AUSTRALIAN
SALTMARSH
ECOLOGY EDITOR: NEIL SAINTILAN
AUSTRALIAN SALTMARSH ECOLOGY
AUSTRALIAN
SALTMARSH
ECOLOGY
Editor: Neil Saintilan
© CSIRO 2009 All rights reserved. Except under the conditions described in the Australian Copyright Act 1968 and subsequent amendments, no part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, duplicating or otherwise, without the prior permission of the copyright owner. Contact CSIRO PUBLISHING for all permission requests. National Library of Australia Cataloguing-in-Publication entry Australian saltmarsh ecology/editor, Neil Saintilan. 9780643093713 (pbk.) Includes index. Bibliography Salt marsh ecology – Australia. Salt marshes – Australia. Coastal zone management – Australia. Saintilan, Neil. 577.690994 Published by CSIRO PUBLISHING 150 Oxford Street (PO Box 1139) Collingwood VIC 3066 Australia Telephone: Local call: Fax: Email: Web site:
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Front cover photos by: Todd Minchinton (main), Pauline Ross, Jordan Iles, John Manger. Back cover photos by (clockwise from top left): Neil Saintilan, Neil Saintilan, John Manger, John Manger, Pauline Ross. Set in Adobe Minion 10/12 and Stone Sans Edited by Janet Walker Cover and text design by James Kelly Typeset by Desktop Concepts Pty Ltd, Melbourne Index by Russell Brooks Printed in Australia by Ligare The book has been printed on paper certified by the Programme for the Endorsement of Forest Chain of Custody (PEFC). PEFC is committed to sustainable forest management through third party forest certification of responsibly managed forests. CSIRO PUBLISHING publishes and distributes scientific, technical and health science books, magazines and journals from Australia to a worldwide audience and conducts these activities autonomously from the research activities of the Commonwealth Scientific and Industrial Research Organisation (CSIRO). The views expressed in this publication are those of the author(s) and do not necessarily represent those of, and should not be attributed to, the publisher or CSIRO.
Contents
List of contributors
vii
Preface
ix
Chapter 1
Australian saltmarshes in global context Paul Adam
1
Chapter 2
Distribution of Australian saltmarsh plants Neil Saintilan
23
Chapter 3
Geomorphology and habitat dynamics Neil Saintilan, Kerrylee Rogers and Alice Howe
53
Chapter 4
The ecology of molluscs in Australian saltmarshes Pauline Ross, Todd Minchinton and Winston Ponder
75
Chapter 5
Ecology of burrowing crabs in temperate saltmarsh of south-east Australia Debashish Mazumder
115
Chapter 6
Fish on Australian saltmarshes Rod Connolly
131
Chapter 7
Saltmarsh as habitat for birds and other vertebrates Jennifer Spencer, Vaughan Monamy and Mark Breitfuss
149
Chapter 8
Ecology and management of mosquitoes Pat Dale and Mark Breitfuss
167
Chapter 9
Protection and management of coastal saltmarsh Pia Laegdsgaard, Jeff Kelleway, Robert J Williams and Chris Harty
179
Chapter 10 Mapping, assessment and monitoring of saltmarshes Jeff Kelleway, Robert J Williams and Pia Laegdsgaard Index
211
231
v
List of contributors
Paul Adam
School of Biological Earth and Environmental Science University of New South Wales Mark Breitfuss
Queensland Bulk Water Transport Authority Rod M Connolly
School of Environment and Australian Rivers Institute Griffith University Pat Dale
Griffith School of Environment Centre for Innovative Conservation Strategies Griffith University Chris Harty
Chris Harty Planning and Environmental Management Alice Howe
School of Engineering University of Newcastle Jeff Kelleway
Rivers and Wetlands Unit NSW Department of Environment and Climate Change Pia Laegdsgaard
Coastal Ecology and Management Debashish Mazumder
Australian Nuclear Science and Technology Organisation Todd Minchinton
Institute for Conservation Biology and School of Biological Sciences University of Wollongong
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Vaughan Monamy
Faculty of Arts and Sciences Australian Catholic University Winston F Ponder
Australian Museum Kerrylee Rogers
Rivers and Wetlands Unit NSW Department of Environment and Climate Change Pauline M Ross
College of Health and Science University of Western Sydney Neil Saintilan
Rivers and Wetlands Unit NSW Department of Environment and Climate Change Jennifer Spencer
Rivers and Wetlands Unit NSW Department of Environment and Climate Change Robert J Williams
Aquatic Ecosystems Research Unit NSW Department of Primary Industries
Preface
As recently as 1990, Peter Fairweather described Australian saltmarsh as the least studied of all marine habitats, and ignorance of the ecological values of saltmarsh had been reflected in the relative lack of protection afforded to the habitat compared to other ecosystems. By way of contrast, mangroves have been recognised as an important fisheries habitat in Australia for nearly a century, and have a long history of protective legislation and regulations. Several decades of sustained research into mangroves through the 1970s and 1980s provided a comprehensive picture of their structure and composition and aspects of their ecology. Over the same period, little attention was give to saltmarshes, in spite of their occupying as much as 16 000 square kilometres of the Australian coastline and supporting more than three times the number of vascular plant species found in mangrove forests. Throughout the 19th and 20th centuries saltmarshes were replaced by playing fields, residential and commercial land and agriculture. We now know that in the closing decades of the previous century, mangroves began replacing saltmarsh from the seaward edge, a trend likely to continue with elevated sea levels as a result of global climate change. The decline of coastal saltmarsh in the southern half of the continent has now come to the attention of policy makers, and in New South Wales coastal saltmarsh has been declared an Endangered Ecological Community under the NSW Threatened Species Conservation Act. Fortunately, the growing awareness of the vulnerability of coastal saltmarsh has prompted more than a decade of research by a number of university and government scientists. While there is still much to be discovered about Australian saltmarshes, the time is ripe to dispel the myth that we know virtually nothing. This book provides the first synthesis of knowledge of Australian saltmarsh ecology. We hope it will stimulate greater interest in this fascinating habitat. The 10 chapters review geomorphology and biogeography, invertebrate ecology, the use of saltmarsh as a habitat by fish, birds and other mammals, and management issues including the control of mosquitos and the threat of invasive species. The picture which emerges is one of a vulnerable habitat which makes a unique and important contribution to the ecology of the coastal zone. Paul Adam’s opening chapter places Australian saltmarsh in a global context. Saltmarshes occur widely on estuarine and sheltered open coasts, and are immediately recognisable through a combination of habitat, vegetation physiognomy and elements of floristics. Australian saltmarshes exhibit patterns of variation at local, regional and continental scales which are similar to those elsewhere, but nevertheless have unique features. The distinctiveness of Australian saltmarshes is strongest in the south. The flora of southern saltmarshes has similarity with that across Gondwana, but with a number of Australian endemic genera and species. Whether patterns in faunal distribution reflect those in the flora is not known at geographic scales, either in Australia or elsewhere. Chapters 2 and 3 explore the biogeography and geomorphology of Australian saltmarshes. The possible impacts of climate change are introduced in these chapters. Saltmarsh diversity increases toward the colder latitudes, and a warming climate may well pose a threat to many species. In Chapter 3, Neil Saintilan, Kerrylee Rogers and Alice Howe present evidence that sea ix
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level rise in the southern and eastern regions of the continent may already be having a detrimental impact, promoting the colonisation of the saltmarsh environment by mangrove. Several chapters then consider the faunal ecology of Australian saltmarsh. Pauline Ross, Todd Minchinton and Winston Ponder provide a comprehensive account of the mollusc fauna of Australian saltmarshes in Chapter 4. The authors describe the unique adaptations of gastropods to the challenges of the saltmarsh environment, and their close association with the saltmarsh flora, both for habitat and food. In Chapter 5, Debashish Mazumder outlines the ecology of grapsid crabs, the dominant crustacean and in many ways the keystone of the saltmarsh ecosystem. Crabs in Australian saltmarshes have limited opportunity to spawn, but on the few occasions the tide inundates the upper intertidal saltmarsh a mass spawning ensues. This event provides a link between the trophic ecology of crabs and fish, and is explored by Rod Connolly in Chapter 6. Many species of fish enter the saltmarsh on the spring tide, including several species of direct commercial importance. Chapter 7 considers the importance of saltmarsh as a habitat for a range of terrestrial species, including birds, bats and other mammals. The significance of saltmarsh for migratory shorebirds has only recently been appreciated in the published literature, partly because the saltmarsh is primarily used as a night-time roost, a time when few ecologists are active. The shallow pools of the saltmarsh afford good protection from many predators, as well as a secondary feeding habitat. Among the other night-time visitors to the saltmarsh are several species of insectivorous bat, including some threatened species. There are numerous species of insects which may be attracting bats to the saltmarsh. One such species is the saltmarsh mosquito, Aedes vigilax. The ecology and management of the saltmarsh mosquito forms the subject of Chapter 8 by Pat Dale and Mark Breitfuss. The saltmarsh mosquito is a biting nuisance in many coastal communities and in some locations a vector of the Ross River virus. There are a number of other viruses which cause disease in humans which could be transmitted by mosquitoes, and with global warming the incidence of infection may increase. Perceptions of this risk will need to be addressed to ensure that public opinion continues to support wetland conservation. Several strategies for mosquito control are discussed and their ecological consequences described. The final two chapters provide an overview of management issues and responses. Pia Laegdsgaard, Rob Williams, Jeff Kelleway and Chris Harty describe the effects of overgrazing, use of off-road vehicles, dumping of waste and reclamation. The legislative and policy responses of the various Australian jurisdictions are discussed and the importance of community awareness is stressed. Implementation of conservation measures for saltmarsh is dependent upon us knowing where it is, and the final chapter provides guidelines which should improve the mapping and monitoring of saltmarsh by natural resource managers. While the book goes a long way towards redressing the common misconception that little is known about Australian saltmarsh, a common refrain in many chapters is that there is still much to discover. Several fruitful areas of research are proposed, notably an improved understanding of the ecophysiology of saltmarsh plants, the study of saltmarsh insects and their trophic dependencies, and a better appreciation of the ecology of saltmarshes in the tropical north and the arid west of the continent. Studies of ecosystem processes have been out of fashion for some time, although there are indications of a resurgence of interest. Saltmarshes, particularly in the USA, were amongst the earliest ecosystems subject to process studies, and these early results have entered textbooks as generalisations applicable to all saltmarshes. Given the differences in floristic composition, climate, tidal regimes and sediment fertility it is likely that quantitatively, Australian saltmarshes will differ from those in the USA, and it would be highly desirable if we had local studies – although these will require multidisciplinary teams
Preface
and substantial budgets. It is the hope of the authors that this book will inspire the next generation of saltmarsh ecologists to answer some of these questions. Finally, there are a number of people who deserve our thanks, not the least the numerous honours and graduate students whose diligent work has contributed so much to our present understanding of Australian saltmarshes, braving hot days, cold nights, mud and mozzies. Thanks are also due to the team at CSIRO Publishing for their enthusiasm and support, including John Manger, and Briana Elwood. Janet Walker did a superbly efficient job with editing, as did Frank Saintilan. Neil Saintilan and Paul Adam August 2008
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CHAPTER 1
Australian saltmarshes in global context Paul Adam
Introduction Coastal saltmarshes are recognised globally as ecosystems of high ecological value which are increasingly under threat (Adam 2002; Valiela 2006). While there is increasing acknowledgement of their importance in Australia, and their ‘Cinderella’ status, demonstrated by Fairweather (1990), has improved over more recent times, they are still relatively unknown compared with the intensively studied marshes of Europe and North America. Coastal saltmarshes can be defined as intertidal communities dominated by flowering plants, principally herbs and low shrubs. They are found on soft substrate shores of estuaries and embayments, and on some open low wave energy coasts. Coastal saltmarsh is also found on the shores of intermittently open saline coastal lagoons. When these lagoons are open to the sea they are tidal (although tidal amplitude may be attenuated in comparison to nearby open shores), but when closed, which is often the majority of times, water level fluctuations are climate driven and lack predictable periodicity. Saltmarsh is distinguished from adjacent communities by both floristic composition and structure. Mangroves are dominated by trees (and amongst the world’s forests are unusual in the virtual absence of an understorey). The boundary between saltmarsh and mangrove is normally sharp, but on temperate coasts there are sites with mosaics of saltmarsh and mangrove where there are groves of Avicennia marina interspersed amongst saltmarsh and at the southern global limit of mangroves in Victoria mangroves are stunted and may be lower than the chenopod shrub Tecticornia arbuscula on adjacent saltmarsh. Seagrass beds are predominantly subtidal and are dominated by various monocots (although none are true grasses). The upper limit of saltmarsh is set by the level of the highest tide, but the nature of the transition to terrestrial vegetation will be determined by topography, and, in urban areas, human activity. Coastal lowlands have been very heavily modified in temperate and subtropical regions, so that natural transitions to terrestrial vegetation are becoming uncommon (Figure 1.1). Swamp forests on coastal flood plains often have an understorey of saltmarshes in the transition zone, which may be inundated with brackish water during storms. Species found in intertidal saltmarsh are also characteristic of seepage zones on seacliffs and rock platforms above the tidal limit, and on some of the most exposed cliffs and headlands, extensive swards (covering hectares) of saltmarsh species are found tens of metres above the sea but subject to high inputs of aerosolic salt. Australia has very extensive areas of saline soil inland – some of these are of natural origin, but salinisation of agricultural and urban land is one of the major environmental problems to 1
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Figure 1.1 Temperate saltmarsh. Newington, Parramatta River NSW. Marsh dominated by Sarcocornia quinqueflora, and fringing stand of Casuarina glauca.
be addressed as a national priority. Some of the species of inland saline areas also occur in coastal saltmarshes, but the majority of the vegetation comprises species in the same genera as those on the coast, but endemic to the inland. At various geological times parts of the inland would have been shallow seas, and, at others, seed transporting birds could have carried propagules between the coast and inland, so that there will have been ample opportunity for gene and species exchange, as well as periods of isolation of local populations and opportunities for speciation. Towards the head of estuaries, conditions may be brackish or fresh, but still subject to tidal influence. Fringing reed and tall sedge communities in the freshwater tidal zone have been very heavily impacted by urban and agricultural development, and by hydrological change as upstream water abstraction reduces freshwater input. There have been few studies of freshwater tidal marshes in Australia. In tropical Australia the upper intertidal, flooded by the tides only infrequently, develops hypersalinity during the dry season. The vascular vegetation of these hypersaline flats is extremely sparse and contains only a small number of mostly succulent species (Batis argillicola, Cressa cretica, Sesuvium portulacastrum and Tecticornia australasica). Although the vascular plant cover is very low, there is a skin of microalgae and cynobacteria amongst the salt crust and extending some millimeters into the underlying sediment. The ecology of these flats has not been extensively studied, but they may make a considerable contribution to estuarine productivity; around the Gulf of Carpentaria considerable quantities of salt and nutrients are released from hypersaline flats during king tides (Ridd et al. 1988). Similar extensive hypersaline flats occur on arid coastal zones elsewhere, and are known as ‘sabkha’ in the Middle East. There is no consensus as to whether these flats should be regarded as saltmarsh or as a separate ecosystem. This uncertainty renders it difficult to determine the
Australian saltmarshes in global context
extent of saltmarsh in Australia, as different estimates have been made on different bases. However, the area of flats is probably roughly the same as the area of fully vegetated saltmarsh. While extensive hypersaline flats are a feature of tropical coasts, smaller bare patches are found within temperate marshes. On the central NSW coasts such patches were formerly frequent on the Parrramatta River, Cooks River, Botany Bay / Georges River (Hamilton 1919; Clarke and Hannon 1967). In the last few decades many of the patches have become vegetated, and the few that remain have been damaged by vehicle use (extensive new bare patches have been created by off-road vehicular use – Kelleway 2005). Whether the revegetation of bare areas is a response to natural environmental change, or whether it reflects human influence (such as greater discharge of stormwater into marshes) remains to be determined.
The saltmarsh environment The saltmarsh environment is a challenging one for many plants, explaining the relatively small flora and its general similarity around the world. Although the flora is not a single taxonomic lineage, and the adaptations necessary to survive the saltmarsh environment have evolved independently on a number of occasions, relatively few families are represented within it. The factor which distinguishes saltmarsh (and mangroves) from other vascular plant communities is tidal inundation. Tides are highly predictable, but the interactions between tides, weather, groundwater influences and vegetation result in complex patterns of environmental variation (Figure 1.2).
Figure 1.2 The interaction between environmental factors and vegetation in saltmarshes. Redrawn and modified from Clarke and Hannon (1969).
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The tidal regime varies considerably around the coastline. In south-west Western Australia and parts of the Gulf of Carpentaria, tides are diurnal, with a single low and high tide a day on the cycle of 24 hours and 50 minutes. Elsewhere tides are semi-diurnal, with two highs and lows a day, or mixed, when the two highs per day differ considerably in height. Tidal ranges are also extremely variable – in southern Australia mostly low (micro- to mesotidal) but with higher ranges in bays and inlets, while in northern Australia tidal ranges are generally high with a maximum of more than 8 m in north-west Western Australia. The tidal range determines the vertical extent of saltmarsh, but the horizontal extent will depend on the local topography and geomorphology and there are laterally extensive stands at sites with low tidal range and narrow fringes on coasts with high tidal ranges, although for a given surface gradient the higher the tidal range the wider the saltmarsh. As a consequence of tidal inundation the soils in saltmarsh are saline; the lower on the shore the more frequent the inundations and the less variable the soil salinity. However, at higher levels of the shore salinities can vary considerably depending on the balance between rainfall and evaporation. Inundation will also result in anaerobic soil, although the duration of waterlogging will depend on sediment type and local drainage. Tidal flooding has other effects on plants. Tidal currents, which increase with tidal range, may dislodge seedlings, so that recruitment may require sufficient long windows of opportunity between inundation to permit germination and development of sufficiently robust seedlings. Estuarine water may be turbid so that after tidal flooding vegetation may be coated with sediment, possibly reducing photosynthesis. Submergence may also change the effective day length and expose plants to a sudden temperature shock. The physiological consequences of these stresses have not been studied. The interaction between the environmental conditions and species results in a general zonation of species (Adam 1990); the more frequently inundated lower marsh providing habitat for fewer species than the higher levels. Communities are also zoned, but at any given level on the shore there is often a mosaic of communities rather than a continuous band of a single community (Zedler et al. 1995). Local microtopographic change to drainage conditions is often reflected in the vegetation mosaic (despite the absence in many Australian saltmarshes of the well developed creek and pan systems which are a feature of saltmarshes elsewhere – Adam 1990, 2000). The zonation of saltmarshes is often interpreted as the spatial expression at one point in time of succession. Conceptual models have been developed in which species colonise mud or sand flats and promote accretion and stabilisation of sediment. As the elevation of the marsh surface rises, frequency of tidal inundation declines and environmental conditions permit the establishment of other species which displace the primary colonists. Continued expansion of primary colonists seawards results in zonation. This basic model of sedimentation driven succession, with various additional complexities to account for variation in relative sea level, is sustained by empirical evidence, but interpretation of zonation as a reflection of succession in Australia is less certain. Pidgeon (1940), influenced by the Clementsian approach which was then one of the major paradigms of ecology, proposed that the zonation of intertidal communities on the New South Wales coast could be interpreted as resulting from succession, and this view has become part of received wisdom. If true, it would be a very atypical successional sequence as it would imply that the primary colonists were trees (mangroves), subsequently replaced by dwarf shrubs and herbs. Pigeon’s model also postulated that the succession continues above the highest astronomical tide level through Casuarina glauca forest to eucalypt swamp forests. In the absence of a drop in relative sea level it is difficult to see that the proposed succession would be driven by allochthonous sedimentation or that autochthonous sedimentation (peat formation) would be sufficient to elevate the surface out of the tidal frame.
Australian saltmarshes in global context
The relationship between saltmarsh and mangrove is complex (see Chapter 3) but it is difficult to accommodate a transition from mangrove to saltmarsh within the standard model of saltmarsh formation. There are few sites where active formation of new saltmarsh is occurring, with the exception of invasion by Spartina anglica. Long-term successional development in Australian Spartina marshes, if it occurs, has yet to be described.
Flora and vegetation of Australian saltmarsh Accounts of saltmarsh flora and vegetation have been published for a number of parts of the Australian coast, including: inter alia by Hamilton (1919); Saenger et al. (1977); Adam (Adam 1981a, b; Adam et al. 1988; Adam and King 1990; Adam 1994); Bridgewater (Bridgewater, Rosser and de Corona 1981; Bridgewater 1982; Bridgewater and Cresswell 1993, 2003; Cresswell and Bridgewater 1998); Craig (1983); Kirkpatrick and Glasby (1981); Thannheiser (2001); Johns (2006); and Kelleway et al. (2007). Saltmarshes occur globally, and most, although exhibiting local characteristics, have immediately recognisable similarities. Australian saltmarshes are no exception; in terms of physiognomy and composition (particularly at generic level) they are similar to saltmarshes elsewhere. Within terrestrial biomes, as a general rule, species richness is highest in the tropics and declines at higher latitudes. This is also the case with mangroves, but saltmarshes show a strikingly different pattern (Adam 1990). Tropical saltmarshes in Australia are extremely depauperate, but species richness increases in temperate latitudes, with the highest number of species being recorded from Tasmanian marshes (Saenger et al. 1977; Bridgewater and Cresswell 2003). Within individual marshes, species richness is generally lowest at low, more frequently tidally inundated elevations and increases higher up the shore, although if freshwater input permits the establishment of tall competitive dominants, such as Phragmites or Typha, in the upper marsh, species richness is again low. The broad geographic scale patterns of variation in species and community distribution within Australian saltmarshes are similar to those elsewhere (Adam 1990). The distinction between tropical and temperature saltmarshes is seen not just in changes in species richness, but also in the distribution of individual species. Some of the tropical species are widespread outside Australia on hot dry shores (Batis, Sesuvium, Cressa) but Tecticornia australasica (Figure 1.3) is an Australian endemic element in the flora. On more temperate shores the flora has a large, widely distributed element (at both generic and species level) as well as a strong Gondwanan element. Adam (1990) has argued that there
Figure 1.3 Distribution of Tecticornia australasica (from literature records, personal observation and records in the Australian Virtual Herbarium).
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Figure 1.4 Distribution of Selliera radicans (from literature records, personal observation and records in the Australian Virtual Herbarium).
is an overall similarity of flora and vegetation in saltmarshes in South Africa, south-western and south-eastern Australia, New Zealand and temperate South America. Links include the upper marsh rush Juncus kraussii, Sarcocornia spp, Triglochin striata, Cotula coronopifolia and Samolus spp. One species common to Australia and New Zealand is Selliera radicans (Figure 1.4), from a family (Goodeniaceae) absent from northern hemisphere saltmarsh floras. (On the central NSW coast there is evidence of a recent decline in S. radicans, only partly explained by habitat loss (Adam et al. 1988; Kelleway et al. 2007). On coasts with a strongly seasonal Mediterranean climate, saltmarsh vegetation is characterised by shrubby chenopods (formerly in the genera Halosarcia and Sclerostegia, but following a recent taxonomic revision by Shepherd and Wilson (2007), Tecticornia spp.). The dwarf subshrub Frankenia is also characteristic of Mediterranean zone saltmarshes. Some of the less common species from brackish upper marsh communities exhibit remarkable transhemisphere disjunctions, which, if they have been correctly identified, may reflect the legacy of past long distance dispersal events, possibly by migratory waders. Examples include Limosella australis, which, amongst other occurrences in Australia, is found in upper saltmarshes on the south coast in NSW, but is an endangered wetland plant in Wales, and Isolepis cernua, widespread, although not abundant, in upper marsh flushes in eastern Australia but which is much rarer in northern Europe. A recent discovery in Australian saltmarshes is the dwarf Eleocharis parvula, one of the most inconspicuous saltmarsh species. Is this a recent introduction, or a cryptic species which had been previously ignored? In the northern hemisphere E. parvula has a circumboreal distribution, but with many disjunctions. Clearly E. parvula has not been deliberately introduced into Australia, and it is difficult to envisage a mechanism for accidental introduction by human agency. The habitat of E. parvula in Australia is similar to that in which it occurs in the northern hemisphere and it is not impossible that it was introduced to Australia by birds, possible a long time ago. We may never be able to determine the origin of E. parvula in Australia, although molecular comparison with northern hemisphere populations may in the future provide insights. Many saltmarsh species have very wide distributions, both at the local scale, within individual marshes, and geographically. This wide amplitude is made possible by the species being made up of many genotypes (Adam 1990). There have been few studies of the genecology of Australian saltmarsh species, but one of the most widespread saltmarsh grasses, Sporobolus virginicus, has been shown to be genetically very variable (Smith-White 1981, 1988). The presence
Australian saltmarshes in global context
of genetic variation may facilitate the response of species to climate change, but also has implications for the use of planting material in rehabilitation or recreation projects. The wide distribution of species may suggest that any material could be used (including commercial cultivars) in these projects, without care being taken to match the genotype to the new environment.
Non-vascular flora The vascular plants are of the visibly dominant component of saltmarshes. Other plants may, however, play important roles in the ecosystem. While there have been many studies of algae in Australian mangroves, saltmarsh algae have been rarely studied. However, they are likely to be as important as algae in saltmarshes elsewhere – contributing to primary productivity, stabilising sediment surfaces, being the food source for filter and surface feeders and, in the case of cyanobacteria, which form part of the algae skin on the sediment surface, functioning as nitrogen fixers. Bacteria and fungi play a major, although in the Australian context largely unquantified, role in decomposition and chemical transformation and provide food for filter feeders. Although it has long been known that a number of European vascular halophytes are vascular-arbuscular mycorrhizal (VAM, summarised in Adam 1990), it is only recently that more detailed studies have been undertaken (Davy et al. 2000). There has been no systematic investigation of VAM in Australian halophytes, although Samolus repens does support VAM (pers. obs). Bryophytes and lichens are not generally considered to be components of saltmarshes, although in some northern hemisphere marshes bryophytes form a characteristic element in the vegetation (Adam 1990). In Australia, bryophytes and lichens are largely absent from saltmarshes, although there are occasional occurrences at the highest driftline or as epiphytes on shrubby chenopods.
Additional factors The environmental factors incorporated in Figure 1.2 are universal in space and time – the interplay of the factors with the pool of Australian halophytic species results in a range of distinctive communities, in the same way that the same factors applied to a different range of species elsewhere would produce a different suite of communities. Human activity results in additional factors coming into play. Climate change Increased concentrations of greenhouse gases, including carbon dioxide and methane, are predicted to lead to global warming and other changes in climate conditions. One consequence of global warming will be a rise in sea level, initially as a result of thermal expansion but in the longer term further contributed to by melting of icecaps. The effects of global sea level rise will not necessarily be translated uniformly into changes in relative sea level, as tectonic movement or increased sedimentation could counter the rise in water level at the local scale. However, the Australian coast is tectonically relatively quiescent, and sediment supply, although increased through catchment erosion since European settlement, is still not great. The rise in sea level is thus likely to result in increased inundation of saltmarsh and the retreat of the tidal limit inland. In southern Australia, where many saltmarshes now abut infrastructure and urban or agricultural development, this will result in coastal squeeze, as the seaward edge of saltmarsh is lost to mangrove invasion or inundation beyond flooding tolerance, and the expansion inland is prevented by lack of habitat. However, in northern Australia, where much of the coast
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is still undeveloped, there are few impediments to the establishment of saltmarsh over what are now terrestrial communities. Although warming is likely to be a universal phenomenon, other aspects of the climate – rainfall amount, intensity and temporal distribution, general storminess and frequency and intensity of major storms – are likely to vary at local and regional scales although currently available mathematical models do not permit detailed modelling of probable changes. However, changes in any of these factors are likely to be reflected in changes in saltmarshes – changes in rainfall regimes will alter the patterns of variation in soil salinity, change in storminess, but particularly in the intensity of major storm events, could result in erosion of saltmarsh vegetation. Warming may result in an expansion southwards of the range of northern species, and through competition this might produce a contraction in southern species. Increased temperatures might favour mangroves at the expense of saltmarsh, and might also favour some weeds over native species. While there remains uncertainty over the details of climate change and the biological consequences, even the most extreme climate change deniers would recognise that the atmospheric carbon dioxide concentration has increased and will continue to do so. This in itself will have profound effects on saltmarshes. Simplistically it might be thought that an increase in carbon dioxide will result in greater photosynthesis and, hence, greater ecosystem productivity. However, an increase in carbon dioxide will lead to greater growth of plants with the C3 photosynthetic pathway, altering the current competitive balance between C3 and C4 species. There will also be an increase in water use efficiency so plants will transpire less and the soil moisture regime will be changed. An increased carbon dioxide concentration is also likely to result in a decreased leaf protein content, which will have flow-on effects through the ecosystem. On current information it is not possible to predict whether lower leaf nitrogen will result in greater herbivory as herbivores need to consume more to achieve the same nitrogen input or lesser herbivory as the decline in resource quality deters herbivores. Experimental studies in the glasshouse, and in the field in American saltmarshes, confirm that carbon dioxide concentrations to the levels predicted over the next century will result in shifts in the relative abundance of C3 and C4 species (in favour of C3), changes in water use efficiency, root-shoot ratio and in nitrogen content (Drake et al. 1989), and it is probable that such changes will be experienced on saltmarshes globally. Pollution Climate change could be regarded as a global consequence of pollution, but there is a range of more local pollution events which could have considerable impacts on saltmarshes. Given that estuaries have been sites for industrial development for centuries, there is a considerable legacy of industrial pollution in estuarine saltmarshes. Metals released into estuaries may be incorporated into saltmarsh sediment, where under reducing conditions their biological availability is lessened. Disturbance to sediments (through dredging or reclamation) may result in oxidation of these metals and their release into the environment in much more biologically available and toxic forms. High levels of heavy metals in saltmarsh sediments have been recorded in many estuaries around the world, and in Australia the association of industry, particularly smelting, with estuaries in the southern part of the continent is reflected in elevated levels of metals in waters, sediments and vegetation (see for example Woods et al. 2007). While the Precautionary Principle would indicate that metal accumulation should be regarded as a concern, and measures taken to prevent new discharges and reduce or eliminate existing sources, the ecosystem-level consequences of metal contamination are less clear (Williams et al. 1994; Valiela 2006) possibly because of the lack of studies which have examined ecosystem processes. Most saltmarsh plants, which are physiologically
Australian saltmarshes in global context
adapted to a stressful environment, may be constitutively tolerant of metal pollution. In Europe, where a number of saltmarsh species are harvested for human consumption, concern has been expressed about possible adverse consequences for human health of eating saltmarsh plants with high metal concentrations (Beeftink et al. 1982), but the effects of metals in saltmarsh plants on the detrital food chain and on direct herbivory by fauna have been poorly studied. One of the most dramatic forms of pollution to saltmarshes is from oil spills. A number of major spills from shipping accidents have affected saltmarshes, but Australia has fortunately not yet experienced a very large spill, although a number of small spills have affected small areas of marsh. Experience around the world suggests that in most cases, as much or more damage, is done in attempting to clean up oil spills in saltmarshes as that incurred directly from the spill (Baker et al. 1994). In general the recommendation is to prevent, as far as possible, oil reaching saltmarsh but if oiling does occur, to permit it to degrade naturally. In tropical regions breakdown is likely to be fairly rapid, but in temperate and polar latitudes residues may remain in sediment for decades and continue to have biological effects throughout this time (Culbertson et al. 2007). In Australia saltmarshes have been identified as ecologically sensitive communities in oil spill contingency planning; if at all possible booms would be deployed to prevent oil reaching saltmarshes, and if oiling of saltmarshes does occur, dispersants would not be used (for example see Carter 1994). As well as tide-borne incursions of oil, terrestrial chemical and oil spills (from road or rail accidents) could reach saltmarshes, and the emergency services would need to manage any oil and chemicals in stormwater drains and waterways so as to minimise impacts on saltmarshes. Runoff of nutrients, herbicides and pesticides into estuaries has impacts on saltmarshes. Eutrophication of waterways has resulted in algal blooms smothering saltmarshes. For the Peel-Harvey estuary (WA), addressing this eutrophication was a very expensive catchment wide process, involving measures to reduce fertiliser application and runoff, and re-engineering a new opening of the estuary into the ocean (Brearley 2005). Herbicide residues (particularly simazine) have been shown to adversely affect microalgae in European saltmarshes (Mason et al 2003). As these algae help stabilise the marsh sediment this is regarded as an important issue to be addressed. Few data are available from which to determine whether there are similar impacts in Australia. A range of other human impacts has been identified on saltmarshes overseas (Adam 2002), many of which are also likely to be relevant in Australia.
Surviving and thriving Plants which characteristically occur in saltmarshes are referred to as halophytes, and in recent decades considerable attention has been given to the mechanisms permitting survival at high salinity. Plants unable to tolerate high salinity are referred to as glycophytes. There is a spectrum of salt tolerance amongst angiosperms, from those intolerant of very low salinities (for example, avocado) to those capable of completing their lifecycle in salinities well above that of seawater. Although many saltmarsh plants can grow in salinities well above seawater it is unlikely that any species has an obligate requirement for high levels of salts (Barbour 1970), and, in the absence of competition from glycophytes, most halophytes grow well under non-saline conditions. The mechanisms of salt tolerance have attracted a great deal of study and we have a good overall understanding of how they operate, although we do not have a complete detailed account of the physiology of every single species.
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For any plant to grow, its cells must be turgid. In order to maintain turgor the water potential (y) within the tissues must be lower (more negative given that, arbitrarily, the y of pure water is defined as 0) than that of the water in the soil surrounding the roots. The problem for halophytes is how low water potentials can be generated without impairing metabolism or survival. In general, halophytes accumulate large quantities of inorganic ions in their tissues sufficient to generate the necessary low water potentials. However, enzymes of halophytes are no more salt-tolerant than these of glycophytes, so there must be internal compartmentation of salt within the cell. If inorganic ions are accumulated in vacuoles (which comprise the majority of cellular volume in most plant tissues) then they must be osmotically balanced by other solutes in the cytoplasm. These so-called compatible solutes must not impair enzyme function even at high concentration, and preferably be of low molecular weight, uncharged and neutral. A small number of types of compound have been identified as functioning as compatible solutes – proline, quaternary ammonium compounds (particularly glycine betaine) and sugar alcohols. Even though halophytes accumulate ions there must be an ability for plants to control the rate of supply of ions to shoots and to adjust tissue salt content. For all halophytes which have been studied there is exclusion of salt from the xylem sap, such that the salt concentration in the transpiration stream is always much less than that at seawater (rarely exceeding 10% seawater); this selective control of ion uptake is one of the major distinguishing differences between glycophytes and halophytes – glycophytes cannot exclude salt, and accumulate lethal concentrations in their tissues. Even with a high degree of exclusion other mechanisms to regulate ion content are required. Reduction in transpiration rate lowers the rate of supply of ions to the leaves – this explains the apparent paradox of many halophytes appearing xeromorphic while growing in an environment with abundant water. A number of species, but far from all, actively excrete salt from stems and leaves through glandular structures known as salt glands which appear to have evolved independently in a number of different flowering plant lineages. Examples of widespread saltmarsh species with salt glands include Frankenia spp., Limonium spp., Samolus repens and S. junceus, and Sporobolus virginicus (Adam 1990; King et al. 1990). Maintenance of tissue salt concentration can also be achieved through becoming succulent – so leaves of species grown in saline conditions are often markedly more succulent than those of the same species grown under non-saline conditions. Succulence is also a feature of plants employing the CAM photosynthetic pathway, but while succulent coastal plants such as Carpobrotus are CAM, no intertidal saltmarsh CAM plant has been reported. There are considerable metabolic costs associated with salt tolerance – including diversion of photosynthate from growth to the synthesis of compatible solutes, and the energy consumption involved in selective ion uptake and secretion. Low transpiration rates inevitably result in low carbon dioxide uptake and hence low growth. These costs are probably the major explanation for the exclusion in the field of halophytes from non-saline conditions. Although the growth of halophytes is maximal under non- or low saline conditions, the maximum growth rates are lower than those of glycophytes under the same conditions. Under non-saline conditions halophytes are likely to be out-competed by more vigorous glycophytes. The other major physiological problem for many plants growing in saltmarsh is the presence, either permanently or temporarily, of water-logged soils, which lack oxygen and may contain phytotoxins. Few angiosperms can survive such conditions for long periods but mostly they do not experience anaerobic conditions because of the presence of efficient internal aerenchyma which permits the passage of oxygen into the root system, and may also permit
Australian saltmarshes in global context
leakage of oxygen into the soil, detoxifying potentially hazardous chemicals, and providing an oxygenated rhizosphere occupied by aerobic microbes and micro and meiofauna. Although there are metabolic costs associated with salt tolerance and surviving flooding, saltmarshes are widely regarded as highly productive, although the data supporting this claim are primarily from North American Spartina marshes. The recognition of the high productivity of Spartina gave rise to interest in the fate of the plant material and led to the ‘outwelling hypothesis’ suggesting that saltmarsh production exported to estuarine and coastal waters sustained the food chain, including commercially and recreationally important fisheries. This in turn led to changes in public perceptions of the value of saltmarshes and, in the USA, introduction of legislation to protect saltmarsh habitat. While aspects of the outwelling hypothesis have been modified (Adam 1990) the high productivity of Spartina on the Atlantic and Gulf Coasts of the USA is firmly established. An assumption that estimates of productivity based on American data can be extrapolated to Australia underpinned early moves to conserve Australian saltmarshes. However, there have been few studies which have measured above-ground standing crops in Australia and even fewer which have estimated productivity. There are reasons to suggest that American results are not directly applicable to Australia: Spartina alterniflora dominated marshes occupy a lower part of the intertidal than most Australian marshes, and have a growth form (a very tall grass) not represented in the lower zones of Australian marshes. In many ways S. alterniflora marshes represent a unique biome, differing not just from Australian saltmarshes, but from saltmarshes in other continents as well. In Australia, the lower intertidal zone is usually occupied by mangroves rather than saltmarsh, except in Tasmania. There has been no study of productivity of introduced Spartina anglica in Tasmania and Victoria. It would be of interest to know whether the species’ productivity is comparable to that overseas, and whether the advent of flowering plants into a zone previously unvegetated has had repercussions throughout the estuarine ecosystem. The studies that report biomass and productivity in Australian saltmarshes are both from temperate coasts (south-west Western Australia, Congdon and McComb 1980; NSW, Clarke and Jacoby 1994), so that again they do not provide a strong basis for extrapolation to the whole coastline of the continent. The significance of algal production in saltmarshes has been little studied anywhere, but algae, which are often abundant on the sediment surface, potentially represent a high quality food for detritivores. The recent studies on tracing the flow of energies between trophic levels in marshes and out into adjacent estuaries (Chapters 5 and 6) will provide us with a more secure basis for understanding the role of saltmarshes in estuaries.
Introduced species Globally the spread of introduced species is regarded as a major threat to biodiversity, second only to habitat loss in its impact (Mooney and Hobbs 2000). While few terrestrial ecosystems are free from weed invasion, saltmarshes would a priori be unlikely candidates to be seriously affected for two reasons. Firstly, the environmental conditions of saltmarsh present a severe physiological challenge which many potential weeds would be unable to overcome, and secondly, many saltmarsh species have propagules which can be dispersed easily, either by flotation or by birds. The very wide distribution of some saltmarsh species, in many cases extending across several continents, reflects the ease of dispersal, and suggests that many species have reached their distributional limits without deliberate or accidental assistance from humans. Long distance transport of viable propagules attached to, or carried within, birds has been demonstrated, and the major migratory pathways of waterfowl and waders are predominantly north-south in orientation. In the mid and high latitudes of the southern hemisphere there is a
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powerful circumpolar atmospheric and oceanic circulation which could potentially carry propagules from west to east between continents (Kloot 1984). It is probable that much of the saltmarsh flora of Australia originated elsewhere, but colonised the continent prior to human occupation. It is important, however, to be able to identify those species which have arrived by human agency, so that management resources are not devoted to conserving them, but rather that attention is given to controlling or preventing further spread. The wide distribution of species may generate debate as to their native status. In some cases the debate can be resolved by closer examination of the species and taxonomic re-evaluation. For example, the common rush of upper saltmarshes in temperate Australia and New Zealand was for many years referred to Juncus maritimus, but systematic investigation by Snogerup (1993) confirmed that it was distinct from the northern hemisphere J. maritimus and that the valid name was J. kraussii. As J. kraussii, it is now clearly seen to be a member of a Gondwanan flora element, rather than having a disjunct bipolar distribution if regarded as part of J. maritimus. Most of the taxonomic issues in the saltmarsh flora have been resolved, but there is still confusion over what species of Spergularia (Caryophyllaceae) occur in Australia and whether or not they are native. Even if the taxonomy and nomenclature are well established there may still be disagreement as to whether species have been introduced by human agency. There has been long running uncertainty in this respect in regard to a group of species characteristic of temperate Gondwanan saltmarshes, including inter alia Cotula coronopifolia, Lobelia alata, Samolus repens and Triglochin striata. Doubts about the native status of members of this group in Australia were first expressed by Mueller (1868) and reinforced by Kloot (1984). However, there is no evidence of deliberate introduction, nor are there indications of spread within the period of historic records; all occupy comparable habitats in the different continents. Cotula coronopifolia may fluctuate in abundance, and can become locally dominant in ephemeral brackish or freshwater accumulations in upper saltmarsh. I would regard this as an opportunistic response to favourable environmental conditions rather than offering support for an hypothesis of recent invasion. Spartina anglica Globally the major exotic invasive species in saltmarshes are grasses in the genus Spartina. Although a number of species have been introduced around the world (for example Ayres et al. 2004 describe the invasion of four exotic Spartina species in San Francisco Bay), in Australia only one species – S. anglica has been introduced (Figure 1.5). Spartina anglica originated in Southampton Water on the south coast of England as a result of spontaneous hybridisation between the native European S. martina and the American S. alterniflora which had been accidentally introduced to Britain. The first collections in 1870 have subsequently been identified as a sterile F1 hybrid and named S. townsendii. In the 1890s there was rapid spread of grass similar to S. townsendii, but which was fertile. Marchant (1963, 1968) showed that the fertile form was an amphidiploid derived from the original sterile hybrid by chromosome doubling. This fertile form was later named S. anglica (Hubbard 1968). S. anglica spread rapidly, and was able to colonise lower levels on the shore than existing saltmarshes and promoted sediment accretion and stabilisation. These properties were viewed as highly desirable for coastal engineering and as fodder, and seeds and transplant material were exported from Britain to countries around the world (Ranwell 1967). The history of the introduction of Spartina to Australia was described by Boston (1981). The first known introduction was a planting in Corner Inlet, Victoria, probably in the 1920s, but this failed to establish. From 1927 and through the 1930s numerous introductions, prima-
Australian saltmarshes in global context
Figure 1.5
Spartina anglica, Tamar Estuary, Tasmania.
rily as seed, were made and plantings occurred in all states, including in the tropics (Boston 1981). Most of the introductions failed; Boston (1981) suggests that lack of care in the treatment of seed meant that much of the material that reached Australia was not viable. S. anglica was successfully introduced into the south-east of the continent, with currently extensive stands in Tasmania and Victoria. It is also widespread in New Zealand. From the perspective of those who promoted the introduction of S. anglica its establishment would be regarded as justifying their enthusiasm, while the failure to spread around more of the coast would be a matter of regret. However, there has been a substantial change in attitudes the focus today is on control. Spread of exotic Spartina is viewed with concern because of the threat to biodiversity posed by expansion into previously unvegetated mudflats, gene flow to native Spartina spp., and low species richness of the new communities dominated by the exotic invader. The change in attitude is well illustrated in Tasmania. S. anglica was planted in the Tamar estuary for the Marine Board of Launceston with the hope that if the mudflats were stabilised by vegetation, ‘it would force the stream flow into the central part of the river, creating a scouring effect and keeping the main channel relatively free of mud’ (unpublished letter from GJ Martin, quoted by Pringle 1993). In terms of vegetating mudflats in the Tamar S. anglica was extremely successfully, spreading explosively in the 1960s (Pringle 1993), and also colonising other intertidal areas in northern and eastern Tasmania. (RGAG 2002). By 1997 S. anglica dominated about 600 hectares of intertidal habitat in Tasmania (RGAG 2002). There was growing concern about adverse impacts on the integrity of native saltmarsh, loss of habitat for migratory waders (several of the sites colonised by Spartina are wetlands of international importance listed under the Ramsar convention), potential for changes to estuarine hydrodynamics and nutrient cycling which could threaten aquaculture (a growing and important component of the Tasmanian economy), and loss of aesthetic values and shoreline access, impacting on coastal tourism.
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The control of Spartina is made more difficult by the nature of its habitat. The soft unconsolidated sediment makes access by machinery difficult, and even pedestrian access can be both difficult and hazardous. Experience in Australia and overseas suggests that physical removal by pulling and digging is only likely to be successful in the earliest stages of colonisation. Once Spartina is well established, physical removal is likely to cause considerable environmental disturbance and promote further spread from the establishment of rhizomes/roots released during the removal process. Herbicides have been used to control Spartina in a number of countries but there have been relatively few studies of their effects on estuarine fauna and non-target flora. There are no herbicides for which application to Spartina is an approved on-label use, and use of herbicides in the estuarine environment will require the issuing of permits by the relevant authority on a case-by-case basis. In Tasmania currently the most cost effective and environmentally least damaging herbicide for Spartina control is Fusilade® (the active constituent being fluazifop-P as butyl ester), a selective post-emergence grass killer which does not affect native saltmarsh species or seagrasses, is rapidly degraded and has very low toxicity to estuarine fauna (RGAG 2002). Smothering with black plastic has been successful in eradicating small patches of Spartina in Tasmania (RGAG 2002) but the technique is not practical for treating large-scale infestations. The long-term hope for the control of environmental weeds is the development of biological control. Spartina in Australia is potentially a good candidate for biological control in that there are no closely related native taxa. Research in the USA has identified the plant-hopper Prokelisia marginata as a potential control agenda for introduced Spartina (Daelher and Strong 1997; Wu et al. 1999). The native habitat of P. marginata is on S. alterniflora on the East and Gulf Coasts of America. Before any biocontrol agent could be released in Australia there would be a requirement for a comprehensive risk assessment. In Victoria the ability of S. anglia to tolerate more tidal submergence than native species has permitted its establishment seaward of stands of the mangrove Avicennia marina. The long-term consequences of this are unknown, but Spartina is very effective in promoting sedimentation and so its presence could lead to reduction in sediment supply to the mangrove zone. Spartina spp. utilise the C4 photosynthetic pathway. As carbon dioxide levels in the atmosphere increase, plants with the C4 mechanisms will be relatively disadvantaged compared to C3 species. However, the greater submergence tolerance of S. anglica is likely to permit its continuing invasion of mudflats, although as sedimentation raises the marsh surface, replacement by C3 native species may be enhanced. While there is still the potential for S. anglica to spread to new sites, an increase in geographical distribution is unlikely, not least because at lower latitudes the lower intertidal habitat is pre-empted by more vigorous mangroves than those in Victoria. Juncus acutus Juncus acutus, sharp rush, is a northern hemisphere species, native to Europe, Asia and North America where it occurs in upper saltmarshes and brackish inter-dune wetlands. In parts of its native range, for example in the United Kingdom, it is a relatively rare species of conservation interest. It has been accidently introduced into Australia where it is now a major weed of pasture and coastal saltmarsh (Figure 1.6). In saltmarsh it can form dense monospecific stands, crowding out the native J. kraussii. A number of trials of potential control methods have been conducted in saltmarsh in Homebush Bay (Paul and Young 2006; Paul et al. 2007; see Figure 1.7). Combinations of cutting and herbicide were effective in eradicating J. acutus and regeneration of natives occurred, although re-establishment of J. acutus seedlings and other weeds needed to be prevented.
Australian saltmarshes in global context
Figure 1.6 Distribution of Juncus acutus (from literature records, personal observation and records in the Australian Virtual Herbarium).
Although eradication has been demonstrated in small plots, the practicality and cost-effectiveness of control over whole marshes remains to be demonstrated. Is Phragmites australis a weed in Australian saltmarshes? Phragmites australis, the common reed, is one of the most widely distributed vascular plants in the world, and is certainly native in Australia. It is primarily a freshwater species, but occurs intertidally in the upper reaches of estuaries and, where freshwater discharge occurs, at the higher elevations in saltmarshes. If there is increased freshwater input, for example associated with stormwater drainage, then Phragmites, and other species such as Bolboschoenus spp. Schoenoplectus spp. and Typha spp. (Zedler et al. 1990) can spread into saltmarsh, forming tall dense stands shading out lower vegetation. In the United States there has been a dramatic spread of P. australis in saltmarshes in recent decades. In part this has been facilitated by clearing and development in the hinterland of marshes, with concomitant increase in nutrient and freshwater inputs. However, the invasive P. australis has been shown to be an exotic European genotype, rather than a native American form (Saltonstall 2002). It has been suggested that P. australis may be spreading in Australian saltmarshes, but there are very few monitoring data to confirm this speculation, nor is it known whether there are non-indigenous ecotypes in Australia. The other introduced species – sleepers or no problem? The examples discussed above are only a few of the many introduced species to be found on Australian saltmarshes. Some of the others are clearly becoming major threats such as groundsel bush Baccharis halimifolia which from its initial introduction in Queensland is now spreading south on the NSW coast. There are active control measures applied because B. halimifolia is implicated as a serious allergy-causing plant for humans. Most of the other species are relative small plants, often annual. They may be locally abundant, but while detracting from the ‘naturalness’ of vegetation are generally viewed as unlikely threats to ecosystem integrity. The majority of these species are found in the upper marsh, often associated with areas disturbed by foot or vehicle passage. Not only are the species not seen as a threat, there are no obvious techniques available for control even if this were thought desirable. Many of the worst environmental weeds in Australia were for long periods (decades) ‘sleepers’ – present in vegetation but not becoming dominant or appearing aggressive invaders
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Figure 1.7 Control of Juncus acutus at Homebush Bay (a) infestation and (b) mounds left following removal.
until they suddenly underwent a population explosion and became recognised as serious threats. Unfortunately it has not been possible to predict which amongst the very large number of introduced species will become the major problems of the future. It is possible that amongst species already present in saltmarshes are some ‘sleepers’, which means that even if control programs are not practical at present we should nevertheless be monitoring changes in the composition and abundance of the weed flora. One species which may be showing signs of making the transition from being benign to being a problem is Aster subulatus, a North American species which has been present in NSW saltmarshes for many decades. Although almost ubiquitous it is rarely abundant, but Keith et al. (2007) recently published a photograph of a dense stand of A. subulatus on a saltmarsh on the south coast of NSW; perhaps a sign of what is to come.
Australian saltmarshes in global context
We also need to reconsider the perception that introduced species, which, while numerically abundant contribute relatively little to stand biomass, are ecologically benign. The largest numbers of ‘minor’ introduced species are found in saltmarsh in the Mediterranean climate zone, and many of these species are also found as weeds in Californian saltmarshes. Recent investigation in California show that there may be adverse interactions between native species and exotics – for example the introduced annual grass Polypogon monspeliensis (also present in Australia) out-competes the native Salicornia virginica (correctly Sarcocornia virginica, Callaway and Zedler 1997). Control of P. monspeliensis invasions requires addressing the environmental changes which facilitate invasion, rather than only the symptom. Callaway and Zedler (1998) showed that alteration of hydrologic regimes, affecting both soil moisture status and salinity, have created conditions favourable for exotic plants, so that controlling existing invasive populations, and reducing the potential for future invasions, will require landscapescale restoration of hydrology. The practicality of this approach in the Australian context remains to be determined, but in order to justify the need, ecological studies to examine interactions between native and introduced species are required.
Discussion Australian saltmarshes resemble those elsewhere in the world in terms of general appearance and physiognomy of the dominant plants. There are fewer data on which to establish whether, faunistically, Australian saltmarshes fit a general global model. Nevertheless, there are some distinctive Australian elements in the flora – Tecticornia spp., Wilsonia spp. – and a number of wider Gondwana links which differentiate Australian saltmarshes from those in the northern hemisphere. Within Australia there is regional differentiation of saltmarsh flora and vegetation which is correlated with climate, and a particularly important feature is the extent of tropical saltmarshes. These have been little studied, but given that much of the tropical coast of Australia is still undeveloped we have unique opportunities and responsibilities to conserve large areas of tropical saltmarshes in association with mangroves and hypersaline flats. Globally saltmarshes are facing a range of threats (Adam 2002) and Australian saltmarshes are not immune from these threats, at both the local, site-specific level and more widely, changes in relative sea level will affect all saltmarshes over time. There is an active need for conservation management now, even if the impacts of some threats will not be apparent for decades hence. The listing of saltmarsh in NSW as an endangered ecological community provides an incentive, at least in that state, for the development of conservation plans. In other states, but most notably in Queensland through provisions of fisheries legislation, saltmarsh is also accorded high conservation status. Many of the saltmarshes in Australia are publicly owned, but despite this there is often no active recognition of this by a relevant management authority. Among the public lands, saltmarsh is frequently an orphan, so a major challenge is to develop a culture in which public agencies are more pro-active in saltmarsh management. While recognition of the importance of saltmarsh can be used reactively to modify or reject proposals which would destroy saltmarsh, our ability to develop proactive management is limited by lack of information. There is still much to be done to document the occurrence and composition of saltmarsh, and we have scarcely begun studies on ecological processes and functions in Australian saltmarshes. The detailed study of the Towra Point saltmarsh by Clarke and Hannon (1967, 1969, 1970, 1971) was, at its time, one of the most detailed in the world, but since then there has been little follow up in Australia. Many aspects of the biology of saltmarsh plants, such as population dynamics, pollination, regeneration biology and responses to
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particular threats have scarcely been investigated. Without information, there will be a very uncertain foundation on which to plan for the future
References Adam P (1981a). Australian saltmarshes. Wetlands (Australia) 1, 8–10. Adam P (1981b). Saltmarsh plants of NSW. Wetlands (Australia) 1, 11–19. Adam P (1990). Saltmarsh Ecology. Cambridge University Press: Cambridge. Adam P (1994). Saltmarsh and mangrove. In Australian Vegetation, 2nd edn. (Ed. RH Groves) pp. 395–435. Cambridge University Press: Cambridge. Adam P (2000). Morecambe Bay saltmarshes: 25 years of change. In British Saltmarshes. (Eds BR Sherwood, BG Gardiner and T Harris) pp. 81–107. Linnean Society of London, Forrest Text: Cardigan, UK. Adam P (2002). Saltmarshes in a time of change. Environmental Conservation 29, 39–61. Adam P and King RJ (1990). Ecology of unconsolidated shores. In Biology of Marine Plants. (Eds MN Clayton and RJ King) pp. 296–309. Longman Cheshire: Melbourne. Adam P, Wilson NC and Huntley B (1988). The phytosociology of coastal saltmarsh vegetation in New South Wales. Wetlands (Australia) 7, 35–57. Ayres DR, Smith DL, Zaremba K, Klohr S and Strong DR (2004). Spread of exotic cordgrasses and hybrids (Spartina spp.) in the tidal marshes of San Francisco Bay, California, USA. Biological Invasions 6, 221–231. Baker JM, Adam P and Gilfi llan E (1994). Biological Impacts of Oil Pollution: Saltmarshes. International Petroleum Industry Environmental Conservation Association: London. Barbour MG (1970). Is any angiosperm an obligate halophyte? American Midland Naturalist 84, 105–120. Beeftink WG, Nieuwenhuize J, Stoeppler M and Mohl C (1982). Heavy-metal accumulation in salt marshes from the western and eastern Scheldt. Science of the Total Environment 25, 199–223. Boston KG (1981). The introduction of Spartina townsendii (s.l.) to Australia. Melbourne State College: Occasional Papers No. 6, 1–57. Brearley A (2005). Ernest Hodgkin’s Swanland: Estuaries and Coastal Lagoons of Southwestern Australia. University of Western Australia Press: Crawley. Bridgewater PB (1982). Phytosociology of coastal salt-marshes in the mediterranean climatic region of Australia. Phytocoenologia 10, 257–296. Bridgewater PB and Cresswell ID (1993). Phytosociology and phytogeography of Western Australian salt marshes. Fragmenta Floristica et Geobotanica Supplementum 2, 609–629. Bridgewater PB and Cresswell ID (2003). Identifying biogeographic patterns in Australian saltmarsh and mangal systems: a phytogeographic analysis. Phytocoenologia 33, 231–250. Bridgewater PB, Rosser C and de Corona A (1981). The Saltmarsh Plants of Southern Australia. Monash University Botany Department: Melbourne. Callaway JC and Zedler JB (1997). Interactions between a salt marsh native perennial (Salicornia virginica) and an exotic annual (Polypogon monspeliensis) under varied salinity and hydroperiod. Wetlands Ecology and Management 5, 179–194. Carter S (1994). Coastal Resource Atlas for Oil Spills from Barrenjoey Head to Bellambi Point. Environment Protection Authority of New South Wales: Chatswood. Clarke LD and Hannon NJ (1967). The mangrove swamp and salt marsh communities of the Sydney district: I. Vegetation, soils and climate. Journal of Ecology 55, 753–771. Clarke LD and Hannon NJ (1969). The mangrove swamp and salt marsh communities of the Sydney district: II. The holocoenotic complex with particular reference to physiography. Journal of Ecology 57, 213–234.
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Clarke LD and Hannon NJ (1970). The mangrove swamp and salt marsh communities of the Sydney district: III. Plant growth in relation to salinity and waterlogging. Journal of Ecology 58, 351–369. Clarke LD and Hannon NJ (1971). The mangrove swamp and salt marsh communities of the Sydney district: IV. The significance of species interaction. Journal of Ecology 59, 535–553. Clarke PJ and Jacoby CA (1994). Biomass and above ground productivity of salt-marsh plants in south-eastern Australia. Australian Journal of Marine and Freshwater Research 45, 1521–1528. Congdon RA and McComb AJ (1980). Productivity and nutrient content of Juncus kraussii in an estuarine marsh in south-western Australia. Australian Journal of Ecology 5, 221–234. Craig GF (1983). Pilbara Coastal Flora. Western Australian Department of Agriculture: Perth. Cresswell ID and Bridgewater PB (1998). Major plant communities of coastal saltmarsh vegetation in Western Australia. In Wetlands for the Future. (Eds AJ McComb and JA Davis) pp. 297–326. Gleneagles Publishing: South Australia. Culbertson JB, Valiela I, Peacock EE, Reddy CM, Carter A and VanderKruik R (2007). Longterm biological effects of petroleum residues on fiddler crabs in salt marshes. Marine Pollution Bulletin 54, 955–962. Daehler CC and Strong DR (1997). Reduced herbivore resistance in introduced smooth cordgrass (Spartina alterniflora) after a century of herbivore-free growth. Oecologia 110, 99–108. Davy AJ, Costa CSB, Yallop AR, Proudfoot AM and Mohamed MF (2000). Biotic interactions in plant communities of saltmarshes. In British Saltmarshes. (Eds BR Sherwood, BG Gardiner and T Harris) pp. 109–127. Linnean Society of London, Forrest Text: Cardigan, UK. Drake BG, Leadley PW, Arp WJ, Nassiry D and Curtis PS (1989). An open top chamber for field studies of elevated atmospheric CO2 concentration on saltmarsh vegetation. Functional Ecology 3, 363–371. Fairweather PG (1990). Ecological changes due to our use of the coast: research needs versus effort. Proceedings of the Ecological Society of Australia 16, 71–77. Hamilton AA (1919) An ecological study of the salt marsh vegetation of the Port Jackson district. Proceedings of the Linnean Society of N.S.W 44, 463–513. Hubbard CE (1968). Grasses. Penguin: Harmondsworth. Johns L (2006). Field Guide to Common Saltmarsh Plants of Queensland. DPI&F Publication: Brisbane. Keith DA, Simpson C, Tozer MG and Rodoreda S (2007). Contemporary and historical descriptions of the vegetation of Brundee and Saltwater Swamps on the lower Shoalhaven River floodplain, southeastern Australia. Proceedings of the Linnean Society of N.S.W. 128, 123–154. Kelleway J (2005). Ecological impacts of recreational vehicle use on saltmarshes of the Georges River, Sydney. Wetlands (Australia) 22, 52–66. Kelleway J, Williams RJ and Allen CB (2007). An Assessment of the Saltmarsh of the Parramatta River and Sydney Harbour. NSW Department of Primary Industries: Cronulla. King RJ, Adam P and Kuo J (1990). Seagrasses, mangroves and saltmarsh plants. In Biology of Marine Plants. (Eds MN Clayton and RJ King) pp. 213–240. Longman Cheshire: Melbourne. Kirkpatrick JB and Glasby J (1981). Salt Marshes in Tasmania: Distribution, Community Composition and Conservation. Occasional Paper No. 8. Department of Geography, University of Tasmania: Hobart. Kloot PM (1984). The introduced elements of the flora of Southern Australia. Journal of Biogeography 11, 63–78. Marchant CJ (1963). Corrected chromosome numbers for Spartina × townsendii and its parent species. Nature 199, 929.
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Marchant CJ (1968). Evolution in Spartina(Gramineae). II. Chromosomes, basic relationships and the problem of S. x townsendii agg. Journal of Linnean Society (Botany) 60, 381–409. Mason CF, Underwood GJC, Baker NR, Davey PA, Davidson I, Hanlon A, Long SP, Oxborough K, Paterson DM and Watson A (2003). The role of herbicides in the erosion of salt marshes in eastern England. Environmental Pollution 122, 41–49. Mooney HA and Hobbs RJ (2000). Global change and invasive species:Where do we go from here. In Invasive Species in a Changing World. (Eds HA Mooney and RJ Hobbs) pp. 425– 434. Island Press: Washington, DC. Mueller F (1868). Fragmenta Phytographiae Australiae 6. Auctoritate Guberni Coliniae Victoriae, ex officina Joannis Ferres: Melbourne. Paul S and Young R (2006). Experimental control of exotic spiny rush, Juncus acutus from Sydney Olympic Park: I. Juncus mortality and re-growth. Wetlands (Australia) 23, 1–13. Paul S, Young R and MacKay A (2007). Experimental control of exotic Spiny Rush, Juncus acutus from Sydney Olympic Park: II. Effects of treatments on other vegetation. Wetlands (Australia) 24, 1–13. Pidgeon I (1940). The ecology of the central coastal area of New South Wales. III. Types of primary succession. Proceedings of the Linnean Society of N.S.W 65, 221–249. Pringle AW (1993). Spartina anglica colonisation and physical effects in the Tamar estuary, Tasmania 1971–91. Papers and Proceedings of the Royal Society of Tasmania 127, 1–10. Ranwell DS (1967). World resources of Spartina townsendii (sensu lato) and economic use of Spartina marshland. Journal of Applied Ecology 4, 239–256. RGAG (2002). Strategy for the Management of Rice Grass (Spartina anglica) in Tasmania, Australia. Department of Primary Industries, Water and Environment: Tasmania. Ridd P, Sandstrom MW and Wolanski E (1988). Outwelling from tropical tidal salt flats. Estuarine Coastal and Shelf Science 26, 243–253. Saenger P, Specht MM, Specht RL and Chapman VJ (1977). Mangal and coastal saltmarsh communities in Australasia. In Ecosystems of the Worlds: Wet Coastal Ecosystems. (Ed. VJ Chapman) pp. 293–345. Elsevier: Amsterdam. Saltonstall K (2002). Cryptic invasion by a non-native genotype of the common reed. Phragmites australis in North America. Proceedings of the National Academy of Sciences 99, 2445–2449. Shepherd KA and Wilson PC (2007). Incorporation of the Australian genera Halosarcia, Pachycornia, Sclerostegia and Tegicornia into Tecticornia (Salicornioideae, Chenopodiaceae) Australian Systematic Botany 20, 319–331. Smith-White AR (1981). Physiological differentiation in a salt-marsh grass. Wetlands (Australia) 1, 20–21. Smith-White AR (1988). Sporobolus virginicus (L.) Kunth in coastal Australia: the reproductive behaviour and the distribution of morphological types and chromosome races. Australian Journal of Botany 36, 23–39. Snogerup S (1993). A revision of Juncus Subgen. Juncus (Juncaceae). Willdenowia 23, 23–73. Thannheiser D (2001). Studien zur Küstenvegetation Victorias (Südaustralien). Bamberger Geographische Schriften 20, 271–285. Valiela I (2006). Global Coastal Change. Blackwell Publishing: Carlton. Williams TP, Bubb JM and Lester JN (1994). Metal accumulation within salt marsh environments: A review. Marine Pollution Bulletin 28, 277–290. Woods JLD, Brown TH, Gangaiya P and Morrison RJ (2007). Water quality in Tom Thumb Lagoon, a highly disturbed urban estuary in Port Kembla, New South Wales, Australia. Wetlands (Australia) 24, 44–66.
Australian saltmarshes in global context
Wu M-Y, Hacker S, Ayres D and Strong DR (1999). Potential of Prokelisia spp. as biological control agents of English Cordgrass, Spartina anglica. Biological Control 16, 267–273. Zedler JB, Nelson P and Adam P (1995). Plant community organization in New South Wales saltmarshes: Species mosaics and potential causes. Wetlands (Australia) 14, 1–18. Zedler JB, Paling E and McComb A (1990). Differential responses to salinity help explain the replacement of native Juncus kraussii by Typha orientalis in Western Australian salt marshes. Australian Journal of Ecology 15, 57–72.
21
CHAPTER 2
Distribution of Australian saltmarsh plants Neil Saintilan
Characteristics of saltmarshes and the saltmarsh environment Coastal (intertidal) saltmarsh has been defined by Adam (1996) as an intertidal plant community dominated by herbs and low shrubs. The plants that comprise coastal saltmarsh are not exclusively intertidal, a characteristic that sets saltmarsh plants apart from mangrove. Most saltmarsh species take advantage of inland saline environments, both aquatic and arid, and their intertidal occurrence is sometimes best understood as a continuation of a network of saline aquatic environments. This is particularly true of the South Australian saltmarsh flora. The characteristics which suit plants to this harsh environment are a tolerance of extreme ranges of salinity and soil water content. The periodic inundation of the intertidal environment by seawater salinises the soil, and salts will subsequently concentrate by processes of evaporation and transpiration. While this phenomenon occurs the world over, in Australia the range of salinity encountered is enhanced by the extremes of climatic variability and, in the tropics, seasonality (Adam 1996). Plants occupying the saltmarsh must be able to withstand periodic soil salinity and inundation. There is a range of strategies amongst the 100 or more species found in Australian saltmarshes. The adaptations to saline conditions are often at the expense of growth rate, and it is this that explains the narrow penetration of saltmarsh plants into upslope freshwater terrestrial environments (Adam 1990). Two-thirds of the 103 Australian saltmarsh plant species listed in Appendix 2.1 belong to five families; Chenopodiaceae; Poaceae; Cyperaceae, Aizoaceae and Asteraceae. While there is a high degree of endemism at a species level amongst Australian saltmarshes, at a generic and family level there are numerous commonalities with the saltmarshes of other continents (Adam 1990; Chapter 1, this volume). In this sense the saltmarshes of Australia are not as distinctly Australian as the terrestrial vegetation communities, though there are unique features of the Australian saltmarsh ecologically.
Description of common saltmarsh plants For those beginning their interest in coastal saltmarsh, there are some excellent keys and field guides available, notably: ●
Field Guide to Common Saltmarsh Plants of Queensland (2006) by Louise Johns, and available through the Queensland Department of Primary Industries. 23
24
Australian Saltmarsh Ecology
●
●
Saltmarsh Plants of New South Wales, by Paul Adam and illustrated by John Barclay. Wetlands (Australia) 1, 11–19. The Saltmarsh Plants of Southern Australia (1981) by P. Bridgewater, C. Rosser and A. de Corona, Monash University, Melbourne.
While numerous plant species can be found in the Australian saltmarsh, particularly at the landward fringe, only a few species dominate. These species are likely to be the most significant ecologically, and their characteristics are described briefly in the following section. Plate 2.1 on pages 45–46 displays colour photographs of these species. Sporobolus virginicus The salt couch Sporobolus virginicus is the most widely distributed saltmarsh plant in Australia, occurring in 33 of the 36 coastal bioregions. Its seeds are primarily airborne, though can be dispersed by water (Naidoo and Naidoo 1992). On the east Australian coast, the species increases in prevalence northward, and is the dominant saltmarsh plant in south-east Queensland. It is known to be tolerant of waterlogged acidic soils (Naidoo and Naidoo 1992) and grows particularly well in sandy locations (Johns 2006). Sarcocornia quinqueflora The family Chenopodiaceae contributes more species to the Australian saltmarsh than any other, and the ‘samphire’ Sarcocornia quinqueflora (beaded glasswort) is the most widely distributed member of the Chenopodiaceae. The plant forms a creeping mat, and spreads primarily through vegetative propagation, though the seeds can be tidally dispersed. The colour of the plant ranges from green to red and purple and may change with environmental conditions. This low-growing plant is found in the wetter parts of the saltmarsh zone, often referred to as representative of the ‘wet’ or lower saltmarsh zone. The species is the dominant saltmarsh plant in southern and central New South Wales, and is found throughout the Australian coastline with the exception of the Northern Territory and the northern half of Western Australia. Juncus kraussii The rush Juncus kraussii grows in fresher conditions than Sporobolus virginicus and Sarcocornia quinqueflora, and may form a landward fringe to the saltmarsh, or a dominant species on upstream saltmarshes within estuaries. The plant forms thick stands generally less than a metre high, though taller stands of up to 2 m have been observed (e.g. Belongil Creek, NSW, Nick Wilson pers. comm.). The species can withstand several months continuous inundation around the margins of lagoons (Adam 1981). It is a plant of the southern Australian saltmarshes and has not been found growing north of 20°S latitude. Samolus repens The creeping brookweed is a low-growing herb that rarely forms a dominant stand, though is commonly found throughout its geographic range, which extends through the southern half of the continent from 23°S latitude. S. repens produces small attractive flowers between September and March which may be white or pink. A more upright form of the genus, Samolus junceus, is found in Western Australia. Suaeda australis Another chenopod, Suaeda australis is a small, woody upright perennial herb, a taller plant than Sarcocornia, with succulent leaves approximately 5 cm long. Commonly known as seablite, S. australis is common on the Australian east coast extending from Cape York Peninsula south
Distribution of Australian saltmarsh plants
to Tasmania, throughout coastal South Australia and the south-western corner of Western Australia. Though some records of the species have been reported from the Northern Territory, it is not common there. S. australis prefers somewhat drier, better-drained conditions than S. quinqueflora, though it relies on water for seed dispersal (Clarke and Hannon 1970). Tecticornia pergranulata Following a taxonomic review by Shepherd and Wilson (2007) the genera Halosarcia and Sclerostegia were incorporated into the genus Tecticornia. The Chenopod genus Tecticornia contributes over a dozen species to the Australian saltmarsh. The succulent stems are similar in appearance to S. quinqueflora, though Tecticornia grows predominantly as a shrub, up to a metre high. T. pergranulata is as common in terrestrial saline environments as on the coast. The species is found in all Australian mainland jurisdictions, though is rare in NSW with a single population in Homebush Bay the only recorded occurrence for the state. Triglochin striata Triglochin striata, the three-ribbed or streaked arrowgrass, is common in less well drained depressions on the saltmarsh plain. The leaves are erect, often in groups of three, and may be up to 30 cm long, though more commonly less than 10 cm. The plant is widely distributed, both in Australia and in other southern continents. Gahnia filum Gahnia filum grows as a clumped sedge, slightly taller than Juncus krausii (approximately 1.5 m) and with a long trailing edge to their slender shoots. Gahnia filum is found almost exclusively in the south-east corner of the continent, including South Australia, Victoria and Tasmania, with some occurrences on the NSW south coast north to the Georges River, where it appears to be spreading (Adam pers. comm.). In South Australia, the plant is the primary habitat of the Yellowish Sedge-skipper Butterfly (Hesperilla flavescens flavia) (Coleman and Coleman 2000).
Saltmarsh structural forms and zonation In their detailed consideration of the distribution and ecology of Tasmanian saltmarshes, Kirkpatrick and Glasby (1981) define a series of saltmarsh structural forms which could validly be applied more widely: 1. Communities dominated by succulent shrubs (e.g. the genera Tecticornia). 2. Communities dominated by grasses (e.g. Sporobolus virginicus, Stipa stipoides, Zoysia macrantha). 3. Communities dominated by sedges and grasses (e.g. Juncus krausii, Gahnia filum). 4. Communities dominated by herbs (low-growing creeping plants such as Wilsonia backhousei, Samolus repens, Schoenus nitens). The distribution of these forms varies across the intertidal zone. Within New South Wales, the lower intertidal is dominated by herbs and grasses which gives way to sedges and rushes in the landward sections of the intertidal zone. Within Victorian saltmarshes, the lower saltmarsh zone is dominated by succulent shrubs of the genera Tecticornia and Sarcocornia. The herbs and grasses are more commonly found in a landward, upper-intertidal zones which are also the most species diverse (see, for example, Schindl 2002). Coleman (2005) described plants characteristic of four elevation zones within the saltmarshes of South Australia. A low marsh community is characterised by Suaeda australis and Sarcocornia quinqueflora. Species characteristic of the mid-marsh are Frankenia pauciflora and species of the
25
26
Australian Saltmarsh Ecology
genus Tecticornia. The high marsh is characterised by a diverse array of species including Mimulus repens (in brackish areas), Puccinellia stricta, Wilsonia humilis, Apium annuum, Samolus repens, Disphyma crassifolium, Spergularia spp., Atriplex semibaccata and Trigoichin striata. A landward community of saltmarsh plants, above the level of normal spring high tide inundation, includes Nitraria billardierei, Distichlis distichophylla and Dianella brevicaulis. Most studies have indicated that a combination of moisture content and salinity explain the distribution of vegetation communities within the saltmarsh. Soil moisture content decreases between the mangrove and terrestrial environments. Soil chlorinity varies less predictably, and will respond to micro-scale hydraulic controls (such as evaporative depressions), as well as plant activity (accumulating salts within the root zone). Temporal variability in salinity may also be high, and related to rainfall, groundwater discharge, and the periodicity of the tides. On the New South Wales coast, for example, spring tides reach their maximum inundation in summer (daylight hours) and winter (night), which are also the times of highest variability in tide height. The periods which inundate the upper-intertidal are also those which least frequently inundate the lower saltmarsh. In summer, the additional evaporative losses resulting from higher temperatures make this the period of maximum soil salinity in the saltmarsh (Clarke and Hannon 1969).
Continental distribution of saltmarsh plants Coastal specialists versus generalists The majority (over 90%) of saltmarsh species are generalists, distributed across a range of saline aquatic habitats both coastal and inland. On the basis of records held within the Australian Virtual Herbarium, there are a number of species which are coastal specialists, though no distinction is made between coastal saltmarsh and coastal dune and headland environments. These species are: Batis argillicola; Baumea teretifolia; Carpobrotus glaucescens; Limonium solanderi; Limonium australe; Suaeda arbusculoides; Sesuvium portulacastrum and Austratipa stipoides (selected examples in Figure 2.1). Several species (over 20) are found predominantly, though not exclusively, in coastal environments. Common species in this category are: Sarcocornia quinqueflora; Sporobolus virginicus; Zoysia macrantha; Triglochin striata; Suaeda australis; Juncus kraussi;, Isolepis nodosa; Selliera radicans; Gahnia filum and Fimbristylis ferruginea. A third category of saltmarsh plants are those where there is no obvious coastal preference to their distribution. Plants loosely described as shrubs appear to fall into this category, including all of the species of Tecticornia and Maireana found in the saltmarsh. Latitudinal patterns of diversity The inverse relationship between saltmarsh species diversity and latitude, mentioned by several authors (Saenger et al. 1977; Specht 1981; Adam et al. 1988) is confirmed in an analysis of the data of Appendix 2.1 (Figure 2.4). The relationship is particularly strong (r2 = 0.64), and contrasts to the trend for mangroves of decreasing diversity with increasing latitude. Two outliers in Figure 2.4 exhibit low diversity in spite of high latitude, and in both cases this is due to rockdominated coastlines (the Gawler and Tasmanian West bioregions). There is little sediment yield from the few rivers that reach the coast in these bioregions. The bioregion is a useful scale of analysis for conservation planning, and Table 2.1 represents the proportion of the total saltmarsh flora present in each of the 36 coastal bioregions. Centres of biodiversity emerge, with the South Australian bioregions particularly diverse. Three-quarters of the 103 listed saltmarsh species can be found within 200 km of Adelaide.
Distribution of Australian saltmarsh plants
(a)
(b)
(c)
(d)
Figure 2.1 Distribution of confirmed records of two coastal specialists (Limonium solanderi (a) and Stipa stipoides (b)) and two generalists (Dissocarpus biflorus (c) and Atriplex semibaccata (d)). In the latter cases the disjunct coastal extent is explained by connection through interior saline lakes.
Fischer (1960) described the correlation between floral and faunal diversity and decreasing latitude as one of the most imposing biogeographic features on the planet. This is a trend to which Australian mangroves conform (Saenger et al. 1977; Wells 1983) but clearly not saltmarsh. Variations in intertidal extent between bioregions provide no explanation; indeed, there is an inverse relationship between intertidal extent and saltmarsh diversity in Australia (Figure 2.5). Northern Australia supports a low diversity of saltmarsh in spite of the large intertidal area available to saltmarsh colonisation (Table 2.2 below). To some degree this may be explained by climatic constrains, with hypersaline conditions developing in upper-intertidal elevations during the dry season in many parts of northern Australia. Wide intertidal areas will be devoid of saltmarsh in these situations. Notwithstanding, the four southern States of Tasmania, Victoria, South Australia and New South Wales contain less than 2.5% of the total saltmarsh/saltpan area yet together support 90% of the saltmarsh flora of Australia. The inability of many saltmarsh species to colonise the intertidal flats of tropical Australia is most probably related to an intolerance of higher temperatures, or a combination of higher temperatures and seasonally higher salinities, which appears to inhibit the germination of
27
Australian Saltmarsh Ecology
Box 2.1
Sources of data and methods of analysis
All confirmed records of species occurrences in Australia are now represented in the Australian Virtual Herbarium, an interactive internet-based tool which can be accessed through the websites of participating herbaria. Bridgewater and Cresswell (2003) used a prototype version of this tool, along with published records and personal observations, to define the distribution of saltmarsh species groups throughout the continent. The Interim Bioregionalisation of Australia (IBRA) was used to provide the fundamental spatial unit for this analysis. Their work presented the first systematic presentation of saltmarsh species assemblage distribution in Australia, and is described later in this chapter. Appendix 2.1 of this chapter is a compilation of the known distribution of Australian saltmarsh plants, utilising the completed records for 103 saltmarsh species now found within the Australian Virtual Herbarium. This list was compiled primarily on the basis of species lists in Adam (1981), Adam et al. (1988), Kirkpatrick and Glasby (1981), Johns (2006), Bridgewater and Cresswell (2003). Some adjustments were made where published accounts expanded the range of some species including, for example, the detailed biogeographic work of Kirkpatrick and Glasby (1981) for Tasmanian saltmarshes. Multivariate statistical analyses, such as cluster analysis, can be used to provide an a priori grouping of bioregions in terms of saltmarsh species occurrence (Saintilan 2009). Such an analysis is presented in Figure 2.2. This cluster dendogram represents the similarity or dissimilarity of bioregions in presence and absence of saltmarsh species (data from Appendix 2.1). The Bray-Curtis 20
40
60
80
100
3 4 5 1 2 35 34 32 33 31 36 28 29 30 14 6 7 8 9 17 21 18 19 10 11 12 13 15 16 25 27 24 26 20 22 23
Bray-Curtis Similarity index
28
Northern Humid East
North
Southern Arid West
Humid East
South
Arid West
Figure 2.2 Clustering of IBRA bioregions on the basis of saltmarsh floristic composition. At the highest level of dissimilarity, the continent can be divided north–south along 23° latitude. These groups may be further divided by coastal orientation.
Distribution of Australian saltmarsh plants
similarity index ranges from zero (no species in common) to 100 (all species in common). The broader divisions on the tree are therefore between groups of sites with a lower proportion of species in common (for example 25% is the first major division). Higher-order branchlets represent similar bioregions in the overall ‘assemblage’ of saltmarsh plants present (up to 90% of species in common in some cases). The abundance of saltmarsh plants (species dominance) in each bioregion was not considered in this analysis. When applied to the Virtual Herbarium data, this divides the continent into a northern and southern region, with a secondary division along coastal orientation (Figure 2.3). Bridgewater and Cresswell (2003) present an alternative analysis, discussed later in this chapter. Coastal Bioregions
34
35
33 31
NW
2
N
32
36
1
3
30
NE 4
29
5
28
6
27 26
23
SW 25
24
20
22
19
S 0
250
500
1000
7 8
21
SE
18 17
9 16
10 15
11
Kilometers
14
12 13
1. Gulf Plains 2. Cape York Peninsula 3. Wet Tropics 4. Brigalow Belt North 5. Central Mackay Coast 6. South East Queensland 7. NSW North Coast 8. Sydney Basin 9. South East Corner 10. South East Coastal Plain 11. Flinders 12. Tasmanian South East 13. Tasmanian Southern Ranges 14. Tasmanian West 15. Tasmanian Northern Slopes 16. King 17. Naracoorte Coastal Plain 18. Murray Darling Depression 19. Kanmantoo 20. Gawler 21. Eyre Yorke Block 22. Nullabor 23. Hampton 24. Esperence 25. Warren 26. Swan Coastal Plain 27. Geraldton Sandplains 28. Carnarvon 29. Pilbara 30. Damperland 31. Northern Kimberley 32. Victoria Bonaparte 33. Darwin Coastal 34. Tiwi Cobourg 35. Arnhem Coast 36. Gulf Coastal
Figure 2.3 Coastal bioregions included in the analysis, with the saltmarsh biogeographic provinces identified by cluster analysis.
some species (Greenwood and MacFarlane 2006). One potential impact of global warming may therefore be a decline in diversity of Australian saltmarsh flora within many bioregions of southern Australia. Clustering of bioregions The northern and southern divisions The primary division is between a northern and southern saltmarsh flora, separated at 23°S, defining a northern division extending from the Central Mackay Coast to the Carnarvon bioregion on the West Coast (see Figure 2.3). The two halves of the content have less than 25% of species in common. Species occurring primarily in the northern bioregions include: Tecticornia australasica; Fimbristylis polytrichoides; Portulaca bicolor; Fimbristylis ferruginea; Batis agrillicola; Xerochloa
29
Australian Saltmarsh Ecology
90 80
Saltmarsh diversity
30
70 60 50 40 30 20 10 0 0
5
10
15
20
25
30
35
40
45
50
Latitude Figure 2.4 Relationship between saltmarsh diversity and latitude for the 36 Australian coastal bioregions. (Saltmarsh plants listed in Appendix 2.1) (r2 = 0.64, P<0.001).
imberis and Sesuvium portulacastrum. Species occurring primarily in the southern bioregions include: Samolus repens; Atriplex cinerea; Apium prostratum; Myoporum insulare and Cotula coronifolia (see Figure 2.6 for examples). Some genera do not show strong overlap in the distribution of member species. For example, the two species of Carpobrotus show mutually exclusive ranges, with C. rossii and C. glaucescens occupying the southern and eastern coastlines respectively. Often, though, and particularly amongst sedges, species within a genus show strongly sympatric ranges, as illustrated in Figure 2.7. Table 2.1
Percentage of the total saltmarsh flora of Appendix 2.1 present in each bioregion.
Bioregion
%
Bioregion
%
Bioregion
%
Gulf Plains
20
Tasm. Southrn Ranges
46
Warren
36
Cape York Peninsula
20
Tasmanian West
33
Swan Coastal Plain
48
Wet Tropics
24
Tasm. Northn. Slopes
43
Geralton Sandplains
49
Brigalow Belt North
26
King
49
Carnarvon
27
Central Mackay Coast
32
Naracoote Coastal Plain
61
Pilbara
19
South-east Queensland
45
Murray Darling Depression
66
Damperland
17
NSW North Coast
39
Kanmantoo
72
Northern Kimberley
16
Sydney Basin
49
Gawler
37
Victoria Bonaparte
14
South-east Corner
50
Eyre Yorke Block
64
Darwin Coastal
12
SE Coastal Plain
59
Nullabor
32
Tiwi Cobourg
13
Flinders
55
Hampton
19
Arnhem Coast
13
Tasmanian South-east
58
Esperence
48
Gulf Coastal
13
Distribution of Australian saltmarsh plants
Table 2.2 The percentage of national saltmarsh area occurring in each state (from Bucher and Saenger 1991), and the percentage of national saltmarsh flora occurring in each state (data from Appendix 2.1). Jurisdiction
Percentage of saltmarsh area
Percentage of saltmarsh flora
New South Wales
0.42
45
Victoria
0.92
55
Queensland
38.15
29
Western Australia
21.81
72
South Australia
0.62
71
Tasmania
0.27
53
36.82
18
Northern Territory
Subgroups based on coastal orientation The cluster analysis represented in Figure 2.2 suggests a tertiary level of division based on contiguious stretches of coastline demarcated by orientation. These subgroups characteristically share between 70% and 80% of species. In the southern division these may be described as eastern, southern and western assemblages. Examples of primarily eastern species include: B. teretifolia; C. glaucescens; Aster australasica; Fimbristylis ferruginea and Sesuvium portulacastrum. Primarily southern species include: Sarcocornia blackiana; Hemichroa pentandra; Tecticornia arbuscula and Distichlis distichophylla (see Figure 2.8). The northern division can be similarly divided geographically. The western coast has been discussed previously, being arid with few estuaries. The tropical northern segment can be divided into a north-facing coast, though including the eastern section of Cape York Peninsula, with a relatively depauperate saltmarsh flora, though containing a few characteristic species including Xerochloa imberbis (see Figure 2.9). The Queensland tropical east coast supports a higher number of saltmarsh species with Aster subulatus, Baumea juncea and Limonium solanderi becoming important. 80
Saltmarsh species
70 60 50 40 30 20 10
-1
0 0
1
2
3
4
Log(10) saltmarsh area (km2) Figure 2.5 Inverse relationship between saltmarsh diversity and the intertidal area within the coastal bioregions of Australia.
31
32
Australian Saltmarsh Ecology
(a)
(b)
(c)
(d)
(e)
(f)
Figure 2.6 Species characteristic of northern and southern continental divisions (a) Tecticornia australasica; (b) Batis agrillicola; (c) Portulaca bicolor; (d) Samolus repens; (e) Atriplex cinerea; (f) Apium prostratum.
Distribution of Australian saltmarsh plants
(a)
(b)
(c)
(d)
(e)
(f)
Figure 2.7 Confirmed distribution of the common saltmarsh rushes Juncus bufonius (a), J. kraussii (d), Isolepis nodosa (b) and I. cernua (e). The coastal distributions largely sympatric. The tropical sedge Fimbristylis ferruginea (c) and F. polytrichoides (f) are sympatric through the northern division.
33
34
Australian Saltmarsh Ecology
(a)
(b)
(c)
(d)
(e)
(f)
Figure 2.8 Species showing distributions confined to southern and eastern coastal segments, primarily within the southern division (a) Baumea teretifolia; (b) Carpobrotus glaucescens; (c) Hydrocotyle bonariensis; (d) Hemichroa pentandra; (e) Carpobrotus rossii; (f) Distichlis distichopylla.
Distribution of Australian saltmarsh plants
Figure 2.9 Examples of species characteristic of the northern and eastern sections of the northern division (a) Xerochloa imberbis; (b) Limonium solanderi.
Biogeographic patterns identified by Bridgewater and Cresswell Bridgewater and Cresswell (2003) identified five coastal/inland saltmarsh phytogeographic groups in their analysis of previous publications and the (then) prototype Australian Virtual Herbarium. Two of these groups, the Sclerostegia (now Tecticornia) tenuis Group and the Halosarcia (now Tecticornia) pergranulata pergranulata Groups were inland associations, being characteristic of the central arid/semi-arid and Murray-Darling basin regions respectively. The remaining three groups covered the major coastal saltmarsh associations and are summarised below. 1. The Sclerostegia (now Tecticornia) arbuscula-Juncus kraussii Group, characteristic of the southern Australian coastline and Tasmania. Species distinguishing this group include: Juncus kraussi;, Selliera radicans; Cotula coronopifolia; Angianthus preissianus; Centrolepis polygyna; Samolus repens; Suaeda australis; Triglichin striatum; Sarcocornia quinqueflora and the introduced Parapholis incurva. This group was further subdivided into five subgroups, again with geographic affiliations: a. a western Tasmanian group comprising: Austrastipa stipoides, Apium prostratum and Atriplex paludosa paludosa. b. an eastern Tasmanian/eastern NSW group consisting of the species of the first subgroup, with the addition of: Limonium australe, Wilsonia backhousei, Lachnagrostis billardieri and Gahnia filum c. a central southern coastal group comprising: Tecticornia halocnemoides-Limonium binervosum, with T. flabelliformis and Maireana oppositifolia, and: d. a south-western coast Tecticornia halocnemoides-Rhagodia baccata subgroup, with Atriplex hypoleuca, Frankenia tetrapetala, T. indica bidens, T. Pterygosperma pterygosperma and Atriplex paludosa. 2. A Western Australian central coast/inland group (corresponding to the arid northern though including the western part of the southern). The species distinguishing this group are largely Chenopod shrubs of the genus Tecticornia, including: Tecticornia doleiformis; T. leptoclada; T. undulata; T. halocnemoides; T. h. catenulate; T. pergranulata pergranulata; T. pterygosperma pterygosperma; T. lylei; T. peltata; Maireana oppositifolia and Tecticornia
35
36
Australian Saltmarsh Ecology
disarticulata. This group was divided into two subgroups, being those of winter-dominated rainfall (Sarcocornia blackiana, T. syncarpa, Triglochin mucronatum, Centrolepis polygyna), and those areas of summer rainfall or highly arid (Tecticornia pruinosa, T. indica leiostachy, Tecticornia calyptrate, T. chartacea, T. pruinosa, Meullerolimon salicorneaceum, Tecticornia arborea and Frankenia hispidula). 3. a group characteristic of the central eastern and northern Australian coastline, termed the Suaeda arbusculoides-Tecticornia indica julacca group. Species distinguishing this group are: Suaeda arbusculoides; Tecticornia indica julacea; T. i. leiostachya; T. pergranulata queenslandica and Tecticornia australasica. The two subgroups of this group include an assemblage confined to the northern and tropical eastern coast (Batis agrillicola, Xerochloa imberbis), and that of the central eastern coastline to the NSW Queensland border (the Juncus kraussii-Suaeda australis subgroup, also characterised by Limonium australe, Sarcocornia quinqueflora and T. pergranulata pergranulata).
Conclusions The increasing saltmarsh species diversity with latitude has been mentioned by several authors. The analysis provided by this chapter provides firm quantitative support, with latitude explaining nearly two-thirds of the variation in saltmarsh species diversity between coastal bioregions. This holds even with the significant variation between bioregions in saltmarsh area, (which is inversely related to latitude), and with three high latitude bioregions virtually devoid of intertidal wetland (Nullabor, Gawler and Tasmanian West). The development of an interim bioregionalisation of Australia is a useful aid to the systematic conservation of Australia’s aquatic resources. The value of this approach is obvious when used as the basis for exploring saltmarsh biogeographic patterns. Clearly, some bioregions are of particular significance in containing a high proportion of the nation’s saltmarsh species, with the South Australian bioregions of Kanmantoo, Murray-Darling Depression and Eyre Yorke Block being good examples, with each containing more then 60% of the 110 saltmarsh species listed in Appendix 2.1. In defining the location of saltmarsh reserves, these would be good places to begin. The highest level of clustering of bioregions on the basis of saltmarsh flora is between a northern and southern division separated at 23° latitude with less than 25% overlap between species. The relatively depauperate northern division corresponds broadly to the Arid Tropical, Dry Tropical, Wet-Dry Tropical, Gulf and Wet Tropical groups of Bridgewater and Cresswell (1999). The diverse southern division corresponds to Bridgewater and Cresswell’s (1999) Warm Temperate, Temperate Tasmania, Mediterranean and Western Transitional groupings. Lower levels of clustering are more equivocal, though particularly arid bioregions (the Arid Tropical region of Bridgewater and Cresswell 1999) and the arid, rock-dominated south Australian bioregions of the Great Australian Bight, separate out from more humid segments of the major divisions. A final level of classification based on coastal orientation is proposed for both divisions. The scheme has strong affinities with that of Bridgewater and Cresswell (2003) though it is not identical. The herbarium and published record data have not to date been analysed on an estuary-byestuary basis. Species lists for Australian estuaries would have several advantages. Firstly, such a list would provide a finer scale of resolution in documenting the occurrence of saltmarsh plants. Further, the lists would provide important guidance to regional NRM bodies, in identifying threatened or endangered species within estuaries and suggesting species lists for appropriate saltmarsh rehabilitation. Little is known of the ecophysiological adaptations and ecological requirements of Australian saltmarsh plants, or of their productivity. Several lines of enquiry are suggested by this
Distribution of Australian saltmarsh plants
chapter. The freshwater and flooding requirement of species, in particular for germination, is an issue of relevance when assessing the impacts of drainage modification and freshwater diversion in estuaries. An understanding of the adaptations of plants to saline conditions has a number of useful adaptations in agriculture, from the rehabilitation of saline lands to the development of salt resistant strains of commercial crops. Australian saltmarsh plants have a valuable role to play in both arenas.
References Adam P (1990). Saltmarsh Ecology. Cambridge University Press: Cambridge. Adam P (1994). Saltmarsh and mangrove. In Australian Vegetation, 2nd edn. (Ed. RH Groves) pp. 395–435. Cambridge University Press: Cambridge. Adam P (1981). Saltmarsh plants of NSW. Wetlands (Australia) 1, 11–19. Adam P (1996). Saltmarsh. In State of the Marine Environment Report for Australia. (Eds LP Zann and P Kailola) pp. 97–105. Department of Environment, Sport and Territories: Canberra. Adam P and Hutchings P (1987). The saltmarshes and mangroves of Jervis Bay. Wetlands (Australia) 6, 58–64. Adam P, Wilson NC and Huntley B (1988). The phytosociology of coastal saltmarsh vegetgation in New South Wales. Wetlands (Australia) 7, 35–57. Backshall DJ and Bridgewater PB (1981). Peripheral vegetation of Peel Inlet and Harvey Estuary, Western Australia. Journal of the Royal Society of Western Australia 4, 5–11. Ball MC and Farquhar GD (1984). Photosynthetic and stomatal responses of two mangroves species, Aegiceras corniculatum and Avicennia marina, to long term salinity and humidity conditions. Plant Physiology 74(1), 1–6. Ball MC (1988). Ecophysiology of mangroves. Trees – Structure and Function 2(3), 129–142. Bucher D and Saenger P (1994). A classification of tropical and subtropical Australian estuaries: Aquatic conservation. Marine and Freshwater Ecosystems 4, 1–19. Bridgewater PB (1982). Phytosociology of coastal saltmarshes in the Mediterranean climatic region of Australia. Phytocoenologia 10, 257–296. Bridgewater PB and Cresswell ID (1999) Biogeography of mangrove and saltmarsh vegetation: implications for conservation and management in Australia. Mangroves and Salt Marshes 3, 117–125. Bridgewater PB (1975). Peripheral vegetation of Westernport Bay. Proceedings of the Royal Society of Victoria 87, 69–78. Bridgewater PB, Rosser C and Corona A (1981). The Saltmarsh Plants of Southern Australia. Monash University Botany Department: Melbourne. Bridgewater P and Cresswell ID (2003). Identifying biogeographic patterns in Australian saltmarsh and mangal systems: a phytogeographic analysis. Photocoenologia 33(2–3), 231–250. Bucher D and Saenger P (1991). An inventory of Australian estuaries and enclosed marine waters: an overview of results. Australian Geographic Studies 29, 370–381. Coleman P (2005). ‘A southern saltmarsh: Life and times between the tides’. Presentation to Australian Saltmarshes 2005, Sydney. Coleman P and Coleman F (2000). ‘Local recovery plan for the Yellowish Sedge-Skipper and Thatching Grass’. SA Urban Forest Biodiversity Program: Adelaide. Congdon RA and McComb AJ (1980). Productivity and nutrient content of Juncus kraussii in an estuarine marsh in south-western Australia. Australian Journal of Ecology 5, 221–234. Clarke LD and Hannon NJ (1967). The mangrove and salt marsh communities of the Sydney district I. Vegetation, soils and climate. Journal of Ecology 55, 753–771.
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Clarke LD and Hannon NJ (1969). The mangrove and saltmarsh communities of the Sydney district II. The holocoenotic complex with particular reference to physiography. Journal of Ecology 57, 213–234. Clarke LD and Hannon NJ (1970). The mangrove swamp and salt marsh communities of the Sydney district III. Plant growth in relation to salinity and water logging. Journal of Ecology 58, 351–369. Clarke LD and Hannon NJ (1971). The mangrove swamp and salt marsh communities of the Sydney district IV. The significance of species interactions. Journal of Ecology 59, 535–553. Clarke PJ and Jacoby CA (1994). Baseline studies of saltmarsh plants in south-eastern Australia; biomass and productivity. Australian Journal of Freshwater and Marine Research 45, 1521–1528. Clarke KR and Warwick RM (2001). Change in Marine Communities: An Approach to Statistical Analysis and Interpretation, 2nd edn. PRIMER-E: Plymouth. Cresswell ID and Bridgewater P (1996). The major coastal saltmarsh association of Western Australia. In Proceedings of the INTECOL V International Wetlands Conference, Perth, Western Australia. (Ed. AJ McComb). Gleneagles Press: Adelaide. Greenwood ME and MacFarlane GR (2006). Effects of salinity and temperature on the germination of Phragmites australis, Juncus kraussii and J. acutus: Implications for estuarine restoration initiatives. Wetlands 26, 854–861. Johns L (2006). Field Guide to Common Saltmarsh Plants of Queensland. Department of Primary Industries and Fisheries: Brisbane. Kirpatrick JB and Glasby CJ (1981). Salt Marshes in Tasmania: Distribution, Community Composition and Conservation. Occasional Paper No. 8. Department of Geography, University of Tasmania: Hobart. Pomeroy L and Wiebe WJ (1988). Energetics of microbial food webs. Hydrobiologia 159(1), 7–18. Naidoo G and Naidoo S (1992). Waterlogging responses of Sporobolus virginicus (L) Kunth. Oecologia 90(3), 445–450. Outhred RK and Buckney RT (1983). The vegetation of Kooragang Island, New South Wales. Wetlands (Australia) 3, 58–70. Saenger P, Specht MM, Specht RL and Chapman VJ (1977). Mangal and coastal saltmarsh communities in Australasia. In Ecosystems of the World I: Wet Coastal Ecosystems. (Ed. VJ Chapman) pp. 293–345. Elsevier: Amsterdam. Saintilan N (2009). Biogeography of Australian saltmarsh plants. Austral Ecology 34 (in press). Sam R and Ridd P (1998). Spatial variations of groundwater salinity in a mangrove-salt flat system, Cocoa Creek, Australia. Mangroves and Salt Marshes 2, 121–132. Schindl TJ (2002) Environmental and physical factors controlling species composition within the Warneet saltmarsh. BSc thesis, Monash University, Australia. Shepherd KA and Wilson PC (2007). Incorporation of the Australian genera Halosarcia, Pachycornia, Sclerostegia and Tegicornia into Tecticornia (Salicornioideae, Chenopodiaceae) Australian Systematic Botany 20, 319–331. Smith-White AR (1988). Sporobolus virginicus (L.) Kunth in coastal Australia: the reproductive behaviour and the distribution of morphological types and chromosome races. Australian Journal of Botany 36(1), 23–29. Spencely AP (1976). Unvegetated saline tidal flats in North Queensland. Journal of Tropical Geography 42, 78–85.
Distribution of Australian saltmarsh plants
West RJ, Thorogood CA, Walford TR and Williams RJ (1985). ‘An estuarine inventory of NSW, Australia’. Fisheries Bulletin No 2. Department of Agriculture, NSW. Wells AG (1983). Distribution of mangrove species in Australia. In Biology and Ecology of Mangroves – Tasks for Vegetation Science. (Ed. HJ Teas) vol. 8, pp. 57–76. Junk Publications: The Hague. Zedler JB, Nelson P and Adam P (1995). Plant community organization in New South Wales saltmarshes: species mosaics and potential causes. Wetlands(Australia) 14, 1–18.
39
*
*
*
*
7
*
8
NSW
*
9
* *
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
Maireana brevifolia
*
*
T. pterygosperma
*
*
*
*
*
*
*
* *
*
*
*
*
* *
*
*
*
*
T. syncarpa
*
*
*
T. peltata
*
*
*
*
*
T. leptidosperma
*
*
*
T. halocnemoides
*
*
*
*
*
*
*
*
*
T. laeptoclada
*
T. flabelliformis
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
* *
*
*
*
Western Australia
T. deleiformis
*
*
South Australia
T. auriculata
T. pruinosa
*
*
*
*
*
*
*
*
*
*
*
* *
*
*
* *
*
*
*
*
*
* *
*
*
*
T. pergranulata
*
* *
*
*
T. indica
*
*
*
*
*
*
*
*
*
*
*
*
Tasmania
T. disarticulata
*
*
Enchylaena tomentosa
*
*
Dysphania littoralis
Tecticornia arbuscula
*
Dissocarpus biflorus
Chenopodium glaucum
A. semibaccata
A. hypoleuca
Vic
NT
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
* *
*
*
*
10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36
*
*
6
*
*
5
A. cinerea
4
Atriplex padulosa
Chenopodiaceae
3
Queensland
2
1
IBRA Bioregion
Geographic range of 110 Australian saltmarsh plants (Data source: Australian Virtual Herbarium, and published accounts)
State
APPENDIX 2.1
40 Australian Saltmarsh Ecology
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
Xerochloa imberbis
*
*
*
*
*
*
*
*
*
*
*
*
*
Sporobolus virginicus
*
*
Spartina anglica
Puccinellia stricta
*
*
*
*
*
*
*
* *
* *
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
Phragmites australis
*
*
*
*
*
*
*
*
*
*
Paspalum vaginatum
*
*
*
*
*
*
*
*
*
*
Hainardia cylindrical
*
*
*
*
*
*
*
*
*
*
*
*
South Australia
*
*
Tasmania
*
*
Vic
Western Australia
NT
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
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*
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*
*
*
*
*
*
10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36
Lachnagrostis billardieri
Distichlis distichophylla
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
Cynodon dactylon
*
*
*
*
*
*
9
*
*
*
*
*
*
*
8
NSW
Austrostipa stipoides
Poaceae (Gramineae)
Tecticornia asutralsica
S. tenuis
S. arbusculoides
*
*
*
7
*
*
*
*
6
*
*
*
*
5
Suaeda australis
*
Sarcocornia quinqueflora
*
4
S. blackiana
*
Salsola kali
R. baccata
Chenopodiaceae (cont.)
Rhagodia crassifolia
Neobassia astrocarpa
M. oppositifolia
3
Queensland
2
IBRA Bioregion
1
State
Distribution of Australian saltmarsh plants 41
*
*
*
*
*
*
*
*
*
*
*
Cotula coronifolia
*
*
* *
*
*
*
*
*
*
*
*
*
*
*
C. spicatum
*
*
A. subulatus
*
*
*
* *
*
Aster australasica
*
Angianthus preissianus
Asteraceae
*
*
Sesuvium portulacastrum
*
*
Lampranthus tegens
Disphyma crassifolium
*
*
*
*
*
*
*
*
C. glaucescens
*
*
*
*
*
*
*
*
*
9
Carpobrotus rossii
Aizoaceae
*
Schoenus nitens
*
*
*
*
*
*
*
Scirpus nodosus
*
Isolepis cernua
*
*
*
*
*
*
*
*
*
*
*
*
Isolepis nodosa
*
*
*
*
*
8
NSW
*
*
*
*
*
*
7
Gahnia filum
F. polytrichoides
Cyperaceae (cont.)
*
Fimbristylis ferrugubea
*
*
*
*
B. teretifolia
B. juncea
*
Baumea acuta
*
*
Cyperaceae
*
6
Zoysia matrella
5
*
4
Zoysia macrantha
3
Queensland
2
IBRA Bioregion
1
State
Vic
Tasmania
South Australia
Western Australia
NT
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
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*
*
*
*
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*
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*
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*
10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36
42 Australian Saltmarsh Ecology
*
*
*
*
*
*
*
*
* *
*
*
*
*
* *
*
Polypogon monspeliensis
*
*
*
*
*
*
*
*
Plantago coronopus
*
*
Myoporum insulare
*
*
*
*
Mimulus repens
*
*
Lotus australis
*
Lobelia alata
L. solanderi
L.binervosum
Limonium australe
*
*
*
*
*
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*
*
*
*
*
*
*
*
*
*
* *
*
*
*
*
*
*
*
* *
*
Lilaeopsis brownii
*
*
*
*
*
*
*
*
*
*
*
*
Leptocarpus brownie
Lawrencia spicata
Juncus kraussii
Vic
Tasmania
South Australia
Western Australia
NT
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
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10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36
*
*
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*
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9
*
*
*
*
8
NSW
Juncus bufonius
*
*
*
7
Hemichroa pentandra
F. tetrapetela
Others (cont.)
F. ambita
Frankenia pauciflora
Cyperus laevigatus
Centrolepis polygyna
Batis argillicola
Apium prostratum
*
*
Others
*
6
Senecio lautus
5
*
4
C. reptans
3
Queensland
2
IBRA Bioregion
1
State
Distribution of Australian saltmarsh plants 43
*
Portulaca pilosa *
*
Tasmania
South Australia
Western Australia
NT
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*
KEY to COASTAL BIOREGIONS (for locations see Figure 2.3) 1: Gulf Plains; 2: Cape York Peninsula; 3: Wet Tropics; 4: Brigalow Belt North; 5: Central Mackay Coast; 6: South-east Queensland; 7: NSW North Coast; 8: Sydney Basin; 9: South East Corner; 10: South-east Coastal Plain; 11: Flinders; 12: Tasmanian South East; 13: Tasmanian Southern Ranges; 14: Tasmanian West; 15: Tasmanian Northern Slopes; 16: King 17: Naracoorte Coastal Plain; 18: Murray Darling Depression; 19: Kanmantoo; 20: Gawler; 21: Eyre Yorke Block; 22: Nullabore; 23 Hampton; 24: Esperence Plains; 25: Warren; 26: Swan Coastal Plain; 27: Geraldton Sandplains; 28: Carnarvon; 29: Pilbara; 30: Damperland; 31: Northern Kimberly; 32: Victoria Bonaparte; 33: Darwin Coastal; 34: Tiwi Coburg; 35: Arnhem Coast; 36: Gulf Coast
Wilsonia humilis
*
*
Wilsonia rotundifolia
*
*
*
Wilsonia backhausii
*
*
*
*
*
* *
Vic 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36
*
*
*
*
*
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9
Triglochin mucronatum
*
*
*
*
8
NSW
Triglochin minutissima
*
*
*
*
*
Triglochin striata
*
*
*
*
*
*
*
*
*
*
*
7
Spergularia media
*
*
*
6
Selliera radicans
Samolus repens
*
Portulaca oleracea
*
*
5
Portulaca bicolor
4
IBRA Bioregion
3
Queensland
2
1
State
44 Australian Saltmarsh Ecology
Australian Saltmarsh Ecology
(a)
(b)
(c)
(d)
Plate 2.1 Common Australian saltmarsh plants: (a) Gahnia filum; (b) Juncus kraussii; (c) Samolus repens; (d) Sarcocornia quinqueflora.
45
46
Australian Saltmarsh Ecology
(e)
(f)
(g)
(h)
Plate 2.1. cont. (e) Sporobolus virginicus; (f) Suaeda australis; (g) Tecticornia pergranulata; (h) Triglochin striata.
Australian Saltmarsh Ecology
Plate 4.1. Some common saltmarsh gastropods found in the Sydney region: (a) Onchidina australis, supralittoral, Careel Bay, Pittwater (x1.3); (b) Cryptassiminea buccinoides, Kurnell (x 3); (c) Ophicardelus ornatus, Careel Bay, Pittwater (x2); (d) Phallomedusa solida, Careel Bay, Pittwater (x1.2); (e) Cassidula zonata, Towra Point, clinging to the top side of a crab burrow (x 2); (f) Ophicardelus sulcatus, O. ornatus and Pleuroloba quoyi, Bayswater, Pittwater (x1). Photos: J. Ponder.
47
48
Australian Saltmarsh Ecology
(a)
(b)
(c)
Plate 4.2. Saltmarsh plant species and microhabitats used by molluscs in Australian saltmarshes: (a) saltmarsh landscape with adjacent mangrove forests in background, (b) mosaic of common temperate Australian marsh plants: the rush Juncus kraussii and the chenopods Sarcocornia quinqueflora and Suaeda australis, (c) bare area of saltmarsh (salt pan) largely devoid of molluscs. Photos: E. Tucker and P. Ross.
Australian Saltmarsh Ecology
(a)
(b)
Plate 4.3. Impacts to saltmarshes with potential to alter molluscan species assemblage: (a) vehicle tracts, (b) the invasive rush Juncus acutus.
49
50
Australian Saltmarsh Ecology
(c)
(d)
Plate 4.3. cont. (c) Phragmites australis along marsh edge, (d) mangrove invasion. Photos: T. Minchinton
Australian Saltmarsh Ecology
(a)
(b)
(c)
(d)
(e)
Plate 5.1. Saltmarsh crab species (a) Helograpsus haswellianus; (b) Paragrapsus laevis; (c) Parasesarma erythrodactyla; (d) Heloecius cordiformis; (e) Scylla serrata.
51
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Australian Saltmarsh Ecology
Plate 5.2. Newly hatched crab larvae from saltmarsh (a) larva of Helograpsus haswellianus; (b) larva of Parasesarma erythrodactyla.
CHAPTER 3
Geomorphology and habitat dynamics Neil Saintilan, Kerrylee Rogers and Alice Howe
Introduction Since the early work of Lugo and Snedaker (1974), geomorphology has been used to organise our understanding of the interactions between coastal wetlands and their habitats. Mangroves and saltmarshes respond to hydrological and geomorphic conditions in consistent ways (Thom et al. 1967; Woodroffe 1983), such that the relationships between hydrological and geomorphic change can be used as a template to predict changing distributions of mangrove and saltmarsh. With the exception of Tasmania, where mangroves are absent, saltmarshes in Australia are restricted to the upper intertidal environment, generally between the elevation of the mean high tide, and the mean spring tide. The distribution of these environments within an estuary or embayment is controlled by patterns of riverine and marine sedimentation, shaped by the major hydrological drivers of river discharge and tidal propagation. The position of intertidal flats within an estuary will also exert profound influences on water salinity, and provide a major control over the suite of saltmarsh species present. The interplay of hydrology and geomorphology in estuaries has been systematised in the work of several Australian authors (Roy 1984; Woodroffe et al. 1989). These models allow predictions of the changing distributions of mangroves and saltmarsh within estuaries as they infill. The accumulation of fine silt in intertidal environments leads eventually to the conversion of wetland to supratidal floodplain, and the channelisation of the estuary. Marine transgression exerts an opposing influence, exposing upper intertidal and supratidal environments to increased tidal inundation. The study of the interactions between mangrove and saltmarshes in estuaries therefore provides insights not only into the influence of geomorphic change, but the possible impacts of changes in eustatic sea level. After a brief review of the geomorphic settings within which saltmarshes are found in Australian estuaries, this chapter considers interactions between mangroves and saltmarshes over a range of timescales, from recent decades to the past several thousand years. These insights, combined with measurements of rates of sedimentation and subsidence in contemporary saltmarshes, makes possible predictions about how the distribution of saltmarsh may change with the onset of accelerated sea level rise.
53
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Australian Saltmarsh Ecology
Geomorphic settings of saltmarsh Tropical northern Australia Tropical northern Australia is drained by several macrotidal rivers, each receiving seasonal floods at the time of the monsoon (Woodroffe et al. 1989). Mangroves are confined to active depositional environments, including channel banks, the edges of mid-channel islands, and prograding coastlines (Woodroffe et al. 1985). The wide upper-intertidal and supratidal flats are bare in lower rainfall areas, or covered in grasses, sedges, and, in more frequently wetted areas, Melaleuca forest (Woodroffe et al. 1989). These wide estuarine flats are known as blacksoil plains. South-east Queensland The coast of south-east Queensland is dominated by large Quaternary sand islands, including Moreton Island, North and South Stradbroke Island and Fraser Island. These islands shelter the coastal mainland and backbarrier sands from wave action. From Moreton Bay to the Great Sandy Strait, wide intertidal flats are fed by coastal rivers including the Brisbane River and the Mary River, with a constant supply of tidally reworked marine sand. Saltmarshes are distributed on the leeward side of sand islands, on the backbarrier tidal deltaic deposits, and in the channels and channel-fringing flats of the large, mesotidal rivers. The coastline experiences a maritime subtropical climate. Moderate rainfall occurs during the summer period due largely to the passage of warm maritime air from subtropical high pressure cells. Lower rainfall occurs in winter, due to the northward movement of high pressure cells and lower seawater temperatures. New South Wales Within New South Wales coastal wetlands are constrained within estuaries. Roy et al. (1984) recognised three estuary types in New South Wales, each distinguished by the estuary entrance conditions. Drowned river valleys are those estuaries with continual open access to the ocean, and an unrestricted tidal range. Most estuaries in the region, however, have entrance conditions restricted by the presence of a sand barrier. Where the barrier is open, the estuary is termed a barrier estuary. Where the estuary is intermittently closed, the estuary is termed a coastal lagoon, or ICOLL (Intermittently Open/Closed Lake or Lagoon). Drowned river valleys are characteristic of the central coast of New South Wales, including the Hawkesbury River, Sydney Harbour (Port Jackson), and Port Hacking. Generally, drowned river valleys are fed by large coastal rivers, and support a diverse range of geomorphic settings for saltmarsh colonisation. For example, in the Hawkesbury River, Juncus saltmarsh plains find their greatest extent on the wide intertidal flats of the main channel and the tributaries in the central reaches of the river, some 20 km from the coast. This pattern is replicated at a smaller scale in the tributary streams downstream, such as Berowra and Marramarra Creeks, where Juncus saltmarsh occupies channel-fringing flats in the central reaches of the tributaries and the landward segments of cut-off embayments. Towards the estuary mouth, in the backbarrier sands within Pittwater and Brisbane Water, chenopod herbfields dominate narrow saltmarshes landward of the mangrove. Barrier estuaries within New South Wales vary considerably in size and in the degree of saltmarsh development, though saltmarsh is a feature common to all barrier estuaries. In estuaries where the mouth is open, or is maintained in an open state artificially, mangroves are able to colonise the intertidal flats, and the interaction between mangrove and saltmarsh may be controlled by estuary entrance conditions. The enhancement of tidal exchange at the estuary entrance may promote the recruitment of mangroves into the estuary and facilitate the dispersal of proagules into previously saltmarsh environments. Under natural conditions in many
Geomorphology and habitat dynamics
ICOLLs, the temporary closure of the estuary mouth may be sufficient to elevate estuary waters above the level of mangrove pneumatophores leading to their widespread dieback due to anoxic shock. In these circumstances, the dominance of saltmarsh is maintained. Victoria and South Australia The saltmarshes of Victoria and South Australia are distributed in three types of estuaries: river estuaries with permanent entrances; barrier estuaries with intermittently open and closed entrances; and wide marine embayments. For example, in Victoria, saltmarshes can be found in the Gippsland Lakes in the east (barrier estuary), Corner Inlet, Westernport Bay and Port Phillip Bay in the central coastal region (embayments), the Barwon River west of Melbourne (river estuary with permanently opened entrance) and the Glenelg River in the west (river estuary with intermittently opened and closed entrance). In South Australia, saltmarshes can be found in the Coorong in the east (barrier estuary lagoon), Spencer Gulf and Gulf St Vincent in the central region of the state (embayments or inverse estuaries) and Davenport River near Ceduna in the west (river estuary with permanently opened entrance). Within the marine embayments, saltmarshes are located landward of mangrove on wide, bay-fringing flats. Of the Victorian embayments, Westernport supports the greatest extent of saltmarsh, being largely protected from wave action by the narrow entrance and the large islands within the bay, most notably French Island. Saltmarshes occupy wide intertidal flats in this mesotidal estuary almost continuously around the shore, and with extensive flats also found on the northern shores of Phillip Island and French Island. In spite of limited fluvial input, the sediments within the bay are fine, and sedimentation rates are relatively high. Sedimentation appears to be driven by tidal reworking, with powerful (3 m amplitude) tides delivering 4 mm of sediment each year to the saltmarsh. This has allowed the saltmarsh to maintain elevation in the context of increasing sea levels. In South Australia, the vast low-lying supratidal areas of Gulf St Vincent and Spencer Gulf provide habitat for some of the largest areas of temperate saltmarsh. These gulf systems are referred to as inverse estuaries in terms of salinity levels increasing with increasing distance from the open sea. The combination of low topographical relief and higher salinity levels, support the large extent of saltmarsh and upper tidal samphire vegetation communities. Tasmania The saltmarsh flora of Tasmania maintains strong affinities with saltmarshes within the embayments of southern Victoria, with the notable exception of the grey mangrove Avicennia marina. The absence of A. marina from Tasmania is most probably due to occasional severe frost, which is more common throughout Tasmania than anywhere in Victoria (Kirkpatrick and Glasby 1981). Saltmarshes are most common in the highly indented south-eastern coast where they occupy the mouths and upper reaches of small estuaries. Saltmarshes are also a notable feature of the western section of the northern coast where a sand plain of low relief has been drowned by sea level rise. Here, sandbars and islands provide sufficient shelter from erosion to allow intertidal saltmarshes to develop (Kirkpatrick and Glasby 1981). Saltmarshes are virtually absent from the wave-dominated western coast, with small pockets within Macquarie Harbour being the exception. Western Australia There are no published overviews of the saltmarsh environments of Western Australia, though saltmarsh and saltpans are extensive, especially in the north of the state. Saltmarshes (grading into saltpans) reach their greatest extent in the macrotidal settings of King Sound, where a narrow mangrove fringe gives way to saltflats more than 10 km wide. The proportion of
55
56
Australian Saltmarsh Ecology
mangrove relative to saltmarsh increases northwards through the numerous inlets of the Kimberly coastline as rainfall increases, though is rarely more than half the spatial extent. Sand dominates the Dampierland coast south of Broome to 80 Mile Beach. Over this section saltmarshes are confined to smaller inlets and back-barrier intertidal depressions. South of Dampierland, in the Pilbara region, rock again dominates, and wide mangrove and saltmarsh flats occur in environments sheltered from wave attack, on the eastern side of sounds or adjacent to offshore islands. In the humid south-east saltmarsh forms in estuaries in geomorphic settings comparable to those in the Mediterranean climates of the south-east Australian coast, including riverdominated (Swan River estuary) and barrier estuaries (Peel-Harvey Inlet). The south coast of the state is arid, and though the saltmarsh flora is diverse, few estuaries occur west of Albany. Wave energy is high along the southern coast, humidity is low and rocky cliffs are the dominant landform.
Mangrove–saltmarsh interactions over the Holocene Northern Australia Longer-term temporal patterns in the distribution of mangrove and saltmarsh can be reconstructed by coring through estuarine floodplain sediments. Preserved within the soil of these wide, periodically inundated flats is a record of the vegetation sequence of the intertidal environment which can be interpreted from palynological and stratigraphic evidence. For example, beneath the blacksoil plains lie extensive mangrove peat deposits. These have been located in King Sound (Semeniuk 1980; 1982), the Fitzroy River (Jennings 1975), the Ord River (Thom et al. 1975) and the Daly and Alligator Rivers (Woodroffe et al. 1985). Radiocarbon dating of these mangrove facies shows a consistency in age across the estuarine plains and also between river systems. In the South Alligator River, 33 radiocarbon dates returned values within a range of 5370 to 6860 years BP with no spatial trends (Woodroffe et al. 1985). These dates corresponded with dates derived from other systems in northern Australia, including the Fitzroy (5800–7500 BP) and the Ord (circa 6700 BP). The widespread occurrence of mangroves in tropical Australian estuaries in the midHolocene has been termed the ‘big swamp’ phase of development. During this phase, the intertidal flats of many rivers were dominated by mangrove. Saltmarsh would have been restricted to the landward limits of the intertidal zone. Just why mangroves were so dominant at this stage has been subject to debate. Earlier work by Jennings (1975) suggested that higher rainfall during this time may have been responsible for the dominance of mangrove. In the higher rainfall areas of tropical Queensland, mangroves still occupy the wider extent of intertidal environments. Jennings proposed that climatic drying led to the dieback of mangrove, their replacement by chenopods and eventually bare flats. Woodroffe et al. (1985) presented an alternative hypothesis based on the geomorphic evolution of the estuary. In this hypothesis, the widespread ‘big swamp’ phase corresponded to the end of the post-glacial marine transgression. The transgressive phase, where the rate of sea level rise exceeds the rate of sedimentation, is a phase which favours mangrove colonisation. Once transgression ceased, the relative contribution of sedimentation to relative sea level increased, and the elevation of the marsh surface aggraded to elevations above that tolerated by mangrove. Mangroves therefore were restricted to channel-fringing environments as the estuary infilled with sediment. In dryer areas, mangroves gave way to hypersaline flats, while in wetter areas, upper intertidal and supratidal environments were colonised by salt-tolerant grasses and sedges.
Geomorphology and habitat dynamics
South-eastern Australia The role of geomorphology in driving habitat dynamics was overlooked in early work on temperate Australian saltmarsh. For example, Pidgeon (1940) followed vegetation succession theory in describing the processes leading to the development of saltmarsh, which was seen as an intermediate community in a serial progression leading to Eucalypt forest. In this model, the plants themselves are the primary drivers of change, preparing the soil for the subsequent vegetative colonisers. Mangroves were the initial colonisers of exposed estuarine mudflat under conditions of extreme salinity and saturation. Under this model, continued accretion led to colonisation by salt-tolerant grasses (Sporobolus virginicus), followed by the rush Juncus kraussii, then Casuarina glauca. As freshwater conditions replace saline, Eucalyptus invades, with the ‘climax’ community appearing as a mixed Eucalypt forest (Pidgeon 1940). Evidence supporting this dynamic succession was seen as the occurrence of ‘relict species in more advanced zones’ and the active invasion of Juncus into the Sarcocornia meadow. Subsequent research using stratigraphic evidence would show the limitations of this model. Mitchell and Adam (1989a) used the stratigraphic approach to reconstruct vegetation history at several sites in the Georges River and Botany Bay. In particular, they sought to test the model of Pidgeon (1940) that saltmarsh replaced mangrove in vegetation succession. If this model is true, then it would be reasonable to suppose that mangrove root systems could be found preserved beneath the contemporary saltmarsh vegetation. Their coring failed to find any evidence of previous occupation of the saltmarsh habitat by mangrove, and suggested a model which has saltmarsh species as primary colonisers, followed by invasion by mangrove. This model accorded with observations of initial colonisation of intertidal flats by saltmarsh on the northern foreshore of Botany Bay, and the spreading of mangroves into the saltmarsh zone at various locations in the Sydney region (Mitchell and Adam 1989b). In spite of the findings of Mitchell and Adam (1989), there are good theoretical reasons to suggest that saltmarsh may replace mangrove as estuaries infill. Roy (1984) presented a model of estuarine infill for south-eastern Australia which describes phases in the availability of intertidal habitat. Under this model, the deeply-incised drowned river valleys characteristic of south-eastern Australia support little in the way of intertidal vegetation immediately following post-glacial marine transgression. As these valleys infill with sediment, the fluvial bayhead delta progrades seaward, supporting wide intertidal flats. It is during this ‘intermediate’ stage of infill that estuaries support the greatest extent of mangrove and saltmarsh. With the completion of infill, floodplains accrete above intertidal elevations, flow is channelised throughout the length of the estuary, and intertidal habitats are restricted to channel fringes and cut-off embayments. In this model, it could be expected that a mangrove phase is followed by a saltmarsh phase on intertidal flats as the estuary infills. In contrast to the model applied to the tropical north, it is envisaged as a gradual transition, and one occurring at different times within the estuary. Initially, the estuary infills in its upstream reaches, then gradually seaward as the channel progrades. The Hawkesbury River estuary near Sydney presents a good example of a drowned river valley in an intermediate phase of infill. Present day intertidal flats support wide-spread mangrove in the central reaches of the estuary. Seaward of these, mudflats exposed at low tide are unvegetated. Headward, intertidal flats are dominated by saltmarsh, principally Juncus kraussii. Using techniques similar to those of Mitchell and Adam (1989b), Saintilan and Hashimoto (1999) found mangrove peats well preserved 20–30 cm beneath the present-day marsh surface, at the approximate elevation of contemporary mangrove root systems. Beneath these peats, estuarine shells dated to approximately 5000 BP. The age of the mangrove peats varied with distance from the edge of wide intertidal flats, from 1200–1700 BP at the fringes of the flats to 500 BP close to the current mangrove/saltmarsh boundary, suggesting gradual infill.
57
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Australian Saltmarsh Ecology
Mangrove root material retrieved from beneath saltmarsh in small creeks in southern NSW told a similar story. Mangrove root material dating to 1300 BP in Currambene Creek and 1900 BP in Cararma Inlet were retrieved from beneath the contemporary saltmarsh (Saintilan and Wilton 2001). The lack of any significant fluvial input into Cararma Inlet suggested tidal reworking of aeolian and washover deposits as the most likely mechanism for infill. Within Currambene, preserved mangrove root material was more sporadic, and the suggestion was made that channel migration had erased evidence of prior mangrove occupation in some locations. The conclusion drawn from these studies and the work of Mitchell and Adam (1989) is that the geomorphology controls the distribution and evolution of intertidal habitats in south-eastern Australian estuaries. In some cases, this is recorded in the stratigraphic record as a replacement of mangrove by saltmarsh. In other locations, such as in the deltaic sands of Towra Point in Botany Bay, saltmarsh may be the initial coloniser. In other places, all record of the history of interaction is erased by the migration of channels. There is some contention about the contribution of mangrove to geomorphic processes (Furukawa and Wolanski 1996). Lear and Turner (1977), for example, argue that mangrove plays only a secondary role in land-building processes. Bird (1986) on the other hand found that Avicennia marina at Westernport Bay did act as a geomorphic agent, which he attributed to the dense network of pneumatophores formed by this species. Bird suggested that mangrove species, such as Rhizophora spp, that do not produce pneumatophores or similar morphological units, may not contribute substantially to sediment accumulation, while those such as A. marina, which produce prolific pneumatophores, are active land builders. Many saltmarsh plants, however, are known to play an active role in sediment accretion (Bouma et al. 2005; Leonard et al. 1995; Leonard and Reed 2002; Marani et al. 2004; Mitsch and Gosselink 2000; Pethick et al. 1990; Pethick 1981; Shi et al. 1995; van de Koppel et al. 2005).
Recent interactions between mangrove and saltmarsh Saltwater intrusion in northern Australia The models outlined in the previous section describe a replacement of mangrove by saltmarsh as the estuary infills. This may occur rapidly following sea level stabilisation, or slowly, where estuarine sediment loads are low compared to the volume of the estuary. In either case, the transgressive phase is dominated by mangrove proliferation, and the progradational phase characterised by the gradual replacement of mangrove by saltmarsh and, in time, supratidal vegetation. The pattern presents a contrast to trends identified from more recent times from air photographic records. Numerous studies from tropical northern Australia and temperate south-east Australia and New Zealand have demonstrated the encroachment of mangrove into upper-intertidal saltmarsh. Saltwater intrusion is an obvious cause of geomorphological and vegetative changes to estuaries and coastal plains in northern Australia over the past 50 years. The gradual extension of tidal influence along stream channels, the expansion of tidal creeks and the formation of new tidal creeks in the Alligator River Region (Winn et al. 2006) and Mary River (Knighton et al. 1991; Mulrennan and Woodroffe 1998) is linked to the encroachment of mangrove and saline mudflats into freshwater vegetation (Finlayson et al. 1998), localised scour and dieback within Melaleuca forests, accretion of sediment on floodplains (Knighton et al. 1991; Woodroffe and Mulrennan 1993; Bell et al. 2001), changes in subsurface hydrology (Jolly and Chin 1992) and land cover changes (Ahmad and Hill 1995; Bell et al. 2001). Changes since 1950 are significant with bare saline mudflats on the East Alligator River exhibiting a ninefold increase and an associated loss of 64% of Melaleuca forests by 2000 (Winn et al. 2006). More than 17 000 hectares of
Geomorphology and habitat dynamics
freshwater vegetation has been adversely affected on the Mary River and a further 35–40% of floodplains are immediately vulnerable to intrusion (Woodroffe and Mulrennan 1998). A single cause for saltwater intrusion has not been identified (Mulrennan and Woodroffe 1998). Rather, it is apparent that factors such as drier-than-average monsoonal conditions, low frequency and low-intensity cyclonic events and above-average ocean water levels (Winn et al. 2006) facilitate extension of tidal influence into freshwater environments, while tributary development, large tidal range, small elevation differences over floodplains, and uncontrolled feral buffalo promote the expansion of tidal influence (Knighton et al. 1991). Due to the dessication of floodplain sediments in the dry seasons, the process of saltwater intrusion now appears to be internally driven and is likely to continue until an equilibrium state is reached between floodplain elevation and tidal influence (Winn et al. 2006; Mulrennan and Woodroffe 1998). Mangrove encroachment in south-eastern Australia Saintilan and Williams (1999, 2000) cited 28 surveys which demonstrated saltmarsh loss to mangrove encroachment over the period covered by archival air photographs (usually 1940s–present). The trend is apparent across all east coast bioregions and a range of geomorphic settings. Within south-eastern Queensland, Pleistocene sand barriers protect wide, shallow backbarrier deposits which support widespread mangrove and saltmarsh. Mangrove encroachment into saltmarsh in these environments is well documented (McTainsh et al. 1988; Morton 1994; Hyland and Butler 1988; Manson et al. 2003). In northern NSW, mangroves and saltmarshes occupy the mouths of large rivers. While losses of mangrove and saltmarsh to agriculture have been extensive (West 1993), mangroves are encroaching saltmarsh and some agricultural pastures (Saintilan 1998). In central-coast NSW, widespread losses of saltmarsh to mangrove have been reported from both shallow coastal lakes (Winning 1990) and drowned river valleys (Williams and Watford 1997; Williams et al. 1999; McLoughlin 2000; Evans 1997; Mitchell and Adam 1989a, b; Williams and Meehan 2004). South coast NSW estuaries, mostly smaller barrier estuaries (Roy et al. 2001) have shown similar trends with a median loss of approximately 40% of the saltmarsh to mangrove encroachment (Chafer 1998; Saintilan and Wilton 2001; Meehan 1997). Within Victoria, saltmarshes and mangroves occupy the shorelines of large coastal embayments. Here, loss of saltmarsh to mangrove encroachment has been consistent though less dramatic. Declines of 5–12% of saltmarsh to mangrove encroachment have been reported for the saltmarshes of Westernport Bay (Rogers et al. 2005b) and losses have also been noted for Corner Inlet (Vanderzee 1988) and the Gulf St Vincent in South Australia (Burton 1982). Within New Zealand, the proliferation of mangrove is more commonly described as a seaward colonisation (Craggs et al. 2001; Morrissey et al. 2003; Parks 2001) though landward encroachment has been noted (Burns and Ogden 1985). Saintilan and Williams (1999) discussed five hypotheses presented by authors to explain the trend of landward mangrove encroachment. The first of these draws parallels with the suggestions of Jennings regarding the big-swamp phase of the tropical northern estuaries, that higher rainfall freshens the upper intertidal environment promoting mangrove colonisation. This is a compelling explanation in south-east Queensland, where the saltmarsh occurs between a seaward and landward mangrove fringe, and where the relative proportion of mangrove and saltmarsh within estuaries is proportional to rainfall (Bucher and Saenger 1994). In New South Wales, the relationship between mangrove/saltmarsh proportion and rainfall breaks down (Saintilan 2004), and there is no landward mangrove fringe in temperate settings. Here there is a consistency in the upper limit of mangrove in spite of spatial variability in rainfall. Further hypotheses concerned nutrients, sediment, and their interaction. Higher nutrients per se do not promote mangrove colonisation of the saltmarsh (Clarke and Myserscough 1993;
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Saintilan 2003), though fresh sediment may (Clarke and Myerscough 1993; Nelson 2006). However, vertical accretion of sediment has also been shown to be associated with mangrove displacement by saltmarsh over longer timescales. Sediment loads in estuaries may have a less direct effect on mangrove encroachment: by increasing the area of mangrove on fresh substrates, the propagule rain on saltmarsh might be increased, increasing the probability of successful establishment. Explanations involving hydrogeomorphic manipulations of this trend have limited geographic explanation, for a trend ubiquitous in south-eastern Australia. We have mentioned above the possibility that increased frequency of estuary opening in NSW may promote mangrove colonisation, primarily by decreasing the frequency of mortality in flooded ICOLLs. Dredging within channels may also increase tidal amplitude, though at a scale lower than that required to explain widespread mangrove encroachment. For example, the permanent sand bypass on the Tweed has not substantially increased tidal amplitude in that system, nor is there any evidence of altered inundation or sedimentation frequency there (Rogers et al. 2007). Recent evidence has implicated relative sea level rise as a possible explanation of mangrove encroachment in south-eastern Australia. The evidence supporting this assertion is developed in detail in the following sections.
Saltmarshes and sea level Models of saltmarsh response to sea level rise Many factors influence the long-term growth, distribution and abundance of mangrove and saltmarsh, including climate, mean sea level, nutrient and sediment addition and subsidence or autocompaction. In the face of predicted sea level rise, an understanding of the processes that influence the marsh surface elevation relative to local water levels is pertinent. Coastal scientists have long recognised the role that sedimentation and vertical accretion has on the stability of wetlands, and recent research of the influence of below-ground processes on marsh surface elevation has highlighted the contribution of marsh subsidence, autocompaction and uplift to the maintenance of mangrove and saltmarsh elevation relative to sea level. Mangrove and saltmarsh typically establish on mud or sand substrates, binding the sediments from further movement. Since the tidal waters inundating saline coastal wetlands are typically turbid (Adam 1990), as the tide moves over mangrove and saltmarsh vegetation, the vegetation creates resistance which reduces flow velocity allowing suspended sediments to settle on the marsh surface (Adam 1990; Lopez and Garcia 1997; Saenger 2002). The degree to which wetland vegetation promotes sedimentation is related to both vegetation morphologic characteristics, particularly stem density, stem diameter and height, and flow conditions including velocity, depth and sediment concentration (Howe et al. 2005). Wetland vegetation typically generates higher resistance than vegetated riparian channels due to higher vegetation densities and shallower flow depths. Although saltmarsh vegetation density is generally higher than that of mangrove (Howe 2008), sedimentation rates are typically lower. This is due to the location of saltmarsh higher in the tidal frame, where it is inundated by fewer tides moving at slower velocity, and hence less sediment is transported there. The relationship between tidal flow, sedimentation and marsh elevation has led to the concept that mangrove and saltmarsh maintain their elevation relative to the sea level through the surface process of sedimentation or vertical accretion. Marsh elevation keeping pace with sea level was first noted in 1858 (Mudge 1858) and marshes have been found to accumulate sediment at rates equal to historic rates of sea level rise (McCaffery and Thomson 1980; BrickerUrso et al. 1989; Oertel et al. 1989).
Geomorphology and habitat dynamics
Figure 3.1
Relationships between sedimentation and elevation in a coastal saltmarsh.
The concept was best described by Pethick (1981) and is referred to as the negative feedback loop. Sedimentation causes the marsh surface elevation to increase, thereby decreasing the depth and period of tidal inundation and subsequently causing sedimentation to decrease (see Figure 3.1). According to this simple model, marshes will continue to develop and increase in elevation until equilibrium is maintained between accretion, inundation and elevation. When equilibrium is maintained sediment accretion is equivalent to sea level rise. However, it is apparent that other factors modify this simple model, namely sea level variation, compaction or subsidence, vegetation morphology and episodic events, such as storms, floods and droughts. Sea level variations alter the tidal regime, thereby altering the relationship between the marsh surface elevation and the water level. Variations in sea level may be mediated by land movement and changes to estuary morphology. Land movement may occur at scales ranging from regional (tectonic plate movement) to local (subsidence due to autocompaction of soils or groundwater abstraction) (Adam 2002; Allen and Pye 1992; Diamante et al. 1987). Where land movement leads to a net decrease in marsh surface elevation it reduces elevation gains achieved by accretion. Estuary morphology may be modified by natural geomorphic processes such as estuary infill or by anthropomorphic activities such as harbour dredging, flood mitigation works, land reclamation and urban development (McDowell and O’Connor 1977; Roy et al. 2001). Vegetation slows the movement of tidal waters across the marsh enabling sediments to settle and become trapped. Storms and floods typically freshen wetlands and carry high sediment loads. Periods of drought, on the other hand, may increase soil salinity so that less salt-tolerant species are excluded. A rise in sea level may be accommodated within the marsh by an increase in the sedimentation rate, a consequence of increased inundation frequency. The marsh may find a new equilibrium in response to low rates of sea level rise facilitated by higher sedimentation, and higher
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Box 3.1 Investigating marsh surface elevation and below-ground processes using Surface Elevation Tables Since mangrove and saltmarsh occupy a small gradient between mean sea level and highest astronomical tides, they are particularly sensitive to sea level changes. Any changes to sea level relative to the land surface may influence the distribution of mangrove and saltmarsh within the intertidal zone. Sedimentation and erosion contributes to the stability of these ecosystems. However, other below-ground processes, including subsidence, soil compaction, plant productivity, groundwater availability and tidal flooding, may also influence marsh surface elevation and the stability of mangrove and saltmarsh environments within the intertidal zone. The influence of below-ground processes on marsh surface elevation may be investigated using Surface Elevation Tables, also known as SETs or sedimentationerosion tables (Figure 3.2). SETs are a precise and non-destructive method for measuring changes in marsh elevation in shallow coastal environments, including mangrove and saltmarsh. SETs were originally developed in the Netherlands in 1982 (Schoot and De Jong 1982), and since this time, the device has been redeveloped and widely distributed by the United States Geological Survey (USGS).
Figure 3.2 Schematic representation of the Surface Elevation Table (SET) and feldspar marker horizon. Adapted by Sarah Imgraben from Cahoon: www.pwrc.usgs.gov/set/.
Geomorphology and habitat dynamics
The SET consists of a vertical benchmark, an aluminium base support pipe and the portable component of the SET. The portable component consists of a vertical arm that fits onto the support pipe, and an accurately levelled horizontal arm that extends across the marsh surface. From a horizontal plate at the end of the arm, pins are lowered to the marsh surface and the length of the pins above the horizontal plate is measured. Measures are repeated over time and changes in mean pin length provide an estimate of relative changes in marsh surface elevation. Vertical accretion or sedimentation is commonly estimated in conjunction with SET measurements using feldspar marker horizons. A visible layer of feldspar is established on the marsh surface and acts as a marker to measure accretion above. Over time, sediment and other matter accumulates above the feldspar marker and by removing a small core from the marker location, vertical accretion can be measured above the visible feldspar layer. Simultaneous measurements of vertical accretion and surface elevation change can provide information about below-ground processes occurring between the bottom of the SET base pipe and below the feldspar marker horizon. Subsidence, autocompaction or uplift may then be estimated as the difference between relative changes in marsh surface elevation and vertical accretion. Currently, three versions of the SET are available; the original SET, (Boumans and Day 1993; Cahoon et al. 2002), the deep-rod SET and shallow-rod SET (Cahoon et al. 2002b). Since 1993 the SET network has spread from the USA and is now used by more than 70 research teams in 17 countries. More information about SETs can be provided by the USGS or accessed online at: www.pwrc.usgs.gov/set/.
plant productivity. However, this negative feedback loop may fall into disequilibrium if rates of compaction are high, the rate of sea level rise is high or sediment supply is insufficient. This disequilibrium is commonly referred to as an ‘accretion deficit’, and is defined as the difference between sediment accretion and sea level rise. By measuring sediment accretion, determining accretion deficits and appreciating the circumstances that cause the negative feedback loop to fail provides important information about the dynamics of mangrove and saltmarsh and the vulnerability of these ecosystems to the effects of sea level rise. Factors influencing marsh surface elevation in Australia Since mangroves are typically located at lower elevations and are inundated more frequently than saltmarsh, accretion is generally greater in mangrove than saltmarsh throughout southeastern Australia. Tidal range can significantly influence sedimentation between marshes, so that, for constant sediment concentration, sites with a higher tidal range exhibit higher rates of sedimentation than sites with a lower tidal range (see Figure 3.3), though the trend is stronger for mangrove than saltmarsh settings. Rates of accretion varied between geomorphic settings in south-eastern Australia, but in saltmarsh were typically half that of the adjacent mangrove (Table 3.1). Tidal conditions within an estuary are generally highly variable, due to factors such as wind-wave set up, tidal circulation, substrate resistance and the built environment. These spatial differences are superimposed on variability with daily, seasonal and decadal timescales. For example, spring tide suspended sediment concentrations are approximately two to three times greater than neap tide concentrations. This trend is enhanced by seasonal tidal variation
63
Jervis Bay Jervis Bay
Hunter River
Minnamurra River Tweed River Western Port Bay Western Port Bay Western Port Bay Western Port Bay Hawkesbury River Parramatta River
Hawkesbury River
Cararma Inlet Currambene Creek
Kooragang Island
Minnamurra River Ukerebagh Island French Island Kooweerup
Marramarra Creek
Berowra Creek Homebush Bay
Quail Island Rhyll
System
Site
Drowned River Valley
Barrier Estuary Barrier Estuary Coastal Embayment Coastal Embayment Coastal Embayment Coastal Embayment Drowned River Valley Drowned River Valley
Barrier Estuary
Barrier Estuary Barrier Estuary
Geomorphic setting Mangrove Saltmarsh Mangrove Mixed Saltmarsh Mangrove Mixed Saltmarsh Mangrove Saltmarsh Mangrove Saltmarsh Mangrove Saltmarsh Mangrove Saltmarsh Mangrove Saltmarsh Mangrove Saltmarsh Mangrove Saltmarsh Mangrove Mixed Saltmarsh Mangrove Saltmarsh
Zone -0.81±1.00 3.25±0.71 0.29±2.02 0.07±1.49 0.14±1.48 1.98±0.54 2.05±0.63 1.92±0.98 0.61±0.44 0.26±0.87 2.40±1.39 0.49±0.68 -2.13±1.66 5.27±0.96 -0.03±2.23 -0.16±0.94 -2.60±2.07 -0.68±1.18 0.92±1.87 0.64±0.75 -1.51±2.68 1.29±1.48 5.64±2.15 4.66±1.16 2.92±1.59 -2.25±1.75 -2.53±1.27
Mean surface elevation change (mm/yr) 3.03±0.41 1.27±0.13 0.65±0.34 1.37±0.48 0.33±0.11 4.72±0.05 4.19±1.25 2.03±0.38 6.64±0.52 5.93±1.21 2.21±0.30 0.50±0.23 9.49±2.69 4.07±0.25 7.20±0.85 2.03±0.32 6.77±0.79 2.35±0.96 5.10±0.72 1.59±0.19 5.47±0.53 5.05±0.72 4.58±0.28 3.33±0.81 2.20±0.29 0.49±0.49 1.79±0.56
Mean sediment accretion (mm/yr) -3.84 1.98 -0.36 -1.30 -0.19 -2.74 -2.14 -0.11 -6.03 -5.67 0.19 -0.01 -11.62 1.20 -7.23 -2.19 -9.37 -3.03 -4.18 -0.95 -6.98 -3.76 1.06 1.33 0.72 -2.74 -4.32
Mean subsidence (mm/yr)
P<0.001
P=0.723
P=0.001
P=0.001
P<0.001
P=0.001
P<0.001
P=0.014
P<0.001
P<0.001
P=0.041
P=0.071
Differences between accretion and elevation (within sites) P<0.001 P=0.004 P=0.105 P=0.287 P=0.429 P<0.001 P<0.001 P=0.281 P<0.001 P<0.001 P=0.030 P=0.100 P<0.001 P=0.176 P=0.002 P=0.022 P<0.001 P<0.001 P=0.003 p=0.110 P=0.003 P=0.165 P=0.840 P=0.654 P=0.905 P=0.031 P<0.001
Differences between accretion and elevation (within zones)
Table 3.1 Mean (standard error) rates of surface elevation change, sediment accretion and subsidence within the mangrove, mixed and saltmarsh zones at study sites over a three-year period throughout south-eastern Australia, p-values indicating whether rates of sediment accretion were significantly different from rates of surface elevation change (Rogers 2004).
64 Australian Saltmarsh Ecology
Geomorphology and habitat dynamics
Mean rate of sediment accretion (mm/yr)
10 9 r 2 = 0.6023 p = 0.0050
8 7 6 5 4
r 2 = 0.1392 p = 0.2585
3 2 1 0 0
0.5
1
1.5
2
2.5
3
3.5
Tide range (m) Mangrove
Mixed
Saltmarsh
Mangrove Trend (r 2 = 0.6023, p = 0.0050)
Saltmarsh trend (r2 = 0.1392, p = 0.2585)
Figure 3.3 Relationship between mean rates of sediment accretion and tide range of study sites (Rogers 2004).
(i.e. with king tides in summer and winter) and stormy conditions, where increased wave energy remobilises sub- and intertidal sediment (Allen 2000). Longer period phenomena, such as the El Niño Southern Oscillation and the 18.6-year tidal return period (known as the lunar nodal regression cycle) also influence both tidal flow conditions and suspended sediment concentrations (Day et al. 2000; Wells and Coleman 1981). Many factors influence the long-term growth, stability and decline of marsh surface elevation and coastal scientists have long recognised the role that sedimentation or vertical accretion has on marsh stability (for example, Curray 1964; Reed 1990; Woodroffe 1992; Reed 1995). This appears to be true over long temporal scales, as suggested by Saintilan and Hashimoto (1999) for the Hawkesbury River in the latter stages of the Holocene highstand and Woodroffe (1990) for the South Alligator River, Northern Territory. However it is now well established that vertical accretion is not a good surrogate for marsh surface elevation change (Kaye and Barghoorn 1964; Reed and Cahoon 1993; Cahoon et al. 1995). Indeed, a significant relationship between vertical accretion and marsh surface elevation is not apparent over interannual periods (Rogers et al. 2005; Cahoon et al. 2006). While vertical accretion, from both organic and mineral matter, do directly contribute to soil volume, marsh surface elevation appears to be influenced by a greater number of variables than vertical accretion alone (Figure 3.4). In particular, it is apparent that below-ground processes, such as compaction and subsidence of soils, the influence of groundwater and belowground productivity significantly contribute to marsh surface elevation change. Storms, floods and changes to catchment characteristics can also be important. Compaction/Subsidence Over both long temporal scales and interannual periods, marsh surfaces may subside due to the weight of accreted material (Cahoon 2003) or compaction of Holocene sediments causing rates of surface elevation change to trail rates of vertical accretion (Kaye and Barghoorn 1964; Cahoon et al. 1995; 1999; 2000). Indeed, it is apparent that the general trend throughout
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Australian Saltmarsh Ecology
Figure 3.4 Relationships between marsh surface elevation and processes influencing marsh surface elevation. Adapted by Sarah Imgraben from Cahoon et al. 1999.
south-eastern Australia is for marsh surface elevation to be dominated by processes of vertical accretion and shallow compaction (Rogers et al. 2006). However, due to the role of other influencing processes, the degree of shallow subsidence or compaction, estimated as the difference between surface elevation and vertical accretion, varied from site to site. Groundwater Groundwater may also influence marsh surface elevation and significant correlations have been found between marsh elevation change and groundwater pressure head, such that when pressure head is high, marsh surface elevation may increase (Whelan et al. 2005). Alternatively subsurface drainage of water has been suggested to cause marsh collapse (Cahoon 2003), while high rainfall and its contribution to groundwater has been suggested to increase soil volumes (Cahoon and Lynch 1997). Similarly, a study of groundwater elevation at Homebush Bay, which was established in response to strong correlations between surface elevation, rainfall and the Southern Oscillation Index, demonstrated the influence of groundwater on marsh surface elevation. It was proposed that negative SOI values, which coincided with a severe El Niño-related drought and reduced rainfall, subsequently caused the depletion of groundwater, decreased pore water storage and resulted in shrinkage of sediments (Rogers and Saintilan 2008). Interestingly, consistent trends throughout south-eastern Australia have been established between marsh surface elevation change and the Southern Oscillation Index (Rogers and Saintilan 2008; Howe 2008), possibly highlighting the regional role of groundwater in maintaining marsh surface elevation over inter-annual periods. In South Australia, wetland reclamation was purported to cause soil compaction as a result of artificial decreases in the water table (Belperio 1993). The results from the NSW experiments would suggest that this is a viable explanation. Below-ground productivity Sub-surface biological processes, such as root growth and peat development caused by significant increases in plant productivity may influence marsh elevation throughout the root zone (Cahoon and Lynch 1997; Cahoon et al. 1999; 2003a; McKee 2004). Application of nutrients to
Geomorphology and habitat dynamics
plots and associated increases in productivity have been observed to directly increase marsh surfaces (Morris et al. 2002; McKee 2004; McKee and Feller 2004) which may be driven either by increases in below-ground biomass and/or high biomass compressibility (McKee 2004). Similarly, the raising of saltmarsh surface elevation at Homebush Bay, which increased in excess of vertical accretion, coincided with an observed increase in the density of mangrove within the saltmarsh zone (Rogers et al. 2005a). Storms and floods Storms and floods generate high energy flows to estuarine wetlands and typically carry high suspended sediment loads. These events may substantially alter the estuarine sediment budget, and sediment deposited during infrequent storms may constitute the bulk of annual sediment deposition (Roman et al. 1997; Stumpf 1983). These storm events are particularly important for microtidal marshes, where sediment supply is otherwise almost entirely restricted to organic matter (Bricker-Urso et al. 1989; Hensel et al. 1999). High energy storm events may also cause extensive erosion of coastal wetlands, particularly in areas with meso- or macrotidal ranges, and when storms coincide with king high tides (Yang et al. 2003). Where soils are weak, storms may also enhance soil compaction (Cahoon et al. 1995). Changes in catchment characteristics Changes to catchment characteristics generally affect estuarine vegetation through modification of either sediment or runoff budgets. Sediment yield to estuaries is often increased by anthropogenic activity in the catchment, particularly as a result of land clearing. However, a deficit in sediment can occur due to construction of major dams or seawalls, extensive abstraction of fluvial flows or deepening of channels for navigation to the extent that fluvial flows no longer reach tidal wetlands (Adam 2002; Syvitski et al. 2005). If excessive sedimentation occurs, such as that due to runoff from land clearing or mining, estuarine vegetation may be smothered within a few days of the sedimentation event (Saenger et al. 1983). Increased runoff, such as through discharge of urban stormwater, increases the contribution of freshwater flows to the estuary. This may result in the replacement of estuarine communities with more brackish species. More commonly, though, freshwater flows are either reduced by upstream abstraction or regulated by construction of instream storages, which may have the reverse effect (Adam 2002). Urban and industrial infrastructure within the estuary can substantially restrict the area available for migration of estuarine communities. At the landward boundary, the built environment creates hard barriers that prevent landward migration of saltmarsh species, while at the seaward boundary, dredging for navigation can substantially reduce availability of suitable sites for seaward progradation of mangrove onto intertidal flats. Possible sea level rise impacts Our consideration of saltmarsh elevation processes in south-east Australia has demonstrated that marsh surface elevation is variable, and that eustatic sea level trends cannot be directly translated into relative sea level trends for coastal marshes. Marsh survival must be understood in relation to relative sea level, which factors in the trends in marsh surface elevation. Further, while surface processes such as vertical accretion do contribute to marsh surface elevation, the factors which influence marsh surface elevation at inter-annual time periods are primarily below-ground, being groundwater and below-ground biotic processes. This conclusion is in concordance with results from a global analysis of SET data (Cahoon et al. 2006). Using the Surface Elevation Table and associated marker horizon technique, Rogers et al. (2006) were able to demonstrate that inter-estuary variability in the rate of saltmarsh decline was correlated with the degree of relative sea level rise, factoring in surface elevation trends and
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regional eustatic sea level trends. Those sites within which surface elevation declined were those experiencing the most rapid rate of mangrove encroachment, when mangrove encroachment was estimated in terms of vertical migration. This implicates relative sea level rise as the primary driver of mangrove encroachment, as opposed to regional climatic trends. However, it is important to note that other factors may be significant. For example, the generally lower rates of mangrove encroachment in the saltmarshes of Westernport Bay in Victoria could equally be due to colder temperatures, inhibiting growth, or the shading effect of dense stands of the saltbush genus Tecticornia. In subtropical Queensland, interactions between mangrove and saltmarsh across the intertidal gradient has been attributed to climatic controls (Duke 2006), which correspond to periods of mangrove encroachment and mangrove dieback from the saltmarsh.
Conclusions A consistent pattern of saltmarsh decline has been demonstrated for the estuaries of south-east Australia. Comparisons of contemporary distributions with those gauged from historical air photography, now extending back five decades for most sites, reveals replacement of saltmarsh by mangrove over a period in which sea levels were higher than the previous first half of the 20th century. The encroachment of saltmarsh by mangrove has led to a decline of important migratory shorebird roosting habitat, and threatens the availability of habitat important for itinerant fish and crustaceans. The trend of mangrove encroachment and saltmarsh decline would be expected to continue with accelerated sea level rise, predicted for the coming century by the IPCC (Solomon et al. 2007). In this sense, the evidence of the past 50 years, and the associated ecological impacts, might be seen as an early indicator of future trends. The rate of sea level rise has been moderate by global standards (1.18 mm yr–1), and to-date Casuarina and Melaleuca forests have shown little sign of dieback in the region, nor has the seaward fringe of mangrove. The subtle gradients separating mangrove from saltmarsh serve as a sentinel for a trend which in time would be expected to cause the movement of adjacent vegetation zones. Whether saltmarsh has extended beneath the canopy of Melaleuca and Casuarina is difficult to determine from air photography. The variability of saltmarshes in the degree of decline is controlled by local factors influencing below-ground processes, such as groundwater flow and below-ground productivity. This trend is consistent with evidence emerging from the global SET network. In a recent global analysis of SET data, Cahoon et al. (2006) demonstrated a poor correlation between surface elevation trajectories and sedimentation, but closer associations with below-ground processes. There are a number of implications for the management of coastal saltmarsh in the prospect of anticipated sea level trends. Firstly, thought must be given to accommodating the landward encroachment of coastal wetland communities. In some places, this would require landward buffers, the size of which might be determined by high resolution elevation modelling, and surface elevation trend modelling (incorporating trajectories derived from SET data). In many of the deeply incised sandstone valleys of the central NSW coast, natural barriers exist to landward encroachment, and the best prospect for saltmarsh accommodation is on headward floodplains. In many situations, landward and headward accommodation of saltmarsh is restricted by agricultural, residential and industrial development. Planning instruments should be invoked to prevent closing options for future movements of these vital wetland communities. At a minimum, residential development and transport corridors should be prohibited within a vertical range of 50 cm of the upper spring tide inundation bordering existing saltmarsh. The management of wetlands also requires attention to the sources of sediment and water which influence the condition of the wetland. For coastal saltmarsh, groundwater is emerging
Geomorphology and habitat dynamics
as an important control for many communities, both in maintaining vegetation composition and also surface elevation. The diversion of groundwater poses challenges for the sustainability of coastal saltmarsh, possibly promoting subsidence, and promoting mangrove encroachment. The relationship between saltmarsh species composition, surface elevation and groundwater is poorly understood. Marshes also require sediment to compensate for autocompaction and maintain surface elevation. The sediment requirements of coastal wetlands should be considered when managing the movement of sediment through the catchment. The prevailing paradigm of trapping sediment in catchments and minimising delivery of sediment to estuaries, and the coastal wetlands which represent the end-point of sediment movement, should be reconsidered in the context of accelerated sea level rise.
Acknowledgements Chris Harty is thanked for reviewing this chapter and making several helpful comments on the Victorian and South Australian saltmarshes. Sarah Imgraben assisted with the preparation of the figures.
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CHAPTER 4
The ecology of molluscs in Australian saltmarshes Pauline Ross, Todd Minchinton and Winston Ponder
Introduction In quiet waters at the interface of land and sea, species of terrestrial and marine origin merge to form the resident assemblage that constitutes intertidal saltmarshes around the globe. The dominant marine animal residents of coastal saltmarsh are benthic invertebrates, including a diverse collection of snails, crabs and bivalves that rely on the sediments, vascular plants and algae as providers of food and habitat across this marginal intertidal landscape. Exposure of the saltmarsh when not submerged by the tides provides sunlight for plant growth, and places for birds and mammals to rest and feed. Inundation of the saltmarsh at high tide brings other transient visitors, particularly fish and prawns that feed on benthic invertebrates, and permits the flux of nutrients and larvae into and out of the marsh. In this dynamic setting, the marine benthic invertebrates of saltmarshes make their living and play a vital role linking the species and food webs of marine and terrestrial ecosystems. This chapter reviews the ecology of the molluscan component of the marine benthic fauna of Australian saltmarshes. The molluscs of saltmarshes comprise a diverse range of interesting species, including snails, slugs, and occasional bivalves and limpets. The most abundant, species-rich and widely studied group of molluscs are the coiled shelled gastropods (snails), and we still know relatively little about the ecology of this group. Indeed, Richardson et al. (1998), in their comprehensive study of crustacean and molluscan assemblages of Tasmanian saltmarshes, stated that: ‘Almost nothing is known of the autecology of the typical saltmarsh gastropods … in Australia apart from habitat notes in taxonomic works …’ (p. 797, 1998). This statement is alarming because it was published only about a decade ago! Indeed, Fairweather in 1990 found that saltmarshes were the least studied (in terms of publication numbers in the primary literature) of all coastal, marine habitats in Australia and yet they are amongst the most endangered (e.g. Adam 2002; Laegdsgaard 2006). Any understanding of the ecology of molluscs living in Australian saltmarshes must be strongly linked to the features of the marsh (e.g. sediments and plants) upon which they depend for food and shelter, as well as the habitats in the surrounding landscape. Saltmarshes of Australia are different from those in many regions of the world because they share the intertidal landscape with mangrove forests, except in Tasmania and parts of Victoria, South Australia and the southern region of Western Australia, where there are no mangroves (Adam 1990). Indeed, most Australian saltmarshes are bordered on their landward sides by terrestrial habitat, which has often been impacted by humans, and their seaward sides by mud flats or
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mangrove forests, the latter typically extending seaward into mudflats, with seagrass beds often in the lower reaches. This configuration of intertidal habitats across the estuarine landscape in Australia has at least two important consequences for the ecology of molluscs. First, in warm temperate and tropical regions, mangroves occupy the lower intertidal area, so saltmarshes are generally constrained to the upper part of the intertidal landscape and thus are not inundated as frequently or as long by the tide as marshes in cool temperate parts of Australia or in other regions of the world (Adam 1990). This prolonged period of emersion can generate harsh abiotic and edaphic conditions (e.g. hypersaline and compact soils) and limit the types and productivity of plants that can take up residence in the saltmarsh. For example, warm temperate, and particularly many of the tropical Australian marshes do not typically contain the large, productive grasses that characterise the lower intertidal areas of marshes in eastern North America. Second, some species of molluscs span the entire intertidal region and use both mangroves forests (or mud flats in cool temperate regions) and saltmarshes as habitat, and, where mangroves are present, they are sometimes scattered throughout the saltmarsh. Therefore, any discussion about factors affecting molluscs of Australian saltmarshes must include their role in and connections to adjacent mangrove forests or mud flats. Consequently, care must be taken when comparing patterns and processes influencing molluscs of Australian saltmarshes to those described in the literature for molluscs of saltmarshes in some other regions of the world.
Scope of this chapter We have three main goals for this chapter with respect to the ecology of molluscs in Australian saltmarshes: (1) to take stock of the state of knowledge; (2) to highlight research needs, and (3) to provide an introduction for non-specialist and international readers. Our focus here is on the dominant molluscs of the intertidal saltmarshes around the coast of Australia. Where there are gaps in the knowledge of Australian saltmarshes, we draw on examples from outside Australia to highlight important areas for future research. We have divided this chapter into five sections. In the first we introduce the major players comprising the molluscan species assemblage of saltmarshes, briefly describing their main characteristics and where they live. In the next two sections we take a population and community ecology approach, first describing the patterns of abundance for the dominant molluscan taxa across various space scales (geographic, among habitats within estuaries, along the intertidal stress gradients, etc.), with habitat features (e.g. plant species, other microhabitats such as crab burrows) and over time, and then detailing the limited number of experimental investigations into how demographic (e.g. dispersal, recruitment) and ecological (abiotic conditions, food resources, species interactions) processes might account for these patterns. The fourth section takes a more ecosystems approach, describing what we know about the ecological role of molluscs as consumers and prey in food webs, and as modifiers of the marsh habitat. Finally, we end with an overview of the dominant human threats to saltmarsh molluscs and conservation and restoration efforts to alleviate them. A conclusion that we can state up front is that there is a remarkable deficiency of knowledge in almost all aspects of the ecology of molluscs in Australian saltmarshes. Much of what we know is confined to the taxonomic descriptions of species and information on where they live, with the few studies of pattern and process restricted to a handful of abundant and common species in temperate saltmarshes, mostly in south-east Australia. Indeed, for the majority of the molluscs, we are still at the stage of determining their basic biology, life histories, and quantifying their patterns of abundance in space and time. Consequently, research opportunities abound – and we identify key areas for future investigation.
The ecology of molluscs in Australian saltmarshes
The major molluscan players of Australian saltmarshes The major molluscs found in saltmarshes include the gastropods and bivalves living on (i.e. epifauna) or in (i.e. infauna) the sediment. There are, however, transient species such as molluscan larvae, and small and juvenile squid, which may enter with the high tide. Epifaunal and infaunal species While molluscs constitute a morphologically and biologically diverse group, including aplacophorans (spicule worms), polyplacophorans (chitons), bivalves (mussels, clams and oysters, etc.), scaphopods (tusk shells), cephalopods (octopus, cuttlefish, squid and nautiloids), and gastropods (snails, limpets and slugs), many groups are not able to tolerate the relatively harsh conditions in saltmarshes. Adapations are necessary to survive major fluctuations in salinity and temperature and also prolonged aerial exposure, or for burrowing taxa, periods of anoxia. Some groups of gastropods, in particular, three families of basal pulmonates (Ellobiidae, Amphibolidae, and Phallomedusidae), one family of caenogastropods (Assimineidae) are particularly successful in these habitats and have a high proportion of their species in them. Another pulmonate family, Oncidiidae, are slugs that mostly live in estuaries and a few species are typically found in saltmarsh. A few other caenogastropods, notably Littorinidae, have a few species that are found in saltmarshes, but also extend into adjacent estuarine habitats, and one genus of Hydrobiidae is found in saltmarshes around the southern half of Australia (including Tasmania). A few infaunal bivalves sometimes extend into saltmarsh habitat, but all are on the fringe of their intertidal distribution. While a very small number of cephalopod, chiton and true limpet (Lotiidae) taxa can live in brackish conditions, only limpets are occasionally marginal members of the saltmarsh community. These snails, slugs and occasional bivalves live on the sediment surface, vegetation, wood and mud of saltmarshes. Although predominantly classified as epifauna, recent studies (Ross et al. 2003, 2006) have found some species (e.g. Ophicardelus and Pleuroloba) buried 5–7 cm below the soil surface and associated with the roots of Sporobolus virginicus and Juncus kraussii saltmarsh vegetation. We still know little about the ecology and biology of molluscs in saltmarshes, including their reproduction and larval settlement, feeding and diet, behaviour, ecophysiology and microhabitat preference (Kaly 1988; Roach et al. 1989; Roach 1998; Roach and Lim 2000). More commonly, studies of saltmarsh molluscs list epifaunal species (Australian Littoral Society 1977; Hutchings and Recher 1982; Robinson et al. 1983; Morgan and Hailstone 1986; Church et al. 1991; Hutchings 1991; CSIRO 1994; Ponder et al. 2000) or revise taxonomy (Ponder et al. 1991; Hyman et al. 2004; Fukuda and Ponder 2005; Hyman et al. 2005; Golding et al. 2007). Gastropods can be classified broadly into two major groups, Orthogastropoda and Eogastropoda (Ponder and Lindberg 1997). Eogastropoda include Patellogastropoda, the true limpets, most of which live in fully marine, hard-shore habitats, although one is occasionally found in saltmarsh. Orthogastropoda include four major groups (subclasses) of gastropods, three of which, the Neritimorpha, Caenogastropoda and Heterobranchia, are found in salt marsh habitats. A subgroup of the Heterobranchia is the pulmonates or ‘air breathers.’ In these gastropods, the original molluscan gill (ctenidium) has been lost and the roof of the mantle cavity is filled with netted blood vessels, which form a ‘lung’. Oxygen passes into the lung via a small contractile opening, the pneumostome. While most air-breathing snails live in freshwater or on land, a few are marine and two of the most primitive pulmonate families are represented in saltmarshes. These are included in Basommatophora in the superfamily Amphiboloidea – the Amphibolidae (Salinator spp.), and Phallomedusidae (Phallomedusa spp.). Other taxa in this superfamily, including the currently monotypic Maningrididae from
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Arnhem Land, are found lower in the intertidal region in Australian estuaries (Golding et al. 2007). Members of a family (Ellobiidae) of another basal pulmonate group, the Eupulmonata, include a number of typical and characteristic saltmarsh taxa. The Onchidiidae are a family of systellommatophoran slugs. These shell-less pulmonates include several genera, including Onchidium and Onchidina, and are marine and semi-terrestrial (Table 4.1). Other marine snails found in saltmarshes are caenogastropods, with three families represented including the Littorinidae (some species of Bembicium and Littoraria), Assimineidae (several genera including Cryptassiminea), and Hydrobiidae (Tatea spp.) (Table 4.1, Plate 4.1 on page 47). Most caenogastropods have a well-developed gill (ctenidium), but in assimineids it is rudimentary or lost. One of the most common and readily visible gastropods on the surface of the sediment in saltmarshes in south-eastern Australia is the snail Phallomedusa solida (Golding et al. 2007), which until recently was known as Salinator solida. Immature specimens of this species can be easily confused with other species of Salinator as these two genera often occur together (Table 4.1, Plate 4.1). The revision of the taxonomy of the Amphiboloidea by Golding et al. (2007) described eight (including six new) Australian species in three families and although several of the species are separated geographically, S. tecta and S. rhamphidia are sympatric in saltmarshes of south-eastern Australia. Records of S. fragilis in ecological studies in New South Wales are either S. tecta and/or S. rhamphidia as S. fragilis is only found in the southernmost parts of Australia (from Perth, Western Australia to Port Phillip, Victoria) and Tasmania (Golding et al. 2007). At high tide, these ‘air breathing’ snails seal their shells with the operculum against water intrusion to survive for an extended period of time buried in the mud (Macpherson and Gabriel 1962). To reproduce, Salinator eggs are deposited in semi-circular masses, covered with substratum from the environment of the species while Phallomedusa has thin string-like coils of eggs (Golding et al. 2007). Eggs take about 14 days to hatch and are carried to the estuary by the receding tide (Macpherson and Gabriel 1962). Ellobiid snails can be readily distinguished from the globose Salinator and Phallomedusa because their shells are elongated, often with a high spire. The common species in south-eastern Australian saltmarsh are Ophicardelus ornatus, Ophicardelus sulcatus, and Pleuroloba quoyi (Hyman et al. 2004, 2005). Another ellobiid, Cassidula zonata, is more typical of mangrove habitats in south-eastern Australia, and is more often found on the lower edges of saltmarsh. Laemodonta typica is also found in mangroves and saltmarsh in south-eastern and eastern Australia. Although similar to C. zonata, species of Ophicardelus and Pleuroloba vary in shape and size and have one or two, rather than three prominent teeth on the columella and outer lip. In general, shell shape and shell grooving distinguishes O. ornatus from P. quoyi and O. sulcatus (Hyman et al. 2004, 2005), which although abundant throughout the saltmarsh, are easily misidentified (Table 4.1, Plate 4.1) and have incorrectly been considered to represent a single variable species by some workers in the past. The ellobiids are even more diverse in tropical Australia with several genera (Cassidula, Ellobium, Melampus and Pythia) present in the upper intertidal areas in estuaries (Smith 1992; Wells 1997). The Onchidiidae represented in south-eastern Australian estuaries by Onchidium damelii and Onchidina australis (Hyman 1999), the latter occurring mainly in saltmarsh habitats, while several other taxa occur in northern Australia (Table 4.1, Plate 4.1). Onchidium damelii graze on the surface of the sediment (Dakin 1947, 1952; Hutchings and Recher 1982; Healy 1986; Kenny and Smith 1987, 1988; Smith and Kenny 1987; Hyman 1999), consuming microscopic algae with a mantle covered in granular papillae some of which bear simple eyes (Bretnall 1919; Dakin 1947; Hyman 1999) and a single pair of tentacles with eyes at the tips. The leathery Onchidina australis is found under stones, logs etc. on the landward edge of the saltmarsh. Tropical members of this family are more diverse but poorly known taxonomically, with the only revision of all Australian taxa being very dated (Bretnall 1919).
The ecology of molluscs in Australian saltmarshes
The littorinind Bembicium auratum, although most commonly found on hard substrata including oysters and pneumatophores in mangrove forests in south-eastern Australia (Kaly 1988; Branch and Branch 1980; Reid 1988; Crowe 1996), can also be found, albeit in smaller numbers, in the saltmarsh. Two related species live in similar habitats in temperate Australia; B. melanostoma occurs in Victoria and Tasmania, and B. vittatum in South and south-west Australia (Reid 1988). Several species of another littorinid, Littoraria, are distributed throughout the Indo-Pacific and are mainly associated with mangrove forests, but in some areas, the association is also with saltmarsh vegetation (Reid 1986) and under sticks and stones in saltmarshes (Hedley 1905; Hutchings and Recher 1982; Ross et al. 2003, 2006; Table 4.1). Several species also occur on sheltered coasts. L. luteola is abundant in mangrove forests and saltmarshes in NSW and tropical Queensland, while Littoraria cingulate pristissini occurs in saltmarshes in the arid region of Shark Bay in Western Australia (Smith and Kershaw 1979; Robinson and Gibbs 1982; Reid 1986; Ponder et al. 2000). Littorinids that inhabit the saltmarsh need to rely on inundation by spring high tides. Interestingly, under dry conditions they withdraw into their shell and attach to the substratum by mucus (Reid 1986). It is thought Littoraria release pelagic egg capsules or veligers after development in the mantle cavity (Reid 1986), on new and full moons (Muggeridge 1979). Although the specificity of their diet is unclear it is likely to include diatoms and microscopic algae (Reid 1986). Members of the Assimineidae are characteristic of saltmarsh and mangroves and nearly all live near the upper part of the intertidal or even supralittoral zone. A new genus, Cryptassimminea, was recently erected for what was previously known as a single species, with seven species now recognised in southern and eastern Australia (Fukuda and Ponder 2005; Table 4.1, Plate 4.1). Two of these species occur sympatrically in much of south-eastern Australia and Tasmania and have been treated as a single taxon in ecological studies as either Assiminea tasmanica or A. buccinoides (Kershaw 1983; Kelaher et al. 1998; Plate 4.1). Very similar taxa occur in South Australia and south-eastern Tasmania (C. adelaidensis and C. kershawi) with only a single species present. C. tasmanica is also sympatric with C. glenelgensis and C. surryensis, both from Victoria and C. insolata from Queensland (Fukuda and Ponder 2005). Additional species of assimineid occur in south-eastern Australia, including Conassiminea spp. (Fukuda and Ponder 2006), Taiwanassiminea affinis (Fukuda and Ponder, in press) and at least one other taxon. In other parts of Australia, a number of additional genera and species are present, many of them undescribed, and include species of Rugapedia (Fukuda and Ponder 2004) and Ovassiminea (Fukuda and Ponder 2005) as well as other genera. The mostly freshwater family Hydrobiidae is represented in saltmarshes and other estuarine habitats around the southern half of Australia by Tatea huonensis and T. rufilabris (Table 4.1). These small, tall-spired dark brown snails mostly live in the upper mangrove forest and saltmarshes on shaded surfaces and sheltering under plant debris (Ponder et al. 1991). In saltmarshes, the abundance of T. huonensis can be highly variable, exceeding 10 000 individuals per m2 (Hutchings and Recher 1982) comparable to the 18 823 per m2 in mangrove forests (Ponder et al. 1991). T. huonensis and Tatea rufilabris can tolerate a wide range of salinities including freshwater (Ponder et al. 1991) and sexes are separate and small cryptic egg capsules are laid (Ponder et al. 1991). These recent taxonomic revisions (see above) mean that a single species of gastropod referred to in earlier ecological studies may now represent several species or even genera or families! To avoid confusion in this chapter we use the current taxonomic classification, pointing out the relationship to the old species where appropriate and flagging possible instances of possible misidentification or taxon lumping. In contrast to the dominance of several species of limpets on rocky shores, Patelloida mimula is the only species of limpet in south-eastern and eastern Australian temperate
79
80
Australian Saltmarsh Ecology
mangrove forests (Ponder and Creese 1980) and, rarely, saltmarshes. Within south-eastern and eastern Australia mangrove forests, P. mimula is most commonly associated with the hard surface of the Sydney rock oyster, Saccostrea glomerata, being infrequently found on other surfaces including the bark of mangrove trees, occasionally on mud (Minchinton and Ross 1999), while in saltmarshes it is infrequently associated with the vegetation (Ross et al. 2003; 2006). Most members of the tropical Potamididae (‘mud creepers’), are confined to the mangroves, but some individuals, especially of the large Telesopium telescopium, often crawl onto the damp or wet flats behind mangroves as do the related batillariids Batillaria and Pyrazus in more temperate areas. Although saltmarsh habitat is unsuitable for most bivalves because they require submersion to feed, representatives can be found within the saltmarsh. Rock oysters (Ostreidae), such as Saccostrea glomerata, are common encrusting hard substrata in mangrove forests and estuarine locations through warm temperate and tropical Australia, and can occasionally be found on the sediment surface, or attached to pneumatophores of mangrove trees which extend into the saltmarsh. Like other marine organisms, oysters spawn into the water where fertilisation occurs. After a period in the water column, they return to the habitat of the adult and settle (Roughley, 1933; Dinamani 1973). An oyster-like anomiid, Enigmonia aenigatica (Yonge 1957, 1977; Morton 1976) that lives mainly on mangrove vegetation in tropical Australia is sometimes seen on the seaward side of saltmarshes behind mangroves. A small mussel (Mytilidae) Xenostrobus securis, most commonly attached to pneumatophores and buried partially in the sediment in the seaward areas of mangrove forests in estuaries where the salinities are low (Wilson 1968, 1969) can also be found infrequently in the saltmarsh attached to the roots of Sporobolus virginicus and Juncus kraussii (Ross et al. 2003, 2006). The small photosynthetic bivalve Fluviolanatus subtortus can sometimes also be found attached to vegetation in pools. It is often confused with mussels but is a member of the family Trapeziidae (Morton 1982). Another small (less than 5 mm) bivalve is Arthritica helmsi commonly associated with algae and vegetation in south-eastern Australia and has been studied in south-western Australian estuaries (Wells and Threlfall 1982a, b). Densities of this species reach 5000 per m2 in the landward areas of mangrove forests (Yerman and Ross 2004), and it can be found at low tidal elevations in saltmarshes (Ross et al. 2003, 2006). An even smaller bivalve sometimes found living together with Arthritica is ‘Montacuta’ nitens. Small pools and swampy areas can also sometimes provide suitable habitat for other taxa that are normal inhabitants of the mangroves or tidal flats, such as the occasional carnivorous naticid (Conuber spp.) in search of its bivalve prey or, where there is some weed, roots or algae in water, the small estuarine hydrobiid Ascorhis (Ponder and Clark 1988). In the southern and western parts of Australia Hydrococcus brazieri, typically an inhabitant of the upper parts of mudflats, can also extend into the saltmarsh (Wells and Threlfall 1982 a, b; Kershaw 1983). This species is the sole member of the endemic family Hydrococcidae (Ponder 1982) and very closely resembles members of the Assimineidae with which it has been confused in the past (Kershaw 1983). In tropical parts other families of small-sized snails can be found in such habitats, notably the rissooidean Stenothyridae and Iravadiidae, and occasionally ectoparasitic pyramidelloideans, neritids (Neritina) and sometimes shelled opisthobranchs (Haminoeidae) (Winston Ponder pers. observ.) Saltmarsh habitats surrounding hypersaline coastal lagoons and inland salt lakes sometimes have members of the pomatiopsid genus Coxiella living in great abundance (Macpherson 1957; Williams and Mellor 1991). Living within the sediment for all or part of their lives are infauna. Such infaunal organisms include the crabs (see Chapter 5), polychaetes (Hutchings and Recher 1982; Hutchings 1999) and some bivalves. In south-eastern and eastern Australia, Glauconome plankta is one of the few burrowing bivalves and this is found on the seaward edge of saltmarshes and extends into the upper mangroves. Wood-boring fauna such as the teredinid bivalves Teredo navalis
Species
Onchidiidae
Ellobiidae
Amphibolidae
Phallomedusidae
East coast and South West Pacific
Oncidina australis Semper 1882
North to Port Curtis (Gladstone) Queensland New South Wales, southern end of range eastern –most of Victoria, also from north-west Tasmania
Pleuroloba quoyi (H & A Adams, 1854)
From Magnetic Island in Queensland to New South Wales and Victoria
Furthest north in Cooktown, Queensland New South Wales, southern end of the range is Lakes Entrance in Victoria. Not found in Tasmania
Ophicardelus sulcatus H & A Adams, 1854
Onchidium damelii Semper 1882
Queensland, Port Curtis (Gladstone) New South Wales, southern end of range in several localities along the Victorian coastline, most westerly being Port Fairy, and north in Tasmania as well as the south-east
Ophicardelus ornatus (Ferussac, 1821)
Around Sydney, but probably a wider distribution
Salinator rhamphidia Golding et al. 2007 Extends in distribution from northern Queensland, MacKay to northern regions in Tasmania
Eastern and southern Australian coast, Adelaide to Brisbane and northern Tasmania
Salinator tecta Golding et al. 2007
Cassidula zonata H & A Adams, 1854
Southern Australia from Perth to Port Phillip Victoria and southern Tasmania
Only in north-western Tasmania in East Inlet Sawyer Bay
Phallomedusa austrina Golding et al. 2007
Salinator fragilis (Lamarck, 1822)
East and south-east coast of Australia from north Queensland to South Australia and the north-west of Tasmania
Phallomedusa solida (Martens, 1878)
Distribution
Hyman 1999; Smith et al. 2002
Bretnall 1919; Hyman 1999
Hyman et al. 2004, 2005
Hyman et al. 2004, 2005
Smith and Kershaw 1979; Hyman et al. 2004, 2005
Smith and Kershaw 1979
Golding et al. 2007
Golding et al. 2007
Macpherson and Gabriel 1962; Smith and Kershaw 1979; Golding et al. 2007
Golding et al. 2007
Macpherson and Gabriel 1962; Smith and Kershaw 1979; Robinson and Gibbs 1982 ; Golding et al. 2007
References
Species distribution and references to the main gastropods typically found in south and south-eastern Australian saltmarshes.
HETEROBRANCHIA: PULMONATA
Family
Table 4.1
The ecology of molluscs in Australian saltmarshes 81
Family
Assimineidae
Littorinidae
CAENOGASTROPODA
Endemic to St Vincent’s Gulf South Australia
Hobart south-eastern Tasmania
From Bundaberg Queensland to Western Port Victoria and southern Tasmania Glenelg River
Surry River, Moyne River (Port Fairy) and vicinity of Geelong, Victoria
Cryptassiminea kershawi Fukuda and Ponder 2005
Cryptassiminea tasmanica (TenisonWoods 1876)
Cryptassiminea glenelgensis Fukuda and Ponder 2005
Cryptassiminea surryensis Fukuda and Ponder 2005
From New South Wales to central coast of Western Australia
Littoraria luteola (Linneaus 1758)
Cryptassiminea adelaidensis Fukuda and Ponder 2005
Port MacDonnell South Australia and west to Houtman Abrolhos Island Western Australia
Bembicium vittatum
From Cape York north-eastern Queensland to Western Port Victoria
Tasmania, Western Port and Port Phillip Bays, Victoria
Bembicium melanostoma (Gmelin, 1791)
Cryptassiminea buccinoides (Quoy and Gaimard 1834)
From Southern Queensland around the coast to Western Australia
Distribution
Bembicium auratum (Quoy and Gaimard 1835)
Species
Fukuda and Ponder 2005
Fukuda and Ponder 2005
Fukuda and Ponder 2005
Fukuda and Ponder 2005
Fukuda and Ponder 2005
Fukuda and Ponder 2005
Reid 1986
Reid 1988
Reid 1988
Reid 1988
References
82 Australian Saltmarsh Ecology
Hydrobiidae
Family
South Western Australia to southern Queensland
Conassiminea zheni Fukuda and Ponder 2006
Tatea rufilabris (A. Adams 1862)
Shallow Inlet on western side of Wilsons Promontory to Wester Port and Port Phillip Bay in Victoria
Conassiminea studderti Fukuda and Ponder 2006
South Western Australia to southern Queensland
Southern New South Wales to Eden and Tasmania
Cryptassiminea insolata Fukuda and Ponder 2005
Tatea huonensis (Tenison Woods 1876)
Distribution Southern Queensland
Species
References
Ponder et al. 1991
Ponder et al. 1991
Fukuda and Ponder 2006
Fukuda and Ponder 2006
Fukuda and Ponder 2005
The ecology of molluscs in Australian saltmarshes 83
84
Australian Saltmarsh Ecology
and Bankia australis (Turner 1971) and the isopod Sphaeroma terebrans, may also be found living in wood washed into saltmarsh areas. Transient species Transient species enter the saltmarsh with the rising tide. There is a lack of studies on the abundance and distribution of larval stages of gastropods and other molluscs (e.g. squid) that arrive in saltmarshes in this way. Larvae of saltmarsh residents must be released and late-stage larvae returned to the habitat of the adult on incoming spring tides. Although larval stages of gastropods have been found in plankton samples collected on saltmarshes and mangrove forests (Ross 2001; Mazumder et al. 2006), much work remains on the identification and quantification of these stages. The presence of juvenile recruits of Ophicardelus spp. (and probably Pleuroloba) associated with Juncus kraussii at high tidal elevations, suggests that larvae are more likely delivered into this vegetation either by passive or active means, rather than recruiting at low tidal elevations and moving upshore to high tidal elevations (Kaly 1988). Virtually nothing is known about the supply of larval stages of gastropods or bivalves into Australian saltmarshes.
Patterns of abundance and diversity of molluscs in saltmarshes Geographic patterns In contrast to the species diversity of mangroves, which is greatest in tropical regions, saltmarsh plants are more diverse in temperate regions (Saenger et al. 1977; Adam 1990). If mollusc diversity and abundance is related to the diversity of plant species, then one would predict the diversity and abundance of molluscs to increase towards higher latitudes. Such a pattern would also be in contrast to the generalised pattern of an increase in species diversity northwards as latitude decreases. Such broad-scale patterns create interesting but as yet untested hypotheses concerning the geographic distribution of molluscs in saltmarshes in Australia. In general the ecology of molluscs in saltmarshes is better known from studies done in the south-eastern states; little has been published on the ecology of molluscs from South Australia and Western Australia (but see Butler et al. 1975; Wells 1984a, b; Wells and Threlfall 1980; and Table 4.1). Ample records of relevant taxa found in saltmarshes exist in museum collections around Australia but await study and analysis. Patterns among habitats within estuaries The configuration of habitats in a landscape has been found to affect the diversity, abundance and growth of organisms within estuarine habitats (Irlandi and Crawford 1997; Yerman and Ross 2004; Jelbart et al. 2007). Yerman and Ross (2004, unpublished data) found that the biodiversity of molluscs in mangrove forests was dependent on the presence of adjacent saltmarsh habitat. Similarly, Jelbart et al. (2007) found that the diversity and abundance of fish in seagrass beds was dependent on their connectivity with adjacent mangrove forests. Historically, saltmarshes in Australia were surrounded on the landward side by terrestrial vegetation ranging from rainforest to grassland, but commonly Melaleuca or Casuarina forests, and on the seaward side by mangrove forests, extensive mud flats or river channels (Plate 4.2, page 48). In recent years, many of the terrestrial habitats have been used for housing, roads and industrial development, and many saltmarshes have been reclaimed for similar purposes. In addition, the encroachment of mangrove forests into the remaining saltmarshes has been raised as a real concern (Saintilan and Williams 1999). We know little about the influence of adjacent terrestrial habitats on the diversity and abundance of molluscs in saltmarshes and the implications for ecological processes within the saltmarsh.
The ecology of molluscs in Australian saltmarshes
Patterns within saltmarshes While several studies have observed the presence or absence of molluscs in saltmarshes (Hutchings and Recher 1974; 1982; Robinson et al. 1983), only a few have measured the distribution and abundance of molluscs among tidal level elevations or within vegetation types and other microhabitats like wood debris and crab burrows (e.g. Kaly 1988; CSIRO 1994a; Richardson et al. 1998; Roach 1998; Roach and Lim 2000; Wong 2002; Ross et. al. 2003, 2006). This lack of descriptive studies limits our understanding of the patterns of molluscs in saltmarshes. In contrast, patterns of molluscs in mangrove forests have been better quantified (Branch and Branch 1980; Kaly 1998; Underwood and Barrett 1990; McGuinness 1990; Clarke and Ward 1994; Skilleter 1996; Minchinton and Ross 1999; Skilleter and Warren 2000; Chapman et al. 2005; Ross 2006). Such studies may be used to infer spatial patterns of molluscs in saltmarshes, as many species of molluscs found in the upper mangrove forest extend in range into the saltmarsh, and in each habitat they are associated with algae or vegetation. Saltmarshes are physically harsh environments, infrequently flooded by tides and depauperate in shade. For molluscs in these arid environments, habitats providing shade and/or the ability to retain moisture, such as litter and vegetation, provide an oasis where the physical extremes of the habitat are ameliorated. Important habitats for molluscs within saltmarshes include saltmarsh vegetation, occasional mangrove trees, macroalgae, disused crab burrows, litter comprising wood and leaves, often derived largely from adjacent mangroves trees, seagrass wrack and man-made litter. Microalgal mats on the surface of the sediment are also important habitats for molluscs in some saltmarshes. Litter, both natural and man-made, is transported in the saltmarsh by successive high tides and deposited in the upper levels of the marsh. The fauna associated with natural and man-made litter is likely to include barnacles, gastropods, amphipods, and polychaetes, as well as terrestrial arthropods such as spiders, insects and centipedes, and an occasional vertebrate. Seagrass wrack has been found to harbour a range of macroinvertebrates (Chapman and Roberts 2004), and such litter and leaflitter is the preferred habitat of some molluscs such as assimineids. Despite the obvious importance of this habitat, little is known about the dynamics and role of natural and man-made litter as habitat in saltmarshes. Saltmarsh vegetation creates a complex mosaic of microhabitats for molluscs. In southeastern Australia it ranges from Juncus kraussii high in the saltmarsh to Sarcocornia quinqueflora low in the saltmarsh. At mid-tidal elevation a mosaic of several saltmarsh species is present: S. quinqueflora; Sporobolus virginicus; Samolus repens; Triglochin striata and Suaeda australis. Although the type of vegetation (succulent or grass) and the mixture of vegetation creating the mosaic are likely to be significant influences on the diversity and abundance of molluscs, there is a paucity of studies which have quantified patterns or measured correlational relationships between molluscs, vegetation and other organisms such as saltmarsh crabs (Table 4.2, Plate 4.2). A comprehensive study of patterns of the distribution and abundance of molluscs in NSW saltmarsh habitats was done by Kaly (1988), who described three main groupings of gastropods in saltmarshes based on tidal elevation: Ophicardelus (and presumably Pleuroloba) spp. and Littoraria luteola, restricted to saltmarsh habitats; Phallomedusa solida, Cassidula zonata and assimineids in the upper mangrove forest and saltmarsh; and Bembicium auratum and Salinator in the lower mangrove forest. Although the identities of the snail species which she investigated are now unclear because of taxonomic revisions, general patterns described in her study may be useful to future investigations. Kaly’s (1988) study found snails were often larger in size in the saltmarsh compared to the upper mangrove forest (Kaly 1988; Roach and Lim 2000). Roach and Lim (2000) found that P. solida, in the upper saltmarsh, were also less numerous, had the lowest growth rate and lived approximately
85
**
Tatea huonensis
Sarcocornia quinqueflora
*
**
***
**
**
**
**
**
***
*
**
**
**
**
**
**
**
*
** **
Saccostrea glomerata
*
***
*
***
***
**
**
**
*
*
**
**
Mangrove forest
No saltmarsh
*
**
***
*
***
***
***
***
***
*
*
**
*
Unvegetated mud
Open saltmarsh
Tatea rufilabris
**
*
***
**
***
Ophicardelus sulcatus
***
*
***
Ophicardelus spp. sub-adults
***
***
Phallomedusa solida
***
Ophicardelus spp. juveniles
***
Salinator tecta
***
Pleuroloba quoyi
*
*
***
Onchidina
Ophicardelus ornatus
*
*
Littoraria luteola
**
*
S. quinqueflora mixed with Sporobolus virginicus
Open saltmarsh
*
**
*
**
Cassidula zonata
Cryptassiminea
*
Sporobolus virginicus
Open saltmarsh
*
Juncus kraussii
Under the cover of C. glauca
Bembicium auratum
Molluscs
Under the cover of C. glauca
Table 4.2 Relative abundance of molluscs typically found in south-eastern Australian saltmarshes from high shore J. kraussii under the cover of C. glauca to landward areas of mangrove forests where there is no saltmarsh (Ross et al. 2003, 2006). Asterisks indicate relative abundance found in this vegetation type and blank cells indicate that this species was not found in this habitat.
86 Australian Saltmarsh Ecology
The ecology of molluscs in Australian saltmarshes
twice as long compared to individuals in the lower and mid saltmarsh which were numerous, more variable and smaller. Other studies have found the diversity and abundance of molluscs within the saltmarsh habitat correlated with saltmarsh vegetation (Ross et al. 2003, 2006), being greater in vegetated than unvegetated habitats such as mud (Ross et al. 2003, 2006), and correlated with the biomass of vegetation (Ross et al. 2003; McLachlan 2004). In general, Phallomedusa solida has been found more commonly associated with Sarcocornia quinqueflora and unvegetated (mud) patches. Ophicardelus ornatus, O. sulcatus and Pleuroloba quoyi are typically found more commonly associated with the rush Juncus kraussii and the grass Sporobolus virginicus, while Cryptassimminea, Tatea huonensis, Bembicium auratum, Littoraria luteola and onchidiids are all found in vegetated and unvegetated (mud) saltmarsh habitats (Ross et al. 2003; 2006; Table 4.2). The abundance of ellobiids and Cryptassiminea increased with the biomass of S. virginicus and J. kraussii respectively, while P. solida was weakly negatively correlated with the biomass of S. quinqueflora (Ross et al. 2003). Stronger correlations were found between taxa, number of molluscs and percentage cover and biomass of S. quinqueflora (McLachlan 2004), than between taxa, number of molluscs and percentage cover and biomass of the grass, S. virginicus (McLachlan 2004). Further, there are potentially complex relationships between the biomass of vegetation, abundance of molluscs and crab burrows. The density of crab burrows was significantly positively correlated with the biomass of S. virginicus and negatively correlated with J. kraussii (Ross et al. 2003). As the biomass of J. kraussii increased, there were fewer crab burrows and more O. ornatus (Ross et al. 2003). Although overall, the number of crab burrows was the best predictor of the abundance of molluscs. In contrast to these studies, Richardson et al. (1998) found the mollusc composition of saltmarshes in Tasmania could not be predicted by their plant assemblages. Ordination of the faunal data grouped the saltmarshes by degree of submersion, whereas the vegetation was grouped by the salinity of the substratum, although there were some single species associations of ellobiid and other snails with saltmarsh plants (Richardson et al. 1998). Saltmarshes in Tasmania, however, have no mangrove forests on the seaward edge, while Kaly (1988) and Ross et al. (2003; 2006) quantified patterns of molluscs in saltmarsh and mangrove forests. It could be that the results of Richardson et al. (1998) might not be applicable to the supratidal saltmarshes found in NSW. Clearly further work on a greater geographic range of saltmarsh habitat is required before meaningful generalisations can be made. At small spatial scales, there are differences in the preferred habitat of species of molluscs inhabiting saltmarsh vegetation. When the sediment is moist more often, Phallomedusa solida is found on the surface of the sediment (Hutchings 1983; Roach and Lim 2000; Ross et al. 2004; 2006), burrowing as the tidal water covers the sediment. When the sediment dries and becomes compact in the saltmarsh during summer, P. solida and Cryptassiminea can often be found clustered at the base of vegetation of S. quinqueflora, in small depressions and in the burrows of grapsid crabs (Roach and Lim 2000; Ross et al. 2003, 2006; Lee and Choy 2004), perhaps to avoid desiccation stress. Similarly, Ophicardelus ornatus, O. sulcatus and Pleuroloba quoyi are commonly associated with the grasses Sporobolus virginicus and Juncus kraussii (Ross et al. 2003; McLachlan 2004), located among the shoots, often centimeters below the surface of the sediment and in crab burrows (Table 4.2). In J. kraussii these snails can be found 5–7 cm below the surface of the sediment (Ross et al. 2003, 2006). Often at low tidal elevations in the saltmarsh, the surface of the sediment is irregular being a mixture of mound and flat areas, perhaps created by the burrowing of mangrove and saltmarsh crabs (Warren and Underwood 1986; McGuinness 1990), water flow and sedimentation around the bases of saltmarsh plants. Such processes may also influence the distribution patterns of snails. If these mound areas are less compact, have more microalgal food, provide shade and access to moisture, then snails may be distributed non-uniformally throughout the area.
87
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Temporal patterns Very few studies have measured patterns of molluscs in saltmarshes over time. More frequently, studies of molluscs in saltmarshes have sampled at one place and one time. Kaly (1988) found that the densities of all NSW saltmarsh species varied with time quantifying peaks of recruitment of Phallomedusa solida, Salinator, Ophicardelus and Pleuroloba in January–April in the upper mangrove forest and in the saltmarsh. In contrast, Roach and Lim (2000) measured relatively small, but definite peaks of recruitment of P. solida in the later part of the year.
Processes affecting patterns of abundance of molluscs in saltmarshes A number of factors have been suggested as determining the distribution and abundance of molluscs in saltmarshes. These include: life history and demographic processes, physicochemical or abiotic conditions, food resources, habitat, species interactions and natural disturbance. We explore each of these processes in more detail below. Life history and demographic processes A significant gap in our understanding is the role of larval dispersal and supply, settlement, post settlement mortality, or movement in determining patterns of abundance of molluscs in saltmarshes. Although several studies in NSW have found the abundance, distribution and size of gastropods to vary with tidal level in the saltmarsh (Kaly 1988; Roach 1998; Roach and Lim 2000), neither post-recruitment mortality, movement or migration has been found to be a sufficient explanation of the abundance and size of the most common saltmarsh snails, Phallomedusa solida, Cryptassiminea spp., Ophicardelus spp. and Pleuroloba quoyi (Kaly 1988). Kaly (1988) found that there was no tendency for these snails to move upshore or downshore in the saltmarsh, discounting migration as an explanation for the size or distribution and abundance of these species. Differential larval supply or settlement seems more probable as factors explaining the abundance and distribution of molluscs in saltmarshes. Although recruitment has been quantified in the few studies done (Kaly 1988; Roach 1998; Roach and Lim 2000), settlement has not been measured. Part of the reason for this is the difficulty of quantifying a settler in the cryptic microhabitats in saltmarshes. The measurement of larval supply also poses substantial sampling challenges. As the majority of molluscs in saltmarsh rely on a larval stage for dispersal, it is reasonable to predict that larval release and return occur infrequently with tidal inundation, perhaps cued with new and full moons (Ross 2001; Mazumder et al. 2006a). Other studies in NSW have found a differential distribution of barnacle larvae with tidal elevation in mangrove forests, with few larvae in the upper mangrove forest (Ross 2001; Sampatumun and Keough 2001), but similar studies are yet to be done on molluscs in saltmarshes. Only two studies have attempted to sample macroinvertebrate larvae in Australian saltmarshes, both in NSW (Kaly 1998; Mazumder et al. 2006a). Kaly (1988) found there were several species of veliger larvae in the plankton, but numbers of veligers were so small, that it was not possible to draw meaningful conclusions. Mazumder et al. (2006a) quantified peaks of zoea and gastropod larvae on spring tides, and found that there were significantly more gastropod larvae on outgoing than incoming tides, with peaks in the summer months (Mazumder et al. 2006a). Given that the reproductive biology and the timing of late-stage larvae of molluscs in the water column is still largely unknown (Roach 1996), the measurement of temporal (new and full moons, and ebb and flood tides) and spatial patterns (low and upper saltmarsh) of larvae will remain a challenging and perhaps profitable area of investigation.
The ecology of molluscs in Australian saltmarshes
Habitat, abiotic stresses and food Experimental manipulations of molluscs and habitat in the upper mangrove forest (Branch and Branch 1990; Underwood and Barrett 1990; Minchinton and Ross 1999; Ross 2006) may provide insight into the significance of habitat for molluscs in saltmarshes. Kaly (1988) proposed that the density and size of Bembicium auratum in the saltmarsh might be due to the absence of oyster habitat. To test this hypothesis, Kaly transplanted oysters into the saltmarsh, but found B. auratum did not recruit on them (Kaly 1988). She concluded that the lack of available substrata was not likely to limit the distribution of B. auratum in the saltmarsh, and other studies in mangrove forests (Underwood and Barrett 1990) have found oysters are a significant habitat determining the mean density and size of B. auratum. In a further experimental manipulation, Kaly (1988) found that species of molluscs in NSW saltmarshes did not respond to crab burrows, but did respond to the cover of vegetation. Kaly (1988) manipulated the cover of Sarcocornia quinqueflora in plots, reducing 100% cover to 50% and increasing 50% cover to 100%. More Ophicardelus or Pleuroloba recolonised plots with 100% cover of S. quinqueflora. Phallomedusa solida did not distinguish between 50% and 100% cover of S. quinqueflora, and Cryptassiminea was less abundant in 100% cover than in 50% cover of S. quinqueflora, except when the cover of S. quinqueflora was 100% as the result of replanting a 50% plot. This suggests that Cryptassiminea may have been responding to an effect associated with the transplant of S. quinqueflora rather than the percentage cover. Kaly’s experiments suggest that molluscs in saltmarshes can make habitat choices at small spatial scales. Once the vegetation within which they live succumbs to the harsh physical environment in the saltmarsh, they may migrate and recolonise other microhabitats. Further experimental work is required to unravel these complex relationships. The harsh physical environment of a saltmarsh may cause abiotic stresses for the molluscs living within, although greater environmental harshness does not imply less stability in diversity and abundance of fauna (e.g. Rainer 1981). Molluscs in saltmarshes are euryhaline, being able to adapt to the salinity changes experienced in estuaries. In experimental investigations, Wilson (1968, 1969) found that the bivalve Xenostrobus securis survived in salinities which ranged from 1–31‰ and it could withstand sudden dilutions from at least 18–1‰, being capable of surviving at 1‰ for many months while salinities of less than six parts per thousand which mimicked flood events were shown to affect survivorship, condition and reburial of Soletellina alba, an infaunal bivalve living on estuarine flats (Matthews and Fairweather 2004). Other estuarine fauna are eliminated by winter salinities with only a small proportion of the population surviving (Wells and Threlfall 1980). However, Ponder et al. (1991) found the gastropods Tatea huonensis and T. rufilabris, could tolerate a wide range of salinity (from 5–35‰) for at least a month. Members of the family Ellobiidae have also been found to be resistant to desiccation. For example, approximately 50% of adult Melampus bidentatus survived 80% loss and reabsorbed 40% of normal body water in 0.75 hours (Price 1980) and survived submergence for at least 10 days in seawater. Similar experimental studies which may provide insight into the processes affecting the distribution and abundance of molluscs in saltmarshes in Australia are yet to be done. The influence of tidal flow on saltmarsh molluscs appears to be important but has not been systematically investigated. However, salt marsh associated with coastal lagoons where there is little or no tidal movement has few or no molluscs. Also, saltmarsh in areas with high salinity, such as in the South Australian gulfs and the Coorong, have low molluscan diversity as do many salt marsh habitats in northern Australia which are often dominated by large, often salty claypans. Virtually nothing is known of the feeding ecology of the molluscs in saltmarshes (Richardson et al. 1998). It is believed molluscs in saltmarshes graze on the surface of the sediment
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(Hutchings and Recher 1982), feeding on microscopic algae (Hutchings and Recher 1974) and plant detritus (Saintilan pers. comm.). Variation in food source has been suggested as a cause of the distribution patterns of Phallomedusa solida in Sarcocornia quinqueflora and Ophicardelus and Pleuroloba, in Juncus kraussii and Sporobolus virginicus (Cochran 1982). Competition Pre-emptive competition in saltmarshes may be a significant factor in locations where molluscs occur at great densities (e.g. up to 18 823 per m2 of Tatea huonensis) or may influence recruitment among species of molluscs. For example, Bembicium auratum may be less abundant in temperate saltmarsh because Phallomedusa solida, Cryptassiminea and ellobiids directly prevent settlement. Competition among those taxa commonly found in the saltmarsh may also directly limit the abundance of other species. In a series of experiments, Kaly (1988) found recruitment of ellobiids increased when Cryptassiminea and P. solida were removed, and the number of recruits of P. solida increased where the assimineids and ellobiids were removed. Kaly’s (1988) work, however, manipulated only those snails found on the surface of the sediment and only at one time and place. Further investigations are required to determine the extent of competition in influencing patterns of diversity and abundance of molluscs in saltmarshes. Predation Predation is perhaps the most investigated factor determining abundance of molluscs in NSW saltmarshes (Kaly 1988; Roach and Lim 2000). Saltmarsh vegetation in general may provide protection against predation, particularly in areas where dense vegetation occurs. There have been reports of Bembicium auratum and other species in the stomach contents of fish, but these studies suggest that gastropods often make up only a small proportion of their diet (Bell et al. 1984; Kaly 1988). Kaly (1988) found that snails formed less than 10% (of volume) of the toadfish Tetractenos hamiltoni and less than 5% for bream Acanthopagrus australis, with the rest being crabs and barnacles. This may imply that fish are either not the dominant predators of snails, had been feeding in the mangrove forest, or that they spit out the indigestible shell component, and the soft bodies of molluscs may be difficult to identify in fish stomachs (Roach 1998), although Hughes (1984) found a large number of fragments of Littoraria luteola in the guts of T. hamiltoni. Further, Roach (1998) demonstrated experimentally using cages that predators can change the size-structure of adult populations by selecting larger individuals as prey and reduce the abundance of Phallomedusa solida in upper mangrove forest of a temperate estuary (Roach and Lim 2000). In contrast, Kaly (1988) found no effect of excluding fish predators on the abundance of the majority of gastropod species, except ellobiids. Roach (1998) observed that toadfish (Tetractenos hamiltoni), bream (Acanthopagrus australis) and eels (Anguilla sp.) were likely predators of P. solida (also Mazumder et al. 2006a, b). It has been suggested that predation pressure is likely to be greater in upper mangrove forest than in saltmarsh because of greater inundation of mangrove habitat or the saltmarsh might have more predation refuges (Roach 1998), but this will be dependent on the relative importance of aquatic predators versus other aerial predators such as birds. Disease Trematode parasites may also be important in determining the abundance, size and longevity of molluscs on the saltmarsh (Sousa 1993), but few investigations have even determined which parasites are found in molluscs of Australian saltmarshes (but see Jamieson 1966; Bell 1988). Some studies have found Austrobilharzia terrigalensis in Batellaria australis, but this species is found in mangrove forests and mud flats, rather than saltmarshes (Walker 1979). Similarly parasites have been observed to reduce the population of the mudflat and seagrass bed
The ecology of molluscs in Australian saltmarshes
scavenging snails Nassarius burchardi and N. jonasi (Borysko pers. comm.), but as yet no studies have determined the impact of this parasite on survival in the field. The most wellknown disease affecting estuarine organisms along the east coast of Australia is QX, or ‘Queensland Unknown’, a single celled protozoan Marteilia sydneyi, which infects Sydney rock oysters, Saccostrea glomerata. The parasite invades the oyster through the gills and migrates to the digestive gland surrounding the oyster’s intestine. Once at this location, it reproduces, destroying the digestive gland and preventing further nutrient uptake by the oyster (NSW DPI 2005). Given the high population densities and their ready access to terrestrial vertebrate predators, the examination of saltmarsh molluscs as intermediate hosts of trematodes and other parasites should be a rewarding field of study. Natural disturbance Natural disturbances in the form of floods caused by cyclones, droughts and fire may also affect molluscs in saltmarshes. Silliman et al. (2005) found that long periods of drought killed off saltmarshes in North America, while fire has been observed (Minchinton pers. comm.) to scorch saltmarsh. Positive interactions The presence of saltmarsh vegetation might ameliorate the harsh environmental conditions of saltmarshes and create positive interactions for molluscs (Bertness and Callaway 1994). There may also be positive interactions between birds, molluscs and saltmarshes. Molluscs may be preyed on by birds, birds may also disturb the sediment surface in a saltmarsh, aerating the soil and facilitating plant growth by supplementing nutrients. Further, it has been suggested that birds may facilitate the dispersal of snails, by carrying them in the mud on their feet (e.g. Ponder et al. 1991).
Role of molluscs in the saltmarsh ecosystem There has been little attention in Australia given to the role of molluscs as consumers or prey within food webs, or how they might influence habitat structure or plant productivity in saltmarshes. Nevertheless, given that saltmarsh molluscs, and particularly the snails, have been shown to be important detritivores in other saltmarshes across the globe (e.g. Adam 1990; Pennings and Bertness 2001), they are likely to perform similar functions in the cycling of nutrients and energy in Australian saltmarshes. Molluscs as consumers Compared to other marine habitats, relatively little is known of the feeding ecology of molluscs in Australian saltmarshes. As the majority of the molluscan fauna in these saltmarshes comprises snails, slugs and occasional bivalves, their role as consumers is largely as herbivores, grazers, and deposit-feeding detritivores (e.g. Hutchings and Recher 1982), with the smaller bivalve component being suspension-feeding planktivores (filter-feeders). Indeed, the benthic macroinvertebrates of Australian saltmarshes are dominated by deposit-feeding gastropods (and crabs; see Chapter 5 in this volume) that consume microalgae and detritus from the breakdown of plant material, small invertebrates and microbes, and thus are likely to play an important role in decomposing organic matter and recycling nutrients through the food web (Cochran 1982; Guest and Connolly 2004; Lee and Choy 2004). Observations reveal that the snails and slugs consume the surface layer of sediment on the saltmarsh, ingesting plant detritus, microalgae and associated forms of organic matter (Hutchings and Recher 1982). Many snails, such as Bembicium auratum, also graze microalgae from
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the surface of any hard object in the marsh, such as the aerial roots (i.e. pneumatophores) and trunk of mangrove trees, stranded or fixed macroalgae such as Hormosira banksii, rocks or wood flotsam (personal observations of authors). Other snails, such as Littoraria luteola, graze the surfaces of leaves of mangroves and other plants in the saltmarsh, ingesting microalgae and associated organic matter (Mazumder et al. in review). Timing of feeding for the majority of snails has not been determined, but the air-breathing snails are largely restricted to feeding on the surface of the marsh. The role of snails in Australian saltmarshes as direct herbivores of saltmarsh plants and macroalgae has not been well established. Cochran (1982) suggested that the snails Phallomedusa and Ophicardelus ornatus feed on macroalgae. Indeed, blooms of ephemeral green macroalgae are common in saltmarshes and can be associated with pulses of freshwater due to rainfall and nutrient pollution (CSIRO 1994; personal observations of authors). Observations in a temperate marsh in south-east Australia revealed that the small snail Tatea huonensis dramatically increased in density and was associated with green macroalgae following heavy rainfall (CSIRO 1994). Snails also congregate on the surfaces of mangrove propagules stranded in the saltmarsh, presumably feeding on organic matter associated with the decomposing propagules (personal observations of authors). Small snails are also commonly associated with red and green algae on the pneumatophores of mangroves (e.g. Bostrychia–Caloglossa species mix) and on the sediment (Catanella complex) (Yerman and Ross 2004; Ross 2006). The distribution of some species of snails correlates with particular species of plants, suggesting that they may be preferred sources of food or habitat. Cochran (1982) studied Phallomedusa solida and Ophicardelus ornatus in a temperate saltmarsh, and hypothesised that the relative distribution of these snails was related to the marsh vegetation and food resources, with P. solida apparently feeding on microalgae and detritus of marsh plants and O. ornatus feeding more on green algae and marsh plants. Irving (2001) indicated that the grass S. virginicus was a food source for P. solida (see also Lee and Choy 2004). Similarly, Guest and Connolly (2004) found that the detritus of S. virginicus contributed to the diet of the slug Onchidina australis in a subtropical marsh. In contrast, Richardson et al. (1998) suggested that because the molluscs and crustaceans of Australian saltmarshes are detritivores they may not be strongly linked to any particular plant for food, suggesting that such plant-animal associations might be more related to the provision of shelter by the particular plant species. Nevertheless, one might expect that plant species identity will affect the quality of the plant material available to detritivores and thus potentially their distribution. Little is known about how the feeding activities of snails might influence the productivity and breakdown of saltmarsh plants, and the subsequent cycling and export of nutrients through the food web within the marsh and into adjacent waterways. Early saltmarsh research in the eastern USA focused on the potential importance of detrital processes breaking down plant material in productive intertidal marshes and exporting organic matter to fuel secondary production in adjacent coastal waterways (‘outwelling’ hypothesis; Teal 1962; Odum 1980; Pennings and Bertness 2001; Connolly and Lee 2007; Chapter 6 in this volume), although the idea that saltmarshes directly export organic matter in this way has received mixed support (Connolly and Lee 2007). Roach (1998) hypothesised that differential and size-dependent predation on Phallomedusa solida among tidal elevations could be an important interaction driving the magnitude of carbon flow across the upper mangrove forest and into the saltmarsh because P. solida processes about 45 kg of carbon per hectare per year in the upper mangrove forest, but, along with the ellobiid snails, 400 kg of carbon per hectare per year in saltmarsh dominated by Sarcocornia quinqueflora (Hunt 1989). More recently, investigations in saltmarshes of the southern USA have revealed that snails may have strong top-down effects on the dominant grasses that constitute the marsh habitat, particularly during times of drought (Silliman et al. 2005), but such dramatic effects of snails
The ecology of molluscs in Australian saltmarshes
have not been observed in Australian marshes where such large grasses are generally absent. Nevertheless, this important study demonstrated that the role of snails in driving saltmarsh plant production and community structure may be more significant than previously thought. Molluscs as prey In addition to breaking down the organic matter of saltmarsh plants for possible export to mangroves and adjacent waterways (e.g. Congdon and McComb 1980; Clarke 1983; Clarke and Jacoby 1994), it is likely that intertidal molluscs, crustaceans and other macroinvertebrates in Australian saltmarshes play an important role as prey in relaying nutrients and energy from intertidal to adjacent coastal waterways. Indeed, Kneib (1997) has proposed the idea of ‘trophic relay’ where organic matter in saltmarshes may be incrementally transferred from productive intertidal saltmarshes to nearby subtidal habitats in the estuary through predator-prey interactions in the trophic web or shifts in habitat when juveniles feeding in intertidal saltmarshes move to subtidal habitats as they develop. Adults and larvae of molluscs of saltmarshes are prey for a variety of marine and terrestrial predators, including eels, crabs, and fish, as well as resident and migratory birds and mammals (Hughes 1984; Hollingsworth and Connolly 2006; Mazumder et al. 2006a; Roach 1998). Although the relative importance of molluscs in the diet of these predators and their effects on molluscan populations has rarely been quantified in Australia (see Predation subsection above), it would seem from stomach content analyses and observational investigations that snails, as well as other macroinvertebrates, including insects, zooplankton and crab larvae, might form an important part of the diet of fish feeding on saltmarshes (Bell et al. 1984; Hughes 1984; Morton et al. 1987, Kaly 1988; Roach 1998; Mazumder et al. 2006a; Hunt 1989; see Chapter 6 in this volume). Indeed, Hollingsworth and Connolly (2006) found that crab larvae were an important component of the diet of glassfish (see also Mazumder et al. 2006a; Connolly and Lee 2007), and Mazumder et al. (2006a) found large numbers of gastropod and crab larvae in the guts of fish feeding on Australian saltmarshes. In contrast, Kaly (1988) estimated that adult snails likely constitute only a small proportion of the diet of fish that commonly visit NSW saltmarshes. Therefore, there appears to be an important role for benthic invertebrates as consumers, particularly the snails and crabs that feed on detritus and microalgae, and their larvae as prey for larger consumers (e.g. fish and other nekton) in relaying nutrients and energy from intertidal saltmarshes to adjacent intertidal mangroves, subtidal habitats and coastal waters. Nevertheless, compared to crabs and other crustaceans, there is relatively little known about the role of molluscs as consumers and prey in Australian saltmarshes, but snails are known to be important herbivores and detritivores in saltmarshes in North America (e.g. Pennings and Bertness 2001). Importantly, because of the high intertidal position of many mainland Australian saltmarshes, they are not frequently inundated and so might not be expected to play as important a role in transferring organic matter to the ocean as other marshes around the globe. Indeed, benthic macroinvertebrates in Australian saltmarshes might be more important in subsidising terrestrial or mangrove productivity, but this speculation awaits further study (Clarke 1986; Guest and Connolly 2004). Molluscs as habitat modifiers There have been no studies in Australian saltmarshes concerning the role of molluscs in providing habitat for other organisms, ameliorating the soil environment or supplying nutrients to the marsh plants. In low intertidal areas of saltmarshes in north-east USA, dense aggregations of mussels have been shown to stabilise the marsh substratum and provide nutrients from their wastes that promote the growth of marsh grasses, and the burrowing activities of crabs have been demonstrated to have similar positive effects on marsh plants (Bertness 1984;
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1985). Mangrove forests of Australia typically occupy similar relative tidal positions to the low marsh in the north-east USA, and a similar habitat-modifying role likely exists in NSW for the oyster Saccostrea glomerata and the mussel Xenostrobus securis. For example, the oyster Saccostrea glomerata, which is prevalent in mangrove forests but only found occasionally in saltmarshes, provides suitable hard substratum for living and feeding and possibly refuges from predation for the snail Bembicium auratum, the limpet Patelloida mimula and other organisms (Underwood and Barrett 1990; Minchinton and Ross 1999; Claus 2004). In addition, the burrowing activities of crabs have been demonstrated in several Australian studies to facilitate the establishment and growth of temperate and tropical mangroves (Smith et al. 1991; Clarke and Allaway 1993; Minchinton 2001). In Australian saltmarshes, molluscs modify the soil through the burrowing and feeding activities of infauna and epifauna, although their effects on the marsh substratum are not conspicuous. Small bivalves, slugs and snails commonly occur and move across the soil surface, but there are few truly infaunal saltmarsh molluscs, although many are found in the upper layer of sediment and in subsurface root mats and stems of vegetation (e.g. Hutchings and Recher 1974; Warren 1990; CSIRO 1994). Even the effects of crabs, which create burrows and modify the topography of Australian saltmarshes, are much less pronounced in the hard marsh soils compared to the soft muds of adjacent mangrove forests (personal observations of authors). Nevertheless, Ross et al. (2003; 2006) have shown a positive relationship between the density of crab burrows and the biomass of the grass Sporobolus virginicus, suggesting that crabs might increase plant production or that they are attracted to these environments. Epifaunal snails and slugs also live and burrow in the upper sediment layers, and snails are commonly found in crab burrows. Ross et al. (2003) discovered that the common and widespread snail Ophicardelus is present in crab burrows and nestled among the roots of marsh plants 5 cm beneath the marsh surface, suggesting a greater role for these snails in modifying the saltmarsh habitat than may be currently appreciated. Indeed, the soils of Australian saltmarshes are harsh environments and, therefore, even minor modifications to the sediments by snails, crabs and infaunal worms should affect the marsh sediments to some extent and have the potential for positive effects on plant production through increased drainage, oxygenation and fertilisation of these soils (e.g. Bertness 1984; Smith et al. 1991). The positive role of molluscs and crabs on saltmarsh plant production is an area for future research.
Human disturbance and conservation of saltmarsh molluscs The dominant anthropogenic threats and impacts to saltmarsh molluscs are the same as those to species in other marine habitats: destruction, fragmentation and degradation of habitat, invasive species, pollution, and climate change (Adam 1990, 2002; Edyvane 1999; Fairweather 1999; Finlayson and Rea 1999; Edgar and Barrett 2000; Hughes 2003; Laegdsgaard 2006). While the harvesting of molluscs is a primary threat in many marine habitats, such extraction activities are rare in Australian saltmarshes. The harvesting of saltmarsh vegetation that provide habitat for molluscs is also not widespread in Australia, but occasional illegal collection activities, such as saltmarsh plants for floral displays and other uses, can indirectly and directly impact molluscs (personal observations of authors). Human disturbances of molluscs in saltmarshes can act directly, for example, through impacts of an oil spill, or indirectly, through degradation of critical habitat. Although direct human impacts pose significant threats, indirect effects on critical habitat, such as the habitatforming saltmarsh plants or sediments, are likely to have greater and more widespread impacts. Therefore, activities that degrade or alter the saltmarsh plant community pose particularly significant threats.
The ecology of molluscs in Australian saltmarshes
Destruction and fragmentation of saltmarsh habitat The most significant impact on the molluscan fauna of Australian saltmarshes has been through past and continued destruction and so-called ‘reclamation’ of saltmarsh habitat for residential, industrial, agricultural and recreational use by humans (Adam et al. 1988; Adam 2002; Laegdsgaard 2006). Loss of habitat naturally increases fragmentation of the remaining habitat, resulting in smaller and more isolated saltmarshes. The impact of fragmentation of the estuarine landscape on saltmarsh invertebrates has not been examined in Australia, but effects on the abundance and diversity of species and trophic interactions are predicted to be considerable (e.g. Debinsky and Holt 2000). Smaller habitats naturally have more habitat edges relative to the size of the habitat (i.e. greater perimeter-to-area ratios), and thus are more greatly influenced by activities in adjacent habitats. Therefore, increasing fragmentation of saltmarshes can accentuate the influence of degradation of the terrestrial landscape on the saltmarsh habitat. Importantly, because most saltmarsh molluscs are sedentary and disperse via planktonic larvae, understanding the processes of dispersal among and settlement into the small patches of isolated saltmarsh habitat is crucial to understanding the population dynamics of the molluscan fauna, and thus conserving them. Connectivity between saltmarshes is important given the fragmentary nature of these habitats – both along and between estuaries. Where large distances occur this may drive genetic divergence over long periods of time and possibly lead to endemism in very isolated estuaries (Ponder 2004), as has been shown to be the case with Cryptassiminea species (Fukuda and Ponder 2005). Restriction of range coupled with habitat reduction or degradation can result in such taxa becoming endangered or even extinct as has occurred in Japan (Wada et al. 1996). The complete destruction of a saltmarsh in an estuarine landscape must also be viewed according to impacts on adjacent habitats, particularly intertidal mangrove forests. Yerman and Ross (2004) provide the only quantitative study in Australia at the landscape scale, examining the effects of the loss of saltmarshes on adjacent mangrove molluscs and crustaceans. They found that the abundance and diversity of mangrove snails was greater in landscapes where mangrove forests were bounded by adjacent saltmarshes than where mangroves had residential parks or seawalls along their landward borders (see also Bozek and Burdick 2005). Similarly, the clearing of mangrove forests seaward of saltmarshes would likely impact the molluscan fauna of saltmarshes. For example, Roach and Lim (2000) found that recruitment of the snail Phallomedusa solida was generally greater in upper mangrove forests than in saltmarshes. Therefore, losses of mangrove forests would be expected to have direct consequences on the local populations of P. solida living in the adjacent saltmarsh. Indeed, the distributions of several dominant members of the molluscan and crustacean species assemblages range throughout the intertidal habitats of mangroves and saltmarsh (see Table 4.2), and the impact of losing one habitat would be strongly felt in the adjacent habitat. Understanding these connections among habitats in the estuarine landscape and their effect on local population dynamics and species interactions is currently an important area of research in marine ecology. Modification and degradation of saltmarsh habitat One the most common human impacts that degrade saltmarshes occurs at the local scale, including dumping of rubbish that kills underlying vegetation and trampling of vegetation and the sediment surface by humans, other animals and recreational vehicles (bicycles, trail bikes, all-terrain and four-wheel drive vehicles) (Clarke 1993; Adam 2002; personal observations of authors; Plate 4.3 on page 49). These activities directly kill molluscs, destroy saltmarsh plants that provide critical habitat for molluscs, and modify the topography of the marsh surface (e.g. mounds, ruts), altering drainage patterns and associated abiotic conditions (e.g. salinity, waterlogging) that are known to affect patterns of abundance of molluscs
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(e.g. McGuinness 1990; Minchinton 2001; McLachlan 2004; Ross 2006). No study has quantified the effects of trampling by humans on molluscs in Australian saltmarshes (but see Anderson 1995). Nevertheless, Ross (2006) found that trampling in Australian mangrove forests reduced the abundance and recruitment of the ellobiid and assimineid snails, which are also common species in adjacent saltmarshes. Saltmarshes have often been used in Australia and across the globe to graze livestock, such as cattle and sheep, resulting in dramatic reductions in the cover and diversity of saltmarsh plants due to the effects of herbivory and trampling, but impacts on molluscs have been less well established (e.g. Adam 1978, 1990; Anning 1980; Bakker 1985; Jensen 1985; Turner 1987; Mitchell and Adam 1989). Bridgewater (1982) described how cattle grazing on the plants Tecticornia sp. in southern Australian temperate marshes could shift dominance of the plant community by these species and increase the colonisation of the marsh by exotic grasses, but made no comments about associated impacts to the invertebrate fauna. Zedler et al. (1995) observed an abundance of the plant Triglochin striatum in a temperate Australian saltmarsh subjected to cattle grazing and hypothesised that unvegetated marsh areas and waterlogged conditions due to cattle hoofprints created a favourable environment for this species. Shuttleworth (2006) quantified differences in environmental conditions, vegetation and macrofauna in intertidal areas landward of mangrove forests in a temperate NSW estuary (a mosaic of unvegetated mud, mangroves and saltmarsh plants) that were either grazed or not grazed by cattle, and found that areas with cattle had less compact soil, fewer mangrove seedlings and pneumatophores, and substantially smaller densities of Phallomedusa solida and Cryptassiminea. The reduced densities of these snails could result from mortality due to burial because of direct trampling by cattle or indirectly from changes to the suitability of microhabitats and sediments due to the increased numbers of depressions from cattle hoofprints that pool water at the soil surface (Shuttleworth 2006; see also Ross 2006). Modifications that alter water drainage, tidal flow, erosion, sedimentation and the general hydrology of saltmarshes and other habitats are widespread across the estuarine and coastal landscape of Australia, including culverts, floodgates, drainage channels and seawalls (e.g. Williams and Watford 1997; Saintilan and Williams 2000). Saltmarsh habitat has also been modified and degraded on a larger scale by powerlines, bridges, rail and roadways, and other infrastructure (e.g. Williams and Watford 1997; Adam 2002). Despite this, little attention has been devoted to studying the effects of these artificial structures and modifications on the species assemblages of molluscs and the ecology of Australian saltmarshes in general. It can be predicted, however, that such alterations of the hydrology and habitat of saltmarshes would dramatically influence abiotic and edaphic conditions (e.g. salinity, waterlogging) and species interactions, favouring some species over others and altering the abundance, composition and diversity of species in the marsh. Perhaps the best example of how modifications to saltmarsh drainage patterns could potentially affect molluscan assemblages is research on the impacts of runnelling (i.e. construction of small drainage channels) for mosquito control in some Australia saltmarshes (see Chapter 8 this volume). Although the impacts of these changes to the drainage and physical structure of marshes have not been quantified for molluscs, runnels affect the habitat structure and abundances of crabs and fish and facilitate dispersal of mangroves into saltmarshes (Breitfuss et al. 2003, 2004; Connolly 2005), and thus would be expected to affect molluscs and other invertebrates (see Connolly and Lee 2007). Pollution Pollution from oil, excess nutrients, chemicals, and other toxic substances has been shown to affect the coastal saltmarshes of Australia directly and indirectly (Adam 2002). The most
The ecology of molluscs in Australian saltmarshes
widely studied pollutant of saltmarshes in Australia is oil, which can arise from leakages and spills from both terrestrial and marine sources. Studies have focused on the effects of oil on saltmarsh species assemblages because their high intertidal position in the estuarine landscape makes them susceptible to the long-term deposition and potentially persistent impacts should a significant spill occur. There have been several documented oil spills in Australia where mangrove forests and saltmarshes have been impacted, resulting in damage and death to some trees and marsh plants (e.g. Allaway 1982; Anink et al. 1985; McGuinness 1988; 1990; Clarke and Ward 1994). The response of saltmarsh species to oil is variable, and recovery of plants can take years (e.g. Allaway 1982). Due to concern over the likelihood of a major oil spill along the Australian coast (e.g. in the Sydney region there is a large oil refinery next to Ramsar-listed saltmarshes and extensive mangrove forests; see McGuinness 1988), experimental studies have been done simulating oil spills by applying weathered crude oil to the saltmarsh and mangrove communities and measuring impacts on the species assemblages, including molluscs, over time (e.g. McGuinness 1988, 1990; Grant et al. 1993; Clarke and Ward 1994). McGuinness (1990) examined the response of macroinvertebrates in a temperate NSW saltmarsh dominated by the chenopod Sarcocornia quinqueflora after applying oil either once or twice, and found that the abundances of assimineid snails and Phallomedusa solida generally declined within a few weeks of oiling, and that the oil killed a substantial number of snails. These negative effects were generally short-lived for Cryptassiminea, with densities rebounding to control levels within a few months after oiling. The same was generally true for P. solida, although the effects of repeated oiling were persistent and densities for this species had not returned to control levels when the experiment ended about a year after the second oiling (McGuinness 1990). Clarke and Ward (1994) did a similar study, including the application of crude oil and diesel fuel, in areas of a temperate NSW saltmarsh dominated by the grass Sporobolus virginicus, but some Sarcocornia quinqueflora was present. The oil and diesel killed almost all of the above-ground cover of vegetation and there was little recovery of plants even 17 months after application. As in the study by McGuinness (1990), the application of oil and diesel killed substantial numbers of snails, and abundances of all species (Phallomedusa solida, Ophicardelus spp., Bembicium auratum, and Littorina luteola) declined dramatically relative to controls (Clarke and Ward 1994). Within about six months (and only about two months for Bembicium auratum and L. luteola), immigration from adjacent areas resulted in the abundances of all gastropod species converging with those in control areas, which was unexpected given the sparse cover of live vegetation. Excess nutrients, particularly nitrogen, from urban and agricultural run-off, have the potential to alter competitive interactions among plants and the relative abundance of plants within saltmarshes because marshes are typically limited by nitrogen (e.g. Valiela and Teal 1974; Minchinton and Bertness 2003). In particular, the influx of nutrients might ameliorate the harsh abiotic conditions of saltmarshes, favouring the colonisation and spread of some species, particularly invasive species not normally suited to saltmarsh conditions (see Invasive species subsection below). Nutrients can also produce blooms of green macroalgae, where snails congregate, apparently to feed (personal observations of authors; see also CSIRO 1994), but how such algal blooms ultimately affect molluscs is not known. Finally, acid runoff from reclaimed coastal land, such as in the Manning River areas of NSW, is another important pollutant that could potentially impact molluscs. Sulphuric acid can form in saline reclaimed land and then run off into drainage ditches and remnant saltmarshes, causing oyster mortality and a range of deleterious impacts to exposed fauna and habitat. The implications for saltmarsh molluscs are yet to be investigated (Sammut et al. 1986; Dove and Sammut 2007).
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Invasive species Saltmarshes are thought to be largely resistant to invasion by exotic and weedy native species because of their stressful environmental conditions (hypersalinity, poorly oxygenated soils, waterlogging, etc.). Nevertheless, over the past century and particularly in recent decades there have been significant invasions of plants and macroinvertebrates in coastal saltmarshes around the globe (e.g. Adam 1990; Ruiz et al. 1999). Invasive species could affect native molluscs in saltmarshes through the introduction of other invertebrates that are superior competitors, predators, parasites or agents of disease, or plants that displace native species and alter critical habitat for molluscs. Australian saltmarshes have been invaded by many exotic plant species, particularly along the along the terrestrial-marsh ecotone (see Adam 1981; Bridgewater et al. 1981; Robinson et al. 1983; Sauer 1988; Clarke 1993; Adam 1994, 2002). While many of the small, herbaceous annual species do not appear to have any obvious impacts on native species assemblages, there are several larger perennials that are more aggressive invaders and cause for concern as they displace native plants (e.g. groundsel bush Baccharis halimifolia, pampas grass Cortaderia selloana, and spiny rush Juncus acutus along the upper border of marshes and rice grass or cordgrass Spartina anglica, the only introduced plant species along the lower marsh and sometimes seaward of mangroves) (Plate 4.3). For many of these species, the process of invasion is likely to be due, at least in part, to physical disturbance (e.g. removal of marsh plants, disturbance of soil) of the saltmarsh along the terrestrial border and amelioration of naturally stressful abiotic conditions (increased terrestrial runoff of freshwater, nutrients and sediment), creating opportunities for colonisation and spread, particularly of plant species that thrive in disturbed areas and have broad tolerances of abiotic conditions (see Minchinton and Bertness 2003). The effects of these species introductions on the molluscs and other fauna of saltmarshes are largely unknown, but an important area of research (e.g. Chambers et al. 1999). In contrast, if we ignore the considerable impacts of domesticated exotic animals, such as sheep and cattle, on saltmarshes (Bridgewater 1982; Adam 1990; 2002), there have been far fewer occurrences of animal invaders that might affect saltmarsh molluscs. This might be changing, however, as the apparently expanding population of the non-indigenous European shore crab Carcinus maenas has been observed in Australian saltmarshes and it preys on molluscs (Thresher et al. 1997; Grosholz and Ruiz 1996). In Australia, there has been a dramatic increase in the human modification of the land surrounding saltmarshes, particularly around urban centres, and recent observations reveal the increased occurrence and seaward encroachment of non-indigenous and native invasive species along the terrestrial border of coastal saltmarshes. In temperate saltmarshes of south-east Australia, the non-indigenous spiny rush Juncus acutus, native to the Mediterranean, is increasingly colonising the upper border of coastal saltmarshes, and is now listed as a noxious weed species in some regions of Australia (e.g. Burchett et al. 1998a; Greenwood and MacFarlane 2006; Harvey 2006). Observations reveal that in some areas this invasive rush is displacing the native rush Juncus kraussii along the upper border of marshes (Greenwood and MacFarlane 2006; personal observations of authors), potentially altering associated invertebrate species assemblages. Harvey (2006), in a detailed study of insect assemblages, revealed that J. acutus hosted a different species assemblage and trophic composition of insects to that of J. kraussii. Moreover, a preliminary study revealed differences in the abundance and diversity of snails on the substratum beneath these plant species (Harvey 2006, unpublished data). Due to the prominent differences in the morphology between these rushes, the displacement of J. kraussii by J. acutus is likely to have significant effects on associated fauna. As with J. acutus, the native grass Phragmites australis is appearing with greater frequency and abundance along the upper border of temperate saltmarshes in south-east Australia,
The ecology of molluscs in Australian saltmarshes
particularly in areas with urban and agricultural development in the adjacent terrestrial landscape (Adam 2002; personal observations of authors; Plate 4.3). The cryptic invasion of an exotic strain or species of P. australis has aggressively invaded coastal marshes throughout eastern North America (Saltonstall 2002), with devastating consequences for native species, including molluscs (Chambers et al. 1999). Indeed, P. australis is expanding its distribution and displacing native plants (e.g. Juncus kraussii, Sporobolus virginicus, Sarcocornia quinqueflora) in some marshes in south-east Australia (personal observations of authors, see also Greenwood and MacFarlane 2006), but effects on the molluscs associated with these plants have not been quantified. Similar changes appear to be occurring in coastal saltmarshes of Western Australia, where increases in freshwater and nutrients from urban stormwater drains has resulted in invasion by the more freshwater cattail Typha orientalis and competitive displacement of the more salt-tolerant native saltmarsh rush J. kraussii (Zedler et al. 1990; Davis and Froend 1999). These shifts in plant species compositions due to changing abiotic conditions have great potential to alter dramatically molluscan species assemblage in coastal saltmarshes, but research in this area is lacking. Less commonly, coastal saltmarshes may be invaded from their seaward borders. In Tasmania and parts of Victoria, as well as other regions of the globe, the non-indigenous grass Spartina anglica has colonised intertidal mudflats and saltmarshes, and even mangroves and seagrass beds (Boston 1983; Daehler and Strong 1996; Hedge and Kriwoken 2000; Kriwoken and Hedge 2000). In Australia, S. anglica was deliberately introduced in the 1930s to reclaim habitat for pastureland and stabilise mudflat sediments (Boston 1983; Hedge and Kriwoken 2000; Kriwoken and Hedge 2000). In some areas, the invasion of S. anglica has resulted in the competitive displacement of smaller native saltmarsh plants, such as Sarcocornia quinqueflora and Samolus repens (and also seedlings of the mangrove Avicennia marina), with possible increases in estuarine productivity and changes to the food web, including molluscan detritivores (Hedge and Kriwoken 2000; Kriwoken and Hedge 2000). Indeed, observations suggest that the physical habitat and entire species assemblage, including plants, invertebrates, fish and birds, is affected by the invasion of S. anglica (Kriwoken and Hedge 2000). Moreover, invasion by S. anglica has the potential to increase substantially the deposition of wrack in saltmarshes, potentially resulting in more bare patches as wrack smothers underlying marsh vegetation (Boston 1983; Minchinton 2002). Hedge and Kriwoken (2000), in a detailed snapshot study in Tasmania, found that the abundances and species richness of benthic macroinvertebrates, which are dominated by molluscs, in areas where Spartina anglica is the main vegetation, were greater than on adjacent unvegetated mudflats, but similar to saltmarshes dominated by the native marsh plant S. quinqueflora and with patches of the rush Juncus kraussii. Despite the similarities in the S. anglica and native saltmarsh communities, there were differences in the abundance of some species as well as the habitat structure and edaphic conditions, and Hedge and Kriwoken (2000) cautioned that considerably more study is needed before firm conclusions can be made about the impact of S. anglica on the molluscs and the other benthic invertebrates in Australian saltmarshes. Another potential threat to molluscs of Australian saltmarshes is the increasing landward encroachment of the native mangrove Avicennia marina into temperate saltmarshes in south-east Australia over the past 50 years (e.g. Mitchell and Adam 1989a, b; Saintilan and Williams 1999; Rogers et al. 2006). The reason why mangroves appear to be colonising saltmarshes is not known, although landward migration due to changes in climate and sea level rise, changes to marsh surface elevation, and amelioration of the harsh abiotic conditions of saltmarshes by human modification of catchment land facilitating establishment and growth have all been implicated (Saintilan and Williams 1999). Evidence does not support one
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explanation over another, and encroachment might be more related to a suite of local environmental conditions (Saintilan and Williams 1999; Saintilan and Wilton 2001; Wilton 2002; Rogers et al. 2006; Saintilan et al. Chapter 3 in this volume; Plate 4.3). Regardless of the mechanism, the molluscan fauna of mangroves dominated by A. marina is different to that of temperate saltmarshes (e.g. Kaly 1988; Roach 1998) and, therefore, the displacement of saltmarsh plants by mangroves is likely to result in dramatic changes to the abundance, composition and diversity of molluscs. There have been few reports of invasive fauna significantly impacting the molluscs of Australian saltmarshes. Nevertheless, there is considerable cause for concern over the widely invasive and non-indigenous European green crab Carcinus maenus, which invaded Australia in the late 1800s but has been expanding its distribution in temperate waters in recent decades (Thresher et al. 1997, 2003). Carcinus maenus has been observed in a variety of coastal habitats in Australia, particularly those with soft sediments and low to moderate wave energy, and although it is not found commonly in intertidal region except during flood tide (Thresher et al. 2003), it is common in coastal saltmarshes around the globe. Moreover, Carcinus maenas is a significant predator of native molluscs, crustaceans and other species, and has demonstrated capacity to alter dramatically entire species assemblages (Cohen et al. 1995; Thresher 1997; Thresher et al. 2003; Grosholz et al. 2000), and therefore has the potential to alter molluscan assemblages of Australian mangroves and saltmarshes. The Pacific oyster Crassostrea gigas, which was deliberately introduced for aquaculture, is now resident in the intertidal areas of many estuaries. While there is concern that C. gigas is out-competing the native oyster Saccostrea glomerata in mangrove forests in some estuaries (e.g. Ponder et al. 2002), it is used extensively for aquaculture and is well established in all intertidal habitats in Tasmania, where mangroves are absent (Kriwoken and Hedge 2000). The effect of this species on molluscs in saltmarshes is not known. Similarly, the mosquito-fish Gambusia holbrooki was deliberately introduced in Australian waters to control mosquito populations (Adam 2002). It can tolerate a broad range of salinities and thus feed on mosquito larvae on saltmarshes, but it also consumes non-target items, such as the eggs and larvae of fish and larvae and adults of crustaceans (e.g. Morton et al. 1988; Ivantsoff and Aarn 1999; Komak and Crossland 2000). Although it has never been assessed, the broad diet of this species suggests that it would also influence the larvae and perhaps adults of saltmarsh molluscs. Climate change and sea level rise Assuming that saltmarshes survive local destruction and degradation by humans over the next century, the most important future impact on saltmarsh molluscs will likely be through the predicted effects of climate change and associated sea level rise. The anticipated consequences for saltmarshes are certainly not clear and range from changes (i.e. changes to temperature, sea levels, storm frequency and sediment dynamics) to conditions that could promote both the landward and seaward migration of this community (e.g. Bryant 1990). Under perhaps the most likely scenario that saltmarshes will migrate landward under sea level rise (e.g. Vanderzee 1988; Woodroffe 1990), the response of molluscs will depend on their ability to colonise and survive at higher levels on shore. For most species this would not appear to be a significant barrier as many disperse via planktonic larvae and have tolerance for a broad range of salinities (e.g. Wilson 1968; 1969; Ponder et al. 1991). More importantly, perhaps, would be predicting whether the saltmarsh plants upon which many molluscs and other species depend as habitat and food would also be successful in colonising these areas. Indeed, much will depend on the rate of sea level rise relative to the life history, physiological tolerances, and degree of local adaptation of the individual species. Also, how mangroves respond to climate change and sea level rise will have important implications
The ecology of molluscs in Australian saltmarshes
for saltmarsh species assemblages, as mangroves can probably respond rapidly through the dispersal of propagules and establishment of seedlings, which can potentially displace saltmarsh plants (Saintilan and Williams 1999). Similarly, the landward spread and increased dominance by the invasive grass Spartina anglica in temperate marshes would be predicted under most climate change scenarios. In addition to sea level rise, increased water and particularly air temperatures may favour some species of saltmarsh plants over others and mangroves over the entire suite of saltmarsh plant species. More frequent storms might also change the sediment dynamics and salinity conditions of saltmarshes, favouring some species over others. Importantly, some of the terrestrial habitat landward of saltmarshes has been modified for human use as urban or agriculture land and thus unavailable for the landward migration of saltmarsh plants and molluscs under sea level rise. Conservation and restoration of saltmarsh molluscs There have been only a few studies either restoring or conducting restoration experiments in Australian saltmarshes (see Laegdsgaard 2006), and we do not know of any restoration project directly targeting specific molluscs or other invertebrate fauna. Indeed, the majority of restoration activities do not involve re-introductions of particular species, but gross changes such as the restoring the natural tidal regime and removing damaging species such as cattle and invasive plants. Nevertheless, these activities and others where saltmarsh vegetation is planted are clearly and sensibly targeting the importance of structural saltmarsh plants as habitat and food for fauna, although the changes to fauna are rarely documented. Importantly, although the majority of saltmarsh molluscs and other benthic invertebrates disperse and settle in new habitats via planktonic larvae, it cannot be assumed that just because habitat is built that invertebrates will come. Chapman and Roberts (2004) provide one of the only experimental studies directly examining how to restore Australian saltmarshes. The study examined the use of seagrass wrack in facilitating the restoration of saltmarsh plants and found that wrack generally increased the cover and biomass of Sarcocornia quinqueflora on an unvegetated sandflat. There were no effects on the benthic invertebrates, including some snails, but few animals were found in plots so the results need to be interpreted cautiously (Chapman and Roberts 2004). To predict the natural recovery of saltmarsh species assemblages following removal or degradation of saltmarsh plants by humans, investigators in Australia have also experimentally simulated disturbance by removing aboveground saltmarsh vegetation and then following patterns of recolonisation (Laegdsgaard 2002; Lee and Choy 2004). Laegdsgaard (2002) cleared small (25 cm × 25 cm) areas in temperate saltmarshes and monitored natural recovery of the dominant plants as estimated by plant cover over 21 months. Recolonisation of the chenopod S. quinqueflora varied with tidal height and occurred primarily via vegetative growth, with low elevations recovering to control levels of about 90% cover within 21 months, but plots at higher tidal elevations attained only about 20% cover. Recolonisation by the grass S. virginicus, which was also primarily through vegetative growth, although seedlings contributed in some plots, was the same across tidal heights and attained only about 30% cover compared to control levels of 90% cover. Laegdsgaard (2002) predicted that full recovery for both species at all tidal elevations could take five years and, although macrofauna were not monitored, we would predict that recolonisation by molluscs that rely on this vegetation as habitat would therefore take longer. Lee and Choy (2004) mowed large circular plots of 5, 10 and 20 m diameter in sub-tropical saltmarshes dominated by the grass Sporobolus virginicus and monitored the macroinvertebrate assemblages for 18 months. In contrast to the study by Laegdgaard (2002), the recovery of S. virginicus was relatively rapid (although spatially variable), with substantial regrowth of
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Box 4.1
Sampling saltmarsh molluscs
While sampling the abundance of saltmarsh molluscs may appear straightforward, there are some important considerations. Sampling the snails of salt marshes can be particularly difficult because many are small and cryptic. Indeed, many snails burrow into the sediment and are often associated with macroalgae, mangroves and saltmarsh plants, particularly the dense tussocks of grasses and rushes. Therefore, non-destructive sampling techniques that estimate abundances by visually observing the surface of the marsh substratum and plants are likely to underestimate numbers (Ross 2006). The most accurate estimate of snail density is likely to result from destructive sampling (Ross, unpublished data), including collection of both above-ground vegetation and the sediments (with the type and size of the sampling unit dependent on the species that you are targeting). It is also important to distinguish live from dead snails, as the shells of dead snails often remain on the marsh surface. Destructive sampling is not always possible or desirable. In such circumstances, because of the strong associations between snails, saltmarsh plants and other microhabitats (Ross et al. 2003, 2006), quantifying snail abundances would benefit from estimates that incorporate the microhabitats present, such as plant identity and abundance (cover, biomass, height, etc.) or densities of crab burrows, to help explain some of the variation. Of course, limitations may remain as molluscs within the sediments and on the vegetation are often not sampled and this should be acknowledged. For some species that burrow under dry conditions (e.g. Phallomedusa solida), counts done when the surface of the marsh is wet might prove more accurate (Roach and Lim 2000), but no study has evaluated this. Finally, because of the sparse distribution of many saltmarsh molluscs (and other benthic macroinvertebrates), choosing appropriate sampling units and sampling scales from pilot studies will be important. We recommend that greater standardisation of methods and protocols are needed for sampling molluscs and other macroinvertebrates in Australian saltmarshes.
shoots from within the plots (although no quantitative data were provided, Lee and Choy 2004). Remarkably, there was no significant effect (for any plot size over the entire sampling period) of the removal of saltmarsh vegetation on the abundance, richness or diversity of macrobenthic invertebrates (which included several species of common snails such as Phallomedusa solida) collected using core and pitfall trap sampling (Lee and Choy 2004). Snails were observed clustered around the bases of plants and in crab burrows. Unfortunately, these two studies used different sampling variables to estimate recovery of vegetation (e.g. cover versus biomass and both are likely important) and fail to report important details of methods to allow comparisons (Laegdsgaard 2002; Lee and Choy 2004).
Conclusions Despite their ecological and economic importance (Adam 1990), and the fact that they are rapidly disappearing, coastal saltmarshes of Australia have attracted far less attention than coral reefs, rocky intertidal seashores and most other marine habitats in Australia. Research into the ecology of saltmarsh molluscs and indeed all aspects of the ecology of coastal salt-
The ecology of molluscs in Australian saltmarshes
marshes has been neglected (e.g. Fairweather 1990; Connolly 1999; Lee and Choy 2004). Nevertheless, with the listing in 2004 of coastal saltmarsh in NSW as an Endangered Ecological Community under the NSW Threatened Species Conservation Act and the increased recognition of saltmarsh as important nursery habitat for fish (e.g. Beck et al. 2001; Minello et al. 2003; Connolly and Lee 2007; Chapter 6 in this volume), there seems to have been an increasing emphasis on research into the ecology of saltmarshes in Australia over the past decade or so. At this early stage, the need for future research is at the most fundamental level, quantifying the distribution and abundance of species in space and time and understanding their life histories and demography. While research-funding agencies in Australia focus on single-species approaches and research needs to sustain fisheries for humans, a progressive and contemporary research agenda for the ecology of Australian saltmarshes will recognise the importance of understanding plant-animal interactions and the dynamics of the saltmarsh plants and invertebrates, which provide food and habitat for these fish (e.g. Kneib 1997). Although saltmarshes occupy greater areas in the tropics than in temperate Australia (Adam 1990; Bucher and Saenger 1994; Zann 1995), research on Australian saltmarshes has largely been done at subtropical and temperate latitudes, which might reflect the increasing diversity of saltmarsh plants (and potential habitat for molluscs) from tropical to temperate regions (Saenger et al. 1977; Adam 1990) or the distribution of the population and research scientists in Australia. Therefore, studies are needed of the molluscs, other macroinvertebrates and the general ecology of tropical saltmarshes (Adam 2002). There continues to be an urgent need for taxonomic studies on the as yet unstudied groups of molluscs, particularly those that inhabit the tropical parts of Australia. It is also vital that researchers lodge vouchers in a museum collection to enable future confirmation of the taxa because of the poor understanding of some groups. We emphasise this because identifying snails in situ in the saltmarsh is not easy because sand and mud often obscure key structural features. The shell and aperture shape, nature of the operculum if present, shell sculpture, shell colour and pattern and presence and shape of an operculum provide useful guides in identification of gastropods (Macpherson and Gabriel 1962; Smith and Kershaw 1979; Ponder et al. 1991; Ponder et al. 2000; Hyman et al. 2004; Fukuda and Ponder 2005; Hyman et al. 2005; Golding et al. 2007; Table 4.1). Other molluscs in the saltmarshes, such as slugs, are identified by the presence or absence of the mantle and position of the pneumatosome, while the shape, external sculpture of valves, the structure of hinge teeth, the position of the ligaments and the shape of the pallial sinus in bivalves are critical in their identification. It should also be emphasised that all of the available identification guides (e.g. Macpherson and Gabriel 1962; Robinson and Gibbs 1982; Wells 1984a, b; Ponder et al. 2000) have outdated taxonomy for many of the saltmarsh molluscan taxa. Without exception there is almost a complete lack of knowledge about the larval stages of saltmarsh molluscs. This should be a priority area for future research because dispersal and recruitment is a key process determining the population dynamics for many species with planktonic larvae (e.g. Underwood and Fairweather 1989) and, moreover, larvae may be an important component of the diet of fish and other nekton (e.g. Hollingsworth and Connolly 2006). Relatively simple information available for other species of benthic invertebrates, such as identification of the larvae and timing and cues for larval release into the plankton and settlement into the benthic habitat, are virtually unknown for most molluscs of Australian saltmarshes. More challenging studies determining the population structure (e.g. open, closed) and connectivity among local populations of molluscs in the isolated and fragmented saltmarshes within and among estuaries of Australia would be extremely desirable. Indeed, any study of the larval stages and recruitment for these species would be valuable.
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There is little doubt that understanding the recruitment processes, population dynamics and maintaining the integrity of the habitat-forming plant species that constitute Australian saltmarshes will be important in sustaining populations of molluscs and other species. Determining the relationships across various spatial scales between the abundance, composition, diversity and productivity of saltmarsh plants and the associated molluscan (and other) fauna will be a valuable area of research, aiding conservation and restoration. Further, understanding the role of molluscs and other macroinvertebrates in relaying nutrients and energy from saltmarshes to adjacent habitats and waterways will help to confirm the significance of saltmarshes as important areas maintaining viable fisheries. Establishing long-term monitoring programs, at the very least for habitat-forming species, should be a priority as they are needed to assess temporal changes in the patterns of abundance of plants, molluscs and other macroinvertebrates and their relationships and thus provide relevant baselines upon which to evaluate future environmental change (e.g. Clarke 2003), particularly response of plants and animals to increases in temperature, sea level rise and carbon dioxide predicted under climate change. We predict that non-indigenous and invasive species are going to become increasing problems within Australian saltmarshes, particularly in temperate regions where marshes are small and invasion will be facilitated by increasing encroachment and modification by humans. We believe that in the short term it will be more important to focus on small-scale impacts to saltmarshes rather than longer and larger scale ones (e.g. climate change, sea level rise). We simply need more research of molluscan and plant assemblages at different places and times, so that any regularities of pattern and similarities of process explaining those patterns can be determined for Australian saltmarshes. This will be costly but there is some obligation to protect endangered ecological communities and their associated fauna in Australia. Otherwise, we may end up with the loss of molluscs from Australian saltmarshes before we even know their ecological significance.
Acknowledgements We thank Neil Saintilan for the invitation to write this chapter and for the opportunity to discover how alarmingly little we know about the molluscs of Australian saltmarshes. We are grateful to numerous colleagues and referees who substantially improved this chapter by providing difficult to find references, taxonomic expertise and helpful comments on various drafts. Support was also provided by the College of Health and Science at the University of Western Sydney, and the Institute for Conservation Biology and School of Biological Sciences at University of Wollongong. Photographs by TE Minchinton, PM Ross, E Tucker and J Ponder. Winston Ponder thanks Ian Loch for discussion and information on saltmarsh molluscs.
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Underwood AJ and Barrett G (1990). Experiments on the influence of oysters on the distribution, abundance and sizes of gastropods Bembicium auratum in a mangrove swamp in New South Wales, Australia. Journal of Experimental Marine Biology and Ecology 137, 25–45. Valiela I and Teal JM (1974). Nutrient limitation in salt marsh vegetation. In Ecology of Halophytes. (Eds RJ Reimold and WH Queen) pp. 547–563. Academic Press: New York. Vanderzee MP (1988). Changes in saltmarsh vegetation as an early indicator of sea-level rise. In Greenhouse: Planning for Climate Change. (Ed. GI Pearman) pp. 147–160. CSIRO: Australia. Wada K, Nishihira M, Furota T, Nojima S, Yamanishi R, Nishikawa T, Goshima S, Suzuki T, Kato M, Shimamura K and Fukuda H (1996). Present status of estuarine locales and benthic invertebrates occurring in estuarine environment in Japan. WWWF Japan Science Report 3, 1–182. Walker JC (1979). Austrobilharzia terrigalensis: a schistosome dominant in interspecific interactions in the molluscan host. International Journal of Parisitology 9, 137–140. Warren JH (1990). Role of burrows as refuges from subtidal predators of temperate mangrove crabs. Marine Ecology Progress Series 67, 295–299. Warren JH and Underwood AJ (1986). Effects of burrowing crabs on the topography of mangrove swamps in New South Wales. Journal of Experimental Marine Biology and Ecology 102, 223–235. Wells FE (1997). A review of the northern Australian species of Ellobium and Cassidula. In The Marine Flora and Fauna of Darwin Harbour, Northern Australia. (Eds JR Hanley, G Gaswell, D Megerian and HK Larson) pp. 213–230. Northern Territory Museum, Darwin and the Australian Marine Sciences Association. Wells FE (1984a). A Guide to the Common Molluscs of South-western Australian Estuaries. Western Australian Museum: Perth. Wells FE (1984b). Comparative distribution of macromolluscs and macrocrustaceans in a north-western Australian mangrove system. Australian Journal of Marine and Freshwater Research 35, 591–596. Canberra. pp. 57–79. Wells FE and Threlfall TJ (1980) A comparison of the molluscan communities on intertidal sand flats in Oyster Harbour and Peel inlet, Western Australia. Journal of Molluscan Studies 46, 300–311. Wells FE and Threlfall TJ (1982a). Reproductive strategies of Hydrococcus brazieri (Tenison Woods, 1876) and Arthritica semen. (Menke, 1943) in Peel Inlet, Western Australia. Journal of the Malacological Society of Australia 5(3–4), 157–166. Wells FE and Threlfall TJ (1982b). Salinity and temperature tolerance of Hydrococcus brazieri (T. Woods, 1876) and Arthritica semen (Menke, 1843) from the Peel-Harvey estuarine system, Western Australia. Journal of the Malacological Society of Australia 5(3–4), 151–156. Williams RJ and Watford FA (1997). Identification of structures restricting tidal flow in New South Wales, Australia. Wetlands Ecology and Management 5, 87–97. Wilson BR (1968) Survival and reproduction of the mussel, Xenostrobus securis (Lam.) (Mollusca: Bivalvia; Mytilidae) in a Western Australian estuary. Part I: Salinity. Journal Natural History 2, 307–328. Wilson BR (1969). Survival and reproduction of the mussel, Xenostrobus securis (Lamarck) (Mollusca: Bivalvia; Mytilidae) in a Western Australian estuary. Pt. II: Reproduction, growth and longevity. Journal Natural History 3, 93–120. Wilton KM (2002). Coastal wetland habitat dynamics in selected New South Wales estuaries. PhD thesis, Australian Catholic University, Australia. Wong P (2002). Distribution and abundance of Salinator solida in microhabitats. BSc (Hons) thesis, University of Sydney, Sydney, Australia.
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Woodroffe CD (1990). The impact of sea-level rise on mangrove shorelines. Progress in Physical Geography 14, 483–520. Yerman MN and Ross PM (2004). Landscape issues for the macrofauna in temperate urban mangrove forests. In Urban Wildlife: More Than Meets the Eye. (Eds D Lunney and S Burgin) pp. 205–210. Royal Zoological Society of NSW: Mosman, NSW. Yonge CM (1957). Enigmonia aenigmatica Sowerby, a motile anomiid (saddle oyster). Nature, London 180, 765–766. Yonge CM (1977). Form and evolution in the Anomiacea (Mollusca: Bivalvia) – Pododesmus, Anomia, Patro, Enigmonia (Anomiidae): Placunanomia, Placuna (Placunidae fam. nov.) Philosophical Transactions of the Royal Society (Ser. B) 276(950), 453–527. Zann LP (1995). ‘Our sea, our future. Major findings of the State of Marine Environment Report for Australia’. Department of Environment, Sport and Territories: Canberra. Zedler JB, Nelson P and Adam P (1995). Plant community organization in New South Wales saltmarshes: Species mosaics and potential causes. Wetlands (Australia) 14, 1–18. Zedler JB, Paling E and McComb A (1990). Differential salinity responses to help explain the replacement of native Juncus kraussii by Typha orientalis in Western Australian salt marshes. Australian Journal of Ecology 15, 57–72.
CHAPTER 5
Ecology of burrowing crabs in temperate saltmarsh of south-east Australia Debashish Mazumder
Introduction Saltmarshes are considered to be important coastal habitats because of their role in filtering surface water prior to its entering the sea, their contribution to coastal productivity (Morrisey 1995), and because they are a source of organic material and nutrients for a wide range of marine communities (Boorman 1999). One important visual feature of a saltmarsh is the presence of a large number of crab burrows, and this indicates an abundance of crabs within the saltmarsh environment. Crabs inhabiting saltmarshes excavate burrows over extensive areas, profoundly modifying the physical structure of the environment (Jones et al. 1994, 1997). The excavation activities of crabs and the resulting burrows may have important ecological significance on ecosystem functioning. Results from a mangrove habitat study found that burial of plant detritus by the excavation activities of sesarmid crabs, or litter directly pulled into their burrows, enhance the heterogeneity and thereby the efficiency of microbial decomposition in subsurface mangrove sediments (Kristensen 2008). Crabs living in the mangrove habitat are relatively well studied compared to those occupying the saltmarsh, and recognised for their role in contributing to the structure and function of mangrove habitats through burrowing and feeding activities (Warren and Underwood, 1986; Smith 1987). Mangrove crabs are also recognised for their role as food for higher-order predators (Robertson 1988) and their contribution to the foodweb through processing of leaf litter into more palatable forms, thereby contributing to nutrient cycling and energy flow (Lee 1995, 1997; Skov and Hartnoll 2002). By contrast very little is known about the ecology of crab species in temperate Australian saltmarsh, and in particular the degree to which saltmarsh crabs support the adjacent estuarine foodwebs.
Sampling of crabs in saltmarsh The most widely employed survey methods used to estimate the diversity and distribution of crabs living in saltmarsh in Australia, include visual census (Nobbs and McGuiness 1999; Golley et al. 1962; Nakasone 1982; Warren 1990; Hagen 1993) and burrow counts (MacFarlane 2002; Kerwin 1971; Aspey 1978; Krebs and Valiela 1978; Breitfuss 2003). Visual census involves an observer counting crabs on the sediment surface. This method is relatively simple and non-destructive, but it has disadvantages because the census is dependent on the emergence of crabs from burrows and their activity on the surface. Factors influencing 115
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the number of crabs observed on the surface include the length of the observation time, the distance of the observer from the crab burrow, the species of crab being observed, environmental variability (including precipitation) and vegetation type (Nobbs and McGuinness 1999). Moreover, the presence of a human observer in the field may have a profound impact on the estimation of the crab population (MacFarlane and King 2002). Those species of crab which are most accurately estimated using this visual census technique, are those that maintain burrows and are visible on the sediment surface of the saltmarsh during daylight hours, and occupy habitat where vegetation is sparse. The semaphore crab, Heloecious cordiformis, is easily observed using this method. Some other species may be excluded from visual census due to their less frequent emergence from burrows. Although Warren (1990) and MacFarlane (2002) both found a good correlation between the density of crab burrows and crab density, others have found the counting of burrows as an estimate of crab density to be limited (Griffin 1968; Christy 1982). Those species of crabs which do not maintain burrows, share, and exchange or use interconnecting burrows are difficult to accurately estimate using the density of burrows. For example, many species of Ocypodid seal their burrows at the end of a period of activity on the surface and may not emerge for days, weeks, or even months (Colby and Fonseca 1984; Warren 1987). In an attempt to standardise sampling strategies to estimate the density and diversity of crabs in a temperate saltmarsh, Mazumder and Saintilan (2003) employed four different sampling techniques including visual census, counting of burrows, pit traps, and artificial substrata (rectangular-shaped hollow brick blocks, placed in the saltmarsh) to estimate crab populations in a temperate Australian saltmarsh. We found that the most effective sampling strategy varied depending on the species of crab and the vegetation type. A visual census was the most effective means of estimating the density of H. cordiformis and Parasesarma erythrodactyla, although this was ineffective in identifying and estimating crabs in heavily vegetated Sporobolus virginicus and Juncus kraussii communities. Although the counting of burrows is quick, this method gives no information on species diversity. The accuracy of employing burrow counts also depends on the separation of used and unused burrows (Breitfuss 2003). Mazumder and Saintilan (2003) found that pit-traps were an effective means of comparing the density of P. erythrodactyla between saltmarsh habitats, although they may have underestimated the density of H. cordiformis in the saltmarsh. In summary, a combination of pit-trap and visual census was thought the most comprehensive strategy for surveying crabs in saltmarsh (Mazumder and Saintilan 2003).
The distribution of crabs in the saltmarsh Two families of shore crabs (Grapsidae and Ocypodidae) are common in the saltmarsh (see Plate 5.1, page 51). The common saltmarsh crab species Helograpsus haswellianus, Paragrapsus laevis and Parasesarma erythrodactyla belong to family Grapsidae while Heloecius cordiformis and Austaloplax tridentata belong to the Ocypodidae family. Grapsidae may be identified by their wide fronts, short eyes and almost square-shaped carapace. Most Grapsids have equalsized chelae (claws). Male Grapsids usually have much larger chelae than the females. The Ocypodids have distinctive stalked eyes close together at the front of the carapace. They are active burrowers and many live in colonies. The male crabs have unequal sized chelae (Edgar 2000). Both Grapsidae and Ocypodidae are generally tolerant of both dry and deoxygenated conditions and are common in shore habitats. Little is known about the biogeography or spatial variability of crabs in saltmarsh. At least five species of shore crabs have been reported as occupying temperate saltmarsh. Morrisey (1995) reported three species of crabs in the saltmarsh and mangrove in Botany Bay and Pittwater, New South Wales, these being Helograpsus haswellianus, Paragrapsus laevis and
Ecology of burrowing crabs in temperate saltmarsh of south-east Australia
Heloecius cordiformis. Mazumder and Saintilan (2003) added Parasesarma erythrodactyla to the species common to the saltmarsh of Botany Bay, and also captured Scylla serrata from the saltmarsh in an unpublished survey. In the Sydney region, H. haswellianus appears to be the most abundant species in saltmarsh and P. erythrodactyla the most abundant species in mangrove, although some studies have reported an abundance of Heloecius cordiformis in the saltmarsh (Warren and Underwood 1986; Warren 1987). Freewater et al. (2007) reported H. haswellianus and P. erythrodactyla to be the most abundant saltmarsh species in Cockle Bay saltmarsh at Brisbane Water, New South Wales. H. haswellianus appears a common intertidal crab in New South Wales (Morrisey 1995; Mazumder and Saintilan 2003), Tasmania (Richardson et al. 1998), and south-east Queensland (Skilleter and Warren 2000; Breitfuss 2003). In South Australia, H. haswellianus was found in saltmarsh and mangrove habitats (Katrak 2007 pers. comm.; Imgraben and Dittman 2008). Similarly, P. erythrodactyla is found in both saltmarsh and mangrove habitats in Queensland (Guest and Connolly 2004), and H. haswellianus and Australoplax tridentata (Ocypodidae) are commonly encountered throughout saltmarsh and mangrove in south-east Queensland, Australia (Snelling 1959; Skilleter and Warren 2000). Portunids crabs (e.g. the mud crab, Scylla serrata), are quite different from the shore crabs and are usually rare in saltmarshes. The mud crab Scylla serrata is typically found in sheltered estuaries, tidal reaches, mud flats and mangrove forests and is an important commerciallyharvested species. However, burrows of Scylla serrata are also found in the saltmarsh of subtropical Australia and Northern Territory. Saintilan (2004) reported that the commercial catches of Scylla serrata correlated more closely with mangrove and saltmarsh areas combined than mangrove area alone for the estuaries of New South Wales. The suggestion was that saltmarsh may represent an important habitat for the species at a state-wide scale. Little information is available concerning introduced crab species. The Green crab, Carcinus maenas was first recorded in Port Phillip Bay, Victoria, in the late 1800s, apparently introduced in the dry blast ballast of wooden vessels from Europe (Fulton and Grant 1902). Since introduction, the species is believed to be distributed widely and found along the south-east coast of Victoria (MRGVM 1984), southern New South Wales (Hutchings et al. 1989), South Australia (Zeidler 1978; Rosenzweig 1984) and Tasmania (Gardner et al. 1994). The European green crab utilises a variety of habitats in its native range including hard and soft substrates, from protected embayments to moderately exposed rocky shores (Grosholz and Ruiz 1996). A port survey by Thresher et al. (2003) reported the presence of C. maenas in south-eastern Australia (Tasmania, Victoria and southern New South Wales) in areas ranging from barren sand and fine silt to heavily vegetated habitats. The distribution and abundance of crabs in a habitat are influenced by many factors including availability of food (Murai et al. 1982; Genoni 1991), grain size of sediment (Griffin 1971; Sasekumar 1974; Waren and Underwood 1986), density of root materials (Ringold 1979; Bertness and Miller 1984), availability of water (Crane 1975; Yates 1978) and tidal flushing (Breitfuss et al. 2004). Some research suggests that the dense impenetrable root systems of saltmarsh plants could limit the burrowing species assemblages (Marsh 1982; Laegdsgaard 2006). To what extent the plant communities in saltmarsh shape the assemblages of crabs is unknown. Richardson et al. (1998) reported that crab assemblages in a Tasmanian saltmarsh were not related to individual plant species. However Mazumder (2004) found that crab abundance varied between vegetation types with H. cordiformis found in greater numbers in saltmarsh dominated by Sarcocornia quinqueflora, and H. haswellianus found in larger numbers in saltmarshes dominated by Sporobolus virginicus and Juncus kraussii. Crabs construct burrows generally for sheltering, protection from predation and dessication as well as breeding and moulting (Dunham and Gilchrist 1988; Morrisey et al. 1999). Burrow aperture is mostly close to the size of resident crabs, and depth of burrow varies with
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species of crabs. For example, burrows of H. haswellianus extended to depths of 30–40 cm (Marsh 1982), while the burrows of Sesarma messa and Alpheus cf macklay extended to a depth of 1.2 m in the mixed Rhizophora stylosa and Ceriops tagal habitat in Queensland (Hollins 2001). Soilwater content of the substrate has been associated with crab burrowing and varied with species. Some species of crabs (e.g. H. haswellianus) prefer well-drained substrates and avoid wet conditions (Marsh 1982, Richardson et al. 1997). Both Parasesarma erythodactyla and Australoplax tridentata prefer wet watery conditions for burrowing (Snellings 1959). Distribution of burrows in saltmarsh generally depends on size classes of crabs and their preferences on soil moisture conditions. Breitfuss et al. (2004) found few small burrows of H. haswellianus in the range of 6–10 mm diameter present in the lower shore, while larger burrows in the range of 21–25 mm were distributed across the habitat. Research on burrow excavation shows that activity of crabs affected sediment topography and biogeochemistry by modifying particle size distribution, drainage, redox conditions and organic matter as well as nutrient availability (Mouton and Fleder 1996; Botto and Iribarne 2000; concept of ecosystem engineering: Jones et al. 1994, 1997). Research also found burrow excavation turns over marsh sediments (Katz 1980; Montague 1982; McCraith et al. 2003), bringing buried material to the marsh surface which was subsequently exported by tidal flushing. Studies of the movement patterns of crabs in the habitat found that the patterns of movement vary among species. In a northern hemisphere example, the non-burrowing crab, Hemigrapsus sanguineus (Grapsidae) was found to move on an average about 7 m over a 24 hour period (Brousseau et al. 2002), and Pachygrapsus marmoratus (Grapsidae) was found to move over an area of 10 m 2 (Cannicci et al. 1999). Research conducted to understand the crab movement by using mark-recapture method in the saltmarsh-mangrove interface in subtropical Australia (Guest et al. 2006) found the majority of crabs (91% for P. erythrodactyla and 93% for A. tridentata) moved less than 2 m from the place of initial capture. This suggests resident crabs most likely depend on autochthonous sources for their diet, as discussed later in this chapter.
Breeding ecology of crabs Although several studies have commented on the breeding period of crabs in the Sydney district, no study has yet examined this systematically. Warren (1987) found ovigerous H. cordiformis in the mangrove during spring and summer, although Mazumder and Saintilan (2003) failed to find any ovigerous H. cordiformis at Towra Point over several summer months, despite it being the most abundant species. The gravid females of different species appear to have different levels of activity. Warren (1987) also found ovigerous P. laevis in the winter months at the seaward zone of mangroves in Sydney, with P. erythrodactyla breeding mostly in the summer months when the higher spring tides coincide with daylight. Saintilan (2005) also reported that P. erythrodactyla sampled from Powells Creek saltmarsh in the Sydney Olympic Park in summer months were all ovigerous females. Observation by Mazumder (2004) over a two-year period at Towra Point saltmarsh in NSW revealed mature H. haswellianus had a prolonged breeding period, from March to October; P. erythrodactyla was ovigerous during November to February, while P. laevis had a short breeding period from July to August. Breeding ecology of these species may be synchronised with temperature and tide. H. haswellianus breeds during autumn/winter when the higher amplitude diurnal tides floods in the saltmarsh occur at night. The greatest numbers of gravid females are found in the coolest months of May, June, July and August (Mazumder 2004).
Ecology of burrowing crabs in temperate saltmarsh of south-east Australia
Larval release It is now well established that the breeding behaviour of crabs inhabiting saltmarshes is synchronised with tidal amplitude. Studies in the 1980s in the northern hemisphere found that newly hatched larvae of fiddler crabs were transported into the main estuary from the tidal creeks (Epifanio et al. 1988). Development of the zoeal stages occurs in the lower regions of large estuaries like Chesapeake and Delaware Rivers (Sandifer 1975; Dittle and Epifanio 1982) or on the adjacent continental shelf (Christy and Stancyk 1982). Crabs living in temperate Australian saltmarshes release larvae when high spring tides flood the habitat (see Plate 5.2, page 52), and larval release from one species or another occurs in almost every month. A transect study in saltmarsh, mangrove, seagrass and open water habitats at Towra Point during spring high tides demonstrated that crab larval densities were substantially higher in the saltmarsh than in other habitats due to the introduction of newly hatched crab larvae from Helograpsus haswellianus which lives primarily in saltmarsh (Mazumder et al. 2008; see Figure 5.1 and Plate 5.2). Density of other zooplankton categories remained constant between these habitats. The lesser crab larval densities in the mangrove, seagrass and open water habitats may be the result of dilution at the point of sampling (Mazumder 2004). A detailed year-round study (Mazumder et al. 2006) examined crab larval abundance between incoming and outgoing tides at Towra Point. Higher numbers of crab larvae were present in the outgoing tide (an average 2125 per cubic metre of water) compared to an average of 4 per cubic metre of water in the incoming flood tide. This was true of all months, though with a strong winter peak and a smaller summer peak (see Figure 5.2). These corresponded to times when H. haswellianus and P. erythrodactyla respectively were ovigerous. Similar patterns were found at Allen’s Creek on the Hawkesbury River (Mazumder 2004; see Figure 5.3). A recent study at a nearby estuary (Freewater et al. 2007) also found that crabs living in the saltmarsh-mangrove complex released significantly higher numbers of larvae in the ebbing tides.
Zooplankton (indiv. m–3)
7000
Other zooplankton Crab larvae
6000 5000 4000 3000 2000 1000 0 A
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B
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Figure 5.1 Mean (+SE) crab larval abundance in different locations from saltmarsh to open water at Towra Point, NSW, Australia.
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100000
No. of crab larvae per m3
10000 1000 100 10 1 0.1
Mar- Apr- May- Jun- Jul- Aug- Sep- Oct- Nov-Dec- Jan- Feb- Mar- Apr- May-Jun- Jul- Aug01 01 01 01 01 01 01 01 01 01 02 02 02 02 02 02 02 02
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Figure 5.2 Mean (± standard error) abundance of crab larvae in flood and ebb tide at saltmarsh, Towra Point, Botany Bay.
Although larval release coincides with high spring tides, it does not necessarily occur on the first spring tide in a month which covers the saltmarsh. It may be that the first tide is a signal for release of larvae on the subsequent high tide. Mazumder (2004) found very few crab larvae present on the first high tide in a series of tides sampled at Towra Point, but a substantial peak in abundance on the second spring tide of the month. Larval release in subtropical Queensland has been correlated with the feeding of fish. Hollingsworth and Connolly (2006) found lower abundance of crab larvae in the stomach of common estuarine glassfish, A. jacksoniensis, on the first night that the marsh was inundated than on subsequent nights. Freewater et al. (2007) also found significantly greater concentrations of crab zoeae on ebb tides on the second and third days of the spring tides in Cockle Bay saltmarsh at Brisbane water.
No. crab larvae per m3
120
1000 900 800 700 600 500 400 300 200 100 0 August
December
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Figure 5.3 Mean (± standard error) abundance of crab larvae in in-coming and out-going waters of saltmarsh at Allens Creek, Hawkesbury River, Spencer, NSW, 2001/02.
Ecology of burrowing crabs in temperate saltmarsh of south-east Australia
Box 5.1 Examining diets of crabs using the Isotope Tracer Technique Analyses of naturally occurring stable isotope ratios of carbon (C), nitrogen (N) and sulfur (S) are commonly used to trace the source of food (Connolly et al. 2004) and the transformation of energy in the ecosystem. Over the last decade, stable isotopes are increasingly used in ecological studies. Conventionally, trophic linkages (who is eating who) have been determined by gut content analyses to identify prey-predator associations. Although gut content analyses provide valuable taxonomic information on dietary items, the method has difficulties as it relies on visual observations of partially digested material, which may be difficult to identify. Furthermore, not all ingested materials are assimilated (Michener and Schell 1994) and some ingested materials are assimilated very quickly and are therefore rarely found in the stomach (Gee 1989). Stable isotope analyses provide a complementary tool, tracing food sources and foodweb analysis using chemically validated information. There are some advantages in both methods: gut content shows the range of ingested items in the stomach; and isotope analysis provides a time-integrated picture of major dietary sources. To understand better resolution of dietary patterns both methods should be used together. Stable isotopes are different naturally occurring forms of elements, varying in the number of neutrons present in the nucleus. There are two naturally occurring atomic forms of carbon (13C and 12C), nitrogen (15N and 14N) and sulfur (34S and 32S). Biota assimilate both forms of C, N and S, and the ratio of 13C to 12C (termed d13C), 34S to 32S (termed d34S) and 15N to 14N (termed d15N) can be determined by an analysis of tissue using an isotope ratio mass spectrometer. The ratios are expressed in parts per thousand (‰) which represents the ratio of heaver to lighter isotopes. This technique provides valuable information on the sources of nutrients and the trophic level (position in the foodweb) of any animal in an ecosystem (Szymczak and Mazumder 2007). Ecological applications of stable isotope analysis rely on the prey’s distinct isotope ratios. These ratios are transferred to consumer tissues upon consumption. There is an increase in the proportion of carbon-13 (13C/12C ratio), sulfur-34 (34S/32S ratio) and nitrogen-15 (15N/14N ratio) of the organism due to preferential metabolic loss of carbon-12, sulfur-32 and nitrogen-14 during food assimilation, excretion and growth. An organism is typically enriched in heavier 13C, 34S and 15N relative to its diet. This small increase in the ratio of the heaver (13C, 15N and 34S) isotope from prey to predator is called trophic fractionation. Trophic fractionation is small for carbon (average + 0.8 to 1.1‰, DeNiro and Epstein 1978) and the trophic fractionation for sulfur is negligible (Peterson and Howarth 1987; Hesslein et al. 1993) though it differs between high-protein (approximately 2.0 ±0.65‰) and low-protein diets (–0.5 ± 0.56‰, McCutchan et al. 2003). Trophic fractionation for nitrogen is higher (3.5‰, Minawaga and Wada 1984). Because of these generalities, carbon and or sulfur isotope composition is used to determine the sources of food for an animal in an ecosystem (Cattaneo et al. 2004) whereas nitrogen on the other hand is used to estimate the trophic position of organisms in the ecosystem.
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Very small amounts of tissue samples (microgram to milligram levels) are used to obtain the isotopic ratios of the sample. In the laboratory, tissue samples are dried and ground to a fine powder. Powdered samples are then loaded into tin capsules, crimped and placed in micro-plates for subsequent isotopic analysis using a continuous flow stable isotope ratio mass spectrometer (EA/IRMS). Stable isotope values are reported in delta (d) units in parts per thousand (‰) relative to the international standard.
Diet of crabs In tropical mangrove environments, research findings showed Grapsidae crabs contributed substantially to nutrient cycling and energy flow through leaf litter degradation (Robertson and Daniel 1989). Research on Indo-Pacific mangroves (reviewed in Lee 1998) found crabs consumed leaf litter and produced leaf litter by-products supporting detritus-based food chain in the estuary. Saltmarsh plants seem to be similarly utilised by crabs. Kreeger and Newell (2000) reported that invertebrates in the saltmarshes of the Atlantic coast of North America utilised carbon directly from Spartina alterniflora. Guest and Connolly (2004), at an Australian east coast subtropical location, found d13C values of two saltmarsh crabs (Parasesarma erythrodactyla and Australoplax tridentata) to closely match the value of salt couchgrass Sporobolus virginicus. However, there is also evidence to suggest that unicellular algae are an important nutritional source for many crab species. Sullivan and Moncreiff (1988) found that benthic microalgae appeared the dominant resources for invertebrate consumers in the Gulf of Mexico coastal marshes. In Australia, Breitfuss et al. (2004) reported P. erythodactyla feeding on algal deposits following daytime high tide. Some isotopic studies investigated the distance that crabs move to derive their dietary nutrition and found that crabs ultimately derive their nutrition from local sources (Marguillier et al. 1997; Deegan and Garritt 1997; Bouillon et al. 2004; Guest and Connolly 2004), although the scale of the local sources measured by these studies varied considerably. In a subtropical Queensland study, Guest et al. (2004) found Australoplax tridentata (Ocypodidae) and Parasesarma erythrodactyla (Grapsidae) obtained their carbon from the immediate vicinity in saltmarsh, within a few metres (Guest et al. 2006). It might be expected, from the evidence of Guest and Connolly (2004) that along the temperate coastline of eastern Australia, S. virginicus would contribute significantly to the diet of saltmarsh crabs. Guest et al. (2006) recently found more depleted d13C values for detritus (-18.5‰, ±0.6‰) than P. erythrodactyla (–16.2‰) and A. tridentata (–15.2‰) suggesting very small detrital fragments along with microphytobenthos which was more enriched in d13C, were assimilated in the crabs tissue in subtropical Queensland saltmarsh. Saltmarsh size may also influence pathways of carbon supply to crabs. Guest and Connolly (2006) used carbon stable isotopes to determine the influence of the size of saltmarsh patches on the trophic contribution of saltmarsh grass S. virginicus to resident crabs (P. erythrodactyla and A. tridentata). They found the size of saltmarsh patches had a significant effect on d13C values of P. erythrodactyla and A. tridentata. Under habitat fragmentation, in patches less than about 0.5 ha, the carbon pool is swamped by non-marsh sources and thus changed the energy pathways to crabs, switching from local to imported sources. Considering the differences of saltmarsh plant communities in the geographical settings as well as the size of saltmarsh patches, the diet of crabs may show regional differences.
Ecology of burrowing crabs in temperate saltmarsh of south-east Australia
Figure 5.4 Stable carbon and nitrogen isotope signatures of crabs and saltmarsh primary producers at Towra Point, 2006.
In temperate NSW, the dominant saltmarsh plant species are S. virginicus and S. quinqueflora, with Juncus kraussii and Suaeda australis also forming extensive stands (Clarke and Hannon 1967; 1969; Adam et al. 1988). Mazumder et al. (submitted) employed stable isotope techniques to understand the dietary relationships between saltmarsh dwelling crabs and their surroundings at Towra Point, NSW. Species of saltmarsh plants and crabs that are common to the temperate saltmarsh in the Sydney metropolitan region, along with fine benthic matter collected through scraping sediment, were analysed for stable isotopes of carbon and nitrogen. The results found that crabs living in east coast Australian temperate saltmarsh are less likely to be herbivores. Unlike the south-east Queensland study by Guest and Connolly (2004), plant detritus contributed very little to the diet of resident crabs in temperate saltmarsh in NSW. The carbon isotope values for H. cordiformis (δ13C –19.85‰), P. erythrodactyla (δ13C –19.87‰) and H. haswellianus (δ13C –21.5‰) were close the value of fine benthic material (δ13C –21.22‰). These results suggested the diet of the three crab species from the saltmarsh at Towra Point appears to rely on microphytobenthos (such as diatoms) and other unknown resources, as a primary food source. However, further detailed investigations are needed and there is evidence to suggest that S. virginicus contributes to the diet of P. erythrodactyla at Towra Point also (Mazumder, unpublished data). Among three common crab species studied, two species, H. cordiformis and P. erythrodactyla had almost identical carbon isotope values which suggested that both crab species access a similar carbon pool in the saltmarsh habitat. To summarise, in subtropical Queensland Australia, S. virginicus appears to be the primary source of dietary carbon for the crab P. erythrodactyla, with carbon isotope ratios (δ13C) of –15.5‰ for S. virginicus and –15.7‰ for P. erythrodactyla (Guest and Connolly 2004). In NSW, saltmarsh plants are less likely to contribute trophically (with the possible exception of
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S. virginicus where dominant: Mazumder unpublished data). However, they may be an important structural habitat and microclimate for crabs.
Conclusion In recent years, saltmarshes have received attention from policy makers and conservationists, and heightened concern is justified given the emerging evidence of their crucial role in supporting estuarine foodwebs (Hollingsworth and Connolly 2006; Mazumder et al. 2006). The extent of the Australian research on subtropical and temperate saltmarsh has answered some important questions regarding the functionality of this ecosystem, but has raised other research questions. The abundances and diversity of crabs resident in temperate Australian saltmarshes suggest saltmarshes are an important habitat for benthic crabs. Besides the provision of crab habitat, part of the conservation value of saltmarshes lies in their role as a net exporter of crab larvae and their contribution to estuarine foodwebs. Limited studies on Australian saltmarshes have found a great deal of similarity between temperate and subtropical conditions in terms of crab species, larval release, and dietary contribution to fish. Crabs living in saltmarsh release their larvae in almost all months. A peak in summer and winter seasons could be related to zooplanktivorous fish recruitment and growth in the estuary, an issue inviting further investigation. Larval release and density in the outgoing tide is also related with tidal velocity and the degree of marsh inundation. This issue also requires further investigation as predicted climate change may alter flow patterns and inundation characteristics. Stable isotope analysis reveals that fine benthic matter most probably consisting of both local and imported sources contributes to the diet of resident crabs in NSW saltmarsh whereas S. virginicus is the primary diet of crabs sampled in subtropical Queensland, implying a regional difference in food chain structure. Further investigation is required to understand the detailed trophic links of crabs with local sources (e.g. saltmarsh plants and others) and imported sources (e.g. tide transported materials). Besides dietary dependency, whether and to what extent saltmarsh vegetation provides other services to resident crabs is also important to know for appropriate protection and conservation of saltmarsh.
Acknowledgements I am grateful to Neil Saintilan, Rob Williams, Rod Connolly and Pauline Ross for reviewing the manuscript and their valuable comments and suggestions for its improvement. My sincere thanks to Suzanne Hollins and Ron Szymczak for their assistance and support. Finally, I would like to thank my wife Heaven and my daughter Srestha for their encouragement.
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CHAPTER 6
Fish on Australian saltmarshes Rod Connolly
Introduction Saltmarshes provide important habitat for fish on all inhabited continents. Fish are a very important aspect of the biodiversity of marsh systems, and the role of saltmarsh in the provision of fish habitat is one of the main reasons why humans value saltmarsh at all. Fish living on marshes or visiting the inundated habitat at high tide are abundant and diverse. Swimming crustaceans such as shrimp and prawns (which together with fish are collectively known as nekton) also occur on saltmarsh, and are included in this Chapter because of the similarities in aspects of their behaviour. This Chapter focuses on Australian saltmarshes as fish habitat, but our early understanding of fish use of saltmarsh came from studies done elsewhere. A review of all saltmarsh nekton research prior to 2000 (Connolly 1999) found the literature to be overwhelmingly North American (90% of the 113 studies), with surprisingly few papers from Europe (7%) given the large number of botanical studies undertaken there (Adam 1990). Only 3% of papers were from southern hemisphere marshes, all from Australia (see Table 6.1). Patterns in the use of saltmarsh by nekton are thus best described for North American marshes (Kneib 1997a). Large numbers of certain small species such as killifish (Fundulus spp.) and grass shrimp (Palaemonetes spp.) are resident on marshes. Numerous other fish and crustacean species visit the inundated marsh as transients (Kneib 1997a). Species using the marsh flat are mainly resident on or near the saltmarsh for their entire lifecycle, while fish congregating around the edge of the saltmarsh are juveniles of species that spawn elsewhere in the estuary or in oceanic waters (Peterson and Turner 1994). A review of the value of saltmarsh as nursery habitat, taking into consideration abundances, growth rates and survival, found that nursery value was greatest for vegetated marsh, particularly at the marsh edge, and lower for unvegetated marsh (Minello et al. 2003). Encroaching human development is resulting in the fragmentation of saltmarshes in many parts of the world (Adam 2002). Where saltmarsh supports major fisheries, the consequences of habitat fragmentation are likely to be large. The marshes of the Gulf coast of the USA, for example, are considered critical nursery habitat for brown shrimp, Penaeus aztecus. A combination of empirical data and numerical modelling of survival rates demonstrates that, initially, brown shrimp productivity increases as saltmarshes decline in extent and fragment into smaller units (Browder et al. 1989; Haas et al. 2004). For a time, these smaller units increase the length of the interface between marsh and water, increasing the linear extent of the marsh edge, the habitat preferred by prawns. However, modelling shows that, ultimately, the amount 131
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Table 6.1 Summary of the geographic effort into research on saltmarsh nekton to year 2000, showing the paucity of Australian studies relative to the area of saltmarsh on this continent (from Connolly 1999). States within USA are ordered by number of studies. Several Australian studies have been published since the review, along with a larger number of recent studies from North America and Europe. Location
Number
% of total
Georgia
14
12
Louisiana
13
12
North Carolina
12
11
Virginia
12
11
Texas
8
7
New Jersey
8
7
Florida
8
7
South Carolina
7
6
15
13
97
86
Other USA states Total USA Canada Total North America
5
4
102
90
Europe
8
7
Australia
3
3
113
100
Total
of marsh relative to open water will decrease to the point where shrimp productivity begins to decline again (Browder et al. 1989). The tidal hydrology of marshes implies strong linkages between marsh and adjacent habitats for mobile animals such as fish and swimming crustaceans (Odum 1995, Rozas 1995). For many species, therefore, saltmarsh is just one of multiple habitats that might be used by individuals over short (one tidal cycle) or long (different parts of the lifecycle) timeframes. The link that nekton provide among habitats has become central to the debate around the outwelling concept. Outwelling describes the transfer of organic matter produced in high intertidal habitats such as saltmarsh to adjacent, deeper-water habitats, where it supports high rates of secondary production (Odum 1968). Outwelling was originally conceived as transfer of particulate or dissolved organic matter (Teal 1962). The emphasis more recently has been on the numerous predator/prey interactions that potentially result in a net transfer of organic matter from intertidal to subtidal habitats, in a process known as trophic relay (Kneib 1997a). Australian saltmarshes typically occur landward of mangrove forests, high in the intertidal zone, and have shorter and less frequent periods of inundation than marshes on the Atlantic and Gulf coasts of the USA, which generally lack mangroves and extend down to the midintertidal zone (Adam 1990). The vegetation of Australian saltmarshes is dominated by succulent herbs and grasses that are considerably shorter than the stands of cordgrass (Spartina spp.) dominating northern hemisphere saltmarshes (Adam 1990). These important physical differences mean that ecological patterns and processes for fish occurring on North American marshes might not apply in Australia (Connolly 1999). Although Australian work remains under-represented in the literature relative to the cover of marsh (about the same extent as in the USA), there have been several local studies since Connolly’s (1999) review, and there are now enough data to form useful conclusions about fish on Australian marshes. This Chapter first reports on fish assemblages of Australian salt-
Fish on Australian saltmarshes
marshes in general, followed by specific sections detailing fish distributions in different marsh microhabitats, their feeding behaviour, how fish can be sampled, and directions for future research.
Species and abundances on Australian saltmarshes Beginning in 1986, patterns in fish abundances associated with Australian saltmarshes have been described in a total of 11 papers, mostly in temperate and subtropical waters rather than tropical waters (see Table 6.2). Early work in Australia sampled water in creeks draining marshes rather than the inundated marsh flats themselves (see Table 6.2). Fish assemblages in tidal creeks in saltmarsh systems include virtually all of the species now known to occur on the marsh flats themselves, but occasionally also include additional, larger species common elsewhere in estuaries (Gibbs 1986; Morton et al. 1987; Davis 1988). The development of the pop net technique for quantitatively sampling nekton from vegetated saltmarsh in the mid-1990s (Connolly et al. 1997) paved the way for several subsequent studies that increased the geographic spread and total amount of information about abundances of fish on saltmarsh. Fish assemblages on inundated Australian marshes are dominated by adults of one or two small species (60–90% of total abundance). These species are usually from the families Ambassidae (subtropical and temperate), Atherinidae (temperate) and Gobiidae (all waters). Very high densities of commercially important species such as banana Table 6.2 Summary of published research on fish assemblages on Australian saltmarshes. Habitats are: inundated flats (flats), intertidal creeks (creeks), and semi-permanent pools (pools). Densities shown only for quantitative sampling of inundated marsh flats, all using pop nets except Crinall and Hindell 2004. Region
State
Habitat
Method
Density (fish.100m –2 )
Temperate
SA
Flats/creeka
Pop/fyke
4
Connolly et al. 1997
Flats
Pop
1–10
Bloomfield and Gillanders 2005
Vic
Flats
Seine
25
Crinall and Hindell 2004
NSW
Creek
Dip
Flats
Pop
Flatsb
Fyke
Subtropical
Tropical
Qld
NT
Reference
Gibbs 1986 56
Mazumder et al. 2005a Mazumder et al. 2006b
Creek
Fyke
Morton et al. 1987
Pools
Dip
Morton et al. 1988
Flats
Pop
2–45
Thomas and Connolly 2001
Flats
Pop
31–64
Connolly 2005
Creek
Fyke
a. Density for flats only, not creek b. Fish collected from retreating tidal waters, no density available
Davis 1988
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Australian Saltmarsh Ecology
prawns, Fenneropenaeus merguiensis, have occasionally been recorded (Connolly 2005). The remainder of the fauna comprises small numbers of up to about 20 other fish species, including juveniles of many economically important species (see Table 6.3). Because Australian saltmarshes drain fully on the ebb tide they have no equivalent to marsh residents such as the killifish of USA marshes. However, toadfish (family Tetraodontidae) are a particularly conspicuous and common component of the fish fauna on marshes around Australia, yet are rare on North American marshes. Toadfish move onto the marsh early on the incoming tide, pushing far onto the marsh in very shallow water. This strategy has the effect of increasing the likelihood of finding major prey items such as snails and crabs (Hughes 1984). The abundances of this family on Australian marshes relative to those on North American marshes might result from the different hydroperiods. The short, infrequent inundation periods on Australian marshes that prevent residency for small fish may create an opportunity for the toadfish to obtain prey relatively easily as they enter the marsh upon inundation. The total density of fish (all species combined) on saltmarsh inundated at high tide differs among studies and among locations (see Table 6.2), ranging from 1–64 individuals per 100 m2 of marsh flat. Fish species diversity on inundated marsh is higher in subtropical waters (23 species, Thomas and Connolly 2001) than in temperate waters (2–10 species at sites along the southern Australian coastline, Connolly et al. 1997; Crinall and Hindell 2004; Bloomfield and Gillanders 2005; and 14–16 species at sites around Sydney, Mazumder et al. 2005a, 2006b). Overall, densities of fish on saltmarsh are lower than in other vegetated estuarine habitats in Australian estuaries. Comparisons among habitats are difficult where different sampling methods are used, but a fair comparison can be made by considering studies using pop nets. Fish densities on saltmarsh are typically less than half that in mangroves in similar estuaries (74–187 individuals 100 m2 in south-east Queensland; Moussalli and Connolly 1998), and relatively lower again compared with densities in intertidal seagrass (600 individuals 100 m 2 in South Australia; Connolly 1994b). Different sampling methods make comparisons with densities on saltmarshes on other continents even more difficult, but the methods most similar to pop nets are drop nets and flume weirs in the USA. The density of fish on Australian saltmarshes is lower than comparable studies in the USA (e.g. 54–114 in Georgia, 100–200 individuals 100 m2 in Texas, Kneib and Wagner 1994; Rozas and Zimmerman 2000, respectively). In fact, nekton densities on USA marshes are similar to densities in Australian mangroves. Given that mangroves in Australia occur at the same height in the intertidal zone as saltmarsh in the USA, this raises the question of whether fish densities are influenced by habitat or merely by elevation. The issue of whether elevation or habitat type is important has been partly addressed using pop net sampling of both saltmarsh and mangrove habitat at the same time in a subtropical estuary (see Figure 6.1). As expected, at high tide, when mangroves have deeper water than saltmarsh, the average number of species in mangroves is higher than in saltmarsh. The important finding, however, is that this difference remains evident when mangroves are sampled before or after the high tide at times when water depth is the same as over saltmarsh at high tide (see Figure 6.1). The conclusion is, therefore, that although elevation is probably important, the habitat type does appear to have some influence, an important finding given the intimate spatial juxtaposition of saltmarsh and mangroves along much of the Australian coastline (see Figure 6.2).
Distributions on inundated marshes One of the main findings from northern hemisphere studies is that nekton are more abundant in vegetated than unvegetated marsh areas. There is no evidence of this in Australia, however,
Fish on Australian saltmarshes
Table 6.3 Occurrence of species and families of fish reported from inundated marsh flats in Australia. C = common, P = present in small numbers. From: subtropical Qld (Thomas and Connolly 2001; Connolly 2005), Temperate NSW (Mazumder et al. 2005a, 2006b), Temperate Vic/ SA (Crinall and Hindell 2004; Bloomfield and Gillanders 2005). Subtropical QLD
Temperate NSW
Ambassis jacksoniensis
C
C
Ambassis marianus
C
Atherinomorus ogilbyi
C
Family
Species
Ambassidae Atherinidae
Atherinosoma microstoma
C
Kestratherina esox
P
Lepatherina presbyteroides
P
Pseudomugil signifer
P
Belonidae
Tylosurus gavialoides
P
Clinidae
Heteroclinus adelaide
Clupeidae
Herklotsichthys castelnaui
Galaxiidae
Galaxia maculatus
Gerridae
Gerres subfasciatus
Gobiidae
C P
P P C
Arenigobius frenatus
P
Calamiana species nova
C
C
Favonigobius lateralis
C
Gobiopterus semivestitus
C
C
Mugilogobius stigmaticus
C
P
Mugilogobius paludis Pseudogobius olorum
C
Arrhamphus sclerolepis
C
Mugilidae
Aldrichetta forsteri Liza argentea
C P
Myxus elongatus
P C
Platycephalus fuscus
Pleuronectidae
Rhombosolea tapirina Gambusia holbrooki
Sillaginidae
Sillaginodes punctata
P P P P
Sillago ciliata
Tetraodontidae
P P
Poeciliidae
Sparidae
C
Mugil cephalus Valamugil georgii
C
P
Hemiramphidae
Platycephalidae
Temperate Vic/SA
P
Sillago maculata
P
Acanthopagrus australis
C
C
Tetractenos hamiltoni
C
P
Torquigener pleurosticta
C
Tetractenos glaba
P
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Australian Saltmarsh Ecology
16 14
Species richness
136
12 10 8 6 4 2 0 Saltmarsh
Mangrove
Shallow
Mangrove
Deep
Figure 6.1 Fish species richness over subtropical saltmarsh is lower than in adjacent mangroves both at high tide (deeper water in mangroves) and on ebbing tides (mangrove water depth same as saltmarsh). Values from Moussalli and Connolly (1998) in south-east Queensland (means, SE, over three consecutive months).
since no differences in species assemblages or densities were detected between vegetated marsh and unvegetated pans in a major sampling program over winter and summer in two estuaries (Thomas and Connolly 2001). It has been suggested that the inundation period on Australian marshes is too short to allow fish to move around according to habitat preferences (Thomas and Connolly 2001). Fish densities over the very extensive unvegetated pans on tropical Australian saltmarshes have not yet been reported, a major gap in our understanding given the potential link with adjacent prawn production and evidence from subtropical studies that banana prawns utilise marsh as juveniles (Connolly 2005). Another finding from northern hemisphere studies is that fish densities on inundated marsh are high near the marsh edge and decline with increasing distance from subtidal water (Kneib and Wagner 1994). In Australia, this has been studied most thoroughly in subtropical
Figure 6.2 Subtropical Australian saltmarsh (marsh grass, Sporobolus virginicus) with distinct transition to mangroves, Avicennia marina. Photo: M. Guest.
Fish on Australian saltmarshes
waters, where pop nets were released at different distances up to 400 m from subtidal water (Thomas and Connolly 2001). No relationship was detected between distance from subtidal water and either fish density or fish species composition. Different species occurred at different distances and many species were found far onto the marshes, with several species caught at the limit of sampling, over 400 m from subtidal water (see Figure 6.3). The factors influencing fish densities on Australian saltmarshes are different to those in the USA. In Australia, the two main influences are water depth and the distance from mangrove-lined feeder creeks. Higher fish densities are found with increasing water depth in pop net samples taken at high tide (see Figure 6.4). Higher densities have also been found alongside (within 20 m) rather than further from (100 m) mangrove-lined feeder creeks linking marshes with subtidal water (Connolly 2005). The importance of intertidal creeks within the marsh system is becoming increasingly clear. Kneib (2003) has shown on the Spartina marshes of Georgia, USA, that fish productivity is much higher at sites having greater than about 2000 m of linear creek edge within a radius of 200 m of the site (see Figure 6.5).
Figure 6.3. Distances fish were caught onto subtropical marsh at two locations (Meldale and Theodolite Creek) in south-east Queensland (Thomas and Connolly 2001).
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Australian Saltmarsh Ecology
Density (fish 100m–2)
138
140 120
r2 = 0.36
100
n = 30
80 60 40 20 0 0
5
10
15
20
25
30
35
Water depth (cm) Figure 6.4 Fish densities increase with water depth of sites sampled at high tide using pop nets. Redrawn from Thomas and Connolly (2001).
Figure 6.5 Productivity of resident and migrant nekton (grams dry weight per m2) is much higher at sites having greater than about 2000 m of intertidal creek edge within 200 m radius. Data from Sapelo Island marshes, by Kneib (2003).
Fish on Australian saltmarshes
Fish feeding on saltmarsh Knowing what fish go where on saltmarshes is obviously a useful first step in understanding saltmarsh nekton ecology. Unfortunately, in Australia the lack of knowledge of basic distributional patterns even until recent times has meant that few researchers have been able to tackle questions of what use fish make of their time on the marsh. We know little, for example, of predator-prey or interspecific competition relationships in saltmarsh systems in Australia. One aspect that is beginning to be better understood, however, is the role of saltmarsh in fish feeding. Early descriptions of fish diets using stomach contents were done on fish caught in creeks draining saltmarsh rather than on the marsh flats themselves. Morton et al. (1987) described the feeding behaviour of fish caught in a small creek draining one of these marshes in southeast Queensland. The marine component of the diet of the six species examined was dominated by benthic invertebrates, predominantly adult shore crabs, although some species also ate planktonic invertebrates (crab larvae and amphipods). The diets also included a range of terrestrial invertebrates, especially a striking diversity of adult insects from eight different orders. It cannot be assumed, however, that the diets described by Morton et al. (1987) are the result of feeding behaviour on the marsh itself, since it has been shown elsewhere that fish can remain in marsh creeks and feed without entering the inundated marsh (Szedlmayer and Able 1993; Le Quesne 2000). In contrast, the feeding activity of fish visiting inundated saltmarsh during high tides has been well studied internationally. Several studies have demonstrated feeding on saltmarsh by comparing stomach fullness and prey composition of fish entering and leaving marsh habitat. Studies in the USA (Rountree and Able 1992; Nemerson and Able 2004; and in brackish marshes, Rozas and LaSalle 1990) and Europe (Lefeuvre et al. 1999; Laffaille et al. 2001; 2002) have detected higher stomach fullness after fish visit marshes. These studies have recorded a range of prey types, dominated by marine invertebrates (e.g. polychaete worms, amphipods) with occasional terrestrial (insect) invertebrates. One early Australian study took a different approach and sampled fish from the brackish, semi-permanent pools that occur high on some subtropical marshes. The small fish that live in these pools feed predominantly on insect larvae that breed there (Morton et al. 1988). In Australia, stomach content analysis of fish leaving saltmarsh habitat has demonstrated feeding activity on temperate water marshes in Victoria and NSW. Fish moving over the edge of narrow marshes in Victoria feed on amphipods and hemipteran insects (Crinall and Hindell 2004). On a marsh in Sydney, NSW, the diets of different fish species varied. Juvenile yellowfin bream (Acanthopagrus australis) ate mainly adult shore crabs, whereas several other species consumed small numbers of crabs and larger numbers of other items including zooplankton and insects. The mangrove goby (Pseudogobius olorum) fed on zooplankton and insects as well as plant material. Perhaps most importantly, the extremely abundant Port Jackson glassfish, Ambassis jacksoniensis, fed predominantly on shore crab larvae (Mazumder et al. 2006). The Port Jackson glassfish has also been the subject of intensive dietary analysis on a subtropical marsh in south-east Queensland (Hollingsworth and Connolly 2006). The glassfish ate mainly crab larvae, but showed a striking temporal pattern of feeding (see Figure 6.6). In winter, the marsh is inundated only at night and only on spring tides. Glassfish visiting the marsh on the first night of a tidal cycle feed only lightly, eating a small number of a range of prey types. This inundation, however, apparently acts as a cue for shore crabs to release larvae, and on subsequent nights, glassfish eat an average of 100–200 crab larvae per fish (see Figure 6.6). Perhaps most importantly, Hollingsworth and Connolly (2006) made a particularly convincing demonstration of the importance of saltmarsh in glassfish diets using a series of other comparisons. As
139
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Australian Saltmarsh Ecology
Figure 6.6 Crab zoea abundances in glassfish (Ambassis jacksoniensis) stomachs after feeding on subtropical saltmarsh (Hollingsworth and Connolly 2006). In each monthly cycle, fish do not feed on zoea on the first night a marsh is flooded but do so on subsequent nights (values are means, SE, scale on LHS). Tidal height is shown for each night of sampling and the night before sampling. Tidal height at which marsh is inundated (2.48 m) is shown by dotted line.
well as comparing stomach contents of glassfish leaving the marsh with those of fish entering the marsh, they also examined fish in two other treatments: 1) fish collected at the same time elsewhere in the estuary that had no opportunity to visit the marsh, and 2) fish collected before and after neap high tides that inundated intertidal mudflats but not saltmarsh. Of all the comparisons, fish that had visited marsh at high tide were the only individuals to have full stomachs; all others had much lower stomach fullness indices and few if any crab larvae. The glassfish research points to a major contribution of saltmarsh habitat in fish diets not available to fish using other parts of the estuary. This could result from the limited time fish have to feed on saltmarsh in Australia. The fish do not get to dine on saltmarsh often, but when they do it seems to be a meal worth waiting for. There is obvious potential for the feeding by glassfish to result in a net transfer of organic matter from the marsh to deeper waters, assuming that a certain amount of predation on glassfish by larger fish occurs when they retreat to deeper habitats at low tide. Such a system of trophic relay has been conceptualised for Australian marshes (Mazumder et al. 2006a; Connolly and Lee 2007), but further work on predation of glassfish is required to demonstrate it.
Future research Research into saltmarsh nekton in Australia is at an early stage, with much still to be learned. The most obvious knowledge gap is in tropical marshes, where basic distributional patterns remain unknown. Several features of tropical marshes make work there more difficult. The extent of inundation is often vast, the inundation is somewhat dependent on unpredictable
Fish on Australian saltmarshes
Box 6.1
Sampling saltmarsh fish
Many theories about saltmarsh function can only be tested with quantitative sampling of nekton on marshes. This presents a real problem, however, because of the short inundation time and erect vegetation (Connolly 1999). The most common collection technique is a fyke net deployed in creeks either draining (e.g. Morton et al. 1987) or flooding (e.g. Davis 1988) a marsh (see Figure 6.7). Although fyke nets catch large numbers of fish efficiently, they cannot usually quantify nekton densities. Research needing quantitative use of fyke nets requires considerable additional effort. The nets can, for example, be deployed in creeks draining a well-defined area of marsh, measured using computerised geo-referencing tools (Connolly et al. 1997). Even in these situations, however, nothing about the distribution of nekton on the inundated marsh itself can be gleaned. Fortunately, with regard to sampling nekton on the inundated marsh flats themselves, necessity has bred invention, and several purpose-specific methods have been developed. Many of the techniques now used to sample fish in estuaries and shallow coastal waters were developed for sampling nekton from the inundated marsh (Rozas and Minello 1997). Each technique has its advantages under different circumstances.
Figure 6.7 Fyke net method of fish collection from large areas of marsh on ebbing tides. Photo: R. Connolly.
Buoyant pop nets with remotely controlled release have become popular, particularly in Australian studies where they are now used more commonly than any other technique (Connolly et al. 1997; Thomas and Connolly 2001; Mazumder et al. 2005a, 2006b; Bloomfield and Gillanders 2005; Connolly 2005). On saltmarsh, pop nets have been shown to catch a slightly different assemblage of fish to fyke nets, typically missing very uncommon species and therefore catching fewer species overall, because of the overall reduced area sampled (Mazumder et al. 2005b). Pop nets are modelled on earlier lift net designs (Rozas 1992), and consist of four mesh walls, buoyant at the top and pegged to the sediment at the bottom (Connolly 1994a). Pop nets can be used to sample small areas accurately (up to 25 m2), and
141
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Australian Saltmarsh Ecology
Figure 6.8. Pop net method of quantitative fish collection from specific areas of inundated marsh. Photo: R. Connolly.
their mobility allows multiple deployments at randomly selected locations (see Figure 6.8). They are labour intensive, however, and cannot be used in tall marsh grass without disturbing vegetation. Another technique, the drop sampler, is popular for sampling nekton on the Gulf coast of the USA (e.g. Baltz et al. 1993; Rozas and Zimmerman 2000). A drop sampler consists of a fibreglass cylinder (usually 1–2 m in diameter), with a metal skirt that cuts through vegetation and into the sediment when it is deployed by dropping swiftly from a boom on the bow of a small boat (Zimmerman et al. 1984). Animals are then removed with small nets and potentially also by pumping the trapped water through fine mesh. Drop samplers can be used in any type of vegetation, but are restricted to marsh edges where a boat can gain access. The largest device used for sampling nekton on marshes is the flume weir (Kneib 1991). A flume weir consists of a series of posts arranged so that when mesh screens are dropped into place on the posts at high tide, the structure forms a polygon sampling 100 m2 of marsh in a single event. Kneib’s (1991) system of carefully constructed walkways to the flume weirs on Sapelo Island, Georgia, USA, allows researchers access without walking on the marsh. The flume weir can be built anywhere on a marsh and, once built and allowed to settle, avoids disturbance of even the tallest marsh grass during deployment. Like the pop net, the flume weir uses the brevity of tidal inundation to advantage. Both techniques rely on fish being caught in a pit on the down-current side as the tide retreats. Although flume weirs are impressively large, they are very labour-intensive to build and cannot, therefore, be deployed easily at multiple locations. One further technique has proven useful for sampling larval and small juvenile fish and crustaceans. Small saucers embedded in sediment successfully collect these small animals that aggregate in any residual water on the marsh as the tide retreats. These small samplers are known as simulated aquatic microhabitats, or SAMs (Kneib 1997b).
Fish on Australian saltmarshes
cyclonic events, most sites require vehicular and vessel access from remote settlements, and the waters also support populations of saltwater crocodiles. It is nevertheless essential that data be collected from these tropical regions since, in Australia, human development is predicted to be most rapid there. There could be considerable excitement in determining the contribution by tropical marshes to Australia’s Northern Prawn Fishery. The harvesting of tiger and banana prawns in the shallow waters of Gulf of Carpentaria and elsewhere in northern Australia is the country’s second most valuable fishery. Juvenile prawns might well be utilising marsh and saltpan habitat, if results from south-east Queensland are indicative (Connolly 2005). Further descriptions of patterns of use of saltmarsh by fish would also be useful in southern Australian marshes, where more data are required to develop a generalised understanding. The most rigorous surveys have been in south-east Queensland, and these have demonstrated major differences in fish use of marshes among estuaries and between seasons (Thomas and Connolly 2001). Another aspect of saltmarsh that remains unknown is the degree to which the shallow water of inundated marshes offers small fish protection from predators. Although this idea has long been held for estuarine systems (Baltz et al. 1993), comprehensive surveys of predators in estuaries suggest that the degree of protection has been overstated (Sheaves et al. 2006). Addressing predation patterns should go hand-in-hand with future work on patterns of movements of fish on and around marshes. Early work by Morton et al. (1987) using fin-clipping to show that certain species tend to be recaptured in the same marsh creek over time could now be done with greater replication and efficacy using modern ultrasonic tracking methods. Our understanding of feeding behaviour of fish on Australian marshes currently relies on data from single estuaries in each of three states (Victoria, New South Wales, Queensland). Additional data from other locations would help to build a broader understanding. The contribution of marsh plants and animals to estuarine food webs will continue to be a major issue in the protection and conservation of marshes. More rigorous data is required on this issue. Chemical tracers such as stable isotopes have helped to distinguish energy (carbon) pathways through food webs in Australian estuaries (e.g. Guest et al. 2004, 2006) and have proven useful in determining fisheries food webs associated with other vegetated habitats (Connolly et al. 2005; Melville and Connolly 2005). Stable isotopes are likely to be useful, too, in confirming the fate of plant and algal production on saltmarshes in food webs on the marsh itself and in adjacent waters. Large areas of saltmarsh have been lost along the more urbanised coasts of Australia. This unfortunate loss might have been scientifically informative had researchers been able to correlate the extent of loss with local changes in fisheries catch statistics. Although great care is required to avoid misinterpreting such correlations (Lee 2004; Loneragan et al. 2005), the approach has been used for saltmarshes in the USA (Boesch and Turner 1984) and would likely be useful in the Australian context. Saltmarshes degraded by urbanisation are beginning to be restored. The science underpinning restoration of Australian saltmarsh is relatively poorly developed, and the number of sites earmarked for restoration is small. There is every likelihood that restoration will become more prevalent in the near future, and it will be important to incorporate the requirements of nekton. International as well as local science has highlighted the importance of intertidal feeder creeks in supporting fish abundances on marshes (Kneib 2003; Connolly 2005). The presence and density of such creeks will need to be carefully matched with original habitat demography for restoration to be fully successful. It will be important, too, to determine natural levels of connectivity with other habitats, so that these can be as nearly as possible emulated in restored marshes. Several studies have measured connectivity among habitats such as saltmarsh and seagrass in the USA (Irlandi and Crawford 1997; Nagelkerken and van
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der Velde 2004), but in Australia data are only now beginning to be compiled (Saintilan et al. 2007), and we remain at the stage of formulating likely theories about connectivity among estuarine habitats (Sheaves 2005).
References Adam P (1990). Saltmarsh Ecology. Cambridge University Press: Cambridge. Adam P (2002). Saltmarshes in a time of change. Environmental Conservation 29, 39–61. Baltz DM, Rakocinski C and Fleeger JW (1993). Microhabitat use by marsh-edge fishes in a Louisiana estuary. Environmental Biology of Fishes 36, 109–126. Bloomfield AL and Gillanders BM (2005). Fish and invertebrate assemblages in seagrass, mangrove, saltmarsh, and nonvegetated habitats. Estuaries 28, 63–77. Boesch DE and Turner RE (1984). Dependence of fishery species on saltmarsh: the role of food and refuge. Estuaries 7, 460–468. Browder JA, Nelson ML, Rosenthal A, Gosselink JG and Baumann RH (1989). Modeling future trends in wetland loss and brown shrimp production in Louisiana using thematic mapped imagery. Remote Sensing Environment 28, 45–59. Connolly RM (1994a). Comparison of fish catches from a buoyant pop net and a beach seine net in a shallow seagrass habitat. Marine Ecology Progress Series 109, 305–309. Connolly RM (1994b). The role of seagrass as preferred habitat for juvenile Sillaginodes punctata (Cuv. and Val.) (Sillaginidae, Pisces): habitat selection or feeding? Journal of Experimental Marine Biology and Ecology 180, 39–47. Connolly RM (1999). Saltmarsh as habitat for fish and nektonic crustaceans: challenges in sampling designs and methods. Australian Journal of Ecology 24, 422–430. Connolly RM (2005). Modification of saltmarsh for mosquito control in Australia alters habitat use by nekton. Wetlands Ecology and Management 13, 149–161. Connolly RM, Dalton A and Bass DA (1997). Fish use of an inundated saltmarsh flat in a temperate Australian estuary. Australian Journal of Ecology 22, 222–226. Connolly RM, Hindell JS and Gorman JD (2005). Seagrass and epiphytic algae support nutrition of a fisheries species, Sillago schomburgkii, in adjacent intertidal habitats. Marine Ecology Progress Series 286, 69–79. Connolly RM and Lee SY (2007). Mangroves and saltmarsh. In Marine Ecology. (Eds SD Connell and BM Gillanders) pp. 485–512. Oxford University Press: Oxford. Crinall SM and Hindell JS (2004). Assessing the use of saltmarsh flats by fish in a temperate Australian embayment. Estuaries 27, 728–739. Davis TLO (1988). Temporal changes in the fish fauna entering a tidal swamp system in tropical Australia. Environmental Biology of Fishes 21, 161–172. Gibbs PJ (1986). The fauna and fishery of Wallis Lake. In Wallis Lake: Present and Future. pp. 1–7. Australian Marine Science Association: NSW. Guest MA, Connolly RM and Loneragan NR (2004). Carbon movement and assimilation by invertebrates in estuarine habitats at a scale of metres. Marine Ecology Progress Series 278, 27–34. Guest MA, Connolly RM, Lee SY, Loneragan NR and Breitfuss MJ (2006). Mechanism for the small-scale movement of carbon among estuarine habitats: organic matter transfer not crab movement. Oecologia 148, 88–96. Haas HL, Rose KA, Fry B, Minello TJ and Rozas LP (2004). Brown shrimp on the edge: linking habitat to survival using an individual-based simulation model. Ecological Applications 14, 1232–1247. Hollingsworth A and Connolly RM (2006). Feeding by fish visiting inundated subtropical saltmarsh. Journal of Experimental Marine Biology and Ecology 336, 88–98.
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Hughes JM (1984). A note on predation by toadfish Spheroides spp. on the mangrove snail Littorina scabra. In Focus on Stradbroke: New Information on North Stradbroke Island and Surrounding Areas 1974–1984. (Eds RJ Coleman, J Covacevich and P Davie) pp. 312–314. Boolarong Publications: Brisbane. Irlandi EA and Crawford MK (1997). Habitat linkages: The effect of intertidal saltmarshes and adjacent subtidal habitats on abundance, movement, and growth of an estuarine fish. Oecologia 110, 222–230. Kneib RT (1991). Flume weir for quantitative collection of nekton from vegetated intertidal habitats. Marine Ecology Progress Series 75, 29–38. Kneib RT (1997a). The role of tidal marshes in the ecology of estuarine nekton. Oceanography and Marine Biology Annual Review 35, 163–220. Kneib RT (1997b). Early life stages of resident nekton in intertidal marshes. Estuaries 20, 214–230. Kneib RT (2003). Bioenergetic and landscape considerations for scaling expectations of nekton production from intertidal marshes. Marine Ecology Progress Series 264, 279–296. Kneib RT and Wagner SL (1994). Nekton use of vegetated marsh habitats at different stages of tidal inundation. Marine Ecology Progress Series 106, 227–238. Laffaille P, Lefeuvre JC, Schricke MT and Feunteun E (2001). Feeding ecology of 0-group sea bass, Dicentrarchus labrax, in salt marshes of Mont Saint Michel Bay (France). Estuaries 24, 116–1125. Laffaille P, Feunteun E, Lefebvre C, Radureau A, Sagan G and Lefeuvre J (2002). Can thinlipped mullet directly exploit the primary and detritic production of European macrotidal salt marshes? Estuarine Coastal and Shelf Science 54, 729–736. Le Quesne W (2000). Nekton utilisation of intertidal estuarine marshes in the Knysna Estuary. Transactions of the Royal Society of South Africa 55, 205–214. Lee SY (2004). Relationship between mangrove abundance and tropical prawn production: a re-evaluation. Marine Biology 145, 943–949. Lefeuvre JC, Laffaille P and Feunteun E (1999). Do fish communities function as biotic vectors of organic matter between salt marshes and marine coastal waters? Aquatic Ecology 33, 293–299. Loneragan NR, Adnan NA, Connolly RM and Manson FJ (2005). Prawn landings and their relationship with the extent of mangroves and shallow waters in western peninsular Malaysia. Estuarine Coastal and Shelf Science 63, 187–200. Mazumder D, Saintilan N and Williams RJ (2005a). Temporal variations in fish catch using pop nets in mangrove and saltmarsh flats at Towra Point NSW, Australia. Wetlands Ecology and Management 13, 456–467. Mazumder D, Saintilan N and Williams RJ (2005b). Comparisons of fish catches using fyke nets and buoyant pop nets in a vegetated shallow water saltmarsh flat at Towra Point, NSW. Wetlands (Australia) 23, 37–46. Mazumder D, Saintilan N and Williams RJ (2006a). Trophic relationships between itinerant fish and crab larvae in a temperate Australian saltmarsh. Marine and Freshwater Research 57, 193–199. Mazumder D, Saintilan N and Williams RJ (2006b). Fish assemblages in three tidal saltmarsh and mangrove flats in temperate NSW, Australia: a comparison based on species diversity and abundance. Wetlands Ecology and Management 14, 201–209. Melville AJ and Connolly RM (2005). Food webs supporting fish over subtropical mudflats are based on transported organic matter not in situ microalgae. Marine Biology 148, 363–371. Minello TJ, Able KW, Weinstein MP and Hays CG (2003). Salt marshes as nurseries for nekton: testing hypotheses on density, growth and survival through meta-analysis. Marine Ecology Progress Series 246, 39–59.
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Morton RM, Pollock BR and Beumer JP (1987). The occurrence and diet of fishes in a tidal inlet to a saltmarsh in southern Moreton Bay, Queensland. Australian Journal of Ecology 12, 217–237. Morton RM, Beumer JP and Pollock BR (1988). Fishes of a subtropical Australian saltmarsh and their predation upon mosquitoes. Environmental Biology of Fishes 21, 185–194. Moussalli A and Connolly R (1998). Fish use of the inundated waters of a subtropical saltmarsh – mangrove complex in south-east Queensland. In Moreton Bay and Catchment. (Eds IR Tibbetts, NJ Hall and WC Dennison) pp. 471–472. School of Marine Science, University of Queensland: Brisbane. Nagelkerken I and van der Velde G (2004). Relative importance of interlinked mangroves and seagrass beds as feeding habitats for juvenile reef fish on a Caribbean island. Marine Ecology Progress Series 274, 153–159. Nemerson DM and Able KW (2004). Spatial patterns in diet and distribution of juveniles of four fish species in Delaware Bay marsh creeks: factors influencing fish abundance. Marine Ecology Progress Series 276, 249–262. Odum EP (1968). Evaluating the productivity of coastal and estuarine water. In Proceedings of the Second Sea Grant Conference. pp. 63–64. University of Rhode Island. Odum WE, Odum EP and Odum HT (1995). Nature’s pulsing paradigm. Estuaries 18, 547–555. Peterson GW and Turner RE (1994). The value of salt marsh edge vs interior as a habitat for fish and decapod crustaceans in a Louisiana tidal marsh. Estuaries 17, 235–262. Rountree RA and Able KW (1992). Foraging habits, growth, and temporal pattern of saltmarsh creek habitat use by young-of-year summer flounder in New Jersey. Transactions of the American Fisheries Society 121, 765–776. Rozas LP (1992). Bottomless lift net for quantitatively sampling nekton on intertidal marshes. Marine Ecology Progress Series 89, 287–292. Rozas LP (1995). Hydroperiod and its influence on nekton use of the salt marsh: a pulsing ecosystem. Estuaries 18, 579–590. Rozas LP and LaSalle MW (1990). A comparison of the diets of Gulf killifish, Fundulus grandis Baird and Girard, entering and leaving a Mississippi (USA) brackish marsh. Estuaries 13, 332–336. Rozas LP and Minello TJ (1997). Estimating densities of small fishes and decapod crustaceans in shallow estuarine habitats: a review of sampling design with focus on gear selection. Estuaries 20, 199–213. Rozas LP and Zimmerman RJ (2000). Small-scale patterns of nekton use among marsh and adjacent shallow nonvegetated areas of the Galveston Bay Estuary, Texas (USA). Marine Ecology Progress Series 193, 217–239. Saintilan N, Hossain K and Mazumder D (2007). Linkages between seagrass, mangrove and saltmarsh as fish habitat in the Botany Bay estuary, New South Wales. Wetlands Ecology and Management 15, 277–286. Sheaves M (2005). Nature and consequences of biological connectivity in mangrove systems. Marine Ecology Progress Series 302, 293–305. Sheaves M, Baker R and Johnston R (2006). Marine nurseries and effective juvenile habitats: an alternative view. Marine Ecology Progress Series 318, 303–306. Szedlmayer ST and Able KW (1993). Ultrasonic telemetry of age-0 summer flounder, Paralichthys dentatus, movements in a southern New Jersey estuary. Copeia 3, 728–736. Teal JM (1962). Energy flow in the salt marsh ecosystem of Georgia. Ecology 43, 614–624.
Fish on Australian saltmarshes
Thomas BE and Connolly RM (2001). Fish use of subtropical saltmarshes in Queensland, Australia: relationships with vegetation, water depth and distance onto the marsh. Marine Ecology Progress Series 209, 275–288. Zimmerman RJ, Minello TJ and Zamora G (1984). Selection of vegetated habitat by brown shrimp, Penaeus aztecus, in a Galveston Bay salt marsh. Fishery Bulletin 82, 325–336.
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CHAPTER 7
Saltmarsh as habitat for birds and other vertebrates Jennifer Spencer, Vaughan Monamy and Mark Breitfuss
Introduction Saltmarshes are highly productive systems (Adam 1990), yet in comparison with other temperate ecosystems they support relatively few species of terrestrial vertebrates (Greenberg et al. 2006). Although the importance of saltmarsh habitat has been documented for faunal species in the northern hemisphere (Sherwood et al. 2000; Greenberg et al. 2006), and for bird species in particular (Goss-Custard and Yates 1992; Ganter et al. 1997; Norris et al. 1998; Norris 2000; Hughes 2004), few studies have investigated the importance of saltmarsh habitat for vertebrate species in Australia. This is beginning to change as coastal saltmarsh in Australia gains recognition as important habitat for bird and mammal species (Adam 1990; Morrisey 2000; Laegdsgaard 2006). In 2004, for example, coastal saltmarsh was listed as an endangered ecological community in three bioregions in New South Wales (NSW Threatened Species Conservation (TSC) Act 1995) which recognised its importance as feeding and roosting habitat for shorebirds and foraging habitat for insectivorous bats.
Birds Bird diversity in saltmarsh Saltmarsh is of direct importance to many avian species by providing habitat in which individuals can breed, feed and roost. In Australia, common colonial waterbird species, such as the Australian White Ibis (Threskiornis molucca), Straw-necked Ibis (Threskiornis spinicollis) and Cattle Egret (Ardea ibis), can be found in large flocks in coastal wetlands when wetlands in inland Australia are dry (Kingsford and Norman 2002). Large numbers of Black Swans (Cygnus atratus), Chestnut Teal (Anas castanea) and Australian Shelduck (Tadorna tadornoides) can also congregate in saltmarshes in order to feed and roost. Coastal saltmarsh may also act as drought refuges for Australian breeding shorebird species such as the Black-fronted Dotterel (Elseyornis melanops), Red-kneed Dotterel (Erythrogonys cinctus), Black-winged Stilt (Himantopus himantopus) and Red-necked Avocet (Recurvirostra novaehollandiae) (Lane 1987; Smith 1991), and some shorebirds, such as the Black-winged Stilt, Masked Lapwing (Vanellus miles) and Red-capped Plover (Charadrius ruficapillus), will also breed in saltmarsh (Marchant and Higgins 1993). 149
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Many migratory shorebird species will roost and feed in saltmarsh. These include the Blacktailed Godwit (Limosa limosa), Common Greenshank (Tringa nebularia), Curlew Sandpiper (Calidris ferruginea), Eastern Curlew (Numenius madagascariensis), Latham Snipe (Gallinago hardwickii), Marsh Sandpiper (Tringa stagnatilis), Pacific Golden Plover (Pluvialis fulva), Rednecked Stint (Calidris ruficollis) and Sharp-tailed Sandpiper (Calidris acuminata). Saltmarsh also provides feeding habitat for terrestrial birds. Birds of prey, such as the Whistling Kite (Haliastur sphenurus) and Swamp Harrier (Circus approximans) are commonly observed over saltmarsh and will hunt small birds and mammals. Small passerines, Zitting Cisticola (Normanton) (Cisticola juncidis normani), Golden-headed Cisticola (Cisticola exilis), Little Grassbird (Megalurus gramineus), Richard’s Pipit (Anthus novaeseelandiae) and Whitefronted Chat (Epthianura albifrons), also frequent saltmarshes. The White-fronted Chat is endemic to Australia and will breed and forage in saltmarsh habitat (Major 1991). Saltmarsh in the wetlands of Sydney Olympic Park, NSW, sustains one of the two remaining populations of White-fronted Chat in Sydney, where their habitat has been lost largely to housing estates and industrial areas (Higgins et al. 2001). In south-eastern Australia, this species roosts communally in saltmarsh and nests in samphire shrubs and grass tussocks (Major 1989; Higgins et al. 2001). Australian saltmarshes support many threatened bird species (see Table 7.1). For example, saltmarshes in Victoria, Tasmania and South Australia sustain the critically endangered Orange-bellied Parrot (Neophema chrysogaster) (Environmental Protection and Biodiversity Conservation (EPBC) Act 1999). During its non-breeding season, this parrot feeds on seeds from saltmarsh species, including seeds of Frankenia, Sarcocornia, Tectocornia and Suaeda (Loyn et al. 1986; Orange-bellied Parrot Recovery Team 1998; Garnett and Crowley 2000; Morrisey 2000). The Capricorn subspecies of Yellow Chat (Epthianura crocea macgregori), which is listed as critically endangered in Australia (EPBC Act 1999), nests and forages in saltwater couch grassland and samphire shrubland in central Queensland. This species is only known from Curtis Island, the Torilla Plain and Fitzroy River Delta in central Queensland, but is seasonally mobile and may occur in other locations (Garnett and Crowley 2000). The Slender-billed Thornbill (St Vincent’s Gulf) (Acanthiza iredalei rosinae) is endemic to samphire shrublands on narrow coastal saline mudflats on northern shores of Gulf of St Vincent and the Spencer Gulf, in South Australia (Garnett and Crowley 2000). Although the Bush Stone Curlew (Burhinus grallarius) is usually found in open woodland, in its coastal range it has also been observed foraging in saltmarsh (Department of Environment and Conservation NSW 2006). This species has suffered dramatic declines in abundance across southern and eastern Australia and is listed as near threatened under risk criteria developed by the International Union for the Conservation of Nature (IUCN) and as an endangered species in NSW (TSC Act 1995) (see Table 7.1). Many threatened waterbird species have been observed feeding in Australian saltmarsh including; Black-necked Stork (Ephippiorhynchus asiaticus), Black-tailed Godwit, Lewin’s Rail (Rallus pectoralis), Painted Snipe (Rostratula australis) and Radjah Shelduck (Tadorna radjah) (see Table 7.1) (Marchant and Higgins 1990, 1993; Higgins and Davies 1996; Garnett and Crowley 2000). Declines in the number and range of many of these threatened species have been associated with the drainage of coastal wetlands (Garnett and Crowley 2000; Olsen and Weston 2004). Saltmarsh as habitat for migratory shorebirds Migratory shorebirds, or waders, depend on coastal and inland wetlands and can occur in large numbers (Lane 1987; van de Kam et al. 2004). During their migration, shorebird species use many wetland sites, spread across several countries, to sustain their energy supplies before
Saltmarsh as habitat for birds and other vertebrates
Figure 7.1 The East Asian-Australasian shorebird flyway stretches from non-breeding sites in Australia and New Zealand to breeding sites in Siberia and Alaska. Source: Australasian Wader Study Group.
they reach their destination on the breeding or non-breeding grounds. Most migratory shorebird species found in Australia breed in Alaska, Siberia, Mongolia, northern China and Japan during June and July of each year, and spend their non-breeding seasons in Australia from September to April (Lane 1987). This migratory route is known as the East Asian-Australasian flyway (see Figure 7.1). In Australia, little is known of shorebird use of saltmarsh habitats, however, saltmarsh has been documented as important habitat for several shorebird species in South Africa (Puttick 1979; Kalejta 1992; Velasquez and Hockey 1992), North America (Bildstein et al. 1982; Erwin et al. 1994) and Europe (Goss-Custard and Yates 1992; Norris et al. 1998; Rosa et al. 2003). Most shorebirds feed on invertebrates in intertidal mudflat and are forced to rest during high tide at roost sites, when their low tide feeding habitat is inundated (Lane 1987). Some species differ from this pattern, by feeding almost continuously in saltmarsh throughout the tidal cycle (Puttick 1979) or switching between habitats (Long and Ralph 2001), for example, from mudflats to saltmarshes during high tides (Yasue et al. 2003). Species that commonly feed in saltmarsh in Australia, throughout the tidal cycle include: the Curlew Sandpiper; Marsh Sandpiper; Red-necked Stint and Sharp-tailed Sandpiper. Smaller flocks of Eastern Curlew and Pacific Golden Plover are often recorded roosting in saltmarsh during daytime high tides (Geering 1995; Loyn et al. 2001). These roosts can be
151
Scientific name
Ephippiorhynchus asiaticus
Limosa limosa
Burhinus grallarius
Rallus pectoralis pectoralis
Neophema chrysogaster
Rostratula benghalensis australis
Tadorna radjah
Black-necked Stork
Black-tailed Godwit
Bush Stone Curlew
Lewin’s Rail (eastern)
Orange-bellied Parrot
Painted Snipe (Australian)
Radjah Shelduck
East Kimberley, Northern Territory to northern Queensland, Cape York Peninsula.
Scarce in south-western Australia, range stable in eastern and northern Australia.
Breeding: south-western Tasmania. Non-breeding: King Island, South Australia and Victoria.
Mostly coastal, through south eastern Australia; Townsville, through Victoria and Kangaroo Island in South Australia.
Scarce in southern Australia, remains common in northern Australia.
Migrant, spends non-breeding season in Australia.
Northern Australia from Pilbara, Western Australia to eastern Queensland to mid New South Wales.
Distribution
Feeds on small invertebrates and seeds from shallow wetland edges.
Inhabits shallow, vegetated, temporary wetlands, recorded occasionally in saltmarsh. Eats invertebrates and seeds.
Non-breeding birds disperse to saltmarsh. Feed on seeds of saltmarsh species.
Permanent to emphemeral fresh to saline wetlands.
Usually found on open woodland and grassland but also recorded foraging in saltmarsh. Feeds on invertebrates, reptiles, vegetation and seeds.
Feeds on intertidal mudflats, saltmarshes and brackish wetlands.
Feed in shallow water, on fish, reptiles and frogs.
Ecology
~
~
CR
~
NT
NT
NT
IUCN
r (QLD) sp (WA)
V
CE
v (VIC)
e (NSW) e (VIC) v (SA)
v (NSW)
e (NSW) r (QLD)
Status
Few threats
Wetland drainage Wetland vegetation clearance
Degradation of grazing habitat
Wetland drainage
Predation by foxes Clearance of vegetation
Loss of wetland habitat Disturbance
Loss of wetland habitat Disturbance
Main threats
Distribution, ecology, status and threats to endangered, threatened and vulnerable bird species reported from Australian saltmarshes.
Common name
Table 1
152 Australian Saltmarsh Ecology
Calamanthus campestris hartogi
Acanthiza iredalei rosinae
Epthianura crocea macgregori
Cisticola juncidis normani
Slender-billed Thornbill (St Vincent’s Gulf)
Yellow Chat (Dawson)
Zitting Cisticola (Normanton)
Distribution
South west Cape York Peninsula, Queensland.
Curtis Island, Torilla Plain and Fitzroy River Delta, central Queensland.
Gulf of St Vincent, Spencer Gulf, South Australia.
Dirk Hartog Island, Western Australia.
Ecology
Breeds in saline coastal grasslands.
Found in freshwater and saline wetlands. Breeds in saltwater couch grassland and samphire shrubland.
Samphire shrublands on narrow coastal saline mudflats. Highly selective of samphire species.
Low sparse heath, saltmarsh or samphire, feeds on insects, spiders and seeds.
~
~
~
~
IUCN
Status
r (QLD)
CE
v (SA)
r (WA)
Main threats
Few threats
Wetland drainage Feral pigs Cattle grazing Industrial development
Residential, saltworks and marina development
Feral animals Fire
IUCN status: CR = Critically Endangered ; EN = Endangered; NT = Near Threatened. Australian Status: National (EPBC Act 1999); CE = Critically Endangered; V = Vulnerable State listings: e = endangered; v = vulnerable; r = rare; sp = specially protected NSW = New South Wales Threatened Species Conservation Act 1995; VIC = Victorian Flora and Fauna Guarantee Act 1988; QLD = Queensland Nature Conservation Act 1992; SA = South Australian National Parks and Wildlife Act 1972; WA = Wildlife Conservation Act 1950; NT = Northern Territory Parks and Wildlife Conservation Act 2000 Sources: Marchant and Higgins 1990, 1993; Higgins and Davies 1996; Garnett and Crowley 2000.
Scientific name
Common name
Rufous Fieldwren (Dirk Hartog Island)
Saltmarsh as habitat for birds and other vertebrates 153
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(b)
(a)
Figure 7.2 Flooded saltmarsh on Kooragang Island, the Hunter estuary, New South Wales. Saltmarsh provides (a) important feeding habitat for migratory shorebirds, such as the Sharp-tailed Sandpiper Calidris acuminata and (b) major roosting habitat for many shorebird species during night-time high tides. Photos: J. Spencer.
important during spring high tides, adverse weather conditions or as a result of high disturbance at their main roost sites (Geering 1995; Lawler 1996). At low tides, Eastern Curlew and Pacific Golden Plover usually feed in intertidal mudflat rather than remain in saltmarsh continuously. Shorebird distributions and densities usually match the distribution of their preferred prey species (Goss-Custard 1970, 1977; Zharikov and Skilleter 2004). Most species segregate themselves in intertidal habitat according to preferences for sediment penetrability and water depth, as shorebirds prefer to feed in shallow water or wet substrates (Dann 1987). Prey accessibility is often determined by the maximum depth at which a shorebird can insert its bill into the substrate and maximum leg length (Dann 1987). This allows a suite of species to co-exist in the same feeding habitat (Dann 1987, 1999). Water depth in saltmarsh habitats is most critical in determining prey availability to shorebird species. The amount of bare substrate or shallow water available is a function of factors such as water level, topography, local rainfall, soil type, wind action and tidal connectivity (Skagen and Knopf 1994). Migratory shorebird species, such as the Sharp-tailed Sandpiper (Figure 7.2a), feed on invertebrates in the bare substrate fringing low-level saltmarsh vegetation. The Sharp-tailed Sandpiper is the most common migratory shorebird species in NSW (Smith 1991) and has been observed in large flocks of >1000 birds in areas of saltmarsh, for example, on Kooragang Island, in the Hunter estuary, NSW. In a study of Sharp-tailed Sandpiper foraging behaviour at this location, 70% of birds were observed foraging during flock scans, while a smaller proportion of birds were observed resting and preening (J. Spencer unpub. data). The Sharp-tailed Sandpiper is an opportunistic feeder (Higgins and Davies 1996) and during this study was observed feeding extensively on adult chironomids in saltmarsh and high numbers of insect parts were also found in faecal samples examined from this species at this site. Many other shorebird species feed extensively on chironomids and this insect biomass has been manipulated in some wetlands to attract shorebirds (Rehfisch 1994).
Saltmarsh as habitat for birds and other vertebrates
Table 7.2 High tide roost characteristics and their importance to shorebirds (after Lawler 1996; Luis et al. 2001; Rogers 2003). Roost characterisitic
Component type
Topography/ elevation
Energetics
Importance to shorebirds Protection from adverse weather
Substrate texture and hardness
Energetics
For cooling/ availability of supplementary foraging habitat
Availability of shallow water
Energetics
For cooling/ preening/ availability of supplementary foraging habitat
Proximity to feeding areas
Energetics
Travelling time to and from feeding areas
Vegetation cover/ type
Energetics/ Predation risk
Windbreak/ camouflage from predators
Distance to tall vegetation (visibility)
Predation risk
Tall vegetation provides cover for predators
Proximity to foreshore
Predation risk
Escape distance from predators
Roost background colour
Predation risk
Conspicuousness to predators
Remoteness
Disturbance
Background noise from machinery/ vehicles
Size of roost
Disturbance/ Energetics
Levels of inter/ intra-specific aggression between birds
Distance to alternative roosts
Disturbance/ Energetics
Time spent in flight
Presence of people/ predators
Disturbance/ Predation risk/ Energetics
Time spent in flight
An additional benefit that may arise for shorebirds feeding in saltmarsh throughout the day is that they can use saltmarsh as night roosting habitat. Small shorebird species need to feed almost continuously throughout the tidal cycle, and often make extensive use of higher flats (Goss-Custard and Moser 1988) or supratidal artificial wetlands (Masero et al. 2000; Masero and Perez-Hurtado 2001; Masero 2003) during high tides. By feeding in saltmarsh throughout the day, Sharp-tailed Sandpipers in this study may limit any extra energy that would be expended travelling to and from a separate roost site. Perceived predation risk is thought to underpin the selection of both feeding and roosting sites by shorebirds (Lawler 1996; Luis et al. 2001; Rogers 2003) (see Table 7.2). Shorebirds usually select different sites to roost in at night than during the day (Lawler 1996; Sitters et al. 2001; Rohweder 2001; Rogers 2003) and these can include areas of flooded saltmarsh. Shorebirds are more vulnerable to ground predators, such as foxes, cats and dogs at night (Rogers 2003) but in daytime high tides will select roosts closest to their low tide intertidal feeding habitat and sites which usually have an open aspect allowing easy detection of birds of prey (Lawler 1996; Luis et al. 2001; Rogers 2003). Avian predators are known to be an important cause of mortality in small shorebird species in the northern hemisphere (Page and Whitacre 1975; Creswell 1996; Dekker 1998; Hotker 2000; Dekker and Ydenberg 2004). Saltmarsh habitat may provide additional benefits to migratory shorebirds, such as the Sharp-tailed Sandpipers, which have non-breeding plumage that is well camouflaged against saltmarsh vegetation. On Kooragang Island, Sharptailed Sandpipers were observed flattening themselves against Sacrocornia quinqueflora and Sporobolus virginicus and moving into this vegetation when birds of prey were overhead or if alarm calls were given.
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In the Hunter estuary, NSW, the main night-time roost used by shorebirds is located in flooded saltmarsh on the north-western portion of Kooragang Island. This night roost is characterised by large pools of open water and low-level sparse vegetation dominated by Sacrocornia quinqueflora (Figure 7.2b). During night-time high tides, this roost can support large flocks of Eastern Curlews (> 100 birds) and Bar-tailed Godwits Limosa lapponica (400–600 birds) (J. Spencer unpub. data). Both species roost on a large sandspit and rock training wall in the lower Hunter River during daytime high tides but these day roosts are deserted at night. Shallow pools of water at the night roost may be attractive to shorebirds as they provide protection from ground predators. Many shorebird species have poor night vision (Rojas et al. 1999), therefore they may well rely on the noise created by a ground predator moving through open water as a form of predator detection during night-time high tides. This night roost also provides supplementary feeding habitat for shorebirds during nighttime high tides, including small numbers of Bar-tailed Godwits, Black-tailed Godwits, Common Greenshanks, Eastern Curlews and Sharp-tailed Sandpipers. The provision of supplementary feeding habitat is important to migratory shorebirds in their pre-migratory period when they need to gain weight rapidly (Kersten and Piersma 1987; Piersma et al. 1999) in order to make a successful migration to their breeding grounds in the northern hemisphere. Some shorebirds may also rely on supplementary feeding at high tide roosts when they have not been successful in meeting their daily energy requirements in the previous low tide period (Caldow et al. 1999; Smart and Gill 2003). Threats to shorebird populations Shorebirds face a number of threats to their populations and habitats in the East AsianAustralian flyway. A total of 20% of shorebird species that use this flyway are listed as critically endangered or near threatened under IUCN risk criteria (Barter 2002). In Australia, migratory shorebirds are listed under international migratory bird agreements that Australia has with Japan (JAMBA), China (CAMBA) and the Republic of Korea (ROKAMBA). Under these agreements these signatory countries have made obligations to protect migratory shorebirds and their feeding and roosting habitat. At a national level, Australia addresses its obligations through the EPBC Act (Commonwealth of Australia 1999), which contains important provisions for the protection and conservation of shorebirds. Any action that will have, or is likely to have, a significant impact on shorebirds and their habitat requires approval under the Act. In their non-breeding range, degradation of habitat and excessive disturbance at roosting and feeding habitats are thought to be the main threats to migratory shorebird populations (Smith 1991; Watkins 1993; Department of Environment and Heritage 2005). Shorebirds can suffer high disturbance rates at their roosting and feeding sites by recreational users, such as fishers, watercraft, walkers and dogs (Burger and Gochfield 1991; Fitzpatrick and Bouchez 1998; Paton et al. 2000; Blumstein et al. 2003) and by machinery and vehicles used at construction sites (Burton et al. 2002; Durell et al. 2005). The effect of cumulative disturbance events, from both avian predators and human-induced sources, can result in decreases in energy stores that are needed for moult and migratory fuelling (Burger and Gochfeld 1991). This has implications for energy conservation as any extra time spent in flight can have significant effects on shorebird body condition and mortality (Durell et al. 2005). In south-eastern Australia, large areas of coastal saltmarsh have historically been drained for agricultural and urban development (Adam 1981; Bucher and Saenger 1991; Zann 1995) and like many habitats, saltmarsh is threatened globally by ongoing development pressures and insensitive use (Adam 2002). In some cases, this has directly impacted the availability of night roosting habitat for shorebirds (Clarke and van Gessel 1983).
Saltmarsh as habitat for birds and other vertebrates
The encroachment of mangrove into saltmarsh habitats (Saintilan and Williams 1999) is a major threat to remaining shorebird habitats in south-eastern Australia (Saintilan 2003; Straw and Saintilan 2005). Most shorebird species prefer open roost sites that allow the detection of potential predators (Lawler 1996; Luis et al. 2001; Rogers et al. 2006). Therefore, most shorebirds will avoid areas with tall vegetation (Lawler 1996; Rogers et al. 2006) such as mature mangroves, as tall vegetation can provide cover for ambushing birds of prey (Dekker and Ydenberg 2004). For example, in North America, capture rates for Peregrine Falcons (Falco peregrinus) were highest near the shore zone where falcons used vegetation as a screen before ambushing shorebirds (Dekker 1998; Creswell 1996). The removal of mangrove in shorebird habitat in the Hunter estuary, NSW, has been successful in restoring habitat for shorebirds but monitoring is required to determine the effectiveness of this technique in the long term. In coastlines of Asia, Europe, New Zealand and North America saltmarsh and unvegetated mudflats have been invaded by the cordgrass Spartina spp. (Callaway and Josselyn 1992; Ruiz et al. 1997; Neira et al. 2006) which often hybridises with local native species (Greenberg et al. 2006). In British estuaries, one species of migratory shorebird, the Dunlin Calidris alpina, has declined by ~50% in response to the spread of cordgrass Spartina anglica, which has reduced the availability of intertidal mudflat (Goss-Custard and Moser 1988). Cordgrass has also invaded sites in the Tamar estuary, in Tasmania (Adam 1981) and Western Port Bay, Victoria (Western Port Ramsar Information Sheet 1999). Although it was originally introduced for reclamation of land and stabilisation of mudflats (Laegdsgaard 2006), invertebrate communities often change following colonisation by hybrid Spartina (Hedge and Kriwoken 2000; Neira et al. 2006) and consequently most shorebirds avoid these areas (Goss-Custard and Moser 1988; Callaway and Josselyn 1992).
Insectivorous bats Many species of insectivorous bats have been recorded feeding in mangrove forests (Hoye 2002) and it is thought that these species also use adjacent habitats, such as saltmarsh, as secondary habitat (Laegdsgaard et al. 2004; Belbasé 2005). Bat species tend to feed in open areas of vegetation and can select riparian zones and forest tracks in dense forests (Law and Chidel 2002; Patriquin and Barclay 2003). They can also range over large distances (Churchill 1998). The range of the Little North-western Bat (Mormopterus loriae coburgiana) is restricted to mangrove forests and adjacent areas and this species can be found roosting in the upper branches of the mangrove Avicennia marina (Churchill 1998; Duncan et al. 1999). Bat species which have been recorded in mangrove habitats and are listed as vulnerable in NSW (TSC Act 1995) include the Yellow-bellied Sheathtail Bat (Saccolaimus flaviventris), Eastern Freetail Bat (Mormopterus norfolkensis), Little Bent-wing Bat (Miniopterus australis), Large Bent-wing Bat (Mi. schreibersii), Fishing Bat (Myotis adversus) and Greater Broad-nosed Bat (Scoteanax rueppellii) (Hoye 2002). Few studies document the direct importance of saltmarsh to insectivorous bats (Mills et al. 1994; Hoye 2002; Laegdsgaard et al. 2004; Belbasé 2005). In a study of bats on Kooragang Island, NSW, calls from 10 species of insectivorous bats were recorded overflying and/or feeding over saltmarsh (Belbasé 2005). These included seven Vespertilionidae species i.e. Gould’s Wattled Bat (Chalinolobus gouldii), Chocolate Wattled Bat (C. morio), Mi. australis, Mi. schreibersii, Little Forest Bat (Vespadelus vulturnus), Lesser Long-eared Bat/Gould’s Longeared Bat (Nyctophilus geoffroyi/N. gouldii) and Greater Broad-nosed Bat (Scoteanax rueppellii), and three species from the Molossidae family; Mo. norfolkensis, Little Freetail Bat (Mormopterus sp.) and White-striped Mastiff Bat (Tadarida australis) (Belbasé 2005).
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In all, 75% of calls recorded over saltmarsh were from V. vulturnus and Mo. norfolkensis and high levels of feeding activity were recorded for both species (Belbasé 2005). Most insectivorous bats are opportunistic feeders (Churchill 1998), therefore these species could be feeding on flying insects from the Orders Homoptera, Hemiptera, Diptera, Coleoptera, Orthoptera, Hymenoptera and Lepidoptera, which are commonly found in or over saltmarsh vegetation (Belbasé 2005; Laegdsgaard 2006). Belbasé (2005) reported high numbers of feeding buzzes from insectivorous bats which coincided with the emergence of mosquitoes following spring high tides. Coastal saltmarshes provide habitat for a number of mosquito species, the most abundant being Aedes vigilax (Diptera: Culicidae). This mosquito is the most important nuisance-biting pest and vector of arboviruses (e.g. Ross River virus and Barmah Forest virus) in coastal areas of NSW (Russell and Dwyer 2000). The larvae of this mosquito are found in temporary pools in saltmarsh areas following inundation by high tides or rainfall. Numbers of saltmarsh mosquitoes can be highly variable, with adults generally living less than three weeks and also dispersing widely (up to 20 km) from larval habitats. Consequently, the abundance of adult Ae. vigilax in estuarine wetlands generally varies in four-week cycles where the greatest abundance of adults occurs approximately 10–14 days following initial inundation of the saltmarsh. The magnitude of population change is dependent on a number of environmental and climatic conditions including tide height, rainfall, rain days, temperature, humidity and predator populations (see Chapter 8). While mosquitoes may be consumed by insectivorous bats, little is known of the contribution mosquitoes make to the diet of bats that forage over coastal saltmarsh. While it is unlikely that bats represent an effective mosquito management tool, they may play a role in integrated pest management strategies. It is also important to identify the significance of mosquitoes to the diet of insectivorous bats and, consequently, the impact broad-scale mosquito control (and subsequent reduction of adult mosquito abundance) may have on coastal bat populations.
Water Mouse The Water Mouse or False Water Rat (Xeromys myoides) is a small native rodent recorded from coastal saltmarsh, mangrove and coastal freshwater wetlands (see Figure 7.3). It is listed as vulnerable under international, Australian and Queensland state legislation (IUCN; Commonwealth of Australia EPBC Act 1999; Queensland Nature Conservation (Wildlife) Regulation 1994 of the Nature Conservation Act 1992). The Water Mouse is distributed in coastal areas of central and south-east Queensland to the Queensland/ New South Wales border (Van Dyck and Gynther 2003, Ball 2004) and the mainland and near-shore islands of the Northern Territory (McDougall 1944, Redhead and McKean 1975, Magnusson et al. 1976, Van Dyck 1997, Woinarski et al. 2000). This rodent is probably entirely nocturnal, sheltering during the day and between tidal cycles in constructed nesting mounds and natural or artificial hollows. The species will consume grapsid and grapsoid crabs, intertidal crustaceans, pulmonate snails and marine gastropods (Van Dyck 1997) which are commonly found in intertidal saltmarsh (Breitfuss et al. 2004). The most important threats to the Water Mouse are the loss, degradation and fragmentation of freshwater and intertidal wetland communities.
Other vertebrates Large animals are often only recorded incidentally in saltmarsh. In southern Australia, kangaroos and wallabies are occasional visitors to upper saltmarsh (Adam 1990). The absence of trees and presence of grasses makes saltmarsh habitats attractive to two species of macropods in particular: the Swamp Wallaby (Wallabia bicolor), and Eastern Grey Kangaroo (Macropus
Saltmarsh as habitat for birds and other vertebrates
Figure 7.3 The rare Water Mouse or False Water Rat Xeromys myoides is a small native rodent found in coastal saltmarsh in southern and central Queensland, and the Northern Territory. Photo: M. Breitfuss.
giganteus). The Eastern Grey Kangaroo has been recorded in NSW saltmarsh on Kooragang Island (P. Svoboda pers. obs) and Cararma and Currambene Creeks in Jervis Bay (N. Saintilan and K. Rogers pers obs.). Rabbits, hares, foxes and rats introduced to Australia are occasional visitors to saltmarsh. The effects of grazing by macropods and rabbits in saltmarsh communities, however, are yet to be investigated (Adam 1990). Reptiles and amphibians are not a normal feature of saltmarsh habitats but they may be more numerous in some brackish situations (Adam 1990). For instance, the vulnerable Green and Golden Bell Frog (Litoria aurea) (EPBC Act 1999) has been recorded in saltmarsh and mangrove areas during periods of high rainfall (P. Svoboda; K. Darcovich pers obs). Incidental sightings of reptile species in saltmarsh include Goannas, or Monitor lizards (Varanus spp.); Red-bellied Black Snakes (Pseudechis porphyriacus) in upper marshes dominated by Baumea, Estuarine Crocodiles (Crocodylus porosus) in higher level flats in the Northern Territory (P. Adam pers obs.); and Eastern Long-necked Turtles (Chelodina longicollis) in the wetlands of Sydney Olympic Park, in Sydney (K. Darcovich pers obs). The Cream-striped Shining-skink (Cryptoblepharus virgatus) is also known from saltmarsh samphire shrubland near Townsville, Queensland. These reptiles may feed on amphibians, small mammals and other reptiles in saltmarsh.
Conclusions and implications for management Despite a general deficiency in the number of studies, there is considerable evidence to support the importance of coastal saltmarsh as habitat for bird, bat and mammal species in Australia. Coastal saltmarsh is used as foraging habitat by several nationally threatened species including the Orange-bellied Parrot, Yellow Chat (Capricorn subspecies), Painted Snipe and Water Mouse. Saltmarsh also provides habitat for numerous shorebird species by supporting breeding for several resident shorebird species, feeding and roosting habitat for resident and migratory shorebirds, and major night-time roosting habitat for many shorebird species. It also provides
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secondary feeding habitat for at least 10 species of insectivorous bats. Saltmarsh probably supplies a range of invertebrate prey items that can be utilised by shorebird and insectivorous bat species. Coastal saltmarshes can act as drought refuges for colonial waterbird species and Australian breeding shorebirds that are more commonly found in inland Australia. These saltmarshes may become increasingly important for waterbird species as a result of increased drought severity and periodicity and river regulation in inland Australia (Kingsford 2000; Olsen and Weston 2004; Kingsford and Porter 2006). The challenge is now to manage remaining coastal saltmarshes sensitively so that they continue to support a high diversity of plant and animal species. The main threat to shorebird species and small mammals, such as the Water Mouse is the loss, degradation and fragmentation of habitat, either through sea level rise, drainage for urban and industrial developments or physical changes to saltmarsh that modify tidal amplitude and frequency of inundation. The encroachment of mangrove and cordgrass into saltmarsh habitats also threatens shorebird feeding and roosting habitats in south-eastern Australia. The removal of standing water by runnelling to control mosquito populations may also conflict with the provision of shorebird feeding and roosting habitat. Bats may also be vulnerable to the secondary effects of some insecticides used to control nuisance-biting mosquitoes (Clark 1988). It is also unclear what contribution the saltmarsh mosquito makes to the diets of shorebird and insectivorous bat species. Ponds are often created within tidal marshes to create waterfowl habitat and to assist in the control of mosquitoes (Erwin et al. 1994). Water levels and salinity regimes of these areas need to be managed to maintain a complex mosaic of moist ground and shallow water with sparse vegetation, as these areas support the largest concentrations of shorebird species. By maintaining natural tidal flows and ensuring that wide shallow edges are incorporated into pools and channels in mosquito runnelling (Lawler 1994), a balance between the provision of shorebird habitat and effective mosquito control could be easily achieved. The next step in the conservation and management of Australian saltmarshes is to address the lack of detailed studies investigating terrestrial vertebrate species use of this habitat and the links between different trophic levels.
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Saltmarsh as habitat for birds and other vertebrates
Smith P (1991). ‘The biology and management of waders (Suborder Charadrii) in NSW’. NSW National Parks and Wildlife Service Species Report No. 9, Hurstville. Straw P and Saintilan N (2005). Shorebird habitat management in Australia – the threat of mangroves. In Status and Conservation of Shorebirds in the East Asian-Australasian Flyway; Proceedings of the Australasian Shorebird Conference. 13–15 December 2003, Canberra. pp. 87–91. Australasian Wader Studies Group and Wetlands International – Oceania. van de Kam J, Ens BJ, Piersma T and Zwarts L (2004). Shorebirds: An Illustrated Behavioural Ecology. KNNV Publishers: Utrecht. Van Dyck, S (1997). Xeromys myoides Thomas, 1889 (Rodentia: Muridae) in mangrove communities of North Stradbroke Island, south-east Queensland. Memoirs of the Queensland Museum 42, 337–366. Van Dyck S and Gynther I (2003). Nesting strategies of the Water Mouse Xeromys myoides in south-east Queensland. Memoirs of the Queensland Museum 49, 453–479. Velasquez CR and Hockey PR (1992). The importance of supratidal foraging habitats for waders at a south temperate estuary. Ardea 80, 243–253 Watkins D (1993). ‘A national plan for shorebird conservation in Australia’. Australasian Wader Studies Group, RAOU Report No. 90, Melbourne. Western Port Ramsar Site Information Sheet (1999). ‘Wetlands International Ramsar Information’ Available at http://www.deh.gov.au/water/wetlands/database/index.html. Woinarski J, Brennan K, Dee A, Njudumul J, Guthayguthay P and Horner P (2000). Further records of the false water rat Xeromys myoides from coastal Northern Territory. Australian Mammalogy 21, 221–223. Yasue M, Quinn JL and Cresswell W (2003). Multiple effects of weather on the starvation and predation risk trade-off in choice of feeding location in Redshanks. Functional Ecology 17, 727–736. Zann LP (1995). ‘Our Sea, Our Future. Major findings of the State of the Marine Environment Report for Australia’. Department of Environment, Sport and Territories: Canberra. Zharikov Y and Skilleter GA (2004). A relationship between prey density and territory size in non-breeding Eastern Curlews Numenius madagascariensis. Ibis 146, 518–521.
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CHAPTER 8
Ecology and management of mosquitoes Pat Dale and Mark Breitfuss
Introduction Saltmarshes form at the dynamic interface between land and sea. This interface experiences fluxes in the biological, physical and chemical processes which determine the density and abundance of species. As a result, saltmarshes provide important ecosystem services but also support the key habitat features necessary for breeding of some species of pestiferous and vector mosquitoes. In Australia, the most common species of mosquitoes that breed in saltmarsh habitats are Aedes vigilax (Skuse) in the warmer tropics and subtropics and Aedes camptorhynchus (Thomson) in the cooler areas. Aedes alternans (Westwood) occurs in both regions, but less commonly. All saltmarsh mosquitoes breed in the free water that pools in shallow depressions following tidal influence or as a result of freshwater inputs.
Mosquito breeding strategies The biology of the mosquitoes inhabiting saltmarsh was described by Russell (1993) for southeast Australia and by Liehne (1991) for Western Australia. The adaptive strategies of different species are reflected in their ability to exploit the dynamic nature of saltmarsh. For example, the larvae of some species develop near saltmarshes during normal tidal cycles and then in them following rainfall, when the salinity of ground pools is lower. These brackish species include Culex sitiens (Wiedemann), Verrallina funerea (Theobold), Aedes procax (Skuse), Culex annulirostris (Skuse) and Aedes notoscriptus (Skuse). Many mosquitoes require water for egglaying. For example Culex sitiens and related Culex species lay a raft of eggs on the water. The Anopheline (genus Anopheles) mosquitoes also lay eggs on water, hence water is needed for both the egg and larval stages. In contrast, aedine mosquitoes (those in the genus Aedes) lay eggs on the ground surface. For the saltmarsh mosquito (Aedes vigilax) the lifecycle requires the varying conditions that occur on the saltmarsh. The eggs are laid, singly, on damp ground or plant stems, and are conditioned by drying out. Then flooding occurs and hatching is stimulated by a decline in dissolved oxygen. Following tidal events, hatching of Aedine species can occur almost immediately and simultaneously. This can result in very high densities of larvae developing under ideal conditions.
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Australian Saltmarsh Ecology
Culex egg raft
Aedes eggs laid singly
Eggs
Larva
and
Pupa
Adult
Figure 8.1 Mosquito lifecycle. Lifecycle drawing adapted from Silver City Vector Control Task Force, http://www.townofsilvercity.org/vector/MosquitolifeCycle.html (accessed 20 October 2006).
Larval development and ecology The lifecycle of the saltmarsh mosquito is shown in Figure 8.1. After hatching, and provided water remains for sufficient time, the larvae will develop through four larval stages. A newly hatched first instar Aedes vigilax is less than 0.9 mm long and a fourth instar is around 4–5 mm. During the larval stages a larva breathes air directly via a siphon or by an air hole, connecting through the water surface. This means that larvae can survive in low dissolved oxygen environments, where aquatic predators may be relatively few. The larvae pupate, ceasing feeding and soon emerge as adults. Males often emerge first to await the females. After mating, females seek a blood meal for protein to help develop their eggs. However some species are autogenous, that is, they can lay one batch of fertile eggs without mating. In summer, when water temperature is warm, the larvae can complete development, pupate and emerge as adults in around five days. In winter development may take three weeks or longer. The life span in the field is probably around two weeks though adults can be kept alive under laboratory conditions for several months (Morris et al. 2002). Aedes vigilax can fly considerable distances, up to 50 km, though this may be influenced by favourable wind (Morris et al. 2002). The biology of Aedes vigilax is described in Sinclair (1976). Detailed research based on remote sensing and field study at Coomera Island in south-east Queensland showed an interesting strategy for Aedes vigilax (Skuse) (Dale et al. 1986). In the Coomera research, oviposition sites were intermediate between low and high marsh in areas of mixed Sarcocornia quinqueflora (Bunge ex Ung.-Stern) and Sporobolus virginicus (L.) Kunth vegetation. Larvae were most numerous on the low marsh, with its Sarcocornia vegetation and on mudflats, but were also plentiful on the upper marsh, in amongst the dense Sporobolus. This is one of the reasons that saltmarsh mosquitoes are numerous: pools in the higher marsh that
Ecology and management of mosquitoes
Table 8.1 Number of Ross River cases in Australia by state and territory, 1993–2007. (Source: Communicable Diseases Intelligence, Australia) Year
ACT
NSW
NT
Qld
SA
Tas
Vic
WA
Aust
1993
4
599
264
2252
773
10
1198
153
5253
1994
1
332
312
2998
28
24
58
95
3848
1995
2
235
369
1643
21
28
32
303
2633
1996
1
1032
137
4881
56
76
152
1445
7780
1997
9
1598
218
2363
635
12
1042
717
6594
1998
6
581
127
1946
67
9
136
288
3160
1999
8
953
157
2305
40
67
223
624
4377
2000
16
750
145
1481
416
8
319
1089
4224
2001
10
716
225
1568
141
13
351
202
3226
2002
0
180
63
886
42
117
37
131
1456 3850
2003
1
494
121
2518
33
4
16
663
2004
6
699
233
2006
53
20
91
1102
4210
2005
6
579
209
1181
155
5
98
311
2544
2006
10
1123
279
2616
328
14
212
820
5502
2007
12
828
300
2145
265
6
92
510
4158
flood on high tides may often be isolated for several days. This means that access for predators such as fish is very limited and the habitat is relatively safe for larvae, though drying up is a risk. On the other hand the lower areas retain water but also have increased risk of predation. Disease transmission Saltmarsh mosquitoes transmit a variety of arbovirus diseases. Ross River (RRv) and Barmah Forest viruses are the two dominant mosquito-borne viruses affecting humans in the vicinity of coastal saltmarsh. They are not fatal but are debilitating and are characterised by arthritis, fever, rash and fatigue (MacKenzie et al.1998). The economic impact of RRv is significant with estimates of the cost of healthcare resources and productivity loss to be AU$1000–2500 per person (Boughton 1996; Harley et al. 2001; Mylonas et al. 2002; Ratnayake 2006). Although RRv is present in all states (Miller et al. 2003), Queensland generally has more cases than any other state (Russell 2002). Tables 8.1 and 8.2 show the annual incidence and rate by state/territory. From the data in Table 8.1, Queensland had on average 53% of the cases, ranging from 35% in 2000 to 78% in 1994. Table 8.2 shows that the Northern Territory has the highest rate, though the numbers are less than in Queensland. Only Tasmania and the Australian Capital Territory have relatively few cases and a low infection rate. Ross River virus is transmitted by female mosquitoes when an infected mosquito takes a blood meal (to enable her eggs to develop). The cycle is shown in Figure 8.2. The mosquito is infected by biting an intermediate host such as a kangaroo, bat, wallaby or other warm-blooded animal. The virus then replicates over a period of around a week within the mosquito and so can be transmitted when next she feeds. Except under epidemic conditions it is thought that humans are a deadend host. That is, the mosquitoes are not reinfected when biting a viraemic person. In some cases transmission may also occur ‘vertically’ when an infected female’s eggs hatch infected larvae, resulting in adults ready to infect. Because of the risk to human health, control of mosquitoes close to human settlement is important and is, for example, mandatory in Queensland under the Public Health Act 2005.
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Australian Saltmarsh Ecology
Table 8.2 Rate of Ross River cases/100 000 in Australia, by state and territory 1993–2007. (Source: Communicable Diseases Intelligence, Australia) Year
ACT
NSW
NT
Qld
SA
Tas
Vic
WA
Aust
1993
1.3
9.9
152.3
70.7
52.7
2.1
26.7
9
29.4
1994
0.3
5.4
175.7
91.8
1.9
5.1
1.3
5.5
21.3
1995
0.6
3.8
202.9
49.2
1.4
5.9
0.7
17.2
14.4
1996
0.3
16.5
73.3
143.7
3.8
16.1
3.3
80.4
42
1997
2.9
25.2
114.8
68.4
42.7
2.5
22.4
39.2
35.2
1998
1.9
9.1
65.8
55.4
4.5
1.9
2.9
15.5
16.7
1999
2.5
14.7
80.3
64.6
2.7
14.2
4.7
33.1
22.8
2000
4.9
11.3
72.5
40.7
27.5
1.7
6.6
57.1
21.7
2001
3.1
10.8
113.3
42.3
9.3
2.8
7.2
10.5
16.4
2002
0
2.7
31.8
23.3
2.7
24.5
0.8
6.7
7.3
2003
0.3
7.3
60.5
64.9
2.2
0.8
0.3
33.4
19.1
2004
1.8
10.3
114.9
50.6
3.4
4.1
1.8
54.8
20.7
2005
1.8
8.3
100.7
28.6
9.7
1
1.9
14.9
12.1
2006
2.9
17.8
129.8
62.6
20.7
2.8
4.1
38.9
26.2
2007
3.5
12
139.6
51.3
16.7
1.2
1.8
24.2
19.8
The responsibility for mosquito control involves the local governments in the states and territories and assistance is provided by the state/territory Health departments, by funding as in WA or by other assistance as in Queensland. Control methods In Australia, control or management of saltmarsh mosquitoes is focussed on the larval stages rather than on adult mosquitoes. Larviciding (control of larval mosquitoes) in saltmarsh
Figure 8.2 Ross River virus transmission cycle. Adapted from a figure provided by the Queensland Institute of Medical Research, mosquito photo provided by Prof Brian Kay.
Ecology and management of mosquitoes
requires a detailed knowledge of the biological and ecological features of the target animal to maximise the efficiency of measures. The main advantage of larval control over adulticiding (control of adult mosquitoes) is that measures are implemented when the target organism is still developing. The concentration of immature mosquitoes on saltmarsh is very high compared to the scattered distribution of adult mosquitoes. Generally, saltmarsh mosquito larvae occur in transient pools located high on the marsh. These pools are filled infrequently by tidal or surface waters and control efforts target the larval habitats. Control can be can be effected via several mechanisms, under three broad strategies: ● ● ●
habitat modification chemical treatment the use of biological agents.
Each has its advantages and disadvantages in terms of effectiveness, costs and environmental impacts. An integrated mosquito management program uses a variety of methods so that there is not complete reliance on any one strategy. Saltmarsh management for mosquito control was reviewed in Dale and Hulsman (1990). They identified serious information gaps in the area of impacts of control and some of these have been remedied at least in part since then. In the area of habitat modification they also identified a changing rationale for saltmarsh mosquito management in Australia. This was one that attempted to modify the habitat only so far as it was needed to reduce mosquito larval populations, rather than adopt the approach that destroying the environment would also destroy mosquitoes, as had been implemented elsewhere earlier in the 20th century. The Australian rationale has been embedded in policy as demonstrated in the Australian Mosquito Control Manual (Morris et al. 2002) and in Queensland, the Mosquito Management Code of Practice (Local Government Association of Queensland 2002). Habitat modification Modifying the larval habitat will affect mosquitoes. The advantage is that this is of long-term effect and reduces ongoing control costs, though it may take several years to recoup the costs in terms of savings by using less chemicals (Dale et al. 1989). The disadvantage is that it may incur a relatively large capital cost at the time of construction, as well as incurring delays because of permitting requirements. There is also the potential for environmental damage, especially with respect to modifying the physical, chemical and biological characteristics of saltmarsh. Habitat modification of saltmarshes for mosquito control was developed early in USA and has evolved into several methods that are widely used. These range from ditching the marsh, via Open Marsh Water Management which retains or restores tidal flooding (Meredith et al. 1985; Ferrigno and Jobbins 1986) to impounding, nowadays allowing for tidal circulation during times of low mosquito populations (Carlson and O’Bryan 1988). The concept of minimal disturbance was developed in Australia and reported in Dale and Hulsman (1990). It led to runnelling, which is a minor form of Open Marsh Water Management. Runnelling was first implemented in New South Wales in 1984 and at Coomera Island in Queensland (S27° 51', E153° 33') in November 1985 (Hulsman et al. 1989). The term runnelling was coined to clearly distinguish it from ditching. Having identified water as the critical variable in the mosquito system, runnelling was designed to connect the isolated pools to the tidal source to increase the flushing potential of otherwise isolated pools and to allow greater predator access to the larvae. At Coomera Island the saltmarsh is usually only flooded by tides exceeding 2.45 m AHD (Australian Height Datum) (predicted at the Brisbane Bar). Thus around 7% of tides each year (around 50) flood the marsh. However, if
171
Australian Saltmarsh Ecology
2.8 2.7
Predicted tidal height (m)
2.6 2.5 2.4 2.3 2.2 2.1 2
ec D
N ov
ct O
g
p Se
Au
l Ju
n Ju
M ay
r Ap
M ar
Fe b
n
1.9 Ja
172
Month
Figure 8.3 Effect of runnelling calculated on the frequency of tidal flooding at Coomera Island (from Breitfuss 2003). Tidal cycle at the Brisbane Bar is shown for 2002. Dashed line represents height of tides (2.45 m) that would normally completely flood the marsh. When runnelled, tidal height required for inundation reduces to 2.20 m (solid line) thereby increasing the number of flooding tides (Queensland Department of Transport 2002).
runnels are constructed to a depth of 0.25 m then the tide will extend into the marsh on a 2.20 m predicted tide. This means that around 28% of tides (n = 196) will flood the marsh after runnelling. This is illustrated in Figure 8.3 above. Natural channels were used as models for the structures (runnels) and so these were designed to be shallow, spoon shaped channels, three times as wide as deep and with a maximum depth of 0.30 m. The depth of runnels was determined with reference to a sandy erodible layer at that depth at the Coomera site. In retrospect, this was a wise decision, as potential acid sulfate soils are encountered throughout coastal Australia at elevations below 5 m AHD. Further, careful site selection and design for runnelling should minimise the potential for encountering and exposing these soils (see Alsemgeest et al. 2005). In addition, runnels were located as much as possible in areas of natural water flow, determined in the field and from colour infrared aerial photographs, in which wet depressions and natural channels appear relatively dark. Figure 8.4 shows a view of a runnel and the runnel layout at Coomera Island. That runnelling reduces mosquito populations is generally accepted and the method is used widely in south-east Queensland, New South Wales and Western Australia. In Western Australia, Latchford (1997) found runnelling to have no significant impact on a number of macro-scale features. However, the study was short-term with less than two years post-runnelling assessment. The impacts of runnelling have been assessed at a range of sites but most research has been on the Coomera site. The 20-year monitoring for that site has been reported in Dale (2008). It concluded that mosquito production was reduced and that there were very small magnitude
Ecology and management of mosquitoes
Figure 8.4
Runnel and layout at Coomera Island (1986).
impacts on the environment including slightly increased substrate moisture lower salinity and some vegetation change (mainly smaller and less dense plants). Other research in south-east Queensland has identified localised impacts associated with the presence of runnels on: ● ● ●
●
●
sediment consolidation and soil water features (Breitfuss and Connolly 2004) seasonal reductions in fish abundance adjacent to runnels (Connolly 2005) patterns in grapsid and grapsoid crab species abundance (Breitfuss et al. 2004; Breitfuss et al. 2005) patterns in the density and size structure of surface-feeding pulmonate snails (Breitfuss et al. 2005) transport of mangrove propagules (Breitfuss et al. 2003).
Many of the patterns associated with the presence of runnels can be explained in terms of natural variability in the characteristics of saltmarshes. For example, there were differences between treatment and controls at runnelled sites in south-east Queensland for some crabs in a trapping experiment (Chapman et al. 2004; Breitfuss et al. 2005) and in a study of crab burrow density (Breitfuss 2003, 2005). The results of the crab burrow study indicated localscale differences in the distribution and abundance of species, between sites with and without runnels, based on their preferences for different substrate conditions. Given the high degree of heterogeneity between these conditions on marshes in south-east Queensland (Breitfuss and Connolly 2004), the results tended to highlight these features rather than impacts associated directly with the presence of runnels. Nevertheless the south-east Queensland sites do have relatively few species and the impacts noted may differ for more species-rich environments at higher latitudes. Concerns have been raised over the potential for runnels to facilitate the invasion of mangroves onto saltmarsh. While runnels will transport mangrove propagules onto saltmarsh (Breitfuss et al. 2003), it is unlikely the propagules are deposited to locations suitable for growth and sustained development. Mangrove invasion of the saltmarsh has not occurred at runnelled sites (Jones et al. 2004) although generally, mangrove density on saltmarsh is increasing and often at the expense of saltmarsh (Saintilan and Williams 2000). Research at the Coomera site has shown that as a mosquito control technique, runnelling results in fewer non-target impacts (noted above) than Open Water Marsh Management (Dale
173
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Australian Saltmarsh Ecology
and Knight 2006). Most of the analyses have indicated that runnelled sites do not differ significantly from the control sites (Dale 2008) and especially when considering saltmarsh processes (Dale and Dale 2002; Dale et al. 1993; Dale and Hulsman 1988). Chemical control Adult mosquitoes may be targeted using a range of non-specific chemicals. In Australia, adulticiding is not the preferred method of control and is generally only implemented under emergency situations when adult mosquitoes pose a direct risk of disease transmission to humans. It is usually considered more effective to control the problem of mosquitoes closer to their source; the larval habitat where the developing mosquitoes are restricted to water bodies. If the saltmarsh habitat is not modified by runnelling, Open Marsh Water Management (OMWM) or some other method, then the most effective control method is to treat larval habitats with a larvicide. The advantage is that because larvicides are used in the isolated pools containing larvae, the effective kill rate may well exceed 95%. The main disadvantages in using larvicides is the volume of product required to be distributed in the field and associated application costs. Most larviciding programs are implemented by using aerial and/or ground-based vehicles and personnel. Monitoring is required to identify the presence of larvae on the saltmarsh, as the location of larval habitat pools is not consistent with all patterns of tidal or surface water inundation. Further, monitoring is necessary to establish both whether an application has been successfully applied to a larval habitat and when the application has ceased to be effective. Studies from overseas and in Australia have identified a number of chemicals suitable for use in Australia. Based on their limited range of non-target impacts and ability to be applied using broad hectare methods, the two most commonly used chemicals are Bacillus thuriegiensis var israelensis (B.t.i, a synthetic bacterium) and (s)-methoprene, an insect growth regulator. Both chemicals are produced in a variety of formulations (liquid, granules, briquettes) that offer options for the type of application and length of activity required. The mode of action of the commonly used larvicide chemicals is either through ingestion or external contact. For B.t.i products, the chemical is ingested by larvae and, due to the pH of the larval midgut, toxic proteins are released from the product that cause cell lysis and death. Only actively feeding larvae are affected by the chemical. The chemical is quite target-specific to Diptera, especially mosquitoes. Further, the potential for broad non-target impacts is reduced because the mode of action requires the release of toxic proteins based on pH. Common formulations are liquids or granules and the activity period is relatively short, less than a week in the field. The insect growth regulator (IGR) (s)-methoprene is delivered to larval habitat pools as a contact chemical. The timing of application for (s)-methoprene is critical for effectiveness on early instars. The mode of action of the chemical is to slow development and hinder the final moult from pupa to adult. Monitoring of (s)-methoprene activity on field populations of mosquitoes is necessary for evaluating mortality and percentage reduction in emergence. Both liquid and solid formulations are available. Solid formulations may provide consistent control for up to a few months in some circumstances. A limited number of organo-chlorine based products are used for larviciding in Australia. The managed use of these chemicals can provide effective control of saltmarsh mosquitoes, though a range of significant non-target impacts on invertebrates and vertebrates have been investigated in the field and laboratory. Although effective at killing larvae the treatments do have adverse impacts on non-targets especially in saline wetlands and on crustaceans, either by direct mortality or by leading to increased mortality from predation (Dale and Hulsman 1990). The products are applied to the areas to be treated either by ground-based application or by aerial spraying. For large and not easily accessible areas, aerial application is most effective and efficient as it can be carried out rapidly. This is important as the window of opportunity may
Ecology and management of mosquitoes
be restricted to only a few days following a tidal or flooding event. This window must also consider the field verification of a hatching event and then the flight must be arranged so that the product is applied at a time relevant to the chemical being used and the stage of development of the larvae. In practice, the control agencies program their operations based on tidal predictions, weather conditions, knowledge of field conditions and often have arrangements with the aerial spraying company for rapid responses. The use of Global Positioning Systems to ensure appropriate coverage and calibration of dispensing equipment minimises both monetary and environmental costs. Biological control Research and experimentation into the effectiveness of biological control agents for saltmarsh mosquito control are in their infancy. To-date, studies have focussed on field observation of natural systems, rather than manipulation. There has been no really effective biological control developed for saltmarsh mosquitoes. While various organisms are known to prey on mosquito larvae, such as copepods, insects and fish, their effectiveness for large-scale control programmes is limited. There is still a dearth of information regarding saltmarsh predators whose lifecycle may be synchronised with that of the larvae. To be effective in reducing very high densities of mosquito larvae, small predators need to be able to colonise the marsh as rapidly as the eggs hatch so as to be able to consume the small first instar larvae. Fish have the potential to predate on larvae when they access the marsh with a flooding tide. There have been year-long studies at the Coomera Island site, conducted prior to runnelling. Morton et al. (1987) recorded some adult mosquitoes in the fish diet, but insects appeared to be a very minor item of diet. Morton et al. (1988), focussing on the mosquito issue at Coomera, found that fish did consume larvae, but not in large quantities and concluded that fish were not likely to control the mosquitoes. Shrimps may also have potential to act as a biological control in saltmarshes but more research is needed on this. A study in south-east Queensland has indicated that they feed on larvae (Morris et al. 2002). In addition, some mosquito larvae prey on smaller larval instars. In the saltmarsh Aedes alternans is one such species, but is not found in sufficient numbers to act as an effective control and moreover is another pest mosquito, although not a known disease vector. The dynamic nature of saltmarsh larval habitats is suited to the breeding strategies of mosquitoes, but not for transient aquatic organisms that require more static conditions for survival. Saltmarsh pools are exposed to irregular periods of drought and flooding, dependant on their location from the tidal front, height on shore and proximity to other sources of water. In addition to the intermittent presence of water on saltmarsh, its salinity characteristics vary considerably, ranging from almost fresh/brackish following rainfall to hypersaline a few days after tidal flooding in summer. Salinities in excess of 40 parts per thousand (ppt) are not uncommon in the isolated pools that provide mosquito larval habitat, compared to seawater of 35 ppt). Water table salinities of up to 80 ppt have been recorded at Coomera Island (Dale, unpublished data).
Conclusion Saltmarshes provide important ecosystem services, including habitat for disease vector mosquitoes. This has human health impacts and possibly longer term consequences for the ecosystems, as a result of attempts to manage the issue. Mosquito management can be carried out via a range of methods and the most developed and best resourced programs are integrated ones, relying on a mix of methods as appropriate to the nature of the problem, areas and species. In
175
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general, in Australia, larval control is preferred to adulticiding and that means that efforts to control saltmarsh mosquitoes focus on saltmarsh larval habitats. Not all saltmarshes are suited to habitat modification, but where they are, runnelling provides long-term and cost-effective control with minor non-target impacts. Chemical control is highly effective as a short-term measure, but requires coordination, monitoring and incurs significant costs. Biological control is unlikely, in the near future at any rate, to offer the extent of management necessary for an effective reduction in vector mosquitoes.
References Alsemgeest G, Dale P and Alsemgeest D (2005). Evaluating the risk of potential acid sulfate soils and habitat modification for mosquito control (runnelling): comparing methods and managing the risk. Environmental Management 36, 152–161. Boughton CR (1994). Arboviruses and disease in Australia. Medical Journal of Australia 160, 27–28. Breitfuss MJ, Connolly RM and Dale PER (2005). Habitat modification: is there significant impact on saltmarsh fauna? Arbovirus Research in Australia 9, 58–63. Breitfuss MJ, Connolly RM and Dale PER (2004). Densities and aperture sizes of burrows constructed by Helograpsus haswellianus (Decapoda: Varunidae) in salt-marshes with and without mosquito control runnels. Wetlands 24, 14–22. Breitfuss MJ and Connolly RM (2004). Consolidation and volumetric soil water content of saltmarsh soils following habitat modification for mosquito control. Wetland Ecology and Management 12, 333–342. Breitfuss MJ, Connolly RM and Dale PER (2003). Mangrove distribution and mosquito control: transport of Avicennia marina propagules by mosquito-control runnels in south-east Queensland saltmarshes. Estuarine Coastal and Shelf Science 56, 573–579. Breitfuss MJ (2003). The effects of physical habitat modification for mosquito control, runnelling, on selected non-target saltmarsh resources. PhD thesis, Griffith University, Queensland, Australia. Breitfuss MJ (2001). Predicting the effects of runnelling on non-target saltmarsh resources. Arbovirus Research in Australia 8, 23–29. Carlson DB and O’Bryan PD (1988). Mosquito production in a rotationally managed impoundment compared to other management techniques. Journal of the American Mosquito Control Association 4, 146–151. Chapman HF, Breitfuss MJ, Dale PER and Thomas P (2004). Influence of saltmarsh habitat modification for mosquito control on shore crab populations in south-east Queensland. Wetlands (Australia) 22, 1–10. Connolly RM (2005) Modification of saltmarsh for mosquito control in Australia alters habitat use by nekton. Wetlands Ecology and Management 13, 149–161. Dale PER (2008). Assessing impacts of habitat modification on a subtropical saltmarsh: 20 years of monitoring. Wetlands Ecology and Management 16, 77–87. Dale PER and Knight JM (2006). Managing saltmarshes for mosquito control: impacts of runnelling, Open Marsh Water Management and grid-ditching in sub-tropical Australia. Wetlands Ecology and Management 14, 211–220. Dale PER and Dale MB (2002). Optimal classification to describe environmental change: pictures from the exposition. Community Ecology 3, 19–29. Dale PER, Dale PT, Hulsman K and Kay BH (1993). Runnelling to control saltmarsh mosquitoes: long-term efficacy and environmental impacts. Journal of the American Mosquito Control Association 9, 174–181.
Ecology and management of mosquitoes
Dale PER and Hulsman K (1990). A critical review of saltmarsh management methods for mosquito control. Reviews in Aquatic Science 3, 281–311. Dale PER, Hulsman K, Easton CS and Kay BH (1989). Recent advances in habitat modification for saltmarsh mosquito control – south-east Queensland and northern New South Wales. Arbovirus Research in Australia 5, 171–177. Dale PER and Hulsman K (1988). To identify impacts in variable systems using anomalous changes: a saltmarsh example. Vegetatio 75, 27–35. Dale PER, Hulsman K, Harrison D and Congdon B (1986). Distribution of the immature stages Aedes of vigilax on a coastal salt-marsh in south-east Queensland. Australian Journal of Ecology 11, 269–278. Ferrigno F and Jobbins DM (1968). Open marsh water management. Proceedings of the New Jersey Mosquito Extermination Association 55, 104–115. Harley DO, Sleigh A and Ritchie SA (2001). Ross River virus transmission, infection, and disease: a cross-disciplinary review. Clinical Microbiology Review 14, 909–932. Hulsman K, Dale PER and Kay BH (1989). The runnelling method of habitat modification: an environment-focused tool for saltmarsh mosquito management. Journal of the American Mosquito Control Association 5, 226–234. Latchford J (1997). The effectiveness and environmental impacts of runnelling, a mosquito control technique. PhD thesis, Murdoch University, Western Australia. Local Government Association of Queensland (2002). Mosquito Management Code of Practice. Local Government Association of Queensland: Newstead. MacKenzie JS, Broom A, Hall RA, Johansen CA, Lindsay MD, Phillips DA, Ritchie SA, Russell RC and Smith DW (1998). Arboviruses in the Australian region, 1990 to 1998. Communicable Diseases Intelligence 22, 93–100. Medical Entomology, University of Sydney and Westmead Hospital http://medent.usyd.edu.au/ Meredith WH, Saviekis DE and Stachecki CJ (1985). Guidelines for Open Marsh Water Management in Delaware‘s saltmarshes – objectives, system designs, and installation procedures. Wetlands 5, 119–133. Miller M, Roche P, Yohannes K, Spencer J, Bartlett M, Brotherton J, Hutchinson J, Kirk M, McDonald A and Vadjic C (2005). Australia’s Notifiable diseases status, 2003 Annual report of the National Notifiable Diseases Surveillance System. Communicable Diseases Intelligence 25, 45–47. Morris CD, Dale PER and Standfast H (Eds) (2002). Australian Mosquito Control Manual, 2nd edn. Mosquito Control Association of Australia: Brisbane. Morton RM, Pollock BR and Beumer JP (1987). The occurrence and diet of fishes in a tidal inlet to a saltmarsh in southern Moreton Bay, Queensland. Australian Journal of Ecology 12, 217–237. Morton RM Beumer JP and Pollock BR (1988). Fishes of a subtropical Australian saltmarsh and their predation upon mosquitoes. Environmental Biology of Fishes 21, 185–194. Mylonas AD, Brown AM, Carthew TL, McGrath B, Purdie DM, Pandeya N, Vecchio PL, Collins LG, Gardner ID, DeLooze FJ, Reymond EJ and Suhrbrier A (2002). Natural history of Ross River virus-induced epidemic polyarthritis. Medical Journal of Australia 177, 356–360. Ratnayake J (2006). The valuation of social and economic costs of mosquito-transmitted Ross River virus. PhD thesis, Griffith University, Queensland, Australia. Ritchie SA and Jennings CD (1994). Dispersion and sampling of Aedes vigilax eggshells in south-east Queensland, Australia. Journal of the American Mosquito Control Association 10, 181–185. Russell RC (1993). Mosquitoes and Mosquito Borne Disease in South-eastern Australia. Department of Medical Entomology, Westmead Hospital: Westmead; and New South Wales and Department of Medicine: University of Sydney.
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Russell RC (2002). Ross River Virus: ecology and distribution. Annual Review Entomology 47, 1–31. Saintilan N and Williams R (2000). The decline of saltmarshes in south-east Australia: Results of recent survey. Wetlands (Australia) 18, 49–54. Sinclair P (1976). Notes on the biology of the saltmarsh mosquito Aedes vigilax (Skuse) in south-east Queensland. Queensland Naturalist 21, 134–9.
CHAPTER 9
Protection and management of coastal saltmarsh Pia Laegdsgaard, Jeff Kelleway, Robert J Williams and Chris Harty
Introduction As awareness of the plight of saltmarsh habitats in Australia increases, so does the imperative to protect and manage these areas effectively. This includes the restoration of degraded habitat, creation of new saltmarsh where possible and ensuring saltmarsh remnants are protected through statutory mechanisms from further loss and degradation. For these activities to occur, it is essential to understand the general ecology of the habitat and the ecological functions of the ecosystem under non-disturbed conditions. Previous chapters of this book provide the first synthesis of this knowledge for Australian saltmarshes. The purpose of this chapter is to discuss management issues and responses for Australian coastal saltmarshes. Management includes protection under planning frameworks as well as on-ground works to improve or maintain saltmarsh. Activities can include ‘do nothing’ scenarios where a saltmarsh is in good condition and requires low-key management measures such as formal identification and isolation from disturbance. Active management activities can be as simple as weed removal, or as complex as restorative efforts depending on resources and need.
Issues demanding protection and management of existing saltmarsh A crucial part of management, and formulation of appropriate management options is recognition of activities that cause disturbance (Laegdsgaard 2001). The history of the site, the cause of problems within a marsh area and the level of perturbation can be the key to effective decision-making regarding the management of a particular area. To protect saltmarsh from loss and effectively manage saltmarsh that is subject to disturbance it is necessary to know the distribution of this type of vegetation (see Chapter 10) and identify the threats to be minimised or eliminated. The main threats to saltmarshes are highlighted in the sections that follow.
Urbanisation Saltmarshes around Australia have been reclaimed as part of agricultural, industrial, port and residential development (Kratochvil et al. 1972; Saenger et al. 1977; Adam 1981; Bucher and Saenger 1991; Zann 1997; Coleman 1998; Davis and Froend 1999; Finlayson and Rea 1999). In 179
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Figure 9.1 Coastal saltmarsh associated with the Merri River estuary located in Warrnambool, Victoria. The saltmarshes are crowded in by urban development and are showing the effects of algal growth.
these circumstances it has been customary to fill or restrict tidal flow to saltmarsh (see Figure 9.1). The construction of new canal-estate suburbs poses a major threat to intertidal wetlands in many parts of Australia (Lee and Choy 2004). The persistence of development pressures on saltmarsh and mangrove wetlands, despite improved knowledge of the sensitivity of tidal wetlands, is demonstrated in Box 9.1. Where saltmarsh remains, the landward progression of saltmarshes, in response to sea level rise, is often restricted by the placement of a road or other hard structures, causing saltmarsh areas to shrink or be fragmented over time. The proximity of saltmarshes to urban settlements has increased the potential for disturbance from recreational use. Access to waterways for fishing and boating, and expansion of 4-wheel drive activities has increased the degree of disturbance.
Box 9.1
Developmental pressures on coastal saltmarsh
In 2007 a development proposal was registered for a 65-berth boat harbour development with 72 residential lots at Port Albert in Victoria covering a 26 hectare saltmarsh and mangrove tidal wetland. The development was referred to the Victorian Minister for Planning who determined in early 2008 that the development was not environmentally acceptable because of environmental effects and applicable policy. It was decided that: • the project was likely to have a range of significant effects on marine and terrestrial environmental assets, including native vegetation, natural coastal processes and ecosystems, migratory bird species protected under international agreements, and the values of the Nooramunga Marine and Coastal Park and the Corner Inlet Ramsar site; and • the project posed significant hazards and involved potential inconsistencies with statutory policy with respect to the management of acid sulphate soils and coastal vulnerability in the context of expected sea level changes. (Department of Planning and Community Development 2008).
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The expansion of residential estates in low-lying coastal land has led to new management challenges relating to the need for increased mosquito control. The saltmarsh mosquito is associated with various diseases and is therefore a priority for management. This is achieved either through pesticide application, which may be harmful to non-target insect populations, or habitat modification (mosquitoes and control methods are discussed further in Chapter 8).
Tidal restriction As part of the installation of agricultural and urban infrastructure, many saltmarshes have had their tidal regime significantly changed. Alterations to drainage and hydrology have come about through the construction of levees, culverts and floodgates, and these structures have had devastating effects on saltmarshes and their faunal communities. These effects range from habitat destruction to modification of the ecology. When estuaries are closed or tidally blocked, water levels rise as a result of localised freshwater runoff, leading to the inundation of saltmarshes for extended periods. Many succulent saltmarsh plants such as Sarcocornia spp. can only withstand short periods of inundation before the plants quickly rot and decompose (Adams and Bate 1994). Naturally-occurring flood events produce a similar effect; however, floodwater does not usually remain to cause permanent damage. Some plants may appear to die because of prolonged submergence; but if the stems of the plant remain alive, despite leaf decomposition, it is possible that the plant will survive and regenerate once water levels drop and the tidal influence is restored. If the tidal movement is not restored, the water table is substantially lowered and there is a relative drop in the surface level of the saltmarsh. This favours the establishment and spread of glycophyte species such as common reed (Phragmites australis), water couch (Paspalum vaginatum) and river clubrush (Schoenoplectus validus), and the loss of saltmarsh species (Roman et al. 1984). P. australis is tenacious and recruits easily to areas that have become tidally isolated. If allowed to persist, it can form extensive stands that restrict the movement of aquatic life and alter the ecology and function of the entire saltmarsh (Adams and Bate 1994; Windham 1995; Weinstein and Balletto 1999). Northern hemisphere studies have found that dense monotypic Phragmites stands can provide unsuitable or less preferred habitat and food for wildlife and waterfowl (Roman et al. 1984), though more recent studies have been equivocal (Chambers et al. 1999; Meyerson et al. 2000; Guntenspergen et al. 2002).
Fragmentation Fragmentation is a major factor contributing to saltmarsh habitat decline. Fragmentation refers to the splitting up of large saltmarsh areas into smaller blocks or disconnection of saltmarsh areas from each other and from other estuarine habitats. Saltmarshes have been broken up within urban areas with the result that some are much smaller than suburban house blocks. A good example of this form of fragmentation can be seen at Davistown near Gosford on the NSW central coast. Residential areas at Davistown have been built on or near saltmarsh and mangrove communities. Areas of saltmarsh have therefore been greatly reduced or bisected by roads, tracks and other hard surfaces. It is not clear, however, what the ecological consequences of fragmentation might be. Fragmentation in other habitats such as forests and grasslands has been shown to cause degradation of habitat quality and decreases in biodiversity (Collinge 1996; Harrison and Bruna 1999; Mazluff and Ewing 2001; Tscharntke et al. 2002). Rare or specialised species and species with lower dispersal capabilities are most affected (Tscharntke et al. 2002). Fragmentation is likely to have a major impact on saltmarsh foodweb dynamics. Larger predator species may require space for prey capture, particularly those that feed on the wing
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Figure 9.2 Fishing access within saltmarsh at high tide at Barwon Heads, Victoria. The access tracks, located in the foreground are completely covered by the high tide.
such as bats and hunting birds. The conversion of forested areas surrounding saltmarsh to urban areas may increase the distance between micro-bat roosting and feeding sites, with detrimental impacts on energetics (see Chapter 7). The migratory birds common to saltmarshes (see Chapter 7) are generally attracted to larger areas of saltmarsh which give better visual protection from predators (Straw 1996; Saintilan 2003). The size of breeding colonies of egrets has been positively correlated to the area of saltmarsh within 20 km radius of the breeding colony (Baxter and Fairweather 1998). Large unfragmented saltmarsh habitats with a diverse array of molluscs and crabs produce large quantities of larvae that is exported with the tide when the saltmarsh is inundated, which is an important food source for fish (Mazumder et al. 2006). In this way, saltmarshes are important in sustaining other estuarine species, many of which are of commercial significance. Crabs and molluscs are generally absent from very small and/or fragmented patches of saltmarsh (Laegdsgaard, personal observation), thereby limiting the supply larval zooplankton to estuarine fish species. Where crabs are present in small or fragmented saltmarsh patches, stable isotope analysis has shown a different trophic function to crabs in larger saltmarsh patches (Guest and Connolly 2006.) With increasing fragmentation and reduced patch size it is predicted that the foodweb will become limited to small predators and few insects (Laegdsgaard 2006). This would be particularly true if the saltmarsh patch also becomes tidally isolated. In grassland communities, fragments can become too small even to maintain viable populations of small birds (Johnson 2001).
Access to waterway Many saltmarsh areas offer a flat easy access point to favourite fishing spots or swimming holes (see Figure 9.2). Unfortunately, foot traffic through saltmarsh can also cause permanent damage. The fragility of the plants within the saltmarsh environment does not withstand trampling, and plants can be damaged leading to the creation of compacted bare areas or defined tracks through the vegetated plain.
Protection and management of coastal saltmarsh
Figure 9.3 Off-road vehicle use can cause the loss of vegetation, compaction of soil and alteration of saltmarsh hydrology.
Unauthorised public access into sensitive saltmarsh areas is a major issue in urban areas which can be very difficult to manage. Informal walking tracks through saltmarshes are quite common, particularly across smaller and more degraded patches, which may be considered by some members of the public as ‘wasteland’ or ‘weeds’. Human trampling has been shown to cause significant reductions in the number and cover of saltmarsh plants (Andersen 1995) and act as a vector for weed species (Lonsdale and Lane 1994).
Off-road vehicles The detrimental effects of all-terrain vehicles on aquatic and terrestrial habitats have been of increasing concern over the past two decades (Slaughter et al. 1990). Off-road vehicles (e.g. mountain bicycles, 4-wheel drive vehicles, trail motorbikes) traversing saltmarsh vegetation can cause localised and widespread damage (see Figure 9.3). Decrease of saltmarsh in areas of NSW and Tasmania has been directly attributed to recreational vehicle use (Kirkpatrick and Glasby 1981; Clarke 1993; Kelleway 2005). In the Sydney region, recreational vehicle use has been responsible for the loss of over 21 000 m2 of saltmarsh along the Georges River (Kelleway 2005) and it has caused major destruction to one of the largest stands of Wilsonia backhousei (listed as ‘Vulnerable’ under the NSW Threatened Species Conservation Act) on the Parramatta River (Kelleway et al. 2007). Low-growing, succulent herbs such as W. backhousei, Lampranthus tegens and more commonly, Sarcocornia quinqueflora, are particularly susceptible to damage by trampling and vehicles. Large reductions in the abundance of the fauna of saltmarsh, including crabs and molluscs, may occur and alterations to saltmarsh hydrology may persist for many years after vehicle use has ceased (Kelleway 2004, 2005). The type of disturbance to the saltmarsh depends greatly on the nature and purpose of the driving. Stem-height reductions and stem breakage are common with light traversing (restricted to a set of wheel ruts) of saltmarsh areas. This type of damage has the potential to recover within a year (Hannaford and Resh 1999). Heavy and constant traversing of saltmarsh by 4-wheel drive vehicles has the potential to create large bare areas of nothing but tyre tracks
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Figure 9.4 Car left in the saltmarsh at Cockle Bay, near Gosford on the NSW Central Coast. Extensive loss of vegetation and damage to the saltmarsh vegetation from vehicles can be seen.
where vegetation may take many years to recover, but only if there is nearby seed stock (Hannaford and Resh 1999, see Figure 9.4). In addition to removal of vegetation, soil compaction and reduction of mollusc and crab populations is likely (Kelleway 2005).
Mowing and watering Where saltmarsh is adjacent to urban development it can be subject to mowing and watering with freshwater (see Figure 9.5). Mowing can disrupt the flowering of grasses and destroy succulent species. Watering of lawns adjacent to saltmarsh reduces salinity within the saltmarsh, which in turn reduces the competitive advantage of saltmarsh plants (Zedler et al. 1990; Genders 1996; Wilson et al. 1996). Common garden plants (e.g. Kikuyu) can escape into nearby saltmarsh areas where they quickly take over in low salinity conditions (Genders 1996; Streever and Genders 1997). The competition between saltmarsh plants and terrestrial vegetation is not restricted to above-ground, as root competition is just as important. In a study on the ability of Sarcocornia quinqueflora seedlings to invade areas where pasture species had been removed, Genders (1996) found that S. quinqueflora were only successful in plots that had been weeded to remove above- and below-ground vegetation. Plots that were mown remained free of saltmarsh plants.
Dumping of litter Associated with public access to saltmarsh areas is the issue of litter dumping. Litter is often observed in and around urban saltmarshes that are out of public view but easily accessible. Commonly dumped items include garden waste and old building materials, but occasionally derelict vehicles have been found. The former is of particular concern with the potential to introduce weeds, including the lawn species buffalo grass (Stenotaphrum secundatum) and kikuyu (Pennisetum clandestinum). Recreational activities such as fishing also have the potential to introduce litter into the saltmarsh environment.
Protection and management of coastal saltmarsh
Figure 9.5 Saltmarsh adjacent to residential properties may be subject to mowing, which can disrupt the flowering of grasses while destroying succulent species.
Litter also enters saltmarshes via the tide and stormwater. Plastics, foams and wooden products easily float in and deposit on saltmarsh plants and animals. Detrimental effects of smothering are common for plants and animals in saltmarsh areas affected by litter.
Stormwater The discharge of stormwater into saltmarshes may alter the salinity and nutrient regimes of a saltmarsh. Direct stormwater discharge onto saltmarsh has promoted mangrove colonisation of the saltmarsh in several urban locations, including Sydney Olympic Park and Careel Bay, Pittwater (Saintilan, personal observation). The spread of freshwater and brackish species, including exotic weeds, may also be facilitated (Wilton et al. 2003). The incursion into saltmarshes of brackish or freshwater reeds such as common reed (Phragmites australis), Cumbungi (Typha spp.) and club rushes (e.g. Bolboschoenus spp.) often occurs at stormwater discharge sites. This phenomenon has been observed in Jervis Bay, southern NSW, (Wilton et al. 2003), the Georges River (Pickthall et al. 2004) and the Parramatta River (Kelleway et al. 2007) of metropolitan Sydney. As the incursion of these reeds is indicative of environmental disturbance, their occurrence should be monitored. Stormwater discharge may also increase the litter load, both natural and anthropogenic, which can smother low-growing plant species and introduce contaminants to the site.
Pollution Roads and tracks through saltmarsh areas can increase the pollutant load within the saltmarsh environment. Runoff of oil and petrol from vehicles into the marsh environment can be detrimental to the fauna that inhabit the area. Excessive vehicle noise may also threaten wildlife such as birds, bats and kangaroos that utilise saltmarsh habitats.
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Industrial developments near saltmarsh have the potential to introduce contaminants into the environment. Of particular concern is the introduction of metals into soft sediments where they are accumulated. Any subsequent disturbance to contaminated sediments allows the heavy metals to become available to organisms for ingestion. Detrital feeders provide a means for contaminants to be passed up through the foodwebs in affected marshes. Oil spills also have the potential to be detrimental to saltmarshes (Burns et al. 2000). Oil can coat plants and animals to the point of death. Additionally, depending on the size of the spill, areas of marsh may ultimately be eliminated.
Invasive species Under natural conditions, saltmarshes are generally less prone to invasion by terrestrial weeds than other habitats due in part to their extreme conditions of salinity and waterlogging. Nevertheless, they are still susceptible, particularly to specialised exotic species. Of concern in estuaries is the sharp rush (Juncus acutus). This species, introduced from the Mediterranean, has become widespread throughout the south-east and south-west of Australia (Auld and Medd 1987; Milford and Simons 2002; Zedler and Adam 2002; Paul and Young 2007; Greenwood and MacFarlane 2006). It has been so successful at invasion that it has been listed as a noxious weed for Australia (NAWC 2003). This species occupies the same niche as the native rush Juncus kraussii but is tougher and more resilient and easily out-competes the native species. The introduction of J. acutus into saltmarshes has altered the structure and complexity of these environments. Once J. acutus becomes established, its sharp tough cylindrical photosynthetic stems form dense impenetrable thickets so that the native rush is displaced or unable to establish. Many gastropods and other invertebrate fauna are believed to depend on J. kraussii for completion of their lifecycle and the same function is not afforded by J. acutus due to its differing structure (Harvey 2006; see also Chapter 4 this volume). Therefore, the ecosystem may be severely impacted by the invasion of J. acutus as it displaces J. kraussii in many saltmarsh ecosystems of coastal Australia. The invasive common cordgrass (Spartina anglica) was introduced into Tasmania and Victoria specifically for reclamation of land and stabilisation of mudflats (see Figure 9.6). One of Australia’s largest infestations (approx. 420 ha) occurs within the Tamar Estuary in Tasmania (Adam 1981; Pringle 1993; Department of Primary Industries and Water 2008). Despite earlier attempts to establish it in NSW, it has not become a problem in this state (Adam and Hutchings 1987). The common cordgrass can have several detrimental effects on natural environments in Australia. These effects include invading mudflats that are rich in invertebrates and producing dense monotypic stands that replace more diverse plant communities. Species such as S. quinqueflora and Samolus repens are particularly prone to competitive exclusion by S. anglica (Simpson 1995; Hedge and Kriwoken 2000). Birds have been observed to avoid S. anglica (Simpson 1995; Hedge and Kriwoken 2000), and species richness and total abundance of fauna are greater in saltmarshes dominated by native plants, compared to those dominated by S. anglica (Hedge and Kriwoken 2000). Other invasive species of note include pampas grasses (Cortaderia spp.), which have considerable salt tolerances, and groundsel bush (Baccharis halimifolia), which is reported to be expanding its distribution in Australia (Adam 2002). Groundsel bush is particularly tenacious once established. It produces many seeds and its seedlings are particularly hardy (Westman et al. 1975). Increased activity by people within saltmarsh areas can also introduce weeds through transport of seeds on clothing, pets and vehicles. Common weed species that occur in saltmarshes around Australia include buck’s horn plantain (Plantago coronopus), rock sea-lavender (Limonium binervosum), grasses (Polypogon monspeliensis) and Aster subulatus. All these
Protection and management of coastal saltmarsh
species have the potential to out-compete native saltmarsh species as P. monspeliensis has been shown to do in northern hemisphere marshes (Callaway and Zedler 1998).
Agricultural practices Many saltmarsh areas in Australia occur on private land in coastal rural districts and are disturbed by agricultural practices. Drainage canals and floodgates have altered tidal inundation and created more agricultural land at the expense of saltmarsh. In addition, some saltmarsh has been used directly for grazing. In general, these practices are not compatible with the conservation of saltmarsh biodiversity. Reclamation In areas that are adjacent to wetlands and have been reclaimed for agriculture, pasture species exclude saltmarsh plants to a point where the pasture species can no longer cope with the salinity. Saltmarsh plants cannot compete with pasture species, and therefore their expansion is limited by competition (Genders 1996; Streever and Genders 1997) in these altered environments. Tidal restriction Many coastal floodplains along the eastern coast of Australia over the last century have been radically modified by drainage and/or the installation of tidal restrictions known as ‘headworks’. The latter in turn form a fixture for floodgates (Williams and Watford 1997). In most cases the gates are hinged to shut as the tide rises and therefore restrict tidal penetration onto land that can be used for crops and grazing land in the absence of saltwater. As a result of being closed or tidally blocked, water levels rise from localised freshwater runoff leading to the inundation of saltmarshes for extended periods. Survival is dependent on individual species responses to the duration of submerged conditions. In times of limited freshwater runoff, a closed floodgate results in a substantial lowering of the water table and relative drop in the marsh surface level that favours pasture species rather than saltmarsh. Floodgates restrict more than just tidal water flow. Nelson (2006) reported on the ‘rapid and progressive’ decline in water quality and native plant communities after the construction of floodgates in the Yarrahapinni Wetland of northern New South Wales. In particular, decomposition and nutrient release is slower on the landward side of floodgates (Dick 1999) than where it is tidal. This affects nutrient exchange with possible impacts on estuarine productivity. In addition, the natural movement of fish and other aquatic life is halted or restricted by floodgates, which may also affect estuarine foodwebs and productivity (Pollard and Hannan 1994). Trampling by livestock Fauna that are native to Australia consist primarily of soft-footed macropods that have limited trampling impact on native ecosystems. However, hard-hoofed ungulates such as cattle and horses have a number of adverse effects on many native ecosystems including saltmarshes (Bridgewater 1982). Trampling by farm animals can have the effect of lowering the ground level and thereby altering the tidal regime and species composition of a given patch of saltmarsh. Constant trampling by hard-hoofed animals can easily disrupt dense vegetation and root systems, often destroying succulent chenopods, such as Sarcocornia spp. and Suaeda spp., and allowing tidal water to pool. Such pools form excellent habitat for biting insects (mosquitoes and midges: see Chapter 8) or other plant species (e.g. Triglochin striata), which are more tolerant of waterlogging and lowered salinity (Zedler et al. 1995). Trampling also introduces gaps where weeds can establish (Bridgewater 1982), which can affect the dynamics of saltmarsh communities.
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Figure 9.6
Cattle in saltmarsh at Breamlea near Torquay, Victoria.
In areas where trampling is high, regeneration of the saltmarsh plants is generally slow (Chabreck 1968; Jensen 1985). Grazing Large scale grazing by introduced farm animals is likely to have a noticeable effect on saltmarsh vegetation dynamics (see Figure 9.6). In particular, grazing may adversely affect rarer species that often live in association with saltmarshes. In a freshwater example, the exclusion of grazers from Lake Hawdon, South Australia, allowed fragile mosses and lichens to re-establish (Haywood 2000) after some years. Grazing can also limit the number of species in a given area as some plants are grazed selectively relative to other unpalatable species (Bakker 1978; Adam 1990; Haywood 2000). For example, the reduction of cover of some species (e.g. Tecticornia arbuscula) has been linked to selective grazing (Kirkpatrick and Glasby 1981; Bridgewater 1982) in saltmarsh areas. There is also anecdotal evidence that cattle will selectively feed on the propagules of the mangrove Avicennia marina (Saintilan and Williams, pers. comm.). In some areas of NSW, farm animals may actively graze on the saltmarsh species, which may reduce the relative cover of vegetation and alter the balance of species in these marshes that are not accustomed to a high level of grazing. Biomass is likely to be reduced by the presence of grazers thereby opening spaces for the proliferation of weed seeds carried in the dung of the grazing animals. Deposits of excrement from large numbers of grazing animals can increase the nutrient load in the saltmarsh that may lead to increased algal levels and oxygen depletion in shallow pools. Areas of Sporobolus virginicus used as grazing land in northern Queensland are also burnt regularly to promote new growth that is more palatable to grazing cattle (Anning 1980). Such practices may be detrimental to the invertebrate fauna of the saltmarsh area. Pollution The use of fertiliser, pesticides and herbicides is a widespread agricultural practice. These compounds can have adverse effects on saltmarsh occurring near agricultural land. Excess nutrients from fertiliser application can lead to eutrophication of nearby saltmarshes. Generally
Protection and management of coastal saltmarsh
Figure 9.7 Where Avicennia marina (White Mangrove) and Tecticornia arbuscula (Shrubby Glasswort) exhibit similar stature such as in Victoria, there is little opportunity for mangrove migration into the saltmarsh environment.
this means an increase in the algal load within the saltmarsh environment. Filamentous algal blooms can smother saltmarsh plants and animals and lower the quality of the environment. Saltmarshes support a large range of insect and spider fauna (Laegdsgaard, unpublished data) that could be adversely affected by the pesticides utilised in agricultural practices. This aspect of saltmarsh ecology is poorly understood but it is likely that the numerous flies and wasps found in saltmarshes would be particularly affected. Herbicide application could have adverse effects on the range of succulent and herbaceous plants found throughout coastal saltmarsh plains. More localised individual plant-based spraying of herbicides and pesticides may have limited effects, but any aerial application would be unable to avoid direct impact to nearby saltmarsh areas. Sea level rise and mangrove incursion Rise of sea level will have a significant impact on the distribution of plants living in and around estuaries (Vanderzee 1988; Williams 1990) with evidence of current impacts already being seen (see Chapter 3). Contingent on sedimentary regimes (e.g. Bruce et al. 2003) terrestrial vegetation, such as Swamp She-oak (Casuarina spp.) and Paperbark (Melaleuca spp.), will be forced further upstream and upslope by the rise in mean sea level and/or the change in the range of sea level within a given estuary. Intertidal vegetation, such as saltmarsh and mangrove should also move further upslope and upriver from their present locations. However, the continued survival of saltmarsh will be limited generally by topography and by structures such as seawalls, roads and buildings. The expansion of mangrove is another threat to the distribution of saltmarsh. The level of threat is heightened in urban areas where saltmarsh is already heavily fragmented (Fenech 1994; Hughes 1998; Mitchell and Adam 1989; Saintilan and Williams 1999, 2000; Evans and Williams 2001). Of interest, in higher latitudes such as in Victoria, not all coastal saltmarshes
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are threatened by mangrove migration especially where A. marina is of a size similar in stature to Tecticornia arbuscula (shrubby glasswort) (see Figure 9.7) the latter species appears to prevent significant mangrove encroachment. Chapter 3 provides further discussion and examples of this phenomenon and likely causes.
Protection through existing legislation and planning Since coastal saltmarshes are considered a unique environment it is important to conserve their biodiversity. In order to achieve this goal, legislative and planning measures that adequately define and protect saltmarsh wetlands are required. Strong legislative and planning tools also need to be supported by funding and resources for effective on-ground management of saltmarsh. The remainder of this chapter discusses the legislative and planning status of coastal saltmarsh in Australia and outlines some of the physical and social management approaches required for effective saltmarsh protection. Planning protection for coastal saltmarsh Legislation is an effective way to protect existing saltmarsh and minimise future disturbance. Australia is generally devoid of legislation that specifically protects saltmarsh. Rather, they are covered under a variety of broader policies that vary considerably between the states (see Table 9.1). These policies aim to support legislation and statutory decision-making by providing an integrated approach to coastal management by addressing catchment, social and economic issues as well as issues specific to particular coastal environments. Some states (e.g. Queensland, South Australia) do provide saltmarsh-specific legislation (i.e. Fisheries) that requires approval to damage or destroy saltmarsh vegetation because it is recognised as valuable fish habitat. In addition, NSW has listed saltmarsh as an endangered ecological community under the Threatened Species Conservation Act 1994. In the past, management efforts to protect coastal environments have failed because they deal only with local coastal issues, and do not consider the coast as well as their catchment areas as a planning unit (Wolanski 2007). Therefore, state governments are moving to more integrated forms of legislative control rather than highlighting specific environments for management. For example, South Australia has moved from specifically protecting tidal vegetation under fisheries legislation to protection of native vegetation under broader native vegetation protection legislation. In some instances statutory controls exist that require a permit or approval to destroy or damage saltmarsh vegetation. Unfortunately, some such controls (e.g. NSW State Environmental Planning Policy (SEPP) 14 – Coastal Wetlands) do not explicitly prohibit clearing or damage to saltmarshes but rather ensure that any activities or development proposals take into account detrimental effects to the wetlands. The development consent authority then considers the proposal against the level of significance of identified impacts in their ruling on the application. Public and private land protection Coastal saltmarsh is found within the intertidal zone landward of mangroves and usually in the upper tidal zone located close to the high-water mark. Such a location tends to see saltmarshes straddle both crown land and private land tenures. In most states, saltmarsh located on Crown Land will generally be protected through some type of reservation status such as a nature reserve, flora and fauna reserve or as conservationoriented public open space. However, the level of protection can vary depending on what is known about the conservation significance of the coastal saltmarsh. In other cases saltmarsh may be protected through a Crown Land consent process whereby a development proposal,
Protection and management of coastal saltmarsh
which may impact on the saltmarsh, will require approval from the relevant state government agency. The statutory planning system is capable of offering protection to saltmarsh located on public land and private land. For example, many planning systems allow statutory planning instruments to extend to low-water mark and some even further offshore (e.g. planning schemes in Port Phillip Bay in Victoria which extend 600 metres beyond low-water mark into the Bay). In addition, coastal saltmarsh occurring on public or private freehold land is afforded potential protection through appropriate zoning, land use, development controls and planning policy directions (Harty 2001, 2004, 2005). Land use and development control systems relate only to new proposals and are very poor at addressing existing land uses and developments. Therefore, coastal saltmarsh occurring on private freehold land is not generally managed for conservation purposes and these saltmarshes face a greater pressure of habitat loss and degradation. In such cases, the landholders need to be encouraged to implement management actions as described in other parts of this chapter. Protection of saltmarshes around Australia, through land use planning, operates through a multi-layered hierarchy from a national and state level through to regional and local levels. Local planning schemes are required to be consistent with state and regional policy. Box 9.2 provides some examples of this hierarchical approach. The decision-making process on individual land use and development proposals, which may impact on coastal saltmarshes is based on a balanced consideration not only of environmental but also social and economic effects. This approach can make decision-making for the
Box 9.2 Heirarchical approach to planning for saltmarsh protection The following are examples of multi-layered hierarchy of state, regional and locallevel planning frameworks that can offer protection for coastal saltmarsh around the country: • In New South Wales, the NSW Coastal Policy, 1997 – Sydney Regional Environmental Plan (Sydney Harbour Catchment) 2005 and Local Environmental Plan template. This has statutory backing under Section 117 of the Environmental Planning and Assessment Act 1979. • In Victoria, the Victorian Coastal Strategy 2002 and draft 2007 – Central West Estuaries Coastal Action Plan 2005 – Local Government Planning Schemes, like the Casey Planning Scheme covering Western Port Bay. This has statutory backing under the Coastal Management Act 1995. • In South Australia, the Living Coast Strategy for South Australia, 2004 – Planning Strategy for Metropolitan Adelaide 2006 – Local Government Development Plans, e.g. Land Not Within A Council Area (Coastal Waters) Development Plan. All the above instruments contain strong policies focussed on the need to protect coastal saltmarsh vegetation and habitat. The benefit of these strategic policy directions founded by legislation is that they must not be compromised at a local planning level. The statutory planning instruments will contain a set of state or council wide planning policies and directives, which are to be advanced and supported in local planning provisions. They are to be considered in the way zones and development controls are applied and when land use and development proposals are assessed and determined.
191
Declaration of coastal saltmarsh an ‘Ecologically Endangered Community’ under the Threatened Species Conservation Act 1995
No
No; but captured within specific marine plant provisions of FA 1994
Northern Territory
Queensland
State Wetlands Policy
DPI Saltmarsh Management Guidelines
Saltmarshspecific policy
Fisheries Act 1994 re protection of all marine plants, irrespective of tenure, below Highest Astronomical Tide
Planning Act 2007 Re. Conservation zoning in the NT Planning Scheme (Mangroves)
State Environment Planning Policy 14 – Coastal Wetlands
Coastal Policy
Draft Marine Protected Areas Strategy
Draft Coastal and Marine Biodiversity Conservation Strategy
NT Coastal Management Policy
Self-assessable codes and/or development approvals required before removal, destruction or disturbance may be undertaken.
No
Local Environmental Plans informed using the Model LEP template.
Sydney Regional Regional Coastal Management Strategy Environmental Plan for Sydney (Sydney Harbour Catchment) 2005
DPI Habitat Management Guidelines
Wetland Management Policy
Statutory controls
General policy
Provisions within DPI&F Fish Habitat Management Policy FHMOP 001 Operational Policy on Marine Plants and other Fish Habitats FHMOP001 (2007)
Territory Parks and Wildlife Conservation No Act 2006, Section 37: Declaration of area of essential habitat
Environmental Planning and Assessment Act 1979 re the ‘7 part test’ Rivers and Foreshores Improvement Act
Local Government Act 1993 re annual state of environment reporting
Fisheries Management Act 1999 re reclamation and dredging
Saltmarsh-specific legislation Other relevant legislation
New South Wales
Domain
Field Guide to Common Saltmarsh Plants of Queensland (2006)
Saltmarsh communities under threat from coastal development and from landward transition of mangrove communities
Contact NT Herbarium for species information
Comments
Table 9.1 State and territory legislation relevant to coastal saltmarsh in Australia. Data from Harty (2001, 2004, 2005) and personal communications 2008 from J Beumer (Qld.), P Coleman (South Australia), A Kitchener (Tasmania), R Lawrie (Western Australia), J Martin (Northern Territory).
192 Australian Saltmarsh Ecology
No
Tasmania
*unless definition includes vegetation such as saline aquatic herbland, then: • Nature Conservation Act 2002 lists Threatened native vegetation communities • Forest Practices Act 1985
No
Saltmarshspecific policy
• Land Use Planning & Approvals Act 1993 • Environmental Management and Pollution Control Act 1999 • Water Management Act 1993 • State Policy & Projects Act 1993 • Threatened Species Protection Act 1995 • Forest Practices Act 1985 (vulnerable land)
Native Vegetation Act 1991 of South Australia. Any clearance of native vegetation that occurs outside the metropolitan area must be conducted with a clearance permit. In recompense offsets, compensation, and other mitigation will be required. The Act also takes over the protection of mangroves since the superceding of the Fisheries Act 1984 with the Fisheries Management Act of 2007
National Parks and Wildlife Act 1972 of South Australia. Protects scheduled plants (including some saltmarsh species) that occur on Crown Lands and in NPWS reserves
No
Commonwealth Environment Protection No and Biodiversity Conservation Act 1999. Protection of Halosarcia flabelliformis, a samphire of the supratidal zone, coastal playas and inland salt lakes
Saltmarsh-specific legislation Other relevant legislation
South Australia
Domain
General policy
Statutory controls
Developments that will have a discharge to a wetland, watercourse or the marine zone are subject to licensing by the Environment Protection Authority
Developments that will necessitate the clearance of native vegetation must be referred to the Native Vegetation Council.
Developments occurring in the Coastal Zone as defined in Council Development Plans (under the Development Act 1993) must be referred to the Coast Protection Board for review.
• Local Government • State Coastal Policy • Waterways & Wetlands Works Manual Planning Schemes 2003 • Marine Protected Areas Strategy • National Reserve System Directions Statement (Targets for reservation of vegetation communities)
Coast Protection Board Policies, specifically those relating to: • development in the coastal zone • conservation of coast and marine habitats including coastal wetlands, rivers and estuaries • heritage and landscape acid sulphate soil development guidelines and risk assessment criteria.
Both of these planning strategies include strategic planning policy directions for the protection of saltmarshes and mangroves.
Planning Strategy for Outer Metropolitan Adelaide, 2006
Planning Strategy for Metropolitan Adelaide, 2006
Wetlands Strategy for South Australia
Draft Estuaries Policy and Action Plan for South Australia
*The definition of saltmarsh being used here is relevant. A broader definition might include saline aquatic herbland (AHS) which is currently listed on the Threatened Native Vegetation Communities List, Nature Conservation Act 2002. The undifferentiated wetland mapping category (AWU) is also listed and may include some saltmarsh.
Comments
Protection and management of coastal saltmarsh 193
No
Victoria
Likewise the National Parks Act 1975 provides for coastal saltmarshes to be included with coastal based national parks and marine national parks.
Crown Land (Reserves) Act 1978 provides for the reservation of crown land for specified purposes including for conservation purposes and can include coastal areas containing saltmarshes that are not within national parks.
Environment Protection Act 1970 provides for the State Environment Protection Policy for Waters of Victoria which protects water quality and regulates pollution discharges on the coast.
Coastal Management Act 1995 provides for the regulation of development on coastal crown land (crown land within 200 metres of the tidal limit) and includes protection of saltmarsh through a consent process.
Flora and Fauna Guarantee Act 1988 provides for the protection of saltmarsh on public land. Protection is provided through a consent process for removal as well as listing of threatening processes which include Spartina as an invasive weed.
Planning and Environment Act 1987 provides for municipal planning schemes which regulate removal of native vegetation including coastal saltmarsh on private land.
Saltmarsh-specific legislation Other relevant legislation
Domain No
Saltmarshspecific policy Western Port Bay and Corner Inlet support the largest stands of saltmarsh but central Victoria west of Melbourne has the most diverse and species rich coastal saltmarsh. Threats to saltmarshes continue from urban development, altered runoff, landfilling. grazing and weed invasion mainly from Spartina. Future threats to saltmarsh relate to sea level rise and ‘coastal squeeze’ with no opportunity to migrate landwards due to obstructions. Although there is some evidence of mangrove expansion into saltmarshes, it is not considered significant.
Comments
Statutory controls State Environment Protection Policy – Waters of Victoria provides for the protection of water quality.
Victorian Biodiversity Strategy recognises saltmarshes, mangroves and Environmental seagrasses as important estuarine Significance Overlays habitat and requiring protection. with Schedules that Regional plans such as the Central West relate to wetlands or Victoria Estuaries Coastal Action Plan, coastal habitat 2005 recognises the importance of including coastal coastal saltmarsh as estuarine habitat saltmarshes under local and provides policy for its protection. government planning Municipal Strategic Statements in all schemes which planning schemes contains planning provides for protection policy for the protection and and regulates enhancement of environmental values development. with some coastal Councils, such as Casey City Council on Western Port Bay including reference to saltmarsh vegetation and habitat.
Victorian Coastal Strategy recognises saltmarshes, mangroves and seagrasses as important estuarine habitat and requiring protection.
General policy
194 Australian Saltmarsh Ecology
Western Australia
Domain
No
Saltmarshspecific policy
• Conservation and Land Management No Act 1984 • Environmental Protection Act 1986 • Planning and Development Act 2005 • Town Planning and Development Act 1928 • Water and Rivers Commission Act 1995 • Waterways Conservation Act 1976 • Western Australian Land Authority Act 1992 • Wildlife Conservation Act 1950
Saltmarsh-specific legislation Other relevant legislation • Wetlands Conservation Policy for Western Australia 1997 (under review) • Environmental Protection of Wetlands, Environmental Protection Authority (EPA) Position Statement No. 4 (2004) • Environmental Offsets, EPA Position Statement No. 9 • Environmental Guidance for Planning and Development, EPA Draft Guidance Statement No. 33 (2005) • Water Resources, WA Planning Commission Statement of Planning Policy (SPP) No. 2.9 • State Coastal Planning, WA Planning Commission SPP No. 2.6
General policy
Statutory controls Environmental Protection (Clearing of Native Vegetation) Regulations 2004
Comments
Protection and management of coastal saltmarsh 195
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Box 9.3
Protection of saltmarsh supported by policy guidelines
An example of strong protective policies for coastal saltmarshes supported in the assessment of a development application is noted in a recommendation in August 2006, from the Development Assessment Commission, South Australia. They decided not to grant concurrence to support a decision by the District Council of Cleve to approve a land division application at Arno Bay (Development Assessment Commission, 2006). In particular, it was considered that the proposal was not appropriate because of the vegetation on the land being a prime example of a mangrove wetland and samphire marsh transitioning into coastal mallee scrub. The proposal would have interfered with environmentally important features of coastal areas, including mangroves, wetlands, stands of native vegetation, wildlife habitats and estuarine areas. An increase in human activity in the locality would lead to further incremental loss of native vegetation. Such occurrences would adversely affect the on shore coastal environment through an increase in pollution via stormwater/wastewater runoff and the damage or depletion of biological and ecological resources.
protection of coastal saltmarsh from the adverse effects of development more difficult when developers argue that their proposals will generate funds and manpower to support environmental improvement. Such arguments often lead to compromises where conditional approval for land use and development proposals is granted. The consideration of individual land use and development proposals on their individual merit often fails to consider cumulative impacts. This can occur where either a single development might have ongoing effects that accumulate over time, and / or a number of combined stresses from different sources can exacerbate each other and have a severe combined impact (South Australian Parliamentary Environment, Resources and Development Committee 2007). The difficulty lies in being able to determine when the degree of cumulative impacts becomes enough to warrant refusal of proposals. The problem is compounded by lack of information on the quality of saltmarsh habitats, their ability to withstand impacts and the lack of legislative authority for consent authorities to pursue the assessment of cumulative impacts. Despite the above shortcomings, the process of decision-making can be effective, as evidenced in Box 9.3.
Management solutions for coastal saltmarsh The remainder of the chapter considers management solutions for degraded or endangered wetlands. These range from direct interventions in the saltmarsh to improved planning and legislative controls. Active management Limit/deny access to vehicles It is important that catchment managers identify saltmarshes experiencing degradation by recreational uses so that measures can be taken to restrict access and remediate these sites. Wetlands most at risk of attracting unauthorised vehicle use include those easily accessed from residential areas (including via fire and access trails) and those with mudflats or unvegetated
Protection and management of coastal saltmarsh
patches (Kelleway 2005). The latter are considered attractive because of the exhilaration caused when churning and splashing mud. Signs highlighting the significance and vulnerability of saltmarsh should be placed within areas accessed regularly by recreational vehicles. Where practical, saltmarsh should be fenced with these structures regularly maintained. In addition, it may be possible to impose fines or charges for driving directly on saltmarsh (e.g. under the Threatened Species Conservation Act 1994 in New South Wales). Where extensive use is unlikely to be excluded it needs to be contained within designated areas so that damage is limited. Access tracks that avoid or minimise impact to saltmarsh areas may need to be established by land managers. In this circumstance, efforts need to be made to ensure vehicle remain on designated tracks. In all cases, fences and gates must be robust and monitored continually as they are likely to be subject to vandalism. A recent example of successful vehicle control at Patonga Creek, New South Wales is further described in Box 9.4. Designated access pathways Where foot traffic is a concern, designated walking tracks may need to be set out in areas that avoid or minimise impact upon sensitive saltmarsh areas (see Figure 9.8). In many cases these tracks can follow the most direct line to the waterway, and walkers and fishers can be encouraged to use the designated track. In appropriate circumstances it may be necessary to establish raised walkways. Such boardwalks have proved successful in many wetland systems and are used by schools, universities and the general public. Often these can be linked to a bird observation point or a recreational area to increase the appeal. Signs to promote an increased awareness of wetland systems (seagrass and mangroves as well as saltmarsh) are desirable. Buffer zones Saltmarsh has been encroached on by mangrove along the south-east coast of Australia from Queensland to South Australia (Saintilan and Williams 1999; Chapter 3 in this volume). A tentative link may exist between the incursion of mangrove and a steady rise in sea level, as measured at the Fort Denison tidal gauge in Sydney Harbour (1.16 mm/year isostatically corrected rise over 81.8 years (Lambeck 2002)). Saltmarshes naturally retreat landward as tidal height increases (as outlined in Chapter 3). If space is unavailable due to local geomorphology or the presence of anthropogenic structures such as parks, buildings and roads, then saltmarsh will disappear. Buffer zones to allow for expansion of saltmarsh should be included in state regional and local planning documents, as well as in the design of new developments adjacent to saltmarsh habitats (see Box 9.5 for more details).
Box 9.4
Effective exclusion of vehicles from saltmarsh
Long-term use of a saltmarsh wetland, predominantly by trail bikes, had resulted in extensive damage to the wetland, with most of the area completely bare of vegetation. With the support of the Hawkesbury-Nepean Catchment Management Authority, Gosford Council installed a gate and fencing at the only entrance point to the wetland and signage stating the importance and sensitivity of the wetland. Importantly, the fencing was heavy-duty and extended down past the high-water mark at one end, and into a rockface at the other. Trail bike access has been successfully denied at this site for more than 12 months, with the saltmarsh showing signs of early, albeit, slow natural regeneration.
197
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Figure 9.8 Educational boardwalk and inappropriate access track at Port Wakefield at the head of Gulf St Vincent in South Australia. Keeping pedestrian traffic on defined tracks is not always possible.
Stormwater management Due to the potential to modify salinity and nutrient regimes and allow the incursion of other species of plants, no stormwater should be allowed to flow into the saltmarsh environment from adjacent urban settlements. Stormwater control measures should ensure adequate stormwater diversion and treatment within the urban development to limit impacts to the saltmarsh habitat. Outlets should be removed from saltmarsh where feasible. Pollution control Runoff from agriculture needs to be minimised with respect to excess nutrients entering the environment. Land uses where fertiliser application is necessary would need to be located so that runoff into saltmarsh is limited. Buffer areas may need to be established to limit runoff into saltmarsh areas. Where possible, aerial spraying of pesticides and/or herbicides should be limited near saltmarsh areas. If it is necessary to apply these with aircraft then it should be restricted to times when weather conditions (e.g. wind) will limit the spread onto non-target habitats. Application of herbicides and pesticides is best limited to localised spraying directly to the plants to minimise the impacts to nearby saltmarsh areas. Past industrial contamination is difficult to manage as clean up of soft sediment is difficult. However, future development should be carefully considered with respect to the potential for contamination of saltmarsh sediments. Oil spills need to be contained with booms if they occur near saltmarsh areas and threaten to wash into the marsh with the tide. Oil and chemical spills in the landward catchment need to be similarly contained.
Protection and management of coastal saltmarsh
Box 9.5
Legislative protection for saltmarsh against sea level rise
Implementation actions under the planning system to protect saltmarsh wetlands particularly from the effects of rising sea level effects and the pressure on these wetland communities to migrate landward could include the following (modified from Haines 2006): • Rezone foreshore buffers (both vertical and horizontal components) for Environmental Protection to prohibit future development. Include provisions in statutory planning instruments to prohibit land zoned for Environmental Protection to be used for urban stormwater management, bushfire management or any other purpose that may diminish the environmental value of the land. • Without changing ownership of the land, easements should be acquired over land within the tidal inundation zone. • The inclusion of ‘inundation easements’ should form part of the conditions of consent for any redevelopment of privately owned foreshore lands. • No new private infrastructure to be permitted within the area predicted to become affected by tidal inundation. • Inundation easements should be acquired over private freehold urban lots within the predicted tidal inundation zone resulting from sea level rise, either as part of development consent, or on a voluntary basis. • For areas included in the predicted tidal inundation zone, the statutory planning instrument should prohibit the construction of any new urban infrastructure. • Subdivision of urban lots within the predicted tidal inundation zone should be prohibited. • Define a tidal inundation zone and line based on the predicted effects of tidal inundation from sea level rise at the end of an appropriate planning horizon (e.g. 100 years). • Define indicative wildlife corridors between mangrove and saltmarsh vegetation and other habitat areas, and provide local provisions that constrain future development within these corridors.
Weed removal Weed removal is particularly important in the maintenance and rehabilitation of many saltmarsh habitats. Generally, the common weeds are simple to eradicate from saltmarsh areas. However, the removal of sharp rush (Juncus acutus) is particularly difficult. In an experimental study of control methods, Paul and Young (2007) found success with several methods: the physical removal of J. acutus and replanting of natives; application of the herbicide Glyphosate (50:1); and mulching of ground cover to prevent growth of J. acutus seedlings. Their study also recommended that a combination of such approaches is likely to provide the most successful outcome. More information is required on the general biology and physiology of J. acutus in order to formulate effective and permanent eradication methods. Controlling access of potential vectors (e.g. public access, vehicles) and the removal (and follow-up monitoring) of problematic weed species are key management actions for weed control. Active weed removal efforts, such as physical and chemical weed suppression, litter removal, and mass planting of native species will be required to control J. acutus in particular (Greenwood and MacFarlane 2006; Paul and Young 2007).
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Australian Saltmarsh Ecology
Mangrove control Mangrove-saltmarsh dynamics are driven by a range of global, regional and occasionally local factors, many of which we are only beginning to understand. The most effective protection of saltmarsh against widespread mangrove incursion will require dealing with sea level rise and promoting a planning system which incorporates its impacts upon tidal areas (see Box 9.5). At the local level, there are often few options available for land managers. In some rare circumstances, saltmarsh may occur where inundation of wetlands is actively controlled. An example is the Wanngal Wetland of Newington Nature Reserve along the Parramatta River, Sydney, where tidal inundation is strategically delivered via a series of weirs and gates. Water is delivered to the wetland, on the basis of detailed hydrological investigations, to favour saltmarsh and mudflat habitats over the growth of mangrove seedlings. The strategic hand-removal of mangrove seedlings up to one metre in height has been used in some urban saltmarshes as a short-term approach to minimise mangrove incursion. For example, in the Hunter estuary of New South Wales, mangrove has been removed specifically to maintain the declining mudflat/saltmarsh habitat for shorebirds at high-tide roost sites (Spencer, in prep. see also Chapter 7). An assessment of the applicability of seedling removal needs to be made on a site-by-site basis. Consideration should be given to the ecological costs (e.g. may reduce fish nursery value for species which rely on mangroves) and benefits (e.g. maintain saltmarsh biodiversity, shorebird habitat). The ongoing financial costs also need to be assessed, as initial seedling removal will only be effective if combined with continual follow-up efforts. Finally, there may also be legislative requirements. In New South Wales the removal of mangroves requires permission under the Fisheries Management Act 1994, and in most instances permits have been granted (Adam 2002). Fencing to exclude stock Actions such as fencing to remove cattle from saltmarsh areas on private land have been successful in allowing saltmarsh to regenerate naturally. Evidence from a large disturbed marsh at Kooragang Island in NSW, where cattle had been excluded to allow rehabilitation, suggests that Sarcocornia quinqueflora is able recover naturally in around five years once the disturbance is removed (Kooragang Wetland Rehabilitation Project unpublished data, Figure 9.9). Areas of saltmarsh on private land should be fenced to exclude cattle. Paddock rotation Often farmers utilise saltmarsh areas as grazing land which, if done at high intensity, can be highly detrimental to the saltmarsh plants. However, if the paddock is allowed a period without grazing activity through a rotation scheme then saltmarsh plants are able to regenerate naturally and maintain their health and biodiversity. Saltmarsh on private land can sustain grazing while also being conserved. Tidal restoration It is possible to convert existing pastureland back to tidal wetland in coastal areas, however this may be dependant upon a range of social, economic and logistical issues (Nelson 2006). Where conversion is warranted, land will often require reshaping in order to restore tidal inundation for saltmarshes to grow and flourish. It is important to understand that saltmarsh species can be sensitive to changes of a few centimetres in elevation and tidal inundation. Zonation of saltmarsh plants requires a specific combination of land gradients (to ensure inundation) and soil salinity (Clarke and Hannon 1970; Adam 1990; Zedler et al. 1995).
Protection and management of coastal saltmarsh
Figure 9.9 A rehabilitation site at Kooragang Island, NSW, five years after intervention. Cattle had reduced this area to bare earth before a fence was put up to exclude cattle and allow saltmarsh to regenerate naturally.
Callaway et al. (1997) established that hydrology and substratum are the key elements in restoration. Porous substratum drains and dries too quickly to be conducive to the growth of dominant saltmarsh species. In addition, soil salinity is a major factor determining seed germination and the ability of plants to mature in saltmarshes so the right balance of tide and freshwater are essential. Saltmarsh areas planted with Salicornia virginica (a species similar to S. quinqueflora) did not thrive in areas that were infrequently tidally flooded but retained water (Callaway et al. 1997). Appropriate drainage may be particularly difficult to achieve but has been successful in a number of circumstances. Several sites in Sydney have been reverted from parks to saltmarsh that is floristically mature and attracting bird species (Eckstein 2004; Sainty and Roberts 2004) and areas of Kooragang Island in NSW have been successfully converted from pasture back to saltmarsh in this way (Nelson 2006; Laegdsgaard, personal observation). It is usually assumed that if hydrology is restored then soils, vegetation and animals will recruit to a modified site. This may not always be the case and assistance may be required. If the site is excavated and the surface soils removed it may take longer for organic carbon levels at depth (where it is needed) to return to similar levels to natural marshes (Havens et al. 2002). It may be necessary to add organic substrate back to the marsh surface at the time of construction when relevelling is complete to speed the development of a sediment profile similar to natural marshes. The success of these suggestions depends in part on an understanding of the location of patches of saltmarsh, which implies that thorough inventories of estuarine foreshore have been carried out. This issue is addressed in the next chapter. Education The value of coastal wetlands as community education resources has long been recognised (e.g. Gilligan 1988). They provide insights into floral and faunal ecology, hydrology, environmental chemistry, and human impact, amongst others. Education should include the importance of saltmarsh to a variety of fauna from birds and bats to fish and invertebrates. The uniqueness of the environment should be emphasised along with the fragility of the vegetation with respect to disturbances. Everyone should be made to feel important in the survival of this habitat by explaining the correct treatment of the local
201
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saltmarsh environment. A good example of active education of saltmarsh environments is the Middle Beach Samphire Trail north of Adelaide in South Australia. This facility consists of a self-guided walking trail/boardwalk through a range of samphire/saltmarsh and mangrove habitats with interpretative signage and a trail brochure. The trail is used by school groups and is strongly supported by the respective local communities. Education can provide the best long-term management solution for saltmarsh conservation especially where they are in close proximity to urban development and schools. Introduction of a wetland awareness programme into community groups and the local school curriculum will raise awareness of the importance of saltmarsh, the need for conservation and the correct protocol for using saltmarsh areas for recreational purposes. Education may ensure that saltmarshes remain part of the environment for future generations. Education also plays an important role in minimising management threats. Adam (2002) identified public and land holder ignorance of environmental legislation as a major factor in failures of saltmarsh protection efforts. Landowners, land managers and the general public need to be made aware of the value of wetlands, the threats wetlands face and the ways in which ecological character can be restored. In specific and appropriate instances, boardwalks through saltmarsh, accompanied passively by interpretive signage or actively by a tour guides (e.g. Sydney Olympic Park), have played an important role in promoting understanding of this habitat type amongst the broader community. Public education is one way in which problems such as litter and unauthorised vehicle use can be addressed. Another part of the education process is the presence of appropriate signage within saltmarsh areas that are accessed by fishers and boat users. If designated tracks or restricted access points are to be adhered to, then education is the key. Rehabilitation, restoration and creation In recent times, some estuarine areas have been rehabilitated to compensate for the loss of these important coastal resources (see Box 9.6 for an example). Unfortunately, many rehabilitation efforts go undocumented and there is little measure of their success. Nevertheless, examples do exist that allow confidence in rehabilitation efforts. A major documentation of the needs and methods of saltwater wetland restoration in NSW is currently underway (NSW Department of Environment and Climate Change, in preparation). This resource will also have application throughout Australia and provide a template for similar documentation in other states. Previously, Streever (1997) identified 10 projects that are specifically aimed at saltmarsh rehabilitation in Australia. It is likely that more have since been instigated. Natural recovery of saltmarsh communities occurs after disturbance through the establishment of seedlings. The rate of recovery (if any) at a site will depend largely on the nature and extent of the original disturbance. The degree of isolation from natural habitats will also affect recolonisation. S. quinqueflora is a prolific producer of buoyant seeds dispersed by the tide (Nelson 1994), but if an entire saltmarsh field is removed and no saltmarsh habitats are nearby, then it is likely that the saltmarsh will be very slow to recover. Seeds would need to be transported from areas further away. This is especially important since S. quinqueflora has been observed as a pioneer in the re-colonisation of intertidal areas allowing for the subsequent growth of species such as S. virginicus (Nelson 2006). Several studies in Australia have examined the active transplantation of saltmarsh plants cultivated in greenhouses or taken from donor populations (Laegdsgaard 2006). Several species of saltmarsh plants can be propagated and grown for transplantation purposes (Burchett et al. 1998). Saltmarsh plants that are transplanted from donor sites into rehabilitation areas survive and spread, although often slowly (Nelson 1996; Dick 1999). The best results from restoration
Protection and management of coastal saltmarsh
are generally achieved where the environment has been prepared for the natural recolonisation or regeneration of saltmarsh plants (Nelson 1996; Burchett et al. 1998; Dick 1999; Nelson 2006). Plants that appear spontaneously in areas tend to grow better than transplanted individuals (Burchett and Pulkownik 1995). In transplantation from natural sites it is important to consider the impacts on the donor sites. The effects of harvesting may take some time to recover also. It has been found that the time required for saltmarsh plants to recover their natural densities in small plots varies with species and location within the saltmarsh. In a study using small, denuded, plots, it was found that access to tidal inundation was important for recovery of S. quinqueflora. It took longer (estimated 4–5.5 years) at higher elevations than at lower positions (14–17 months) (Laegdsgaard 2002). For S. virginicus, recovery was estimated at 4–5 years regardless of position (Laegdsgaard 2002). Saltmarsh areas that are restored using transplants from donor sites may establish a compliment of fauna faster as some may be transported in with the transplant. This is effectively inoculating the site with fauna. This has been shown to be an effective way to speed up the colonisation of sites by fauna in the USA (Brady et al. 2002). It may also be possible to stock the site with some of the invertebrates known to occur in natural saltmarshes. This requires collection of fauna from nearby natural sites if appropriate. It may be useful where a created or restored saltmarsh area is far removed from other natural areas and therefore the recruitment capacity of fauna may be limited. Direct stocking of target taxa has been successful and can allow restored areas to function like natural systems more quickly by facilitating the establishment of key species (Bradshaw 1996).
Legislative options The need for integrated planning Coastal saltmarsh is influenced by activities that occur in the adjoining estuary and the impacts of land use changes on saltmarshes and other estuarine wetlands have been identified (McLoughlin 2000; Wilton 2002; Williams and Thiebaud 2007, Wolanski 2007). Focussing planning and management efforts solely on a particular component of the estuary fails to take
Box 9.6
Rehabilitation of saltmarsh in practice
A good example of a saltmarsh restoration project where these principles have been employed is the Kooragang Wetland Rehabilitation Project on the central coast of NSW. The project was initiated in 1993 to compensate for the loss of estuarine habitat due to 200 years of clearing and filling (Buckney 1987). The project covers three sites: Tomago, Ash Island and Stockton, and it is one of the major environmental projects in NSW. The main area for rehabilitating estuarine habitats has been Ash Island where tidal flushing is being restored through the removal of culverts. In areas where tidal restriction has been removed there has been an observed reversal of the habitat degradation brought about by the tidal restriction and reclamation for agriculture (Streever et al. 1996), with a return and expansion of native saltmarsh and mangrove species. One of the reports from the project (Williams et al. 2000) showed a large-scale loss of saltmarsh from the 1950s to the 1990s.
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into account processes occurring elsewhere that would affect the ecosystem function of tidal wetlands. For example, around Australia many of the commercial port facilities have or plan to be deepened through dredging. Port dredging may alter tidal regimes and water circulation patterns and in turn may affect seagrass, mangrove and saltmarsh wetlands. Likewise, major land use changes in catchments from forested areas to cleared agricultural land or urban areas with large areas of impermeable surfaces could change runoff patterns and rates of sedimentation and nutrient loads which are likely to affect mangrove and saltmarsh wetlands. It is evident that there is no single planning instrument that can achieve all the desired environmental outcomes for saltmarsh wetlands. Regulating new developments only through the planning process will not effectively deal with existing resource management problems because planning controls have little power to control existing unsustainable land uses and minor variations in those uses. Therefore to conserve saltmarsh wetlands, a range of measures needs to be implemented that deals with the sustainable long-term use of land and water resources. The consideration of a ‘catchment to coast’ concept has, therefore, been recognised in the planning and management of tidal wetlands. These planning and management tools need to be organised in a manner that identifies and protects core saltmarsh habitats, provides an anticipatory framework to guide future land development and integrates this with sustainable land management methods. An appropriate approach could involve an overall management framework for saltmarsh protection based on core and buffer habitats. Core saltmarsh areas can be protected through zoning and development control; performance standards and monitoring; retention of land and buffer areas through the subdivision process; voluntary conservation agreements; land acquisition and property management planning. Land use can be guided and managed in buffer areas, and beyond, to ensure that coastal saltmarsh protection outcomes are achieved. Table 9.1 suggests individual states have made varying degrees of progress to these ends. Establishing reservations A more effective way to conserve saltmarsh could be achieved through inclusion within the conservation estate (examples of where saltmarsh has been included in reservations are provided in Box 9.7). The creation of new conservation areas should include saltmarsh areas, particularly where large, pristine areas remain. An appropriate means of protection for saltmarsh involves identifying and declaring marine protected areas (MPAs). State and territory governments are slowly moving towards marine protected area planning. The process usually involves undertaking a strategic statewide assessment to identify candidate areas incorporating marine geomorphic and biological values. Saltmarsh protection can also be afforded to wetlands on public or private land through international agreements, such as the Ramsar Wetland Convention. Some of the most ecologically important saltmarsh wetlands in Australia are already listed under the Ramsar Convention (e.g. Moreton Bay in Queensland, Kooragang Island in New South Wales, Western Port Bay in Victoria, Eighty Mile Beach in Western Australia, Cobourg Peninsula in Northern Territory and Moulting Lagoon Nature Reserve in Tasmania).
Conclusion Australian saltmarshes have been subjected to a range of damaging human impacts since the time of European settlement. The current restricted distribution of saltmarshes, particularly in urbanised estuaries is very much a result of human modifications. Today, many of the same impacts continue to threaten the extent and health of saltmarshes, while the impacts such as global warming and sea level rise will continue into the future.
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Box 9.7
Saltmarsh inclusion in reservations
Examples of no-take MPAs where saltmarsh habitat located on Crown Land have been protected are the Marine National Parks, and Marine Sanctuaries in Victoria, with sites located in Western Port Bay, Corner Inlet and Port Phillip Bay. Port Gawler Conservation Park was declared under the South Australian National Parks and Wildlife Act 1972 and there are aquatic reserves declared under the Fisheries Management Act 2007 such as the Barker Inlet – St Kilda Aquatic Reserve or the Towra Point Aquatic and Nature Reserve at Botany Bay, Sydney declared under the NSW Fisheries Management Act 1994. The Adelaide Dolphin Sanctuary has been declared covering the Barker Inlet/Port River estuary in the north of Adelaide under the Adelaide Dolphin Sanctuary Act 2005, and although very little of the extensive saltmarsh/samphire habitat is directly protected, it represents an example of a protected area where saltmarsh is included because it contributes to the habitat protection of a marine species.
The protection of saltmarsh requires strong, effective legislative and planning measures that adequately define and protect saltmarsh wetlands. This needs to be supported by funding and resources for land managers to effectively combat the threats facing saltmarsh. The next chapter details three of the most important aspects of saltmarsh management and protection: saltmarsh mapping, assessment and monitoring.
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CHAPTER 10
Mapping, assessment and monitoring of saltmarshes Jeff Kelleway, Robert J Williams and Pia Laegdsgaard
Introduction Saltmarsh wetlands are naturally dynamic ecosystems (Boorman 2003), with their character also influenced by human-induced modifications such as the exclusion of tide, extraction of freshwater, vegetation clearing in the catchment, discharge of effluent, mining of sand and most importantly for saltmarsh, the filling or ‘land claim’ of tidal areas. Such changes may limit the development of new saltmarshes (Adam 2002) and impact on the biological diversity within surviving saltmarshes (e.g. Bozek and Burdick 2005). Land use changes in and adjacent to saltmarsh have caused considerable losses, particularly within urban waterways. Also of concern is the degree to which surviving saltmarsh stands have become fragmented. The construction of levees, seawalls, jetties, tracks and roads, combined with the proliferation of mangroves has fragmented many large patches of urban saltmarsh into smaller units (the impacts of such fragmentation on saltmarsh ecology are discussed in Chapter 9). The main causes of the poor state of many saltmarshes are fragmentation, the growth of exotic species, presence of infill and litter, evidence of mangrove incursion and physical disturbances, particularly in urban areas. In a study of the saltmarsh of the Parramatta River/Sydney Harbour (Kelleway et al. 2007), these factors led to the categorisation of more than half of the estuary’s saltmarsh patches as being in ‘poor’ condition, with few patches considered of ‘good’ or ‘excellent’ condition. The sensitivity of saltmarsh to such disturbances has lead to the development of a range of tools by which condition assessment (e.g. Sainty and Jacobs 1997; Kessler 2006), rehabilitation (e.g. Streever 1998; Laegdsgaard 2002, 2006; Paul and Young 2005; NSW DECC 2008) and monitoring (e.g. Finlayson and Mitchell 1999; Chapman and Underwood 2000) can be undertaken to assist land managers in the conservation of saltmarsh. Despite being sensitive to many of the changes associated with urbanisation, saltmarsh in some circumstances has demonstrated a level of robustness that allows it to survive. Sydney Olympic Park, one of the largest areas of urban saltmarsh in Australia, provides an example, with extensive saltmarsh growth in areas that have been highly modified by land reclamation, the dumping of waste and un-expended munitions. Such modifications may have even favoured colonisation by opportunistic species such as Suaeda australis, which is relatively common in disturbed areas (see Adam et al. 1988). In appropriate conditions, saltmarsh plants 211
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have also been seen to colonise seawalls as well as in some instances the land behind seawalls (Kelleway et al. 2007), highlighting their occasional tolerance of human modifications. The need for a strategic planning approach for the identification, mapping and assessment of saltmarsh and mangrove wetlands has been identified (Harty 2001, 2005). At the national level, an Australian framework for natural resource management standards and targets is in place (Natural Resource Management Ministerial Council 2002). This framework has identified that critical assets, including estuarine habitats, must be identified and protected. The Australian government has developed a number of resources to assist land managers in the identification and protection process. These include map production guidelines, data management principles and an outline of best practice guidelines for natural resource assessment (Australian and New Zealand Spatial Information Council (ANZLIC) 1999, National Land & Water Resource Audit and ANZLIC 2003). To promote the effective protection of saltmarsh, natural resource managers and development assessors need to locate, map and prioritise the value of saltmarsh, to conserve existing saltmarsh, and to rehabilitate degraded wetlands. This chapter sets out some background to the mapping, assessment and monitoring of saltmarsh wetlands, and provides details of the methodologies available to carry this out.
Mapping saltmarsh The classification and mapping of vegetation is one of the most important scientific tools in modern land management (Keith 2004). It is vital to land use and conservation planning, monitoring and assessment works and rehabilitation efforts. This applies especially in the intertidal zone, where the physical, chemical and biotic conditions (as well as legislative and policy frameworks) of the aquatic and terrestrial zones merge. Ideally, vegetation mapping comprises three components: 1) recognition of vegetation type (e.g. community type, structural assemblages, species presence/cover) 2) vegetation extent (e.g. area, location) and 3) vegetation or community condition (e.g. health, threats). Such a comprehensive approach is unlikely across the whole of Australia at the present time due to the lengthy coastline, lack of sensitivity of data capture facilities and financial constraints. Estimates of the length of the Australian coastline and the cover of coastal saltmarsh are shown in Table 10.1. Examples of saltmarsh mapping efforts across Australia are shown in Table 10.2. Due to the limitations referred to above, most of these projects have concentrated on mapping either the extent or type of saltmarsh. Why map saltmarsh? 1. To fulfil legislative and compliance requirements Policies and legislation which aim to protect saltmarsh wetlands exist in most states and the Northern Territory (see Harty 2001, 2005). Part of the legislative protection may require Government agencies or land managers to map saltmarsh. For example, in New South Wales, saltmarsh is listed as an ‘Endangered Ecological Community’ (NSW Threatened Species Conservation Act 1995). This listing has two major implications with regard to mapping. First, local governments with estuarine frontage are required to report annually on the status of saltmarsh within their State of the Environment reporting (NSW Local Government Act 1993 (Section 428(2)(c)). Also, the proponent of any development that may impact upon saltmarsh is required to carry out the ‘7 part test’ as outlined under s5A of the NSW Environmental Planning and Assessment Act 1979. These legislative requirements apply to all coastal saltmarsh communities, regardless of their size and floristic composition.
Mapping, assessment and monitoring of saltmarshes
Table 10.1
Estimates of cover of Australian coastal saltmarsh. Length of shoreline (km) (Australian Bureau of Statistics 2007)
Area of saltmarsh (km2 ) (Bucher and Saenger 1991)
Queensland
7400
5322
Northern Territory
6200
5005
Western Australia
12 500
2965
Victoria
1800
125
South Australia
3700
84
New South Wales
1900
57
Tasmania
3200
37
36 735
13 595
Total
2. To inform the planning process The production of comprehensive and accurate maps of saltmarsh distribution will greatly aid in the planning and management of estuaries and their catchments. Land use planning and development decisions need to be informed of the location of sensitive habitats to ensure that ecologically sound choices are made. Recovery planning for threatened species and communities will also be developed on the basis of mapping of past, present and future projected distributions. Example – Planning for sea level rise: Depending on local sediment budgets, rise in sea level may cause saltmarsh to move upstream and upslope of its present distribution. It is therefore important that planning measures are implemented to create buffers that will provide upslope refuge for saltmarsh communities throughout an estuary. This process requires detailed and comprehensive mapping of the present-day distribution of saltmarshes. LiDAR (Light Detection And Ranging) has been identified as a powerful tool for modelling sea level rise in coastal areas (Hudson and Douglas 2006). Unfortunately, in some cases the upslope migration can be interrupted by natural topographic features as well as infrastructure including roads and buildings. 3. To inform the management process Knowledge of the distribution of saltmarshes is also required for management purposes such as controlling public access, removing weeds, and resolving ecological emergencies. Example – Oil spill planning: The day-to-day use of urban estuaries by commercial and recreational vessels, vehicular traffic on roads in an estuarine catchment, as well as industrial operations, enhance the potential for spills of oil and other toxic substances. Prior recognition of resources of ecological and social value is one key to rapid response. In Carter’s (1994) categorisation of ecological sensitivity for the Parramatta River/Sydney Harbour in relation to oil spills, he identified saltmarsh as a resource of extreme sensitivity. The distribution of saltmarsh needs to be repetitively assessed in order to ensure an appropriate response to a spill of oil or other materials. 4. To inform ecological research Various types of classification schemes have been designed for estuaries. A scheme for the estuaries of south-east Australia was devised (Roy 1984; Roy et al. 2001) in which the principal types of estuaries were identified as drowned river valleys, barrier estuaries and intermittently
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Heap et al. (2007)
Danaher (1995a) Danaher (1995b) Danaher and Stevens (1995) Bruinsma et al. (1999) Bruinsma (2000) Bruinsma (2001) Bruinsma and Danaher (2000) Bruinsma and Danaher (2001) Bruinsma and Duncan (2000) Queensland EPA (online)
Northern Territory
Queensland
South Australia
West et al. (1985) West et al. (in prep.)
Study
Whole-of-coast saltmarsh mapping
New South Wales
Domain
SI SI SI SI SI SI SI SI SI SI
SI
API API
Method
Bridgewater (1982) Adam et al. (1988) South Australian Department of Environment & Natural Resources (1994) Canty and Hille (2002) Fotheringham (1994)
Hyland and Butler (1988) Morton (1993) Ebert (1995)
Coast and Wetlands Society (1985) Adam et al. (1988) Saintilan (1997) Williams and Watford (1997) Saintilan (1998) Williams and Watford (1999) Evans and Williams (2001) Wilton (2002) Pickthall et al. (2004) West et al. (2004) Williams and Meehan (2004) Kelleway et al. (2007) Williams and Thiebaud (2007)
Site-specific saltmarsh mapping
API API
PS PS API, PS
API API API
PS PS API API API API API API API, PS API API API, PS API
Method
Table 10.2 Saltmarsh mapping projects in the Australian states and territory (derived in part from Wilton and Saintilan 2000). PS = Pedestrian Survey; API = Aerial Photograph Interpretation; SI = Satellite Imagery.
214 Australian Saltmarsh Ecology
‘Conserving Victoria’s Saltmarsh’ state-wide mapping project (commencing 2008)
Victoria
Western Australia
TASVEG state-wide vegetation mapping (underway) (see Harris & Kitchener 2005)
Whole-of-coast saltmarsh mapping
Tasmania
Domain
–
–
Sauer (1965) Backshall and Bridgewater (1981) Congdon (1981) Pen (1983) Tauss (2002) Tauss (2005)
Bridgewater (1975) Calder (1980) Bridgewater (1982) Adam et al. (1988) Vanderzee (1988) Ball (1998)
Kirkpatrick and Glasby (1981)
Site-specific saltmarsh mapping
PS PS API, PS API API, PS API, PS
PS PS PS PS PS API
API
Mapping, assessment and monitoring of saltmarshes 215
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Australian Saltmarsh Ecology
opening lagoons. Further, and irrespective of type of estuary, four types of geomorphic zone were recognised as potentially present in any given estuary. These zones are known, from upstream to downstream, as: the Riverine Channel; Fluvial Delta; Central Mud Basin, and Marine Tidal Delta. Attempts have been made to use these zones to assess the likely distribution of saltmarsh plants, and hence in conservation management (Pickthall et al. 2004, Kelleway et al. 2007). It is relatively simple to delimit the zones with the assistance of aerial photographs, topographic maps, and bathymetric contours. For conservation and management purposes it is advantageous to know whether there is a greater likelihood of encountering species of innately limited distribution or abundance in one portion of an estuary compared to another part. Mapping of saltmarshes to the species level provides data that can be used to inform our understanding of saltmarsh biogeography (e.g. Adam et al. 1988, Chapter 2 in this volume). Such knowledge has particular importance for historical and future comparisons in regard to climate change and human impact. Maps of saltmarsh need to be continually updated. Estuarine plant communities are highly dynamic and their distribution and biotic composition is likely to change, even over short periods of time. Natural events such as storms and variations in local rainfall may cause such changes, as can anthropogenic activities.
Assessment of type In order to begin the vegetation mapping process, a clear and comprehensive definition needs to be made of the resource being assessed. This will take into account the purpose of the mapping project (e.g. ecological study, to inform planning and management) and should have a scientific basis. For example, recent mapping in New South Wales (Kelleway et al. 2007) has used the definition of Coastal Saltmarsh classified by the NSW Scientific Committee which is contained within the NSW Threatened Species Conservation Act 1994. Similarly, the mapping units used in Tasmania’s TASVEG mapping project have been defined on the basis of workshops with a ‘Scientific Advisory Committee’ and a ‘Mapping Users Reference Group’ (Harris and Kitchener 2005). Such an approach will ensure that the type assessment will begin in an ecologically-sound, useful and usable format. In many cases the type assessment will extend beyond the general term of ‘saltmarsh’. Historically, many of the site-specific field surveys conducted in Australian saltmarshes have provided detail on saltmarsh phytosociology and biogeography (e.g. Bridgewater 1982; Adam et al. 1988). These have allowed some distinctions in saltmarsh structure and type to be made, generally on the basis of elevation and floristics. The most common distinction that is made is between the higher marsh, usually dominated by Juncus kraussii rushlands, and the lower marsh, usually dominated by the genera Sporobolus, Sarcocornia, Triglochin and Samolus, and in the southern saltmarshes, shrubby chenopods (Tecticornia spp.). In Tasmania, where saltmarsh plant and structural diversity is high, saltmarsh community mapping units have been divided to a finer level of detail as part of the current statewide vegetation mapping (TASVEG) (Harris and Kitchener 2005). These mapping units, derived on the basis of their phytosociology are: 1) Lacustrine Herbland, 2) Saline Aquatic Herbland, 3) Saline Sedgeland/Rushland, 4) Succulent Saline Herbland, and 5) Saltmarsh (undifferentiated). As in this example, it is important that type assessment reflects the diversity of plant species and community structure which are likely to be encountered within a mapping area. Where information on the species assemblage and distribution within a mapping area is required, greater resources will generally be required. Recent advances in the resolution and
Mapping, assessment and monitoring of saltmarshes
accessibility of aerial photography have made aerial photograph interpretation (API) a common and important tool in saltmarsh mapping (e.g. Tables 10.2 and 10.3), however, the intensity of field checks (ground truth) associated with API now needs to be assessed in relation to the capture of information on species composition. It is vital that the field sampling conducted as part of the API process to inform type assessments, be appropriate to research questions and management issues. Several API mapping efforts have employed field surveys to collect such information. The study of the Georges River (NSW) by Pickthall et al. (2004) provides more detail than most other API-derived saltmarsh maps in NSW in that it set out the species assemblage in every meadow of saltmarsh identified. Similarly, the Surveying Western Australia’s Land Edge (SWALE) project used vegetation transects to identify species composition in saltmarshes, amongst other plant communities (Tauss 2002, 2005). Kelleway et al. (2007) determined the species composition (percent of total cover) of each patch of saltmarsh in the Parramatta River/Sydney Harbour estuary. Advances in LiDAR mapping of elevation and plant structure may provide an important tool in the future with which to identify and map such distinctions in type of saltmarsh including the distribution of individual species. Again, such efforts will require an adequate level of field checking, especially in the development phase.
Assessment of extent The mapping of saltmarsh entails complications that are general to vegetation mapping. These include variations in the classification (definition) of vegetation units and the accumulation of spatial error into the exercise at the georeferencing/orthorectification and image analysis phases (Wilton and Saintilan 2000). There are also complications in saltmarsh mapping that are not necessarily seen in other terrestrial or aquatic communities. This section identifies the problems faced when mapping saltmarsh and discusses the merits of several mapping methods. 1. Challenges in mapping saltmarsh There is a range of challenges faced in the planning of saltmarsh mapping projects. The complexity of wetland habitats generally (including vegetated and bare areas, creeks and saltpans), as well as the presence of typical saltmarsh plant species in a range of habitats (e.g. intertidal flats, intertidal rock platforms, sea-cliffs), requires that mapping units be appropriately defined. Further to this, issues such as gaining access to intertidal areas can cause complications in the planning stage of a mapping project. Challenges in the mapping of saltmarsh also include the obscuring of saltmarsh by canopy and adjacent species, and the location and estimation of size of small patches. Saltmarsh patches are often located partially, and in some cases, entirely, under the canopy of mangrove species, swamp oak (Casuarina spp.) and terrestrial plants. In Sydney Harbour, for example, almost half the total area of saltmarsh occurs under mangrove canopy (Kelleway et al. 2007). This makes it impossible for API alone to provide an accurate representation of saltmarsh extent. The blocking of sight by mangroves (seaward) and terrestrial species (landward) will also limit the accuracy of distant observations in field based mapping efforts. The size range over which single units of saltmarsh can grow is immense, and this has implications for the scale at which mapping is done, the budget that can be applied, and consequently the size of saltmarsh area which can be detected. Kelleway et al. (2007) highlighted the variation in size of patches in Sydney Harbour / Parramatta River where 757 distinct saltmarsh patches were mapped, ranging in size from 59 776.05 m2 to 0.20 m2. A high degree of field inspection will be required in any mapping program that seeks to locate the smallest of patches,
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particularly as in some cases the small patches are the locations of rare species (e.g. Wilsonia backhousei, Kelleway et al. 2007). 2. Generalised mapping of historical distribution In urban estuaries, where significant changes have been made throughout the catchment, it may be advantageous to know where saltmarsh occurred prior to, or in the early stages of, urbanisation. The use of oblique photographs and/or anecdotal accounts may be practical. Historical studies based on diaries, reports and artworks of early settlers (e.g. McLoughlin 1987, 2000, 2002) have been used to highlight significant changes in the distribution of estuarine vegetation, including saltmarsh, in an urban estuary. A review of early scientific literature of an area may provide information on the distribution of saltmarsh prior to some human-induced modifications. A comparison between historical vegetation surveys (Hamilton 1919) and recent mapping (Kelleway et al. 2007) in Sydney Harbour highlighted significant changes to the structural and floristic character of many saltmarshes. These include the disappearance of large, unvegetated hypersaline areas within saltmarshes and a major decline in plant species such as Wilsonia backhousei and Selliera radicans. 3. Aerial photograph interpretation (API) mapping In Australia, systematic aerial photographic surveys commenced in the 1930s, although availability is likely to vary greatly between states, regions and estuaries. The first project to map the whole of the distribution of coastal saltmarsh for any one state was initiated by the then NSW State Fisheries in the late 1970s. Using aerial photographs taken in the late 1970s and early 1980s, a compendium of maps showing the extent of saltmarsh, as well as seagrass and mangrove in over 130 estuaries, was produced (West et al. 1985). These maps had a wide range of uses as planning (e.g. for oil spills, Carter 1994), and management tools (e.g. assessment of developments proposed in the intertidal zone). West et al. (1985) used the camera lucida technique, an analogue method in which a series of optics facilitated the tracing of a vegetation boundary from an aerial photograph to a reference map, (in their case a tracing of a 1:25 000 scale topographic map). More details on the method are available in Williams et al. (2003). More recently, and in conjunction with satellite imagery or aerial photography, geographic information systems (GIS) have been used to map saltmarsh and other coastal wetlands in Australia (see Table 10.2). Wilton and Saintilan (2000) made a series of recommendations for overcoming some of the challenges faced during the API mapping of estuarine wetlands. These include the georectification of photo images using ground control points, clear distinctions in the definition of mangrove, saltmarsh, mixed-habitat and Casuarina communities, and the employment of a scale of 1:5000 or finer to differentiate habitats. Consequent studies have shown inadequacies even at this level of resolution, with a scale of 1:1000 or finer now recommended (see Table 10.3). An issue related to accuracy of time-series mapping is the evolution of mapping technology. Williams et al. (in press) provide a useful example in referring to three studies of the cover of saltmarsh of the estuary of the Parramatta River/Sydney Harbour done in the past 20 years. API showed the cover of saltmarsh was seen to increase in each assessment, particularly between the second and third surveys (see Table 10.3). The progressive increase in the number of patches and area of saltmarsh identified across the three studies highlights an increase in the accuracy of the methods used, rather than an on-ground increase in distribution. Indeed, it is possible that in some locations saltmarsh may have disappeared or become markedly reduced in extent. Note that from the first study to the second, a large increase had occurred in the resolution of aerial photographs, but that there
Mapping, assessment and monitoring of saltmarshes
Table 10.3 Comparison of cover of saltmarsh in the estuary of the Parramatta River/Sydney Harbour over two decades (from Williams et al. in press). Air photo resolution (mm)
Tracing/ Digitising Scale
West et al. (1985)
2000
West et al. (2004) Kelleway et al. (2007)
Study
Primary locations in which saltmarsh was identified
Number of patches
Area of saltmarsh (ha)
1:25 000
Homebush Bay and upstream
3
7.3
300
1:1500
Homebush Bay and upstream
45
9.6
150
1:700 (based on pedestrian survey)
entire estuary
757
37.3
was a relatively small increase in total area and number of patches. From the second to the third survey, the resolution was marginally enhanced, but there was a four-fold increase in measured area and five-fold increase in the number of patches. The third study, while initiated with API, was a pedestrian survey designed to discover as many patches as possible regardless of size, and revealed many hundreds of patches, many of which were small in size (see Table 10.4) and/or below canopy. This highlights the importance of employing a pedestrian survey if a comprehensive map of the extent of saltmarsh in a locality, estuary or region is to be produced. 4. Pedestrian survey – extent of cover The most effective mapping approach involves a combination of API and a comprehensive pedestrian survey. Currently, the only estuary in NSW for which the presence and cover of saltmarsh has been determined by this technique is the Parramatta River/Sydney Harbour (Kelleway et al. 2007). This study identifies three mapping stages: a) Preliminary GIS mapping A preliminary assessment is undertaken with API, using the most recent digital, ortho-rectified, aerial photo-mosaic coverage of the catchment. An experienced GIS operator digitises all Table 10.4 Size distribution of field-determined saltmarsh patches in Sydney Harbour / Parramatta River (Kelleway et al. 2007). Patch area (m2 )
No. of patches
% of total patches
Cumulative total (%)
Area of patches (ha)
Cumulative area (ha)
<1
34
4.49
4.49
0.002
0.002
1–5
152
20.08
24.57
0.046
0.048
5–10
102
13.47
38.04
0.080
0.128 1.085
10–100
245
32.36
70.40
0.957
100–500
131
17.31
87.71
3.066
4.151
500–1000
34
4.49
92.20
2.366
6.517
1000–5000
43
5.68
97.88
8.431
14.948
5000–10 000
10
1.32
99.20
6.954
21.902
>10 000 (>1 ha)
6
0.79
100.00
15.404
37.306
757
100.00
100.00
37.306
37.306
Total
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Australian Saltmarsh Ecology
saltmarsh patches discernible at a scale of 1:700 or better. An additional class, ‘possible saltmarsh’ may be digitised for areas requiring field verification. A ‘presumptive map’ is produced to take into the field. b) Pedestrian survey A comprehensive field survey is required in order to locate and map all saltmarsh patches regardless of size, canopy cover and shading, saltmarsh plant density or species composition. Access to the shoreline may be gained primarily from the landward side (pedestrian), and in some instances from the water (boat). It is important that access arrangements with land managers and owners are made as early as possible prior to field inspection. Field mapping can be used to check four types of saltmarsh sites in relation to the preliminary map: ●
●
●
●
locations correctly located on the presumptive map. Field observations are used to correct the boundaries of saltmarsh on A3-sized segments of the presumptive map. locations too small to be shown accurately on the presumptive map. A detailed sketch is drawn on a separate piece of waterproof paper (with approximate dimensions, GPS locations, landmarks) to improve accuracy during the correction process. locations which were inadvertently omitted from the presumptive map. A sketch is made with dimensions, shape, orientation, GPS location(s) and proximity to landmarks. spurious or erroneous locations. These are noted and removed from the presumptive map.
Extensive pedestrian field surveys also ensure that saltmarsh growing in non-typical areas (i.e. other than intertidal flats) are identified. For example, saltmarsh was unexpectedly found on intertidal rock platforms in the bays and ‘headlands’ of the lower and occasionally upper parts of the Parramatta River/Sydney Harbour estuary (Kelleway et al. 2007). The locating of small patches is especially important where rare species such as Wilsonia backhousei might occur. c) Re-digitising of preliminary map Based upon corrections, additions and omissions noted during the field survey, the presumptive map is re-digitised to reflect the true extent of saltmarsh. This re-digitising is done at an appropriate scale (1:700 or finer) in order to capture the detail set out by the pedestrian survey. The final map can be used to calculate statistics on the number, size and distribution of saltmarsh within the estuary, and can include recordings made in the field such as plant species cover and results of condition assessments. The limitations of API in the mapping of saltmarsh became apparent in the Kelleway et al. (2007) study. These were later quantified in a comparison of presumptive (API only) and final (field derived) mapping efforts of the field-accessible saltmarsh of the Parramatta River/Sydney
Table 10.5 Correspondence between preliminary GIS mapping and pedestrian survey of Parramatta River/Sydney Harbour. Analysis restricted to field-accessible saltmarsh areas (Williams et al. in press). Category
Area (ha)
Proportion of final area (%)
Mapped: correctly identified and located
9.3
37%
Mapped: size underestimated
8.5
34%
Unmapped
7.3
29%
Spurious locations
4.9
n/a
Mapping, assessment and monitoring of saltmarshes
Harbour (see Table 10.5). This analysis found that during the initial API process, 4.9 hectares of the shoreline was incorrectly interpreted as saltmarsh. Ground-truthing identified these incorrectly mapped features to include non-tidal reedlands dominated by Phragmites australis and Typha spp., mud, water, lawn and anthropogenic structures. API was able to correctly identify and locate 37% of saltmarsh by area and correctly locate but underestimate the size of 34% of saltmarsh area, but completely missed the remaining 29% of saltmarsh area (Williams et al. in press). Much of the saltmarsh area which was missed or underestimated by API is likely to have been hidden under canopy vegetation – almost half of the estuary’s entire saltmarsh area was under mangrove, swamp oak or terrestrial tree canopies. In this case the pedestrian field survey made a substantial contribution to the correct location of saltmarsh and correct determination of patch size. 5. LiDAR mapping Light Detection and Ranging (LiDAR) technology, also known as airborne laser scanning, provides an effective tool in the production of topographic maps. Whereas traditional topographic maps generally define the lowest elevation as 2 m or 10 m, LiDAR-derived Digital Elevation Models (DEMs) generally have a vertical resolution better than 0.3 m Root Mean Square (RMS), with a horizontal resolution better than 0.6 m RMS (Hudson and Douglas 2006). As saltmarsh topography changes of the order of centimetres, it is important that the highest possible resolution data are obtained. Future improvements in LiDAR technology may assist in this regard. The greatest potential of LiDAR mapping for coastal saltmarsh wetlands will be in relation to the assessment and forecasting of climate change. When analysed in conjunction with API and pedestrian surveys, LiDAR DEMs will be provide detailed, baseline data on the status of saltmarsh wetlands relative to current sea level and allow for the modelling of future rises in sea level.
Assessment of condition As part of the saltmarsh mapping process, further assessments can be made of the condition of the saltmarsh to allow appropriate management options to be determined. Condition relates not only to ecological health and function but also to disturbance and its effects. A better understanding of condition may be achieved through the development of indicators which reflect the effect of processes on key aspects of saltmarsh structure and function. To date, the paucity of knowledge of saltmarsh ecosystem function in Australia has hindered the development of appropriate indicators. Instead, other indicators are used as surrogates of ecosystem function. In NSW, non-destructive rapid assessment methodologies have been developed in the Hawkesbury (Sainty and Jacobs 1997) and Parramatta River/Sydney Harbour (Kessler 2006; Kelleway et al. 2007) estuaries. The methodologies allow for the condition of saltmarsh wetlands to be compared between sites and over time. They also aim to identify when management actions or further study are required. By way of example, Tables 10.6 and 10.7 provide lists of the indicators and rating methodologies proposed by Laegdsgaard (unpublished data) and Kessler (2006), respectively. Condition is largely related to the disturbances and activities that affect the flora and fauna within a saltmarsh site. As described in the previous chapter there are a variety of activities that can alter the saltmarsh environment. There are a number of pressure categories that can be utilised to assess the level of disturbance within a given saltmarsh (see Table 10.6). These may be used in conjunction with flora and fauna measures to reflect an overall condition.
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Table 10.6 Visual assessment indicators relating to disturbances for condition of saltmarshes (Laegdsgaard, unpublished data). Indicator
Rating
Proximity of development
Pristine – No development nearby Near Pristine – >1 km away Slightly disturbed – within 500 m Highly impacted – adjacent or <10 m away
Evidence of disturbance (e.g. trampling, vehicle use)
Pristine – No evidence of disturbance Near Pristine – Low Slightly disturbed – Medium Highly impacted – Heavy
Presence of filamentous algae
Pristine – No filamentous algae present Near Pristine – Low (<10% marsh surface) Slightly disturbed – Medium (10–60%) Highly impacted – High (>60%)
Stormwater load
Pristine – None Near Pristine – Low Slightly disturbed – Medium (roads & tracks) Highly impacted – High (roads, outlets & development)
Reducing the pressures within a saltmarsh may ultimately be reflected in the measurements of flora and fauna indicators producing an improved condition rating.
Saltmarsh monitoring Monitoring trends in condition and extent Appropriate management of Australia’s estuarine areas and the saltmarshes within those estuaries is confounded by an absence of adequate information. Since monitoring addresses the extent of change within the environment, there is a need for scientific rigour in monitoring studies. This does not mean that an effective monitoring program is necessarily complex or expensive, but at the very least it is thorough and well-designed with clear objectives. Lack of clear objectives in past monitoring programs has led to a failure to deliver useful information when required (Finlayson and Mitchell 1999). When monitoring large, near-pristine saltmarshes it is important to have condition indicators that can be monitored over time to assess deterioration. If a decline in saltmarsh health is detected early it should be possible to remove the identified threat and halt or reverse adverse effects. The continual mapping of saltmarsh extent within a given locality, estuary or region is one way to monitor saltmarsh distribution. Given the dynamism of estuarine wetlands, and expected impacts of future sea level rise, spatial monitoring is an important tool. Wilton (2002) highlighted the spatial changes that have already occurred to many coastal saltmarshes in New South Wales. Considerable changes were identified through aerial photograph analyses, with the proportion saltmarsh decline over the study period as high as 97% (Careel Bay, Pittwater). The API method allows for the detection of spatial changes and the nature of these changes (e.g. transformations from saltmarsh to mangrove or terrestrial communities or to anthropogenic structures) to be assessed. Surface Elevation Tables (SETs) provide another useful tool for monitoring structural changes in intertidal wetlands. SETs allow for the detection of change in intertidal environments by taking high precision measurements of surface elevation over time (Rogers et al. 2005; see also Chapter 3 this volume). The monitoring of erosion and sedimentation processes
Mapping, assessment and monitoring of saltmarshes
Table 10.7 saltmarsh.
The score sheet of Kessler (2006) used for rapid condition assessment for urban
Indicator
Method of scoring
Physical site characteristics Area of site
Total 6/30
0 = <1ha 1 = 1-2ha 2 = >2ha
Tidal Flushing
0 = Debris line located in the vegetation community closest to the marine environment 1 = Debris line located between vegetation communities 2 = Debris line located in the vegetation community further from the marine environment
Evidence of edge erosion
0 = >20% of boundary length affected by erosion 1 = 5–20% of boundary length affected by erosion 2 = <5% of boundary length affected by erosion
Anthropogenic impacts Limits to site expansion
6/30 0 = >20% of boundary length unable to expand 1 = 5–20% of boundary length unable to expand 2 = <5% of boundary length unable to expand
Anthropogenic structures found within the site
0 = >10% covered with anthropogenic structures 1 = 5–10% covered with anthropogenic structures 2 = <5% of site covered with anthropogenic structures
Presence of rubbish
0 = >10% covered with anthropogenic rubbish 1 = 5–10% covered with anthropogenic rubbish 2 = <5% of site covered with anthropogenic rubbish
Fauna characteristics Evidence of crab populations
4/30 0 = <30% covered by crab burrows 1 = 30–70% covered by crab burrows 2 = >70% covered by crab burrows
Evidence of snail populations
0 = <3 snails present (per m2) 1 = 3–5 snails present (per m2) 2 = >5 snails present (per m2)
Vegetation characteristics Community distribution
14/30 0 = <3 vegetation communities present 1 = 3–4 vegetation communities present 2 = >4 vegetation communities present
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Australian Saltmarsh Ecology
Indicator
Method of scoring
Detailed species composition
0 = <4 species present
Total
1 = 4–10 species present Vegetation cover on site, including algae
Threatened species present on site Site productivity using Species Biomass Index
Evidence of mangrove intrusion
Evidence of introduced species
2 = >10 species present 0 = <30% covered by vegetation 1 = 30–60% covered by vegetation 2 = >60% covered by vegetation 0 = No 2 = Yes 0 = <30 1 = 30–60 2 = >60 0 = >10% covered by mangroves 1 = 5–10% covered by mangroves 2 = <5% covered by mangroves 0 = >10% covered by introduced species 1 = 5–10% covered by introduced species 2 = <5% covered by introduced species
in saltmarshes, as well as impacts of catchment changes and increasingly, changes associated with sea level rise, will benefit from the strategic use of SETs. Monitoring success of rehabilitation Ecosystem rehabilitation is becoming increasingly important in the modern world as more of the remaining patches of natural habitat become degraded (Chapman and Underwood 2000). However, success is rarely gauged and where some attempt to monitor success has been made it often suffers from inadequate or inappropriate sampling design (Grayson et al. 1999). Rehabilitation exercises should be treated as experiments with relevant studies to look at recovery and success built into the restoration plan. Deciding whether an area has been successfully ‘rehabilitated’ is difficult. What is the reference condition that would allow this decision to be made? There are problems with using historical condition as this implies that no changes would have occurred naturally over time. The likelihood that a site has not varied over time is low given saltmarshes are naturally dynamic systems. Comparisons are probably best made to a number of reference sites that resemble what the goal is for the restored area, that is, the end point for restoration. It is important to include reference and control sites in the sampling regime by which to track the degree of natural variability in the areas of interest and similar areas. When considering restoration, the control and reference sites have distinct purposes. Control sites can be chosen to represent the ‘before’ condition of the potential restoration site in order to track the changes occurring at the restoration site with some degree of confidence. The reference site should be chosen to approximate the condition that should be reached by the restoration site once fully restored. It is desirable to have more than one reference site to ensure a realistic level of natural variability is achieved. Control and reference sites will provide information on natural variability and should be located so that they will be exposed to the same natural stresses (floods, winds, storms) as the restoration site. In addition, the control and reference sites should be free of impacts that are not present at the restoration site.
Mapping, assessment and monitoring of saltmarshes
Monitoring needs to continue even after the geomorphic and vegetation structure of the site has been restored to ensure that ecological function is also restored. For example, a study of saltmarsh recovery in the United States showed that while vegetation structure may respond reasonably quickly (several years), it can take much longer for benthic fauna (15–25 years) and soil attributes (still dissimilar after more than 25 years) to resemble reference sites (Craft et al. 1999) It has been suggested that such time frames need to be considered in the monitoring of rehabilitation success in Australia (NSW DECC 2008). Robust conceptual models of the varying types of saltmarsh that occur around Australia should be developed to detail the trophic interactions and the links in process and function within saltmarshes, as well as between a saltmarsh and the adjacent estuary. Such models will allow for a comprehensive consideration of restoration success. The measurement of several indicators such as zooplankton dynamics and fish utilisation may be required to demonstrate the restoration of ecological function.
Conclusions There is a pressing need for the comprehensive management of saltmarsh wetlands, given their environmental sensitivity and the threats they face within the urban estuaries. Management is often best initiated, or at the very least is complemented, by mapping the extent, type and condition of the resource in question. Over extensive coastlines where urban disturbance is minimal, satellite imagery, while of low resolution compared to aerial photography, offers the advantage of inexpensive data sets that reveal large sections of coastline. In urban environments, where human disturbance is concentrated and fragmentation is likely to be high, highresolution mapping is required. The latter is best achieved by a combination of GIS mapping and intensive field surveying. The continued mapping of saltmarsh wetlands and technologies such as LiDAR and SETs have been identified as key mapping and monitoring tools, especially with the onset of sealevel rise. Assessment of type and condition of saltmarsh and the monitoring of restoration efforts will also become of increasing importance as human impact upon wetlands expands. The conservation of Australian saltmarshes will be greatly enhanced by the information that these tools provide for future planning, management, rehabilitation and monitoring efforts.
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McLoughlin LC (2000). Changes in estuarine wetlands distribution along the Parramatta River, Sydney, 1788–1940: Implications for conservation and planning. Cunninghamia 6, 579–610. McLoughlin LC (2002). Questioning assumptions and using historical data in developing an information base for estuarine management. Proceedings of the Coast to Coast 2002 Conference, pp. 281–285. Morton RM (1993). Fluctuations in wetland extent in southern Moreton Bay. In Future Marine Science in Moreton Bay. (Eds JG Greenwood and NJ Hall) pp. 145–147. School of Marine Science, University of Queensland. National Land & Water Resources Audit and ANZLIC (2003). Natural Resources Information Management Toolkit: Building Capacity to Implement Natural Resources Information Management Solutions. National Land & Water Resources Audit and ANZLIC – the Spatial Information Council on behalf of the Australian Government, Canberra, Australia. Natural Resource Management Ministerial Council (2002). National Framework for Natural Resource Management Standards and Targets. Natural Resource Management Ministerial Council: Canberra. New South Wales Department of Environment and Climate Change (2008). Saltwater Wetlands Rehabilitation Manual. NSW Department of Environment and Climate Change: Sydney, New South Wales. Paul S and Young R (2005). Experimental control of exotic spiny rush, Juncus acutus from Sydney Olympic Park: 1. Juncus mortality and re-growth. Wetlands (Australia) 23, 1–13. Pen LJ (1983). Peripheral vegetation of the Swan and Canning estuaries 1981. Department of Conservation and Environment Bulletin 113, 1–43. Pickthall J, Williams RJ, Adam P and Connolly D (2004). Part 3, Estuarine vegetation. In Shaping the Georges River Catchment; Georges River Catchment Environmental Study No. 2-Biodiversity. Volume III-Aquatic Biodiversity. (Eds RJ Williams, A Bryant and D Ledlin). Planning NSW: Sydney. http://www.planning.nsw.gov.au/plansforaction/georges_biodiversity.html Queensland Environment Protection Authority (online). Queensland Wetlands Programme. Australian Department of Environment, Water, Heritage and the Arts & Queensland Environmental Protection Agency. http://www.epa.qld.gov.au/wetlandinfo/site/PPL/ QldWetlandProgramme.html Rogers K, Saintilan N and Cahoon D (2005). Surface elevation dynamics in a regenerating mangrove forest at Homebush Bay, Australia. Wetlands Ecology and Management 13, 587–598. Roy P (1984). New South Wales estuaries-their origin and evolution. In Developments in Coastal Geomorphology in Australia (Ed. BG Thom) pp. 99–121. Academic Press: New York. Roy PS, Williams RJ, Jones AR, Yassini I, Gibbs PJ, Coates B, West RJ, Scanes PR, Hudson JP and Nichol S (2001). Structure and function of south-east Australian estuaries. Estuarine, Coastal and Shelf Science 53, 351–384. Saintilan N (1997). Tweed River Estuary: Photogrammetric Survey of Wetland Distributions, 1930–1994. Report prepared for the Tweed River Project of the Department of Land and Water Conservation, June 1997. Saintilan N (1998). Photogrammetric survey of the Tweed River Wetlands. Wetlands (Australia) 17, 74–82. Sainty GR and Jacobs SW (1997). ‘Hawkesbury-Nepean Saltmarsh Assessment’. Report to Hawkesbury Nepean Catchment Management Trust: Windsor, New South Wales. Sauer J (1965). Geographic reconnaissance of Western Australian seashore vegetation. Australian Journal of Botany 13, 39–69.
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South Australian Department of Environment and Natural Resources (1994). A Vegetation Survey of Barker Inlet, Gulf St Vincent, South Australia. Department of Environment and Natural Resources: Adelaide, South Australia. Streever WJ (1998). Kooragang Wetland Rehabilitation Project: opportunities and constraints in an urban wetland rehabilitation project. Urban Ecosystems 2, 205–218. Tauus C (2002). Surveying Western Australia’s Land Edge: Reference transects in coastal vegetation at Geraldton, Port Kennedy, Bunbury and Esperance, Western Australia. The Western Australian Herbarium (W.A. Department of Conservation and Land Management); Volunteers of the WA Herbarium’s Regional Herbaria; and Coastwest/Coastcare. Tauus C (2005). Surveying Western Australia’s Land Edge 2: Reference transects in coastal vegetation at Kalbarri, Mandurah and Albany, Western Australia. The Western Australian Herbarium (WA Department of Conservation and Land Management); Volunteers of the WA Herbarium’s Regional Herbaria; and Coastwest/Coastcare. Vanderzee MP (1988). Changes in saltmarsh vegetation as an early indication of sea-level rise. In Greenhouse: Planning for Climate Change. (Ed. GI Pearman) pp. 147–160. CSIRO Publishing: Melbourne. West G, Gallen C, Thiebaud I and Williams RJ (in prep.). A Second Inventory of Estuarine Macrophytes for NSW, Australia. NSW Department of Primary Industries. West G, Williams RJ and Laird R (2004). ‘Distribution of estuarine vegetation in the Parramatta River and Sydney Harbour, 2000’. Final report to NSW Waterways Authority. NSW Fisheries Final Report Series No. 70. West RJ, Thorogood CA, Walford TR and Williams RJ (1985). An Estuarine Inventory for New South Wales, Australia. Fisheries Bulletin 2. Department of Agriculture: New South Wales. Williams RJ, Kelleway J and Allen CB (in press). Saltmarsh of the Parramatta River: A comparison of API and pedestrian survey for the determination of cover and species composition. Cunninghamia. Williams RJ and Meehan AJ (2004). Focusing management needs at the sub-catchment level via assessments of change in the cover of estuarine vegetation, Port Hacking, NSW, Australia. Wetlands Ecology and Management 12, 499–518. Williams RJ, Meehan AJ and West G (2003). Status and trend mapping of aquatic vegetation in NSW estuaries. In Coastal GIS 2003: An Integrated Approach to Australian Coastal Issues. (Eds CD Woodroffe and RA Furness) pp. 317–346. Wollongong Papers on Maritime Policy, No. 14. Centre for Maritime Policy: Wollongong. Williams RJ and Thiebaud I (2007). ‘An analysis of changes to aquatic habitats and adjacent land-use in the downstream portion of the Hawkesbury Nepean River over the past sixty years’. Final report to the Hawkesbury-Nepean Catchment Management Authority. NSW DPI – Fisheries, Final Report Series No. 91. Williams RJ and Watford FA (1997). ‘Change in the distribution of mangrove and saltmarsh in Berowra and Marramarra Creeks, 1941–1992’. Final report to Hornsby Shire Council. Williams RJ and Watford FA (1999). ‘Distribution of seagrass, mangrove and saltmarsh in the Cowan Creek Catchment Management Area’. Final report to SHURE and the Cowan Creek Catchment Management Committee. Wilton K (2002). Coastal wetland habitat dynamics in selected New South Wales estuaries. PhD thesis, Australian Catholic University: North Sydney, Australia. Wilton KM and N Saintilan (2000). Protocols for Mangrove and Saltmarsh Habitat Mapping. ACU Coastal Wetlands Unit Technical Report 2000/01, produced for the Estuaries Branch, NSW Department of Land and Water Conservation, Sydney.
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Index
Acanthiza iredalei rosinae (Slender-billed Thornbill) 150, 153 Acanthopagrus australis (bream) 90, 135, 139 aerial photograph interpretation (API) mapping 218–19 agricultural practices 187–90 Aizoaceae 23, 42 algal production 11, 143 Ambassis jacksoniensis (Port Jackson glassfish) 135, 139, 140 Amphibolidae 77, 81 Apium annuum 26 Apium prostratum 30, 32, 35, 43 Arthritica helmsi 80 assessment of saltmarsh 216–22 Assiminea tasmanica 79, 82 Assimineidae 77, 78, 79, 80, 82 Aster Australasica 31, 42 Aster subulatus 16, 31, 186 Asteraceae 23, 42–3 Atriplex cinerea 30, 32, 40 Atriplex hypoleuca 35, 40 Atriplex paludosa 35, 40 Atriplex semibaccata 26, 27, 40 Australian Virtual Herbarium 5, 6, 15, 26, 28, 35, 40 Austratipa stipoides 26 Austrobilharzia terrigalensis 90 Australoplax tridentata 117, 118, 122 Avicennia marina 1, 14, 55, 58, 99, 136, 157, 188, 189 Baccharis halimifolia (groundsel bush) 15, 98, 187 Bankia australis 84 Batellaria australis 91 Batis agrillicola 2, 26, 29, 32, 36, 43 bats, insectivorous 157–8, 160 Baumea juncea 31 Baumea teretifolia 26, 34 Bembicium auratum 79, 82, 85, 86, 87, 89, 90, 91, 94, 97 Bembicium melanostoma 79, 82
Bembicium vittatum 79, 82 Berowra Creek (Hawkesbury River) 54, 64 birds and saltmarshes 149–57, 159–60 diversity 149–56 habitat for migratory shorebirds 150–6 population threats 156–7 threats to shorebird populations 156–7 bioregions 24, 26–7, 28, 29–35, 36, 40–4, 59, 149 bryophytes 7 Burhinus grallarius (Bush Stone Curlew) 150, 152 buffer zones 197 Cararma Inlet (Jervis Bay) 58, 64, 159 Carcinus maenus (European green crab) 98, 100, 117 Carpobrotus glaucescens 26, 34 Carpobrotus rossii 34 Cassidula zonata 47, 78, 81, 85, 86 Casuarina 68, 84, 218 Casuarina glauca 2, 4, 57 Chenopodiaceae 23, 24, 40–1 Chenopodium glaucum 40 Cisticola juncidis normani (Zitting Cisticola) 150, 153 climate change 7–8, 100–1, 104, 124, 216, 221 coastal saltmarsh 1–3, 179–205 agricultural practices 187–90 buffer zones 197 defined 1, 23 dumping of litter 184–5 education 201–2 fencing to exclude stock 200 fragmentation 181–2 grazing 188 invasive species 186–7 land protection public/private 190–1 management solutions 196–204 mangrove incursion/control 189–90, 200 mowing 184 off-road vehicles 183–4 paddock rotation 200 pathways, access 197
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Australian Saltmarsh Ecology
planning, integrated 203–4 planning protection 190, 191 pollution 185–6, 188–9, 198 protection and management 179 protection through legislation/ planning 190–6, 199, 203–4 reclamation 187 rehabilitation, restoration and creation 202–4 reservations 204, 205 sea level rise 189–90, 199 stormwater 185, 198 tidal restoration 200–1 tidal restriction 181, 187 trampling by livestock 187–8 urbanisation 179–81 vehicle access 196–7 watering 184 waterway access 182–3 weed removal 199 conservation, saltmarsh 17, 36, 75, 95–6, 103–4, 149, 190, 192, 196, 212 Corner Inlet (Vic) 12, 55, 59, 180, 194, 205 Cotula coronopifolia 6, 12, 35 crabs in saltmarsh 115–24 breeding ecology 118 diet 121–4 distribution 116–18 isotope tracer technique, examining diet using 121–2 larval release 119–20 sampling 115–16 Cryptassiminea buccinoides 47, 79, 82, 87 Currambene Creek (Jervis Bay) 58, 64, 159 Cyperaceae 23, 42 Dianella brevicaulis 26 Disphyma crassifolium 26, 42 Dissocarpus biflorus 27, 40 Distichlis distichophylla 26, 31, 34, 41 diversity, latitudinal patterns of 26–9 Dysphania littoralis 40 education 201–2 El Niño 65, 66 Eleocharis parvula 6 Ellobiidae 77, 78, 81, 89 ellobiid snails 78, 87, 90, 92, 96 Enchylaena tomentosa 40 endangered ecological community 17, 36, 75, 95–6, 103–4, 149, 190, 192, 196, 212 Enigmonia aenigatica 80 environment, saltmarsh 3–5, 7–9, 23
environmental factors 3–5, 23 climate change 7–8 pollution 8–9 Ephippiorhynchus asiaticus (Black-necked Stork) 150, 152 Epthianura crocea macgregori (Yellow Chat) 150, 153 fencing to exclude stock 200 field guides 23–4 Fimbristylis ferruginea 26, 29, 31, 33 Fimbristylis polytrichoides 29, 33 fish on Australian saltmarshes 131–44 distribution on inundated marshes 134–8 feeding 139–40 future research 140–4 sampling 141–2 species and abundances 133–4 flora of Australian saltmarshes 1–3, 5–7, 17 non-vascular 7 Fluviolanatus subtortus 80 Frankenia 6, 10, 150 Frankenia hispidula 36 Frankenia pauciflora 25, 43 Frankenia tetrapetala 35 French Island (Western Port Bay) 55, 64 Gahnia filum 25, 26, 35, 42, 45 gastropods 47, 75, 77–8, 81, 84, 85, 88, 89, 90, 91, 103, 186 geomorphic settings of saltmarsh New South Wales 54–5 South Australia 55 south-east Queensland 54 tropical northern Australia 54 Victoria 55 Western Australia 55–6 geomorphology 53–69 Glauconome plankta 80 glycophytes 9–10 Gondwana saltmarshes 5, 12, 17 grazing 188 Great Australian Bight 36 groundwater 3, 26, 61, 62, 65, 66, 67, 68–9 Gulf of Carpentaria 2, 4, 143 habitat destruction and fragmentation of 95, 181–2 modification and degradation of 95–102 molluscs as habitat modifiers 93–4 saltmarsh for migratory birds 150–6
Index
halophytes 7, 9–10 Halosarcia 6, 25, 35, 193 Hawkesbury River 54, 57, 64, 65, 119, 120 Heloecius cordiformis 51, 116, 117 Helograpsus haswellianus 51, 52, 116, 119 Hemichroa pentandra 31, 34, 43 herbicides 9, 14, 188, 189, 198 Homebush Bay 14, 16, 25, 64, 66, 67, 219 human disturbance of saltmarsh molluscs 94–102 Hydrobiidae 77, 78, 79, 83 Hydrococcus brazieri 80 Hydrocotyle bonariensis 34 hypersaline flats 2–3, 17, 56 introduced species 11–17 invasive species 98–100, 186–7 Isolepis cernua 6, 33, 42 Isolepis nodosa 26, 33, 42 isotope tracer technique, examining crabs diet using 121–2 Juncus acutus 14–15, 16, 49, 98, 186, 199 Juncus bufonius 33, 43 Juncus kraussii 6, 12, 24, 33, 35, 36, 43, 45, 48, 57, 77, 80, 84, 85, 86, 87, 90, 98, 99, 116, 117, 123, 186, 216 Juncus maritimus 12 Kanmantoo 29, 30, 36, 44 Kooragang Island (Hunter River) 64, 154, 155–6, 157, 159, 200, 201, 204 Kooweerup (Western Port Bay) 64 land protection public/private 190–1 latitudinal patterns of diversity 26–9 legislation 190–6, 199, 203–4 lichens 7, 188 LiDAR mapping 213, 217, 221, 225 Limonium australe 26, 35, 36, 43 Limosa limosa 150, 152 Limosella australis 6 Limonium solanderi 26, 27, 31, 35 Littoraria cingulate pristissini 79 Littoraria luteola 79, 82, 85, 86, 87, 90, 92, 97 Littorinidae 77, 78, 79, 82 livestock fencing to exclude 200 trampling by 187–8 Maireana brevifolia 40 management 17, 196–204 mangrove and saltmarsh 53
encroachment in south-eastern Australia 59–60 incursion/control 189–90, 200 recent interactions 58–60 relationship 1, 5 mangrove–saltmarsh interactions over the Holocene northern Australia 56–7 south-eastern Australia 57–8 mapping saltmarsh 212–16, 217–21 aerial photograph interpretation (API) mapping 218–19 historical distribution 218 legislative and compliance requirements 212–13 LiDAR mapping 221 pedestrian survey 219–21 re-digitising of preliminary map 220–1 Marramarra Creek (Hawkesbury River) 54, 64 Melaleuca 54, 58, 68, 84 Melampus bidentatus 89 Mimulus repens 26, 43 Minnamurra River 64 molluscs, ecology in Australian saltmarshes 75–104 climate change 100–1 competition 90 conservation of 94–102 as consumers 91–3 disease 90–1 epifaunal and infaunal species 77–84 geographic patterns 84 habitat, abiotic stresses and food 89–90 habitat, destruction and fragmentation of saltmarsh 95 habitat, modification and degradation of saltmarsh 95–102 as habitat modifiers 93–4 human disturbance 94–102 invasive species 98–100 life history and demographic processes 88 monitoring programs 104 natural disturbance 91 patterns among habitats within estuaries 84 patterns of abundance and diversity 84–8 patterns within saltmarshes 85–7 pollution 96–7 positive interactions 91 predation 90 as prey 93 processes affecting patterns of abundance 88–91 restoration 101–2
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Australian Saltmarsh Ecology
role in saltmarsh ecosystem 91–4 sampling 102 sea level rise 100–1 temporal patterns 88 transient species 84 monitoring saltmarsh 222–5 mosquitoes 167–76 biological control 175 breeding strategies 167–8 chemical control 174–5 control methods 170–1 disease transmission 169–70 habitat modification 171–4 larval development and ecology 168–75 Murray-Darling 35, 36 Myoporum insulare 30, 43 Nassarius burchardi 91 Nassarius jonasi 91 Neophema chrysogaster (Orange-bellied Parrot) 150, 152 New South Wales geomorphic settings of saltmarsh 54–5 State Environmental Planning Policy (SEPP) 14 – Coastal Wetlands 190 Nitraria billardierei 26 non-vascular flora 7 Onchidiidae 78, 81 Onchidina australis 47, 78, 81 Onchidium damelii 78, 81 Ophicardelus ornatus 47, 78, 81, 87, 92 Ophicardelus sulcatus 47, 78, 81, 86 paddock rotation 200 Parasesarma erythrodactyla 51, 52, 116, 117, 122 Paragrapsus laevis 51, 116 Parramatta River (NSW) 2, 64, 183, 185, 200, 211, 213, 217, 218–19, 220, 221 Patelloida mimula 79–80, 94 Peel-Harvey estuary (WA) 9, 56 Penaeus aztecus 131 pesticides 9, 188, 189, 198 Phallomedusidae 77, 81 Phallomedusa austrina 81 Phallomedusa solida 47, 78, 81, 85, 87, 88, 89, 90, 92, 95, 96, 97, 102 Phragmites australis 15, 41, 50, 98–9, 181, 185, 221 planning protection for coastal saltmarsh 190, 191 Pleuroloba quoyi 47, 78, 81, 86, 87, 88 Poaceae 23, 41
pollution 8–9, 96–7 Polypogon monspeliensis 17, 43, 186 Portulaca bicolour 29, 32, 44 Pseudogobius olorum 135, 139 Puccinellia stricta 26, 41 Quail Island (Western Port Bay) 64 Queensland, south-east geomorphic settings of saltmarsh 54 Rallus pectoralis (Lewin’s Rail) 150, 152 Rhyll (Western Port Bay) 64 Rostratula australis (Painted Snipe) 150, 152 Saccostrea glomerate 80, 86, 91, 94, 100 Salicornia virginica 17, 201 Salinator fragilis 81 Salinator rhamphidia 81 Salinator solida 78 Salinator tecta 81, 86 saline soil, inland 1–2 saltmarsh(es) agricultural practices 187–90 algal production 11, 143 assessment of condition 221–2 assessment of extent 217–21 assessment of type 216–17 buffer zones 197 characteristics 23 crab species 51 defined 1–3 distribution 53 dumping of litter 184–5 education 201–2 endangered 17, 36, 75, 95–6, 103–4, 149, 190, 192, 196, 212 environment 3–5, 7–9, 17, 23 fencing to exclude stock 200 field guides 23–4 flora 1–3, 5–7, 17 fragmentation 95, 181–2 global 1, 5, 8, 11, 12, 17 grazing 188 habitat, destruction and fragmentation of 95 habitat, modification and degradation of 95–102 and insectivorous bats 157–8, 160 introduced species 11–17 inland 1–2 invasive species 98–100, 186–7 land protection public/private 190–1 latitudinal patterns of diversity 26–9 management 17
Index
management solutions 196–204 mangrove incursion/control 189–90, 200 and mangrove, recent interactions 58–60 monitoring 222–5 mowing 184 off-road vehicles 183–4 paddock rotation 200 pathways, access 197 planning protection 190, 191 planning, integrated 203–4 pollution 185–6, 188–9, 198 pollution 8–9, 96–7 protection and management 179 protection through legislation/ planning 190–6, 199, 203–4 reclamation 187 rehabilitation, restoration and creation 202–3 reservations 204, 205 response to sea level rise 60–3 sea level rise 189–90, 199 stormwater 185, 198 structural forms 25–6 surface elevation 63–7 surviving and thriving 9–11 tidal range 4 tidal restoration 200–1 tidal restriction 181, 187 trampling by livestock 187–8 urbanisation 179–81 vegetation 1–3, 5–7, 17 vehicle access 196–7 and vertebrates 149, 159–60 and water mouse 158 watering 184 waterway access 182–3 weed removal 199 weeds 11–17 zonation 4, 25–6 see also birds and saltmarshes; coastal saltmarshes; crabs in saltmarsh; fish on Australian saltmarshes; mapping saltmarsh; molluscs, ecology in Australian saltmarshes saltmarsh plants 5–7 biogeographic patterns identified by Bridgewater and Creswell 35–6 clustering of bioregions 29–35 coastal specialists versus generalists 26 continental distribution 26–36 description of common 23–5 environment 3–4 geographic range of Australian 40–4 latitudinal patterns of diversity 26–9
species 23 sources of data and methods of analysis 28–9 structural forms 25–6 subgroups based on coastal orientation 31–5 surviving and thriving 9–11 zonation 4, 25–6 saltmarshes and sea level 60–8, 189–90, 199 below-ground productivity 66–7 changes in catchment characteristics 67 climate change 100–1 compaction/subsidence 65–6 factors influencing marsh surface elevation in Australia 63–8 groundwater 66 models 60–3 sea level rise impacts 67–8, 100–1 storms and floods 67 saltwater intrusion in northern Australia 58–9 Samolus repens 7, 10, 12, 24, 25, 26, 30, 32, 35, 44, 45, 85, 99, 186 Sarcocornia quinqueflora 2, 24, 25, 26, 35, 36, 41, 45, 48, 85, 86, 87, 89, 90, 92, 97, 99, 101, 117, 168, 183, 184, 200 Sarcocornia blackiana 31, 36 Sarcocornia virginica 17 Sclerostegia 6, 25, 35 Scylla serrata 51, 117 sea level see saltmarshes and sea level Selliera radicans 6, 26, 35, 44, 218 Sesuvium portulacastrum 2, 26, 30, 31, 42 South Alligator River (NT) 56, 65 South Australia geomorphic settings of saltmarsh 55 Spartina alterniflora 11, 12, 14, 122 Spartina anglica 5, 11, 12–14, 41, 98, 99, 101, 157, 186 Spartina townsendii 12 species diversity 36 epifaunal and infaunal 77–84 intertidal saltmarsh 1 introduced 11–17 invasive 98–100, 186–7 saltmarsh plant 23, 36 transient 84 Sphaeroma terebrans 84 Sporobolus virginicus 6, 10, 24, 25, 26, 41, 46, 57, 77, 80, 85, 86, 87, 90, 94, 97, 99, 101, 116, 117, 122, 136, 155, 168, 188 stormwater management 185, 198 structural forms of saltmarshes 25–6 Suaeda arbusculoides 26, 36
235
236
Australian Saltmarsh Ecology
Suaeda australis 24–5, 26, 35, 36, 41, 46, 48, 85, 123, 211 surface elevation, factors influencing marsh 63–7 below-ground productivity 66–7 changes in catchment characteristics 67 compaction/subsidence 65–6 groundwater 66 storms and floods 67 Surface Elevation Tables (SETs) 62–3, 67, 222 Tamar Estuary (Tas) 13, 157, 186 Tasmania geomorphic settings of saltmarsh 55 Tatea huonensis 79, 83, 86, 87, 89, 90, 92 Tatea rufilabris 79, 83, 86 Tecticornia 35, 68 Tecticornia arbuscula 1, 31, 35, 40, 188, 189, 190 Tecticornia auriculate 40 Tecticornia australasica 2, 5, 29, 32, 36 Tecticornia calyptrate 36 Tecticornia disarticulate 35–6, 40 Tecticornia indica julacea 36, 40 Tecticornia pergranulata 25, 35, 40, 46 Tecticornia pruinose 36, 40 Tecticornia tenuis 35 Telesopium telescopium 80 Tetractenos hamiltoni 90, 135 Threatened Species Conservation Act 1994 190 tidal range of saltmarsh 4 tidal restoration 200–1 Tadorna radjah (Radjah Shelduck) 150, 152
Towra Point saltmarsh 17, 47, 58, 118, 119–20, 123, 205 trampling by livestock 187–8 Triglochin striata 6, 12, 25, 26, 44, 46, 85, 187 tropical northern Australia geomorphic settings of saltmarsh 54 Ukerebagh Island (Tweed River) 64 urbanisation 1–2, 7, 61, 143, 179–81 vegetation of Australian saltmarshes 1–3, 5–7, 17 vehicles, off-road/access 183–4, 196–7 Victoria geomorphic settings of saltmarsh 55 water mouse 158 watering 184 waterway access 182–3 weed removal 199 weeds 11–17 Western Australia geomorphic settings of saltmarsh 55–6 Wilsonia humilis 26, 44 Xenostrobus securis 80, 89, 94 Xerochloa imberbis 29–30, 31, 35, 36, 41 Xeromys myoides (Water Mouse/False Water rat) 158, 159 zonation of saltmarshes 4, 25–6 Zoysia macrantha 25, 26, 42