarsh in New South Wales, but only ~2.5% of that in Queensland or the Northern Territory (Table 1.6). The total area of saltmarsh in Australia was estimated by Bucher & Saenger (1991) as 13,595 km2.

Table1.6: Estimates of the extent of Australian saltmarsh, based on Bucher & Saenger (1991). Source: Kelleway et al. (2009, Table 10.1).

Jurisdiction Length of shoreline (km) Area of saltmarsh (km2)

Northern Territory 6,200 5,005

Queensland 7,400 5,322

New South Wales 1,900 57

Victoria 1,800 125

Tasmania 3,200 37

South Australia 3,700 84

Western Australia 12,500 2,965

Total 36,735 13,595

In contrast to the values given by Kelleway et al. (2009), estimates of the area of Victorian coastal saltmarsh, based on mapping of Orange-bellied Parrot habitat by Carr (1979), indicate that < 60 km2 of coastal saltmarsh remained in the state. The OzCoasts database reports that there were 68.5 km2 of coastal saltmarsh in New South Wales, out of an Australian-wide area of 13,029 km2 (OzCoasts 2009). There is thus quite good agreement between the OzCoasts (2009) and Bucher & Saenger (1991) estimates of nationwide saltmarsh area. No values were available in the OzCoasts database for the area of coastal saltmarsh in Victoria. According to vegetation mapping currently available from the Department of Sustainability and Environment, the area of EVC 9 Coastal Saltmarsh Aggregate is ~132 km2. The results of the mapping and inventory component of the current project show that the area of coastal saltmarsh of all types in Victoria is currently 192.12 km2 (see Chapter 5, especially Table 5.5).

By comparison with the area in Victoria, the most recent saltmarsh mapping in New South Wales (as with that for mangroves, based on interpretation of 1:25,000 aerial photographs and extensive field reconnaissance), indicates that the area of coastal saltmarsh in New South Wales is ~72 km2 (Williams et al. 2006). In Tasmania, the total area of saltmarsh is 77 km2, accounted for by roughly equal areas of saline grassland, saline aquatic herbland and succulent saline herbland (Felicity Faulkner, Department of Primary Industries, Parks, Water and Environment, pers. comm.; see also www.dpiw.tas.gov/tasreserveestate, internet resource, accessed 10/05/2010).


types of coastal saltmarsh in victoria

A distinction has often been drawn between two general forms of saltmarsh on the Victorian coast (Barson & Calder 1981). The first, a ‘dry’ type, is present in the central-western parts of the state, where low summer rainfall and high temperatures lead to intensely hypersaline conditions in the more elevated sites. Here the vegetation is dominated by Tecticornia pergranulata and Tecticornia halocnemoides. The second is a ‘wet’ type, and is found where rainfall is higher; here the vegetation is dominated by samphires such as Sarcocornia spp. and Tecticornia arbuscula. Figure 1.11 shows an example of the two types.

Figure1.11: Comparison of ‘dry’ coastal saltmarsh (left) and ‘wet’ coastal saltmarsh (right) at Hastings. Note that there is a range of different types of ‘dry’ and ‘wet’ coastal saltmarsh and these two images are indicative only. The ‘dry’ saltmarsh is dominated by Disphyma crassifolium; ‘wet’ saltmarsh is dominated by Tecticornia arbuscula in the shrub layer and Sarcocornia quinqueflora as groundcover.

The geographic differentiation of the ‘dry’ and ‘wet’ forms is a direct consequence of patterns in rainfall and evaporation across Victoria. Annual rainfall ranges from ~1000 mm in far south-west and parts of eastern Victoria, to 400–600 mm in the orographic rain-shadow areas of the western shores of Port Phillip Bay, from Corio to Torquay, as well as in a relatively small area near Loch Sport in Gippsland. A small low-rainfall area (< 600 mm) also occurs near Lake Reeve, Gippsland (Bureau of Meteorology & Walsh 1993). In these Mediterranean-climate regions, summers are typically hot and dry, which leads to hypersaline conditions developing over summer in the mid to upper saltmarsh, relieved only by rainfall in the wetter seasons of autumn, winter and spring.

Additionally, Victorian authors have often distinguished between two types of saltmarsh on the basis of elevation and tidal inundation: ‘upper’ saltmarsh and ‘lower’ saltmarsh (Barson 1977; Carr 1982; Barson & Calder 1981). Figure 1.12 shows ‘upper’ and ‘lower’ saltmarsh at The Spit Nature Conservation Reserve, on the western shore of Port Phillip Bay.


As the ‘lower’ saltmarsh receives daily, or at least frequent, tidal inundation, it is often termed ‘wet’ saltmarsh. The terminology unfortunately coincides – and leads to potential confusion – somewhat with the geographic-climatic differentiation between ‘dry’ saltmarsh of the central-west of the state and the ‘wet’ saltmarsh of the east. In contrast to the ‘lower’ saltmarsh, the ‘upper’ saltmarsh is subject to less frequent tidal inundation, mostly under rare conjunctions of tidal, meteorological and other factors which lead to unusually high tides or storm surges. As shown in Table 1.2, a range of brackish-water wetlands or dryland vegetation may occur landward of ‘upper’ saltmarsh in Victoria. EVC 842 Saline Aquatic Meadow, also dominated by submerged perennial or annual monocots, occurs at various spatial scales where shallow ponds are formed, either permanently or seasonally (Figure 1.6). Water in these ponds may become hypersaline in summer, which leads to salinities that are lethal to vegetative growth of even highly salt-tolerant macrophytes. The plants regenerate from a soil-stored seed bank or dormant underground organs with the restoration of hyposaline conditions, usually following rainfall in autumn, winter or spring.

floristics and life forms

Coastal saltmarsh across the world is dominated by plants belonging to relatively few families, presumably due to the evolutionary constraints of overcoming the physiologically harsh conditions. The vast majority of saltmarsh plant species belong to the families Chenopodiaceae, Cyperaceae, Juncaceae and Poaceae, with some representatives also from the Amaranthaceae, Apiaceae, Frankeniaceae and Plumbaginaceae (Adam 1990). Globally, the range of common genera is similarly quite limited: Salicornia and Sarcocornia, Juncus, Spartina and Plantago are the common genera in coastal saltmarshes of the Northern and Southern Hemispheres (Chapman 1977). Even so, there are some marked floristic differences, especially at the species level, in saltmarshes of different regions (Day et al. 1989; Kennish 1990; Adam 2002). Along the

Figure1.12: Comparison of lower saltmarsh dominated by Tecticornia arbuscula (background) and upper saltmarsh dominated by Frankenia pauciflora, Disphyma crassifolium and Sarcocornia quinqueflora, The Spit Nature Conservation Reserve.


north Atlantic coast of North America, for example, Spartina alterniflora dominates the zone between mean sea level and mean high-water level. This species then grades landwards into mixed Juncus-Salicornia communities, with some Distichlis spicata and Spartina patens present as well. In contrast, saltmarshes along the south Atlantic coast of the USA are dominated by Spartina alterniflora, Salicornia bigelovii, Sarcocornia europaea, Sarcocornia virginica, Distichlis spicata and Juncus roemerianus. The northern Pacific coast of the USA supports a saltmarsh flora that differs again, and is more diverse than that of the American Atlantic coast. South American saltmarshes are frequently dominated by taxa such as Spartina, Limnonium, Distichlis, Juncus and Heterostachys. In Europe, Atlantic saltmarshes are typically dominated by Festuca and Puccinellia; saltmarshes in southern England by Spartina anglica and Spartina townsendii; and along the North Sea by species of Armeria, Limonium, Spergularia and Triglochin, with some Plantago maritima. Along the Mediterranean coast, low shrubby vegetation dominates the saltmarsh flora. Arctic saltmarshes are generally dominated by Puccinellia and Carex species. In southern Australian coastal saltmarsh, Plantago spp. are rare and opportunistic, and Austrostipa, Disphyma, Distichlis, Frankenia, Limonium, Puccinellia, Sarcocornia, Spergularia, Suaeda, Tecticornia and Triglochin are common genera. The floristics and life forms of Victorian coastal saltmarsh are discussed in detail below (see Tables 1.7 & 1.8 and accompanying text; see also Appendix D).

Saenger et al. (1977) provided the most exhaustive list of Australian saltmarsh plant species, but King et al. (1990) argued that it was incomplete because criteria for inclusion of species in the upper saltmarsh were inconsistently applied. Saenger et al. (1977) employed a three-way geographic split of Victorian taxa into east, central and west areas, and this system fails to acknowledge the rainshadow/non-rainshadow component of the central region of the state. ‘Dry’ saltmarsh is found in the western parts of, for example, Port Phillip Bay, and ‘wet’ saltmarsh around the more easterly, and wetter, Western Port. The two regions are thus conflated in Saenger et al.’s (1977) compilation, which severely limits the utility of the list. Moreover, we have concerns about some erroneous (or, at least, debatable) inclusions in the species list. The floristics of Victorian saltmarshes are described by Bridgewater (1975, 1982) and Bridgewater et al. (1981), and those of New South Wales by Adam (1981b), Adam et al. (1988) and Saintilan (2009b). Saintilan (2009a) reported a nationwide assessment of the biogeography of Australian saltmarsh plants.

Seen together, these studies indicate that two families are widely represented in Australian saltmarshes: Chenopodiaceae and Poaceae (Love 1981). Although chenopods are a conspicuous feature of Australian coastal saltmarsh, as a family they are not at all restricted to saltmarshes or saline environments. Leigh (1981) noted that a total of 280 species in the family Chenopodiaceae are endemic to Australia, and another 17 species had become naturalised; the 21 species found in saltmarshes represent less than 10% of the total number of chenopod species found in Australia. The non-saltmarsh chenopods generally occur in low rainfall regions or in sites that suffer water stress due to other factors (e.g. osmotic stress or rockiness). That observation highlights the fact that salt stress and drought stress are physiologically similar in some respects, and that the Chenopodiaceae are genetically capable of exploiting many environments that suffer water deficit, regardless of its causes. As discussed below, this is often achieved by succulence. Note that the taxonomy of an important group of chenopods in Australian saltmarsh – the genera Halosarcia, Pachycornia, Sclerostegia, Tecticornia and Tegicornia – has been recently revised (Shepherd & Wilson 2007). These genera are now placed in an expanded circumscription of Tecticornia.


In view of the limitations of the species list prepared by Saenger et al. (1977), we prepared a more comprehensive floristic checklist of the Victorian vascular flora of coastal saltmarsh (Appendix D). It is summarised in Tables 1.7 and 1.8. The inventory of the Victorian vascular flora of coastal saltmarsh and mangrove lists a total of 140 indigenous taxa (species, infraspecific taxa and informal names) that have been recorded in upper and lower saltmarsh in the state. Of the total, 68 species are halophytes, being confined to saline environments or able to tolerate moderate salinity if not confined to saline areas. In other words, we calculate that halophytic taxa constitute 48% of the total indigenous saltmarsh flora. For the inventory, species of coastal saltmarsh were defined as those occurring in upper and lower saltmarsh and associated Saline Aquatic Meadows landward of and including the communities (or EVCs) dominated by Gahnia filum and Austrostipa stipoides. These taxa are indicated in bold in Appendix D; excluded is a fairly large suite of glycophytic species occurring opportunistically in coastal saltmarsh and which are characteristic of vegetation of non-saline habitats. Only 14 taxa, including Avicennia marina, occur in lower saltmarsh and only three of these (Avicennia marina, Atriplex paludosa subsp. paludosa and Tecticornia arbuscula) are largely confined to the lower elevations.

Table1.7: Biogeography of the Victorian indigenous flora of saltmarsh, mangrove and saline aquatic meadows.

Biogeographic origins Total flora including glycophytes

Species confined to saline environments

1 Southern temperate Australian endemic 59 (42%) 29 (42%)

2 South-east Australian endemic 29 (21%) 15 (22%)

3 Southern/south-eastern Australian – New Zealand endemic

17 (12%) 10 (15%)

Subtotal 105 (75% of saltmarsh flora) 54 (79% halophyte flora)

4 Southern Hemisphere temperate element 3 (2%) 1 (1%)

5 Cosmopolitan temperate element 7 (5%) 2 (3%)

6 Southern/eastern Australian – Malesian/Western Pacific/New Zealand endemic

12 (9%) 6 (9%)

7 Old World – Western Pacific tropical/subtropical element

2 (1%) 1 (1%)

8 Southern Australia – South Africa element 2 (1%) 0

9 Pantropical element 9 (6%) 4 (6%)

Total 140 taxa 68 taxa

75% Southern Hemisphere elements

79% Southern Hemisphere elements


Table1.8: Life forms in the Victorian saltmarsh flora.

Indigenous flora

Exotic flora

Floristic diversity

Total plant species/taxa in saltmarsh, mangrove and saline aquatic meadows (excludes marine submerged angiosperms)

131 118

Total species/taxa in lower (wet) saltmarsh/mangrove 14 2

Species/taxa confined to lower saltmarsh/mangrove environments 1 2

Total species/taxa in upper (dry) saltmarsh 131 115

Species/taxa confined to upper saltmarsh or other saline environments 64 c. 30

Life forms

A annual (non-succulent) 34 52

B biennial 2 10

Ea emergent aquatic 2 2

Gb bulbous geophyte 0 1

Gc cormous geophyte 1 4

Gt tuberous geophyte 1 0

Ls large shrub 4 1

Pa parasite (herbaceous) 1 1

Pr perennial herb (rhizomatous or stoloniferous) 29 13

Pt perennial herb (tufted or tussock-forming) 27 26

S small to medium shrub 4 0

Sa submergent aquatic (macrophyte: stoloniferous/rhizomatous) 14 0

Ss subshrub 4 0

T tree 2 2

V vine 1 0

X-A succulent annual 1 2

X-S succulent shrub/subshrub 4 0

X-P succulent rhizomatous/stoloniferous herbs 5 1

At the family and generic levels, there are strong affinities with saltmarshes elsewhere in the Northern and Southern Hemispheres. At the species and infraspecific levels, however, there is a major endemic southern and south-east Australian element in the total Victorian saltmarsh flora (including halophytes and opportunistic glycophytes): some 60% of taxa are endemic in temperate Australia. A similar figure for endemism (64%) also applies at the specific and infraspecific levels to the more strictly halophytic flora of Victorian coastal saltmarsh (Table 1.7). In respect of extra-Australian affinities of the saltmarsh, mangrove and saline aquatic meadow flora, the strongest affinities are with the New Zealand flora at family, generic and species levels (Wardle 1991; Haacks & Thannheiser 2003). Indeed, a total of 12% of the saltmarsh, mangrove and saline aquatic meadow flora and 15% of the more strictly halophytic flora is shared with New Zealand. The combined figures indicate that 73% of the Victorian saltmarsh, mangrove and saline aquatic meadow flora consists of a southern Australian or shared New Zealand endemic element.


A cosmopolitan temperate and pantropical floristic subset of the Victorian saltmarsh, mangrove and saline aquatic meadow flora runs to about 13% of the total. Another significant element of wide distribution (9% of the flora) is shared by southern and eastern Australia, Malesia, the Western Pacific and New Zealand (Table 1.7). Surprisingly, only very weak affinities at the species level (3 species, 1% of flora) are evident between Victoria and South Africa. At the generic level there is also an important endemic element, and strictly Australian endemic genera in the Victorian marine and saltmarsh flora include Amphibolus, Angianthus, Hemichroa, Hyalosperma, Lawrencia, Rhagodia and Wilsonia. Many genera are near endemic in Australia, for example, Tecticornia (Wilson 1972, 1980; Shepherd & Wilson 2007).

Table 1.8 shows the variety of life forms in the saltmarsh flora of southern Australia. The diversity in life forms translates to a startling diversity in the structure of Victorian coastal saltmarsh, as discussed later. The life forms observed in saltmarsh species relate to the suite of life-history strategies available to plants to tolerate saline, waterlogged and anoxic environments. These characteristics are discussed in detail below (Chapters 1.7 & 1.8), from physiological and reproductive perspectives. Essentially, species must either possess specialised tolerance mechanisms, or else they must opportunistically exploit available niches when environmental conditions are transiently suitable. In the latter case, they effectively avoid exposure to the more extreme of the already highly stressful conditions that typify most coastal saltmarsh environments. Although few in number, many of the structural dominants in coastal saltmarsh are succulent shrubs.

Contrary to the claims of Bridgewater & Kaeshagen (1979), life-form statistics for the Victorian saltmarsh flora indicate that obligate annuals, as well as a few facultative annuals, are well represented in the flora. Indeed, of the total Victorian saltmarsh flora of 131 taxa (species and infrataxa), 33 (or 25%) are obligate annuals. All except one of the annual species – the submerged aquatic macrophyte Lepilaena cylindrocarpa – are confined to upper or dry saltmarsh.

The total flora includes some opportunistic glycophytes, but of the essentially halophytic flora more or less confined to such saline environments – 67 taxa – no fewer than 19 (28%) are more-or-less obligate halophytes. The annuals of the upper saltmarsh are best represented in the floristically richest coastal saltmarshes located west from Melbourne to Torquay, where rainfall is relatively low. Many such species likewise occur in drier north-western Victoria in non-coastal saltmarshes. Examples include Angianthus preissianus, Cotula vulgaris var. australasica and Senecio halophilus. In these Mediterranean climates the annuals establish after the autumn break when salt is leached out of the upper soil profile.

Coastal saltmarsh-type vegetation in Victoria can also occupy saline environments outside the influence of tides. Two general situations apply. The first is cliffs and headlands on high-energy coasts, where the vegetation is floristically and structurally equivalent to saltmarsh but is influenced by wave splash and spray (the spray zone) and atmospheric salt deposition rather than by tides. The second is coastal saltmarsh formerly connected to the sea but now no longer under tidal influence because of falling Holocene sea levels. In the case of the halophytic vegetation of cliffs and headlands, examples may occur anywhere along the Victorian coast, but perhaps the best development is seen at Lady Julia Percy Island off the western Victorian coast. That vegetation type was termed Beaded Glasswort (Sarcocornia quinqueflora ssp. quinqueflora) Halophytic (Saltspray) Cliff Herbfield by Carr & Mathews (2004). Beaded Glasswort herbfields were described and mapped also at Portland, ~200 m from the sea, on an elevated headland (Carr & Beauglehole 1980). These communities differ from those of cliffs because they occur in endorheic seasonal wetlands. Halophytic spray-zone vegetation has also been described for Tasmania (Kirkpatrick & Glasby 1981; Kirkpatrick & Harris 1999) and New South Wales (Keith 2004).


Saltmarsh vegetation no longer connected to the sea occurs in various locations on the Victorian coast, but is perhaps best developed in the Lonsdale Lakes area. Here very extensive saltmarsh occurs at Freshwater Lagoon and in the Lake Victoria system (North et al. 2007). These saltmarshes are influenced by the residual salinity of former tidal regimes and by atmospheric salt deposition from the high-energy Bass Strait coast 0.5–2.4 km to the south. Atmospheric deposition can contribute large amounts of salt to coastal areas; for example at Kingston in South Australia, atmospheric salt deposition 120 m from the ocean has been measured at the rate of 368 kg ha–1 yr–1 (Cole et al. 2003). Atmospheric salt deposition in Victorian coastal saltmarsh may be much more important as a source of salinity than generally recognised.

Adam (1990) argued that grasses were relatively unimportant in Victorian saltmarshes, with the exception of local areas dominated by Distichlis distichophylla, Austrostipa stipoides and Puccinellia stricta. Austrostipa stipoides can be a conspicuous feature of upper saltmarshes in Victoria (Figure 1.13), Distichlis distichophylla can form localised swards (Figure 1.14), and Sporobolus virginicus may form extensive swards on some New South Wales and Queensland marshes. Nevertheless, Adam’s conclusion does indicate an important floristic difference between the saltmarshes of southern Australia and those of the Northern Hemisphere: those in the Northern Hemisphere are frequently dominated by tall grasses (e.g. Spartina spp. in temperate saltmarsh of North America, and Puccinellia spp. in arctic and temperate European saltmarsh). The difference has been noted in many other reviews of Australian saltmarsh (e.g. Love 1981; Morrisey 1995) and occasionally acknowledged by Northern Hemisphere writers (e.g. Kneib 1997), but is important enough to restate here. The introduced grasses *Spartina anglica and *Spartina x townsendii are present in some Victorian coastal saltmarshes, although they seem not to have spread into coastal areas of New South Wales. Chapter 1.12 addresses *Spartina spp. in more detail.

Figure1.13: Austrostipa stipoides grassland at the edge of upper saltmarsh with Sarcocornia quinqueflora herbfield and scattered Tecticornia arbuscula shrubs, The Spit Nature Conservation Reserve.


Figure1.14: Distichlis distichophylla at the landward edge of upper saltmarsh, The Spit Nature Conservation Reserve.

Although saltmarsh vegetation is commonly thought to be species-poor, quantifying the actual number of species present depends on the definition of what constitutes a saltmarsh (see Table 1.1) and the equally variable definition of what constitutes a ‘saltmarsh plant’. These issues are reflected in divergent estimates of the number of plant taxa in Australian saltmarshes in the literature. Turner et al. (2004, page 47) included a diagram that showed the temperate coastal saltmarsh of southern Australia contained 10−30 plant species, in comparison with < 10 species for the northern parts of the continent. In contrast, Morrisey (1995) argued that more than 200 saltmarsh plant species had been recorded in Australia. Keith (2004), however, reported that more than 200 species of flowering plants had been recorded in saltmarshes merely between Forster and Bega on the central coast of New South Wales. Saenger (1994) stated that between 50 and 60 species, belonging to 16 families, had been recorded from saltmarshes of southern Australia. That estimate is difficult to reconcile with Barson & Calder (1981) noting that 49 plant species (including exotic taxa) were found in the saltmarshes of Western Port alone. Saintilan (2009b) recorded nearly 80 species of plants in Australian coastal saltmarsh, and Saintilan (2009a) documented the distribution of 93 species of saltmarsh plants across the country. As shown in Table 1.8, our tally of the number of vascular plant species in Victorian coastal saltmarsh, mangrove and associated saline aquatic meadows (but excluding submerged marine angiosperms) is 249, of which 47% are exotic.

Notwithstanding such disagreement about their floristic diversity, it is clear that individual saltmarshes along the south-eastern coast of Australia contain, in absolute terms, few species. In saltmarshes around Sydney, for example, most contain four species or fewer (Morrisey 1995). Species diversity increases with higher elevation within a given saltmarsh, and the upper saltmarsh can be considered species-rich (Adam 1990). The increase in species richness with increasing elevation is likely to be caused by the increasing environmental heterogeneity of the upper marsh, given the impact of freshwater inflows or deficit (caused by seepage, rainfall, etc.), and varying soil types away from the lower marsh terraces.


Although Victorian coastal saltmarsh may contain, in absolute terms, a low number of plant species, it is floristically more diverse than coastal saltmarsh in more northerly parts of the country. A pattern of decreasing floristic diversity in saltmarshes with decreasing latitude has been commented on by a number of reviewers (e.g. Turner et al. 2004; Saintilan 2009a,b; Saintilan et al. 2009) and is not limited to Australia (Adam 1994). Specht & Specht (1999), for example, identified about 30 species of plants in Australia saltmarshes at 40oS, compared with < 10 species at latitudes of 10−20oS (Figure 1.15). Similarly, Saenger et al. (1977) identified only seven species (belonging to four angiosperm families) for saltmarshes in tropical Australia, but 36 species (belonging to 16 angiosperm families) for Victorian coastal saltmarshes at a latitude of about 35oS. The reasons for such strong latitudinal patterns are not well understood (Adam 1994), but the analysis published recently by Saintilan (2009a) hints strongly at mean minimum temperature as a controlling factor.

Figure1.15: Relationship between species richness in saltmarshes and mangroves with latitude along the coastline of eastern Australia. Source: Specht & Specht (1999, Figure 17.10).

The greater floristic diversity of temperate coastal saltmarsh compared with tropical coastal saltmarsh does not translate to the relative extent of the habitats in the two regions. Table 1.9 (cf. Table 1.6) shows the percentage by area of saltmarsh in different Australian jurisdictions and the comparable saltmarsh flora, as reeported by Saintilan (2009b). The greater diversity of plant species in southern parts of the country (e.g. Tasmania, Victoria) over northern Australia is evident. Adam (1994) argued that tropical saltmarshes were far more extensive than those in higher latitudes, because of the greater tidal amplitude on tropical coasts and the often flat topography, which combined to allow for the development of extensive plains of coastal saltmarsh.

Table1.9: Percentage of national saltmarsh area and comparable percentages of the national saltmarsh flora occurring in each jurisdiction. Source: Saintilan (2009b, Table 2.2).

Jurisdiction Saltmarsh area (%) Saltmarsh flora (%)

Northern Territory 37 18

Queensland 38 29

New South Wales < 1 45

Victoria < 1 55

Tasmania < 1 53

South Australia < 1 71

Western Australia 22 72


structure

While many saltmarshes in the Northern Hemisphere are dominated exclusively by grasses and/or herbs, many Victorian marshes are structurally far more complex. The complexity arises from the range of life forms present in the flora (Table 1.8; Appendix D), and the fact that species with different life forms can be assembled into complex patterns within a given marsh. Succulent shrubs such as Tecticornia spp. form one structural type. Large tussock-forming monocots (e.g. Austrostipa stipoides and Gahnia filum) make up one form, and low rhizomatous/stoloniferous grasses another (e.g. Distichlis distichophylla and Sporobolus virginicus). Succulent herbs (e.g. Sarcocornia spp., Hemichroa pentandra, Disphyma clavellatum) and prostrate shrubs (e.g. Frankenia pauciflora) constitute a different structural form. Local patches dominated by annuals are also not uncommon. Submerged macrophytes in the shallow pools of low-lying depressions are another structural form (Figure 1.6). Structural and floristic distinctions are used in our revised typology for coastal saltmarsh (Chapter 4), and are likely to be important ecologically, for example, in providing heterogeneous habitat for different faunal species (Chapter 1.9).

exotic species

The weed flora of southern Australia is derived largely from agriculture and horticulture, neither of which include many species tolerant of saline mudflats (Carr et al. 1992). Nonetheless, wetlands seem particularly susceptible to weed invasions (Zedler & Kercher 2004) and those species capable of invading saltmarsh may often pose a serious threat. The uppermost zone of saltmarshes was posited by Adam (1990) to be particularly sensitive to invasion by exotic plant species. The claim is borne out by the data presented in Table 1.8 for Victorian coastal saltmarsh, which shows 115 exotic taxa are found in the upper dry saltmarsh, in comparison with only 2 for the lower wet saltmarsh.

King et al. (1990) argued that, while the upper saltmarsh provided habitat for many alien plant species, the potential for them to spread was limited. It would seem that this was an overly optimistic assessment, as more than 100 exotic species have been reported for the upper dry saltmarsh in Victoria alone. Adam (1994) noted that coastal saltmarshes in New South Wales were invaded by a number of aggressive weeds, including *Juncus acutus (introduced from the Mediterranean and/or California: Paul Adam, University of New South Wales, pers. comm.) and *Cortaderia selloana (from South America). In northern New South Wales and southern Queensland, *Baccharis halimifolia is an important weed of coastal saltmarsh. Other exotic taxa in New South Wales coastal saltmarsh include a suite of small annuals or other short-lived species, including *Plantago coronopus, *Limonium binervosum, the composite *Aster subulatus and the grasses *Parapholis incurva and *Polypogon monspeliensis (Morrisey 1995).

In Victoria the most problematic weed species in upper saltmarsh include *Lophopyrum ponticum, *Parapholis incurva, *Hordeum marinum and *Juncus acutus. Other common exotic species include *Atriplex prostrata, *Parapholis strigosa and *Hordeum marinum. The very lowest levels of Victorian coastal saltmarsh can be invaded by *Spartina anglica and *Spartina x townsendii. Upper levels are commonly invaded by a much larger suite of weeds, often agricultural or garden escapees. Weeds are discussed in greater detail in Chapter 1.11. Chapter 1.12 has been allocated specifically to *Spartina. An inventory of the exotic (weed) flora of saltmarsh in Victoria is given in Appendix D.


non-vascular vegetation

Many non-vascular plants occur in coastal saltmarsh, although they are often inconspicuous, and many are poorly understood in terms of their ecological significance and interactions. Merely assessing their abundance and diversity is difficult, given the lack of taxonomic information or local expertise for many groups.

Algae are often diverse and diatoms, in particular, are likely to play a critical role in supporting food webs and stabilising sediments that is not reflected by their low biomass and inconspicuous nature (Sullivan & Currin 2000). Unfortunately the information on algae in Australian saltmarshes and mangroves is extremely fragmentary (Saenger et al. 1977). The multicellular Enteromorpha and Rhizoclonium are common on sediment surfaces, and in some cases Hormosira banksii may be found growing among mangroves (Adam 1994). In North American coastal saltmarsh, a diverse suite of benthic algae has been recorded on bottom sediments and on plant surfaces, including motile pennate diatoms (e.g. Navicula, Nitzschia and Amphora) and diverse cyanobacteria, including mats of coccoid forms (e.g. Chroococcus, Anacystis), as well as colonial (e.g. Merismopedia) and filamentous forms, both with and without heterocysts (e.g. Anabaena, Nostoc, Calothrix, Oscillatoria, Lyngbya, Microcoleus, Phordium and Schizothrix). Further details on the great diversity of algae found in North American saltmarshes are available in Sullivan & Currin (2000).

There is also extremely limited knowledge of saltmarsh fungi and lichens, but such taxa are known to exist. For example, numerous lichen taxa grow on the branches of Tecticornia shrubs and the leaves of Avicennia marina. It is not known whether mycorrhizae occur in mangroves or coastal saltmarsh of southern Australia, but vesicular-arbuscular mycorrhizae are known from saltmarshes of tropical Australia and have been reported on Samolus repens (Adam 1994, 2009a). We are aware of only one survey of lichens in coastal saltmarsh: Filson (1980) reported a moderately species-rich suite of lichens, particularly on Tecticornia spp., basalt rocks and artificial structures within saltmarsh at Point Wilson on the western shores of Port Phillip Bay.

Mosses are known to be common in Arctic and Scandinavian saltmarshes, as well as locally common in marshes of the United Kingdom (Adam 1990). Current records of Australian saltmarshes indicate few or no bryophytes, except for the mosses Funaria salsicola and Pottia drummondii and the liverworts Riella halophila and Carrpos sphaerocarpus, which are found on saline flats (Catchside 1980) and are perhaps the most halophytic of all bryophytes (Paul Adam, University of New South Wales, pers. comm.). Tortula papillosea was reported by Bridgewater (1975) to occur as an epiphyte on Tecticornia (Sclerostegia) arbuscula.

There are no ferns in saltmarsh or mangrove vegetation in Victoria, which seems also to be the case globally. Interestingly, this has not always been the case and the fossil record is rich in ferns that apparently grew in intertidal environments (Adam 1990). Note that the fern Acrosticum aureum is common in tropical and subtropical mangroves around the world, including in Queensland (Paul Adam, University of New South Wales, pers. comm.; see also Bunt 1982).


1.6 Zonation,sedimentationandsuccessionalchange

zonation in victorian coastal saltmarsh

A conspicuous feature of many coastal wetlands is the strong zonation in vascular plant communities; indeed nearly three-quarters of a century ago, Pidgeon (1940) reported on spatial zonation in mangroves and saltmarshes along the New South Wales central coast. The zonation in vegetation is most evident when the full tidal range, from subtidal to truly terrestrial plant communities, is visible (Figure 1.16).

Figure1.16: Zonation of plant communities at Rhyll Inlet, Phillip Island, showing a progression of mangroves (Avicennia marina) on the seaward side, saltmarshes in the upper intertidal zone, Swamp Paperbark (Melaleuca ericifolia) on the terrestrial fringes and, behind them, the truly terrestrial flora dominated by Eucalyptus viminalis subsp. pryoriana.

Although the early studies of Victorian saltmarsh by Bird (1971), Bridgewater (1975) and Saenger et al. (1977) reported a clear elevational zonation of plant taxa over the full tidal range, elevation exerts strong control on the gradation of individual plant species within a saltmarsh not only at this gross scale but also at a range of much finer spatial scales. The lower elevations of Victorian coastal saltmarsh, for example, are typically dominated by taxa such as Sarcocornia quinqueflora, and the upper, more elevated parts of the marsh by a tussock grassland of Austrostipa stipoides, the sedge Gahnia filum and, where there are freshwater inputs, rushlands of Juncus kraussii (Figure 1.17).


Bridgewater (1975) provided what is probably the most detailed description of plant zonation within an Australian saltmarsh. Behind the most seaward zone of Avicennia marina, nine vegetation complexes in Western Port saltmarshes were identified on the basis of floristic and structural criteria:• Introduced *Spartina• Extensive Salicornia (now Sarcocornia) dominated zone• Extensive Arthrocnemum (now Tecticornia) dominated zone• Suaeda complex• Puccinellia complex• Juncus complex• Stipa (now Austrostipa) complex• Schoenus-Cotula complex• Melaleuca zone.

In some cases, where there was a clear and simple elevational gradient, the various zones followed a consistent pattern with distance from the sea; in other cases they were intermixed, depending on local changes in topography and drainage caused by minor depressions or raised areas (e.g. hummocks) or creeks and tidal runners. Similar results were reported in the classic studies of saltmarshes around Sydney by Clarke & Hannon (1967), who noted that ‘…zonation of the vegetation throughout most of the area is striking…but in localised patches no clear zonation is discernable’. Although elevational zonation is often striking, the presence of fine-scale features (tidal channel, shell mounds, etc.) can cause individual zones to become so small that they are below the resolution of most mapping scales. Moreover, zonation in coastal saltmarsh can be manifest at a scale far finer than vegetation patterning in most other systems, because of the strong physical gradients

Figure1.17: Idealised zonation of plant taxa with elevation and tidal influence at Corner Inlet: a) an idealised sequence with species characteristic of each zone; b) Avicennia marina encroaching into saltmarsh; c) erosion of saltmarsh back to the Melaleuca ericifolia zone; and d) a typical sequence in response to sea-level rise. Source: Vanderzee (1988, Figure 2).


that sort vegetation over a small scale (i.e. metres) in intertidal systems. Pratolongo et al. (2009, page 95) summarised the factors controlling plant zonation in coastal saltmarsh in the following terms:

The result of the interaction between hydrodynamics, elevations, and vegetation is to produce a shore-parallel zonation of plants which, in reality, is made more complex and spatially variable by the microtopography of the marsh surface.

biotic versus abiotic factors controlling plant zonation

Four spatial models have been proposed to explain how vegetation changes along environmental gradients, such as the gradient in elevation and physico-chemical conditions from the sea to land (Dale 1999). In the first model, distinct groups of species occur with sharp boundaries between each group; in this case the various taxa within a spatial group conform most closely to what is understood as a ‘plant community’. In the second, there can be sharp exclusion boundaries between the various species but there seems to be no natural groupings into plant communities as occurs in the first model. In the third, groupings can occur that are not exclusive, again with little evidence of plant community structure. Finally, there may be no obvious groupings or exclusion, and species then appear to behave more of less independently of each other.

As in other systems, the spatial distribution of organisms in saltmarsh is determined not only by their physiological requirements but also by biotic interactions with other organisms, especially competition and predation. This process gives rise to the ‘realised’ vs ‘fundamental’ niche (Hutchinson 1957). Competitive interactions are particularly well understood for plants in saltmarshes of the Atlantic coast of the USA (Nybakken 2001). Four vegetation zones have been recognised in these saltmarshes: an upper, Marsh Elder zone vegetated with Iva frutescens; a Black Rush zone with Juncus gerardii; a Salt Hay zone with Spartina patens; and the most seaward Cordgrass zone, vegetated with Spartina alterniflora (Nybakken 2001). The zonation is controlled by a suite of competitive interactions and different plant adaptations. At the highest elevations, Iva frutescens out-competes Juncus gerardii and displaces it to a lower elevation in the upper marsh. Juncus gerardii out-competes other saltmarsh taxa because its dense root mats preclude colonisation by rhizomatous species such as Distichlis spp. and Spartina alterniflora that have below-ground runners. In principle, Spartina patens could compete well with Juncus gerardii as it also has a dense root mat, but Juncus gerardii emerges from winter dormancy early in spring, before Spartina patens commences growth. Thus Juncus gerardii has a critical ‘head start’ in colonising areas that are theoretically available to both taxa. In turn Spartina patens can out-compete Spartina alterniflora, but it is limited to the zone immediately below Juncus gerardii because it cannot tolerate anoxia and high H2S concentrations, which are found in the most waterlogged sediments where Spartina alterniflora grows. A similar suite of competitive interactions has been reported for the zonation of Salicornia virginica and Arthrocnemum subterminalis in saltmarshes of southern California (Pennings & Callaway 1992).

In a particularly elegant set of manipulative experiments, Crain et al. (2004) investigated the relative importance of physical and biotic factors in controlling the distribution of plant species along salinity gradients in estuaries on the New England coast of the USA. Saltmarsh species transplanted to freshwater locations thrived in the absence of neighbouring plants, but were quickly suppressed when freshwater taxa were present. In contrast, freshwater taxa when transplanted to saline sites quickly died. Crain et al. (2004) concluded that the spatial segregation of plants along salinity gradients in estuarine wetlands was driven by


competitively superior freshwater taxa displacing halophytic taxa to saltmarshes, and the freshwater taxa being prevented from colonising saline sites by physical factors, especially high salinities.

The North American examples outlined above may allow some generalisations to be made about saltmarsh vegetation in south-eastern Australia. It seems that physical stressors generally set the seaward limit and competitive interactions the upper limit. The generalisation makes intuitive sense: in areas with benign conditions, success involves out-competing others; in harsh environments, the main task is survival. Thus species must either specialise to occupy a vacant, but extreme, niche alone, or compete with others in more habitable places. Bertness & Pennings (2000, page 42) concluded as follows:

…there is a trade-off in marsh plants between competitive ability and the ability to deal with physical stress. This trade-off typically leads to competitively subordinate marsh plants dominating physically stressful habitats, while their competitive dominants monopolize physically benign habitats…This translates into competitively subordinate plants living in regularly-flooded low marsh habitats, while competitively dominant plants monopolize less frequently-flooded, high-marsh habitats.

Aside from the early seminal research by Clarke & Hannon (1967, 1969, 1970, 1971), little work comparable to that undertaken in the USA has been done in Australian saltmarshes to disentangle the complexity of possible interactions among saltmarsh taxa and their relationships with edaphic conditions. Nonetheless there is some good, if generalised, understanding of the interacting roles that waterlogging and salinity play in controlling plant distributions in some other saltmarsh areas of southern Australia. Kirkpatrick & Harris (1999), for example, elucidated the relative role of drainage and salinity in controlling the distribution of plant taxa in Tasmanian coastal saltmarsh.

sedimentation

Saltmarshes and mangroves are usually perceived as developing on muddy, low-energy, tide-dominated coasts. The term ‘tidal flat’ describes the range of environments found on these coasts. The tidal flat consists of several distinct sedimentary and vegetation zones, typically shore-parallel and related to the duration of tidal submergence, tidal current regime (flood- or ebb-dominated) and wave energy, sediment source and grade and input of fresh water. Sediment source(s) can be a combination of offshore, alongshore, adjacent cliffs and river inputs (Woodroffe 2002). As tidal flats have zones with different inundation regimes (see Chapter 1.3), the genesis and character of sediments in each of the zones also differs (Amos 1995). Unlike sandy coasts, where sediment size becomes coarser landward, muddy coasts display a reverse gradation, with sand in the subtidal zone giving way to mud in the intertidal and supratidal zones (Woodroffe 2002). The muddy zones may, however, contain low sand or shell ridges resulting either from storm surge deposition or marking the former position of the shoreline.

The tidal waters that inundate coastal saltmarshes are often turbid, and vegetation along the least elevated parts of the marsh reduces tidal turbulence and flow rates, and allows suspended particles to settle out (Adam 1990). The small hummocks created around clumps of saltmarsh vegetation facilitate further sediment deposition, and vegetation either assists sediment accretion or inhibits sediment erosion (Long & Mason 1983). Similarly, mangrove pneumatophores are important in trapping and retaining sediment (Bird 1980b).


The vegetation of mangroves and saltmarshes, therefore, can facilitate sediment deposition and stability in four ways: • The fine mat of surface roots and deeper rhizomes bind sediment and prevent erosion.• Emergent shoots and roots (e.g. pneumatophores) decrease current velocities and encourage the

deposition of fine particles. • Plants add organic matter to sediments via primary production, thus building up the sediment surface.• Plant roots and rhizomes form a dense, sometimes impenetrable mat, which deters burrowing

invertebrates that would otherwise rework sediments.

Rapid rates of sedimentation in the lowest parts of a saltmarsh imply that saltmarshes may extend laterally (i.e. seawards), as well as accreting vertically. The widely held view of saltmarshes and mangroves as land builders, however, may not be as valid as is often assumed. Adam (1990) argued that saltmarshes were better seen as taking advantage of sites where sediment deposition was occurring already. In this view, the capacity of saltmarsh vegetation to stabilise sediment against subsequent erosion is the critical process, rather than any putative land-building ability. The process was explained by Morrisey (1995, pages 206–207):

Saltmarshes begin to form when sediment deposited by rivers or the sea…accumulates to heights above the average level of neap high tides. Under these circumstances, plants start to colonise the sediment, which their roots bind and stabilise. The aerial parts of the plants also retard the movement of water over the sediment, causing sediment to be deposited at an increasing rate. Thus, the mud- and sand-flat becomes higher until eventually it is no longer flooded by even the highest tides.

Some researchers, however, suggest that saltmarshes that become elevated because of sedimentation also become increasingly steeply sloping at their seaward margins, and therefore increasingly susceptible to fresh erosion. Models of hypothetical saltmarshes confirm that complex feedback patterns are possible, with saltmarsh terraces forming and collapsing over time (van de Koppel et al. 2005). Certainly, there are many cases along the Victorian coast where saltmarsh is being eroded (Figure 1.18).

Typically saltmarshes show a net accretion of sediment with each tide, although erosion may occur in localised spots and more generally during storms (Figure 1.18; see also Wolters et al. 2005 for examples in the United Kingdom). The two contrasting processes of deposition and erosion vary in time and space, which results in a spatially and temporally rich mosaic of transitory physiographical features in saltmarshes and mangroves.

Consistent with this spatial and temporal variability, rates of sediment accretion in four Northern Hemisphere saltmarshes have been reported to vary from 2 to 150 mm year–1 (Long & Mason 1983). In some cases (e.g. Spartina-dominated marshes in Connecticut, USA) there has been a clear impact of human settlement on deposition rates; in one saltmarsh 24 cm of sediment had been deposited over the period 1900 to the 1970s, in comparison with 10 cm in the preceding century (Long & Mason 1983).

Figure1.18: Erosion of coastal saltmarsh, Belfast Lough, Port Fairy.


There are few reports of changes in surface elevation in Australian saltmarshes, but the exhaustive study of Rogers et al. (2006) gave rates varying from -0.68 to +5.27 mm year–1 in various saltmarshes of south-eastern Australia. Rates for saltmarshes and mangroves at four sites in or near Western Port are shown in Table 1.10.

Table1.10: Mean (± standard errors) rates of change in surface elevation at four sites near Western Port. Source: Rogers et al. (2006, Table 3).

Site Vegetation type Change in surface elevation(mm year–1)

French Island Mangrove -2.13 + 1.66

Saltmarsh 5.27 + 0.96

Koo Wee Rup Mangrove -0.03 + 2.23

Saltmarsh -0.16 + 0.94

Quail Island Mangrove -2.60 + 2.07

Saltmarsh -0.68 + 1.18

Rhyll Mangrove 0.92 + 1.87

Saltmarsh 0.64 + 0.75

A commonly held view is that mangroves stabilise and build up coastal land. In his monograph The Coast of Victoria, Bird (1993, page 196), for example, argued that ‘The building of a mangrove-fringed salt marsh terrace around the northern shores of Westernport Bay during the past 6000 years was the outcome of vegetation colonising and stabilising foreshore areas as muddy sediment accreted’. Figure 1.19 shows the accompanying figure that outlined the proposed mechanism.

The role of mangroves as land stabilisers, however, has long been controversial (e.g. see Bird 1971, 1980b; Ellison 1998). Bird & Barson (1982) argued that a mangrove system can be considered stable if it occupies the same area of intertidal land, and unstable if its boundaries advance or retreat. In cases where there is a seaward advance, there is often evidence of landward senescence, due to either the replacement of mangroves with paperbark swamps or even terrestrial plant communities, or, under highly salinised conditions, saltmarshes or unvegetated saline flats. Colonisation by mangroves can occur quickly on newly deposited sediments, if permitted by the prevailing hydrological and wave exposure regimes. Such colonisation is likely to interact synergistically with reduced flows and increased sediment deposition, providing an additional feedback loop for the deposition of yet higher sediment loads in high runoff catchments and consequently greater opportunity for mangrove development and/or expansion within the tidal zone. Blasco et al. (1996) suggested that mangroves are reactive opportunists and rapidly colonise newly deposited sediment and in doing so help to consolidate that sediment and may promote further sedimentation. Spenceley (1997), however, argued that while mangrove pneumatophores may promote sedimentation under low-energy conditions, medium to high-energy conditions can cause eddy currents to develop which initiate localised scour and erosion.

Woodroffe (2002) showed that as mangroves and saltmarsh occur in a variety of geomorphological settings, including the estuarine, embayment, back-barrier and deltaic, they are subject to a wide range of morphodynamic processes. Backbarrier environments, such as Lake Reeve (eastern Victoria) are low-energy environments and are buffered from the extreme conditions of wave or flood that can be experienced in more exposed estuarine and embayment sites. They thus experience very different sedimentary regimes which influence the rate and type of sedimentation. Tidal creeks that cross the marsh form a network of channels


resembling a river system with tributary and distributary channels. At higher tides, channels overflow and the marsh becomes a floodplain receiving muddy sediment. Mini levee banks develop at the channel margins and impede the return flow of tides and contribute to the development of saline pools with little vegetation. Channel margins also slump and disintegrate, further contributing sediment for tidal distribution.

temporal change in saltmarsh: environmental and successional processes

Successionalmodelsforcoastalvegetation

The zonation that occurs from mangroves to saltmarsh to terrestrial vegetation along an elevational gradient (Figures 1.16, 1.17 & 1.19) has sometimes been interpreted as evidence for seral succession. Such an explanation assumes that the elevational pattern repeats in space a succession that actually takes place in time.

Figure1.19: Proposed mechanism for the evolution of the mangrove-saltmarsh terrace on the northern shores of Western Port: a) the sandy coast at the end of the Late Quaternary marine transgression; b) with Holocene mud accretion, a mangrove fringe begins to spread seawards; and c) as the muddy terrace is built up to mean high tide level, mangroves are displaced by saltmarsh, backed by swamp scrub vegetation. Source: Bird (1993, Figure 135).


The succession is proposed to commence with pioneer mangroves, which colonise bare mudflats. As sediment accumulates, the surface becomes elevated and less frequently inundated, a process which creates conditions suitable for invasion by less flood-tolerant species. Because sediments have become progressively elevated, the original pioneer species can move forward towards the sea. Eventually a balance is reached between accretion and erosion at the most seaward side of the mangrove swamp (Mitchell & Adam 1989b).

The strongest proponent for a seral succession in mangrove-saltmarsh-terrestrial vegetation was Chapman (1974), but the interpretation has a much longer history (e.g. see Saintilan et al. 2009). It was proposed, for example by Pidgeon (1940), to occur in mangrove and saltmarsh communities along the central New South Wales coast. The successional model commenced with saline mudflats, then progressed sequentially through Avicennia/Aegiceras mangrove forest, Sarcocornia/Suaeda saltmarsh, Sporobolus virginicus grassland, Juncus kraussii rush meadow, Baumea juncea sedge meadow, Casuarina glauca she-oak forest, Eucalyptus robusta mahogany forest and culminated in a mixed eucalypt forest. Two Melaleuca species (M. ericifolia and M. quinquenervia) could be involved as well.

Given our current understanding of geomorphological processes, however, some successionalist interpretations of saltmarsh now sound naïve (e.g. Pidgeon’s 1940 model commencing with saline mudflats and culminating in a mixed eucalypt forest). Adam (1994) strongly and rightly criticised that model, noting (among other objections) that it was difficult to see how seral succession could be driven by sediment accretion above the tidal limit. The more plausible and restricted case of whether vegetation within coastal saltmarsh exhibits a regular pattern of clear successional change is also controversial. Chapman (1941, 1974) argued that the spatial zonation of saltmarsh plants provided strong evidence for a clear successional pattern. Bird (1986) proposed such a process to operate in Western Port: he argued that mangroves controlled the pattern of sedimentation in near-shore areas and resulted in the building of a depositional terrace at the level of spring high tides, which was then progressively colonised by, and succeeded with, saltmarsh vegetation. Figure 1.19 showed the process as outlined in Bird (1993).

Unfortunately, evidence from temporal and spatial changes in saltmarsh vegetation fails to provide strong arguments for or against succession or seral development in saltmarsh (Saintilan et al. 2009). Some evidence from long-term observations suggests that boundaries between different plant communities in saltmarshes can be stable for decades, unless the marshes are subject to altered management practices, in which case vegetation boundaries can change very rapidly (Adam 1990; see also Rogers et al. 2006). Other evidence counters some of the proposed successionalist models. Adam (1990), for example, argued that the preferential invasion of the seaward side of saltmarshes by *Spartina did not reflect a successional process. Vanderzee (1988) provided an overview of the various models that have been proposed to account for the colonisation of bare mudflats by saltmarsh taxa, and applied them to the case of Corner Inlet. If a simple succession model from mangroves to saltmarshes to terrestrial vegetation were to hold, remnants of earlier mangrove material should be found at depth in sediments under existing saltmarsh vegetation. Morrisey (1995) argued that there was no palaeobotanical evidence to support this inference, at least from Australian coastal wetlands. Mitchell & Adam (1989b) similarly reported that sediment cores taken in saltmarshes around the Sydney region showed no indication of the previous occurrence of mangroves at sites that currently support saltmarsh vegetation. Adam (1994) did, however, cite the work of Herbert (1951), who recorded Avicennia stumps in saltmarshes around Brisbane. Moreover, Saintilan & Hashimoto (1999) detected mangrove root systems preserved


~0.3 m beneath current saltmarsh along the Hawkesbury River in central New South Wales. The roots were 500–1,700 years old, and their presence provided evidence for the replacement of mangroves by saltmarsh at that site. In the most recent study, Saintilan et al. (2009) reviewed the topic and could find palaeobotanical evidence for and against the proposition.

It is, of course, simplistic to pit successionist against non-successionist models of vegetation dynamics. Clearly, examples of succession exist (witness the colonisation of newly emerged volcanic islands) and, just as clearly, examples exist where patterning is caused by other factors (e.g. environmental stress: see Crain et al. 2004), or where any successional patterns are constantly made redundant by environmental changes. But successionist models do remind us that saltmarshes are dynamic; there is no doubt that saltmarshes can change with time.

Mangrove–saltmarshtransgressions

In comparison with the putative successional transition from mangrove to saltmarsh discussed above, it is very clear that many coastal areas in temperate Australia have experienced a converse shift in vegetation: the transgression of mangroves into saltmarsh (Saintilan & Williams 1999; Rogers et al. 2005). Using a suite of historical maps dating back to 1831 and photographs from the early 20th century, McLoughlin (1987) concluded that there had been a major period of expansion of mangroves along the Lane Cove River, near Sydney, from about 1900 to the 1940s. Mitchell & Adam (1989a) reported a similar encroachment of mangroves into saltmarsh at Towra Point at Botany Bay, Lime Kiln Bay and Mill Creek along the Georges River and at Cabbage Tree Basin, Port Hacking in New South Wales. Wilton (2001) likewise reported an encroachment of mangroves into saltmarsh at Careel Bay, in Pittwater near Sydney, and Haworth (2002) similar pattern for mangroves and saltmarshes at every site that was examined along the Georges River for the period 1930–1970. Evans & Williams (2001) reported a 79% loss of saltmarsh area and a 33% increase in mangrove area at Kurnell Peninsula in Botany Bay. Saintilan & Wilton (2001) found that the area of saltmarsh in Currambene Creek and Cararma Inlet ( Jervis Bay, southern New South Wales) had decreased by 53% and 35%, respectively. In the case of Currambene Creek the loss was due primarily by the landward encroachment by mangroves, whereas in Cararma Inlet it was due to the seaward encroachment by Melaleuca and Casuarina spp.

Saintilan & Wilton (2001) argued that sea-level rise could not explain the loss of saltmarsh in most of these cases, and proposed instead that altered patterns of freshwater and nutrient input to intertidal areas were responsible. Saintilan & Hashimoto (1999) commented on the recent landward transgression of mangroves into saltmarshes along the Hawkesbury River in central New South Wales. As noted above, these shifts are contrary to the vegetation dynamics that would be expected under a traditional seral interpretation of mangrove-saltmarsh succession. Similarly, Vanderzee (1988) reported that southerly parts of Corner Inlet in Victoria were subject to erosion, accompanied by the regression of the Sarcocornia quinqueflora zone, and the landward colonisation of mangroves (see Figure 1.16). He recorded a shoreline recession of up to 30 m over the 42 years for which aerial photographs were available. The erosion may be a consequence of relative sea-level rise as a result of the ongoing tectonic subsidence of Corner Inlet.


Otherfactorsdrivingvegetationchangeincoastalwetlands

One aspect not often considered in the analysis of vegetation change in coastal vegetation is that successional patterns may occur that are not related to the traditional view of mangroves and saltmarshes as land-builders, and thus not strongly predicated upon elevation. In the high marsh of North American saltmarshes, for example, a range of alternative successional patterns may develop according to the intensity of disturbance. In these wetlands, taxa such as Distichlis spicata and Salicornia europaea are present because periodic disturbance creates hypersaline conditions conducive to colonisation by the highly salt-tolerant Salicornia europaea and clonal Distichlis spicata that can resource-share using ramets located in less saline parts of the marsh (Nybakken 2001). In this example, the clonal characteristics of many wetland plants becomes important (see also Hatton et al. 2008). The simple pattern of elevational zonation and control of plant distributions by an interacting waterlogging-salinity regime could thus be supplemented by the inclusion of the more complex competitive interactions shown in Figure 1.20.

Figure1.20: Alternative successional patterns in bare areas of saltmarshes in the New England region of North America. Source: Nybakken (2001, Figure 8.25).


1.7 Waterandsaltrelations

water uptake by plants

All plants, both halophytes and glycophytes, obtain water by maintaining a lower (i.e. negative) water potential in their tissues than occurs in their surrounding environment (Etherington 1982). In order for plants to take up water, there must be a progressively more positive gradient in water potential from the soil or water environment to the atmosphere, so that water flows from the water or soil pore space into the cells in the roots of the plant, thence into cells in the leaves, and finally into the atmosphere through stomata. Saltmarsh and mangrove soils present particular difficulties for plants because the high concentrations of salt mean that water in the soil has a very low (i.e. highly negative) water potential and thus plants have difficulty in extracting water into their root tissues. Note that plants are incapable of actively moving water from the soil into their tissues: water has to flow passively along the water-potential gradient. Mangroves and saltmarsh plants must thus maintain a very low water potential, and they do this via a suite of specialist adaptations.

Water potential consists of three components: i) an osmotic or solute component contributed by dissolved salts; ii) a matric component contributed by the attraction of water to surfaces, especially clays; and iii) a pressure component, which can be either positive (i.e. hydraulic pressure) or negative (i.e. suction). Halophytic plants maintain a low water potential mostly by accumulating osmotically-active substances in their cells to lower the osmotic component of water potential. These substances, called osmotica, are usually either inorganic compounds, such as salt, or organic compounds, such as sugars or nitrogen-containing compounds such as amino acids and glycine betaine (Cain & Boon 1987; Boon & Cain 1988). The overall water potential of plant cells is a balance between the negative solute potential caused by the presence of these osmotically active substances and the (usually) positive pressure potential caused by turgor pressure. If the cells of saltmarsh plants lose so much water that they become flaccid, the pressure component of water potential falls to zero and the only factor driving water into the cells is the osmotic component.

In saltmarsh and mangrove soils, the solute component becomes progressively more negative as salinities increase. In order to maintain a lower (i.e. more negative) water potential, the plants must accumulate osmotica in their cells to continue to take up water. There is a limit to the amount of salts that even halophytic plants can accumulate since sodium and chloride are inhibitory to most plant enzymes. Many plants, therefore, accumulate organic compounds that do not interfere as greatly with physiological processes; the problem is, however, that there is a limit to how highly concentrated these organic compounds can become before they also start to affect cellular processes. In any case, the diversion of material to cellular osmotica means that carbon compounds are not available for other processes such as growth and reproduction. A diversion of materials, especially carbon and nitrogen, that could be used otherwise to support vegetative growth may be one of the reasons why saltmarsh plants are uncompetitive with the glycophytes that grow in terrestrial areas to the landward side of the marsh (e.g. see Crain et al. 2004).

Many halophytes accumulate organic compounds to maintain a suitably low water potential in the cells. These compounds contain large amounts of nitrogen, so dealing with osmotic stress creates an additional problem for the plant by having also to deal with nutrient limitations created by this additional demand for nitrogen (Cain & Boon 1987; Crain 2007). It is widely recognised that nitrogen is the element likely to be limiting plant growth in Northern Hemisphere saltmarshes (e.g. Valiela & Teal 1979) for Spartina-dominated marshes


in Massachusetts, USA; Nedwell (1982) and Abd. Aziz & Nedwell (1979, 1986a,b) for saltmarshes on the east coast of England). The topic of nutrient-control of plant growth and the effects of eutrophication of coastal saltmarsh is discussed later.

Saltmarsh plants can experience even more severe water-relation problems when the soils in which they grow dry out. Saltmarsh soils, especially in the ‘upper’ marsh inundated only by extreme tides, are often subject to more-or-less complete desiccation (see Chapter 1.5). Similarly, plants growing in the ‘dry’ saltmarshes of central Victoria, where the Mediterranean-type climate imposes severe climatological stresses on coastal vegetation, will be expected to commonly experience periods of extreme water stress. When soils dry out and become hypersaline, the matric component of water potential becomes increasingly important because water is retained on the surface of soil particles, especially clays, by capillary action. Also, salts become further concentrated in the soil profile, leading to even more severe osmotic and direct-ion stresses.

impacts of flooding on soils

Because of their intertidal position, the soils of mangroves and saltmarshes are episodically inundated with saline water and, less commonly, with fresh water. It is well known that flooding has a wide range of effects on soils (Ponnamperuna 1984). The replacement of air spaces with water decreases the rate at which oxygen can diffuse from the atmosphere into the soil. Soil becomes anoxic (i.e. free O2 is absent) if flooding is prolonged, since soil microbes (mainly bacteria and fungi) consume oxygen at a faster rate than it can diffuse from the atmosphere or from plant roots and rhizomes. The replacement of the air-filled spaces between soil particles by water after flooding also affects fluxes of the carbon dioxide produced by microbial, plant and animal respiration; the diffusive loss of carbon dioxide to the atmosphere also decreases, causing an accumulation of carbon dioxide in the soil and a resultant decrease in soil pH.

The loss of free O2 from soil pores causes a reduction in the redox potential. Changes in redox potential have wide-ranging effects on many biogeochemical processes. The changes occur because the sequence in which various redox reactions take place in flood-prone soil is controlled largely by their relative reduction potentials and the related free-energy change of their oxidation-reduction reactions (Boon 2006). When O2, the oxidant with the greatest free energy yield, is exhausted, oxidation continues using the next most efficient oxidant, and so on until the supply of organic material has been exhausted or possible oxidants are no longer available. In practical terms what this means is that aerobic (i.e. oxygen-mediated) decomposition is progressively replaced by anaerobic processes, such as fermentation, denitrification, sulfate reduction, and methanogenesis. Of these various anaerobic processes, the two of greatest importance in saltmarshes are denitrification and sulfate reduction. Denitrification is important because it results in the conversion of nitrate to nitrogen gas, causing a net loss of biologically available nitrogen from the ecosystem (Kemp et al. 1982). It has been long recognised that the limited availability of nitrogen frequently limits the growth of plants in estuarine systems. Sulfate reduction is critical in marine-influenced wetlands because it results in the production of hydrogen sulfide, a potent plant poison and, ultimately, in the possible production of acid-sulfate soils. The role played by sulfate in saltmarsh energy flows has been examined by a range of researchers, e.g. see Howarth & Teal (1980) for a seminal study.

Redox potential controls also the availability to plants of metals in soils. Iron and manganese, for example, are reduced to more soluble Fe2+ and Mn2+, respectively, from the largely insoluble Fe3+ and Mn4+ forms that


occur in aerated soils. Redox potential can also control the availability of phosphorus, since under aerobic conditions phosphorus is usually bound strongly to iron hydroxides and oxyhydroxides. Under anoxic conditions, free orthophosphate can become available for uptake by plants and/or diffusion from sediment pore-waters into overlying or tidal flood waters. As noted above, redox controls nitrogen availability, because biologically available forms of nitrogen such as nitrate are converted by denitrifying bacteria to nitrogen gas under anoxic conditions (Boon 2006).

Soil flooding affects also the production of toxic compounds and plant hormones (Etherington 1982; Colmer et al. 2008). The production of toxic hydrogen sulfide during bacterial sulfate reduction has been mentioned earlier. The production of reduced forms of metals, especially the more soluble, reduced forms of iron and manganese, may inhibit plant growth if concentrations are too high. Finally, a range of organic compounds are produced in response to soil anoxia, such as ethanol and cyanogenic (i.e. cyanide-producing) compounds, which may be plant toxins. Some of these compounds (e.g. ethylene) are produced by plants undertaking aerobic respiration; others (e.g. hydrogen sulfide) as a consequence of microbial activity. Of the organic compounds, ethylene is perhaps most important because it is a potent plant hormone (Devlin 1975). Ethylene is involved, for example, in the ripening of fruits and the development of geotropism in seedlings, as well as being a potent inhibitor of bud growth and thus in the development of apical dominance.

The combination of these factors make the saltmarsh environment extremely hostile to plant growth. The ecological consequences are clear: all species growing in saltmarsh must specialise, and specialisation involves evolutionary change and comes at a competitive cost. Hence saltmarshes are dominated by few species from few families, and often contain species that do not grow elsewhere (Chapter 1.5).

impacts of flooding on plants

Flooding during the growing season affects almost all aspects of the physiology and growth of plants (Kozlowski 1997; Colmer & Flowers 2008). Some of the most important effects on flood-intolerant species are:• Inhibition of seed germination and seedling establishment• Inhibition of vegetative growth• Changes in plant morphology and anatomy• Promotion of early senescence and mortality.

These negative impacts on plants seem ironic in the saltmarsh context: an environment that is inherently hostile because of high concentrations of salt causing physiological water deficits is, in fact, often flooded with water.

The injury and inhibition of growth that occurs to flood-intolerant plants, or to flood-tolerant plants that have been inundated for too long, results from a wide range of physiological dysfunctions (Kozlowski 1997). It is difficult to unravel each of the processes because flooding affects almost all aspects of plant physiology, including carbohydrate, protein, organic-acid and lipid metabolism. Flooding typically also decreases the uptake of the major plant macronutrients, especially nitrogen, phosphorus and potassium. The loss of soil nitrate, via denitrification, following soil waterlogging has been mentioned above. A decrease in the uptake of nutrients may be a result also of the death of fine roots and mycorrhizae under the anaerobic conditions created in flooded soils. In contrast, the uptake of iron may increase after soils are flooded, because the


reduced, ferrous form of iron (i.e. Fe2+) is more soluble than the oxidised, ferric form (i.e. Fe3+). Under some circumstances, the reduced forms of iron and manganese may be so biologically available that they become toxic to plants experiencing prolonged flooding. A wide range of phytotoxic compounds also may be responsible for poor plant growth after flooding; the accumulation of organic compounds, such as ethanol, organic acids and aldehydes, during anaerobic plant respiration can cause injury to flooded plants. Soil microbes, especially sulfate-reducing bacteria, also may produce plant toxins: the generation of hydrogen sulfide has been mentioned earlier.

Flood-tolerant plants possess a number of adaptations that allow them to survive periodic inundation. The most important adaptation for many wetland plants is that they can aerate their rhizosphere (i.e. root zone) and thereby maintain aerobic conditions around the roots and rhizomes, even in permanently flooded soils (e.g. see Sorrell et al. 1994). It is unclear which taxa of saltmarsh plants can aerate their rhizosphere effectively; some emergent wetland plants, such as Phragmites australis, are excellent root aerators whereas other taxa, such as Melaleuca ericifolia, are incapable of aerating their rhizosphere (Morris et al. 2008). The ability of mangroves, especially Avicennia marina, to aerate their rhizosphere is well known (Saenger 1982; Clough et al. 1982). The pneumatophores (breathing roots) of Avicennia marina are one such adaptation for life in waterlogged, anoxic sediment (Figure 1.21). Other than such broad generalisations, however, it is difficult to be more specific about the nature of responses to flooding, as little is known of the response or tolerance of halophytic plants to sediment anoxia (Colmer & Flowers 2008).

Figure1.21: Partly submerged aerial breathing roots (pneumatophores) of Avicennia marina, Western Port.


impacts of salinity on plants

Dealing with the adsorption of water is only part of the problem that saltmarsh and mangrove plants face by living in a saline environment; there is also the issue of dealing with salt. The stresses are exacerbated even further in the case of halophytes, which must cope simultaneously with low physiological water availability, high salinity and waterlogged and anoxic sediments (Colmer & Flowers 2008).

Even when salt is excluded at the roots (e.g. as occurs in mangroves: see Clough et al. 1982), some inevitably finds its way into the xylem and ultimately into above-ground organs. The issue for the plant is that when salt moves into photosynthetic organs, such as leaves, primary production is usually inhibited. Salinity exerts an inhibitory effect on plants in two main ways (Bailey et al. 2006):• General osmotic effects, whereby dissolved salts decrease the availability of water to plant roots• Specific toxic effects, arising mostly from the accumulation of sodium and chloride by plant tissues, and

to a lesser extent by the disruption of the uptake of other essential ions (e.g. calcium and magnesium) in the presence of high external salt concentrations.

The response of plants to salinity varies greatly during their life cycle. Kozlowski (1997) concluded that woody plants are usually relatively salt tolerant during seed germination, much more sensitive during seedling emergence and establishment, and progressively more salt tolerant as plants mature. There is good evidence, however, that salt tolerance within a single species can vary from region to region and from genotype to genotype (e.g. see Hanganu et al. 1999 and Zhao et al. 1999 for Phragmites australis). As with responses to flooding, there is a wide range of morphological and physiological responses to exposure to salinity by woody glycophytes. In terms of visually obvious injury, toxic effects appear first as chloride toxicity: leaf margins become chlorotic, then leaf blades become scorched. Sodium toxicity becomes progressively apparent as leaf mottling and the creation of necrotic patches; eventually the leaves are shed and the twigs die back (Kozlowski 1997). Metabolic impacts may be evident as well, often as suppression of leaf initiation and expansion, and accelerated rates of leaf abscission from salt-affected plants. Leaves may become thicker in response to salinity, as can epidermal cell walls and cuticle (Longstreth & Nobel 1979; Saenger 1982).

Salinity results also in a wide range of physiological responses. The accumulation of high concentrations of salt in salt-affected tissues results in enzyme inhibition, with consequent effects of photosynthesis, uptake of nutrients, and protein and carbohydrate metabolism. The combination of physical and metabolic impacts (e.g. stomatal closure due to water stress, combined with disruptions to enzyme activity) results in salt-affected plants showing decreased rates of primary production compared with those growing under freshwater conditions.

Different species of halophytes have different mechanisms of dealing with salt. Salt glands, which allow the excretion of excess salt, are known in several Australian taxa, notably Frankenia spp., Limnonium spp. Samolus repens, S. junceus and Sporobolus virginicus (Adam 2009a); studies have been made also of salt secretion by salt glands on the leaves of Australian Avicennia marina (Boon & Allaway 1982, 1986). Other taxa excrete salt via epidermal trichomes (e.g. Atriplex spp: Winter et al. 1981) and some partition it into non-photosynthetic organs such as the stem, bark or older leaves (Saenger 1982). Waisel (1972) reported that European Salicornia spp. (close relatives of Australian Sarcocornia spp.) translocated salt back to the roots via the phloem, and Love (1981) reported that Juncus shed their stems when salt contents increase to an excessive level; some


Salicornia spp. behave similarly with respect to leaf shedding. Many taxa of saltmarsh plants are succulent, and succulence is generally thought to be an adaptation whereby salts that have not been excluded at the roots, or cannot be excreted by salt glands, are partitioned within parts of the leaves where they do little physiological damage. Leaf succulence seems to be a response to the accumulation of sodium in the leaf and is caused by enlargement of cells in the hypodermal and mesophyll cells, at least in mangroves (Clough et al. 1982). The relationship between succulence, sodium, and other physiological processes such as photosynthesis, however, is complex ( Jennings 1968).

Salt is not, however, always negative in its effects on halophytes. Some plant species have a physiological requirement for sodium, especially among those with the C4 photosynthetic pathway. Atriplex vesicaria was the first such plant to be discovered (Brownell 1979; Adam 1993), and it is possible that other Atriplex species have the same requirement. Furthermore, sodium at concentrations of up to 50–200 mM NaCl induces positive growth responses in most halophytes (Flowers et al. 1986; O’Leary 1995).

1.8 Reproduction

sexual versus asexual reproduction

Flowering plants can reproduce via either sexual or asexual means. Sexual reproduction involves the exchange of gametes (i.e. the movement of pollen from flowers on one plant to the stigmas of flowers on the same or another plant) and the production and dispersal of seeds, their germination, and eventual recruitment of young plants into the population. In contrast, asexual reproduction involves clonal (or vegetative) expansion and dispersal of plants via the production of an assortment of organs and life strategies, including rhizomes, stolons, asexually produced seeds, tubers, turions and plantlets, or fragmentation of the plant body. Because of the wide range of mechanisms for vegetative expansion and dispersal, clonal plants can spread across large areas as a population of semi-independent parts of what was originally a single seedling. In these cases, the visually obvious ‘plants’ are called ramets and are components of a single, but larger, and genetically homogeneous spatial unit called the genet (Harper 1977). As noted earlier for American Distichlis spicata, the ability of clonal plants to resource-share in heterogeneous environments can have significant ramifications for interspecies competition (Nybakken 2001).

It has been shown repeatedly in freshwater and brackish-water wetlands that recruitment from seed is irregular and highly unpredictable, and the success of germination is often dictated by subtle changes in hydrological conditions (Britton & Brock 1994; Rea & Ganf 1994; Kellogg et al. 2003; Robinson et al. 2006). These factors mean that conditions suitable for sexual recruitment for wetland plants may occur only at few times of the year or, in the case of wetlands with variable hydrological regimes, only in a few years every decade. In contrast, the period suitable for plants to grow and propagate vegetatively, though still largely dependent on the presence of water, is not restricted to a single season or a short time of year that has the requisite water and salinity regimes. Moreover, as seed production is largely a function of biomass, long-term persistence through clonal offspring offers a greater probability of continual seed output, so that viable seed will be present within the wetland system if suitable conditions for germination do occur (Warwick & Brock 2003).

Many studies from across the world have shown that freshwater and some brackish-water wetlands are often dominated by clonal plants (e.g. see Mühlberg 1982; Cook 1990; Grace 1993; Oborny & Bartha


1995; van Groenendael et al. 1997; Crow & Barre Hellquist 2000; Hatton et al. 2008). Clonality may be an important mechanism for colonisation and spread in saltmarshes too: examples of taxa that can spread clonally in Australian estuarine and brackish wetlands include Baumea, Bolboschoenus, Distichlis, Juncus, Phragmites, Schoenoplectus and Sporobolus. Neverthless, almost all the dominant species in coastal saltmarsh (e.g. Sarcocornia, Tecticornia, Gahnia, Austrostipa) reproduce primarily from dispersive (and presumably sexually generated) propagules. The ecological causes and consequences of a ‘split’ between sexual and clonal reproduction in coastal wetlands remains largely unstudied. They may relate to the extreme dynamism of intertidal systems, which may discourage the growth of extensive ramets. Sexual reproduction may allow better colonisation of newly available substrate. Alternatively, the pattern may be an indirect consequence of the evolutionary origins of the taxa that occupy Victorian saltmarshes for other reasons.

sexual reproduction in saltmarsh plants

Adam (1990) noted that the level of understanding of sexual reproduction in saltmarsh plants, from a global perspective, varies from very little (for processes such as pollination and seed dispersal) to a great deal (for processes such as germination and seedling establishment). Sadly, little information has been obtained for Victorian, or even south-eastern Australian, saltmarsh plants.

Pollination

Pollination in the Chenopodiaceae, although generally thought to be by wind (anemophily), has been shown in many cases to be mediated by insects (entomophily) (Blackwell & Howell 1981). In Western Australia, Keighery (1979) documented Honey Bees Apis mellifera and syrphid flies collecting nectar on Sueada australis and effecting pollination. He regarded that species as being primarily wind pollinated. Nevertheless, some information on various aspects of sexual recruitment in saltmarsh plants can be gleaned from simple observations of the plants themselves: for example the floral morphology of Tecticornia halocnemoides s.l. suggest strongly that it is wind pollinated (Figure 1.22).

Figure1.22: Tecticornia halocnemoides in female phase of flowering with exerted stigmas, The Spit Nature Conservation Reserve.


Although scant information is available on the pollination of the Victorian or Australian saltmarsh and mangrove flora, extrapolations for other species can be made from observations of the morphology and hence floral biology/pollination syndromes can be inferred (Faegri & van der Pijl 1979; see also Figures 1.23 & 1.24). On these bases, the dominant pollination syndrome seems to be anemophily; almost all members of the Centrolepidaceae, Chenopodiaceae, Cyperaceae, Juncaceae, Juncaginaceae and Poaceae are thus pollinated, regardless of whether they are saltmarsh or non-saltmarsh taxa.

Figure1.23: Frankenia pauciflora, Disphyma crassifolium and Distichlis distichophylla, The Spit Nature Conservation Reserve.

Figure1.24: Flowering Disphyma crassifolium with Distichlis distichophylla, Point Lonsdale.

The remainder of the saltmarsh flora is presumably entomophilous, unless the species are predominantly self-pollinating, which seems likely to be the case for many of the annuals of upper saltmarsh, given their small or very small flowers and high seed-set (e.g. Apium annuum, Brachyscome perpusilla, Cotula vulgaris var. australasica, Daucus glochidiatus, Gnaphalium indutum, Hornungia procumbens, Hydrocotyle capillaris, H. medicaginoides, Sebaea albidiflora and Sonchus hydrophilus). Species with a pronounced floral display with larger flowers or inflorescences, and those with floral fragrance (e.g. Avicennia marina, Cuscuta tasmanica, Frankenia pauciflora var. gunnii, Samolus repens, Selliera radicans and Wilsonia spp.; Geoff Carr, pers. obs.) are likely to be entomophilous. These flowers offer insects a floral reward: nectar, pollen or both, and they are likely to attract a range of generalist pollinators. Recent observations at Ocean Grove (Geoff Carr, March & April 2010) indicate that Avicennia marina, which potentially offers pollen and nectar rewards from its


fragrant flowers, is pollinated by a suite of small to large flies (Diptera), several wasp species, solitary bees, and the Saltbush Blue Butterfly Theclinesthes serpentata serpentata. All insect species were consuming nectar.

The Museum of Victoria Bioinformatics website has published relevant data on native bees that have been recorded on a suite of saltmarsh plants (see www.museumofvictoria.com.au/bioinformatics). Collection of insects on particular flowers does not necessarily imply that the insect species is an effective pollinator of the plant species, but it seems very likely that they are. The records – plants and insects – are as follows: Angianthus preissianus – Colletid bees Neopasiphae simplicior (WA); Disphyma crassifolium ssp. clavellatum – Halictid bees Lasioglossum erythrurum, L. veronicae and L. mesembryanthemi (SA); Frankenia pauciflora var. pauciflora – Anthophorine Apid bees Amegilla murrayensis and Thyreus waroonensis (WA), Megachilid bee Megachile chrysopyga (WA); Limonium spp. – Anthophorine Apid bees Amegilla murrayensis (WA), Halictid bees Lasioglossum cognatum (SA), L: sp-chi (WA); Megachilid bees Megachile captionis (WA), M. macularis (WA); Muehlenbeckia florulenta – Halictid bees Lasioglossum cognatum, L. veronicea (SA); Samolus repens var. repens – Halictid bees Lasioglossum cognatum (SA).

Other Victorian saltmarsh genera or species that are assumed or known to be entomophilous, based on floral structure, are Avicennia marina, Cuscuta tasmanica, Hemichroa pentandra, Lawrencia spicata, Lawrencia squamata, Lobelia irrigua, Mimulus repens, Schenkia australis, Selliera radicans, Senecio halophilus (perhaps also self-pollinating), Spergularia spp. (perhaps also self-pollinating) and various Wilsonia species. Wilsonia humilis and Wilsonia rotundifolia are often visited by small flies, and Selliera radicans is visited by flies and solitary bees. It is likely that butterflies have a role in pollination of entomophilous upper saltmarsh plants, particularly the Painted Lady Vanessa kershawi and the threatened Altona Skipper Hesperilla flavescens, which is dependent on Gahnia filum as a larval food plant (Braby 2000; Crosby 1990). In South Australia, studies of pollination in Frankenia pauciflora have recorded Syrphid flies visiting flowers (Molly Whalen, Flinders University, pers. comm.).

Observations by Geoff Carr at The Spit Nature Conservation Reserve on the western shores of Port Phillip Bay (November 2009 to January 2010) revealed several insect species visiting flowers of Frankenia pauciflora, which was flowering prolifically at the time. The most abundant – exotic Honey Bees – were avidly collecting pollen and nectar; also common were the exotic Cabbage White (Butterfly) Pieris rapae rapae feeding on nectar, while occasional visitors were a native Halictid bee (pollen and nectar) and the Saltbush Blue (Butterfly) Theclinesthes serpentata serpentata (feeding on nectar). The latter species’ larval food plants are Atriplex spp. (Braby 2000). These observations indicate a generalised entomophilous pollination syndrome for Frankenia pauciflora and that native solitary bees and butterflies are involved in pollination. There were no other entomophilous plant species flowering in the saltmarsh at the time, only wind-pollinated Chenopodiaceae (Sarcocornia quinqueflora) and *Plantago coronopus.

Very little seems to be known about the biology and ecology of the saltmarsh flora that depends on invertebrates for pollination, or on any possible co-dependence between the flora and insects. It appears that entomophily in the saltmarsh flora is most likely of a generalist nature, but that may not be true for all plant species. Nor is it known if pollination biology aspects of saltmarsh function have been compromised by destruction of adjoining non-saltmarsh vegetation (e.g. by stock grazing, and residential and industrial development); some pollinators may be dependent on the intactness of both systems for their survival. Disruption of ecosystem function may also have arisen because of the advent of exotic pollinators (e.g. Apis


mellifera) and an exotic entomophilous flora in or near saltmarshes; consequences for different species in the indigenous flora and invertebrate pollinating fauna may be positive or negative because of the exotic biota.

Seedgerminationandseedlingestablishment

One aspect of the sexual reproduction of saltmarsh plants, however, has been reasonably well studied: environmental conditions required for seed germination and seedling establishment (Ungar 1978). A number of conclusions can be drawn from the diverse but patchy studies on this topic.

First, the seed of halophytic saltmarsh plants can generally survive temporary, and in some cases even prolonged, exposure to seawater. Boorman (1968), for example, showed that seeds of Limonium vulgare could survive at least 21 weeks of immersion in seawater. In contrast, seeds from glycophytes are usually killed by exposure to salt water. Indeed incubation of seed in seawater may enhance germination, as it is similar to osmotic conditioning (seed priming), a horticultural technique that improves germination rate and percentage in many vegetable and flower crops. Although germination in some, perhaps many, succulent halophytes is stimulated by NaCl (Ungar 1962), most plants, including halophytes, have reduced and/or delayed germination when subject to constant salinity above 1% w/v NaCl (Chapman 1976; Sen & Mohammed 1994). The effect is mainly due to osmotic pressure preventing imbibition, rather than direct-ion toxicity (Osmond et al. 1980), but effects of the mobilisation of food reserves by NaCl may contribute as well (Iyengar & Reddy 1994). Once the osmotic stress is removed, germination is enhanced in many species (Khan 1992). Seeds in priming solution become partially rehydrated, allowing metabolic processes to commence but preventing radical elongation and visible germination (Taylor & Harman 1990). Several inorganic salts including NaCl are used in seed priming (Khan 1992; Chong & Bible 1995) and primed seeds may germinate more rapidly and uniformly under a wider range of temperatures than untreated seeds (Chong & Bible 1995). Osmotic conditioning also occurs in soil. Seeds restricted from germinating by moderate to high soil salt concentrations remain viable for extended periods and may show higher germination rates and percentages when salinity is reduced (Boorman 1968; Khan 1992).

Second, despite an ability to withstand at least some exposure to salty water, seed from almost all saltmarsh and related taxa that have been studied to date show an improved rate of germination in freshwater conditions or after temporary exposure to freshwater (Ungar 1978). Naidoo & Kift (2006), for example, showed that seed of Juncus kraussii in South Africa germinated best at salinities < 20% seawater. Similarly, Greenwood & MacFarlane (2006) showed that *Juncus acutus and Juncus kraussii from the Hunter River region of New South Wales germinated best in fresh water and failed to germinate in seawater. Indeed, seed from many species of saltmarsh plants is incapable of germinating in salinities that commonly exist in the field, leading to the suggestion that episodic reductions in salinity are required for many species to recruit sexually (Adam 1990). Seed from some species of Salicornia spp., however, is particularly tolerant of high salinity and can readily germinate in salinities that exceed those of seawater.

Third, it is possible also that periods of high salinity are required for some species of annual halophytes to flower, although the same species may require fresher conditions to germinate. Kingsbury et al. (1976), for example, reported that reproductive activity in Lasthenia glabrata increased markedly within weeks of exposure to salinity stress, and that lengthy exposure to high salinity caused a shift towards reproductive development over vegetative growth.


Less well understood is the effect of inundation on seed germination. Such a knowledge gap is especially unfortunate, since saltmarshes are characterised not only by episodically high salinity but also by poor drainage and waterlogging: many landward areas of coastal saltmarsh remain flooded even after rainfall has lowered salinities to the point where seed germination is no longer inhibited by high salinity. Seed from some species of saltmarsh plants can germinate under water (Clarke & Hannon 1970) and in some cases, such as Juncus spp., germination is promoted by inundation (Rozema 1975). It may be that seed from some other saltmarsh taxa shows a similar response to flooding as that of Melaleuca ericifolia, seeds of which do not germinate other than on moist – but not flooded – substrata (Ladiges et al. 1981; Robinson et al. 2006).

Phenologyofsexualrecruitment

The phenology of recruitment by seed also appears to be poorly studied in Australia saltmarshes, even though it has obvious and potentially important ramifications for the dynamics of vegetation patterning in time and space, as well as the fortunes of populations of individual plant species. The indigenous and exotic saltmarsh flora would seem to be largely partitioned between those species that recuit from seed in spring-summer, versus those that recruit in autumn-winter (Geoff Carr, pers. obs.). A smaller subset is likely to be much more opportunistic in their germination responses, perhaps especially in the perennial and annual exotic flora.

The cool-season flora, germinating following the autumn break, includes all indigenous annual species, including the submerged aquatic Lepilaena cylindricarpa and excepting the annual chenopods Atriplex australisica and Chenopodium glaucum. Cool-season germinating perennials, notably the structural dominants/co-dominants Disphyma crassifolium and Frankenia pauciflora, generally exhibit massed recruitment in autumn. By contrast, all the indigenous chenopods (e.g. Atriplex, Sarcocornia, Suaeda and Tecticornia) of major structural and functional significance in Victorian saltmarshes appear to be warm-season recruiters, germinating in spring and summer. Heavy summer rain may stimulate massed recruitment in these perennials. This warm-season recruitment phenology in the indigenous perennial Chenopodiaceae is precisely mirrored in the exotic annual chenopod flora (e.g. Atriplex, Chenopodium and Suaeda). In the indigenous saltmarsh grass flora, there appears to be a division between C3 autum-winter recruters (e.g. Puccinellia) and C4 spring-summer recruiters (e.g. Distichlis distichophylla, Eragrostis parriflora, Sporobolus virginicus and Zoysia macrantha). A similar pattern is evident in the exotic Poaceae between C3 and C4 grasses, whether they be annuals or perennials.

Seedbanks

The role of seedbanks in saltmarshes deserves particular attention because of the possible importance of buried seed in allowing saltmarsh plants to re-establish after disturbance, especially after the relief of hypersaline conditions by heavy rainfall or freshwater flooding. Many studies of the vegetation of freshwater wetlands, for example, have stressed the role of sediment-based seedbanks in the recovery of the plant community, especially after water-level manipulations (Brock & Britton 1995; Brock & Casanova 2000; Capone & Brock 2006). Adam (1990) concluded that little was known of the seedbank of saltmarshes, with the exception of studies on Salicornia communities in British and North American saltmarshes: most of these studies have reported low numbers of viable seed and low plant diversity in soil-seed banks.


Seeddispersal

The extensive literature on seed dispersal of saltmarsh species in the Northern Hemisphere has documented various dispersal syndromes and mechanisms. There are clear implications for the evolutionary and functional significance of these traits in saltmarsh floras, which suggests that they are present in the Australian saltmarsh flora, though we are unaware of such studies to confirm dispersal mechanisms here. European and North American studies on seed dispersal in relation to the saltmarsh flora include dispersal by tidal movement (e.g. Bakker et al. 1955; Chang et al. 2007, 2008; Huiskes et al. 1995) and dispersal externally (epizoochory) and internally (endozoochlory) by birds (waterfowl) (e.g. Vivian-Smith & Stiles 1994; Figuerola & Green 2002; Chang et al. 2005).

Major saltmarsh species in south-eastern Australia, such as Sarcocornia quinqueflora, are likely to be dispersed by tides, and in this species masses of buoyant seeds can be seen deposited on bare saltmarsh substrates by receding tides (Geoff Carr, pers. obs.). Sarcocornia can be dispersed also as newly germinated seedlings, which might survive if carried by tides into moist areas rather than into the (drier) drift line (Paul Adam, University of New South Wales, pers. comm.). The occurrence of saltmarsh species (including Sarcocornia quinqueflora and Sebaea albidiflora – Geoff Carr, pers. obs.) as colonisers of secondary salinisation areas (Allen 2007), remote from naturally saline sites carrying potential source populations, indicate bird dispersal (probable epizoochory rather than endozoochory) (Geoff Carr, in Kinhill Stearns 1986, pages 10–16). Further studies are required on this important topic, especially in view of the fragmentation of saltmarshes and the consequences of different dispersal mechanisms for colonisation of new intertidal areas following climate change.

reproduction in mangroves

As a group, mangroves commonly exhibit a strange and wonderful suite of reproductive strategies, and Avicennia marina is no exception to this pattern (Saenger 1982). Avicennia marina reproduces by seed and is incapable of clonal growth. Harty (1997) reported that, in Victoria, Avicennia marina flowered from May to June, in contrast to the summer flowering of Avicennia marina at lower latitudes along the Australian coast. At Ocean Grove, however, this species has been observed to flower in February to April (Geoff Carr, pers. obs.) It is thought that bees pollinate Avicennia marina flowers (Clifford & Specht 1979); certainly the flowers are fragrant and attractive to bees and, as described earlier, a range of flies, wasps and solitary bees have been observed visiting mangrove flowers at Ocean Grove. In some cases mangroves are commercially valuable honey producers (Lear & Turner 1977), but their economic role in this capacity in Victoria is probably very limited (see Chapter 1.10).

The fruit of Avicennia marina is a fleshy capsule with a single seed. In many mangroves, the fruits contain seeds which develop precociously. In other words, the seed germinates while still attached to the parent tree and the embryo develops into a seedling without any dormant period. In some mangrove taxa (e.g. Rhizophora), the embryo ruptures the pericarp and grows beyond it into a substantial seedling while still attached to the tree; in Avicennia, however, the embryo does not enlarge sufficiently to rupture the pericarp (Figure 1.25). Saenger (1982) termed these sorts of mangroves ‘cryto-viviparous’, in comparison with the viviparity shown by Rhizophora. Viviparity and cryto-viviparity are thought to be adaptations to allow improved buoyancy for wider dispersal, rapid establishment of the seedling once it settles to the substratum, and an early development of salt and water balance within the young plant. The propagules of Avicennia


marina are initially buoyant and well adapted to dispersal by water (Clarke & Myerscough 1991; Clarke 1993). There is some evidence that initial dispersal occurs during low tides (Clarke & Hannon 1969). The propagules of Avicennia marina sink soon after losing the pericarp. The shedding of the pericarp is controlled by the salinity of the water, so propagules in brackish water disperse less than those in higher or lower salinities (Saenger 1982). Clarke & Hannon (1969) reported that it took about five days for Avicennia marina propagules to root once they had settled onto a sediment surface.

Figure1.25: Young plants and germinants of Avicennia marina.

1.9 Ecologicalvalues

primary production

The central element of many claims for the exceptional ecological value of coastal saltmarsh, especially in the older literature, is that saltmarsh plants are highly productive (e.g. see Kormondy 1976). The claims are based almost entirely on one type of saltmarsh: Spartina-dominated systems of the mid North American Atlantic coast (Day et al. 1989). The problem is that generalisations about high rates of primary production are unlikely to hold for all coastal saltmarshes, and the claims are especially not likely to be true for the taxonomically and structurally different saltmarsh of south-eastern Australia. As noted in Chapter 1.5, Victorian coastal saltmarsh is not dominated by the highly productive grasses that dominate saltmarshes along the mid Atlantic coast of the USA. Moreover, although the early studies of North American Spartina-dominated saltmarshes stressed their extremely high primary production, it now seems that many of the


early estimates were obtained with flawed methods and the values should be revised downwards (Morrisey 1995; Nybakken 2001; see Good et al. 1982 for a comprehensive review of the early data on saltmarsh primary productivity). In contrast, it seems that rates of net productivity in mangroves have been markedly underestimated by most methods used in the past and bona fide productivity is likely to be far higher (up to an order of magnitude) than considered formerly (Alongi 2009).

Knox (1986) assessed the range of biomass-based techniques that can be used to quantify annual net primary production in saltmarshes. In general, five techniques are available: i) methods which assume that maximum live biomass corresponds to primary production; ii) measurements based on maximum standing crop (i.e. live biomass plus dead material); iii) methods that assume that primary production equals maximum live biomass minus minimum live biomass within any one year; iv) the method used by the International Biological Program, which assesses primary production as a net change in weight between successive sampling periods; and v) Smalley’s method, which quantifies annual net primary production as discrete positive or negative ‘gains’ over a series of given intervals.

The different techniques can yield productivity estimates that vary several fold. Knox (1986) reported, for example, that net annual production for a Puccinellia-dominated marsh in eastern England estimated on the basis of differences between maximum and minimum biomass was 254 g DW m–2 year–1, whereas the maximum-standing-crop method returned an estimate over three times higher, at 800 g DW m–2 year–1. Moreover, the methods commonly used with vascular plants quantify the above-ground production only, and the less-often-measured production of below-ground components can easily exceed that of the leaves and shoots. In coastal Spartina alterniflora saltmarsh of Georgia (USA), for example, net above-ground primary production varied from 1,350–3,700 g DW m–2 year–1 whereas comparable below-ground values varied from 2,020–4,780 g DW m–2 year–1 (Day et al. 1989).

Primary production by non-vascular plants is usually omitted from such whole-of-marsh assessments, and the productivity of benthic algal mats or algae associated with emergent plant surfaces can often equal or exceed that of the visually obvious angiosperms. Sullivan & Currin (2000), for example, showed that primary production by benthic algal ranged from ~10–60% that of the overstorey vascular plant community in Atlantic and Gulf coastal saltmarshes in the USA, and from ~75–140% in a saltmarsh in southern California. The significance of algal productivity becomes even greater when it is recognised that algal (especially diatom) biomass has a ‘better’ C:N:P ratio (Atkinson & Smith 1983), and is more readily incorporated into metazoan food webs than the highly lignified tissues of vascular wetland plants.

The seminal studies by Teal (1962) of Spartina-dominated marshes of Georgia (USA) indicated a net primary production of about 1,600 g C m–2 yr–1 (= ~3,200 g DW m–2 year–1 or ~32 tonnes ha–1 yr–1). That value is commonly cited to support claims for very high saltmarsh primary productivity. Subsequent studies in different regions of North America, perhaps not unexpectedly, have returned a range of different values (Mendelssohn & Morris 2000). In the Spartina-dominated marshes in New Jersey, net primary production is only 325 g C m–2 yr–1 (~6 tonnes ha–1 yr–1: Good 1965) and, on the Gulf coast of the USA, typical values of 300–3,000 g C m–2 yr–1 (~6–60 tonnes ha–1 yr–1) have been reported (Turner 1976). Saltmarshes around San Francisco on the mid Pacific coast have a net primary production of about 50–1,500 g C m–2 yr–1 (~1–30 tonnes ha–1 yr–1: Atwater et al. 1979; cited in Nybakken 2001) and Juncus gerardii marshes along the coast of Maine in eastern USA of between about 200 and 600 g DW m–2 year–1 (~120–750 g C m–2 yr–1 or ~2.4–15


tonnes ha–1 yr–1), depending on the method used to estimate production (Knox 1986). Figure 1.26 shows the strong geographical effect on primary productivity of coastal saltmarsh in North America.

Figure1.26: Annual above-ground productivity of North American coastal saltmarsh, sorted according to location. Source: Mendelssohn & Morris (2000, Figure 1).

There are few estimates of biomass or primary production in Australian saltmarshes. In the most detailed study to date, Congdon & McComb (1980) quantified the above-ground biomass and primary productivity of a Juncus kraussii marsh in southern Western Australia as 3.7–14.1 tonnes ha–1 and 3–13 tonnes ha–1 year–1, respectively. Clarke & Jacoby (1994) reported the biomass of Juncus kraussii near Jervis Bay (south coast of New South Wales) ranged from 96–4,400 g DW m–2 (~0.1–44 tonnes ha–1) and estimated the above-ground net primary productivity at ~800 g DW m–2 yr–1 (~8 tonnes ha–1 yr–1). Laegdsgaard (2006) made available some previously unpublished values for above-ground production of saltmarsh plants: 148–1,600 g DW m–2 year–1 (~1.5–16 tonnes ha–1 yr–1) for Sporobolus virginicus; 88–800 g DW m–2 year–1 (~0.1–8 tonnes ha–1 yr–1) for Sarcocornia quinqueflora; 400 g DW m–2 year–1 (~4 tonnes ha–1 yr–1) for Samolus repens; and 1,100 g m–2 year–1 (~11 tonnes ha–1 yr–1) for Bolboschoenus caldwellii.

Considerably more information is available on the biomass and/or primary production of Australian mangroves (e.g. Briggs 1977; Bunt 1982; Clough & Attiwill 1982; Clough 1992). Information on Victorian mangroves is mostly quite old, often having been obtained during the Western Port studies of the mid-1970s (Shapiro 1975a,b; Ministry for Conservation 1977). It is likely that bona fide rates of primary production were seriously underestimated in the older studies (Alongi 2009). Notwithstanding this caveat, a comparison of productivity and/or biomass for a selection of saltmarsh and southern-Australian mangrove communities is shown in Table 1.11. Mangroves along the south-eastern coast of Australia would seem to have slightly higher net productivity than saltmarshes, and values ranging from about 2 to 15 tonnes ha–1 yr–1 have been reported in the relatively meagre dataset.


Table1.11: Comparison of above-ground biomass and net primary production for a range of mangrove and saltmarsh communities in the Southern Hemisphere. Comparative values are shown for selected Northern Hemisphere systems as well. Source: modified from Saenger (1994, Tables 13.1, 13.3 and 13.4).

Species Country and latitude Above-ground biomass(tonnes ha–1)

Net primary production(tonnes ha–1 year–1)

Mangroves

Avicennia marina Australia: 18oS 8

Australia: 27oS 2–6

South Africa: 30oS 7

Australia: 32oS 7–104 ~15

Australia: 34oS 24–128 6–7

Australia: 38oS 86 2

New Zealand: 37oS 7–104 3–7

Saltmarshes

Juncus kraussii Australia: 32oS 12

Australia: 34oS 4–4 3–13

Australia: 34oS 3–7

Mixed saltmarsh Australia: 34oS 41

Sarcocornia marsh Australia: 34oS 5–8

Samolus marsh Australia: 34oS 4

Northern Hemisphere

Spartina marsh USA 8–66

Typha marsh USA 4–23

Mangroves USA 12–73

secondary production and food-web structure

It is widely acknowledged that the vascular-plant based food webs in mangrove and saltmarsh ecosystems are detritus-driven (Long & Mason 1983; Por 1984b). In other words, rather than live vascular plant material being consumed by grazing herbivores, plant material enters food webs only after it has died and has been colonised by aquatic microbes. The dead plant material is first leached of dissolved organic carbon by rain while still standing or by periodic tidal submergence; bacteria and fungi then colonise the particulate material that remains. Grazing invertebrates, such as snails, scrape off the nutritionally rich microbial biofilms and perhaps gradually remove also parts of the underlying vascular plant substrata. Eventually, the vascular plant material is either buried (e.g. as peat) or supports metazoan food webs via this detrital pathway. A portion may be exported as dead plant material out of the saltmarsh, either into surrounding estuarine waters or stranded along the drift line; this still-controversial topic is discussed later.


The classic study demonstrating the importance of detrital food webs in saltmarshes is that of Teal (1962), who worked on Spartina-dominated high saltmarshes in Georgia (USA). Table 1.12 shows the energy fluxes reported in that study. The salient points are: i) the dominance of bacterial respiration as a fate of vascular plant detritus; ii) the relative insignificance of herbivory; and iii) the large amount of material that seemed to exported from the marsh. Less than 5% of primary production was directly consumed by herbivores in the saltmarsh.

Table1.12: Summary of the main energy pathways through a Spartina-dominated high saltmarsh in Georgia, USA. Source: Teal (1962).

Component Percentage of flux

Net primary production by Spartina 100

Net primary production by algae 2.5

Consumption by herbivores < 5

Consumption by bacteria 47

Export ~45

The conclusions reached by Teal (1962) about the fate of saltmarsh plant material have been, by and large, substantiated by more recent investigations. Figure 1.27 shows, as an example, a model proposed by Newell & Porter (2000) for the fate of Spartina detritus in North American coastal saltmarsh. The critical role played in the processing of vascular plant material by hitherto ignored microbiota – bacteria, fungi and meiofauna – is clearly evident.

Figure1.27: Conceptual model of the fate of vascular plant material in Spartina-dominated saltmarshes of North America. Source: Newell & Porter (2000, Figure 4).


There are at least three reasons why direct herbivory of vascular plant material is a relatively minor fate of primary production in saltmarshes, at least for Spartina-dominated saltmarshes of the Northern Hemisphere. First, saltmarsh plants contain high concentrations of structural tissue, which is difficult to digest by non-ruminant herbivores. The high concentrations of structural cellulose, hemi-cellulose and lignin mean that vascular plant material usually has high C:N and C:P ratios (Atkinson & Smith 1983); in other words, the vascular plant material is not very nutritious.

Second, the nitrogen present in plant tissues may not be readily assimilated by herbivores. Studies of the effects of fertilisers on Spartina-dominated saltmarshes in north-eastern USA have shown a strong impact of nitrogen enrichment on the consumption of saltmarsh plants by herbivorous invertebrates, with enriched plots often having plants that are more readily consumed by herbivores (e.g. Deegan et al. 2007 and Sala et al. 2008; but see also McFarlin et al. 2008 for contrary evidence). Although chenopods often have comparatively high total-nitrogen contents, it is not clear whether the nitrogen contained in the cellular osmotica that makes up much of this tissue nitrogen is as available to grazing herbivores as nitrogen contained in protein. In the case of the amino acids such as proline, organic osmotica are probably bioavailable: it is not clear whether other nitrogen-containing cellular osmotica (e.g. glycine-betaine) are incorporated as effectively.

Third, like all halophytes, saltmarsh plants contain high concentrations of salt, and salt can be a strong feeding deterrent for many vertebrate grazers. Leigh (1981), for example, noted that the high salt content of chenopod shrubs in inland Australia presented serious difficulties for graziers trying to provide sufficient drinking water for sheep. It is not clear whether high internal salt concentrations are a strong deterrent for invertebrate herbivores which often suck plant juices and may be able to selectively tap into the phloem. However, even when the phloem is used, the water will have a low water potential and that may act as an effective deterent to invertebrate feeding.

The relatively minor importance of direct herbivory means that degradation by microbes (bacteria and fungi) is by far the major fate of the vascular plant material produced in saltmarshes (Figure 1.27). The early data collected by Teal (1962) have been amply supported by more recent studies: Newell et al. (1983), for example, showed suspended bacteria were responsible for oxidising 86–91% of the carbon that entered planktonic food webs in saltmarshes of North Carolina (USA). Bacterial production is likely to be even more important in saltmarsh sediments (Tibbles et al. 1992). Newell & Porter (2000) provide an excellent summary of more recent information on the role of microbes in food webs and biogeochemical cycles in North American coastal saltmarsh.

Consumption by grazing waterbirds is one exception to the generalisation that saltmarsh primary production is mostly not channelled directly into herbivory-based food webs. Snow Geese Anser caerulescens, for example, removed nearly 60% of the biomass of Spartina and Scirpus in a North Carolina saltmarsh, and other reports show Snow Geese ‘eating out’ entire marshes in Louisiana and Texas (Good et al. 1982). In the Victorian context, the only information we are aware of in relation to birds consuming large amounts of saltmarsh plants is the case of Cape Barren Geese Cereopsis novaehollandiae, which have been reported feeding on leaf material of Disphyma crassifolium ssp. clavellatum (Rogers 1990). The Orange-bellied Parrot Neophema chrysogaster is a saltmarsh specialist which feeds on seeds of a wide range of coastal halophytes (Mondon et al. 2009). This topic is discussed later.


Although herbivory is likely to be an insignificant fate of vascular plant material in grass-dominated saltmarshes of the Northern Hemisphere, the same cannot be said about the algal components of the saltmarsh or mangrove flora. Algae are not only more nutrient-rich than vascular saltmarsh or mangrove plants, but have a lower content of recalcitrant structural compounds such as cellulose, hemicellulose and lignin (Atkinson & Smith 1983). They are thus likely to be more attractive as food sources to invertebrates that feed by scraping biofilms from leaf surfaces, sediments or hard material such as rocks and submerged wood. Few studies have addressed the topic of the relative palatability of vascular plants versus algae and microbes but, in one of the earliest, Peterson & Howarth (1987) showed that macrophytes (in this case, Spartina) and algae were about of equal importance in supporting food webs, and the relative importance of each varied with the animal concerned, size, location and trophic position of the various consumers. A more recent study (Galvan et al. 2008) examined the dietary contribution of different types of primary producer in a Spartina-dominated marsh near Massachusetts (USA) and concluded that phytoplankton and benthic algae were the dominant food sources for saltmarsh invertebrates. The role of vascular plant detritus varied among species and habitats. In his recent review, Alongi (2009) concluded that algae, not only vascular plant detritus, were important in fuelling mangrove food webs.

In many cases the relative contributions of different primary producers has been unravelled by analysing the stable isotopes of carbon (δ13C) and nitrogen (δ15N), sometimes supplemented with sulfur (δ34S). The method is based on differences in δ13C being used to identify ultimate food (i.e. carbon and energy) sources, and differences in δ15N being used to identify trophic position (e.g. see Bunn & Boon 1993; Boon & Bunn 1994). In coastal systems, seagrasses, algae and the C4 plant Spartina are often unable to be separated on the basis of their carbon and nitrogen isotopic signatures, and δ34S analysis then becomes useful to differentiate between potential food sources (Connolly et al. 2004).

Many studies have used isotopic methods to elucidate the structure of food webs in Northern Hemisphere saltmarshes (Haines 1977, 1979; Jackson et al. 1986, Créach et al. 1997; Eldridge & Cifuentes 2000; Winemiller et al. 2007; Galvan et al. 2008), and the method has found increasing application in Australian saltmarshes and mangroves (e.g. Boon et al. 1997; Sheridan & Hays 2003; Connolly 2003; Guest & Connolly 2004, 2006; Melville & Connolly 2005; Svensson et al. 2007; Hindell 2008; Spilmont et al. 2009; Abrantes & Sheaves 2010). In possibly the first application of the method to Australian saltmarshes, Boon et al. (1997) used δ13C and δ15N analyses to determine the relative importance of mangroves, saltmarsh and seagrasses as food for burrowing callianassid shrimp in intertidal mudflats on the northern shore of Western Port. Seagrasses and algae were the most important source of carbon and energy, and saltmarsh plants (mostly Sarcocornia) were relatively unimportant as a food source. Svensson et al. (2007) used stable-isotope analyses to examine the role of coastal saltmarsh in two estuaries in south-western Western Australia; seston, benthic microalgae and, to a lesser extent seagrasses and saltmarsh, were the main contributors to food webs in these estuaries. Saltmarsh played a greater role at sites near rivers and close to dense fringing vegetation: other detrital sources were more important at other sites. In the site where coastal saltmarsh had been removed as a result of coastal development – Leschenault Inlet – saltmarsh played a lesser role in supporting food webs than it did in a site where saltmarsh remained more intact. (As an aside, we note that the δ15N signature of saltmarsh plants can be used also to indicate the degree of eutrophication: see Castro et al. 2007 for an example of this application in a Portuguese estuary.)


Results obtained with the use of stable isotopes have been largely confirmed by studies using other techniques, especially gut analysis. The critical difference between stable-isotope analyses and gut analyses, however, is that the former indicates what foods are actually assimilated by the animal, whereas the latter indicates only what is eaten. Some foods, especially those that are difficult to digest (e.g. brown and red algae), may be ingested but not digested, and so do not contribute meaningfully to the nutrition of the herbivore. Mazumder et al. (2006) used gut-content analyses to show that itinerant fish leaving a saltmarsh in Botany Bay (e.g. Glassfish Ambassis jacksoniensis, Flat-tail Mullet Liza argentea and Blue Eye Pseudomugil signifer) had fed on crab larvae while they were in saltmarsh. Most recently, Platell & Freewater (2009) reported a similar importance of crab zoeae as food for small fish in saltmarshes in the Brisbane Water estuary on the central coast of New South Wales.

export of detritus and support of offshore food webs

Early (1960s to 1970s) studies of saltmarshes and mangroves suggested strongly that coastal environments were net exporters of particulate organic material (especially vascular plant detritus) to their surrounding estuary (see Long & Mason 1983 for a typical analysis, and Table 1.12 for the seminal data of Teal 1962). The conceptual model developed at the time viewed saltmarshes as essential provisioners of coastal food webs: they exported detrital vascular plant material that fuelled adjacent food webs, resulting in increased fish production in estuaries and ultimately to substantial human economic benefit. Such an export of particulate organic detritus provided the rationale for protecting estuarine plant communities, especially saltmarshes, from destruction and other types of anthropogenic alteration.

In a major paradigm shift, Haines (1977, 1979) suggested that, rather than being a source of vascular detritus for nearby food webs, the importance of coastal saltmarsh was more as nursery and feeding areas into which fish could move on a high tide to feed (on algae mostly) and avoid predators. The critical role of saltmarsh then became the export of living animals, especially larval stages of invertebrates and fish, rather than the export of dead particulate detritus derived from vascular plants into waters of the nearby estuary. The concept of saltmarshes, at least those in the Northern Hemisphere, as nursery areas rather than an obligatory exporters of particulate detritus was tested by Nixon (1980). He examined all available mass-balance studies of coastal saltmarsh, and concluded that many of the commonly held beliefs about the functioning and importance of coastal wetland communities were not supported by empirical evidence. Even if broad patterns in export could be discerned, site-specific differences often made it impossible to generalise across all saltmarshes with regard to the amount of material exported as particulate detritus and the role that such exports had in supporting coastal food webs. It is now clear that not all saltmarshes, even all Spartina-dominated systems, export nutrients (Nixon et al. 1976) nor even organic matter (Chalmers et al. 1985), as was previously thought.

Over 25 years ago, Long & Mason (1983, page 124) concluded that ‘It is clearly very unwise to generalise about the role of salt marshes as either sources or sinks of materials to neighbouring coastal waters’. The topic continues to generate debate: for more information see the recent reviews by Teal & Howes (2000), Valiela et al. (2000), Zimmerman et al. (2000) and Kreeger & Newell (2000). Recent research, however, lends support to the notion that coastal saltmarsh is particularly important in terms of providing habitat for small invertebrates and fish, which then move into the estuary on out-going tides or when they are mature (e.g. see Stevens et al. 2006 for fish; Mazumder et al. 2006 for crabs), rather than as supporting offshore food webs by exporting vascular plant detritus.


The caveat that limits the extrapolation of these findings to Victorian saltmarsh is that, again, the critical studies to test the various explanations have been undertaken almost entirely on North American, Spartina-dominated systems, and little, if at all, for other systems (e.g. see critiques for mangroves by Manson et al. 2005 and Alongi 2009). Saltmarsh along the Victorian coast differs greatly from that of the Spartina-dominated Atlantic coast in at least three ways, and all could have an impact on the role of saltmarsh as supporters of offshore food webs. First, south-east Australian saltmarsh is generally subject to less frequent tidal inundation: descriptions of Northern Hemisphere saltmarshes stress their intertidal location and regular tidal inundation, whereas large parts of Victorian saltmarsh are very nearly supratidal, especially the upper marsh areas. The lack of frequent tidal inundation means there is less opportunity for plant material to be washed out of the saltmarsh twice a day on the ebbing tide. It has implications also for the use of saltmarshes by nekton (Kneib 1997).

Second, it would seem that the high primary productivity attained in some, or even many, North American Spartina-dominated saltmarshes is not always matched in saltmarshes of south-eastern Australia (see Table 1.11), even allowing for possible overestimation of net primary production in some early studies.

Third, chenopods dominate much of the vegetation in the saltmarshes of south-eastern Australia, and such highly succulent plants may decompose at different rates than grasses such as Spartina. Although there is an abundance of information on primary production and decomposition processes in Northern Hemisphere saltmarshes (e.g. see Valiela et al. 1985), little research has been undertaken on decomposition processes or rates in Australian saltmarshes. The only published study we could locate was van der Valk & Attiwill (1983), who found that above-ground litter of Tecticornia (Sclerostegia) arbuscula and Sarcocornia quinqueflora decreased by 61% and 50%, respectively (by mass), after about three-quarters of a year in Western Port saltmarshes. Below-ground litter lost only 35–43% of its weight over the same period.

provision of habitat

There is a substantial literature on the value of mangroves and coastal saltmarsh as habitat for birds and fish, and it continues to be updated regularly by new research and periodic reviews (e.g. Stephens et al. 2006; Faunce & Serafy 2006; Wilson et al. 2007; Visser & Baltz 2009; Spalding et al. 2010). Much less is known about the use of coastal wetlands by mammals or reptiles, although the little evidence that is available suggest that the utilisation of intertidal habitats by terrestrial mammals is intermittent or cyclical (Moore 2002; Visser & Baltz 2009).

Birds

Large numbers and diverse species of birds use mangroves and coastal saltmarsh, including foraging rails, crakes, plovers, stilts, avocets, ibis, egrets and ducks, and roosting swans, cormorants and pelicans (Land Conservation Council 1993; see Figures 1.28–1.31).


Figure1.28: Royal Spoonbills Platalea regia feeding and preening in the interface between mangroves and saltmarsh.

Figure1.29: Masked Lapwings Vanellus miles in coastal saltmarsh.

Figure1.30: White-faced Heron Egretta novaehollandiae in coastal saltmarsh.

Figure1.31: White-fronted Chats Ephthianura albifrons feeding on the edge of coastal saltmarsh.

In terms of specific bird species, it is known that Victorian coastal saltmarsh provides over-wintering habitat for Eastern Curlew Numenius madagascariensis, Common Sandpiper Actitis hypoleucos, Red-necked Stint Calidris ruficollis, Common Greenshank Tringa nebularia, Marsh Sandpiper Tringa stagnatilis, Double-banded Plover Charadrius bicinctus, Sharp-tailed Sandpiper Calidris acuminata and Latham’s Snipe Gallinago hardwickii. Blue-winged Parrot Neophema chrysostoma and Orange-bellied Parrot Neophema chrysogaster feed in saltmarshes along the Victorian coast. The Orange-bellied Parrot is critically endangered in Australia and Victoria (Department of Sustainability and Environment 2007c) and is listed under both the Victorian Flora and Fauna Guarantee Act 1988 and the Commonwealth Environment Protection and Biodiversity Conservation Act 1999 (see Chapter 2 for detail on legislation relevant to the Australian coast). The role of Australian saltmarsh in supporting colonial-breeding and migratory bird species was reviewed by Spencer et al. (2009). They concluded that saltmarsh provided feeding and roosting habitat for a number of colonial-breeding waterbird species, including White Ibis Threskiornis molucca, Straw-necked Ibis Threskiornis spinicollis and Cattle Egret Ardea ibis, especially when inland wetlands were in drought. Occasionally large


numbers of Black Swan Cygnus atratus, Chesnut Teal Anas castanea and Australasian Shelduck Tadorna tadoranoides fed and roosted in coastal saltmarsh.

Not only shorebirds make use of coastal saltmarsh: small passerines such as White-fronted Chat Ephthianura albifrons, White-browed Scrubwren Sericornis frontalis, Horsfield’s Bronze-cuckoo Chrysococcyx basilis, Southern Emu-wren Stipiturus malachurus, Striated Fieldwren Calamanthus fuliginosus and Brown Thornbill Acanthiza pusilla are regularly found in saltmarshes. Raptors such as Swamp Harrier Circus approximans and Whistling Kite Haliastur sphenurus and, less commonly, White-bellied Sea Eagle Haliaeetus leucogaster and Peregrine Falcon Falco peregrinus also have been observed in coastal saltmarsh. At The Spit Nature Conservation Reserve on the western shores of Port Phillip Bay, Brolga Grus rubicunda has made several attempts to nest in the upper saltmarsh in recent years; they also utilised tidal mudflats at The Spit Lagoon as feeding habitat (Geoff Carr, pers. obs.).

While many of these bird taxa use saltmarsh episodically or seasonally rather than as residents, in some areas of coastal Victoria where the clearance of hinterland vegetation has been particularly severe more sedentary birds (in particular passerines) depend on coastal saltmarshes for most or all of their life cycle. Resident passerines, sometimes breeding, including White-fronted Chat (see Figure 1.31), Southern Emu-wren and Brown Thornbill, have been recorded year-round in some coastal saltmarshes (Glenn Ehmke & Chris Tzaros, Birds Australia, pers. comm.).

Spencer et al. (2009) concluded that little was known of the role played by coastal saltmarsh for migratory bird taxa in Australia, but did note overseas studies that showed a pivotal role of coastal saltmarsh for migratory shorebirds in South Africa, North America and Europe. They proposed that many of the bird species that migrate to Australian saltmarshes fed in the marshes during the day and roosted in them at night. The use of the same habitat for feeding and roosting might provide substantial energetic benefits to migratory birds, as they would not have to fly to and away from feeding grounds each day. Moreover, saltmarsh may provide important habitat with reduced exposure to predators; bird species that use saltmarshes as a roosting sites, for example, do not tend to use wooded areas (including mangroves) because the latter can hide land-based predators and, more importantly, restrict landing and take off areas. Waders also may prefer saltmarsh as night-time roosts because of the safety that the open vegetation structure provides from predators (Spencer et al. 2009). Additionally, many saltmarshes contain ponds of water that are flooded by the occasional high tide (see Figure 1.6): these ponds contain aquatic macrophytes (e.g. Ruppia or Lepilaena spp.), as well as small fish and invertebrates that remain after the tide has receded, and can be valuable food resources for shorebirds (Laegdsgaard 2006). Table 1.13 shows the variety of habitat attributes possessed by coastal saltmarsh that may be taken advantage of by shorebirds.


Table1.13: Properties of coastal saltmarsh that make them attractive roosting sites for shorebirds. Source: modified from Spencer et al. (2009, Table 7.2).

Property Importance to shorebirds Mechanism

Topography Protection from adverse weather Energetics

Substratum Availability of supplementary foraging habitat Energetics

Shallow water Availability of supplementary foraging habitat; preening Energetics

Proximity to feeding areas Travelling time to/from feeding areas Energetics

Vegetation Protection from adverse weather; camouflage from predators Energetics; predation

Distance to tall vegetation Protection from predators Predation

Proximity to foreshore Escape distance from predators Predation

Roost background colour Camouflage from predators Predation

Remoteness Freedom from disturbance to feeding/roosting Disturbance; energetics

Size of roost Inter-bird aggression Disturbance; energetics

The Orange-bellied Parrot Neophema chrysogaster deserves special mention because of their critically endangered status in Victoria. The basis of the following text was provided largely by Glenn Ehmke from Birds Australia. Orange-bellied Parrots are winter migrants from Tasmania to mainland south-eastern Australia, and are generally present in Victoria from April to September. The Victorian range of the species extends from east of Corner Inlet in Gippsland to the South Australian border. Recent (2006–2008) habitat modelling has elucidated patterns of regional-scale habitat preferences within the species’ mainland range: in general Orange-bellied Parrot are found around low-lying, low-energy coastlines, particularly bays, estuaries, inlets and coastal saline lakes. Although birds forage also in nearby pastures on a range of exotic dicots and grasses, in Victoria coastal saltmarsh is the principal over-wintering foraging habitat for Orange-bellied Parrots. These coastal communities support many of the species of food plant needed by the species, including Beaded Glasswort Sarcocornia quinqeflora, Shrubby Glasswort Tecticornia arbuscula, Austral Sea-blite Suaeda australis and other less commonly used species such as Grey Glasswort Tecticornia halocnemoides and Southern Sea Heath Frankenia pauciflora (Loyn et al. 1986).

Orange-bellied Parrots display a number of consistent habitat preferences within Victorian coastal saltmarsh. First, they prefer wetter, more frequently inundated saltmarshes and, in fact, are rarely observed foraging more than 10 m from a saltmarsh shoreline. Second, they tend to prefer saltmarshes with complex shorelines that are close to fresher waterbodies or watercourses. Thus many of their preferred foraging sites are made up of small patches of saltmarsh surrounded by water, on small islands and spits. Third, they forage in relatively open saltmarshes where their preferred food plants are not totally dominant. In Victoria, saltmarshes are often densely vegetated with potential food-plant species, such as Beaded Glasswort. The birds, however, avoid these areas and prefer areas with a more open structure and some bare ground. The bare ground present in foraging plots is often expressed as narrow avenues dissecting areas where the preferred food plants are found. Such patterning allows the parrots to move around freely on the ground while feeding on the seed of saltmarsh plants. The implication is that, as far as the foraging needs of Orange-bellied Parrot are concerned, high-quality saltmarsh habitat is quite restricted in Victoria.


The phenology of seed set is an important factor controlling the distribution of Orange-bellied Parrot. As no single known food plant produces seed over the entire over-wintering season, the birds depend on a mosaic of different food resources and habitat types to satisfy their foraging requirements. Habitat selection in response to the different seeding times of the most important food plants Sarcocornia quinqeflora and Tecticornia arbuscula is the best recorded example of this phenomenon. Sites such as Lake Connewarre in Victoria that have abundant and highly productive Sarcocornia quinqeflora are often abandoned after July, when the plant has finished seeding, presumably because the late-seeding Tecticornia arbuscula is rare or absent from this location. In contrast to the case with Lake Connewarre, Orange-bellied Parrots are more often recorded at Swan Bay, a site rich in Tecticornia arbuscula, later in the year (i.e. after July), when that plant species has the most seeds.

Mangrove shrubs are well used by roosting birds, and foraging waders use the muddy substrata between the trees as feeding areas (Land Conservation Council 1993). Birds known to use mangroves include cormorants (e.g. Little Pied Cormorant Phalacrocorax melanoleucos), ibis (e.g. White Ibis Threskiornis molucca), spoonbills (e.g. Royal Spoonbill Platalea regia), egrets (e.g. Little Egret Egretta garzetta), herons (e.g. Mangrove Heron Butorides striatus) and oystercatchers (e.g. Pied Oystercatcher Haematopus longirostris). Grey-tailed Tattler Heteroscelus brevipes and Terek Sandpiper Xenus cinereus roost in mangroves, and Common Sandpiper Actitis hypoleucos and Greenshank Tringa nebularia are known to forage in mangrove mud; Azure Kingfishers Alcedo azurea are often seen in mangrove trees. Because of their high productivity and complex food webs, mangroves are intimately linked with the productivity of commercially important species such as finfish and penaeid prawns (Saenger 1994) and these high rates of secondary production may provide a ready source of food to these species.

Fish

There is a clear hierarchy in current knowledge about the roles of different types of coastal aquatic vegetation as nursery areas for fish and the support they give to commercial fisheries: for seagrasses the relationship is well established (see Green & Short 2003 for a global overview; Saintilan 2004 for an example from south-eastern Australia); for mangroves the evidence is more equivocal and additional research needs to be undertaken (Manson et al. 2005; Alongi 2009); and for coastal saltmarsh (especially in areas other than the Spartina-dominated systems of North America) the evidence is scant.

There are many reasons for this lack of knowledge, but sampling difficulties are perhaps paramount. Coastal saltmarshes are not easy environments to study fish, as water on the marsh is usually shallow and temporary and often tangled with vegetation. Even the most permanent water (in creeks) is tidally influenced and comes and goes. It is, therefore, not surprising that fish populations of Australian saltmarshes have been little examined, and Connolly (2009) could identify only 11 studies on the topic. In an early study of fish in an Australian saltmarsh, Morton et al. (1987) reported the presence of 19 species, from 14 families, in a creek that drained coastal saltmarsh in south-eastern Queensland. The recent implementation of the pop-net technique for sampling fish from vegetated areas in coastal saltmarsh has allowed sampling to be undertaken in areas other than the easily-sampled creeks (Kneib 1997; Connolly 2009). Sampling undertaken with this new approach has shown that the small ponds common on saltmarsh flats are important habitat for small fish, including juvenile forms of large species (e.g. Silver Biddy Gerres ovatus and Yellow-finned Bream Acanthoprus australis) and adults of small species such as Blue Eye Pseudomugil signifer and a range of gobies. Connolly


(2009) argued that fish assemblages on inundated saltmarsh were dominated by adults of one or two small species, usually from the families Gobiidae across Australia, Ambassidae in subtropical and temperate regions, and Atherinidae in temperate regions.

In one of the few Victorian studies, Crinall & Hindell (2004) used fyke and seine netting to study the fish fauna and their feeding habits in a saltmarsh of Western Port. Ten fish species were identified, of which the Hardyhead Atherinosoma microstoma was the most abundant. A number of commercially important species (Yellow-eye Mullet Aldrichetta forsteri, Spotted Whiting Sillaginoides punctata, and Greenback Flounder Rhombosolea tapirina) also were recorded. Gut-content analysis showed that most fish ate gammaridean amphipods and hemipteran insects.

If mangroves act as nursery areas for juvenile fish and support commercial fisheries, it is mainly because mangroves provide refuge from predators and shelter from physical disturbance, and are high-nutrient environments that provide abundant food and thus allow rapid growth rates of secondary consumers (Laegdsgaard & Johnson 2001; Manson et al. 2005; Alongi 2009). Hindell & Jenkins (2004) concluded that temperate mangroves supported an assemblage of fish that was less species-rich, but just as abundant in terms of total fish numbers, as those for mangrove systems in the tropics. They suggested that mangroves provide fish, especially at young stages, with resources not available on nearby mudflats, and especially refuge from predation and better access to food. There appears to be a dearth of refuges for small fish in estuaries at low tide (e.g. permanent water holes within the mangroves or mudflats) and, except for several species of goby (e.g. Gobiopterus semivestitus and Pseudogobius olorum) that appear to be sedentary and are small enough to shelter within grapsid crab burrows at low tide, fish must be migrating up to hundreds of metres during flood tides, using shallow unvegetated habitats to feed and/or avoid predation in mangroves areas. Their study demonstrated that, for many species of fish, temperate mangroves provided valuable nursery habitat, but the pattern of usage was highly species-specific and depended on where, at what time and how fish were sampled. Not entirely similar results were reported by Clynick & Chapman (2002) for fish in small patches of mangrove in Sydney Harbour: in this case it was concluded (page 669) that the mangrove patches ‘…may not be particularly important habitat for fish because similar species in similar abundances are found on adjacent mudflats’.

Payne & Gillanders (2009) recently reported on fish assemblages in mangroves of South Australia. They concluded that mangrove habitats in temperate Australia did not support a greater diversity or abundance of fish than did adjacent mudflats. Some species (e.g. Yellow-eyed Mullet) were strongly associated with mangrove habitats, and the total fish abundance was positively correlated with the density of pneumatophores. What this study does corroborate is the concept that it is the maintenance of connectivity between estuarine habitats that should be the focus for estuary management, not only the preservation of individual habitats or species, as it is the complex interplay between mangroves, saltmarshes, mudflats and open-water areas than is important for estuarine fish populations (see Melville & Connolly 2005; Meynecke et al. 2007).

Othervertebrates

Few native mammals are encountered in coastal saltmarsh, and there are no Victorian species dependant on it for their survival. Water Rats Hydromys chrysogaster sometimes live in coastal saltmarsh, and some individuals may have home ranges that are largely composed of saltmarsh. Eastern Grey Kangaroo Macropus giganteus occasionally venture into Victorian saltmarsh, and feed on several saltmarsh plant species, probably including


monocots and succulent chenopods. Saltmarshes along the Georges River near Sydney provide important regional habitat for Swamp Wallaby Wallabia bicolor. Introduced Hares also are found on saltmarshes near Sydney (Paul Adam, University of New South Wales, pers. comm.).

No amphibian or reptile species are restricted to coastal saltmarsh or mangroves in Victoria, but saltmarshes form an important part of their habitat for a number of reptile species. Saltmarshes are particularly important for the Swamp Skink Lissolepis (Egernia) coventryi, which is considered vulnerable in Victoria (Department of Sustainability and Environment 2007c) and is listed under the Flora and Fauna Guarantee Act 1988. This species of diurnal lizard feeds predominantly on invertebrates, supplemented by plant material. It inhabits dense vegetation on lowland swamp margins, particularly saltmarsh dominated by Gahnia filum, Tecticornia arbuscula, Atriplex paludosa and Austrostipa stipoides, along with brackish near-coastal wetlands dominated by a range of species including Gahnia trifida and Poa spp. Saltmarsh provides its core habitat, and is where this species is most abundant (Schulz 1985; Clemann 1997). The Glossy Grass Skink Pseudomoia rawlinsoni (considered near-threatened in Victoria) frequently co-exists with Swamp Skink in saltmarsh habitat. Several other reptile species with generalist requirements occasionally venture into coastal saltmarsh, such as Common Blue-tongue Tiliqua scincoides, Blotched Blue-tongue Tiliqua nigrolutea, Tiger Snake Notechis scutatis and Lowland Copperhead Austrelaps superbus (Peter Robertson, Wildlife Profiles, pers. comm.).

Invertebrates

Saenger et al. (1977) provided a comprehensive – but now dated – list of the invertebrate animals of south-eastern Australia that live in saltmarshes. In her review of Australian saltmarshes, Laegdsgaard (2006, page 383) noted that the invertebrate fauna was ‘…probably the least studied component of saltmarshes in Australia and information is lacking’. She concluded that the sediments of saltmarshes supported few infauna, and only three groups (oligochaetes, polychaetes and bivalves) have been identified with saltmarsh-specific taxa. Crustaceans and molluscs were dominant surface-dwelling invertebrates; a study of Tasmanian saltmarshes revealed their use by 50 species of macro-invertebrate, of which eight species were restricted to saltmarshes alone. Figure 1.32 shows the Mottled Shore Crab Paragrapsus laevis in wet coastal saltmarsh; as noted earlier, crab zoeae are an important food source to many small fish that enter coastal saltmarsh during high tide to feed.

A general absence of aquatic invertebrates from saltmarshes can be attributed to a mixture of the hostile physico-chemical nature of saltmarshes and the presence of intense predation. Recent studies have shown that coastal saltmarsh can be sites of severe invertebrate predation; the common gastropod Salinator solida, for example, is subject to such intense predation by toad fish, bream and eels in saltmarshes of Towra Point near Sydney that excluding predators reduced snail mortality from 82% to < 1% (Laegdsgaard 2006).

As an ecotone between terrestrial and aquatic environments (Traut 2005), saltmarshes should be expected to provide important habitats for terrestrial invertebrates, especially insects. The mosquito fauna of Australia saltmarshes is relatively well known because of the public health and economic importance of mosquitoes as vectors of diseases of humans (notably the viral Ross River fever and Barmah Forest Disease), domestic animals and livestock, and because of the nuisance factor (Russell 1993). The nuisance factor looms large in the negative public perception of saltmarsh as mosquito-infested wastelands (Chapters 1.11 & 3.3). Indeed, a recent outbreak of Barmah Forest Disease near Batemans Bay, on the south coast of New South Wales, was linked with coastal saltmarsh (Paul Adam, University of New South Wales, pers. comm.).


Figure1.32: Mottled Shore Crab Paragrapsus laevis.

There are ~120 mosquito species known for south-eastern Australia , almost all indigenous (Russell 1993) and about five of the species of greatest public-health and nuisance-factor significance (all indigenous species) occur in saltmarshes, mangrove and brackish environments in coastal Victoria. The larvae occur in brackish, saline, and hypersaline water; most species have a wide distribution in coastal and non-coastal environments (Russell 1993, 1996). The great majority of mosquito species in south-eastern Australia occur in non-saline environments (Russell 1993) and by implication relatively few species are adapted in the larval stage to higher salinities. The mosquito species of greatest concern for human populations in saltmarsh, mangrove and brackish environments are Aedes australis (Victoria, New South Wales, Tasmania, South Australia), Aedes alternans (Victoria, New South Wales, South Australia, Queensland, Northern Territory, Western Australia), Aedes camptorhynchus (Victoria, New South Wales, South Australia, Western Australia), Aedes vigilax and Culex annulirostris (Russell 1993, 1996). Of these, Aedes vigilax is particularly well studied, because of its link with diseases such as Ross River fever.

A casual inspection of saltmarshes shows the dense populations of spiders and insects that live in saltmarsh vegetation. More detailed information on the use of saltmarshes by terrestrial insects is not available, although Clarke & Miller (cited in Laegdsgaard 2006) did note that saltmarshes of the Hunter River, central New South Wales, supported 47 species of insects, compared with 13 species in adjacent mangroves. There are undoubtedly many other reports of the terrestrial insect fauna of Australian saltmarshes and mangroves scattered in published and unpublished sources. We have not systematically collated the data here, except for the Australian butterfly fauna monographed by Braby (2000).


Butterflies use saltmarsh flora as food plants: larvae eating leaves and other organs, and adults feeding on nectar as described earlier for plant pollination (Chapter 1.8). Several butterfly species may be saltmarsh specialists, as the larvae feed on halophytic saltmarsh plants in Victoria, though mostly not exclusively. Braby (2000) recorded the following species feeding as larvae on saltmarsh plants:• Hesperilla donnysa Varied Sedge-skipper: Gahnia filum (among others) page 152• Hesperilla flavescens Yellow Sedge-skipper: Gahnia filum (exclusively) page 155 (see also Crosby 1990)• Hesperilla chrysotricha cyclospila Golden-haired Sedge-skipper: Gahnia filum (among others)• Theclinesthes serpentata serpentata Saltbush Blue: Atriplex australasica, A. cinerea, A. paludosa, A.

semibaccata, A. suberecta, *Chenopodium album (among others)• Theclinesthes sulpitius Samphire Blue: Sarcocornia quinqueflora, Suaeda australis (among others).

Although there have been few scientific studies of the infauna of temperate mangrove communities (Chapman & Underwood 1995), it is known anecdotally that these communities provide a sheltered, muddy habitat for a wide range of animals (Harty 1997). Pulmonate gastropods and crabs are dominant in the landward edges but in the seaward zones, where epiphytic algae are more abundant, a rich microcrustacean fauna dominated by amphipods and isopods is found, as well as abundant polychaete worms. There is good evidence that the distribution of different species of snails within a given mangrove swamp varies also with elevation, at least for mangrove communities around Sydney (Chapman & Underwood 1995).

Grapsid crabs (Paragarpsus gaimardii and P. laevis), Semaphore Crab Heloecius cordiformis, Mud Crab Helograpsus haswellianus and Red-fingered Crab Sesarma erythrodactyla are common in mangroves of south-eastern Australia. Molluscs also are abundant and include gastropods (e.g. Bembicum spp. and Austrocochlea spp.) as well as whelks (Velacumantus australis and Nassarius pauperatus). In his survey of the macrobenthos present in mangrove stands in four areas in Nooramunga, in southern Victoria, Morgan (1986) reported a fauna of 23 species, dominated by corophiid amphipods, small bivalves and oligochaetes; the visually-obvious taxa were gastropods and barnacles. Certainly mangrove pneumatophores provide a solid substratum for colonisation by marine invertebrates (Figure 1.33).

Figure1.33: Colonisation of Avicennia marina pneumatophores by barnacles.


1.10 Economicandsocialvalues

a global perspective

Coastal wetlands have rarely received a good press. Nearly three centuries ago Catesby, in his Natural History of Carolina, Florida, and the Bahamas Islands (1731/1754), described mangrove swamps in the following terms:

…these watery woods are also plentifully stored with ravenous fish, turtles and other animals which prey continually one upon another, the alligator on them all: so that in no place have I seen such remarkable scenes of devastation as amongst these mangroves in Andros, one of the Bahama mangroves where carcasses of half-devoured animals are usually floating in the water (cited in Lear & Turner 1977, page 45).

Queen (1977), Doody (2008), Silliman et al. (2009) and Weis & Butler (2009) provide detailed reviews of how humans have used saltmarshes across the globe for a wide range of activities; Walsh (1977), Lear & Turner (1977) and Spalding et al. (2010) produced a similar overview for mangroves. Saltmarshes have long been diked and drained in order to create additional agricultural land, especially for grazing; The Netherlands provides the best example where this process has occurred. In North America, Spartina patens marshes have been used for grazing since colonial times and are often still cut to make hay, and shoots of Spartina alterniflora were cut for thatching. Reed Phragmites australis cutting offered similar opportunities in European saltmarshes (Doody 2008). Coastal saltmarsh has been reclaimed for rice cultivation across the USA, parts of the Mediterranean, and in India and Malaysia. Hunting is an important use of saltmarshes in the Northern Hemisphere: Queen (1977) noted that not only waterbirds but fur-bearing species, such as muskrat, racoons, mink and nutria, were hunted in saltmarshes across northern Europe and the Americas. Oysters and prawns have been collected for centuries by both indigenous and colonial communities from tidal creeks in American saltmarshes, and more recently large expanses of Asian mangrove and saltmarsh have been destroyed for aquaculture (Adam 2002; Vivay et al. 2005). Mangroves and saltmarshes were often proposed in the 1970s as ideal places for the treatment of sewage and stormwater, in the hope that such otherwise ‘useless’ areas would serve as treatment areas for urban waste-waters, not only to remove nitrogen and phosphorus but also to reduce loads of organic pollutants (see Nedwell 1974 for tropical mangroves; Valeila et al. 1973, 1975 for North American coastal saltmarsh; Valeila et al. 1976 for wetlands more generally). Melbourne’s Western Treatment Plant, where sewage from the northern and western suburbs of Melbourne is treated, is located in part on former saltmarsh (see Carr & Kershaw 2009), although in this case the treatment method was grass infiltration rather than direct disposal into coastal wetlands.

Some saltmarshes, particularly in the UK, are currently used as military training grounds (Doody 2008). Important areas of coastal saltmarsh and mangrove are similarly used for military purposes in Australia and, in some cases, have extremely high biodiversity and conservation value because of limited access provided for other uses. In northern Queensland, for example:

The mangrove, saltmarsh and seagrass communities of SWBTA [Shoalwater Bay Training Area] remain in their natural state. Unlike communities elsewhere, they have never been cleared or significantly altered, and are isolated from the elevated sediment and nutrient influences that characterise developed catchments (O’Neill 2009, page 180).


indigenous values

It is clear that many coastal wetlands have profound cultural value for indigenous people. What little information that is available on aboriginal uses of Victorian coastal saltmarsh and mangroves suggests that the littoral zone of Victorian coast had a patchy, but overall low, productivity for Aborigines. Instead, most food resources were concentrated in the freshwater or brackish swamps and estuarine lagoons; food included shellfish, fish, eels and birds such as swans and pelicans (Gaughwin 1981; Freslov & Frankel 1999).

There is no historical evidence of Aboriginal use of saltmarsh plants in Victoria, although that may mean only that any such use was not recorded. As noted below, Avicennia marina seeds were cooked and then leached to remove toxins in Queensland, and Disphyma crassifolium leaves were eaten raw in South Australia, but there is no evidence of their use in Victoria. The River Mangrove Aegiceras corniculatum was used to treat ear ache by aborigines in New South Wales, Queensland and the Northern Territory, although curiously only by women (Lassak & McCarthy 2001). Rhizomes and young shoots of Phragmites australis (a component of Estuarine Wetland rather than Victorian coastal saltmarsh) were eaten in Tasmania, and in Victoria the leaves were used for baskets and bags, and the stems for spears, necklaces (cut into beads), knives and nose-reeds (Gott 1993).

There is considerably more information on the use of coastal wetlands by indigenous people in Queensland (Queensland EPA 2009a). Most wetland sites along the Queensland coast known to be used by Aborigines date from the mid-to-late Holocene period, i.e. less than ~4,000 years BP. The high abundance of late Holocene sites reflects the increased usage of coastal areas as sea levels stabilised ~3,7000 years BP. Presumably sites that were used earlier (e.g. mid-to-early Holocene and late Pleistocene, 5,500–12,000 years BP and 12,000–30,000 years BP, respectively) were submerged as sea levels rose during the Holocene marine transgression.

Queensland EPA (2009a) noted that there were more than 950 Aboriginal cultural sites associated with coastal wetlands in Queensland, including petroglyphs, burial sites, ceremonial stone and earth arrangements, scarred trees, hearths, quarries, middens, wells, fish traps, grinding grooves and food and fibre resources. Although the analysis referred to coastal wetlands, rather than mangroves or saltmarshes specifically, the results are broadly useful for our purposes. In Queensland, coastal wetlands would have provided a wide range of natural resources to Aboriginal people, including shellfish (e.g. oysters, whelks and cockles), prawns and fish. A wide diversity of wetland plants provided useful material, including food from the fern Blechnum indicum, meristems from cabbage palms Livistona australis and L. decipiens, nectar from paperbarks Melaleuca spp., seed from Sporobolus spp., succulent shoots from samphires such as Tecticornia spp, and leaves and shoots of Phragmites australis. Roots and tubers of the sedge Cyperus rotundus were eaten, as were fruits of Avicennia marina after they had been treated to remove toxins.


economic values

There is now widespread recognition that wetlands, in general, provide valuable ecosystem services to human populations and thus have high, if somewhat indirect, economic value (Woodward & Yui 2001). One of the most exhaustive analyses of the economic value of wetlands was undertaken by Ghermandi et al. (2008), who examined 385 estimates of economic value, obtained from 167 studies worldwide. They, however, did not focus on coastal wetlands, and most of the studies were undertaken in North America, Europe and Asia, with very few from Australia.

Historicalaspects

At present saltmarshes and mangroves have little direct economic value (as opposed to the provision of broader ecosystem services) to most of Australian society. Their greatest use is for the production of salt and shell-grit, both of which have the potential to grossly modify coastal saltmarsh. In the past, however, they were considered valuable as areas of land and for other resources, especially by agricultural and early industrial societies. The role played by saltmarshes for grazing, especially after their full or partial loss through land-claim (often euphemistically called ‘reclamation’), is discussed later.

In the late 19th century, Victorian mangrove and coastal saltmarsh plants were heavily exploited for their soda-ash content (Bird 1978, 1981). Since tallow was abundant during the early settlement of Australia (because of the large number of sheep) but soap was relatively scarce, many types of native vegetation were burnt to yield the ash needed for soap making (Adam 1990). Coastal saltmarsh plants were an excellent source of ash for this purpose, since they accumulated inorganic salts as part of their physiological response to living in a saline environment (see Chapter 1.7). In Europe, Salsola spp. were harvested from saltmarshes for ash production and soap making, but in southern Australia mangroves (particularly Avicennia marina) were the main plants used. Carr & Carr (1981) and Whinray (1981) noted that Tecticornia arbuscula, Sarcocornia quinqueflora, Atriplex cinerea and Rhagodia candolleana (the latter being a common coastal species which is rare and only opportunistic in saltmarshes) were used in the same way. Of the non-mangrove species, Sarcocornia quinqueflora had the highest soda-ash content (Whinray 1981).

The utilisation for ash production was so intense that Bird (1978, 1981) argued that clearing mangroves to generate ash for soap making has had long-term impacts on their distribution in Victoria. In his review of the natural history of French Island (Western Port), Lacey (2008) noted that the first Europeans recorded as living on the island (~1843–1844) were engaged in burning mangroves to produce barilla ash for soap making. Ash is required also for the production of glass, and it would seem that glassworts (Salicornia spp.) were used across Europe for this industry (Adam 1994).

There is some evidence that saltmarsh plants have been harvested for food, especially by indigenous peoples and near-starving early settlers in early Australia. Certainly in Europe the glasswort Salicornia europaea has been described as an ‘excellent vegetable’ that is ‘delicious boiled until soft, with the water in which has come to be boil poured away to reduce the saltiness. It is then eaten with melted butter like asparagus’ (Phillips & Rix 1993, pages 78-79). In Australia, Cribb & Cribb (1975) noted that the samphire Sarcocornia quinqueflora is edible (indeed it was used prominently on the ABC’s The Cook and the Chef); Lear & Turner (1977) make the same comment for cooked Avicennia marina fruit. Harvesting such plant taxa today, however, would be of negligible economic or social significance in Australian saltmarshes and mangroves. (Even so, Paul Adam


[University of New South Wales, pers. comm.] reports of a case in New South Wales of the illegal collection for restaurants of Salicornia from coastal wetlands near Sydney.)

In New South Wales, the mangrove Aegiceras corniculatum is a valuable producer of nectar for honey (Lear & Turner 1977); although Avicennia marina is thought to be bee-pollinated (Chapter 1.8), it is not evident that it is economically important in Victorian apiculture. Aegiceras corniculatum has antihelmintic properties due to the presence of the chemical rapanone, but its medicinal potential has not yet been implemented (Lassak & McCarthy 2001). Avicennia marina is recorded by Jones (1971) as having been used to make boat keels and elbows, as well as tools such as mallets. Lear & Turner (1977) noted that Avicenna marina wood had been used for tools and boat-building, its bark for tanning, ash for washing cloth and seeds to make ointments to treat ulcers.

Saltandshell-gritproduction

Large expanses of wet and dry (upper and lower) saltmarsh and associated areas were alienated for commercial or industrial salt production in the 19th century between Melbourne and Geelong on the western shores of Port Phillip Bay and Corio Bay. In these areas – at Moolap and Point Henry, Avalon and Laverton – rainfall averages only about 600 mm per year and evaporation greatly exceeds precipitation (Bureau of Meteorology & Walsh 1993), so allowing salt production by evaporation of seawater. Repeated flooding of the extensive systems of pondages with seawater permits the accumulation of a crust of salt several centimetres deep, and this crust is harvested at the end of summer for processing and refinement. Despite the modification of the former saltmarsh and the creation of pondages, which cover many hundreds of hectares, these areas provide hypersaline wetlands for a wide range of shorebirds and specialists of ephemeral lakes (e.g. Red-necked Avocets Recurvirostra novaehollandiae and Banded Stilts Cladorhynchus leucocephala) that feed on brine shrimp. Thus, the destruction of saltmarsh has – through luck alone – not been an unmitigated disaster for flora and fauna.

Shell-grit mining has been undertaken in other coastal saltmarshes around Victoria. Extensive areas of shell-grit – marine bivalve and gastropod shell deposits, laid down in the shallow seas of the Holocene marine incursions – have been strip-mined from the 19th century up until the present time. The sedimentary material occurs in beds one to two metres deep and has been used as a source of commercial lime, for road and building construction, and as a source of calcium carbonate for the poultry industry. Most mining has occurred in areas adjacent to or supporting saltmarsh, predominantly on the western shores of Port Phillip Bay and Corio Bay from Melbourne to Point Lonsdale, but also in the Point Lonsdale–Collandina area. Shell-grit mining has been very destructive of saltmarshes (Figure 1.34) and it is likely that hundreds of hectares of upper and lower saltmarsh were destroyed.

Notwithstanding this destruction, shell-grit mining has opportunistically created extensive, hyposaline and hypersaline wetlands that support rare Saline Aquatic Meadow vegetation, dominated by the submerged annual and perennial macrophytes Ruppia spp. and Lepilaena spp. (Figure 1.35). These sites, often fringed by saltmarsh regrowth, can be important habitats for a suite of waterbird species, including Little Egret Egretta garzetta, Great Egret Ardea alba ssp. nigripes, Whiskered Tern Chidonias hybridus ssp. javanicus, and Royal Spoonbill Platalea regia (Carr et al. 2005; North et al. 2007).


Figure1.34: Damage caused by shell-grit mining, Point Lonsdale.

Figure1.35: Extensive areas of saltmarsh mined for shell-grit can provide highly significant flora and fauna habitat, Point Lonsdale.

Ecosystemservices

One topic that is not well researched in Australia, but may lead to the better identification of the economic value of coastal wetlands, is the role played by saltmarsh and mangroves in providing ecosystem services (see Woodward & Yui 2001). The dependence of coastal fisheries on estuarine wetlands, in at least some parts of the world, has been noted above. The large number of birds, including migratory species, that are attracted to mangroves and saltmarshes may have implications for regional communities in terms of bird-watching. OzCoasts (2008) argued that saltmarshes provided an important ecosystem service in terms of water purification. There have been few recent reports to extol the value of saltmarshes for treating effluent, but the report by Coleman (1996) is noteworthy. Coleman (1996) proposed that ‘…saline wetlands have the decided advantage in their ability to reduce the phosphates in their discharge waters’ and that there were many positive attributes of saltwater wetlands for treating stormwater in the Adelaide region, specifically at the mouth of the Little Para River (South Australia).

Among the greatest of the economic services provided by mangroves and coastal saltmarsh is likely to be their role in coastal protection. Coastal wetlands provide appreciable – if often totally unrecognised – services in terms of the prevention of coastal erosion. As noted by French (1997), coastal saltmarsh provides valuable ecosystem services by protecting land against erosion and many of these benefits are lost when sea walls and


other ‘sea defence’ structures are used in their place. Doody (2008) showed that saltmarshes attenuated wave energy in various sites in The Wash estuary in the UK by 72–97%; in contrast, bare mudflat attenuated waves by only 10–56%. The irony of ‘protecting’ hinterland by building sea walls and alienating coastal saltmarsh was explored by Doody (2008) also in economic terms. He reported that the retention of an 80 m wide strip of coastal saltmarsh could reduce by a factor of 12 the cost of sea defences in south-east England. A sea wall with a crest height of 12 m was required where there was no saltmarsh to front the coast: with an 80 m wide saltmarsh, the height of the required structure required was reduced to 3 m. A wide saltmarsh zone with an active tidal creek system is a means of initially buffering sea-level rise as the channels accommodate and distribute the increased tidal flows.

social values

In the State of the Marine Environment Report for Australia, Zann & Dutton (2000) concluded that, of all the variety of environments in the country, Australians most cherished the coastal zone. The great social value of the coastal zone is reflected in the high real estate value of land that offers even a glimpse of the sea, as well as by the markedly coastal nature of our population centres; ~90% of the Australian population lives in urban areas, which themselves are mostly the large cities located on the coast (Tiffin & Gittens 2004). Indeed, about one-quarter of the Australian population is now located within a 3 km wide strip of the coast, and the movement of people from inland areas to coastal areas, which started in the early 20th century, continues to this day (Zann & Dutton 2000).

Despite our coast-centric population, there is a dearth of information on values and attitudes held by the general public about marine environments (Zann & Dutton 2000). Some studies have been undertaken of community perceptions of the coast near Perth and of the Great Barrier Reef (Fenton & Syme 1989 and Hundloe 1990, respectively, both cited in Zann & Dutton 2000). More recently, a social marketing approach has been applied to elucidate the environmental values and foster more sustainable behaviour in the Corner Inlet region of Victoria (Ipsos-Eureka 2008). The Corner Inlet region is of particular relevance to the present investigation because of its extensive saltmarshes and mangroves.

Ipsos-Eureka (2008) reported that there was an overall lack of awareness of the unique features of Corner Inlet among the local population, as well as a disconcerting lack of awareness of the impacts created by human activities on coastal and marine ecosystems. Because there was an existing, albeit rudimentary, understanding of environmental issues in the region, it was proposed that a communication strategy intending to foster more sustainable behaviours should aim to further develop this nascent understanding. Interestingly, the Corner Inlet study reported:

…there is not a particularly high level of trust in natural resource management agencies within the community, which appears to boil down to a perceived lack of local presence and knowledge, and a degree of transience among individuals who are contracted to work for various agencies. There is also a perception that individuals within agencies lack respect for the existing knowledge and connections that landholders have in relation to their own properties and the area (Ipsos-Eureka 2008, page 12).

Chapter 3.3 reports on the variety of assessments we undertook, as part of this project, on attitudes of the Victorian community to mangroves and coastal saltmarsh.


1.11 Threateningprocesses

land-claim, habitat destruction and fragmentation

Because of their position along the coast where land is required for human settlement, saltmarshes have experienced a long history of modification and destruction (Gedan et al. 2009; Silliman et al. 2009). In recognition of the magnitude of the changes imposed, Gedan et al. (2009, page 117) concluded that ‘since the Middle Ages, humans have manipulated saltmarshes at a grand scale, altering species composition, distribution, and ecosystem function’. In some cases, such as across much of northern Europe, coastal saltmarshes have been ‘reclaimed’ for agricultural purposes since at least the 17th century and possibly even the 13th century (Adam 1990; French 1997). In other cases, it is known that physiographic changes to estuaries since pre-Roman times have had marked impacts on the distributions of coastal saltmarsh (Steers 1977). Impacts continue to this day: Pontee (2004), for example, pointed out that maintenance dredging of estuaries in the United Kingdom continues to have the potential to adversely affect coastal saltmarsh.

Saltmarsh clearing is still being proposed in coastal Victoria. The most outstanding example is the Lonsdale Lakes residential development near Point Lonsdale, where 40 ha (8 ha of saltmarsh and 32 ha of Saline Aquatic Meadow, formerly saltmarsh subjected to shell-grit mining) are proposed for conversion to a canal residential estate (Carr et al. 2005; Mueck & Smales 2007). The vegetation complex has outstanding biodiversity and conservation values documented by Carr et al. (2005) and North et al. (2007); despite these values, and in apparent contradiction of its own Coastal Strategy, the Department of Sustainability and Environment approved clearing (see the fuller exposition of this case in Chapter 2). The creation of canal estates in New South Wales has been illegal for over a decade (Paul Adam, University of New South Wales, pers. comm.).

‘Reclamation’ is the word often used to describe such activities, especially the filling-in of coastal wetlands for human use. But we feel it is clearly an inappropriate descriptor, since it implies the re-gaining of land that was originally terrestrial and ‘ours’. In recognition of this problem, some authors (e.g. Strong & Ayres 2009, page 11) have enclosed the word ‘reclamation’ in quotation marks to indicate its inappropriateness. We propose that the term ‘land-claim’ is better suited to what has been euphemistically termed ‘reclamation’ in the past. Interestingly, and quite independently, the same word was coined by Thomsen et al. (2009, page 367) to describe the process of wetland loss in Australasian coastal marshes. In many ways ‘land-claim’ overlaps with what is often called ‘habitat destruction’, but is often preferable because it indicates that the land has not been merely modified but, in fact, completely cleared and converted to other uses. There is , of course, a risk that the term may be confused with Aboriginal land claims, so we use the hyphenated form (land-claim) to indicate our use of the term to replace ‘reclamation’.

Because of the country’s more recent colonisation by Europeans, Australian mangroves and saltmarshes do not share the centuries-old history of large-scale modification that characterise so many European saltmarshes. Nevertheless, almost all coastal areas of south-eastern Australia have experienced some degree of alienation and habitat destruction over the past 150–200 years. In the Koo Wee Rup area bordering the northern shore of Western Port, for example, large areas of saltmarsh, mangrove and swamp paperbark (Melaleuca cricifolia) wetlands were ‘reclaimed’ for agriculture, as described enthusiastically by East (1935). Yugovic & Mitchell (2006) reported on historical changes to the Koo Wee Rup swamp since European colonisation. Bird


(1980a,b) reported that much of the mangrove fringe of Western Port has been reduced by clearing, land-claim, drainage and dieback. Changes to the shoreline from the early surveys of 1842 and 1865 were evident by 1939 (Bird 1980a,b) and changes have continued apace since then. A boat channel, for example, was cut into mangroves and saltmarsh in 1967 at Yaringa (Bird 1971) and, as described below, even more extensive marina developments have since taken place. Land claim similar to that of Koo Wee Rup in particular, and Western Port in general, has taken place elsewhere along the Victorian coast, as shown in Figure 1.36 for Shallow Inlet in South Gippsland. Note that this figure shows the extent of all ‘reclaimed’ swampland, not merely that of saltmarsh.

Figure1.36: Extent of swamp lost by land-claim at Shallow Inlet, South Gippsland. Source: Bird (1993, Figure 167).

Saintilan & Williams (2000) reviewed the records of saltmarsh loss in eastern Australia and concluded that, based on 28 surveys employing historical aerial photographs, there had been a widespread decline of saltmarsh from estuaries since the 1930s–1940s (Table 1.14). Not included in this analysis is the finding by Harty & Cheng (2003) that the extent of saltmarsh in the Brisbane Water area of the mid central coast of New South Wales had decreased by 78% since 1954. One result evident in Table 1.14 is the dominance of reports for New South Wales coastal saltmarsh, and the dearth of information on changes in Victorian coastal systems. In one of the few Victorian studies, Ghent (2004) compared past and present distributions of coastal saltmarsh in Port Phillip Bay and found that ~65% of pre-European saltmarsh had been lost, mostly before 1978. Gullan (2008) estimated that about 30% of Victorian coastal saltmarsh had been permanently cleared for coastal or marine development.


Table1.14: Change in extent of saltmarshes for various estuaries of the Australian east coast. Source: Saintilan & Williams (2000, Table 1).

Location Period Saltmarsh lost (% or area)

Queensland

Hinchinbrook Channel 1943–1991 78%

Oyster Point 1944–1983 75%

Moreton Bay 1944–1988 65 ha

Coolangatta-Caloundra 1974–1987 11%

New South Wales

Tweed River 1947–1986 72%

Tweed River 1930–1994 “local increases”

Clarence River 1942–1986 15%

Macleay River 1942–1986 35%

Hunter River excl Hexam 1954–1994 67%

Lake Macquarie 1954–1986 25%

Berowra-Marramarra Creek 1941–1994 25%

Careel Bay 1938–1994 92%

Couranga Point 1954–1994 30%

Homebush Bay 1930–1983 > 80%

Botany Bay 1950–1997 30–100 ha (different sections)

Minnamurra River 1938–1997 49%

Shoalhaven River 1949–1996 “increase”

Currembene Creek 1944–1989 14 ha

Merimbula Lake 1948–1994 30%

Pambula Lake 1948–1994 40%

Victoria

Corner Inlet 1941–1985 “extensive”

Shores of Westernport Bay 1920s–1970s “extensive”

South Australia

Gulf of St Vincent 1935–1979 865 ha

The proportions of habitat loss for Australian saltmarshes are not dissimilar to those reported by Doody (2008) for saltmarshes in other parts of the world, despite our shorter history of land-claim. In the USA, ~50% of the original saltmarsh of Narragansett Bay has been filled in; 73% of the original saltmarsh lost from Puget Sound and 95% from San Francisco Bay (Doody 2008). The area of coastal saltmarshes near New York City has decreased by ~12% since 1959, with low marshes and saltmarsh on islands showing the greatest losses, at 38% and up to 78%, respectively (Hartig et al. 2002). Using historical maps, Bromberg & Bertness (2005) concluded that the average loss of New England (USA) marshes since the early 1800s was 37%; Rhode Island had lost 53% of its coastal saltmarshes since 1832, and the Boston area had lost a staggering 81% since the late 1700s.


Land-claim of mangroves and coastal saltmarsh is often prompted by a desire to create new land for agricultural or industrial/urban uses, and the consequences vary according to which aim is pursued. Certainly, large areas of the Victorian coast have been destroyed for urban housing (Figure 1.37) and, even in areas where coastal saltmarsh remains, individual houses can very nearly extend into the remaining marsh (Figure 1.38).

Figure1.37: Land-claim of coastal ecosystems, including parts that were previously saltmarsh, for urban housing, Cheetham, western shore of Port Phillip Bay.

Figure1.38: Housing extending close or into coastal saltmarsh, Corner Inlet.

In the case of land-claim for agricultural and some industrial purposes, it may be possible for the modified saltmarsh to retain significant ecological values. The value of Port Phillip Bay saltmarshes that had been modified for the production of salt or shell-grit has been referred to previously. Land-claim for any purpose, however, usually involves infilling to raise the soil profile to prevent flooding and the construction of levees, bunds or sea walls to prevent future ingress of seawater; in these cases the modified landscape usually retains few or none of its original ecological values. Figure 1.39 shows an example of this type of severe land-claim: parts of the original coastal saltmarsh at Hastings (Western Port) were originally used as a council rubbish tip and then used for a sporting ground.

Figure1.39: Areas of former coastal saltmarsh which are now sporting grounds and, previously, were used as a council rubbish tip, Hastings.


Losses may not be as severe as shown in Figure 1.39. Figure 1.40, as an example, shows soil and other material dumped on a saltmarsh near Tooradin. Degradation following land-claim for industrial/urban use (Figures 1.37 & 1.39) or the simple dumping of material (Figure 1.40) is often compounded by the use of poor-quality materials for fill (e.g. contaminated rubble and building waste) with the possibility of their leaching of noxious substances into tidal waterways, and the later imposition of polluting industries (e.g. petrochemical, steel-making etc.) on the reclaimed land.

Figure1.41: Fragmentation of saltmarsh after road construction, The Spit Nature Conservation Reserve.

Figure1.40: Dumped spoil and rubbish, near Tooradin.

In many cases, the claiming of land in coastal areas, especially for urban or industrial uses, involves the creation of roads and other access structures; road building can have a direct impact in terms of lost land as well as an indirect impact in terms of fragmentation of saltmarsh habitat and alteration to tidal flows (Figure 1.41).


French (1997) provided an extensive review of the effects of bunds and sea walls on ecosystem processes in coastal areas, with a strong emphasis on English examples and on coastal geomorphology. Originally the sea walls were built to reclaim inundated coastal land, for agricultural purposes. Such ‘land claims’ had ecosystem-wide impacts, including increased tidal range upstream, increased back-scour along the barriers, prevention of natural habitat adjustment in the most landward areas, and changes to currents, wave propagation and sediment movement (and deposition). Ironically, the construction of sea walls to defend land further landward against seawater ingression may be an inappropriate response to ‘protect’ such lands. As noted by French (1997, page 90):

a wide, shallow intertidal flat leading into a saltmarsh is much better at dissipating energy than a narrow, steep profile. The added advantage here is that the marsh, covered with vegetation, has a high bed roughness and so its frictional drag is higher than that of a mudflat. Hence, the best defence for a piece of land is a mudflat and salt marsh.

The construction of roads and levees within coastal saltmarsh leads inevitably to habitat fragmentation unless ameliorating structures are positioned carefully to allow tidal inundation. Studies of non-coastal ecosystems across Australia have shown how fragmentation adversely affects ecological values of the remaining system (see Hobbs & Yates 2003). The same argument may hold for coastal saltmarsh. Laegdsgaard (2006) noted that, in New South Wales, large saltmarshes provided habitat for substantial populations of crabs and molluscs that were are absent from smaller saltmarshes. As the patch size was reduced, food webs were proposed to become increasingly limited to small predators and involve fewer taxa of insects. Laegdsgaard (2006) further argued that the migratory birds attracted to large saltmarshes probably chose those sites because, in part, they offered good visual protection from predators. Connolly (2009) argued that fragmentation would have adverse impacts on the fish fauna of saltmarsh along the New South Wales and southern Queensland coasts. Studies of brackish tidal marshes in Connecticut (USA) have shown that fragmentation adversely affected the abundance of bird species that selectively used short-grass meadows. On the basis of palaeobotanical studies of coastal systems in the United Kingdom, Long et al. (2006) argued that coastal marshes’ resilience to sea-level rise would be contingent in large part on the size of the system affected.

Although Laegdsgaard (2006) argued that fragmentation would have adverse impacts on coastal saltmarsh vegetation, there are some reasons for believing that saltmarshes may fare better than many terrestrial ecosystems if fragmented. Coastal saltmarsh is naturally fragmented, geomorphologically young, its biota strongly constrained by a suite of environmental factors (cf. complex competitive interactions which contribute more to pattern and process in other ecosystem), structurally simple, and consists of relatively few species, many of which are presumably highly dispersible. Clearly, the likely impacts of fragmentation on the ecological condition of coastal saltmarsh requires further research.

While complete destruction is the most severe form of change, other forms of habitat modification can occur in saltmarshes. Saltmarshes are commonly used, for example, as dumping grounds, often for vehicles (Figures 1.42 & 1.43).


Figure1.42: Dumped motor vehicle, Western Port.

Figure1.43: Remnants of dumped car bodies, Western Port. This material has since been removed.


fire

Fire is another potential habitat-modifying process, but is a rarely reported phenomenon in saltmarsh and the vegetation is generally thought to be more-or-less resistant to fire because of its succulence. The saltmarsh flora in general, however, is not well fire-adapted (Figure 1.44). In this aspect it is dissimilar to much other Australian vegetation, which has a wide range of fire-adaptation strategies including canopy-stored or soil-stored seedbanks and a capacity of adults to resprout after fire (Gill 1981). Fire is apparently lethal to many saltmarsh species, particularly the succulent chenopods (e.g. Sarcocornia and Tecticornia). There are few studies of the impacts of fire on coastal saltmarsh in the international literature, although we note the report by Isacch et al. (2004) that fire in an Argentinian saltmarsh resulted in changes to both vegetation and the bird fauna, and that recovery responses varied greatly across different species.

Peat fires, when the organic-rich substrate ignites and burns, are well known in south-western Australia (Horwitz et al. 1999, 2003). They are also not uncommon in south-eastern Australia, particularly in Melaleuca squarrosa swamp scrub. Peat ignites when it is exceptionally dry, so peat fires could be expected to be more common under drought conditions. A fuel reduction burn on northern French Island in the early 2000s resulted in spotting of fire into saltmarsh where several areas of peat burnt (M. Douglas, Parks Victoria, pers. comm.). The fires were extinguished by Parks Victoria. Although similar events are probably very rare in saltmarsh, it is possible that the frequency of peat fires may increase under climate change. Interestingly, the peat fire described by Horwitz et al. (1999) also arose from a management fire.

Figure1.44: Fire in coastal saltmarsh, Salt Swamp, Barwon River estuary.


mangrove encroachment

Many studies have reported the progressive encroachment of mangroves into saltmarsh, particularly since the turn of the 20th century (Mitchell & Adam 1989a,b; Saintilan & Hashimoto 1999; Saintilan & Williams 1999, 2000; Saintilan & Wilton 2001; Wilton 2001; Hawroth 2002; Rogers et al. 2005; Saintilan et al. 2009). In her review of Australian saltmarshes, Laegdsgaard (2006, page 389) referred to mangrove incursion as an ‘imminent threat’ to coastal saltmarshes. That review had a strongly New South Wales–southern Queensland content, and it has yet to be shown that the situation is as critical in Victoria as it is in these northern states. Rogers et al. (2005), for example, concluded that there was less mangrove encroachment into coastal saltmarsh in Western Port than has been reported for comparable sites along the New South Wales coast. Even so, it is not unusual to find mangrove seedlings establishing into areas of coastal saltmarsh in Victoria (Figure 1.45), although in some locations altered hydrological and salinity regimes can also result in the death of mangroves (Figure 1.46).

Figure1.45: Mangrove encroachment into saltmarsh, Stony Creek Backwash, Melbourne. Mangroves were planted in the area in the 1980s.

Figure1.46: Death of mangroves after ponding created by sand movement, southern French Island.


Wilton (2002) investigated the encroachment of mangroves into saltmarsh along the Georges River near Sydney. Minor changes to tidal inundation, caused by upstream and local engineering works, were one explanation. Slight rises in sea level since the 1950s were also invoked, as were small but significant changes in climate, especially increases in night-time temperatures. Until the mid 1940s, cattle were grazed on a number of the affected saltmarshes, and the removal of cattle may have resulted in encroachment by Avicennia marina.

The topic was extensively reviewed by Rogers et al. (2006), who concluded that the landward encroachment of mangroves into saltmarsh along the coast of south-eastern Australia was facilitated by local factors which contributed to the compaction of saltmarsh soils during drought periods. Data generated under the auspices of the first National Land and Water Resources Audit (NLWRA) provide circumstantial evidence for the importance of local factors on the encroachment into saltmarshes by mangroves. The OzCoasts (2008) information sheet, making use of NLWRA data, showed that the proportion of saltmarsh:mangrove area in tide-dominated deltas and tidal creeks progressively decreased in favour of mangroves as one moved from near-pristine, to largely unmodified, to modified, then severely modified systems. The progressive change was interpreted in terms of the effect of increased sediment load and consequential rapid infilling in modified estuaries.

excessive freshwater inputs

Since the zone immediately behind saltmarshes is commonly developed for urban uses, it often receives large amounts of stormwater and other types of freshwater runoff. Although most analyses of pollution refer to contamination with particular chemical substances, in the case of coastal saltmarsh an argument can be mounted that freshwater flows themselves are potential pollutants.

Increased freshwater inflows can change saltmarsh geomorphology, the biogeochemistry and bioavailability of contaminants, and the productivity and reproductive success of the vegetation. Increasing freshwater flows, for example via channelisation, have been reported to cause scouring and resuspension of sediments, which result in the oxidation and release of toxic metals (Boesch et al. 1999). Channelisation also results in more localised tidal inundation, which can bring tidally-borne contaminants deeper into the wetland. Decreased salinity has been reported to increase selenium bioavailability; increased salinity can increase the uptake of lead by estuarine biota. Although the uptake of copper seems not to be directly affected by salinity, copper bioavailability is affected strongly by interactions with dissolved organic compounds, and these are often brought into coastal environments with inflows of fresh water (Boesch et al. 1999).

The input of stormwater may change species composition in the vegetation. Phragmites australis, Typha orientalis and Bolboschoenus caldwellii, for example, have displaced more salt-tolerant taxa in south-western Western Australian saltmarshes that had been flushed with stormwater (Adam 2000, 2002). Beare & Zedler (1987) reported a similar process for Californian saltmarshes, where Typha domingensis was the invasive, and salt-intolerant, species. Boesch et al. (1999) provided a further example of possible impacts arising from increased freshwater inflows: in the example of Charlotte Harbour (Florida, USA), a change from sheet to channelised flows following the construction of channels within a saltmarsh resulted in Typha replacing mangroves as the dominant vegetation. Note that Phragmites australis is a serious weed in coastal saltmarshes of eastern North America, where it can invade into estuarine areas with salinities up to about one-half those of seawater (Meyerson et al. 2009).


In Victorian coastal saltmarshes, increased freshwater inputs, mostly from residential and industrial sources, have caused significant changes in vegetation in a number of locations. Coastal saltmarsh on the western shores of Port Phillip Bay and the Bellarine Peninsula has been transformed into brackish and freshwater wetlands dominated particularly by Typha domingensis and T. orientalis, Phragmites australis, Schoenoplectus tabernaemontani and Bolboschoenus caldwellii, often to form monospecific stands (see Carr et al. 2001; North et al. 2007; Carr & Kershaw 2009). Perhaps the most spectacular example is at Point Henry near Geelong, where several hectares of saltmarsh have been transformed into a permanent freshwater Cumbungi (Typha spp.) swamp which has been subsequently colonised by willows *Salix cinerea (Carr 1992). The freshwater input is from runoff over several hectares from a nearby industrial complex.

Carr & Kershaw (2009) described and mapped vegetation changes in saltmarsh on the western shore of Port Phillip Bay that was subject to freshwater inputs from adjoining sewage-treatment ponds (elevated ~1–2 m above the saltmarsh) at the Western Treatment Plant near Werribee. Over ~40 years, large areas of lower saltmarsh have been transformed to herbaceous communities dominated by Phragmites australis, Typha domingensis, Schoenoplectus tabernaemontani and Bolboschoenus caldwellii in particular, as well as a few exotic species, notably *Atriplex prostrata and *Puccinellia fasciculata. Since recent decommissioning and draining of the sewage ponds for conservation purposes, tidal inundation has killed these glycophytic species and sites have been colonised by Suaeda australis, an opportunistic halophyte common in the unmodified saltmarsh vegetation. The trajectory for the vegetation now is recovery to the former floristic composition and structure of the lower saltmarsh vegetation. These results indicate that reinstatement of saltmarsh vegetation indirectly destroyed by freshwater inputs is readily achievable.

nutrient enrichment and eutrophication

Nutrientsinestuaries

Estuaries are generally productive systems, even under pristine conditions (McComb 1995). When in its original state, the vegetation in a catchment is highly effective in trapping and recycling nutrients and the streams that leave the catchment and flow through estuaries into the sea tend to have low concentrations of plant nutrients such as nitrogen and phosphorus. Estuarine biota are therefore well adapted to trapping the nutrients that enter via stream flow, and it is likely that nutrient availability controlled the growth rate and biomass accumulation of plants in many estuarine settings before the catchments were disturbed by human activity (McComb 1995). Indeed, Harris (2002) argued that the export of nutrients from a catchment would remain low as long as more than about 50% of the original native vegetation was retained. Once clearing exceeded ~50%, horizontal and vertical movements of water and nutrients across the landscape increased sharply, and the export of nutrients to the sea then increased exponentially.

As noted by McComb (1995) ‘…from this perspective it is not so surprising to find that increased nutrient loading has such profound consequences [in estuaries]’. Given their position on the coast, commonly near large population centres, many estuarine systems in both the developed and developing worlds are severely polluted with excess nutrients. In the USA, for example, it has been estimated that ~60% of coastal rivers and bays are moderately to severely degraded by nutrient pollution (Howarth et al. 2002). In fact, reviews examining the relative importance of various anthropogenic factors in degrading Northern Hemisphere estuarine systems have repeatedly shown that eutrophication is the most important single stressor (e.g. Valiela


et al. 2000; Carpenter et al. 1998; de Jong et al. 2002; Howarth et al. 2002; Kennish 2002; Smith 2003; Serveiss et al. 2004; Howarth & Marino 2006; Bricker et al. 2007). The limited amount of information available for Australian estuaries suggests that many, especially those near large urban centres, also are adversely affected by excessive nutrients (Davis & Koop 2006). In some, perhaps many, cases the adverse effects of nutrient enrichment became evident mostly in the 1950s and 1960s, when population started to rise rapidly and industrial land use become more common (see Kemp et al. 2005).

Nutrientsandtheconceptofeutrophication

Nutrient enrichment is often linked with the concept of eutrophication and it is worth teasing out the two notions. Reynolds (2003) argued that the term eutrophication confounded two interrelated components: • Enrichment of biological systems by nutrients, especially nitrogen and phosphorus• Enhanced production of algae and higher plants that often results from the increased concentration of

plant nutrients.

There is little practical value in defining eutrophication solely in terms of an increase in nutrient concentrations (e.g. as done by Charlier & Lonhienne 1996) unless there is a clear management implication of these increases, viz algal blooms or undesirable shifts in dominant plant communities. Conversely, a discussion of eutrophication of aquatic systems cannot refer to increases in primary production alone (e.g. as proposed by Nixon 1995) without a clear link being made also to the factor overwhelmingly responsible for these changes, viz increased nutrient loads.

Eutrophication has many negative effects on coastal aquatic systems (see reviews by Carpenter et al. 1998; de Jong et al. 2002; Smith 2003). The increased growth of undesirable algae and aquatic weeds interferes with commercial and recreational fishing, as well as other recreational uses of the water body. Plants of most concern are filamentous green algae and phytoplankton, especially cyanobacteria and dinoflagellates, the latter being responsible for red tides and sometimes contamination of shellfish. Oxygen shortages caused by the death and decomposition of excessive plant growth can result in the death of fish and macro-invertebrates. Eutrophication also contributes to general decreases in the biodiversity of receiving waters. It may interfere with biogeochemical processes in the sediments, particularly with increase rates of nutrient flux under anoxic conditions, which leads to a positive-feedback loop that exacerbates the original problem with higher external nutrient loads.

Nutrientlimitation

The nutrients implicated most commonly in eutrophication of aquatic systems are nitrogen and phosphorus. A nutrient is regarded as limiting if additional amounts result in an increased rate of net primary production (Howarth 1988). The paradigm developed in the 1970s suggested that nitrogen was the limiting nutrient in marine systems and phosphorus the limiting nutrient in freshwater systems (see Ryther & Dunstan 1971; Schindler 1977). The position of estuarine systems was ambiguous (Howarth & Marino 2006), and it was usually left unstated that the generalisation referred to planktonic algae and rarely, if ever, to aquatic macrophytes, either submerged or emergent.

Although nitrogen or phosphorus can be limiting alone or even simultaneously in different estuaries and temporal shifts can occur in the limiting nutrient across seasons or years in a single estuary (Howarth 1988; Cloern 2001; Davis & Koop 2006), it remains mostly true that planktonic primary productivity in temperate


estuaries is nitrogen-limited (Howarth & Marina 2006). Whether nitrogen or phosphorus is the limiting nutrient in a given circumstance is thought to be depend on three main factors: i) the ratio of nitrogen to phosphorus in incoming sources, especially riverine waters; ii) preferential losses from the water column or sediments, especially denitrification; and iii) rates of N2 fixation (Howarth 1988). The fact that nitrogen tends to be the limiting nutrient in temperate estuaries is generally ascribed to low rates of N2 fixation when salinities exceed ~10–12 g L–1 (Howarth & Martin 2006) and rapid rates of denitrification in the sediments (Seitzinger 1988; Harris 2001). Few studies have considered the potential role played by groundwater in estuarine nutrient dynamics (e.g. Valiela et al. 1990), but at least one (Krest et al. 2000) showed that saline aquifers could contribute significantly to both the import and export of nutrients from a coastal saltmarsh. Moreover, the role of atmospheric deposition of nutrients has been neglected until recently. Evidence is now accumulating which indicates that aerial deposition of nitrogen, especially on estuarine systems near large industrial and population centres, can be a significant component of the total nitrogen budget (see Hicks et al. 2000; Howarth et al. 2002; Tappin 2002).

Nitrogencyclingincoastalwetlands

The following review is based largely on Boon (2001, 2006); Joye et al. (2009) and Tobias & Neubauer (2009) reviewed the biogeochemistry of tidal flats and saltmarsh, respectively, and their articles provide additional information specific to these two types of estuarine environment. The biogeochemistry of mangroves, mostly from the perspective of tropical systems, has been covered by Alongi et al. (1992) and Twilley & Rivera-Monroy (2009). The topic of biogeochemistry is not merely of interest to environmental chemists but affects the depth of understanding of how these ecosystems function at almost every scale and ecological level. As Alongi (2009, page 628) argued: ‘Highly evolved and energetically efficient plant-soil-microbe relations are a major factor in explaining why mangroves can be highly productive in a typically harsh environment’.

Organic nitrogen (i.e. nitrogen in the form of proteins, nucleic acids and cellular osmotica) is mineralised to ammonium (NH4

+) by a diverse group of heterotrophic bacteria when organisms die. As noted earlier, halophytic plants often contain large amounts of nitrogen-containing cellular osmotica in order to maintain a suitable water balance. The ammonium liberated by heterotrophic bacteria and other benthic microbes (especially flagellates and ciliates) can then be taken up (assimilated) by plants and bacteria in order to synthesis nitrogen-containing chemicals in the cell. Another portion is reversibly adsorbed to inorganic particles, particularly to negatively charged clays in sediments. Since the adsorption is a reversible process, ammonium may be released from sediment particles to the overlaying water along a simple concentration gradient, depending on the relative concentration differences between interstitial waters and the water column. Under alkaline conditions the ionised ammonium ion (NH4

+) is converted to un-ionised ammonia gas (NH3) and may be lost to the atmosphere as a diffusive flux of NH3 from the water column. Under aerobic conditions some ammonium is oxidised to nitrite (NO2

–) by autotrophic nitrifying bacteria, which use the energy liberated by oxidising nitrite to fix CO2 into cellular materials. Another type of nitrifying bacteria oxidise nitrite to nitrate (NO3

–); both groups require O2 for the oxidation.

The nitrate and nitrite produced by nitrifying bacteria can then be used as an oxidant (an electron acceptor) by another group of bacteria, the denitrifiers. Under anaerobic conditions, denitrifying bacteria use nitrate in lieu of oxygen to oxidise the organic substrates provided by dead plants and animals. The coupling of nitrification and denitrification takes place in the interfaces between aerobic and anaerobic sites in wetland sediments, such as around the roots of emergent plants and the holes of burrowing animals. The process


occurs in these locations because the conversion of ammonium to nitrate by nitrifying bacteria requires oxygen, whereas the conversion of nitrate to N2 gas takes place mostly under anaerobic conditions. Coupled nitrification-denitrification can account for the loss of large amounts of nitrogen (as N2 gas) from aquatic systems. A small amount of nitrous oxide (N2O) is produced during bacterial denitrification; nitrous oxide is a potent greenhouse gas and so this conversion has implications for greenhouse-gas emissions from wetlands. Counteracting the loss of nitrogen through coupled nitrification-denitrification is the fixation of N2 by prokaryotes such as cyanobacteria and bacteria.

Nitrate that is not denitrified can be taken up by plants as a nitrogen source. Ammonium, however, is the preferred source for most plants because it is in the same reduction state (–5) as the amino acids that plants use to build proteins. Nitrate not assimilated is readily lost from aquatic systems because it is negatively charged, unlike the positively charged ammonium. Because of its charge, nitrate is not adsorbed readily onto negatively charged clay particles and so can easily diffuse into the overlaying water or be leached downwards deeper into the soil profile. Moreover, nitrate is a strong oxidant and is quickly used by denitrifying bacteria. Thus the concentration of ammonium is almost always far greater than that of nitrate or nitrite in wetland sediments.

Because nitrate can be used as an alternative oxidant by many bacteria when oxygen is not freely available, the nitrogen cycle cannot be separated conceptually from the carbon cycle. It is an important consideration when evaluating the potential for wetlands to intercept nutrients, because wetland plants and the holes of burrowing animals (e.g. crabs and callianassid shrimps) provide the aerobic-anaerobic interfaces in which coupled nitrification-denitrification can take place, as well as the carbon substrates used by the denitrifying bacteria (see Bird et al. 2000 for an example from Western Port).

It should be clear that the nitrogen cycle in coastal wetlands is extremely complex and involves multiple changes in oxidation state across different forms of nitrogen (e.g. from highly reduced ammonium to fully oxidised nitrate). Nitrogen cycling is linked closely with carbon cycling (e.g. the oxidation of complex organic substrates and the production of reduced forms of carbon, such as methane) and sulfur cycling. There are a number of processes by which nitrogen can be lost from coastal wetlands, including: i) denitrification and the loss of N2 gas (and, to a lesser extent, N2O); ii) ammonia volatilisation and the loss of NH3 gas to the atmosphere; iii) leaching of nitrate; and iv) exports of organic nitrogen in the form of plant or animal tissue during tidal exchange or the migration of shore-feeding birds.

Phosphoruscyclingincoastalwetlands

Phosphorus is present in wetlands in a wide variety of forms, including organic and inorganic forms, and dissolved, adsorbed and particulate forms. Unlike nitrogen, phosphorus does not have a number of oxidation states; the significance of this is that none of the transformations in the phosphorus cycle involve reductions or oxidations. Phosphorus in wetland sediments occurs almost entirely in the +5 valency state, and all lower oxidation states are thermodynamically unstable and readily oxidise to PO4

3– even under highly reducing conditions (Vymazal et al. 1998). This is a critical point of difference between the phosphorus and nitrogen cycles; rather than alternating oxidation and reduction as occurs in nitrogen cycling, almost all steps in the phosphorus cycle are shifts between inorganic and organic forms, or between dissolved forms, particulate forms and adsorbed forms.


Because of the lack of alternating oxidation-reduction steps, phosphorus is not lost from mangroves or coastal saltmarsh sediments via metabolic gas-producing processes such as denitrification or ammonium volatilisation. The possible exception is via the production of phosphine gas, PH3, at extremely low redox potential, but this is unlikely to be a major pathway for phosphorus loss from most wetlands (Boon 2006). Thus there is no obvious route for permanent removal of phosphorus from coastal wetlands other than by adsorption to inorganic particles, burial of organic detritus in the sediments, or export with the tides or with migrating animals. Biological uptake by plants, biofilms and microbes in the sediments are three short-term removal processes for phosphorus in wetlands (Brown et al. 1998). Long-term removal processes can be effected by chemical precipitation, adsorption to sediments and burial in litter or peat.

The two main types of mechanism regulating uptake and release of phosphorus in wetlands are: i) physico-chemical reactions dominated by adsorption/desorption processes and controlled largely by pH and redox conditions; and ii) biological processes, dominated by bacterial activity, responsible for shifts between inorganic and organic forms (McComb & Qiu 1998). Adsorption/desorption processes are important for two reasons: i) phosphorus tends to precipitate in the presence of divalent cations (e.g. Ca2+, Mg2+) or ferric iron (Fe2+) at neutral or alkaline pH; and ii) the reduced (ferric) form of iron is far more soluble than the oxidised ferrous (Fe3+) form. The reaction of phosphorus with iron has particular significance for phosphorus cycling in wetlands, since phosphorus complexed with Fe3+ is largely biologically unavailable. Only when wetland sediments become anaerobic can phosphorus be released; under aerobic conditions, the phosphorus is held tightly within the sediment or soil in iron oxyhydroxides. The black colour of mangrove and saltmarsh sediments indicates that they are anaerobic and that iron exists in them in the reduced (Fe2+) form.

Effectsofnutrientenrichmentoncoastalsaltmarsh

Almost all studies of nutrient enrichment and eutrophication of estuarine systems have been concerned with phytoplankton growth. Some work has addressed the problem of excessive growth of macro-algae and death of seagrasses consequent to nutrient enrichment (e.g. Harlin 1995; Fletcher 1996; Raffaelli et al. 1998; Burkholder et al. 2007), but few studies have examined whether nutrient enrichment will have adverse effects on coastal saltmarsh or mangroves. That neglect is somewhat surprising, as coastal saltmarshes are ‘particularly vulnerable to eutrophication because estuaries concentrate nutrients from the entire upstream watershed’ (Sharitz & Pennings 2006, page 209).

The research undertaken to date has been concerned largely with nitrogen cycling in Spartina alterniflora marshes of the Northern Hemisphere, and it was done mostly in the late 1970s and 1980s (e.g. Haines et al. 1977; Teal et al. 1979; Valiela & Teal 1979; DeLaune et al. 1980; Whitney et al. 1981; Smart 1982; Peterson et al. 1987). Some early studies addressed nitrogen cycling in European marshes (e.g. Abd. Aziz & Nedwell 1979, 1986a,b; Henriksen & Jensen 1979; Jensen et al. 1985) but more recent work has started to examine the effects on nutrient enrichment on other saltmarsh plant species and in other locations (e.g. Boyer et al. 2001). The early studies commonly showed that primary productivity and biomass accumulation in Spartina alterniflora was limited by nitrogen availability, which led to the generalised notion that saltmarshes could absorb excess nutrients because they would simply be tied up in increased plant biomass. Even in the early 2000s it was not uncommon to find reports arguing that coastal saltmarshes could have a role in ameliorating the catchment-derived loading of nutrients into estuaries (e.g. see Davis et al. 2004).


More recent work has not supported that assumption, and it is now clear that nutrient enrichment affects much more than the mere standing crop of vascular plants in coastal wetlands. Deegan et al. (2007), for example, showed that even mild nutrient enrichment resulted in ecological responses at several trophic levels in a Northern Hemisphere coastal saltmarsh. Similarly, Crain (2007, page 26) concluded that eutrophication by nitrogen or phosphorus had the:

…potential to greatly reduce the characteristic high diversity of oligohaline marshes. Inputs of both nutrients in coastal watersheds must be managed to protect the diversity and functioning of the full range of estuarine marshes.

Moreover, almost all studies of the effect of nutrients on coastal saltmarsh have examined only the response of above-ground biomass. In a noteworthy study, Turner et al. (2009) showed that nutrient enrichment of a North American coastal saltmarsh resulted in a reduction in the amount of below-ground biomass and carbon accumulated in the sediments. They concluded that:

…the net effects of 36 yr of nutrient enrichment in replicated field experiments do not lead to higher organic or inorganic accumulation. Enrichment reduces organic matter belowground and may result in a significant loss of marsh elevation equivalent to about half the average global sea level rise. Sustaining and restoring coastal emergent marshes is more likely if they receive less, not more, nutrient loading (Turner et al. 2009, page 1634).

The early emphasis on effects of nutrient enrichment on primary production also failed to recognise that eutrophication was likely to have effects on a wide range of ecological interactions in coastal saltmarsh. Three types of ecological effects stand out: i) competitive interactions between plant species; ii) rates of herbivory and the relative roles of bottom-up versus top-down control of plant biomass; and iii) effects on biota other than plants, especially on ecosystem-scale biogeochemical processes and microbial dynamics.

Competitive interactions between plant species

Nybakken (2001) argued that nutrient enrichment would cause changes in the competitive ability of different species in North American coastal saltmarsh via effects on the allocation of above-ground and below-ground biomass and the variable response of different taxa to physiological stress (Figure 1.47). Other studies have confirmed that nutrient enrichment does affect competitive relationships between different plant species in coastal saltmarsh: see Pennings et al. (2002) for the relationship between Spartina alterniflora and Distichlis spicata in Atlantic coast and Gulf coast saltmarshes of the USA; Rickey & Anderson (2004) for Phragmites australis and Spartina pectinata in Illinois; Tyler et al. (2007) for Spartina alterniflora, hybrid Spartina and Salicornia virginica in a west-coast USA saltmarsh; and Hunter et al. (2008) for Distichlis spicata and Salicornia bigelovii in Mexico.


Figure1.47: Changes in competitive interactions among plant species in North American saltmarshes with increasing nutrient availability and environmental stress. Source: Nybakken (2001, Figure 8.26).

Rates of herbivory and top-down control of biomass

Nutrient enrichment also alters the rate of herbivory and the relative role of top-down (herbivory) versus bottom-up (nutrient limitation or soil aeration effects) on biomass accumulation by saltmarsh plants. Not unexpectedly, plants in nutrient-enriched sites are often more palatable to herbivores than those in nutrient-poor sites. Sala et al. (2008), for example, used a factorial design to show that nutrient enrichment so increased insect herbivory on Spartina alterniflora in a New England (USA) marsh that the standing biomass was reduced by nearly one-half compared with control sites. Moreover, there was a greater incidence of secondary fungal infections in grazer-induced wounds in the nutrient-enriched sites. They concluded that eutrophication could trigger top-down regulation of biomass even though bottom-up (nitrogen limitation) was the principle controlling factor under non-fertilised conditions. Similar results were reported by Bertness et al. (2008) but somewhat contradictory findings by McFarlin et al. (2008). In a study of top-down control of saltmarsh


biomass not involving nutrient manipulations, Silliman & Bertness (2002) concluded that rapid rates of grazing by invertebrate herbivores (a periwinkle in this case) could ‘…convert one of the most productive grasslands [a Spartina-dominated marsh] in the world into a barren mudflat within 8 months’ (page 10,500).

Ecosystem-scale biogeochemical impacts

Almost all the early studies of nutrient impacts on coastal saltmarsh were concerned with effects on net primary production and standing crops of the visually obvious vascular plants. Attention later shifted to include below-ground biomass and to the relative importance of bottom-up versus top-down control. Even so, until recently studies were still overwhelmingly concerned with plant responses. There is no reason to assume that other saltmarsh biota will be affected by nutrients in the same way as vascular plants. In one of the few studies on biogeochemical and microbial effects of nutrient enrichment, Sundareshwar et al. (2003) demonstrated that bacterial productivity in a North American coastal saltmarsh was limited by phosphorus availability, even though primary productivity by the vascular plants was nitrogen-limited. Thus individual trophic groups may well respond differently to nitrogen and phosphorus enrichment in coastal wetlands. Most recently, Bowen et al. (2009) showed that bacterial productivity was not increased by nitrogen fertilisation in high-marsh Spartina alterniflora sites in North America.

Whatisan‘acceptable’nutrientloadtocoastalwetlands?

An issue that often confronts those charged with managing coastal wetlands is the setting of nutrient targets or maximum acceptable loads. The US Environmental Protection Agency released its Draft Nutrient Criteria. Technical Guidance Manual for nutrients in wetlands in 2006 (United States Environmental Protection Agency 2006), which updated an earlier report of 1998 (United States Environmental Protection Agency 1998). Unlike almost all other approaches previously developed for wetlands, the 2006 review analysed sediment nutrients as well as nutrients in the water column, adopted a bioregional perspective, and stressed the role of nutrient loading rates as well as mere concentrations in the receiving wetland. Unfortunately, it did not set any guidelines for nutrient concentrations or loadings.

The situation with ‘acceptable’ nutrient loads to Victorian estuaries is similarly undeveloped. The Victorian EPA has recently released draft water-quality guidelines for riverine estuaries (Environment Protection Authority Victoria 2009). Preliminary guideline values for annual median concentrations of Total Nitrogen in estuary waters were set at 0.5 and 0.6 mg N L–1 for surface and bottom waters, respectively; concentrations for Total Phosphorus were set at 0.05 and 0.07 mg P L–1, respectively.

When ‘acceptable’ nutrient loads to coastal wetlands are discussed, the concept of assimilative capacity is often invoked to show that systems such as mangroves and saltmarsh can ‘process’ such inputs on an indefinite basis. Assimilative capacity is the assumption that aquatic systems can assimilate wastes without their ecological community being damaged or changed to an unacceptable degree. The assumption has a number of theoretical and practical limitations (Campbell 1981, 1986). First, it requires a value judgement as to what constitutes ‘unacceptable change’. Campbell (1986, page 154) was a rigorous critic of the idea and argued that ‘…by disguising value judgements as technical or scientific decisions, public criticism of controversial resource management decisions can be deflected’.

Second, it assumes that coastal wetlands are able to assimilate additional nutrients on a long-term basis. There are some ecological systems that can assimilate external nutrient loads on a long-term basis. In Australia,


such systems might include that set of estuaries and coastal lagoon systems that have very rapid rates of denitrification in the sediments, which will lead to the efficient loss of bio-available nitrogen and strong nitrogen-limitation of plant growth on a long-term basis (Harris 2001). Nitrogen cycling in Port Phillip Bay provides a good example of such a process: in this coastal embayment denitrifying bacteria, associated with the holes of burrowing callianassid shrimps in the sediments, reduce ammonium and nitrate from the catchment to N2 gas at rates sufficient to maintain phytoplankton in a state of chronic and severe nitrogen limitation (CSIRO 1996). Such long-term processing of an external nutrient source, however, is possible only because bacterially mediated denitrification converts biologically available nitrogen to biologically unavailable N2 gas, a biogeochemical transformation that results in a permanent loss of large amounts of nitrogen from the ecosystem. The process is highly sensitive to both the load of inorganic nitrogen and organic matter. If disrupted, it is unlikely to recover and the consequences will be the uncoupling of nitrification from denitrification, resulting in nitrogen accumulating in the water column, the release of algae from chronic nitrogen-limitation, and severe and probably irreversible algal blooms (Graham Harris, CSIRO, pers. comm. 15/06/2006).

Similar long-term, sustainable processes of nutrient removal is much more difficult to achieve with phosphorus because there is no gaseous form analogous to N2 (other than phosphine PH3). In other words, there is no biogeochemical process in the phosphorus cycle that equates to denitrification in the nitrogen cycle. Instead, the capacity of mangroves and coastal saltmarsh to ‘assimilate’ phosphorus is due almost entirely to the capacity of sediments to bind incoming inorganic phosphorus to iron or deposit it within peat. With long-term loading of substantial amounts of phosphorus, the ability of sediments to bind phosphorus eventually will become saturated and sediments will switch from being a phosphorus sink to becoming a phosphorus source (Havens & Schelske 2001; Bailey et al. 2002).

Conclusions

Coastal saltmarsh is often subject to nutrient enrichment because it is located in or near estuarine systems that, since the 1950s at least, have become progressively eutrophied. The growth of saltmarsh vascular plants is often limited by the availability of nitrogen, but it is unclear whether the growth of other biota (e.g. algae, microbes) is also nutrient-limited and, if so, whether nitrogen is the element of concern. Early proposals that coastal wetlands could be used to ‘assimilate’ nutrients and thus protect estuarine waters from the adverse effects of excessive nutrient loading are largely naive. Nutrient enrichment, far from being a benign event, is likely to have far-reaching impacts on many aspects of saltmarsh ecology.

toxicants

The low position of saltmarshes in the landscape means they are often depositional environments and, as such, highly susceptible to contamination with pollutants, especially particle-associated chemicals that may be deposited onto, then buried within, the sediment surface. The sedimentation of particles in coastal wetlands is enhanced not only by the presence of dense vegetation, but also by the mixing of fresh and saline water: when particles suspended in fresh water come into contact with seawater they are rapidly precipitated and fall to the sediments. Moreover, the position of saltmarshes near favoured areas for modern industrial activities, including petroleum refining, steel mills and ports (e.g. western shores of Western Port and Port Phillip Bay), further increases their susceptibility to pollution (Boesch et al. 1999).


Oilpollution

Oil spills are one of the most obvious forms of pollution in saltmarshes and mangrove vegetation, and a number of factors conspire to make these types of estuarine wetland particularly susceptible to oil pollution. First, saltmarshes and mangroves are often found in areas that have been partially reclaimed for industrial uses, including petroleum refining. Second, saltmarshes and mangroves often surround ports or other areas where ships berth, and thus are exposed to the small but not uncommon spills of hydrocarbons that occur when ships are refuelled (Figure 1.48). Third, saltmarshes and mangrove vegetation are particularly susceptible to impacts from rare but catastrophic large oil spills: Adam (1990) noted that saltmarshes were chronically exposed to oil following the infamous Torrey Canyon, Amoco Cadiz and Metula catastrophes. Finally, low-energy coastal environments, especially saltmarshes, can be highly efficient traps for oil. Their dense vegetation and fine-grained sediments are particularly effective in retaining hydrocarbons, which can then be released progressively after the pollution event.

The noxious effects of oil pollution on estuarine systems are well documented (see review by Kennish 1998). Crude oil is a complex mixture of thousands of different compounds, many of which are toxic and some of which bioaccumulate. About 75% of the weight of crude and refined oil is made up of hydrocarbons with a molecular weight of 16 to 20,000. The remaining 25% is made up of various non-hydrocarbon compounds, typically containing oxygen, sulphur, nitrogen and various metals.

The toxicity of these various constituents varies greatly. Of the four main types of hydrocarbon in crude oil (straight-chain alkanes, branched alkanes, cycloalkanes and aromatics), the straight-chain alkanes tend to be the most acutely toxic and the aromatic compounds the least. Aromatic hydrocarbons, however, can be taken up rapidly by the biota and, because of their high solubility in lipids, may persist for long periods of time in marine animals (Kennish 1998). In contrast, the more water-soluble compounds in crude and refined oil (e.g. benzene, toluene and xylene) are highly toxic to plankton but, because of their low molecular weight, are often rapidly lost to the atmosphere via evaporation after an oil spill. The remaining components that find their way to saltmarshes are subject to photochemical oxidation, often with the creation of polar oxidised compounds from the high molecular-weight aromatic precursors. Oil-in-water and water-in-oil emulsions can be formed following the agitation by waves and currents of the spilt oil, and the viscid masses of water-in-oil emulsions (‘chocolate mousse’: Kennish 1998, page 33) can persist for months along the shore and entangled in vegetation. Finally, the heaviest fractions, consisting of crude-oil tar balls, can be exported to the coast and

Figure1.48: Small port facility adjacent to saltmarsh and mangroves, Port Franklin.


washed up along saltmarshes and mangrove shrublands and forests. If oil penetrates deeply into sediments, the anoxic conditions present in most coastal marine sediments will hinder bacterial and fungal decomposition of the oil. Even so, there is some evidence that bacterial and fungal populations in saltmarsh sediments adapt to oil spills, with petroleum-degrading taxa being selected after exposure to the hydrocarbons (Albrechtsen 1991).

It would seem that multiple oil-pollution events are more detrimental to intertidal marine plants than single but larger events, although short-lived annual plants can easily be killed by a single exposure to oil (Kennish 1998). Kennish (1998) proposed that within five years most organisms in a saltmarsh would have recovered from a catastrophic oil spill, with the exception of long-lived plants. The final stages of recovery should occur within about 10 years, although it may take 20–100 years for full recovery if the affected area was very large and long-lived plants were killed. The speed with which saltmarshes could be expected to recover from an oil spill is longer than that expected for similar events on beaches, tidal flats or rocky shores, for which a full recovery is expected within about five to 10 years. The impacts of oil pollution on coastal saltmarsh are not limited to the vascular vegetation: complex interactions and changes may result because of effects on invertebrates and wetland food webs. Carman et al. (2000), for example, showed that diesel-contaminated saltmarsh showed increased rates of production by benthic algae, due to the mortality of algal grazers and long-term effects on N2 fixation on nitrogen availability to plants.

There is almost no information on the impact of oil spills on Australian mangroves and saltmarshes, and what little information that is available is drawn mostly from the Sydney area (e.g. Anink et al. 1985; McGuinness 1990; Grant et al. 1993; Clarke & Ward 1994). Most information on hydrocarbon pollution comes from Northern Hemisphere research (Baker 1979; see Shin et al. 2001 as an example). Generalisations are difficult to make, because impacts vary according to the type of oil spilt, season of the accident and its relationship to plant life histories, floristic composition of the affected saltmarsh and type of remediation involved. Most work, however, shows that the lighter, more volatile fractions are most toxic to saltmarsh plants and that subsequent clean-up activities can result in further damage to the affected sites (e.g. via toxic effects of detergents used to disperse or emulsify the oil). Impacts vary according to the species of saltmarsh plant exposed to oil: some taxa, such as Salicornia spp. in Europe seem extremely sensitive to oil and, as these species do not have long-lived seed banks, temporary local extinctions could occur if adult plants are killed (Adam 1990).

Willis (1951) reported that an oil spill killed the small population of mangroves in the estuarine Kororoit Creek, Altona. Even so, mangroves were able to recolonise the affected area, probably from the extensive mangrove population at Williamstown, ~1.5 km to the south-east at Jawbone Marine Reserve (Carr et al. 1987; Carr 1988). Hoff (2002) documented the impacts of an oil spill on mangroves in South Australia. In August 1992, the tanker Era released ~300 tonnes of heavy bunker oil, a blend of diesel and heavy residual, into Spencer Gulf, South Australia. Approximately 57 tonnes of oil washed up among the mangroves south of Port Pirie. Due to an extremely high tide, oil penetrated ~50 m into the mangroves, with the result that 75–100 ha of mangrove were oiled. Although clean-up attempts were undertaken near the jetty, no such attempts were made for the mangroves. By November 1992, 2.3 ha of mangroves had suffered extensive defoliation; the affected area expanded slightly to 3.2 ha by 1995 and then stopped increasing. Trees that were totally defoliated did not recover during the four-year period of subsequent monitoring.


Polycyclicaromatichydrocarbons

Often linked with contamination due to oil spills is the contamination of saltmarshes with polycyclic aromatic hydrocarbons (PAH). These compounds are among the most ubiquitous organic pollutants in coastal marine systems (Kennish 1998). Examples of PAH include naphthalene, biphenyl, anthracene, pyrene and creosote. They are often carcinogenic, mutagenic, or both, particularly after metabolic activation by marine organisms. Because they are relatively insoluble in water, PAH adsorb strongly to particles and accumulate in sediments, where benthic organisms take them up. Polycyclic aromatic hydrocarbons originate from a wide range of human activities, but mostly from oil spills, in municipal and industrial effluents, fossil fuel combustion, asphalt production and waste incineration. Although PAH affect organisms through direct toxic effects, they have also a wide range of secondary and indirect impacts on marine life: some organisms (e.g. mussels) cannot metabolise PAH, so the compounds accumulate in their tissues, whereas other organisms (e.g. fish, annelids, some crustaceans) have well developed enzyme systems that deal with the pollutant. Even so, fish exposed to PAH often develop tumours and lesions and suffer from other developmental malformations (Kennish 1998). Meudec et al. (2006) showed that saltmarsh plants can bio-accumulate PAH from contaminated sediments.

Halogenatedhydrocarbons

Halogenated hydrocarbons are among the most persistent, ubiquitous and toxic of all marine pollutants (Kennish 1998). These lipophilic, high molecular-weight compounds can biomagnify through marine food webs and pose direct health risks to humans: examples include DDT, chlordane and polychlorinated biphenyls (PCB). They enter coastal marine systems through urban and industrial runoff, sewage disposal and atmospheric deposition. Most halogenated hydrocarbons are broad-spectrum toxicants that poison entire biotic communities. They readily sorb to particles and therefore usually accumulate in bottom sediments, where residues of the most recalcitrant forms can persist for decades and perhaps centuries (Kennish 1998). The fact that coastal wetlands provide such valuable nesting and foraging sites for wading birds means that contamination with persistent halogenated hydrocarbons is a major risk in these areas: Boesch et al. (1999) gives examples of where pelicans and colonial wading birds have been seriously affected by DDT and heavy metals in North American coastal wetlands.

Heavymetals

There are three broad classes of heavy metals: i) the transitional metals; ii) the metalloids; and iii) organometal compounds (Kennish 1998). The transitional metals (e.g. copper, cobalt, iron, manganese) are often required in trace amounts by plants, animals and microbes for enzyme function, but are generally toxic at relatively high concentrations. In contrast, metalloids (e.g. arsenic, cadmium, lead, mercury, selenium and tin) are not required for metabolic processes and are usually toxic at low concentrations. Organometal pollutants (e.g. tributyl tin, alkylated lead, methylmercury) are highly toxic to marine organisms. The behaviour of metals in coastal marine ecosystems is particularly complex, and processes of metal mobility, chemical speciation and biological availability are influenced by complicated interactions among redox potential, salinity and pH (Williams et al. 1994; Borch et al. 2010).

Heavy metals pose a pollution threat to saltmarshes because of their extreme persistence, high toxicity, tendency to bioaccumulate, and ability to sorb onto suspended particles and accumulate in depositional environments. Moreover, heavy metals adsorb preferentially onto organic matter (Harland et al. 2000) and so


the organic-rich sediments of coastal saltmarshes and mangroves are ideal sites for heavy-metal accumulation. In their review of the behaviour of heavy metals in saltmarshes, Williams et al. (1994, page 277 ) concluded that ‘…salt marshes act as a very efficient sink for metal contaminants although metal concentrations in halophytes do not generally reflect environmental contamination levels’. Because of the threat posed to saltmarshes by heavy metals, there is a reasonably diverse literature on the topic (e.g. see Giblin et al. 1980; Ragsdale & Thorhaug 1980). Some information is available also on heavy-metal concentrations in Australian saltmarsh plants (Foster et al. 2005; Thomson et al. 2007). One conspicuous source of heavy metals in coastal saltmarsh and mangroves may be the leaching of copper, chromium and arsenic from CCA-treated wooden walkways; Weis & Weis (2002) have shown that concentrations of these three heavy metals was markedly higher underneath, and up to 10 m away, from walkways through a saltmarsh in New Jersey (USA).

acid sulfate soils

Acid sulfate soils are soils that contain sulfidic materials and produce sulfuric acid (H2SO4) when exposed to the air (Department of Sustainability and Environment 2009d). In Australia, potential and/or actual acid sulfate soils are found along almost the entire coastline, with the main exception being the steep limestone cliffs of the Great Australian Bight (White et al. 1997). Acid sulfate soils are especially common along the eastern seaboard in estuaries, mangroves, saltmarsh and paperbark (Melaleuca spp.) swamps (Figure 1.49). There are many examples where their disturbance has created severe environmental problems in eastern Australia (e.g. White et al. 1997; Wilson et al. 1999; Johnston et al. 2003).

Figure1.49: Distribution of land that has the potential to contain coastal acid sulfate soils(shown in red) in the South Gippsland region. Source: Department of Sustainability and Environment (2009d, page 36).


The critical component of acid sulfate soils is pyrite (FeS2), a highly insoluble crystalline form of iron sulfide produced (usually within the past 10,000 years) by the reaction of ferrous sulfide (FeS) with sulfur. In coastal areas the ferrous sulfide had been produced in earlier saline or brackish-water swamps, such as mangroves, paperbark swamps and saltmarshes, by sulfate-reducing bacteria oxidising the abundant organic material produced in these highly productive environments. The overall reaction by which sulfuric acid is produced in acid sulfate soils is as follows:

2 FeS2 + 7.5 O2 + 7 H20 2 Fe(OH)3 + 4 H2SO4

The sulfuric acid produced when acid sulfate soils are activated moves through the soil, stripping iron, aluminium and manganese, as well as dissolving, in the worst cases, heavy metals such as cadmium. This noxious mixture makes the soil highly toxic and, combined with the very low pH (< 3), renders the growth of most plants impossible. The rate of production of sulfuric acid is often between 100–300 kg ha–1 year–1 (White et al. 1997), and sufficient acid can be produced from highly acid sulfate soils that it seeps into adjacent waterways, resulting in drastic reductions in pH, massive fish kills and the death of estuarine invertebrates, including economically important species such as shellfish. Fish kills linked to the disturbance of acid sulfate soils have been reported frequently for large estuarine rivers in northern NSW. In the Richmond River, for example, a large flood in 1994 liberated over 1,000 tonnes of sulfuric acid, 450 tonnes of dissolved aluminium and 300 tonnes of dissolved iron from 4,000 ha of inundated catchment. The event acidified a 90 km reach of the river for seven weeks, and the pH fell to less than 3 in some locations. In another example, Johnston et al. (2003) reported that extensive fish kills in the Clarence River estuary of northern NSW were sometimes caused by hypoxic conditions which, in turn, were a result of anoxic, iron-rich surface waters draining from two back swamps with active acid sulfate soils.

Acid sulfate soils generally do not present a serious management problem as long as they remain waterlogged. They become problematic when wetlands are drained, for example when ditches cause the watertable to drop rapidly and surface soils to dry out and oxidise (Department of Sustainability and Environment 2009d). Large spoil heaps, often raised along the edges of such drains, also can produce acid for many years after the drain has been excavated. The release of sulfuric acid from these spoil dumps typically occurs after drought-breaking rains, which raises the watertable back to its original (pre-drought or pre-drainage) level and washes the acid and dissolved metals out of the surface layers of the soil. In many cases, reverting to the earlier hydrological regime is not sufficient to cure the problem, as large volumes of acid may have accumulated in the soil and there may have been irreversible changes to the soil structure due to drying, acidification and oxidation.

The distribution of acid sulfate soils on the Victorian coast beyond the present high tide is determined largely by the extent of submergence by a mid-Holocene sea level that was 1.0 to 1.5 m higher than present. No complete inventory of this potential has been conducted but there is sedimentary and geomorphological evidence for such a higher level in Gippsland, Western Port, Port Phillip and western Victoria (Gill 1971; Gill & Amin 1975; Rosengren 1987, 1988). Potential acid sulfate soils may therefore occur inland of existing saltmarsh on land that has been cleared and drained for agriculture. Indeed, there may have been past episodes of acid-sulfate activation that have not been recognised or recorded.


introduced plants

Wetlands seem particularly susceptible to invasions by exotic plants; in their review of the causes and consequences of weed invasions into wetlands, Zedler & Karcher (2004) noted that wetlands accounted for < 6% of the earth’s area but 24% of the world’s most invasive weeds were wetland species. Weed invasions into coastal saltmarsh, at least from the evidence presented in North American studies, seem to be related closely to anthropogenic disturbance, especially shore-line development, increasing eutrophication and the removal of native vegetation (Bertness et al. 2002).

The presence of exotic plants or exotic genotypes, especially in the upper levels of coastal saltmarshes, has been noted earlier in this report (Chapter 1.5) and is certainly not a feature peculiar to Australian saltmarshes. An aggressively invasive genotype of Phragmites australis, for example, is a serious weed of saltmarshes in North America (Lynch & Saltonstall 2002; Meyerson et al. 2009) and, as noted later in this report, various forms of Spartina are problematic weeds of temperate coastal areas across the world. It is convenient to differentiate between the weed threat posed by Spartina from that posed by other taxa, if only because Spartina is one genus that infests mangroves as well as the very lowest levels of coastal saltmarsh. A separate section of this review is devoted to Spartina (Chapter 1.12).

Numerous exotic plant species have colonised the saltmarshes of south-eastern Australia. As outlined in Table 1.8, of the total saltmarsh flora of 249 taxa found in Victorian mangroves and saltmarsh, 118 (47%) are exotic. Of these, approximately 30 taxa are essentially halophytes, more or less confined to saline environments. The life-form statistics for the total exotic flora are given in Tables 1.7 and 1.8, which shows that 44% are obligate annuals. The exotic annual and perennial invaders have had catastrophic, although largely undocumented, impacts on upper saltmarsh in south-eastern Australia and the situation will worsen.

Table 1.15 lists what we consider to be the 20 worst exotic invasive plant species in Victorian coastal saltmarsh. Tall Wheat Grass *Lophopyrum ponticum (Figure 1.50), currently widely promoted by government agencies for establishment on saline lands as stock fodder, is unquestionably the most serious invader of upper

saltmarsh because of its very broad ecological amplitude and robust life form: a very large tussock grass (Booth et al. 2009). It is by no means confined to saline or moist environments. Other grasses, for example, *Hordeum marinum and *Lolium rigidum, are also problematic weeds, especially of upper saltmarsh (Figure 1.51). Invasion of upper saltmarsh, particularly by annual grasses, was investigated by Carr et al. (2002) at The Spit Nature Conservation Reserve on the western shore of Port Phillip Bay.

Figure1.50: *Lophopyrum ponticum invading saltmarsh vegetation dominated by indigenous Suaeda australis, Puccinellia perlaxa and *Hordeum marinum, near Colac.


Table1.15: The 20 most seriously invasive (and potentially serious) exotic plant species currently in Victorian coastal saltmarshes, August 2008. Species underlined are assessed as the worst invasive species. Life-form abbreviations are as per Table 1.8.

Species Common name Life form Current distribution and abundance†

*Cortaderia jubata Pink Pampas-grass Pt 4

*Cortaderia selloana Pampas Grass Pt 4

*Dactylis glomerata Cock’s Foot Pt 1

*Elytrigia pungens Sea Couch-grass Pr 5

*Festuca arundinacea Tall Fescue Pt 1

*Gladiolus undulata Wild Gladiolus Gc 2

*Hordeum marinum Sea Barley-grass A 1

*Juncus acutus ssp. acutus Spiny Rush Pt 1

*Lophopyrum ponticum Tall Wheat Grass Pt 1

*Limonium hyblaeum Sicilian Sea-lavender Pr 5

*Lolium rigidum Wimmera Rye-grass A 1ØMelaleuca halmaturorum ssp. halmaturorum Salt Paperbark T 6

*Moraea flaccida One-leaved Cape-tulip Gc 6

*Oxalis pes-caprae Soursob Gb 2

*Parapholis incurva Coast Barb-grass A 1

*Parapholis strigosa Slender Barb-grass A 1

*Phalaris aquatica Phalaris Pt 1

*Plantago coronopus Buck’s-horn Plantain Pt 1

*Paspalum vaginatum Salt-water Couch Pr 5

*Spartina spp. (*S. anglica and *S. x townsendii) Cord-grass Pr 3

* Exotic speciesØ Dual origin species (see Walsh & Stajsic 2007) that are indigenous in Victoria as well as invasive outside their natural geographic

distribution† Following codes in Carr et al. (1992): 1 widespread, medium to large populations 2 widespread, small populations 3 limited distribution, medium to large populations 4 limited distribution, small populations 5 rare or localised, medium to large populations 6 rare or localised, small populations

Figure1.51: Upper saltmarsh being invaded by exotic annual grasses, particularly *Lolium rigidum, Lake Victoria.


grazing

Domesticstock

Doody (2008) reviewed the effects of grazing on Northern Hemisphere coastal saltmarsh. It is clear that grazing affects saltmarshes in a large number of ways and, in some cases, may increase plant richness and diversity (see Bouchart et al. 2003 for French coastal saltmarsh). In almost all studies of European coastal saltmarsh, however, the abundance of succulent taxa (e.g. Salicornia europaea, Suaeda maritima) decreased in grazed areas. Although most studies have examined the impacts of grazing on vegetation, a few have addressed impacts at a broader, ecosystem-wide scale. Levin et al. (2002), however, found that grazing by feral horses on Spartina-dominated marshes affected fish abundances, predation and behaviour.

The impacts of grazing by stock and feral animals on terrestrial and riparian vegetation in Australia are well known (Staton & Sullivan 2006). In general, stock grazing causes physical changes to the soil (erosion, compaction, pugging, alterations to microtopography), damage to the vegetation (trampling), nutrient enrichment from animal dung and urine, localised death of plants after smothering by dung or scalding by urine and a shift in species composition, in part due to the ecological changes noted above, in part due to the replacement of palatable species by less-palatable species. In her review of Australian saltmarshes, Laegdsgaard (2006) noted that many Australian saltmarshes were on private land and subject to stock grazing. That is certainly the case also for Victoria, and in Victoria saltmarsh is routinely grazed by cattle (particularly in western Victoria and Gippsland) and sheep (e.g. Port Phillip Bay), as well as suffering grazing by other introduced animals such as several species of feral deer and goats (e.g. French Island) – see Figure 1.52.

In principle, trampling by hard-hoofed animals such as cattle and sheep would have potential for most impact on the brittle-stemmed and succulent chenopod shrublands and herbfields that typify many coastal saltmarshes of southern Australia. This type of shrubland is not only easily damaged, but is made up of slow-growing species. It is well known that too heavy a stocking density can cause extensive pugging of wetland soils. Grazing would be expected to open areas between individual shrubs and other life forms, disturb the soil surface and allow invasion by alien plant species, such as *Polypogon monspeliensis and *Parapholis incurva.

Figure1.52: Cattle in coastal saltmarsh, Breamlea.


Selective herbivory seems to be responsible for the absence of Limonium australe (a species listed as vulnerable in Victoria) from grazed areas in Tasmanian saltmarshes (Kirkpatrick & Glasby 1981). Laegdsgaard (2006) argued that grazing led also to the creation of pools that could be quickly colonised by biting insects and other plant species (e.g. Triglochin striata) more tolerant of flooding. Figure 1.53, for example, shows a cross-fence comparison of a cattle-grazed and an ungrazed saltmarsh near the Barwon River estuary.

Figure1.53: Comparison of a cattle-grazed (right-hand side of fence) and an ungrazed saltmarsh, Barwon River estuary.

Theoretical analyses suggest that grazing should have a wide range of impacts also on saltmarsh fauna and the ecological processes that operate in saltmarshes. For example, impacts on nesting birds can be expected, as well as on benthic invertebrates via the compaction, trampling and nutrient-enrichment of soils. Impacts on patterns of nitrogen cycling – with associated effects on invasions by opportunistic (r-selected) weeds – are also highly likely, but to our knowledge have not been examined. As noted above, studies of the impacts of grazing by feral horses have revealed impacts on fish populations in Northern Hemisphere saltmarshes.

There are very few detailed empirical studies of grazing impacts on Australian saltmarsh. Bridgewater (1982) and Kirkpatrick & Glasby (1981) noted that grazing by cattle and sheep had an adverse impact on Victorian and Tasmanian saltmarshes. Kirkpatrick & Harris (1999) argued that grazing and fire had eliminated Tecticornia (Sclerostegia) arbuscula heath from some coastal saltmarshes in Tasmania. Mondon et al. (2009) concluded that grazing had had adverse impacts on the food value of Sarcocornia quinqueflora seed in saltmarshes along the Victorian coast, with possible implications for feeding by the endangered Orange-bellied Parrot. Lee & Choy (2004) mimicked some of the physical effects of grazing by removing surface vegetation from a Sporobolus virginicus saltmarsh in southern Moreton Bay (Queensland). They reported that mowing the surface vegetation did not have statistically significant impacts on macrobenthic fauna, mobile epibenthos, or juvenile nekton and plankton. Lee & Choy (2004) concluded that Sporobolus virginicus saltmarshes were highly resilient to disturbances such as removal of surface vegetation, and in this response were quite unlike nearby mangroves. It is unclear whether this conclusion can be applied to coastal saltmarsh in other, more southerly, regions of Australia that are floristically and structurally more diverse.

In one of the few detailed Victorian studies, the impacts of sheep grazing on saltmarsh and non-saltmarsh grasslands plant species were examined using a 3.2 ha enclosure over 18 months at The Spit Nature


Conservation Reserve (Carr & McMahon 2009). Some plant species were not grazed at any stage; others were intensively grazed at all times, while others were grazed lightly or moderately grazed. It was predictably evident that a hierarchy of preference by sheep for various plant species existed on the site, but there was some variation over the duration of the trial in grazing of different taxa. For example, Gahnia filum was not grazed in the early stages but plants were moderately grazed later. The changing feeding preferences were explained by the relative availability of palatable species on the site: those species that were initially and consistently preferred became less available as they were eaten out, forcing the attention of the sheep onto other, less preferred, species.

Some plant species, though palatable, were not heavily grazed; common examples included plants with rosettes of leaves pressed flat to the ground, such as *Hypochaeris radicata, *Leontodon taraxacoides and *Plantago coronopus. In these species, inflorescences were eaten even if the leaves were not. A suite of smaller annuals also were not grazed, or suffered minimal grazing because their leaves were mixed with a sward of Sarcocornia quinqueflora ssp. quinqueflora; the latter species, with its very salty tissue, was not grazed at all. This empirical observation supports the earlier hypothetical contention that the high salt content of saltmarsh halophytes should act as a feeding deterrent for many vertebrate herbivores (Chapter 1.9).

Sheep exerted an impact on the saltmarsh in ways other than direct grazing on plants. Mechanical damage by trampling, and in some cases by pulling plants out but not consuming them, had effects on the soil. The succulent chenopods Sarcocornia and Tecticornia spp., as well as Disphyma crassifolium are fragile and particularly vulnerable to trampling damage by hard-hoofed grazers such as sheep. Of these species, only Disphyma crassifolium was grazed but, in this species and in Sarcocornia, plant cover was greatly reduced by the mechanical damage of trampling sheep; both species occupy the Sarcocornia herbfield zone where sheep obviously spent considerable time. Sheep stayed out of the two Tecticornia zones, as evidenced by the absence of physical damage to Tecticornia halocnemoides and Tecticornia pergranulata (both very fragile) and general absence of sheep footprints in the bare soils of these zones.

Carr & McMahon (2009) concluded that, from the observations of grazing and other impacts by sheep on the site at The Spit Nature Conservation Reserve, a number of recurring effects of grazing could be seen for coastal saltmarsh. First, sheep concentrated grazing in the elevated non-saltmarsh parts of the site that carried grassland vegetation. There they preferentially consumed exotic and indigenous grasses and dicot herbs (annuals and perennials), driving some species to virtual extinction, or at least eliminating the standing crop of the species.

Second, sheep did not graze most of the perennial indigenous and exotic annual halophytes in saltmarsh areas. They consistently avoided Tecticornia halocnemoides, Tecticornia pergranulata, Frankenia pauciflora, Hemichroa pentandra, *Atriplex prostrata, Sarcocornia quinqueflora and Suaeda australis. It is possible that high salt contents in the leaf tissues of these species prohibited grazing. The salt-accumulating succulent Disphyma crassifolium was patchily, in some cases heavily, grazed; all inflorescences, however, were consumed.

Third, mechanical damage to plants and soils was a major impact of sheep grazing, especially in the Sarcocornia quinqueflora zone. At the most extreme, sheep removed all vegetation, as seen along their tracks which were conspicuous in some locations. Fourth, sheep had a major impact on the structure and floristic composition of vegetation on the site but impacts ranged from negligible in the Tecticornia zones to major in the Austrostipa stipoides –*Phalaris aquatica grassland zones.


Observations of any grazing trials, however, must always be interpreted within the climatic conditions that operated at the time. In the case of the Carr & McMahon (2009) study, the period of observation and monitoring (2007–2009) was in a time of serious drought. The effects of drought were obvious, and included widespread death or partial death of mature Frankenia pauciflora plants, premature death of some annual species before plants could achieve reproductive maturity (e.g. Apium annuum, *Senecio vulgaris, and *Hordeum marimum), failure of germination and/or survival in the indigenous and exotic, cool-season annual flora, which generally germinated during autumn, and very poor flowering and fruiting in the perennial indigenous and exotic flora generally.

There are a limited number of earlier and less detailed studies of grazing on Victorian saltmarsh, for example those by Carr et al. (1991) and Lane & Kinhill Planners (1979). Carr et al. (1991) found that sheep grazing was generally deleterious to saltmarsh at Murtcaim and Point Wilson, and reported that it had the ability to substantially reduce or eliminate Sarcocornia quinqueflora, Sclerostegia arbuscula, Tecticornia halocnemoides and Tecticornia pergranulata from coastal saltmarsh. Frankenia pauciflora was the only saltmarsh dominant that was not eliminated by sheep grazing. Carr et al. (1991) found also that the reproductive performance of Sarcocornia quinqueflora, the most important food plant for the Orange-bellied Parrot (Chapter 1.9), was compromised by sheep grazing since sheep grazed strongly on its inflorescences and infructescences. The impact was noted also by Lane & Kinhill Planners (1979, page 7), who reported that:

Sheep grazing and rabbit grazing are potentially detrimental to the existence of ample food for the Orange-bellied Parrot in the Point Wilson area…Sheep, the most destructive grazers of saltmarsh, should be excluded from all areas of saltmarsh of importance to the parrot.

Feralanimals

Although attention has focussed on the deliberate grazing of saltmarsh by domestic stock (particularly cattle and sheep), the impacts of introduced feral mammals, though largely unstudied, is pervasive in south-eastern Australia and by imputation has major consequences for vegetation floristics and structure and ecological processes and soil stability. European Rabbit Oryctolagus cuniculus and Brown Hare Lepus capensis are very abundant in southern Victoria, and the former probably occur in all upper saltmarshes and even lower saltmarsh across the state (e.g. the Spit Nature Conservation Reserve). Impacts arising from feral animals are probably profound, and along with invasions by exotic plants, feral animals are probably instrumental in the train of degradation of south-eastern Australian saltmarshes more generally. It is likely that rabbits have historically so profoundly affected the floristic composition and structure of upper saltmarsh in Australia that we are unaware of the full extent and nature of their impacts. Other feral mammals may have serious impacts on saltmarsh vegetation, but the scale or intensity of the impacts have not been studied. Of particular concern are Sambar Deer Cervus unicolor (Peel et al. 2005) and Hog Deer Cervus porcinus (Quin et al. 2007), as well as goats Caprea hircus, which graze on coastal saltmarsh on French Island. Introduced rats may frequent the drift line, but their ecological impact is unknown.

Feral cats Felix catus are known to inhabit saltmarsh and mangroves, as radio-collar studies on French Island have recently shown that individual cats are capable of living entirely within the intertidal zone (Michael Johnston, Department of Sustainability and Environment, pers. comm.). There are currently no data on the ecological impacts of feral cats in this environment, but impacts are likely to be substantial given that cats are presumably consuming many native species including birds and crabs.


Exoticinvertebrates

Invertebrates, especially exotic gastropods, can sometimes be a major negative influence in coastal saltmarsh. As noted below, grazing by gastropods can denude large areas of coastal saltmarsh in the Northern Hemisphere.

Five species of exotic terrestrial gastropods have been recorded in upper saltmarsh in Victoria. These herbivorous species, all European, are Cochlicella barbara, Cornu aspersum, Theba pisana, Cernuella virgata and Cernuella vestita. Very high biomass can be achieved locally, particularly of Cochicella barbara in species-rich Sarcocornia quinqueflora herbfield, for example at The Spit Nature Conservation Reserve on the western shores of Port Phillip Bay. All gastropod species are serious pests of agricultural and horticultural systems (Coupland & Baker 2007). Their impacts in saltmarsh vegetation have not been documented but, as they are novel animals in these environments, it is unlikely that the indigenous flora has evolved defensive strategies against their herbivory. Since the study by Carr (1982), who documented 12 saltmarsh vegetation communities in the area, there has been a major decline in the quality of Sarcocornia quinqueflora herbfield vegetation: massive invasion of exotic annuals, especially grasses, and the decline of all (and extinction of many) of the 15 indigenous annual species, all of which are cool-season species germinating after the autumn break (Carr et al. 2002). The changes occurred before the current drought.

It is likely that the indigenous flora, including the annuals, is particularly vulnerable to consumption by gastropods at the seedling stage; it may also be the case that the exotic invasive flora, mostly consisting of European annuals, has co-evolved with these snails, and such plant species have developed physical, chemical or temporal defence strategies against snail predation. The only direct observations on the impacts of exotic snails were on the flowers of Disphyma crassifolium. In several locations around Corio Bay, the abundant Theba pisana have been observed eating staminoids, stamens and styles (Geoff Carr, pers. obs.), with unknown effects on the plants’ reproduction.

dieback

In 1995, a large area (> 10 ha) of Tecticornia arbuscula shrubland at The Spit Nature Conservation Reserve was observed to have died back. The cause is unclear, but a generic dieback has been noted also for coastal saltmarsh at Torquay (Carr et al. 2000) and at Point Wilson (Geoff Carr, pers. obs.). Localised dieback, of unknown cause, in Avicenna marina was recently observed in the estuary of Hovells Creek in the north-west corner of Corio Bay (Geoff Carr, pers. obs., February 2010).

Dieback has been observed in many coastal areas in the Northern Hemisphere that have been invaded by exotic Spartina spp. (see Chapter 1.12), as well as in coastal saltmarshes in a number of other parts of the US eastern seaboard (Osgood & Silliman 2009). In the latter case, a suite of causal factors have been identified, including increased salinity (arising from drought), fungal pathogens, invasive gastropods and extensive herbivory by estuarine crabs.

inappropriate mosquito control

Stagnant pools, both permanent and temporary, in mangroves and on the upper levels of coastal saltmarsh are often ideal sites for mosquitoes to lay their eggs (see earlier section on the habitat value of coastal saltmarsh: Chapter 1.9). The main problem species of mosquitoes are Aedes vigilax, Aedes camptorhynchus, Aedes alternans, Anopheles hilli and Culex sitiens (Russell 2008). Of these species, Aedes camptorhynchus is the main


problem in Victoria. Different species of mosquito utilise mangroves and coastal saltmarsh in different ways: Aedes vigilax, Aedes camptorhynchus and Aedes alternans, for example, lay their eggs on drying soil at the base of plants in the higher levels of saltmarsh and at the edge of pools in mangroves; the eggs hatch when the sites are later inundated by rain or tides. Other species, for example, Anopheles hilli and Culex sitiens, are associated with permanent pools.

Mosquito control in saline coastal wetlands can be a sensitive issue, particularly for urban populations living close to these sites (e.g. the various towns and settlements around the Gippsland Lakes). Local councils are often under pressure to spray for mosquito control because of fears associated with disease, especially Ross River fever (spread by Aedes camptorhynchus in southern Australia: see Russell 2002) and possibly also Barmah Forest Disease. One mechanism often used in south-east Queensland and northern New South Wales to control mosquito numbers – runnelling – involves cutting drains into the marsh. It causes obvious structural damage and has the potential to accelerate mangrove encroachment into coastal saltmarsh (Breitfuss et al. 2003). Jones et al. (2004) reported no effect of runnelling on mangrove encroachment into Sporobolus virginicus saltmarshes in south-eastern Queensland but a potential for impacts in Sarcocornia-dominated marshes. Runnelling is most often undertaken in species-poor saltmarshes in south-eastern Queensland and different types of impact may be expected in the species-rich marshes of south-eastern Australia (Paul Adam, University of New South Wales, pers. comm.). Moreover, given the incidence of insect-driven pollination in Victorian coastal saltmarsh (see Section 1.8), it is possible also that mosquito-control measures (especially those using insecticides) could have impacts on sexual recruitment by saltmarsh plants.

As control of mosquito populations is difficult, expensive and not always successful, it is essential that planning policies ensure an adequate buffer distance between urban settlements and known mosquito habitats such as saltmarshes. Some mosquito taxa, however, can disperse widely from mangroves and coastal saltmarsh: Russell (2008) noted that Aedes vigilax can disperse > 50 km with assistance by the wind. Dale & Breitfuss (2009) provide an overview of the ecology and management of mosquitoes in Australian saltmarsh.

recreation

It is often difficult to separate recreational impacts from other forms of habitat destruction and fragmentation in coastal wetlands. For example, the destruction of coastal saltmarsh or mangroves to create marinas is a threat that could be discussed under a heading of land-claim, habitat destruction or recreation (Figure 1.54). Moreover, many recreational pursuits have impacts that include oil pollution (e.g. marinas and recreational boating) and the liberation of heavy metals (e.g. leaching of anti-fouling compounds from boat hulls, digging of contaminated mangrove sediments for bait).

Even passive forms of recreation can have a range of subtle impacts on mangroves and coastal saltmarsh. The construction of elevated boardwalks through mangroves may save the mangroves from trampling but does involve the permanent loss of mangrove habitat during construction (Figure 1.55). Walkways that are not elevated interfere strongly with patterns of tidal inundation of the landward side of saltmarsh, as apparent with the walkway through saltmarsh at Hastings (Figure 1.56). As noted earlier, CCA-treated wooden walkways may be a conspicuous source of copper, chromium and arsenic (Weis & Weis 2002). In the absence of walkways, visitors may walk through mangroves and saltmarshes, often with evident ecological impacts. Trampling has been reported to have adverse impacts on algal mats and benthic macrofauna in mangroves near Sydney (Ross 2006).


Figure1.54: Large marina created on areas that were previously mangrove and coastal saltmarsh, Western Port.

Figure1.55: Boardwalk through mangroves, Hastings.

Figure1.56: Poorly designed path constructed on fill through coastal saltmarsh, Hastings.


A large number of coastal saltmarshes in Victoria are within State Game Reserves, where hunting (for waterfowl) is permitted. Examples include Connewarre State Game Reserve and Jack Smith Lake. Hunting of Hog Deer is permitted in some saltmarshes (e.g. Snake Island in Corner Inlet). Finally, bait digging is undertaken in many mangrove areas across the coast of south-eastern Australia (Figure 1.57). It can cause not only habitat destruction in the mangroves themselves, but also the disturbance of nearby areas during access. There is some evidence that bait digging can result in the liberation of heavy metals from contaminated sites (NSW Fisheries and Hawkesbury Nepean Catchment Management Trust 1998). In these cases, the oxidation (and shift in redox potential) that occurs when buried sediments are brought into contact with the air presumably has an important role in altering the bioavailability of pollutants (e.g see Borch et al. 2010 and earlier discussion on toxicants).

Figure1.57: Bait digging in mangroves, Gosford (New South Wales central coast).

In many cases, entirely inappropriate forms of recreation are undertaken in coastal saltmarsh, including the driving of vehicles through the marsh (Figure 1.58). Vehicular access is a major issue in some parts of the Victorian coast, such as in Western Port, where hectares of saltmarsh have been destroyed through such recreation. Laegsdsgaard et al. (2009) noted the impact of off-road vehicles on coastal saltmarsh in New South Wales and southern Queensland. Not only 4WD vehicles are implicated, but a range of mountain bicycles, trail bikes and quad bikes. Horse riding may take place in some Victorian coastal marshes too. Damage from these types of recreational access can range from the relatively subtle (e.g. stem breakage) to long-lasting and severe (e.g. wheel ruts: see Figure 1.59).

Laegsdsgaard et al. (2009) argued that recreational vehicle access had contributed to the loss of threatened Wilsonia backhousei communities along the Parramatta River. Kelleway (2005) examined the impact of recreational vehicles on saltmarshes along the Georges River in Sydney, and reported that vehicle access had been responsible for the complete loss of ~2 ha of saltmarsh. Impacts were severe in Sarcocornia communities and less evident in Juncus communities. In both cases, however, wheel ruts were prone to waterlogging and their persistence may alter composition of the vegetation.


Figure1.58: Evidence of inappropriate recreation in a saltmarsh, Old Tyabb.

Figure1.59: Close-up view of vehicular damage, Old Tyabb.

inappropriate rehabilitation

In response to the century-plus history of land-claim and degradation of mangroves and coastal saltmarsh, a number of projects are have been commenced to rehabilitate degraded sites, particularly in urbanised areas (Field 1999; Greenwood & MacFarlane 2006; Green et al. 2009; Adam 2009b). Laegdsgaard (2006) noted that the rehabilitation of coastal saltmarsh was a relatively new undertaking in Australia, that mistakes have been made in the past and that lessons are still being learnt. Two errors are commonplace: i) the large amounts of damage done when inappropriate techniques are used to remove invasive plants (Figure 1.60); and ii) the planting of inappropriate species when degraded saltmarshes are revegetated (Figure 1.61).

There are many ongoing problems with the rehabilitation of coastal ecosystems, especially in terms of gauging the effectiveness of the rehabilitation (Chapman & Underwood 2000). Zedler (1988) reviewed the state of saltmarsh restoration in California, and updated that review with an analysis of wetland restoration in


Zedler (2006). It would seem that many of the lessons identified in Zedler’s review of 1988 had yet to be fully incorporated into restoration practices nearly 20 years later: piecemeal planning, rather than regional co-ordination, was still common; off-site mitigation still took place and species requirements were still very poorly understood. Given the recent implementation of saltmarsh restoration in Australia, it is highly likely that similar mistakes are being made here also. Adam (2009b) summarised the range of factors that influenced the effectiveness of saltmarsh rehabilitation: hydrology, weeds, introduced fauna, grazing and pollution. Hughes & Paramor (2004), Elliot et al. (2007) and Weis & Butler (2009) have reviewed the suitability of a range of approaches commonly invoked during the rehabilitation of coastal wetlands.

Figure1.60: Extensive areas of coastal saltmarsh damaged by inappropriate methods used to remove the invasive Spiny Rush *Juncus acutus, Grahams Wetland Reserve, Werribee South.

Figure1.61: Inappropriate planting of non-saltmarsh shrubs within coastal saltmarsh, Point Cook Coastal Park.


1.12 Spartina

spartina: an estuarine weed around the world

Sixteen species of the grass genus Spartina are known worldwide (Nightingale & Weiller 2005) and a number have become problematic weeds of saltmarshes and near-shore coastal areas across the globe. Spartina alterniflora, the perennial grass native to the eastern coast of North America, for example, is now on the list of the nine most harmful invasive alien plant species in China (Wang et al. 2006). Spartina alterniflora was introduced to China from the USA in 1979 in order to check soil erosion and protect dykes; it currently covers more than 112,000 ha of coastal land and continues to spread along the eastern coastline (Wang et al. 2006; An et al. 2007). Four species of Spartina have been imported into China, commencing with Spartina anglica from England in 1963. Spartina anglica caused a major weed problem in coastal China soon after it was introduced, and by 1985 it covered over 36,000 ha. The area covered by Spartina anglica has since decreased to < 50 ha and only Spartina alterniflora currently presents a serious weed threat in Chinese coastal areas (An et al. 2007). In fact, in North Jiangsu, Spartina alterniflora currently covers ~410 km of coast (out of a total coastline of 954 km) and in some cases the infestations can be 4 km wide (Zhang et al. 2004).

Spartina alterniflora poses a threat to saltmarshes on the western coast of the USA, into which it has colonised from the eastern (Atlantic) coast of North and South America (Doody 2008). Hacker et al. (2001) and Hedge et al. (2003) reported that Spartina alterniflora, Spartina anglica and Spartina patens were serious weeds in coastal areas of Washington State in north-western USA. Spartina anglica was introduced into Puget Sound (Washington State) in 1961 by an agronomist working for the Research and Extension Unit of Washington State University in order to stabilise dyke banks and provide forage for cattle; it has successfully invaded 73 sites, affecting 3,311 ha, after its introduction 36 years earlier (Hacker et al. 2001). Between 1995 and 2000, the area infested with Spartina spp. in Washington State increased 2.5 fold; over 8,000 ha of intertidal land are now vegetated with various Spartina species in Washington State (Hedge et al. 2003).

Spartina alterniflora affects coastal areas of California: in the saltmarshes of San Francisco Bay, Spartina has been reported to expand at rates of up to 2.5 m per year (Rosso et al. 2006). In fact, the history of Spartina invasion in California is significantly different to that in Washington State, as Spartina alterniflora was introduced in 1975 to California by the Army Corps of Engineers to ‘restore’ coastal marshes. It then hybridised with the native Spartina foliosa to produce a highly invasive hybrid (Williams & Grosholz (2008). The invasive nature of the Spartina alterniflora-Spartina foliosa hybrid has uncanny similarities with the invasive and hybrid *Spartina x townsendii that infests southern Australia. In New Zealand, Spartina x townsendii was introduced in 1913 and had expanded to cover 40 ha of the New River Estuary by 1952. Spartina anglica was then introduced in 1973 and in some places has replaced native coastal saltmarsh (Doody 2008).

Two species of *Spartina (see details below) are currently found in estuarine areas and saltmarshes of southern Australia, including Tasmania, Victoria and South Australia (Bird & Boston 1968; Fotheringham et al. 1996; Lane 1996; Wells 1996; Williamson 1996; Kriwoken & Hedge 2000). They have the potential to invade into estuaries of south-western Western Australia (Keighery 1996). As shown below, *Spartina is very seriously invasive in mangrove and lower saltmarsh vegetation over, potentially, many areas of coastal Victoria. Presently no other weed species is invasive in these environments, and *Spartina, with its tolerance to salinity and inundation, is unique in the exotic flora (Figures 1.62 and 1.63).


Figure1.62: *Spartina anglica invading tidal mudflats and mangrove shrubland, Anderson Inlet.

Many factors make Spartina spp. potentially ferocious weeds of intertidal, and in some cases even subtidal, coastal areas. Spartina is capable of asexual propagation from small vegetative fragments and these fragments are easily dispersed by waves and tides. Fertile forms can disperse readily by sexual means as well (Bridgewater 1996). Adults are resistant to many chemical herbicides and the location of infestations in aquatic areas (e.g. difficulty of access, risks with adverse side-effects from spraying etc.) generally makes them difficult to control. Moreover, Spartina anglica is more tolerant of tidal submergence than any other species of European saltmarsh plant and thus can colonise mudflats that are frequently and deeply inundated by the tides, and which could not otherwise support vascular plants (Adam 1990). In fact, Spartina anglica is so tolerant of tidal inundation that it has replaced the seagrasses Zostera noltii and Zostera marina along parts of the British coast. Finally, the spread of Spartina spp. across the globe has often been enhanced by purposeful plantings by various agencies (e.g. port authorities) who, even as late the end of the 20th century, aimed to use the plant to stabilise estuarine mudflats and prevent shoreline erosion.

Figure1.63: *Spartina anglica in flower, Anderson Inlet.


Nightingale & Weiller (2005) recognised two species of Spartina in Australia: *Spartina anglica C.E.Hubbard and *Spartina x townsendii H.Groves & J.Groves. A third species, Spartina maritima (Curtis) Fernald, was previously thought to be present in Australia but that was due to a misidentification of some specimens of *Spartina x townsendii. All *Spartina x townsendii are sterile males, whereas *Spartina anglica is fertile, with bisexual florets. Three main characteristics separate *Spartina anglica from *Spartina x townsendii: i) length of the ligule: 2–3 mm long in the former and 1–2 mm in the latter; ii) anthers dehiscent in the former but indehiscent in the latter; and ii) pollen fertile in the former but infertile in the latter.

the strange origin of *spartina x townsendii and *spartina anglica

*Spartina anglica and *Spartina x townsendii are both species which were inadvertently created by human activity. Neither existed naturally, and their emergence as new species took many decades to understand. This unusual phenomenon is worth tracing. Adam (1990) and Gray et al. (1991) provide detailed overviews of the taxonomy of Spartina spp., and much of the original pioneering research was undertaken by Marchant (1967, 1968).

In 1870, a novel, rapidly growing Spartina was collected from mudflats in the Hythe, Hampshire, southern England (Adam 1990). It was soon known colloquially as the ‘Hythe Spartina’. Before its discovery, two species of Spartina had been recorded in England: Spartina maritima, a native restricted to southern England; and Spartina alterniflora, which was believed to have been introduced in ships’ ballast from North America and was first observed in the River Itchen near Southampton in 1829 (Adam 1990). Interestingly, following its introduction from North America the exotic Spartina alterniflora had become widespread across southern England in the late 19th and early 20th centuries, but is now nearly extinct there (Adam 1990; Bridgewater 1996).

At first the Hythe Spartina spread slowly from its original location but ‘…towards the end of the eighties [1880s], something occurred that favoured the spreading of the grass’ (Stapf 1913, cited in Gray et al. 1991). The new Spartina was observed to spread rapidly across the central coast of England over ensuing decades, undoubtedly assisted by the efforts of port authorities to plant it in estuaries to stabilise mudflats, reduce coastal erosion and ‘reclaim’ coastal land (Adam 1990). It was subsequently realised that the Hythe Spartina was a hybrid between the native Spartina maritima and the exotic Spartina alterniflora, occasioned by human-induced movement of plants. To complicate the story further, it is now known that there are in fact two types of Spartina derived from the hybridisation of Spartina maritima and Spartina alterniflora. The story is complex, and was difficult to unravel.

The initial hybrid between the parental species (the F1 hybrid) is sterile, and spreads only by vegetative means. Its sterility is due to its hybrid chromosome set, which comes from two parents with chromosomes that cannot pair properly during meiosis (i.e. cell division to create haploid sex cells, which requires chromosome pairing). The sterile hybrid is now known as Spartina x townsendii. The sterility suffered by S. x townsendii is not, however, the end of the story. The process of chromosome doubling can circumvent this genetic block: if all the chromosomes in the unbalanced set are, for some reason, replicated and retained in a single cell, each chromosome has a pair (its double), and meiosis can take place, leading to the formation of viable pollen and seeds. Chromosome doubling, although ‘abnormal’, happens from time to time in plant evolution. The resultant plants are genetically distinct (with extra chromosomes), fertile among themselves, and may have


ecological or morphological characteristics that distinguish them from the F1 plants which bore them. For these reasons they constitute a new (fertile) species, and in the case of Spartina they have been called Spartina anglica. Moreover, a series of subsequent back-crosses (some sterile) may result from the suite of taxa described here.

Astonishingly, it took over a century from the Hythe Spartina’s appearance to discover the relationships between the plants and the other Spartina spp. present in south-eastern England, and to erect valid botanical names for each entity. The story is as follows.

In 1879, H. and J. Groves described the unusual Hythe Spartina and grouped it with the native Spartina maritima, then known as Spartina stricta. In 1880 and 1882, they revised their earlier description and identified the plant as a new species, Spartina townsendi (originally with one i, subsequently two), which grew in a tidal position between Spartina alterniflora and Spartina maritima and had morphological characteristics intermediate between them. Gray et al. (1991) noted that, for many years, it was thought that there was only one form of the hybrid between Spartina maritima and Spartina alterniflora, the fertile amphidiploid now known as Spartina anglica. In 1957, however, C.E. Hubbard showed that the original F1 hybrid still existed and was the form identified by the Groves in 1880. That finding meant that, under the laws of botanical nomenclature, the name Spartina townsendii should apply only to the initial hybrid and not to the amphiploid derivative. This, however, left the amphiploid form without a valid name, and in 1968 Hubbard called it Spartina anglica. The naming, however, was invalidly published; but the error was rectified in 1978 and Spartina anglica has become the legitimate name for the amphiploid form.

Gray et al. (1991) reviewed the range of evidence that Spartina anglica arose as a crossing of the Old World Spartina maritima and the North American Spartina alterniflora to produce the sterile F1 hybrid Spartina x townsendii, which then underwent chromosome doubling to form the fertile amphidiploid Spartina anglica. In turn, Spartina anglica can cross with Spartina alterniflora to produce a series of backcross progeny, as shown in Figure 1.64. Strong evidence for the amphiploid interpretation is provided by chromosome numbers in the various forms of Spartina: Spartina maritima and Spartina alterniflora have chromosome counts of 2n=60 and 2n = 62, respectively, whereas Spartina x townsendii counts are 62 for the sterile form and 120, 122 and 124 for the fertile Spartina anglica (Marchant 1968).

On the basis of extensive analyses of isoenzymes and seed proteins, Gray et al. (1991) concluded that Spartina anglica from south-east England was almost completely lacking in genetic variation. The lack of within-species variability may be the result of a single origin for the species or a multiple origin from uniform parents, and is maintained by the clonal spread of populations from few, or even single, progenitors. They argued that evolutionary history of Spartina anglica gave rise to a number of important ecological consequences. The relatively narrow ecological niche of the species, for example, is likely to be a function of its genetic uniformity. The lack of within-species variation probably accounts also for its occurrence as dense, monospecific stands that are susceptible to disease: the ergot fungus Claviceps purpurea has been recorded to rapidly and extensively infect Spartina anglica in Poole Harbour, United Kingdom.

For some time, Spartina x townsendii and Spartina anglica co-existed in estuaries of south-eastern England. Spartina x townsendii, however, has slowly been replaced by Spartina anglica across much of its original range, largely as a result of the greater vigour and tolerance to tidal inundation of the latter species. Gray et al. (1991)


noted that, in the late 1960s, there were only ~20 ha of Spartina x townsendii in Britain, compared with ~12,000 ha of Spartina anglica. Before this decrease in area, Spartina x townsendii had extended its range in the late 20th century to include parts of Holland, the extreme south-west of France and north-east Spain. It remains in Holland, especially at more stable and higher elevations of the intertidal zone, but its Spanish extent has been decreased by extensive destruction of coastal saltmarsh. By 1978, the occurrence in Spain of Spartina x townsendii was limited to a single site near San Sebastian airport (Gray et al. 1991). Complex temporal patterns of expansion, dieback and regrowth have been reported for Spartina x townsendii in Denmark by Doody (2008): large swards were formed from 1954–1964, dieback was observed from 1976–1988, and recolonisation by 1995.

Figure1.64: Genomic relations in Spartina. Source: redrawn from Gray et al. (1991), based originally on Marchant (1968).


It is unclear exactly when Spartina anglica arose in England or how it spread in the first few decades. The first record is from 1892, from Lymington, Hampshire, but subsequent records do not always differentiate between the sterile F1 hybrid and the fertile amphiploid. It seems, however, that Spartina anglica spread rapidly after the 1880s and 1890s; it was present in the Isle of Wight by 1893 and on the north coast of France by 1906 and in the saltmarshes of the River Seine by 1915. It spread through French estuaries very rapidly in the early-mid 20th century: there were between 4,000 and 8,000 ha present on the north coast of France by the mid 1960s. The French occurrences seem to be the result of natural transport, as there are no records of deliberate introductions into French estuaries (Gray et al. 1991). Doody (2008), however, notes the possibility that the occurrence of a Spartina alterniflora x Spartina maritima hybrid in south-west France occurred as a second, independent hybridisation event (see Baumel et al. 2003) and provides a date of 1906 for the introduction of Spartina anglica into France (see Doody’s Table 20).

The rapid spread of Spartina anglica and its capacity to form dense swards on the seaward edges of estuaries later led to the species being introduced deliberately across much of Europe and, eventually, to the Southern Hemisphere. The first record of it being deliberately planted in Britain is for the Beaulieu Estuary in 1898, followed by estuaries in Norfolk in 1907 and in Lincolshire in 1910. Large numbers of British estuaries were deliberately planted with Spartina anglica in the 1920s, and by 1967 it was estimated that Spartina anglica was present in 86 sites and covered 12,205 ha in Great Britain (Hubbard & Stebbings 1967). It was introduced in Ireland in 1925, Denmark in 1931, Germany in 1927, New Zealand in 1913 and the USA in 1960 (Doody 2008, using data from Ranwell 1967). The introduction of Spartina anglica to China in 1963 was noted above: the Chinese plants were brought into the country from Britain as a batch of seed in thermos flasks by the eminent saltmarsh ecologist D.S. Ranwell. Oddly, the evident ability of Spartina spp. to create severe weed infestations after the 1963 introduction of Spartina anglica was not fully acknowledged when, in 1979, Spartina alterniflora was introduced to China from North America, again to stabilise estuarine shores and ‘reclaim’ intertidal lands.

There are some areas where attempts to introduce Spartina anglica have not been successful, or where rapid expansion has not always followed successful introductions of alien Spartina species. As shown below, early attempts to introduce Spartina spp. to southern Australia were often unsuccessful. The eastern coast of the USA, Caribbean islands, India and South Africa also have seen unsuccessful attempts at its introduction. It is likely that winter temperatures are the factor limiting successful colonisation of these areas, as in the colder areas Spartina anglica is damaged by frost and, in the warmer areas, the warm winters impede vegetative development (Gray et al. 1991). *Spartina spp. are not present in New South Wales, and Adam (1990) argued that its restriction to cool temperate regions may be a function of seeds requiring a cold pre-treatment for germination. In California, Spartina anglica has spread little beyond the boundaries of its original introduction in the 1970s, and Spartina densiflora (originally from South America) has spread to cover only ~5 ha near San Francisco in the past ~30 years (Doody 2008).

The arrival of a brand-new species, un-tuned by interactions with its environment and its neighbours, is an interesting ecological phenomenon. Although Spartina anglica spread rapidly across British and European estuaries in the late 19th and early 20th centuries, there is good evidence that dieback started to occur in many sites in the mid 1920s (Gray et al. 1991). At Poole Harbour, where Spartina anglica was first recorded in the late 19th century, the area of infestation had expanded to 775 ha by 1924 but then declined to 415 ha by 1980. Similar declines have been reported for populations across south-eastern and south-western England,


as well as in northern France and south-west Holland. The cause has not been unequivocally identified, but extreme sediment anoxia and increased organic-matter contents have been proposed: ‘Die-back can therefore be viewed as a ‘natural’ process in which a newly evolved species has dramatically altered the sedimentary and drainage characteristics of the marshes, and created the anaerobic, waterlogged conditions which led to its own destruction’ (Gray et al. 1991, page 46-47). Doody (2008) provides a more detailed examination of dieback of various Spartina infestations around the world.

other spartina hybrids

The vigour and rapid expansion of so-called Spartina x townsendii across English estuaries, including its growth into areas that had earlier been vegetated with the native Spartina maritima, would today suggest it was a hybrid. The sharing of morphological characters with Spartina maritima and Spartina alterniflora would have reinforced this idea. Moreover, hybridisation is not uncommon among different Spartina species: along the eastern coast of the USA, Spartina alterniflora readily hybridises with Spartina patens, and along the western coast the native Spartina foliosa hybridises with the introduced Spartina alterniflora (Doody 2008). These Spartina alterniflora x Spartina foliosa hybrids are taller, more productive and can tolerate a far wider range of inundation conditions than the native Spartina foliosa (Brusati & Grosholz 2006).

Another unusual form of Spartina was found at the end of the 20th century in Europe, this time in south-west France. Plants were collected in 1892 by Neyraut and named Spartina x neyrautii in 1894 by Foucaud (Gray et al. 1991). Subsequent analysis showed this species also to be a Spartina maritima x Spartina alterniflora hybrid, and Gray et al. (1991) argued that, under the international laws of botanical nomenclature, it must be regarded as a synonym of Spartina x townsendii. More recently, Doody (2008) discussed the evidence obtained by Baumel et al. (2003) that a second, independent hybridisation event took place in south-west France involving the two parent species and giving rise to Spartina x neyrautii , but with different genotypes than those of south-east England.

history of *spartina in australia

There are no native *Spartina spp. in Australia (Doody 2008) and Boston (1981) provided a detailed history of the introduction of *Spartina into the country. What makes *Spartina infestations different from many other coastal-verge weed problems is that the plant was almost always deliberately, not accidentally, introduced into Australian estuaries. Although in some cases the plants were introduced by government agencies, many of the introductions seem to have been made without quarantine clearance or government sanction (Boston 1981).

The first known introduction to Australia was a planting in Corner Inlet, probably in the 1920s, by Professor A.J. Ewart from the University of Melbourne (Boston 1981). Many subsequent attempts at introducing it, mainly as seed, were made in the late 1920 and 1930s in all states, including the tropics (Adam 2009a). In fact, between 1927 and 1929 alone, there were at least 14 attempts to introduce *Spartina into Australia (Boston 1981). Most of the earlier attempts to establish *Spartina in South Australia and Tasmania were not successful (Fotheringham et al. 1996; Wells 1996), and their failure has been ascribed to poor treatment of the seed material, with the result that most was not viable. Following the first attempts at introducing *Spartina to South Australia in the 1930s using seed and plant material from England, a repeat attempt was made in 1957, this time using Tasmanian material, at Port Pirie (Fotheringham et al. 1996). In Tasmania, *Spartina was introduced deliberately in 1947 by the Department of Agriculture for the Port of Launceston Authority


to stabilise 20s (Wells 1996). Such later colonisation attempts were successful, and in the 1960s the area under *Spartina increased markedly. It is now found across much of the northern and eastern coastlines of Tasmania (Kriwoken et al. 2000) and especially in the Rubicon and Tamar estuaries (Lane 1996). By 1997, *Spartina anglica had covered 600 ha of intertidal land in Tasmania (Adam 2009a).

The identity of the various *Spartina species introduced into Australia has not always been clear. Nightingale & Weiller (2005) noted that *Spartina anglica has been reported for estuaries, inlets and lagoons in southern Gippsland, as well as north and south Tasmania; in contrast, *Spartina x townsendii has been recorded for tidal mudflats and mangrove swamps in South Australia as well as some estuaries in South Gippsland. In South Gippsland, the more vigorous *Spartina anglica is thought to have replaced the original populations of *Spartina x townsendii.

Williamson (1996) provided an overview of *Spartina introductions into Victoria and his review has yet to be updated. As reported by Bird & Boston (1968), the earliest plantings were at Lake Connewarre, Bass River, Corio Bay, Andersons Inlet and several locations in Corner Inlet in the 1920s and 1930s. At this time, for example, more than 3,000 plants were introduced by the Geelong Harbour Trust to ‘reclaim’ a shallow lake bed for eventual use as grazing land (Boekel 1996). Bird & Boston (1968) note that, like attempts in Tasmania and South Australia, many of the early attempts to introduce *Spartina into Victoria failed. Moreover, some that were initially successful did not evolve into large-scale infestations: the original plantings at Lake Connewarre, for example, spread little, possibly because they were limited by rabbit grazing. A survey in 1967 could find no trace of the plants at Lake Connewarre. Similarly, the original plantings in Corio Bay have since disappeared (Bird & Boston 1968) and recent visits suggest it has not re-appeared.

In other parts of Victoria, however, plantings of *Spartina did start to expand. The spread of *Spartina through Anderson Inlet started to be noticed in the 1980s and the Department of Conservation and Natural Resources began to investigate what was now considered a problem in 1991 (Williamson 1996). Mapping in 1993 showed there to be between 150 and 280 ha of *Spartina in Victoria, with most in Anderson Inlet (108 ha confirmed, 130 ha suspected), Corner Inlet (42 ha) and small infestations in Shallow Inlet, Westernport and the Barwon River (Williamson 1996). Bird & Boston (1968) provide earlier historical information on the presence of *Spartina in the Gippsland Lakes and Corner Inlet and discussed the possibility that most of the plantings in these areas were of less vigorous forms of *Spartina x townsendii, or perhaps even of the original English *Spartina maritima. As noted above, it is likely that this was an erroneous identification (see Nightingale & Weiller 2005).

physiology and phenology of *spartina

*Spartina anglica and *Spartina x townsendii are erect, halophytic grasses which grow to about 1 m high (Figures 1.50 and 1.51). They are both deep-rooted perennials with a large underground biomass of extensive fleshy rhizomes. As a fertile species, *Spartina anglica can spread via both sexual and asexual (clonal) mechanisms, whereas *Spartina x townsendii is limited to clonal growth and asexual means for dispersal.

*Spartina is unusual in Victoria – and indeed the world – in that it is a weed of the infra-tidal and lowest tidal locations: it colonises areas below the level of even wet saltmarsh, often into the mangrove zone and sometimes even into areas that previously supported the growth of seagrasses (see Figure 1.51). The niche is similar to that described for *Spartina anglica in other regions of the world (Gray et al. 1991).


Gray et al. (1991) attributed the ability to grow in the most inundation-prone parts of the coast to the possession by Spartina of the C4 photosynthetic pathway, the plants’ ability to oxidise phytotoxins such as Fe2+ and hydrogen sulfide, and its vigour and tolerance of high rates of sediment accretion. Adult plants, for example, can grow under a very wide range of sedimentation regimes, even in systems that are accreting at rates of up 8–10 cm year–1. Gray et al. (1991) cited Chinese cases where Spartina anglica has withstood short-term sediment accretion of up to 26 cm. In all of these attributes, Spartina anglica performs better than its sterile progenitor Spartina x townsendii or its parent taxa. Such hybrid vigour is not unusual: Doody (2008) notes that the Spartina foliosa x Spartina alterniflora hybrid in San Francisco is more robust than the native Spartina species.

The reasons for plants not establishing in a given area, despite deliberate attempts at their introduction, are not well understood. As noted above, few of the early attempts to introduce *Spartina spp. to Australia were successful, even though the climate and general environment seemed appropriate. Gray et al. (1991) proposed that a suite of factors interacted to prevent successful colonisation of new areas; important factors could include the presence of unstable or sandy substrata, too cold (especially the incidence of frost) or too warm winters, and lack of sexual recruitment in the case of Spartina x townsendii. It seems that, for most species of Spartina, a narrow set of requirements need to be met for successful seed germination and the establishment of young plants, which is then followed by aggressive invasion of surrounding areas by clonal expansion. In north-west Spain however, Spartina maritima seems to colonise new areas from rhizome fragments (Doody 2008). Rhizome tillering is an effective mechanism for both species to maintain rapid expansion into new areas in the absence of successful sexual recruitment. The conditions required for successful seed germination are not well understood.

The phenology of Spartina anglica is well understood, at least for English populations, and in those cases the plant shows a strongly seasonal pattern of growth. Leaf axils produce overwintering buds in late autumn, which then quickly grow and flower in summer to early autumn (Doody 2008). The flowering culms die as new buds are formed the following autumn. Rhizomes develop during winter in response to short daylength (Gray et al. 1991).

It is likely that the strong effect of temperature on Spartina anglica results from the C4 metabolism of the plant. Plants with the C4 photosynthetic pathway have tropical origins and are usually poorly adapted to cool climates. Gray et al. (1991) noted that photosynthetic CO2 assimilation in Spartina anglica ceases at 7–9oC and significant canopy development does not occur until air temperatures exceed 9oC. As discussed later in the section on climate change impacts (Chapter 1.13), C4 plants have greater water-use efficiency than C3 taxa, and this may contribute to the ability of Spartina to tolerate salinity stresses.

The above-ground biomass and rates of primary productivity of Spartina anglica are variable, again probably in response primarily to temperature and secondarily to inundation regimes. In English marshes, above-ground biomass of Spartina anglica varies from ~400 to ~1,200 g m–2, with an average of ~1,150 g m–2. Estimates of above-ground net primary production range from 475–1,850 g m–2 year–1 (Gray et al. 1991); comparison with the data shown in Chapter 1.8 (e.g. Figure 1.26) indicates that rates of primary production are not dissimilar to those of Spartina alterniflora in the USA. As before, though, below-ground production is not included in these estimates, and root and rhizome production could easily exceed that of above-ground organs. Moreover, the increase in vascular plant production has to be set against the inevitable decrease in the productivity of algae, especially of benthic algae.


environmental impacts and the need for control

Little is known of the environmental impacts of *Spartina in southern Australia, so the following review draws from both local and overseas experience. Early work in Victoria includes the The sea has weeds too seminar in Inverloch (Department of Conservation and Natural Resources 1992b) and the collation of a set of photographic transparencies of Spartina in South Gippsland and Western Port (Phillips 1992). An Australasian conference on *Spartina control, How green is your mudflat? was held at Yarram in 1995 (Rash et al. 1996). Much of the more recent Australian work on *Spartina, however, has been undertaken as part of student BSc(Honours) projects and the results are available only in the form of the students’ theses; details are provided below on these investigations.

The impacts of *Spartina infestations encompass ecological, social and economic effects, and the responses by different community or industry groups to the plant are not always negative. Kriwoken & Hedge (2000) analysed the impacts of *Spartina anglica on Tasmanian estuaries, and noted the variety of responses of different interest groups to infestations. In the River Tamar of northern Tasmania, the plant is welcomed by some residents and agencies since ‘Spartina infestations significantly improve the navigability of shipping channels by stabilising sediments’ (Kriwoken & Hedge 2000, page 575). Some residents also prefer the ‘green meadows’ of Spartina over the original brown mudflats. Conversely, others consider it a serious nuisance, since it limits public access to the water for recreation and competes for space with aquaculture, especially of the Pacific Oyster Crassostrea gigas. Table 1.16 shows the contrasting perspectives held by different groups towards *Spartina infestations in Tasmania. Note that mangroves do not occur in Tasmania, and different responses may be elicited on the mainland, where there is often a band of mangrove vegetation between saltmarshes and the sea.

Table1.16: Range of positive and negative perceptions of *Spartina in Tasmania. Source: modified from Kriwoken & Hedge (2000, Table 1).

Group Positive perceptions Negative perceptions

Tourism and recreation Aesthetic: attractive green swards replace unattractive brown mudflats

Infestations impede public access to the shore for boating or fishing; opportunities for bird watching reduced

Primary industries Alternative fodder crop Competition for coastal land with aquaculture

Engineering Land-claim; stabilisation of channels for shipping and navigation; mitigation of coastal erosion

Increased rates of sediment accretion; alteration to existing geomorphology

Conservation Increased rate of primary production* Drastic modification of intertidal habitats

* Note that this assessment of positive impacts is i) yet to be empirically established and ii) has little to do with conservation values

In a BSc(Honours) thesis, Sheffield (2002) examined community attitudes to *Spartina in Anderson Inlet in Victoria. She found a diversity of community attitudes to the plant, and most (84% of respondents) were aware of the presence of *Spartina in the estuary. One sector of the community was opposed to removal of *Spartina; in stark contrast, another believed that total eradication was necessary. In total, 68% of respondents indicated that the negative impacts of *Spartina outweighed any positive effects the plant may have.

Callaway & Josselyn (1992) examined the ecological impacts of introduced east coast Spartina alterniflora on estuaries of the west coast of the USA, and their summary is adapted for Table 1.17. (In some ways the


invasion of coastal marshes by exotic Spartina spp. is mimicked by the invasion of some North American Spartina marshes by the introduced Phragmites australis; see Weis & Weis 2003 for an analysis of the ecological impacts of Phragmites infestations.)

Table1.17: Potential ecological effects of Spartina alterniflora on estuaries of the west coast of the USA. Source: modified from Callaway & Josselyn (1992).

Potential impact Likely mechanisms

Competitive replacement of native taxa Higher seed production and germination; more rapid clonal growth

Increased rate of sedimentation Greater stem densities; larger and more rigid stems

Impacts on food-web structure Changes in quality and quantity of detritus

Decreased benthic algal productivity Shading under dense Spartina canopy

Impacts on upper saltmarsh Increased production of wrack and deposition in upper marsh

Impacts on habitat quality Greater stem densities

Impacts on benthic fauna Greater root/rhizome densities; colonisation of subtidal zones

Loss of foraging areas for shorebirds and waders Colonisation of bare mudflats; colonisation of subtidal zones

In contrast to some community and agency perspectives, the position held almost unanimously by conservation biologists and natural resource managers is that *Spartina is a serious weed that should be eradicated. In Victoria, the introduction and spread of *Spartina to estuaries has been listed as a potentially threatening process under the Flora and Fauna Guarantee Act 1988. The environmental impacts can be grouped into four main categories: • Competition with native plants• Effects on fauna habitat• Effects on sediment deposition and accretion• Biogeochemical effects.

Interactionswithotherplants

Little research has been conducted on the interactions between *Spartina and native Victorian plants. In Victoria, mangrove beds seem particularly susceptible to *Spartina infestations and the dense swards that result may prevent the establishment of young Avicennia marina seedlings. Spartina may also replace several other native vascular species of the lower marsh, including Juncus krausii and Sarcocornia quinqueflora.

Effectsonfauna

Most information on the effects of *Spartina on fauna are related to birds, although the effects on crustaceans and other invertebrates are likely to be great. *Spartina infestations can affect the use of coastal saltmarsh and other intertidal habitats by shorebirds in at least three ways (Gray et al. 1991; Doody 2008). First, *Spartina can readily colonise unvegetated mudflats and destroy a crucial feeding habitat for wading and shorebirds. Second, the increased rate of sediment accretion (see below) leads to more elevated sediments, with a reduction in tidal inundation and consequently a reduction in the time that shorebirds can feed on alternately exposed and inundated shores. Third, the presence of *Spartina may have direct effects on invertebrates (both taxonomic composition and abundance) and thus alter food sources available to birds.


A number of English studies show that Spartina anglica inhibits feeding by waterbirds by limiting opportunities for foraging (Gray et al. 1991) and most concern is directed to effects on overwintering shorebird species (Doody 2008). In Europe, the major concern with Spartina anglica is the colonisation of mudflat that was previously vegetated with macroalgae such as Enteromorpha, or of seagrass beds vegetated with Zostera spp. Invasion of these habitats has impacts on foraging opportunities by birds such as geese (e.g. Branta bernicla), Widgeon Anas penelope and a wide range of waders. The mechanism proposed to account for the impact of Spartina anglica on shorebirds is that it removes feeding areas and increases the time needed to feed on those mudflats and seagrass beds that do remain; both factors exert energetic costs on the birds (Gray et al. 1991; Doody 2008).

Decreases in the abundance of Dunlin Calidris alpina, in particular, have been attributed to Spartina anglica infestations in England. This shorebird species feeds at the tidal edge and along mudflats and, in one particular estuary, the Dyfi in Wales, it has been shown that decreases in bird numbers in the 1970s coincided with marked expansion in the area of the estuary colonised by Spartina anglica. Subsequent studies of other estuaries have shown the greatest decline in shorebird numbers has occurred in those estuaries where Spartina anglica had expanded the most (Gray et al. 1991; Doody 2008). Such correlational studies, of course, cannot prove a causal link between decreased bird abundance and Spartina infestations. Experimental approaches are required to establish such links, and might be provided by observations of bird numbers after Spartina anglica had been removed from infested areas. When control has been undertaken, the abundance of Dunlin has sometimes increased but sometimes has not been affected. In the latter case, it may be that the prior infestation permanently altered sediment characteristics and birds may not the able to return to their former feeding grounds.

The effect of Spartina anglica on invertebrate populations has received little attention, even in the south-east of England where the species arose (Gray et al. 1991). The available studies indicate that there is little insect diversity in the canopy of Spartina anglica marshes: ‘this rather skeletal insect food-chain mirrors, but is less rich than, S. alterniflora marshes’ (Gray et al. 1991, page 43). Some evidence suggests an impact on benthic surface and infauna. Neira et al. (2007) similarly reported that, for areas affected by the Spartina foliosa x Spartina alterniflora hybrid in San Francisco Bay, there was a substantial reduction in the species richness of macrofauna and an increase in dominance by a limited number of species. The new fauna also had different feeding modes than that of the original macrofauna, with a shift from surface feeders that ate benthic algae to sub-surface feeders, principally oligochates and polychaetes, that fed on Spartina detritus.

Broadly similar results have been reported for the west coast of the USA. Levin et al. (2006) reported that the invasion of sandflats in San Franciso Bay by the Spartina alterniflora x Spartina foliosa hybrid shifted food webs away from algal-based to detrital-based systems. The shift in fundamental supplies of carbon and energy for estuarine food webs was reflected in a shift away from invertebrates that consumed surficial algae (e.g. amphipods and bivalves) to those that could consume vascular plant detritus (e.g. polychaetes and oligochates).

Since Spartina anglica is a C4 plant, it should be easy to use stable-isotope analyses to trace its incorporation into food webs. On the basis of δ13C analysis, Jackson et al. (1985) reported that few canopy species of insect incorporated Spartina carbon, with the inference that few taxa fed on the plant. The reasons for a lack of incorporation of Spartina anglica material into food webs in English infestations may result either from the


relative indigestibility of C4 plants or the recent origin of the plant in English estuaries. Most Spartina anglica production in these marshes seems to be processed by benthic invertebrates, particularly polchaetes such as Nereis diversicolor.

Hindell (2008) undertook a stable-isotope analysis of the role played by *Spartina anglica in providing carbon and energy sources for Estuary Perch (Macquaria colonorum) in Anderson Inlet, an embayment on the Victorian coast heavily infested with Spartina. While δ13C and δ15N analyses could differentiate clearly among a number of potential food sources (e.g. red algae; non-*Spartina saltmarsh; Avicennia-Melaleuca-Phragmites; and brown algae-green algae), the *Spartina signature could not be separated from that of seagrasses on the basis of these two elements. Analysis of δ34S, however, did allow differentiation of *Spartina from seagrasses. The isotopic analysis showed that Estuary Perch fed on a wide range of food sources, and a mix of seagrass and *Spartina could contribute up to 70% of the food of larger fish. Although Anderson Inlet contains a relatively small area of seagrass (< 1 km2), seagrasses made a surprisingly large contribution to the food of Estuary Perch, especially to adult fish. *Spartina, however, seemed to contribute significantly to the food of juvenile fish.

Sedimentationrates

The dense sward of above-ground shoots and thick network of underground rhizomes created in Spartina infestations readily traps suspended particles in incoming tides and results in accelerated sediment accretion. Doody (2008) provided a short overview of the impact of Spartina infestations on sedimentation rates. In south-west England sedimentation rates of 100–120 mm year–1 were reported by Ranwell (1964) and, under exception circumstances, an accretion rate of 200 mm year–1 was recorded (Ranwell 1967). These rates are an order of magnitude greater than the rates commonly observed of coastal saltmarsh (Doody 2008).

The introduction of Spartina also has the capacity to change the geomorphology of saltmarshes by causing infilling of small tidal creeks and covering the surface of open pools, thereby simplify the tidal drainage system and surface hydrology of the marsh. The extensive litter formed by the grass also affects ecosystem processes as the marsh becomes a sediment sink. Larger tidal channels invaded by Spartina become narrower and concentrate tidal flows so when there is dieback of the plant there is rapid and active erosion of the channel margins occur.

Biogeochemicalimpacts

Coastal saltmarsh and mangroves are often natural depositional zones, where contaminants accumulate because of their adsorption to suspended particles in the water column and the subsequent deposition of these particles when they contact seawater and water velocity decreases within vegetated zones. An implication of such deposition is that, should plants and sediments be disturbed, accumulated contaminants may be released into the general environment. Sheehan (2008) investigated that possibility in the Tamar Estuary of northern Tasmania, a location with not only extensive *Spartina anglica infestations but also a long history of heavy industry in the headwaters and catchment. He found that concentrations of cadmium, chromium, copper, lead, nickel and zinc in *Spartina sediments were within the acceptable range outlined by ANZECC/ARMCANZ (2000) for sediment quality. Concentrations of a number of organic pollutants (e.g. PCBs and PAHs) also were within sediment-quality guidelines. Sheehan (2008) concluded that sediments under *Spartina anglica beds did not act as a sink for organic pollutants or that the accumulation of contaminants was highly localised within different parts of the estuary.


Studentprojectson*Spartinainsouth-easternAustralia

A number of student projects have examined the impact of *Spartina anglica in estuaries in south-eastern Australia (Graham 2000; Cowling 2001; Hamilton 2001; Sullivan 2001; Sumby 2001; Sheffield 2002; Sheehan 2008; Cutajar 2009). With the exception of Sheehan (2008), a PhD thesis, all were short-term studies (as a direct function of most being undertaken as part of a BSc(Honours) thesis) but even so they do provide some indications of the effects *Spartina has on estuarine habitats in the south-east of Australia.

Cowling (2001) compared macrofauna in four habitats near the Bass River in Western Port: unvegetated mudflats; native saltmarsh; and two forms of *Spartina anglica infestation. She found no depletion of macrofaunal diversity in *Spartina infestations and an increase in the *Spartina that grew in slightly elevated mounds (as opposed to extensive flat swards). In a set of transplant experiments, it was shown that the native pulminate gastropod Ophicardelus ornatus, which was abundant only in mounded *Spartina, grew more slowly when moved to nearby mudflats. Conversely, Salinator fragilis, which was abundant on unvegetated mudflat, did not survive when confined to flat expanses of *Spartina sward. Cowling (2001) concluded that some native invertebrate taxa in Anderson Inlet would be benefited by *Spartina and others disadvantaged.

Hamilton (2001) reported on some aspects of *Spartina anglica in Western Port. He found that plants were fertile (hence *Spartina anglica not *Spartina x townsendii), had spread into a band about 3 km long either side of the Bass River, and on the basis of transplant experiments, could grow out onto the open mudflats of the embayment.

Sullivan (2001) examined fish populations in the Tamar Estuary in northern Tasmania. He reported the four most common fish detected in beds of *Spartina anglica were Hardyhead Atherinosoma microstoma, Tasmanian Smelt Retropinna tasmanica, Yellow-eye Mullet Aldrichetta forsteri and the goby Tasmanogobius lordi. Some fish taxa (e.g. Atherinosoma microstoma, Aldrichetta forsteri) were more abundant in Spartina than on adjacent mudflats, whereas others showed no preference between the two habitats.

Sumby (2001) reported that there was no difference in the depth of the oxic zone, particle size, organic-matter content or algal pigments (chlorophyll a and phaeophytin) between estuarine mudflats in Western Port and areas vegetated with *Spartina. The macrofauna of both habitats was dominated by a single species of polychaete worm, Nephtys australiensis.

Cutajar (2009) compared the structure of macrobenthic animal communities in *Spartina-affected areas of Anderson Inlet with that in adjacent mudflats and saltmarsh. He found that macrofauna in Spartina anglica patches were depauperate compared with animal populations in native saltmarsh and nearby mudflats, and that their species assemblages were unique. That finding suggests that *Spartina anglica had an overall negative impact on macrobenthic biodiversity and significantly altered the native macro-invertebrate community.

control of spartina infestations

Roberts & Pullin (2008) recently undertook a meta-analysis of the effectiveness of various methods used to control Spartina infestations around the world. They found that the effectiveness of herbicides varied greatly from location to location, and with type of chemical used, the presence or absence of wetting agents and application technique. Cutting can be effective if combined with smothering for the control of Spartina anglica, but it is not clear whether it is effective for other Spartina species. There is some potential


for biological control of Spartina infestations (Hedge et al. 2003; Roberts & Pullin 2008): for example, the hemipteran leafhopper Prokelisia marginata has been shown to stress or kill Spartina alterniflora in greenhouse experiments (Doody 2008). Table 1.18 outlines the effectiveness, advantages and disadvantages of various control measures for Spartina infestations.

Table1.18: Comparison for the effectiveness, advantages and disadvantages of various control measures for Spartina infestations. Source: expanded from Doody (2008, Table 22).

Method Effectiveness Advantages and disadvantages

Herbicides Variable Dalapon® can be difficult to obtain; glyphosate can be ineffective; costly; requires repeated (maybe continual) application; suitable for small- and large-scale infestations

Digging Variable Best for small-scale infestations; best for seedlings; unlikely to be effective with adult plants; sediment disturbance

Inundation May be effective in minimising spread Costly; damaging to other saltmarsh plant taxa

Bulldozing Ineffective Massive disturbance to sediments; habitat destruction

Harrowing Counterproductive Greater propagation from broken rhizomes

Burying Effective if plants fully covered Difficulty of access for required ploughing machinery

Crushing Partially effective Requires repeat treatments; difficulty of access of required machinery

Burning Ineffective Impractical

Grazing Can prevent seedling establishment Cheap; increased shoot density of remaining plants

Mowing Can prevent seedling establishment Difficulty of access of required machinery; increased shoot density of remaining plants; requires repeat treatments

Covering with black plastic Variable Damage to plastic sheets by waves and tides; suitable only for small infestations

Biological control Possibly highly effective Not currently demonstrated to be practicable; may involve introduction of alien species; avoids use of hazardous chemicals; limited need to ongoing intervention

In England, control of Spartina anglica is based on a mixture of physical removal (digging or bulldozing) and herbicide spraying. Gray et al. (1991) reported that the most commonly used herbicide was glyphosate (Roundup®) but the most frequently recommended was the sodium salt of dichloropropionic acid, Dalapon®. Doody (2008) reviewed the effectiveness of various herbicide treatments for Spartina control in the UK and concluded that Dalapon® could achieve a 99% kill after three applications in some locations. In other areas, however, Spartina continued to grow and expand over subsequent years. Experience in the UK suggests that glyphosate is less effective (e.g. 50 % kill in initial applications to the Spartina infestation in Lindisfarne National Nature Reserve), a result confirmed by experience in the USA (Doody 2008) and Australia (Bishop 1996). Hammon & Cooper (2003) also found that glyphosate was relatively ineffective as a control agent, this time for Spartina anglica infestations in Northern Ireland.


In Australia, Fusilade® is used for *Spartina control. The active ingredient in Fusilade® is 212 g L–1 fluazifop-P as the butyl ester. In New Zealand, Gallant® is widely used for Spartina control (Shaw & Gosling 1996; Doody 2008). Palmer et al. (1996) reported on the toxicity of Fusilade® to seagrasses and some animal species of Victorian coastal macrofauna. Field trials showed the herbicide did not affect growth of Zostera muelleri at Toora Beach, South Gippsland, but a combination of laboratory tests and field trials showed it was toxic to some aquatic invertebrates at concentrations above ~1 mg L–1. The active ingredient showed little persistence in the marine environment, and most of the fluazifop-P butyl ester had degraded to the less toxic, acid form within three days.

Pritchard (2006) reviewed the effectiveness of aerial and ground applications of Fusilade® in the control of *Spartina in Anderson Inlet. It was concluded that aerial (helicopter) and ground applications of the herbicide both gave ‘…very good control nine months after application’ (Pritchard 2006, page 3). Ground applications seemed to give slightly more consistent control (regrowth < 1%) than the aerial application (regrowth 0.5 to ~3%) but the small size of the experiment precluded sufficient statistical power to detect small differences between the two treatments. A risk with aerial application is that the herbicide may drift into non-target areas: in the case of the Anderson Inlet trials, the drift was 1–2 m outside the target plots. Aerial applications, however, could not control *Spartina growing under non-target plants such as mangroves. Anecdotal reports suggest that mangroves have been adversely affected by aerial spraying at the Bass River Mouth, and have suffered some defoliation (Steve Sinclair, pers. obs.).

Bishop (1996) undertook a series of trials to assess different control methods for *Spartina in Tasmania. Glyphosate was ineffective. Slashing also was ineffective as a long-term control measure because the plants rapidly regrew, but a combination of slashing and smothering with black plastic was highly effective. Attempting long-term smothering of plants with black plastic, of course, will present considerable logistical problems for *Spartina in subtidal and intertidal areas. Lane (1996) reported results broadly similar to those of Bishop (1996) for *Spartina control in the Rubicon Estuary of northern Tasmania

Two student theses have examined removal methods for *Spartina in Anderson Inlet, Victoria. Graham (2000) compared physical removal and slashing as control methods, and found that mechanical removal was more effective than slashing. In fact, there was some evidence that slashing led to a subsequent increase in growth. Sheffield (2002) compared the effectiveness of herbicide applications versus physical removal for the control of *Spartina anglica. For the herbicide treatment, a Fusilade® and D-C Trate® mixture was applied by hand in a well-replicated experimental design. The contrasting treatment was physical removal to 5–10 cm depth with a spade. Over a subsequent three-month monitoring period, Sheffield (2002) found that physical removal led to a significant decrease in *Spartina density, whereas herbicide-treated quadrats did not differ significantly from controls. Like many student projects, the time scale for follow-up observations was short and it is not surprising that physical removal led to an immediate reduction in plant density, as plants were presumably extracted during the removal process.

As with most invasive plants, control of Spartina is difficult and expensive, may be only partially effective, and almost certainly involves repeated interventions (Reeder & Hacker 2004; Taylor & Hastings 2004; Li & Zhang 2008; Roberts & Pullin 2008; Williams & Grosholz 2008). A suite of different strategies can be invoked to control infestations, and each approach is associated with a different set of risks. Taylor & Hastings (2004), for example, reported that, although the best strategy for controlling Spartina alterniflora infestations


in California was to remove a large fraction of infested areas annually, starting with the main invasion, that approach assumed that funds would be available in subsequent years for ongoing control. If resources were more limited, a better strategy was to control fast-growing but low-density plants. Subsequently, Grevstad (2005) argued that a strategy of controlling outlying patches first, then moving to control of dense meadows was more cost-effective than a strategy that adopted the reverse order. This latter approach has obvious parallels with the widely-known ‘Bradley’ method sometimes recommended for other (mainly terrestrial) systems (Bradley 1988).

management of *spartina in victoria

The following section on the control of *Spartina on land managed by Parks Victoria was provided by Steffan Howe (Parks Victoria). *Spartina is considered by Parks Victoria to be a major environmental weed and significant threat to bays and estuaries and on the habitats of migratory bird species and near-shore marine fauna. It causes siltation and impacts on the biodiversity and integrity of native saltmarsh and seagrass communities. Parks Victoria manages ~70% of the coast and 5% of the marine environment across the state, with many of the *Spartina infestations in Victoria concentrated within these areas. In particular, *Spartina infestations occur in Corner Inlet Ramsar Site (including Corner Inlet Marine National Park, Corner Inlet Marine and Coastal Park and Nooramunga Marine and Coastal Park), Western Port Ramsar Site (including Bass River mouth and estuarine reaches, and Moody’s Inlet), San Remo, Shallow Inlet Marine and Coastal Park and Anderson Inlet Wildlife Management Cooperative Area.

Parks Victoria manages pest plants such as *Spartina in accordance with the objectives of a range of treaties, conventions, Acts, regulations and policies, including: i) ensuring that threats to all indigenous flora and fauna occurring in the parks and reserves system are managed, in accordance with the Flora and Fauna Guarantee Act 1988; and ii) the eradication or control of exotic flora and fauna in the parks and reserves is in line with the National Parks Act 1975 and the Catchment and Land Protection Act 1994.

Parks Victoria has invested considerable time and resources in efforts to eradicate *Spartina : to date ~218 hectares of *Spartina has been treated, mostly since 2006, which includes multiple follow up treatments for a number of locations. Parks Victoria has also worked in partnership with a range of other land managers including Catchment Management Authorities, Melbourne Water and Birds Australia in an attempt to eradicate the plant.

*Spartina control in Victoria originally commenced in Corner Inlet in 1995. Spraying trials indicated that Fusilade® applied by hand boom spraying, when the plants are uncovered by tides, was the most cost effective method of control. *Spartina control later involved use of boats, quad bikes and four-wheel-drive vehicles and then progressed to the use of hovercraft as spray platforms. Significant challenges for successful eradication of *Spartina include tidal restrictions, the inaccessible nature of the tidal mudflats and the size and scale of the infestation. New infestations are continually being found up to 3.5 kilometres from the shoreline which highlights the urgent need to eradicate all infestations over a short time frame. Parks Victoria is continuing its *Spartina eradication program and continues to seek opportunities to expand partnerships and additional funding for the eradication program, and develop research projects to better inform *Spartina management.

Chapter 5.5 describes the mapping of *Spartina infestations undertaken as a part of the inventory and condition component of the current project: *Spartina was found in five sections of the Victorian coast and had a combined area of ~275 ha (Table 5.8).


1.13 Climatechange

Mangroves and coastal saltmarsh are among the vegetation types of Victoria most likely to be affected adversely by climate change. In large part their sensitivity is a function of coastal location and susceptibility to sea-level rise. Saltmarsh vegetation is particularly sensitive to subtle changes in waterlogging and salinity regimes, and both variables will be affected by rises in sea levels. In pre-colonial times coastal vegetation could often respond to sea-level rise by retreating into the terrestrial hinterland, but only where that land was sufficiently flat (Woodroffe & Davies 2009). Steep hills or dunes discourage the formation of wetlands and, in a number of places along the Victorian coast (e.g. around the Otways and other areas of western Victoria), saltmarshes abut steep terrain and thus have nowhere to retreat should sea levels rise. Critically, the intensity of post-colonisation development on coastal plains has meant that there is no room left for a landward retreat of coastal vegetation. Figure 1.65 shows one example of the ‘squeeze’ facing mangroves and saltmarshes should sea levels continue to rise.

Figure1.65: Saltmarsh squeeze: the limited capacity of saltmarshes to retreat landwards with further rises in sea level, eastern shore of Western Port.

climate change and australia

The Intergovernmental Panel on Climate Change (IPCC) published its 4th assessment of global climate change in 2007 (Intergovernmental Panel on Climate Change 2007). It is the most comprehensive global review and analysis of the scientific research on climate change to date, and concluded that:• The global atmospheric concentration of CO2 has increased from about 280 parts per million (ppm) in

1750 to 379 ppm in 2005.• The rate at which CO2 concentrations in the atmosphere is increasing is, itself, increasing; concentrations

increased at a rate of 1.9 ppm year–1 from 1995–2005, compared with a rate of 1.4 ppm year–1 from 1960–2005.

• Of the 12 warmest years in the instrumental record of global surface temperatures (which dates from 1850), 11 occurred in the period from 1995–2006.

• Not only has there been an increase in surface temperatures across the world, but similar to the case with CO2 concentrations, the rate of increase has increased in the past half century. The 100-year linear warming trend for the period 1906–2005 was 0.74oC per century, whereas the trend from 1956–2005 was nearly twice as great, at 1.3oC per century.


In the light of these findings, IPCC (2007) concluded that:• There is unequivocal evidence that the global climate system is warming.• Humans are very likely to have caused (and will continue to cause) most of the warming that has occurred

since 1950. • It is very likely that changes to climate will continue well into the future and that such future changes will

be larger than those experienced to date.

The detailed findings of IPCC (2007) have been summarised in a number of reports tailored specifically to Australia. For the purposes of this review, among the most useful are Australian Greenhouse Office (2008), Gardiner (2008), Garnaut (2008), Pittock (2009) and Steffen et al. (2009a,b). The two detailed reports by Steffen et al. (2009a,b) have been collated into a shorter, more-accessible book (Steffan et al. 2009 c). Gardiner (2008) provided a succinct overview of the evidence for climate change and its likely effects on Australia. He concluded that:• Over the period 1910–2004, the average maximum temperature in Australia has risen by 0.6oC and the

average minimum temperature by 1.2oC.• Since 1950, temperatures have risen by 0.4–0.7oC and have been accompanied by more heatwaves, fewer

frosts, less rain in southern and eastern Australia and an increase in drought intensity.• Snow depths in October in alpine areas have decreased by 40% over the past 40 years.• Sea levels have risen by 70 mm since 1950 and have shown an average increase of 1.2 mm year–1 from

1920-2000.

Gardiner (2008) noted the following projections for Australia, based on IPCC (2007) data and assuming a 1990 base year:• A mean warming of 0.1–1.3oC will take place by 2020 for those areas within 800 km of the coast; average

warming of 0.3–3.4oC and 0.4–6.7oC is projected for 2050 and 2080, respectively.• Rainfall will decrease over most of southern and subtropical Australia.• Annual stream flow in the Murray-Darling Basin will decrease by 10–25% by 2050.• There will be an increase of up to 20% in droughts over most of Australia, and an increase of up to 80% in

south-western Australia. • The frequency of very high and extreme bushfire days will increase by 4–25% by 2020, and 15–70% by

2050.

Steffen et al. (2009a,b) drew a distinction between three types of drivers which influenced the susceptibility of Australia’s biodiversity to climate change. First were proximate drivers, factors associated with Australian society that directly affect the biota. Examples include direct removal of animals by hunting and plants by land clearing, ecological impacts arising from river regulation and water extraction, and the introduction of exotic species. The second family of drivers were called ultimate drivers. These indirectly affect the biota, mostly through socio-economic forces or existing institutional arrangements. Examples of these drivers include population growth and the size of the ‘ecological footprint’ of the population. The third set of drivers were global drivers, which includes processes that have a worldwide occurrence, such as globalisation, ozone depletion and international fisheries.

All three sets of drivers conspire to affect Australia’s susceptibility to climate change, but the differentiation between proximate, ultimate and global drivers fails to completely acknowledge a number of the fundamental aspects of the Australian landmass that make it especially susceptible to anthropogenic climate change


(Pittock 2009). Since mainland Australia lies in the lowest latitude band of any large developed country, it is located in a warm region of the world and thus is highly sensitive to further temperature increases. Moreover, the impact of increased temperature will be exacerbated by the large size of the continent, with little opportunity over much of the inland for temperatures to be moderated by proximity to the ocean. In contrast, small island states (e.g. New Zealand, United Kingdom) will be buffered to some extent by maritime influences. Finally, most of Australia is already either arid or semi-arid (49% and 20%, respectively: see Williams 1998), and so has little capacity to withstand further reductions in rainfall.

These natural factors are conflated by a suite of social factors which further increase Australia’s susceptibility to climate change. Agriculture, for example, is overwhelmingly based in the southern parts of the country, and it is these regions that are expected to suffer most from reduced rainfall and stream runoff (Pittock 2009). Already surface- and ground-water resources in south-eastern Australia are severely over-allocated and water shortages are already apparent for urban, commercial and industrial uses, as well as for agricultural production let alone biodiversity conservation (Pittock 2009). There are many examples of counter-adaptive trends within the Australian community that further complicate attempts to respond effectively to anthropogenic climate change. Population growth, for example, is not only among the greatest in the developed world but focused along the coastal fringe which is most susceptible to sea-level rise (Pittock 2009; Steffen et al. 2009a). Perverse economic incentives – for example, for ongoing clearing of native vegetation – add additional stresses to the maintenance of biodiversity in the light of further climate change (Steffen et al. 2009a).

current and projected climate for victoria

CurrentclimateincoastalVictoria

Table 1.19 shows some climate statistics for four sites along the Victorian coast. These data provide a baseline for interpreting projected changes in climate for 2030 and 2070.

Table1.19: Climate statistics for four sites in coastal Victoria. The meteorological station number is shown below each site. Source: Bureau of Meteorology (www.bom.gov.au/climate/averages/tables/shtml, accessed 24/04/2009). The length of record, in years, is shown in brackets. na = not available.

Climate variable Portland (090070)

Melbourne (086071)

Wonthaggi (086127)

Gabo Island(084016)

Mean annual maximum temperature (oC) 17.8 (80) 19.8 (154) 18.7 (39) 17.9 (130)

Mean annual minimum temperature (oC) 9.7 (80) 10.2 (154) 9.5 (39) 12.1 (129)

Median annual rainfall (mm) 832 (126) 645 (154) 919 (94) 889 (143)

Mean number of days of rain (> 1mm) per year 115 (128) 100 (154) 122 (96) 102 (141)

Mean number of cloudy days per year na 179 (54) 204 (40) 152 (52)

Mean annual relative humidity at 9 am/3 pm (%) 77/66 (86/35) 69/53 (54/54) 80/67 (34/34) 80/78 (97/97)


Climateprojectionsfor2030and2070

Detailed climate projections are available for a number of coastal sites in Victoria, including Melbourne (Australia Greenhouse Office 2008), Western Port (Macadam et al. 2008) and Gippsland (McInnes et al. 2005a,b,c). Table 1.20 shows, as an example, projected changes for the climate of Melbourne in 2030 and 2070 over a range of emission scenarios (Australian Greenhouse Office 2008). Annual temperatures in Melbourne are projected to increase by 0.6–1.2oC by 2030 and up to 3.8oC by 2070. The number of days over 35oC could increase by 10–13 days per year by 2030, and by up to 26 days per year in 2070. Note that an increase in mean annual temperature of 1.2oC is about the same as the present-day difference in average maximum temperatures between Melbourne and Wonthaggi (Table 1.19).

Projections for rainfall are highly variable, and range between -9% to +1% (relative to the base year of 1990) by 2030. In general though, the projected decrease in rainfall is always considerably greater than any projected increase; as an example, spring rainfall in 2030 could decrease as much as 16% (for the 10th percentile estimation) but the 90th percentile estimation is a barely significant increase of only 1%. The median projection (i.e. the 50th percentile) for rainfall in 2030 was always negative, and indicated a reduction of 1–7%. For 2070 under emission scenario A1FI, the median projection was a reduction in mean annual rainfall of 11%, ranging from a possible reduction of 21% in spring and 4% in summer. The worst-case scenario indicates a reduction in rainfall of 25–43% by 2070, depending on season. By way of comparison, the difference in mean annual rainfall between Melbourne and the two locations in the extreme west and east of the state (Portland and Gabo Island) is of the order of 30–40% (Table 1.21). Changes in rainfall are not expected to be spread evenly across the four seasons. Winter and spring are expected to show the greatest reduction and the least potential for an increase in rainfall. The seasonal pattern is broadly consistent with the recent modelling of rainfall in south-eastern Australia; Timbal & Jones (2008) predicted that the greatest decline in rainfall in south-eastern Australia was likely to occur in early winter.

Associated with the higher temperatures and (generally) decreased rainfall is a projected increase in evaporation of up to 5% by 2030 and possibly 16% by 2070. Relative humidity is projected to generally fall, by somewhere between 1% in 2030 and 2–4% by 2070. Solar radiation will increase by up to 1.8% in 2030 and 5.7% in 2070. Projections for changes in wind speed are highly variable.


Table1.20: Projected climate changes for Melbourne in 2030 and 2070 under a range of scenarios. Values shown encompass the range over the 10th, 50th and 90th percentiles for 2030 (emission scenario A1B), 2070 (emission scenario B1) and 2070 (emission scenario A1Fl). Source: Australian Greenhouse Office (2008, Appendix B, Table B9).

Climate variable Season 2030(A1B)

2070(B1)

2070(A1Fl)

Temperature (oC) Annual 0.6–1.2 1.0–2.0 1.9–3.8

Summer 0.6–1.4 1.1–2.4 2.1–4.5

Autumn 0.6–1.2 0.9–2.1 1.8–4.0

Winter 0.5–1.0 0.8–1.7 1.5–3.2

Spring 0.6–1.3 1.0–2.1 1.9–4.1

Number of days > 35oC Annual 10.6–12.8 11.9–16.8 15.4–25.9

Rainfall (% change) Annual -9 to +1 -14 to +1 -25 to +3

Summer -11 to +9 -17 to +14 -30 to +27

Autumn -9 to +6 -14 to +10 -25 to +19

Winter -10 to +2 -17 to +3 -30 to +5

Spring -16 to +1 -25 to +1 -43 to +2

Potential evaporation (% change) Annual +1 to +5 +1 to +8 +2 to +16

Wind speed (% change) Annual -6 to +4 -9 to +6 -18 to +12

Relative humidity (%) Annual -1.5 to -0.1 -2.4 to -0.2 -4.7 to -0.3

Solar radiation (% change) Annual +0.1 to +1.8 +0.1 to +3.0 +0.3 to +5.7

Caveatsaboutclimateprojections

Climate projection always involves a number of caveats. Indeed, the term projection is used instead of prediction as an explicit recognition that the modelling does not represent an attempt to forecast the most likely evolution of climate in the future (Anon 2000). We identify five caveats that need to be stated explicitly.

First, climate projection entails a suite of emission scenarios that first need to be developed, and each is associated with its own inherent uncertainties (Anon 2000). The emission scenarios are then applied to a range of global climate models, which yield a suite of projections that vary in projected climate variables according to their climate sensitivity. Outputs are then downscaled to generate modelled regional climates and projections (Whetton 2001).

Second, in addition to inherent mathematical uncertainty in global projections, a range of social factors affect the accuracy of the model outputs and their reporting. For example, IPCC projections are inherently cautious because they are subject to review by representatives of various governments, who may exercise a veto power over the more contentious scientific conclusions (Pearce 2007; Lovelock 2009; Pittock 2009).

Third, there is an inevitable time lag between data collection and release of the IPCC reports and so the most up-to-date information cannot be included in the analysis or reflected in the final reports (Spratt & Sutton 2008; Pittock 2009). Indeed, almost all recent evidence suggests climate changes are occurring at rates faster than those indicated by IPCC (2007). The Lawrence Livermore National Laboratory (Department of the


Environment 2008), for example, recently reported that ocean temperatures and sea level rises between 1961 and 2003 were about 50% larger than estimated in IPCC (2007). Similarly, Rahmstorf et al. (2007) calculated that recent rates of sea-level rise were considerably greater than any of those projected by the 2007 IPCC report.

Fourth, regional projections do not take into account local topographic effects (Australian Greenhouse Office 2008). As noted below, this caveat applies also to the consideration of sea-level rises, where local land subsidence or elevation can exacerbate or mask, respectively, the direct effects of changes in sea level.

Fifth, the choice of base year is crucial, as percentage changes are highly dependant upon which year is chosen for comparison. The year 1990 is commonly used as the base year for regional climate projections. Because of the inherent variability in projections, it is best to look on the values as indications of possible patterns of future climate change rather than as deterministic predictions (Dr Rae Moran, Department of Sustainability and Environment, pers. comm. 22/12/2008).

a model for investigating climate change impacts on coastal wetlands

Although Steffen et al. (2009a,b) differentiated among proximate, ultimate and global drivers as factors controlling the impact of climate change on Australia’s biodiversity, for the purposes of our analysis it is more useful to distinguish between those factors that directly influence the biota, usually via physiological impacts, and those that exert indirect effects, often mediated through species-species interactions or from secondary changes to the environment.

Figure 1.66 shows an approach that can be used to predict changes to mangroves and coastal saltmarsh that result from climate change. It is important to note that envelopes of uncertainty are associated with each step. The model commences with the recognition that different emission scenarios can be envisaged, ranging from ‘business as usual’ options to almost complete isolation of carbon emissions from human activities. Emission scenarios feed into global climate models to yield climate projections at a global scale; global projections are then downscaled to generate regional climate scenarios. Both steps have uncertainties associated with them.

Regional climate projections then feed into an analysis of ecological consequences via two main pathways: i) the direct effect of altered climate on wetland biota (mostly via physiological mechanisms); and ii) a suite of indirect effects. These indirect effects are often associated with secondary changes to the environment, such as altered hydrology and changes to runoff. They can be associated also with complex ecological interactions among species (e.g. rates of herbivory, competition among species sharing a similar niche, or the availability of insect pollinators for flowering plants), effects on major ecological processes (e.g. rates of primary production and decomposition) and the complex patterns of feedback between climate, hydrology and vegetation. The last step, which seeks to analyse the ecological functions of coastal saltmarshes and mangroves, is likely to be the most difficult to complete. With current information it is virtually impossible.

In principle the model could be used to predict climate change impacts at the level of genetic responses, species-level responses (e.g. extinction of a threatened species, changes in the relative abundance and distribution of a common species, or the spread of exotic species) or even at the community or ecosystem scale. Most commonly, though, impacts are considered at the species level.


Figure1.66: Model developed to predict likely climate change impacts on coastal wetlands.

direct impacts on biota

Direct impacts of climate change on the biota of mangroves and coastal saltmarshes are likely to be mediated through two main types of disturbance: i) higher temperatures; and ii) increased ambient CO2 concentrations.

Temperature

Bonan (2002) argued that temperature influences the biota via a wide range of mechanisms, including:• Phenology (the timing of onset of different phases of a plant’s or animal’s development e.g. flowering,

seed germination and establishment of seedlings)• Allocation of resources to above- and below-ground components (e.g. shoots and leaves versus roots and

rhizomes in plants)• Allocation of resources to reproductive versus maintenance activities (e.g. investment in seeds by plants,

success of reproduction in animals)• Patterns of life history and longevity (e.g. shortened life spans due to heat stress or drought)• Variations in competitive or other interactions between organisms.

Emission scenarios

Global climate models

Global climate

projections

Regional scenarios

and projections

Direct impacts: 1

o physiological

effects on biota

Impacts on wetland structure and ecological

function

Indirect impacts: 2o

environmental changes &

ecological interactions Feedback


One consequence of increasing temperatures in south-eastern Australia is that the incidence of frosts is likely to decrease. Sjerp (2007), for example, estimated that there would be at least a 40% reduction in the number of frost days in West Gippsland, and a total loss of frost days in East Gippsland, by 2070. Low winter air temperatures and/or the high frequency of frosts have been suggested as factors that control the southerly distribution of the mangrove Avicennia marina in Victoria (Oliver 1982). Relieved of their current limitation by cold and/or frost, mangroves could expand their distribution, and possibly also their productivity, across southern Victoria.

Increased temperatures may have consequences also for the spread of exotic taxa in mangroves and coastal saltmarsh. Loebl et al. (2006) attributed the recent spread of Spartina anglica in parts of the Wadden Sea (The Netherlands) to increasing temperature. Their study showed that Spartina anglica, after being introduced to the Wadden Sea in 1927, spread rapidly across sheltered shorelines in ensuing decades. It was anticipated that rising sea levels and increased storm activity would limit its spread as niches available for further colonisation narrowed. However, many new sites have become infested since 1985 with dense monotypic swards of Spartina anglica, and the renewed spread coincided with a shift in local temperatures after ~1987. The critical physiological thresholds of 4oC for seed germination and 7oC for photosynthesis were exceeded often after 1987 and the warmer springs, in particular, were thought to have been responsible for increased rates of germination, vegetative growth and geographic spread. In a more recent paper, Nehring & Hesse (2008) came to similar conclusions.

Carbondioxideconcentration

There is currently only a weak understanding of how elevated concentrations of CO2 in the atmosphere will affect coastal plant communities (Mckee & Roth 2008). Even so, higher concentrations can be expected to affect marine and coastal ecosystems by at least two direct mechanisms: i) acidification of marine waters; and ii) greater availability of atmospheric CO2 for plant photosynthesis. The first pathway involves acidification of water as the partial pressure of CO2 increases in the atmosphere and fundamental changes occur in the carbon cycle of the oceans; the resultant fall in ocean-water pH is expected to have major impacts on phytoplankton (Poloczanska et al. 2007; Steffen et al. 2009a,b) and on organisms that build calcareous shells. Nelson (2009), for example, recently reviewed the likely impacts of altered CO2 concentrations on calcified macroalgae.

The second mechanism is more relevant to mangroves and coastal saltmarsh, since their vegetation is overwhelmingly emergent and changes in ambient CO2 concentration will directly affect the growth of emergent plants with different photosynthetic pathways (Bonan 2002). Plants fix atmospheric CO2 in different ways using different metabolic pathways. The different photosynthetic systems are commonly grouped into three main kinds, each of which has a competitive advantage in different environments: • C3 photosynthesis, the pathway used by most plants for photosynthesis • C4 photosynthesis, notably common in tropical grasses, and advantageous in warmer climates and under

water stress• CAM photosynthesis, which occurs in relatively few taxa but is strongly advantageous under extreme

drought stress. It is often associated with succulence.

Plants possessing the C3 photosynthetic pathway (as shown below, most saltmarsh taxa in south-eastern Australia, plus all the mangrove species) have high rates of photorespiration and a variable photosynthetic capacity (Table 1.21). In contrast, plants with the C4 photosynthetic pathway (e.g. grasses such as *Spartina


and Distichlis) show little photorespiration and at full sunlight can be twice as productive as C3 plants. Moreover, because of their more efficient use of CO2, C4 plants use less water to achieve the same rate of primary production as C3 plants. As the optimal temperature for C4 plants is usually greater than that of C3 plants, a warmer and drier climate thus might be expected to favour the former over the latter. The complication is that the C3 plants require higher CO2 concentrations than do C4 plants. Thus we can expect a complex interaction – and possibly replacement of taxa (and plant life forms) with different photosynthetic pathways – in wetlands subject to climate change.

Adam (2008) concluded that it was almost certain that there would be changes in the balance between C3 and C4 plants in coastal saltmarshes as a result of climate change. Earlier, Gray & Mogg (2001) reported that higher temperatures and ambient CO2 concentrations favoured the C3 Puccinellia maritima over the C4 Spartina anglica in Northern Hemisphere saltmarshes. In a more recent study of the importance of photosynthetic pathways on competition between C3 and C4 plants, Mckee & Rooth (2008) compared the responses of Avicennia germinans (a C3 mangrove) with Spartina alterniflora (a C4 saltmarsh grass) to elevated atmospheric CO2, and nutrients in the northern Gulf of Mexico. While monospecific stands of mangroves grew more quickly with elevated CO2, competition from Spartina suppressed growth of the mangrove trees and increased rates of herbivory greatly reduced the survival of mangrove seedlings. The authors concluded that increased CO2 concentrations alone would not result in mangroves replacing Spartina in the region, and that concomitant changes in climate, environmental stress and disturbance would be needed to alter the competitive relationships between the two species. Even so, increased CO2 concentrations could favour the growth of young mangrove saplings where competition and herbivory were low, as well as increasing the biomass (both above- and below-ground) of the mangroves.

Table1.21: Comparison of photosynthetic and productivity characteristics of C3 and C

4 plants. Source: Bonan (2002,

Table 9.2).

Characteristic C3 plants C4 plants

Photorespiration High Low

Photosynthetic capacity Low to high High to very high

Light saturation Intermediate light intensities None or extremely high light intensities

Water use efficiency Low (1–5 g kg–1 H2O) High (3–5 g kg–1 H2O)

Optimum temperature for photosynthesis (oC) 15–25 30–45

CO2 compensation point (ppm) 30–50 0–10

Direct experimental evidence of which photosynthetic pathway is employed by particular plant species is not always available, but what information that is available shows that most plant families only contain one type and only a few families contain multiple types of photosynthetic pathways. These variable families tend to have been better studied (see inter alia Ting 1985; Monson 1989; Ehleringer & Monson 1993; Akhani et al. 1997; Kalapos et al. 1997; Sage 2004, 2007). The literature thus enables us to assign most species to a given photosynthetic pathway with a reasonable degree of confidence, based simply on taxonomic patterns.

The most notably variable families are Poaceae, Cyperaceae and Chenopodiaceae (Amaranthaceae), all of which are well represented in the saltmarsh flora of south-eastern Australia. Thus it would seem that


the Victorian saltmarsh flora (and possibly that of other regions) possesses a particularly wide range of photosynthetic pathways among a relatively narrow group of species. Among the Victorian flora in general, ~94% of species are C3, 5% C4, and 1% CAM. If we examine the Victorian saltmarsh flora in particular, we find that, among those 18 species able to dominate saltmarshes on the coast, many more than expected from the general flora at large are non-C3 species: ~25% (5 species) are C4 and ~5% (1 species) is CAM (Table 1.22). This phenomenon has not, to our knowledge, been the subject of any research but it is likely to be related to the extreme range of environmental conditions tolerated by saltmarsh species. Despite CAM photosynthesis being often associated with succulence, the succulent chenopods in Victorian coastal saltmarsh seem largely to possess the C3 photosynthetic pathway, for reasons unknown.

Table1.22: Likely photosynthetic pathways of some species in the Victorian saltmarsh flora.

Photosynthetic pathway

C3 C4 CAM

Angianthus preissianus Atriplex cinerea Disphyma clavellatum

Austrostipa stipoides Atriplex paludosa

Gahnia filum Distichlis distichophylla

Lawrencia squamata Sporobolus virginicus

Puccinellia spp. Suaeda australis

Sarcocronia quinqueflora

Tecticornia arbuscula

Tecticornia halocnemoides

Tecticornia pergranulata

Wilsonia backhousei

Wilsonia humilis

Wilsonia rotundfolia

indirect impacts: mean sea-level rise and extreme events

Coastal wetlands are susceptible to geomorphological changes arising from rise in mean sea levels and from changes in storm frequency and intensity (Day et al. 2008). Sea levels exert an almost overwhelming influence on coastal wetlands (Woodroffe & Davies 2009) and it is this topic that has received most attention by studies examining likely impacts of climate change on coastal plant communities. Although the information base is still rather small, there are a number of studies that have addressed the role of sea-level rise in mangrove development and others that make some predictions about the likely impact of future rises on these types of ecosystems (e.g. Gilman et al. 2007, 2008; Woodroffe & Davies 2009). It is clear from the stratigraphic record that mangroves have survived rapid rises of sea level in the past and such knowledge provides a basis for predicting how coastal wetlands will respond to sea-level rises that will occur as a result of climate change. The predicted responses, however, are not always in agreement: in the case of mangroves on small oceanic islands, there is some controversy as to whether or not they will survive in the face of rises in mean sea level (Woodroffe & Davies 2009). The situation with mangroves on the shores of larger land masses is


in less dispute, although it seems clear that the main response in the past has been for mangroves to migrate landwards. The issue today, of course, is that the extensive development of the hinterland (see Chapters 1.11 and 6.2) will often preclude such migration.

With regard to other types of coastal wetlands, the early examinations of likely sea-level rise impacts suggested that there would be a large-scale loss of wetland habitat (e.g. see the review by Pratolongo et al. 2009). More recent studies have taken a more measured approach and it has been realised that the issue is not so much sea-level rise per se but changes in relative sea levels that will be the fundamental determinant of responses. In other words, if coastal wetlands can maintain their elevation, they can keep up with – as opposed to catch up with (Woodroffe & Davies 2009) – sea-level rise (Pratolongo et al. 2009).

Projectedriseinmeansealevels

The 4th assessment of global climate change (Intergovernmental Panel on Climate Change 2007) presented a number of projections for future sea-level rises. Under the high emissions scenario (A1FI), a rise in mean sea levels of 0.26 to 0.59 m by 2100 was projected. The estimate is based on thermal expansion of the oceans and contributions from melt-water from mountain glaciers and increased ice flow from Greenland and Antarctica (Pittock 2009). The projected increases in sea level, however, do not take into account uncertainties in carbon-climate feedback loops, nor the full effects of changes in ice-sheet flow. If such changes were to be incorporated, sea levels could rise by an additional 0.1 to 0.2 m.

There is growing evidence that the IPCC (2007) estimates for future sea-level rise are too low. Steffen et al. (2009a, page 90), for example, concluded that ‘…the IPCC sea level rise projection of up to about 1 m by 2100 may be far too low in the light of its own projected temperature change.’ Among the most compelling evidence for accelerating sea-level rise is that presented by Rahmstorf et al. (2007), who argued that the sea-level rise by 2100 would likely be of the order of 0.5 to 1.4 m. Other authors have provided even more dire predictions, including those by James Hansen and co-workers that rises of up to 5 m may occur by the end of the century (Table 1.23). The higher estimates of sea-level rise are based on a linear (Rahmstorf et al. 2007) or an exponential (Hansen 2007, cited in Pittock 2009) relationship between sea-level rise and average warming.

Table1.23: Projections of global sea-level rise in the 3rd and 4th IPCC assessments, compared with some more recent projections. Source: Pittock (2009, Table 6).

Case Warming to 2100 (oC) Sea-level rise to 2100 (m)

3rd IPCC assessment (2001) 1.4–5.8 0.09–0.88

4th IPCC assessment (2007)• Low case (Scenario B1)• High case (Scenario A1Fl)

1.1–2.92.4–6.4

0.18–0.380.26–0.59

Rahmstorf et al. (2007) 1.4–5.8 0.5–1.4

Hansen (cited in Pittock 2009) Up to 5

The most recent Victorian Coastal Strategy proposed that planning for a rise of sea level of not less than 0.8 m by 2100 should be implemented until better national benchmarks for coastal vulnerability had been developed (Victorian Coastal Council 2008).


Regionalvariationsinsealevel

The values shown in Table 1.23 refer only to the incremental component of sea-level rise and to mean sea levels. Although there is considerable variation in the magnitude of projected rises, that there will be a rise in mean sea levels over coming decades is one of the most confident projections of climate change science (Walsh 2004).

It should be noted, however, that the rise in sea levels will not be limited to a 2100 time frame. Even if CO2 emissions were to cease immediately, sea levels would continue to rise for centuries because of the slow but continual warming of the oceans and contraction of the Greenland ice sheet (Pittock 2009). Nor will sea levels rise uniformly in different regions; isostatic and tectonic land movements will contribute to regional differences in sea levels (Pittock 2009). Local changes in sea levels may be caused also by modifications to the mouths of estuaries and coastal embayments, through geomorphological changes and engineering works, for example by dredging (Bird 2006). Walsh (2004) provides, as an example of regional variation, tidal records since the 1940s in Port Pirie and Port Adelaide. Sea levels at these two sites show divergent trends, owing to local land subsidence near Port Adelaide. Other sites with more stable geologies and long historical records (e.g. Fort Denison in Sydney, Fremantle in Western Australia, Port Arthur in Tasmania) show that average sea-level rises in Australia are broadly similar to the global average of about 1.7 mm per year. The instrumental record shows that sea levels have risen an average of 0.86 ± 0.12 mm year–1 at Fremantle (1915–1998), 1.38 ± 0.18 mm year–1 at Sydney (1897–1998) and 1.2 ± 0.2 mm year–1 at Port Arthur (1890–1998) (Abbs 2002). As noted in Chapter 6.2, rises in the relative mean sea level around the Australian coastline for the period 1920–2000 has been estimated to be around 1.2 mm year–1 (Church et al. 2006).

Importanceofextremeevents

Perhaps as important as the incremental rise in mean sea levels, however, is the potential impact created by extreme events. Extreme events include those rare storms or floods that temporarily increase further the height of the sea and, combined with increased wave action, result in greatly increased erosion and penetration of the ocean inland than would otherwise be the case (Walsh 2004). Bryant (1990) outlined the range of factors which contribute to variations in extreme sea level, including king tides, storm surges, wave set-up and run-up, shelf waves and seiching. Water levels are likely to be affected also by large variations in river discharge arising from changed precipitation, runoff or storm frequency. Marine intrusions can be exacerbated by such riverine floods, in which case increased river discharge causes estuarine waters to ‘back up’ into the swollen rivers and inundate adjacent lands. Cahoon (2006) reviewed the likely impact of major storms on coastal wetlands and concluded that they exerted geomorphological effects by a diverse suite of processes.

Extreme events may cause not only acute impacts, however, but initiate also a suite of long-term effects. For example, storm surges may so effectively attack coastal sand dunes that the dunes are permanently breached, with the result that lagoons that were formerly separated from the ocean by a line of sand dunes become more-or-less permanently open to the sea (Walsh 2004). Such an event is considered likely for the Gippsland Lakes, where the single line of sand dunes along Ninety Mile Beach could be easily breached by an extreme event; under these circumstances, Lake Reeve would shift from an intermittently wetted lagoon with extensive areas of saltmarsh to one experiencing near-permanent inundation by ocean water.


The incidence of extreme events is related to climate change in at least two ways. The first involves a clear physical relationship between altered meteorological conditions and changes in weather. For example, an increase in mean atmospheric temperature increases the saturation vapour pressure of water and thus the water content of the atmosphere; other factors being equal, this leads to an increase in the intensity of precipitation. The increase in atmospheric water increases also radiative warming and, when combined with radiative cooling of the free atmosphere, tends to reduce the stability of the atmosphere and also leads to more intense precipitation (Mitchell et al. 2006).

The second is more statistical in nature, and results from simultaneous changes in the mean and variance of meteorological variables. The combination of a slight change in mean and a slight change in variation can result in a marked increase in the frequency of extreme events (Figure 1.67).

Figure1.67: Relationship between an increase in slight mean and a slight increase in variance, giving rise to a marked increase in the incidence of extreme events. Source: Walsh (2004, Figure 2).


ExtremeeventsandtheVictoriancoastline

Detailed investigations have been undertaken of possible impacts of extreme events and storm surges for two parts of the Victorian coast: i) Western Port (Western Port Greenhouse Alliance 2008); and ii) eastern Victoria, including Corner Inlet and the Gippsland Lakes (McInnes et al. 2005b,c; McInnes & Hubbert 2006; Sjerp 2007). Both sets of analyses were preceded by broader projections for regional climate (Macadam et al. 2008 for Western Port; McInnes et al. 2005a for eastern Victoria). Both regions have extensive areas of mangrove and/or coastal saltmarsh.

Western Port

Mean sea-level rises of 0.17 m and 0.49 m are projected for Western Port by 2030 and 2070, respectively (Table 1.24). These are relatively small values when compared with the heightened sea levels that are expect to arise from those extreme events that are considered likely under even the most reasonable climate change scenarios. Storm tides at Cowes (on Phillip Island), for example, could reach 2.29 m by 2030 and 2.74 m by 2070. Not only will storm surges be higher than those currently experienced, but they will occur more often. Indeed, storm surges with a current return interval of 1:100 years would have a new average return interval of only 1:40 or even 1:6 years by 2030, and 1:20 or 1:1 years by 2070. In other words, what is currently a severe storm that occurs only once a century could become an annual event by 2070. Linked with the increase in the severity and frequency of storm surges is a projected increase in extreme rainfall and extreme winds (Table 1.24).

Table1.24: Projected changes in the incidence of extreme events in Western Port for 2030 and 2070. ARI = Average return interval. Source: Western Port Greenhouse Alliance (2008, Table A).

Event 2030 2070

Sea-level rise (m) 0.17 m 0.49 m

Storm tide at Cowes (maximum height in m for 1:100 ARI) 2.29 m 2.74 m

Storm surges (change to current 1:100 ARI frequency) Decrease to 1:40–1:6 Decrease to 1:20–1:1

Extreme rainfall (12 hour, % change) +3 to +22 +17 to +61

Extreme winds (% change) -1 to +5 -3 to +14

The practical implications for the coastal wetlands of Western Port are that the area of land subject to inundation during a 1:100 year storm surge may increase by 4–15% by 2030 and by 16–63% by 2070. As shown in Figure 1.68, such increases in the incidence of severe storm surges will have implications not only for coastal saltmarsh but also for the residential housing that commonly abuts it. If engineering solutions are implemented to ‘protect’ urban infrastructure (e.g. sea walls), coastal saltmarsh will be affected by major changes to hydrological and salinity regimes arising from human efforts to defend the coast, as well as from more direct climate change induced impacts. Such human interventions designed to ‘adapt’ to climate change may constitute a wide family of proximate or ultimate drivers (Steffen et al. 2009a) that exacerbate impacts on coastal vegetation.


Figure1.68: Housing abutting coastal saltmarsh and at high risk of sea-level rise, Hastings.

Gippsland and eastern Victoria

In the case of the eastern Victorian coast, the projected increase in heights of storm tides is relatively small (up to ~0.2 m), because storm surges along this part of the coast are currently quite small (McInnes et al. 2005b). The anticipated mean rise in sea levels, therefore, is likely to have a greater impact than storm surges on extreme sea levels over much of eastern Victorian coast. Only under high-wind scenarios would storm surges contribute more than about 20% of the anticipated total rise in sea levels. Table 1.25 shows the projected storm surge heights for various locations along the Gippsland coast under a range of climate change scenarios (e.g. different wind-speed projections and mean sea-level rises).

The values shown in Table 1.25, however, are likely for two reasons to be underestimates. First, McInnes et al. (2005b) noted that land subsidence was common across many parts of the Gippsland coast; thus the change in relative sea level would be greater than the simple increase in mean sea levels would otherwise indicate. Second, the effect of an increase in the ferocity of floods in the rivers that discharge into the Gippsland Lakes was not taken into account: McInnes et al. (2005b) argued that the flood risk at locations such as Lakes Entrance would be increased by the combination of mean sea-level rise, increased storm surges and increased river discharge during floods leading to ‘back up’ within the lakes.

Table1.25: Projected changes in total sea level (i.e. combination of mean sea-level rise and storm tides) for various locations along the eastern Victorian coast. The values indicated are for a 1:100 year average return interval, over a range of climate change scenarios. Source: McInnes et al. (2005b, Table A3).

Location Storm surge height (m)

Current 2030 2070

Venus Bay 1.74 1.76–1.96 1.77–2.40

Tidal River 1.73 1.74–1.95 1.75–2.40

Port Welshpool 1.71 1.73–1.92 1.74–2.35

Lakes Entrance 1.40 1.41–1.61 1.41–2.06


A subsequent study (McInnes & Hubbert 2006) examined the likely future incidence of extreme sea levels at Corner Inlet and in the Gippsland Lakes. The predicted 1:100 year storm tide heights of 2.08 m, 1.65 m and 0.98 m estimated in this later study for Stony Point, Port Welshpool and Lakes Entrance, respectively, agreed well with estimates of 2.00 m, 1.60 m and 0.96 m made in the previous report (McInnes et al. 2005b).

For Corner Inlet, the greatest inundation arising from increased storm surges and a rise in mean sea levels was expected across the islands and along the northern coastline. It was estimated that inundation at Port Franklin, Port Welshpool and Port Albert would increase by between 15–30% by 2070 under the combined high-wind speed/high mean sea-level rise scenario. These areas are fringed by some of the most extensive saltmarsh and mangrove areas in Victoria (see Chapter 5).

For the Gippsland Lakes, the greatest impact of storm tides is expected to be in the existing wetland areas around the main lakes and in Lake Reeve, the latter currently an extensive (and Ramsar-listed) coastal saltmarsh. The greatest increase in extreme sea levels is expected near Lakes Entrance (e.g. 1.57 m under the 2070 high-impact scenarios), compared with 1.18 m at Metung and 0.89 m at Paynesville. The total area expected to be inundated in the Gippsland Lakes by 2030 ranged from 25 km2 (under the most parsimonious conditions) to 50 km2 (under the high-wind speed/high mean sea-level rise scenario). A further 25% increase in the area inundated was projected for 2070.

The most recent analyses of climate change assume more severe pre-conditions and predict greater and more widespread impacts than those that were undertaken only 10 years earlier. This shift is illustrated by a comparison of the most recent reports by McInnes and co-workers (2005, 2006) and Sjerp (2007) with the decade-older analysis of the climate change vulnerability of the Gippsland Lakes by Lawson & Treloar (1996). The 1996 study assumed a mean sea-level rise of only 0.3 m and concluded that there was little likelihood of the frontal dunes being breached. Sjerp (2007), in contrast, concluded that a worst-case scenario of 1:100 year ocean storm, high storm surge, 0.8 m sea-level rise and flooding in Lake Reeve would see a breach of the sand dunes at The Honeysuckles, a site just east of Seaspray. Other breaches could occur in Bunga Arm at the Blowholes and in the channels east of Rotamah Island.

Implicationsofincreasedsealevels(meanandextreme)forcoastalwetlands

Before the likely impacts on mangroves and coastal saltmarsh are explored, it is worth re-stating that the current position and extent of coastal saltmarsh and mangroves is a result of relatively recent geomorphological events. Knox (1986), for example, noted that saltmarshes along the coast of the USA appear to have originated mostly within the past 3,000 years, when sea-level rise slowed sufficiently to allow this type of coastal wetland to develop; Woodfroffe & Davies (2009) stressed that mangrove ecosystems have responded in the geological past to increasing sea levels by migrating into the previously terrestrial hinterland. Tropical Australia provides an example of such a change: Crowley (1996, cited in Hobday et al. 2006c) showed that the mangrove swamps that are now extensive in northern Australia developed only some 6,000–8,000 years ago, as sea levels rose. In Victoria, the Gippsland Lakes and their associated saltmarshes are only of Holocene origins (Bird 1993).

Historical studies such as these indicate that changing sea levels can result in the destruction, re-organisation or movement of coastal ecosystems. Bryant (1990), for example, cited the work of Jones et al. (1979) which showed that mangroves at Bulli (Sydney) were able to colonise new sediments and keep pace with sea-level


rises of approximately 15 mm year–1 over the recent Holocene. Such rates of sea-level rise are within the predicted range of increases under most climate change scenarios to 2030 or 2070. In other cases, however, palaeobotanical studies have shown that sea-level rises have caused widespread changes in coastal plant communities. Walker & Singh (1981), for example, showed that, at Wilsons Promontory, a marine incursion converted a formerly freshwater swamp to saltmarsh about 7,000 years ago which, after the accumulation of sand, changed into the current Melaleuca ericifolia scrub (see Ladd 1979 for details).

Notwithstanding these studies, it is clear that sea-level rise will exert a major influence on nearly all aspects of the ecology of coastal saltmarshes and mangroves. In fact, Gilman et al. (2008) concluded that, of all the likely consequences of climate change, a rise in relative sea level was the greatest threat to mangroves. The most obvious impact is that coastal wetlands will be inundated more frequently and more deeply with seawater. The most low-lying areas may become permanently inundated (e.g. Lake Reeve, as discussed above). Under these conditions, the simplest expectation is that saltmarshes would be replaced by mangroves, and mangroves by seagrasses. In more elevated positions, the frequency of inundation from either storm surges or floods will increase and would be expected to have profound effects on the water and salinity regimes in the wetlands. As an example of a site-specific impact on coastal wetlands, Sjerp (2007) predicted that many of the freshwater and brackish-water wetlands that currently fringe the Gippsland Lakes (e.g. Tucker Swamp, Clydebank Morass) would become inundated with saline water as a result of increasing sea levels and other factors related to climate change. Lake Reeve, currently a large saltmarsh along the seaward side of the Gippsland Lakes, was projected to ‘…become wetter for significantly longer periods of time, if not permanently inundated’ (Sjerp 2007, page 11). Increased inundation would be expected to favour also *Spartina, as it is commonly found on the seaward side of even mangrove communities.

Williams (1990) examined the possible impact of climate change on some New South Wales saltmarshes, and concluded that a 0.5 m sea-level rise would create an additional 243 ha and 411 ha of mangrove and saltmarsh at Minnamurra and Crooked Rivers (130–150 km south of Sydney). A critical value in making such assessments is knowledge of the upper and lower height datums for mangrove and saltmarsh communities; as few relevant topographic data existed at that time, Williams (1990) assumed that mangroves would grow no deeper than mean tide and saltmarshes no higher than 1 m above mean sea level.

Erosion of coastal shorelines would be increased by higher mean sea levels and by an increase in storm surges and other extreme events. There is a simple relationship between sea levels and landward erosion, known as Bruun’s Rule (Walsh 2004). Although highly simplified by comparison with what is known to occur as coasts erode (Abuodha & Woodroffe 2006), Bruun’s Rule predicts there will be about 100 m of coastal erosion for every 1 m of sea-level rise. In other words, the land will retreat about 100 m for every metre the sea level rises, and the mean sea-level rises of 0.5–1.4 m indicated by Rahmstorf et al. (2007) translate to the loss of a strip of coastal land 50–140 m wide.

Increased erosion will result not only from a rise in mean sea levels but also via storm surges and more general increases in wind speed and tidal action. Erosion will be most severe during extreme events, and it would seem that particularly susceptible to storm surges are sandy coastlines (Walsh 2004). The erosion of sandy beaches could have particularly strong impacts on the breaching of coastal barriers that currently protect coastal lagoons from the ocean. Such sandy dunes do not normally survive overtopping by high-intensity storms (Nott & Hubbert 2003) and, at least for the Gippsland Lakes, it has been predicted that the seaward line of


dunes could be breached by the combination of higher mean sea levels and storm surges (Sjerp 2007). In his analysis of the impacts of higher sea levels on Port Phillip Bay, Bird (2006) argued that increased sea levels, combined with the associated increase in wave action and erosion, would see the loss of the Mud Islands (and their saltmarshes) near the entrance of the Bay. The mouths of the creeks and rivers that discharge into Port Phillip (e.g. Yarra River, Werribee River) also are likely to experience increased erosion (i.e. become wider and deeper) as tides become higher.

It is clear that storm surges and other extreme events will have severe impacts on Victorian coastal saltmarshes and other types of coastal wetland because of their erosive capacity as well as effects on saline water penetrating deep into formerly terrestrial zones. An analogy is provided by studies of erosion in Northern Hemisphere saltmarshes: studies of saltmarshes along the south-east coast of England have shown that the combination of strong winds, high tides and increased wave height has contributed to the extensive loss of saltmarsh since the 1970s (Wolters et al. 2005).

other indirect impacts

Increasedevaporationandcreationofhypersalineconditions

Although the most immediate of the hydrological impacts on saltmarshes and mangroves will be caused by rises in mean sea levels and extreme events, other indirect effects can be expected to arise also from changed patterns of rainfall, wind speed and evaporation. In Western Port, for example, climate change projections indicate increases of 0.3–5.3% in solar radiation by 2070 and a decrease in relative humidity by up to 3.2% (Macadam et al. 2008). Projections for wind speeds are highly variable, but mean wind speeds are more likely to increase than to decrease across coastal Victoria and Bass Strait in summer, winter and spring (McInnes et al. 2005a). The combined effect of increased insolation, lower relative humidity, higher wind speeds and an increase in the number of hot days (Table 1.20) will result in increased rates of evaporation, leading to a more rapid onset of hypersalinity in saltmarshes and thus to exacerbated water and salt stress for the resident biota. In this case, an altered climate will directly result in altered salinity regimes within coastal wetlands and, in particular, in saltmarshes that do not receive the ameliorating effect of daily inundation by the tides.

Effectsonfreshwaterrunoffandriverdischarge

Impacts on both salinity and flooding regimes are likely to occur as a result also of changed patterns of freshwater runoff. The effects could be experienced either as a direct effect of decreased local runoff onto saltmarshes, or as an indirect effect arising from altered base flows and patterns of over-bank flooding from nearby rivers. In south-eastern USA, coastal saltmarsh has experienced widespread die-off over the past decade or so, and drought is thought to be one of the major causes of the losses. Drought not only leads to reduced freshwater inflows, and thus hypersalinity in coastal environments, but at least in North American saltmarshes has been shown to allow increased grazing pressure (by gastropods), which causes a cascade of ecological disruption (Silliman et al. 2005; Osgood & Silliman 2009).

Changes to rainfall (a general but unseasonable decrease) and evaporation (an increase of 5–16% by 2070: Table 1.20) are expected to result in reduced runoff across much of south-eastern Australia. The relationship between reduced rainfall and reduced runoff will not be 1:1; an increasingly dry catchment will yield proportionally less runoff than a simple calculation based on reductions in rainfall would suggest. In the case


of south-western Western Australia, for example, empirical observations have shown that a decline of 10–20% in rainfall has resulted in a 40–60% decline in runoff over the past three decades (www.climatechange.gov.au/impacts/water.html, internet resource accessed 3/02/2009). Similar responses are predicted for the rainfall:runoff relationship in the Melbourne region; Timbal & Jones (2008) predicted that the percentage reduction in runoff was likely to be about twice as great as any reduction in rainfall in Melbourne’s water-catchment areas.

Changes in groundwater behaviour and other catchment-scale processes will also contribute to the direct effects of reduced rainfall on runoff. The base-flow of rivers in south-eastern Australia is derived largely from groundwater, and groundwater recharge is expected to be lower under conditions of higher temperature, reduced rainfall and greater evaporation. Over the longer term, freshwater runoff is expected to decrease, both in terms of reduced base-flow due to the lowered groundwater tables and in terms of an overall decrease in runoff due to less rainfall (Williams et al. 2001). Jones & Durack (2005) estimated the effect of climate change on runoff from Victorian catchments. On a statewide basis, it was predicted that decreases in runoff would vary from, at best, 0–20% in the east and south of Victoria and, at worst, to a decrease of 5–45% in the west by 2030. By 2070, East Gippsland could experience an increase in runoff of up to 20%, but changes across the rest of Victoria were, at best, a reduction of 5–10% and, at worst, a reduction of more than 50%.

Although it is likely that freshwater runoff and stream discharge will generally be lower in a future climate change world, it is also highly likely that the incidence of extreme events will increase. The possibility of an increase in storm surges and other extreme oceanic events has been discussed earlier; a growing body of evidence that suggests that extreme events will also become more common with freshwater ecosystems. Schreider et al. (2000), for example, predicted an increase in the magnitude and frequency of floods in south-eastern Australia, although conditions would vary from place to place. In the case of the Upper Parramatta River (Sydney), it was predicted that a 1:100 year flood under current conditions would become a 1:44 year event if CO2 concentrations were to double. The impact would be greater in the Hawkesbury-Nepean River system, where a current 1:100 year event would become a 1:35 year event. Similarly, the incidence of the other type of extreme freshwater event – drought – is expected to increase markedly with climate change (Nicholls 2004; Lake 2008).

Increasedincidenceofbushfiresandpeatfires

Bushfires could be responsible for two types of climate change induced impact on coastal wetlands: i) direct effects on the burning of wetland vegetation; and ii) indirect effects arising from the loss of terrestrial (forest) vegetation and related changes in runoff and river discharge. The topic of fires in coastal saltmarsh was discussed in Chapter 1.11, and it is possible that an increased incidence of peat fires may occur subsequent to climate change. Williams et al. (2001) examined likely impact of climate change on bushfire risk in terrestrial environments in Australia. As outlined in Chapter 1.11, saltmarshes are occasionally subject to fires (e.g. see Figure 1.44) and further drying of their peat substrata would likely increase the fire risk.

Stream flows typically increase immediately following the destruction of mature Eucalyptus forest by fire or logging, due to both a reduction in interception and in rates of evapo-transpiration. Between 2 and 20 years later, however, stream flow decreases as the forest trees regrow. Evapo-transpiration from eucalypt-dominated forests in south-eastern Australia is typically at a maximum, and stream flow at a minimum, for forests of 20–30 years age; after this time stream flow slowly increases again as the forest trees approach maturity. Even


so, reductions in stream flow remain evident for at least a century after logging or bushfire (e.g. see Kuczera 1985; Watson et al. 1999). Recent analyses have been undertaken of the likely impact of the 2003 and 2006 bushfires on freshwater flows into rivers that discharge into the Gippsland Lakes (Tilleard et al. 2009). Under current climate conditions, discharge for individual rivers affected by the 2006 Gippsland fires is expected to fall by up to 42% compared with pre-fire conditions. For the Gippsland Lakes, the maximum reduction in inflows as a result of the 2006 bushfires is expected to be around 735 GL year–1, or about 26% of mean annual inflow from the most affected rivers.

Ecologicalinteractions:weedinvasions

Although evidence was presented earlier on the likely role played by climate change induced increases in temperature on the invasion of European saltmarshes by Spartina anglica, it is probably unlikely that the composition of the current weed flora of Victorian coastal saltmarsh will change greatly in respect of potential climatic envelopes (i.e. temperatures and rainfall). Saltmarsh and mangrove habitats are naturally restricted geographically and new territory currently unoccupied by weeds will not become available as a result of migration along a climatic temperature gradient.

That conclusion, however, does not allow the inference that smaller scale changes will not occur. The terrestrial weed flora of upper saltmarsh or the aquatic macrophyte *Spartina in lower saltmarsh and mangrove, for example, may change regionally in floristic composition and/or structure as the current weed flora – which is mostly European annuals and perennial herbs – is sifted ecologically in response to declining rainfall and higher temperatures. Moreover, tolerance to drought varies considerably among the current weed flora, and it is possible that some species may achieve competitive advantage because of their relative drought tolerance. For example, different species in the genus of exotic annual grasses Parapholis, which are among the most ubiquitous and abundant of exotic saltmarsh weeds (Chapter 1.11), have different moisture tolerances: *Parapholis incurva appears to be more drought tolerant than *Parapholis strigosa and the former species appears to be largely replacing the latter, at least in saltmarshes along the central Victorian coast (Carr 1982).

Increasing temperatures may particularly favour one weed species in the current saltmarsh flora, *Paspalum vaginatum. This species is a member of a genus with a predominantly warm-temperate and tropical distribution. In Victoria it is considered exotic (Walsh 1994) whereas it is considered to be native in New South Wales ( Jacobs & Wall 1993). The climate in western Victoria may become more amenable and its range could extend much more to the west of the state. It is currently rare outside of East Gippsland.

A second species that may respond in a similar way is the shrub *Baccharis halimifolia, which is a major threat to saltmarsh vegetation in more northerly coastal saltmarsh (see Chapter 1.11 and Saintilan 2009a,b). It was introduced in Queensland, where it has since become widely naturalised, and is now naturalised also on the north and central coasts of New South Wales (Porteners 1992). It is reported by Saintilan (2009b) to be extending its distribution south. The natural distribution of *Baccharis halimifolia is eastern USA (New York, Arkansas, Florida, South Carolina and Texas), areas that are generally warmer than southern Victoria. Whether or not Victorian saltmarshes are temperature-limiting for this species is not known, but if it is limited currently by low temperatures a spread further southwards is not impossible.


prior studies of climate change impacts on coastal wetlands

The 10th meeting of the Ramsar Conference (Changwon, Republic of Korea, November 2008) examined the possible impacts of climate change on the ecology and human-use values of wetlands (Ramsar Convention 2008). The analysis, however, is too general for the specific case of saltmarshes and mangroves along the Victorian coast.

Of more use in identifying the elements likely to be affected is the analysis undertaken by Meith (1991) of the impacts of climate change on Mediterranean wetlands. He concluded that droughts, floods, water shortages and catastrophic weather events would exacerbate an ongoing degradation of regional landscapes around the Mediterranean coast of Europe. A lengthening of the dry summer period and a shift northwards of rainfall patterns were predicted to expand arid zones around the European Mediterranean; salinisation problems would intensify, especially around irrigated landscapes. Positive feedback loops might easily be established, whereby irrigation expanded under the drier climate and the problems of secondary salinisation became even more pressing. Stream flows were predicted to be reduced, especially as evapo-transpiration increased as a result of rises in air temperatures. Soil structure could suffer, mainly as a result of a lowering of the already low organic-matter contents, poor water-retention ability and decreased cover from sclerophyllous scrublands (such as the maquis), leading to severe erosion and an increase in the input of sediments into wetlands. Meith (1991) argued that lakes and wetlands would likely experience longer dry periods and become increasingly filled with eroded sediment.

Simas et al. (2001) undertook a modelling study of the likely impacts of rising sea levels on saltmarshes associated with the Tagus Estuary in Portugal. They concluded that only the worst case sea-level rise scenarios were likely to have adverse impacts, and that areas with high tidal ranges were likely to be buffered somewhat by greater sediment transport and accretion. Given, however, that the most recent IPCC reports suggest more severe climate change (e.g. even higher air temperatures and greater rises in mean sea level) than the early ones that were presumably used by Simas et al. (2001), their conclusions may have to be tempered with the recognition that climate change induced disturbances will be greater than were used in their models.

Scavia et al. (2002) assessed the likely impact of some climate change variables (e.g. sea-level rise, altered patterns of freshwater discharge, etc.) on coastal and marine systems of the USA. They concluded (page 153) that ‘Natural biological and geological processes should allow responses to gradual changes, such as transitions from marsh to mangrove swamp as temperatures warm, as long as environmental thresholds for plant survival are not crossed’. Although sea-level rise was believed to be the greatest risk factor, other factors posed significant threats to coastal wetlands, including altered precipitation, climate change induced alterations to land use in catchments that surrounded coastal wetlands, increased CO2 concentrations and increased air temperatures.

Morris et al. (2002) examined a particular case of changes in relative sea level for Spartina-dominated wetlands in the USA. They developed a model that predicted coastal saltmarsh dominated by Spartina alterniflora would be relatively stable against changes in relative sea level when surface elevations were greater than those required for the highest rates of primary production by the plants. However, when surface elevations were less than the critical level, ecosystems should become unstable. High rates of sediment deposition in many estuaries along the south-east coast of the USA may be sufficient to offset the observed rates of relative sea-


level rise. The marshes predicted to be most resilient to sea-level rise were those above their optimal elevation, as they would be less vulnerable to monotonic rises in relative sea level as well as to variations in mean sea level. A more pessimistic conclusion was reached by Chmura et al. (1992), who predicted that coastal saltmarshes along the Louisiana coast would probably not be resilient to rises in mean sea level. In a subsequent paper, Morris (2007) proposed a planting regimen that could maintain Spartina-dominated coastal saltmarsh in the face of sea-level rise.

A number of Australian studies have investigated the possible impacts of climate change on coastal wetlands. Unfortunately, there is little or no empirical information on ecological changes that have occurred so far in response to climate change (Hobday et al. 2006c). In one of the first analyses of the potential for climate change impacts on Victorian coastal vegetation, Vanderzee (1988) argued that the well-defined pattern of plants in Victorian saltmarshes reflected strongly the extent, timing and duration of inundation by seawater. Thus even small changes in tidal inundation would be expected to have large consequences for saltmarsh floristics and, probably, productivity. He concluded that recent changes in saltmarsh communities at Corner Inlet included the landward migration of plant species in response to tectonic submergence of the coastline. Such changes were proposed to be a good model for predicting possible impacts arising from climate-induced rises in sea levels. A decade-long hiatus seems to have taken place before the topic of climate change induced impacts on coastal wetlands was again studied.

Walsh (2004) reviewed the possible effects of sea-level rises on saltmarshes and mangroves. He reported that, because of their vulnerable position, saltmarshes would be extensively inundated as the sea level rose yet there had been little or no scientific studies of the likely impacts of sea-level rise on saltmarshes in Australia.

A series of reports was published by the Australian Greenhouse Office in 2006 on the impacts of climate change on coastal and marine ecosystems. Voice et al. (2006) concluded that, although the vulnerability of saltmarsh and mangrove communities varied around the coast and with species, the productivity and range of these plant types could be negatively affected by climate change. They noted that predicting the impacts of climate change on coastal wetlands was made more difficult because the ecological information necessary to make any judgements was scattered across diverse and often hard-to-obtain sources. Even so, Voice et al. (2006) tabulated what they saw as the likely impacts of climate change on saltmarshes and mangroves: a summary is shown in Table 1.26.


Table1.26: Likely impacts of climate change on coastal saltmarsh and mangroves. Source: Voice et al. (2006, Table 4a).

Climate change driver Likely impact Sensitivity and confidence

Known thresholds

Rise in mean sea levels Vegetation loss High& Good

Unknown

Extreme storms Reduction in vegetation cover Low-Medium& Good

Tolerant of storms unless threat is combined with other stressors

Increased waves and wind Reduction in vegetation cover Medium-High& Moderate

Impact greater when combined with sea-level rise

Increased CO2 concentration in atmosphere

Increased primary productivity Low& Good

Increase in productivity up to 30%; limited by water stress and salinity

Increased sea temperature Increased respiration Low-Medium&Good

Unknown

Increased air temperature Altered productivity and changes in species composition

Low-Medium &Good

Impact depends on latitude

Decrease in humidity Altered productivity and changes in species composition

High&Moderate

Unknown

Decreased rainfall Reduced productivity, invasion of mangroves into saltmarsh, hypersalinity in saltmarshes

High&Good

Unknown

Increased rainfall Increased productivity and diversity

Low&Good

Unknown

Hobday et al. (2006a,b) concluded that the major climate-induced threat to mangroves was a rise in mean sea level. There were musings about whether increased CO2 concentrations would lead to increased mangrove productivity, but the authors felt that other factors – such as reductions in nutrient availability and adverse salinity regimes – could offset any CO2-related productivity increases. Hobday et al. (2006c) provided a more detailed analysis of the possible impacts of climate change on mangroves, from an Australia-wide perspective. Of the indirect processes controlling the performance of individual wetland taxa (see Figure 1.65), sea-level rises are among the most critical for mangroves. Rising sea levels were proposed to be likely to stress existing mangrove plants and, depending on the relative rates of sea-level rise versus sedimentation, allow mangroves to track rising sea level by colonising more landward areas (Hobday et al. 2006c). Similar migration in southern Australia will be limited severely by the scale and intensity of coastal development in the hinterland. (A similar situation applies worldwide: Gilman et al. (2007) reported that any landward migration of mangroves in response to sea-level rise in Samoa was often constrained by coastal development.) The main issue here is whether the spatial and temporal scales associated with climate change are shorter than those available for ecosystems to respond to the new sets of conditions (Day et al. 2008).

Hobday et al. (2006c) noted also the sensitivity of mangrove ecosystems to even slight changes to hydrological regimes, induced by either shifts in tidal inundation or fluxes of freshwater from the hinterland and nearby rivers. Thus shifts in freshwater penetration into mangroves would be expected to have profound impacts


on rates of primary productivity, and perhaps even the ongoing persistence of mangroves in some areas. For this reason, the discussion of climate-related effects on freshwater runoff and stream discharge in southern Australia could have ramifications for the performance of coastal wetlands. As mangroves are dispersed by floating seeds or seedlings (for the viviparous taxa), altered patterns of tidal inundation could affect mangrove recruitment. Since Avicennia marina propagules can establish successfully only within 4–5 days of dropping from the parent trees, colonisation of new areas may be limited by the ability of the seedlings to reach new areas.

Hobday et al. (2006c) concluded their analysis by proposing that climate change would have the following impacts on mangroves:• Possible increase in primary production due to increase in atmospheric CO2

• Reduction of area under mangroves due to sea-level rise• Changes in mangrove distributions, with likely southwards direction as air and ocean temperatures

increase• Shifts in the timing of flowering, due to altered temperature, rainfall and humidity• Increased damage to mangrove areas due to severe storms and flooding.

Sjerp (2007) undertook an analysis of the likely impacts of climate change on the Gippsland Lakes in eastern Victoria, and many of the conclusions are applicable to the wider case of coastal saltmarshes and mangroves along the Victorian coastline. Table 1.27 shows a summary of the conclusions.

Table1.27: Summary of impacts of climate change on the Gippsland Lakes, classified according to a) impacts related to altered weather patterns and b) impacts related to sea-level rise. Source: Sjerp (2007, Table 4).

Impact Likely change Certainty

Altered weather patterns

Average temperatures Increase in temperatures High

Dominance of south-westerly frontal synoptic patterns

Increased dominance Moderate

Precipitation Decreased annual rainfall Moderate–Likely

River discharge Decrease of up to 50% Moderate–Likely

Storms and heavy rainfall Increase in severity and frequency Moderate

Bushfires Increase in number of fire-danger index days Moderate

Change in river discharge related to fires in catchment

Immediate increase in runoff, followed by long-term decrease

High

Erosion of rivers, floodplains and lakes Increase High

Sedimentation and water-column turbidity

Increase in sedimentation and decrease in water clarity Moderate

Sea-level rise

Mean sea levels Rise High

Coastal erosion Increase Low–Moderate

Penetration of tidal prism into estuaries Increased penetration up rivers High

Salinity in main lakes Increase High


Most recently, the Sydney Olympic Park Authority undertook an education and training program on wetland adaptations to climate change (Sydney Olympic Park 2008). Cox (2008) argued that existing wetlands would shrink with reductions in freshwater runoff, and that an increase in the frequency of extreme events (with associated increased discharge and increased sediment load) would have impacts on sedimentation in open-water areas and in fringing beds of aquatic plants. Mazumder (2008) concluded that food-web structure in estuaries, saltmarshes and mangroves would be modified by climate change. Field (2008) argued that mangroves would be affected by changes in relative sea level, CO2 concentrations, temperature and precipitation, of which sea-level changes were most likely to have the greatest impact. Saintilan (2008) concluded that the response of coastal plant communities to rises in mean sea level was complex and dynamic, as sea level controlled rates of sedimentation and erosion, plant primary production and food-web structure.

Gilman et al. (2008) reviewed the global threat posed to mangroves by climate change. They concluded that mangrove ecosystems were seriously threatened by climate change, and the threat would be imposed mostly through rises in relative sea level. The mangrove ecosystems most at risk were those currently experiencing a net lowering of sediment elevation and where there was little or no room for landward migration. Mangroves associated with the Pacific Islands were highly at risk of inundation by rising relative sea levels.

The most recent analyses of likely climate change impacts on Australian mangroves and coastal saltmarsh is found in Steffen et al. (2009a,b,c). Their studies addressed biodiversity on an Australia-wide scale and thus little space could be allocated to mangroves and saltmarsh alone. It was concluded that rise in mean sea levels could result in landward migration of mangroves and salinisation of upstream wetland habitats, and that storm surges would result in increased erosion and sediment slumping and possibly also changes to the dynamics of wrack deposited on the shoreline. Human ‘adaptive’ responses, especially the construction of sea walls, could interfere with species migration in the face of rising sea levels, and altered patterns of river discharge would affect inputs of detritus and nutrients into estuaries.

summary of likely impacts of climate change on mangroves and coastal saltmarshes

Climate change encompasses future variation in a wide range of climate variables (Table 1.21). It will manifest as impacts on coastal wetlands via both direct effects on the biota, mostly mediated through the effect of higher temperatures and higher CO2 concentrations, as well as a wide suite of indirect effects. These indirect effects will be mediated through changes to ecological interactions among species and a range of secondary changes to the bio-physical environment, such as higher mean sea levels, increased incidence and severity of extreme events such as storm surges, altered salinity regimes, changes to freshwater runoff and more frequent and severe wildfires. Although all stages in estimating climate change impacts are associated with uncertainty (Figure 1.64), the following general conclusions seem to be warranted with respect to likely changes in the climate.

Temperature

An increase in air temperature of 0.6–1.2oC by 2030 and up to 3.8oC by 2070 is projected for southern Victoria (Table 1.21). Such an increase could have the following ecological effects:• Modify the phenology of all wetland biota, with particular impacts on flowering and germination of

plants and the breeding success of invertebrates, fish and birds


• Facilitate the invasion of saltmarshes by mangroves, which could be relieved of their existing winter low temperature and/or frost limitation in southern Victoria

• Disrupt life histories of stenothermal invertebrates and fish, with consequences for growth, mortality and secondary productivity

• Increase rates of evaporation from saltmarsh pools, leading to more saline conditions and perhaps the creation of hypersaline areas within saltmarshes

• Increase rates of primary production by saltmarshes and mangroves, unless other factors (e.g. hypersalinity, lowered nutrient availability, etc.) intervene

• Facilitate further invasion by weeds that are currently temperature-limited (e.g. *Baccharis halimifolia or *Paspalum vaginatum) or will be advantaged by increased inundation (e.g. *Spartina anglica).

Carbondioxideconcentration

Increased concentrations of atmospheric CO2 will affect competitive interactions between C3 and C4 plants. In the case of C3 plants, rates of primary production may be increased, but it is not possible to make definitive conclusions because a wide range of other important environmental variables (e.g. soil water content) co-vary with changes in CO2 concentration. Higher atmospheric CO2 concentrations have already affected the pH and alkalinity of ocean water, and may affect standing water in saltmarshes, with implications for shell-building by marine invertebrates and productivity by benthic algae, especially calcified species.

Rainfall

Although projections for rainfall are highly variable, it is expected that rainfall will decrease in south-eastern Australia with ongoing climate change. Moreover, changes in rainfall are not expected to be spread evenly across seasons, and winter and spring rainfall will decrease the most. Decreased rainfall will result in a disproportionate decrease in runoff; reductions in runoff of perhaps greater than 50% for streams in parts of Victoria are projected for 2070. Reductions in runoff of 20–30% are possible by 2030. Decreases in rainfall and freshwater runoff could have the following effects:• Impacts on all aspects of the wetting and drying cycles of coastal saltmarsh, with longer periods between

freshwater inundation – a reduction in freshwater inundation over winter or spring could see impacts on the germination of seed of saltmarsh plants, as well as on the successful recruitment and establishment of young plants

• Exacerbate the development of hypersaline conditions in saltmarshes, with possible conversion of saltmarsh to hypersaline flats

• Replacement of areas of coastal wetlands currently vegetated with glycophytic (e.g. Phragmites australis) or brackish-water plants (e.g. Melaleuca ericifoia) by saltmarshes as a result of increased salinity

• Affect primary productivity and distributions of mangrove communities, albeit in unpredictable directions

• Modify patterns and rates of coastal sedimentation and erosion, with effects on the area suitable for mangrove or saltmarsh colonisation.

Sealevel

Rises in sea levels will be a function of ongoing and incremental increases in mean sea levels, combined with short-term and unpredictable increases due to storm surges and other extreme events. Modelling of sea levels


in the Western Port region has indicated a rise in mean sea level of up to ~0.5 m by 2070, but additional inundation by storm surges of the order of 2 m or more. Increased sea levels will have severe ecosystem-wide impacts on mangroves and coastal saltmarshes, including:• Partial or complete inundation (and possible loss) of existing saltmarsh and mangrove communities• Replacement of areas currently vegetated with saltmarshes by more inundation-tolerant mangroves• Replacement of areas currently vegetated with glycophytic (e.g. Phragmites australis) or brackish-water

plants (e.g. Melaleuca ericifolia) by saltmarshes• Changes in the floristic distribution of saltmarsh taxa along the elevational gradient from the sea • Affect the distribution of plant propagules, via altered currents and tidal patterns• Modify patterns and rates of coastal sedimentation and erosion, with effects on the area suitable for

mangrove or saltmarsh colonisation.

Frequencyandseverityofextremeevents

The incidence ofextreme events is projected to increase with future climate change. The number of days over 35oC, for example, could increase by 10–13 days per year by 2030, and by up to 26 days per year in 2070. An increase in the frequency and severity of floods is predicted for rivers in south-eastern Australia, as well as increases in storms. The increased incidence (and possibly increased severity) of severe weather events could have the following ecological impacts:• Increased severity and incidence of storm surges, which could breach coastal sand dunes, lead to increased

rates of erosion and subject landward areas to additional inundation by seawater• Increased incidence and severity of riverine floods, with impacts on the inundation by freshwaters of

coastal wetlands and, in some locations, ‘back-up’ of marine or brackish water into other wetlands• Increased incidence and severity of droughts, with potential impacts on freshwater inundation,

evaporation and creation of hypersaline conditions in saltmarshes• Increased incidence of flashy discharge, with implications for sudden and rapid inundation of coastal

wetlands, increased erosion in the catchment and sedimentation, increased nutrient loads, and increased scouring of channels and wetland basins.

1.14 Ecologicalcondition

Chapters 1.11–1.13 outlined the wide range of threats faced by mangroves and coastal saltmarsh in Victoria. The following section investigates the methods that can be used to quantitatively assess the ecological condition of these types of estuarine wetlands. Chapter 5.4 outlines two approaches, consistent with existing state policy, developed during the present project to assess ecological condition in Victorian coastal saltmarsh.

what is ecological condition?

The condition of wetlands is notoriously difficult to measure. As noted in Chapter 5.4, the term ‘ecological condition’ is used in a wide range of senses, usually without a single universally understood meaning (Andreasen et al. 2001; Scholes & Biggs 2005; Gibbons & Freudenberger 2006; Parkes & Lyon 2006). There are at least seven, sometimes interrelated, reasons for the confusion. First, as is often the case with natural ecosystems, it is not clear what constitutes a wetland in ‘good’ condition, let alone which factors can be used to quantify a decline in condition (Fairweather 1999a,b).


Second, almost all methods for assessing the condition of floodplains or wetlands make use of some sort of benchmark, often derived from a reference site (Brinson & Rheinhardt 1996). It is often difficult or impossible to find coastal wetlands in south-eastern Australia that have not been subject to human interference and can act as suitable reference sites (see Sinclair & Sutter 2008).

Third, the attributes that can be used to independently ascribe ‘pristineness’, and thus reference condition, are not always clear or uncontroversial. In many cases, even expert ecologists differ markedly in their assessments of ‘healthy’ or ‘pristine’ sites (see Wood & Lavery 2000 for seagrasses in Western Australia).

Fourth, a wide range of variables are available to be monitored but there is sometimes little agreement among experts as to the ‘best’ (however defined) variables to measure. Some assessment protocols, for example, quantify hydrological components, others water quality, or various aspects of the biota, which in turn can include measurement of aquatic macroinvertebrates (frequently used to monitor ecological condition of freshwater streams), fish, reptiles, waterbirds, frogs or plants.

Fifth, almost all protocols employ some sort of scoring and/or weighting of different variables, and results are heavily dependent upon the weighting systems devised. Weighting systems are often used when benchmarks from pristine sites are not available; the problem then becomes one of choosing the scoring or weighting to be given to different indicators or variables. For example, are bird abundances ‘worth’ twice the floristic diversity of the vegetation? What cut-offs should be used for the individual variables? Is a 2 ha saltmarsh ‘worth’ a score of 1 and a 100 ha saltmarsh a score of 3? In most cases, there is no independent basis for choosing the bin sizes of the different categories.

Sixth, because many wetlands are naturally subject to large changes over time, it is often difficult to distinguish catastrophic changes that signify real decline from large-scale changes that are within normal temporal variation and would be followed by recovery. This is particularly the case where vegetation cover and plant health are used as indicators.

Finally, it is difficult to separate the notion of ecological condition from public perceptions of a preferred state for wetlands. It means that the concept of ecosystem health or ecological condition is especially vulnerable to political manipulation. Rapport et al. (1998), for example, provided a North American example of the logging industry promoting un-logged, natural forests as ‘unhealthy’ because they were subject to fire and other natural disturbances. Similar prejudices may apply to many types of coastal wetland, which are variously described as ‘stagnant’. Chapter 3 of this report addresses, in detail, some of the widely-held community perceptions of mangroves and coastal saltmarsh, not all of which are heartening or complimentary.

Notwithstanding these difficulties, it is generally acknowledged (Cairns et al. 1993; Bunn 1995; Schofield & Davies 1996) that the most complete assessments of the condition of aquatic systems make use of three sets of attributes:• Analysis of structure• Quantification of ecological function• Analysis of ecosystem values or ecosystem services, usually from a perspective of human benefit or

utilisation.

The term ‘structure’ is used in two different ways by different sorts of biologists and it is essential to differentiate among them. Botanists commonly use the term to mean the architecture or physical diversity


of habitat, for example, the physical structure and types of habitats provided by a grass-dominated wetland are quite different from those of a shrub-dominated system. Under this meaning, the term is interpreted without any reference to the taxonomic composition of the wetland. In almost complete contrast, ecologists interested in rehabilitation and monitoring commonly use the term to refer to the taxonomic composition of an area, without any reference to the ecological interactions or functions of those species. Under this meaning, structural assessments are undertaken in terms of animal or plant abundances at a given taxonomic level, for example, in the compilation of species lists.

Ecological function refers to the fundamental ecological processes that operate in a wetland. Examples of important processes include rates of primary productivity, ratio of production by algae and by vascular plants, pathways for the decomposition of organic matter, food-web structure, rates of recruitment and patterns of ecological succession. Unlike structural analyses, functional analyses are often expressed in units of rates or fluxes, for example, the rate of primary production in g C m–2 day–1, fluxes of nitrogen through an ecosystem by N loading, N2 fixation and denitrification, typically in units such as gN m–2 day–1 or kgN ha–1 year–1.

Ecosystem values or services refer to the benefits that human populations receive from mangroves or coastal saltmarsh, a topic addressed in Chapters 1.8 and 1.9. Thus assessments of ecosystem values or services address the extent to which coastal wetlands provide the range of ecosystem services expected of them by the human population. Saltmarshes, for example, can provide opportunities for passive recreation (e.g. bird watching, painting and photography), active recreation (e.g. hunting and fishing), as well as flood mitigation and nutrient interception.

The inclusion of a ‘value’ component into ecological assessments tends to shift the analysis away from ecological condition towards a more nebulous assessment of wetland ‘health’ (e.g. see Myer 1997; Boulton 1999). The shift in emphasis has its own serious problems, not the least of which is the meaning of the word ‘health’ when applied to natural systems (Calow 1992, 2002; Suter 1993). Moreover, the inclusion of human needs renders assessments of ecological condition dependent on geographic position; the idea that an identical wetland would be considered in better or worse condition depending on whether it were near or far from a town does not sit well with current scientific views on ecological ‘condition’.

Once condition has been measured, defining what constitutes an unacceptable condition or, even more problematically, detecting an unacceptable change in condition in the face of considerable spatial and temporal variability, presents a difficulty for all condition-assessment protocols (Fairweather 1999a). The problem is particularly critical for wetlands. It has been confronted in recent reports on the Ecological Character Description of Ramsar-listed wetlands, which must set limits of acceptable change for important wetland services and benefits (Department of the Environment and Water Resources 2007). The success of this approach is yet to be determined.

protocols for assessing wetland condition in south-eastern Australia

There is a plethora of protocols which purport to assess the condition of wetlands (e.g. see review by Department of Sustainability and Environment 2007a), including a number of procedures developed specifically for south-eastern Australia. Almost all emphasise structural elements (i.e. taxonomic issues), and only a few (e.g. Spencer et al. 1998) make some assessment of ecological function or process as well.


IndexofWetlandCondition

An Index of Wetland Condition (IWC) was developed by the Department of Sustainability and Environment in 2005 (Department of Sustainability and Environment 2005b, 2006, 2007a, 2009b; Papas et al. 2009). It adopts a definition of ‘wetland condition’ based on the Ramsar Convention definition of ecological character. For the purposes of the IWC, wetland condition is defined as ‘the state of the biological, physical, and chemical components of the wetland ecosystem and their interaction’ (Department of Sustainability and Environment 2005b, page i). The method is designed to be applicable to all lentic Victorian wetlands, including billabongs, but not to those that have a marine hydrological influence. In other words, it is not suitable for coastal saltmarsh or mangroves.

The IWC assesses condition on the basis of six sub-indices, which in turn make use of various indicator variables:• Wetland catchment oadjacent land use obuffer width and continuity• Physical form oreduction in wetland area ochanges in bathymetry• Hydrology oseverity of change of the water regime • Water properties oactivities leading to nutrient enrichment oevidence of a change in salinity• Soils ophysical soil disturbance extent and severity• Biota ocritical life forms oweeds o indicators of altered processes ovegetation health and structure.

Each of the six sub-indicators is scored out of 20, and a standardised final condition score (out of 10) is calculated for a given wetland. Five condition categories, adopted from the Index of Stream Condition, are used for reporting the final score: very poor, poor, moderate, good and excellent. The sub-indices are weighted on the basis of outcomes of the an earlier testing program for the IWC (Papas et al. 2009). Some sub-indices are more intricate than others. For example, the Hydrology sub-index has three possible score outcomes, whereas the Biota sub-index has a more complex scoring system based on an EVC typology. The Biota sub-index is initially scored out of 100, then standardised to a score out of 20 to match the other sub-indices. The IWC is currently being used to assess wetland condition across the state.

HabitatHectares

Also relevant to the current project is the ‘Habitat Hectares’ approach (Parkes et al. 2003; Department of Sustainability and Environment 2004b). It is well established for use with terrestrial vegetation in Victoria and a limited number of wetland types. As outlined in Chapter 5.4, it does not explicity define ecological


condition, although a view of condition is implicit and it seems that ecological ‘condition’ is essentially synonymous with ecological ‘quality’.

The approach has two components: i) a site score; and ii) a landscape-context score. The site score relies on a comparison between the observed vegetation and a pre-1750 benchmark (benchmarks are available from www.dse.vic.gov.au). The landscape-context score assesses the size and connectivity of the vegetation, in relation to nearby patches of native vegetation. The combination of the site and landscape context scores provides a habitat score, which is taken to be a measure of condition. It can be weighted by the size of the patch in question to give a Habitat Hectares score, which may be used for many purposes (such as calculating offsets according to the Native Vegetation Management - A Framework for Action: see Department of Natural Resources and Environment 2002b, 2004b).

The development of a condition-assessment tool specifically tailored to Victorian saltmarsh (Chapter 5.4) addresses in detail the relevant aspects of the Habitat Hectares approach.

Estuarine-specificprotocols

Colman et al. (1991) proposed a set of indicators for the environmental monitoring of Victorian coastal and marine environments. Among the biotic indicators proposed were extent and area of seagrass beds; extent of intertidal algal beds; abundance, age structure and breeding success of shorebirds; number of introduced species of plants and animals; size, structure and abundance of intertidal animal populations; and catch-effort data on commercial fish. For marine wetlands in particular, they recommended the monitoring of area and period of inundation, salinity of inundating waters, and community structure and condition of wetland vegetation. The latter was to be undertaken in terms of local on-ground surveys ‘…indicating health and composition of plant communities’ (Colman et al. 1991, Table S1).

Arundel et al. (2009) recently produced a set of recommended themes and measures for a forthcoming Victorian Index of Estuary Condition. The approach is broadly consistent with that used for the past ~10 years by the Victorian Index of Stream Condition, and is based on measurement of six sub-indices: i) physical form; ii) hydrology; iii) water quality; iv) sediment; v) flora; and vi) fauna. In total, 18 measures were proposed: four for physical form; three for hydrology; two for water quality; three for sediment; four for flora; and two for fauna. For the flora sub-index, the proposed indicators were:• aquatic macrophytes: percentage change from historical, as well as percentage cover or, for algae, number

of blooms• fringing macrophytes: extent and condition• microphytobenthos: pigments• phytoplankton: chlorophyll a.

We are aware of only four protocols that have been developed specifically for, or used in, coastal saltmarsh or mangroves of south-eastern Australia.

First, Kessler (2006) developed an assessment protocol for New South Wales coastal saltmarsh that was rapid, quantitative and non-destructive. It made use of four sets of indicators, each with a quantitative measure and measured according to a scoring scheme. The various sets of indicators received different weightings. The system is shown in Table 1.28.


Table1.28: Final recommendations for selection of rapid indicators of saltmarsh condition. Source: Kessler (2006, Table 6).

Indicator Measure Scoring

Physical site characteristics (6 out of 26)

Site area Area of site in ha 0 = < 1 ha1 = 1–2 ha2 = > 2 ha

Tidal flushing Area of saltmarsh undergoing most regular tidal flushing

0–2 on the basis of position of debris line

Evidence of edge erosion % boundary affected by erosion

0 = > 20%1 = 5–20%2 = < 5%

Anthropogenic impacts (6 out of 26)

Limits to site expansion % boundary with barriers 0 = > 20%1 = 5–20%2 = < 5%

Anthropogenic structures % site covered with structures

0 = > 20%1 = 5–20%2 = < 5%

Presence of rubbish % site covered 0 = > 20%1 = 5–20%2 = < 5%

Fauna (2 out of 26)

Crab populations % area covered by burrows 0 = < 20%1 = 20–60%2 = > 60%

Vegetation (12 out of 26)

Community distribution Number of communities 0 = < 31 = 3–42 = > 4

Species composition Number of vascular plant species present

0 = < 31 = 3-72 = > 7

Threatened species Presence 0 = No2= Yes

Mangrove intrusion % site covered by mangroves

0 = > 10%1 = 5–10%2 = < 5%

Introduced species % site affected 0 = > 10%1 = 5–10%2 = < 5%

Second, Pacific Wetlands (2008) used a modified version of the Kessler (2006) protocol to assess condition of coastal saltmarsh at the Sydney Olympic Park site. Indicators were scored as 0, 1 or 2, and included saltmarsh area, vegetation condition, Wilsonia condition, area of introduced species, abundance of mangroves seedlings or adults, abundance of crab holes, presence of mosquito larvae (scored only as 0 or 2), erosion or sedimentation, presence of litter and evidence of trampling.


Third, coastal saltmarsh in parts of western Port Phillip Bay was monitored, as part of the channel-deepening project, by Horlock & Houtgraaf (2008). Condition was assessed by two criteria: i) assessment of canopy ‘health’ of shrub life-form components, using a categorised grading based on percentage cover and the ratio of living:dead foliage along transects; and ii) assessment of cover ‘health’ of herb life-form components, using a categorised grading of percentage cover and ratio of living:dead foliage within quadrats. The canopy ‘health’ index was based on protocols originally devised for terrestrial vegetation, including studies at Chowilla in semi-arid South Australia near the River Murray.

Fourth, Sinclair & Sutter (2008) attempted to assess vegetation condition in a large number of estuarine wetlands in western Victoria with the Habitat Hectares approach, and found that in some cases the available ‘Habitat Hectares’ benchmarks were inappropriate.

conclusions

The assessment of ecological condition of mangroves and coastal saltmarsh in south-eastern Australia is in its infancy. There are formidable, perhaps insurmountable, problems in devising protocols that can assess quantitatively ecological condition in coastal wetlands. The four protocols that have been devised or trialled with coastal saltmarsh all make use of scoring systems (often arbitrary) and differential weighting across indicators (also often arbitrary). Indeed, one method that was used in Victorian saltmarshes was based on protocols developed and trialled with semi-arid terrestrial vegetation, not with wetand or coastal vegetation.

The biggest problem with developing a method to assess ecological condition in coastal wetlands, however, is that the concept itself of ‘ecological condition’ has no agreed definition (see Andreasen et al. 2001; Scholes & Biggs 2005). Moreover, the commonly used ‘ecological health’ approach is a minefield of confused and intermingled concepts and many ecologists harbour grave concerns about the term being used in scientific analyses. Some important unresolved questions when attempting to quantify ecological condition in coastal wetlands are:• Are ‘natural’ states the most desirable? If so, can they even be defined?• Do (natural) cyclic changes in wetlands represent fluctuations in condition?• Does small size equate to low condition? What if the wetland has only ever been small? Because of site-

specific geomorphological limitations, not all wetlands have the capacity to be large. The same restriction applies to other concepts commonly used to indicate condition, such species diversity: what if the wetland only ever supported a low number of species? Does that make it in poorer condition than one that supports many species?

• Should one system always be comparable with others? Should condition assessments, indicators or variables avoid comparing ‘apples with oranges’?

One of the tasks of this project is to develop a method for assessing the ecological condition of coastal saltmarshes. That task is undertaken in Chapter 5 (see also Appendices I and J).


1.15 Priorclassification,mappingandinventorystudies

In order to effectively manage wetlands, it is necessary first to describe the different types of wetlands that exist in a given area and then to quantify their extent and distribution. The former activity is termed wetland classification and the latter wetland inventory. Mapping is commonly the intermediate step between these two activities: an appropriate classification system is first devised, wetlands are mapped according to that system and the extent of different wetland types quantified in the final, inventory phase.

Classification, therefore, is a necessary precursor to mapping which, in turn, is a necessary precursor to the preparation of an inventory. The different activities are intimately interlinked, but not all studies have addressed all three aspects. Some studies have been content to prepare a classification system for wetlands in general or saltmarshes in particular; others have developed a classification system, or utilised an existing system, to map wetlands or saltmarshes. Only a few studies have used classification and mapping to prepare an inventory. Adam et al. (1988), for example, developed a classification system for saltmarshes in New South Wales but there was no intent to provide a map or inventory (although it could be subsequently used for these purposes).

why classify and map coastal vegetation?

There are many reasons why coastal vegetation needs to be classified and mapped. Some include: • Describe areas of coastal vegetation to help managers apply appropriate management strategies and tools• Reveal underlying pattern and process in the marsh and lead to improved understanding of how coastal

plant communities are structured and evolve• Help managers communicate unambiguously about vegetation areas• Help policymakers know what they are dealing with• Fulfil legislative directives for natural resource management• Allow inventory statistics to keep track of change, especially anthropogenic changes• Underpin scientific sampling design and description.

Note that classification is a human need and a human activity. Thus there is rarely a single ‘correct’ answer to a classification or mapping problem, and sometimes utility and breadth of application can be as important as scientific rigour.

the problem of saltmarsh classification

Classification systems tailored specifically for saltmarshes and other coastal wetlands are necessarily more restricted than classification systems developed for wetlands as a broad group and in which saltmarshes or mangroves form merely one of a number of categories. The development of suitable classification systems is not easy, and Adam (1990, page 63) argued that:

the nature of saltmarsh vegetation poses particular problems in designing an appropriate classificatory framework. The wide ecological amplitude of certain saltmarsh species makes it possible to trace floristic links between most saltmarsh communities.

Under the Water Framework Directive, European countries are obliged to classify and assess the condition of aquatic systems, and marine angiosperms (i.e. seagrass beds and saltmarsh) have been identified as one of the


biological elements to be used in the classification (Best et al. 2007). The current approach taken in the United Kingdom, for example, to saltmarsh classification is based on habitat extent, zonation and floristic diversity. It has been proposed that a typology based on such structural analysis is inadequate, and the classification system under development for the Directive may include attributes such as the function of coastal saltmarsh as fish nurseries (Best et al. 2007). Such a shift from classifications based on taxonomic components to ones based on ecological processes mirrors similar developments in the assessment of the ecological condition of aquatic systems, as noted previously.

phytosociological approaches to classification

Vegetation analysis in continental Europe has traditionally centred on phytosociology, and the approach still finds widespread application in many parts of the world (Best 1988). As shown in the examples below, many recent analyses of saltmarsh vegetation have been based on phytosociology.

The phytosociological method is based on a hierarchical approach to vegetation classification that parallels the hierarchical Linnaean taxonomic system to classify plants, animals and bacteria. Three distinct schools of phytosociological analysis developed in Europe in the early 20th century: i) the Raunkiaerian (mostly in Denmark); ii) the Uppsala (mostly in Sweden); iii) and the Zürich-Montpellier school of Braun-Blanquet. The last has had by far the most impact (Shimwell 1971; Randall 1978; Westhoff & van der Maarel 1978).

The basic unit of phytosociology is the association. Associations are expressed as a two-way table which tabulates species composition against quadrats (also known as relevés or sample plots); in other words, an association is an abstraction which illustrates not only the consistent occurrence of those species which define the association but also any variation in the occurrence of other taxa (Adam 1990). Associations are usually named after one or two species that are present (but not necessarily assumed to be dominant) within the identified plant community. These character species are defined as those that are restricted to a given community and which clearly identify it and the environmental conditions in which the association is found. Communities are subsequently ordered into a hierarchical classification of units (called syntaxa) with each level distinguished by a different suffix: sub-association (-etosum); association (-etum); alliance (-ion); order (-etalia) and class (-etea). The process is outlined in Collinson (1988).

The classification of coastal saltmarsh has often been based on the phytosociological tradition (e.g. see Beeftink 1977). The most detailed classifications of regional Australian saltmarsh vegetation also have been based on this approach to vegetation analysis (e.g. Bridgewater 1975; Adam et al. 1988), and the analysis of 3,000+ km coastline of southern Australia for patterns in saltmarsh vegetation undertaken by Bridgewater & Cresswell (2003) similarly used a phytosociological approach. Phytosociological approaches continue to be used widely to classify and/or map saltmarsh vegetation: recent examples of their application include mapping of coastal halophytic vegetation in Greece (Korakis & Gerasimidis 2006); vegetation in abandoned salt-evaporation pans in France (Bouzille et al. 2001), vegetation of coastal Spanish marshes (Molina et al. 2003) and a study of saltmarsh vegetation along the Adriatic coast (Pandza et al. 2007). All these studies were of European saltmarshes, perhaps in keeping with the European origins of the Zürich-Montpellier school. The approach, however, has been applied also outside of Europe: Mucina et al. (2003) used phytosociological methods to study plant zonation in South African saltmarshes; Peinado et al. (2008) applied them to coastal vegetation along Baja California in Mexico; and Haacks & Thannheiser (2003) used the Braun-Blanquet


approach to describe saltmarsh vegetation in New Zealand. In the latter study, 27 associations were identified, of which three (Leptinelletum dioicae, Plangianthetum divaricati and Puccinellietum walkeri) were endemic to New Zealand.

Unlike in continental Europe, phytosociological approaches to vegetation classification found little favour in America or the British Isles, where pioneering researchers objected to its static interpretations (e.g. see Clements 1928; Tansley 1929). Kellman (1975, cited in Randall 1978, pages 44–45) describes the drawbacks of phytosociology as follows:

In essence, the system appears to possess most of the undesirable attributes of an organisation and few of the desirable features of a summary. It is highly selective of the data it treats, employs ill-specified strategies and has proven impossible to apply in areas of floristically complex vegetation. Above all, it relegates the user to a role which is little more than that of a descriptive technician. Its continued use in vegetation studies appears to reflect more the inertia of the system than its intrinsic value.

Many current-day botanists also make a number of objections to the phytosociological approach to vegetation classification. First, there is a large element of subjectivity in choosing the positions of the relevés and the higher categories are necessarily less well defined than the lower categories of the hierarchy (Dalby 1987; Adam 1990). Moreover, the hierarchical classification of associations into higher levels is necessarily one-dimensional, whereas the relationships between and among different associations is complex and likely to be multidimensional (Webb 1954). Adam (1990) noted that the difficulty in identifying phytosociological character species in saltmarshes is that many saltmarsh plants have very wide ecological tolerances, thus the number of communities that could be identified by discrete character species is rather limited. He proposed instead that saltmarsh communities would be better defined not by the behaviour of a limited number of individual character species but rather by the occurrence of a total assemblage of plant species, known as the character combination. The character combination was defined as ‘a group of taxa whose joint occurrence is exclusive to a particular vegetation type without any of the taxa necessarily being a character taxon’ (Adam 1990, page 63).

Second, the primacy of the taxon renders the approach subject to rapid obsolescence. The issue is particularly relevant to saltmarsh, where the taxonomy of many taxa requires substantial revision, and indeed to the Australian flora more generally, as it remains subject to taxonomic change and uncertainty. In other words, as the taxonomy dates so does the classification.

Third, not only do classifications become dated with taxonomic revision, but new survey data cannot be readily analysed with data from previous decades.

Fourth, phytosociological approaches ignore the issue of species turnover arising from geographic distance, and conflates it with environmental distance. As such, saltmarshes in Western Australia will have more in common floristically with other Western Australian estuarine vegetation (if they were also included in a hypothetical analysis) than they would with Victorian saltmarshes. This may be an undesirable outcome.

Finally, the approach takes no account of structural elements: a tree is ‘equivalent’ to a small ephemeral herb in the analysis, and this conflation is also often undesirable.


whole-of-nation classification for australian coastal saltmarsh

We are aware of only one whole-of-nation classification of Australian saltmarsh, that of Bridgewater & Cresswell (2003). They identified five main saltmarsh groups:• Tecticornia (Sclerostegia) arbuscula-Juncus kraussii group found along the coastline of southern Australia

and Tasmania• Tecticornia (Halosarcia) doleiformis-Tecticornia (Halosarcia) leptoclada group found in coastal and inland

Western Australia• Suaeda arbusculoides-Tecticornia (Halosarcia) indica ssp. julacea group found in central eastern and

northern Australian coastlines • Tecticornia (Sclerostegia) tenuis group found in central arid and semi-arid regions• Tecticornia (Halosarcia) pergranulata ssp. pergranulata group found in the Murray-Darling Basin.

This system is clearly inadequate for describing saltmarsh patterning within Victoria, given the floristic and structural complexity of Victorian systems outlined earlier in Chapter 1.5.

mapping and inventory protocols

Wilton & Saintilan (2000) developed a set of recommendations for mapping Australian mangroves and coastal saltmarsh. They distinguished between mapping undertaken in order to prepare resource inventories and mapping undertaken for the purposes of environmental monitoring. When maps were used to construct resource inventories, serious errors could occur if the mapping were undertaken at too coarse a scale (e.g. 1:100,000 or 1:250,000). In these cases, small or linear areas of a feature would not be well represented and large areas would not include small internal gaps or patches.

Wilton & Saintilan (2000) identified 45 coastal areas that had been mapped at scales ranging from 1:5,000 to 1:100,000, but with most studies having been undertaken at scales of 1:10,000 to 1:30,000 (Table 1.29). They proposed that mapping should be undertaken at a scale of 1:10,000 or finer for detecting habitat change and, ideally, a scale of 1:5,000 or finer to differentiate mangroves from saltmarshes in ecotonal systems. Of the 45 study areas, five were in Queensland, three in Western Australia, two in South Australia, and one in Tasmania. Only two Victorian studies were cited, of which only one (Vanderzee 1988, for Corner Inlet) provided maps. An important conclusion to be gained from the Wilton & Saintilan (2000) report is that almost all of the mapping of coastal intertidal and coastal wetland vegetation has been undertaken in New South Wales (with 32 of the 45 study areas).

Table1.29: Analysis of scales used in Australian saltmarsh and mangrove mapping. Source: Wilton & Saintilan (2000).

Scale range Number of studies

1:5,000 or finer 1

1:5,001 to 1:10,000 8

1:10,001 to 1:30,000 17

1:30,001 to 1:50,000 3

1:50,000 or coarser 3


The often-confronted difficulty of separating mangrove from saltmarsh habitats when the two interdigitate was noted earlier, in Chapter 1.2 of this report. Wilton & Saintilan (2000) proposed that such ecotonal areas be identified as mangrove habitat when the mangrove canopy had gaps of < 10 m, as mixed habitat where the gap between nearest neighbour mangrove canopies was 10–20 m, and as saltmarsh when the gap between individual mangroves was more than 20 m.

With regard to the potentially troublesome saltmarsh-terrestrial boundary, Wilton & Saintilan (2000) recommended that the vegetation types characterising the landward habitats (Casuarina glauca in New South Wales) should be mapped at the same time as other intertidal wetlands were mapped, in order to provide supplementary information on future mangrove-saltmarsh dynamics.

state-specific studies: victoria

Bridgewater (1975) provided the most detailed classification of plant communities in a Victorian coastal saltmarsh. Taking a strongly phytosociological approach, he identified ten vegetation complexes in the peripheral vegetation of Western Port. A suite of plant communities was identified within each vegetation complex: within the Sarcocornia (Salicornia) complex, for example, two communities were identified, a Triglochin striata community and a Sarcocornia quinqueflora community. The Avicennia complex and the *Spartina complex were each represented by single plant communities, typified by Avicennia marina and *Spartina x townsendii, respectively. While this system may apply fairly well to Western Port, it is inadequate for statewide applications (see Chapter 1.5).

The first statewide mapping of coastal saltmarsh was undertaken by Carr (1979), as part of a study into the habitat requirements of Orange-bellied Parrot. Three types of saltmarsh were mapped across the state: Sarcocornia (Salicornia) quinqueflora with no Arthrocnemum (now Tecticornia) arbuscula present; Tecticornia arbuscula plus Sarcocornia quinqueflora; and Tecticornia halocnemoides + Sarcocornia quinqueflora and Tecticornia arbuscula. Sites that were suspected of containing saltmarsh, but for which access was unavailable, were also mapped. The scale of mapping was 1:100,000. Figure 1.69 shows an example of the maps generated in that project.

Subsequently, a number of mapping and inventory studies have been undertaken at finer (i.e. regional or local) scales in Victoria. Carr (1982), for example, described and mapped in detail the vegetation of the Murtcaim saltmarsh (The Spit Nature Conservation Reserve) near Point Wilson, at a scale of 1:4,000. A total of 12 vegetation units were mapped, based on extensive quadrat sampling, air-photo interpretation and ground truthing. The vegetation units were recognised according to the structural dominants/co-dominants of the communities. Carr (1982) recognised a vegetation unit in the upper saltmarsh ecotype of Tecticornia (Halosarcia) pergranulata. Following the taxonomic revision of the Salicornieae by Wilson (1980), Yugovic (1984) recognised that this ‘dwarf form’ of Tecticornia pergranulata was in fact Tecticornia halocnemoides. The mapping of the Murtcaim saltmarsh by Carr (1982) is probably the most detailed mapping undertaken to date of coastal saltmarsh in Victoria; of particular interest is that it shows the complicated and small-scale patterns that can exist within saltmarsh vegetation. These patterns can sometimes be interpreted as a function of tidal regimes in combination with variations in macrotopographic and microtopographic relief (see also the discussion in Chapter 1.3 on inundation regimes of coastal wetlands). The soil and hydrological environments in the study area were described by Adams (1982) and Kinhill Pty Ltd (1982), respectively.


Figure1.69: Distribution of saltmarsh of Western Port in the report by Carr (1979). Saltmarsh is shown in green.

Yugovic (1985) mapped the vegetation of Lake Connewarre, the estuarine area near the mouth of the Barwon River that forms the basis of the management template developed in Chapter 7 of this report. The vegetation map units used by Yugovic (1985) were:• Wilsonia humilis herbland• Avicennia marina shrubland• Juncus kraussii rushland• Samphire herbland and samphire shrubland• Gahnia filum sedgeland• Poa poiformis grassland• Distichlis distichophylla grassland• Muehlenbeckia cunninghamii (now M. florulenta) shrubland• Eleocharis acuta sedgeland• Schoenoplectus pungens reedswamp• Phragmites australis reedswamp• Melaleuca lanceolata scrub• Exotic vegetation and bare areas

A number of fine-scale surveys have been undertaken of saltmarshes in various parts of coastal Victoria. Yugovic & Mitchell (2006), for example, undertook an ecological review of the Koo Wee Rup swamp, located on the northern shore of Western Port. Yugovic (2008) later reported a flora survey of The Inlets, a part of Koo Wee Rup swamp. Figure 1.70, a map taken from Yugovic (2008), shows the detail achievable in studies of relatively small areas of coastal saltmarsh, and such levels of detail are often required for the preparation of site management plans (see Chapter 7).


Figure1.70: Detailed map of vegetation, The Inlets, Koo Wee Rup. Source: Yugovic (2008).

BBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBB

BBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBB

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

CL

EA

R E

D

E22665

E22666

C L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E DC L E A R E D

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Moo

dys

Inle

t

Cardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek DrainCardinia Creek Drain

DDDDDDDDDDDDDDDDDDDDDDDDDDDDDeep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

Deep Creek Drain

LegendEcological vegetation class

Mangrove Shrubland

Saline Aquatic Meadow

Coastal Saltmarsh

Estuarine Wetland

Estuarine Reedbed

Estuarine Scrub

Estuarine Flats Grassland

Brackish Sedgeland

Brackish Grassland

Plains Grassland

Tall Marsh

Swamp Scrub

Adventive vegetation / bare area

Figure 2: Vegetation, The Inlets Waterway Reserve, Koo WeBiosis Research Pty. Ltd.38 Bertie Street(PO Box 489)Port MelbourneVICTORIA 3207

DATE: 6 October 2008

Checked by: JY

Location: MRG 6700s\6781\Mapping\Final Figures\6781 Figure 2.wor

File number: 6781Drawn by: RMF

Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.Acknowledgement: VicRoads, Melbourne Water.

ee Rup.

Scale: 50 100 150 200 2500

metres

BBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBBB Quadrat

Study area


Most recently, some investigations have been undertaken on extent and condition of Victorian coastal saltmarsh as part of the Channel Deepening project (Horlock & Houtgraaf 2008). Sinclair & Sutter (2008) mapped estuarine wetlands of the western coast of Victoria, including saltmarsh in the Glenelg River, Fawthrop Lagoon, Belfast Lough and Lake Yambuk. Mapping was undertaken using EVCs as the map units; hence there was no differentiation within the Saltmarsh category.

Some student theses have mapped Victorian mangroves and coastal saltmarsh or attempted some aspect of an inventory. Calderwood (1998), for example, mapped changes in extent of Avicennia marina in Western Port from 1974 to 1994. It was difficult to draw firm conclusions about changes over the two-decade period, because maps for different parts of the coast were not made spatially congruent across the dates. Ghent (2004) compared past and present distributions of coastal saltmarsh in Port Phillip Bay, and found that ~65% of pre-European saltmarsh had been lost, mostly before 1978. A non-metric multidimensional scaling ordination was used to differentiate floristic categories of saltmarsh in Port Phillip Bay and Western Port, but with little success. Ross (2000) summarised the historical mapping of shorelines, mangroves and saltmarsh in Western Port, commencing with the 1842 surveys of George Smythe.

state-specific studies: non-victorian

Queensland

In 2003 the federal and Queensland governments established the Queensland Wetlands Program to develop a statewide wetlands inventory and to map the distribution of wetlands at a scale that could guide future management decisions. The Program makes use of existing (primarily remotely sensed) data, without the reliance on detailed field assessments. A rapid wetland classification and mapping approach was employed, based on the identification of Regional Ecosystems defined by the Queensland Herbarium (Queensland EPA 2005, 2009c). The scale of mapping was 1:100,000 for the entire state, except for specific areas such as the Great Barrier Reef, coastal areas, and the densely populated south-east. For these specific areas, mapping was undertaken at 1:50,000 or, in a few cases 1:25,000 (Queensland EPA 2005, 2008a,b, 2009c). The program was run through the Environmental Protection Agency and the Queensland Parks and Wildlife Service. Maps for the entire state have recently been made available (Queensland EPA 2009c).

An hierarchical classification system was developed to map wetlands in the Queensland program, based on the approach developed by Cowardin et al. (1979) for the USA but modified to provide for the rapid classification process (Queensland EPA 2005; Queensland EPA 2009c). Wetlands were classified into one of five ecological systems (after Cowardin et al. 1979): marine, estuarine, riverine, lacustrine or palustrine. Classification was then modified according to salinity (< 0.5 g L–1 = fresh; 0.5–30 g L–1 =brackish; > 30 g L–1 = salt), and dominant vegetation. A full description of the approach is available in Queensland EPA (2005, 2009c).

More specific mapping has been undertaken for coastal wetlands of south-east Queensland (Queensland EPA 2008a) and the Great Barrier Reef (Goudkamp & Chin 2006; Queensland EPA 2008b). For the specific area between Maroochy Shire and the Queensland–New South Wales border, coastal wetlands below the 2.5 m contour (i.e. those assumed to be subject to tidal influence) were mapped at 1:25,000 by the Queensland Herbarium (Queensland EPA 2008a). Over 37,000 ha of vegetation were mapped, using the following map units:• Mangrove• Samphire


• Claypan• Sporobolus virginicus grassland associations• Casuarina glauca associations• Sedgelands• Melaleuca quinquenervia associations• Heathlands.

Coastal saltmarsh was not differentiated into classes, although a distinction was made between samphire and other types of coastal wetlands.

In the case of mapping of coastal wetlands associated with the Great Barrier Reef Marine Park, Goudkamp & Chin (2006) reported that there were 1,660 km2 of coastal saltmarsh in and adjacent to the World Heritage Area, which represented over 40% of the combined area of mangrove and saltmarsh along the entire Great Barrier Reef coast. It is not the intention of this review to identify all examples of mangrove and coastal saltmarsh mapping in Australia, but the mapping undertaken of the Shoalwater Bay military training area (O’Neill 2009) does deserve mention for its completeness and detail.

NewSouthWales

A number of wetland-specific classification schemes and associated mapping-inventory studies have been undertaken in New South Wales. In one of the first, Goodrick (1970) devised a classification system that was later used widely to prepare an inventory of coastal wetlands in New South Wales. It identified 14 different wetland types:• Coastal bog• Open fresh waters• Coastal Lepironia swamp• Tea-tree swamp• Fresh meadow• Seasonal fresh swamp• Semi-permanent fresh swamp• She-oak swamp• Salt meadow• Reed swamp• Salt flat• Mangrove swamp• Shallow estuarine waters• Shallow saline lagoons.

The Goodrick system is non-hierarchical and different criteria, which conflate vegetation, geomorphology and hydrology, are used to define the different wetland types. Adam et al. (1985) proposed that the Salt meadow and Salt flat categories in the Goodrick system could be merged into a single type: Saltmarsh.

West et al. (1985) undertook a detailed mapping and inventory study of mangrove, saltmarsh and seagrass vegetation in 133 estuaries and embayments along the New South Wales coast. Aerial photographs, of various scales and dating from the 1970s and 1980s, were used to map the distributions of the three vegetation types. Field surveys, undertaken in 1981–1984, were then used to check the aerial photography interpretations.


This information was transferred to base topographic maps at a scale of 1:25,000 and the area of different vegetation types calculated on the basis of dot grids.

Adam et al. (1985) used a simple classification to identify wetland areas along the entire coastline of New South Wales. Seven vegetation types were identified as ‘coastal wetland’: • Mangrove• Saltmarsh• Melaleuca forest• Casuarina forest• Sedgeland• Brackish and freshwater swamp• Wet meadow.

Vegetation types not mapped by Adam et al. (1985) as ‘coastal wetland’ included terrestrial Melaleuca forest, terrestrial Casuarina forest, Melaleuca-Eucalyptus forest, littoral rainforest, dune thicket, wet heath, and disturbed or modified wetlands. Colour aerial photographs (1:25,000) were used to identify areas of coastal wetland; maps were drawn at this scale also (since it was the scale used for topographic mapping) then reduced to a scale of 1:100,000 for reporting. Different types of estuarine wetland were not mapped individually, nor was there a differentiation among different types of coastal saltmarsh.

Subsequent to that mapping study, Adam et al. (1988) developed a floristically based classification which followed classical phytosociological principles to devise a hierarchy of saltmarsh communities in New South Wales. Relevés (2 m x 2 m quadrats) were collected for sites along the New South Wales coast from Jervis Bay in the south to Port Macquarie in the north, and supplemented with published species lists for saltmarshes around the Sydney region. Saltmarsh was defined as ‘…intertidal and dominated by herbaceous species or low shrubs’ (Adam et al. 1988, page 39). In addition to two mangrove communities (characterised by Avicennia marina and Aegiceras corniculatum), 25 plant communities (some consisting of exotic plant species as dominants) were identified under Sarcocornietum quinqueflorae:• Sarcocornietosum• Triglochinetosum• Sporoboletosum• Wilsonia backhousei community• Samolus repens community• Sporoboletum virginici• Selliera radicans community• Juncus kraussii dominated community• Juncus kraussii – Sporobolus virginicus community• Juncus kraussii – mixed marsh community• Juncus kraussii – Suaeda australis community• Gahnio – Juncetum kraussii• Juncus kraussii – brackish pastures• Baumea juncea communities• Paspalum vaginatum community• Hydrocotyle bonariensis community


• Cotula coronopifolia community• Mimulus repens community• Cyperus laevigatus community• Reedswamps• Zoysia macrantha grassland communities• Plantago coronopus fringe community• Suaeda australis community.

Three other assemblages were recognised but not placed into the saltmarsh classification: Cynodon dactylon grasslands; Tecticornia arbuscula; and cliff crevice communities. The study by Adam et al. (1988) was primarily an exercise in saltmarsh classification, and there was no intent to apply it to map areas of saltmarsh or prepare an inventory. Of course, it could be used to those ends later if desired.

The mapping by Adam et al. (1985) was updated in the Comprehensive Coastal Assessment: Estuarine Resources of NSW, a program which commenced in 2005. The intention was to create digitised maps of estuarine aquatic habitats, primarily seagrass, saltmarsh and mangrove environments, from Port Stephens north to the Queensland border and from Lake Illawarra south to the Victorian border (West et al. 2006). Estuaries in the most politically sensitive central coast region (i.e. from about Wollongong to Newcastle) were omitted from the study. The results are available in West et al. (2006) and Williams et al. (2006). The typology was based on aquatic habitat types (seagrass, mangrove and saltmarsh), and estuaries were then categorised on the basis of whether there had been gains, losses or no change over the areas reported in the earlier study by West et al. (1985). Spatial layers for aquaculture leases, recreational fishing havens, aquatic reserves, Marine Parks, SEPP 14 wetlands and National Parks were also incorporated, followed by data on fish species diversity.

The Comprehensive Coastal Assessment program was given the specific task of providing and analysing information to support decision-making and planning on the New South Wales coast, and was underpinned by the planning objectives derived from the 1997 New South Wales Coastal Policy. The publication of the classification and inventory component by Williams et al. (2006) was followed by the production, in 2007, of a double-DVD toolkit containing information on the various studies, including those relating to estuaries with mangroves and saltmarsh.

As with Victoria, parts of the New South Wales coast have been subject also to more fine-scale mapping and/or inventory studies. Clarke (2003), for example, reported on a long-term (1990–2003) analysis of mangroves and saltmarsh in Jervis Bay, using a mixture of photoplots and on-ground qualitative observations. An earlier report (Clarke 1993) provided a floristically based classification for mangroves and saltmarsh for that section of the New South Wales south coast. Kelleway et al. (2007) mapped saltmarshes of the Parramatta River and Sydney Harbour, using aerial photographs and comprehensive ground truthing. Over 750 patches of saltmarsh were identified, with a combined area of 37 ha. About 50% (18 ha) of saltmarsh occurred as an understorey of other vegetation types and was not readily discernable from aerial photographs. As well as extent and location, the Kelleway et al. (2007) study addressed the topic of saltmarsh condition; the majority of saltmarsh patches were in poor condition. Other than those aspects of the study that addressed the presence of rare or endangered species, there was no attempt to differentiate among floristically different types of saltmarsh. Saltmarshes of the Sydney Olympic Park were mapped by Pacific Wetlands (2008), with an emphasis on management implications and the presence of Wilsonia backhousei, a vulnerable species in New South Wales.


Tasmania

As part of a statewide assessment of the vegetation of Tasmania, Kirkpatrick & Harris (1999) identified 16 structural-dominance communities in Tasmanian saltmarshes:• Communities dominated by succulent herbs 1. Tecticornia (Sclerostegia) arbuscula open-heath (also closed-heath low shrubland) 2. Suaeda australis open to closed-heath 3. Sarcocornia quinqueflora low open-heath 4. Sarcocornia blackiana low open-heath 5. Hemichroa pentandra low open- to closed-heath 6. Disphyma crassifolium low open-heath• Communities dominated by grasses 7. Stipa stipoides tussock grassland to closed-tussock grassland 8. Distichlis distichophylla closed-grassland 9. Puccinellia stricta open-grassland 10. *Spartina anglica grassland to closed-grassland 11. Deschampsia caespitosa tussock grassland • Communities dominated by sedges and rushes 12. Gahnia filum - Gahnia trifida tussock sedgeland to closed-tussock sedgeland 13. Juncus kraussii open-rushland 14. Leptocarpus brownii open-rushland• Communities dominated by herbs 15. Wilsonia backhousei herbfield to closed-herbfield 16. Samolus repens ± Schoenus nitens closed-herbfield.

Saltmarsh communities 1, 3, 12 and 14 were the most widely distributed in Tasmania; communities 1, 7, 12 and 13 commonly intergraded in variable mixtures in coastal saltmarshes. Communities 1, 2, 4, 5, 9, 10 and 14 were almost totally restricted to areas that were tidally inundated whereas, in contrast, communities 3, 6, 7, 8, 11, 12, 13 and 16 were often found outside the saltmarsh environment. The classification provided by Kirkpatrick & Harris (1999) is identical to that of Kirkpatrick & Glassby (1981) except for the addition of Community 11 (Deschampsia caespitosa tussock grassland). Note that Community 10 is dominated by the exotic *Spartina; compare with the classification of New South Wales saltmarsh communities by Adam et al. (1988) in which the Hydrocotyle bonariensis and Plantago coronopus communities are dominated by exotics.

SouthAustralia

Coastal saltmarsh and mangrove are currently being mapped for the entire coastline of South Australia (Canty et al. 2006). The classification is based first on broad landform classes, then on a dichotomy between estuarine and non-estuarine systems, then on patterns of tidal inundation (intertidal; supratidal; non-tidal; stranded tidal; intermittent tidal) and on cover. The cover category is effectively a habitat class, and consists of habitats such as sand, seagrass, mangrove, samphire, paperbark and sedges. Finally, an assessment is made of condition, divided into categories such as intact, uniform, patchy, dieback, degraded and prograding.


Units mapped in the South Australian investigation are the major habitat classes, developed into a five-tier coding system based on the categories of landform, tidal class, vegetation cover and vegetation intactness or condition. The category definitions allowed for 99 discrete codes, each of 8 numerals, to identify different types and condition of coastal wetlands. Different types of saltmarsh were not fully differentiated: all were considered together under ‘samphire’, although there is a differentiation between intertidal and supratidal vegetation.

Documents

1 literature review - Victoria University · chapter 1: literature review 5 1 literature review 1.1 Background to the study A primary driving force behind this project is the recognition