Australia’s NationalScience Agency Ecologicalassetsofthe Southern Gulfcatchmentstoinform waterresourceassessments A technical report from the CSIROSouthern GulfWaterResource Assessmentfor the National Water Grid Linda Merrin1,Danial Stratford1, Rob Kenyon1, Jodie Pritchard1, Simon Linke1, RocioPonceReyes1,Rik Buckworth1,2, Pascal Castellazzi1, Bayley Costin1, RoyAijun Deng1, Ruan Gannon1, Steve Gao1,Sophie Gilbey1,ShelleyLachish1,HeatherMcGinness1, NathanWaltham3 1CSIRO,2Charles DarwinUniversity,3James CookUniversity A group of logos with a sun and waves Description automatically generated A black background with purple text Description automatically generated ISBN 978-1-4863-2066-0 (online) A black background with purple text Description automatically generated ISBN 978-1-4863-2065-3 (print) Citation Merrin L, Stratford D, Kenyon R, Pritchard J, Linke S, Ponce Reyes R, Buckworth R, Castellazzi P, Costin B, Deng R, Gannon R, Gao S, Gilbey S, Lachish S, McGinness H and Waltham N (2024) Ecological assets of the Southern Gulf catchments to inform water resource assessments. A technical report from the CSIRO Southern Gulf Water Resource Assessment for the National Water Grid. CSIRO, Australia. Copyright © Commonwealth Scientific and Industrial Research Organisation 2024. To the extent permitted by law, all rights are reserved and no part of this publication covered by copyright may be reproduced or copied in any form or by any means except with the written permission of CSIRO. Important disclaimer CSIRO advises that the information contained in this publication comprises general statements based on scientific research. The reader is advised and needs to be aware that such information may be incomplete or unable to be used in any specific situation. No reliance or actions must therefore be made on that information without seeking prior expert professional, scientific and technical advice. To the extent permitted by law, CSIRO (including its employees and consultants) excludes all liability to any person for any consequences, including but not limited to all losses, damages, costs, expenses and any other compensation, arising directly or indirectly from using this publication (in part or in whole) and any information or material contained in it. CSIRO is committed to providing web accessible content wherever possible. If you are having difficulties with accessing this document please contact Email CSIRO Enquiries . CSIRO Southern Gulf Water Resource Assessment acknowledgements This report was funded through the National Water Grid’s Science Program, which sits within the Australian Government’s Department of Climate Change, Energy, the Environment and Water. Aspects of the Assessment have been undertaken in conjunction with the Northern Territory and Queensland governments. The Assessment was guided by two committees: i.The Governance Committee: CRC for Northern Australia/James Cook University; CSIRO; National Water Grid (Department of Climate Change, Energy, the Environment and Water); Northern Land Council; NT Department of Environment, Parks and Water Security; NT Department of Industry, Tourism and Trade; Office of Northern Australia; Queensland Department of Agriculture and Fisheries; Queensland Department of Regional Development, Manufacturing and Water ii.The Southern Gulf catchments Steering Committee: Amateur Fishermen’s Association of the NT; Austral Fisheries; Burketown Shire; Carpentaria Land Council Aboriginal Corporation; Health and Wellbeing Queensland; National Water Grid (Department of Climate Change, Energy, the Environment and Water); Northern Prawn Fisheries; Queensland Department of Agriculture and Fisheries; NT Department of Environment, Parks and Water Security; NT Department of Industry, Tourism and Trade; Office of Northern Australia; Queensland Department of Regional Development, Manufacturing and Water; Southern Gulf NRM Responsibility for the Assessment’s content lies with CSIRO. The Assessment’s committees did not have an opportunity to review the Assessment results or outputs prior to their release. This report was reviewed by Dr Adam Liedloff (CSIRO, Australia) and Tom Vanderbyl (Principal Badu Advisory Pty Ltd, Australia). The ecology team received great support from people in the Northern Territory Government and associated agencies. They provided access to files and reports, spatial and other data, species and habitat information and they also provided the team with their professional expertise and encouragement. For the Northern Territory - Simon Cruikshank and Thor Sanders. People in private industry, universities, local government and other organisations also helped us. They include Lindsay Hutley, Clement Duvert, Keller Kopf, Erica Garcia and Colton Perna. Acknowledgement of Country CSIRO acknowledges the Traditional Owners of the lands, seas and waters, of the area that we live and work on across Australia. We acknowledge their continuing connection to their culture and pay our respects to their elders past and present. Photo Leichhardt Falls on the Leichhardt River. Source: CSIRO Director’s foreword Sustainable development and regional economic prosperity are priorities for the Australian, Queensland and Northern Territory (NT) governments. However, more comprehensive information on land and water resources across northern Australia is required to complement local information held by Indigenous Peoples and other landholders. Knowledge of the scale, nature, location and distribution of likely environmental, social, cultural and economic opportunities and the risks of any proposed developments is critical to sustainable development. Especially where resource use is contested, this knowledge informs the consultation and planning that underpin the resource security required to unlock investment, while at the same time protecting the environment and cultural values. In 2021, the Australian Government commissioned CSIRO to complete the Southern Gulf Water Resource Assessment. In response, CSIRO accessed expertise and collaborations from across Australia to generate data and provide insight to support consideration of the use of land and water resources in the Southern Gulf catchments. The Assessment focuses mainly on the potential for agricultural development, and the opportunities and constraints that development could experience. It also considers climate change impacts and a range of future development pathways without being prescriptive of what they might be. The detailed information provided on land and water resources, their potential uses and the consequences of those uses are carefully designed to be relevant to a wide range of regional-scale planning considerations by Indigenous Peoples, landholders, citizens, investors, local government, and the Australian, Queensland and NT governments. By fostering shared understanding of the opportunities and the risks among this wide array of stakeholders and decision makers, better informed conversations about future options will be possible. Importantly, the Assessment does not recommend one development over another, nor assume any particular development pathway, nor even assume that water resource development will occur. It provides a range of possibilities and the information required to interpret them (including risks that may attend any opportunities), consistent with regional values and aspirations. All data and reports produced by the Assessment will be publicly available. Chris Chilcott Project Director C:\Users\bru119\AppData\Local\Microsoft\Windows\Temporary Internet Files\Content.Word\C_Chilcott_high.jpg The Southern Gulf Water Resource Assessment Team Project Director Chris Chilcott Project Leaders Cuan Petheram, Ian Watson Project Support Caroline Bruce, Seonaid Philip Communications Emily Brown, Chanel Koeleman, Jo Ashley, Nathan Dyer Activities Agriculture and socio- economics Tony Webster, Caroline Bruce, Kaylene Camuti1, Matt Curnock, Jenny Hayward, Simon Irvin, Shokhrukh Jalilov, Diane Jarvis1, Adam Liedloff, Stephen McFallan, Yvette Oliver, Di Prestwidge2, Tiemen Rhebergen, Robert Speed3, Chris Stokes, Thomas Vanderbyl3, John Virtue4 Climate David McJannet, Lynn Seo Ecology Danial Stratford, Rik Buckworth, Pascal Castellazzi, Bayley Costin, Roy Aijun Deng, Ruan Gannon, Steve Gao, Sophie Gilbey, Rob Kenyon, Shelly Lachish, Simon Linke, Heather McGinness, Linda Merrin, Katie Motson5, Rocio Ponce Reyes, Jodie Pritchard, Nathan Waltham5 Groundwater hydrology Andrew R. Taylor, Karen Barry, Russell Crosbie, Margaux Dupuy, Geoff Hodgson, Anthony Knapton6, Stacey Priestley, Matthias Raiber Indigenous water values, rights, interests and development goals Pethie Lyons, Marcus Barber, Peta Braedon, Petina Pert Land suitability Ian Watson, Jenet Austin, Bart Edmeades7, Linda Gregory, Ben Harms10, Jason Hill7, Jeremy Manders10, Gordon McLachlan, Seonaid Philip, Ross Searle, Uta Stockmann, Evan Thomas10, Mark Thomas, Francis Wait7, Peter Zund Surface water hydrology Justin Hughes, Matt Gibbs, Fazlul Karim, Julien Lerat, Steve Marvanek, Cherry Mateo, Catherine Ticehurst, Biao Wang Surface water storage Cuan Petheram, Giulio Altamura8, Fred Baynes9, Jamie Campbell11, Lachlan Cherry11, Kev Devlin4, Nick Hombsch8, Peter Hyde8, Lee Rogers, Ang Yang Note: Assessment team as at September, 2024. All contributors are affiliated with CSIRO unless indicated otherwise. Activity Leaders are underlined. 1James Cook University; 2DBP Consulting; 3Badu Advisory Pty Ltd; 4Independent contractor; 5 Centre for Tropical Water and Aquatic Ecosystem Research. James Cook University; 6CloudGMS; 7NT Department of Environment, Parks and Water Security; 8Rider Levett Bucknall; 9Baynes Geologic; 10QG Department of Environment, Science and Innovation; 11Entura Shortened forms For more information on this figure or equation or table, please contact CSIRO on enquiries@csiro.au Units UNIT DESCRIPTION cm centimetre g gram ha hectare kg kilogram km kilometre (1000 metres) m metre mAHD metres above Australian Height Datum mm millimetre mS milliSiemens ppt parts per thousand t tonne Preface Sustainable development and regional economic prosperity are priorities for the Australian, NT and Queensland governments. In the Queensland Water Strategy, for example, the Queensland Government (2023) looks to enable regional economic prosperity through a vision that states ‘Sustainable and secure water resources are central to Queensland’s economic transformation and the legacy we pass on to future generations.’ Acknowledging the need for continued research, the NT Government (2023) announced a Territory Water Plan priority action to accelerate the existing water science program ‘to support best practice water resource management and sustainable development.’ Governments are actively seeking to diversify regional economies, considering a range of factors, including Australia’s energy transformation. The Queensland Government’s economic diversification strategy for North West Queensland (Department of State Development, Manufacturing, Infrastructure and Planning, 2019) includes mining and mineral processing; beef cattle production, cropping and commercial fishing; tourism with an outback focus; and small business, supply chains and emerging industry sectors. In its 2024–25 Budget, the Australian Government announced large investment in renewable hydrogen, low-carbon liquid fuels, critical minerals processing and clean energy processing (Budget Strategy and Outlook, 2024). This includes investing in regions that have ‘traditionally powered Australia’ – as the North West Minerals Province, situated mostly within the Southern Gulf catchments, has done. For very remote areas like the Southern Gulf catchments (Preface Figure 1-1), the land, water and other environmental resources or assets will be key in determining how sustainable regional development might occur. Primary questions in any consideration of sustainable regional development relate to the nature and the scale of opportunities, and their risks. How people perceive those risks is critical, especially in the context of areas such as the Southern Gulf catchments, where approximately 27% of the population is Indigenous (compared to 3.2% for Australia as a whole) and where many Indigenous Peoples still live on the same lands they have inhabited for tens of thousands of years. About 12% of the Southern Gulf catchments are owned by Indigenous Peoples as inalienable freehold. Access to reliable information about resources enables informed discussion and good decision making. Such information includes the amount and type of a resource or asset, where it is found (including in relation to complementary resources), what commercial uses it might have, how the resource changes within a year and across years, the underlying socio-economic context and the possible impacts of development. Most of northern Australia’s land and water resources have not been mapped in sufficient detail to provide the level of information required for reliable resource allocation, to mitigate investment or environmental risks, or to build policy settings that can support good judgments. The Southern Gulf Water Resource Assessment aims to partly address this gap by providing data to better inform decisions on private investment and government expenditure, to account for intersections between existing and potential resource users, and to ensure that net development benefits are maximised. Preface Figure 1-1 Map of Australia showing Assessment area (Southern Gulf catchments) and other recent CSIRO Assessments FGARA = Flinders and Gilbert Agricultural Resource Assessment; NAWRA = Northern Australia Water Resource Assessment. The Assessment differs somewhat from many resource assessments in that it considers a wide range of resources or assets, rather than being a single mapping exercises of, say, soils. It provides a lot of contextual information about the socio-economic profile of the catchments, and the economic possibilities and environmental impacts of development. Further, it considers many of the different resource and asset types in an integrated way, rather than separately. The Assessment has agricultural developments as its primary focus, but it also considers opportunities for and intersections between other types of water-dependent development. For example, the Assessment explores the nature, scale, location and impacts of developments relating to industrial, urban and aquaculture development, in relevant locations. The outcome of no change in land use or water resource development is also valid. The Assessment was designed to inform consideration of development, not to enable any particular development to occur. As such, the Assessment informs – but does not seek to replace – existing planning, regulatory or approval processes. Importantly, the Assessment does not assume a given policy or regulatory environment. Policy and regulations can change, so this flexibility enables the results to be applied to the widest range of uses for the longest possible time frame. It was not the intention of – and nor was it possible for – the Assessment to generate new information on all topics related to water and irrigation development in northern Australia. Topics For more information on this figure please contact CSIRO on enquiries@csiro.au not directly examined in the Assessment are discussed with reference to and in the context of the existing literature. CSIRO has strong organisational commitments to Indigenous reconciliation and to conducting ethical research with the free, prior and informed consent of human participants. The Assessment allocated significant time to consulting with Indigenous representative organisations and Traditional Owner groups from the catchments to aid their understanding and potential engagement with its requirements. The Assessment did not conduct significant fieldwork without the consent of Traditional Owners. CSIRO met the requirement to create new scientific knowledge about the catchments (e.g. on land suitability) by synthesising new material from existing information, complemented by remotely sensed data and numerical modelling. Functionally, the Assessment adopted an activities-based approach (reflected in the content and structure of the outputs and products), comprising activity groups, each contributing its part to create a cohesive picture of regional development opportunities, costs and benefits, but also risks. Preface Figure 1-2 illustrates the high-level links between the activities and the general flow of information in the Assessment. Preface Figure 1-2 Schematic of the high-level linkages between the eight activity groups and the general flow of information in the Assessment Assessment reporting structure Development opportunities and their impacts are frequently highly interdependent and, consequently, so is the research undertaken through this Assessment. While each report may be read as a stand-alone document, the suite of reports for each Assessment most reliably informs discussion and decisions concerning regional development when read as a whole. For more information on this figure please contact CSIRO on enquiries@csiro.au The Assessment has produced a series of cascading reports and information products: • Technical reports present scientific work with sufficient detail for technical and scientific experts to reproduce the work. Each of the activities (Preface Figure 1-2) has one or more corresponding technical reports. • A catchment report, which synthesises key material from the technical reports, providing well- informed (but not necessarily scientifically trained) users with the information required to inform decisions about the opportunities, costs and benefits, but also risks, associated with irrigated agriculture and other development options. • A summary report provides a shorter summary and narrative for a general public audience in plain English. • A summary fact sheet provides key findings for a general public audience in the shortest possible format. The Assessment has also developed online information products to enable users to better access information that is not readily available in print format. All of these reports, information tools and data products are available online at https://www.csiro.au/southerngulf. The webpages give users access to a communications suite including fact sheets, multimedia content, FAQs, reports and links to related sites, particularly about other research in northern Australia. Executive summary This activity (the ecology activity) seeks to determine the relative risks between different water resource development scenarios in the Southern Gulf catchments using a set of prioritised water- dependent assets. Environmental assets are selected from freshwater, marine and terrestrial habitats. The key questions that this activity seeks to address in the catchments include: • What is the main environmental context of the catchments that could influence water resource development? • What are the key environmental drivers and stressors that are currently occurring or likely to occur in the catchments (including key supporting and threatening processes such as invasive species, water quality and habitat changes)? • What are the known linkages between flow and ecology? • What are the key ecological trade-offs between different water resource developments considering impacts from potential changes in flow on species and habitats? This report provides a synthesis of the prioritised ecology assets including developing asset knowledge bases, conceptual relationships, and evidence narratives, including flow–ecology relationships, in the context of the Southern Gulf catchments. Contents Director’s foreword i The Southern Gulf Water Resource Assessment Team ii Shortened forms iii Units iv Preface v Executive summary ix Contents x Figures .............................................................................................................................. xii Tables ............................................................................................................................ xvii Part I Ecology of the Southern Gulf catchments 1 1 Introduction 2 1.1 Ecology and water resource development ......................................................................... 2 1.2 Water resource development and ecological changes ...................................................... 3 1.2.1 Flow regime change............................................................................................... 4 1.2.2 Altered longitudinal and lateral connectivity ........................................................ 4 1.2.3 Habitat modification and loss ................................................................................ 4 1.2.4 Increased invasive and non-native species ........................................................... 4 1.2.5 Synergistic and co-occurring processes both local and global .............................. 5 1.3 Ecology asset-based approach to modelling and assessment ........................................... 5 1.3.1 Identifying and prioritising assets ......................................................................... 5 1.3.2 Reviewing, conceptual modelling and developing evidence narratives ............... 7 1.3.3 Mapping the location and distribution of the assets ............................................ 8 1.3.4 Understanding flow–ecology relationships and modelling ................................... 8 2 Ecology of the Southern Gulf catchments 11 2.1 Ecology of the Southern Gulf catchments ........................................................................ 11 2.1.1 Southern Gulf catchments and its environmental values ................................... 11 2.1.2 Protected, listed and significant areas of the Southern Gulf catchments .......... 13 2.1.3 Important habitat types and values of the Southern Gulf catchments .............. 16 2.1.4 Significant species and ecological communities of the Southern Gulf catchments ............................................................................................................................. 17 2.1.5 Current condition and potential threats in the Southern Gulf catchments........ 19 3 Ecological assets from the Southern Gulf catchments and marine region 20 3.1 Fish, sharks and rays ......................................................................................................... 20 3.1.1 Barramundi (Lates calcarifer) .............................................................................. 20 3.1.2 Bull sharks (Carcharhinus leucas) ........................................................................ 28 3.1.3 Catfish (order Siluriformes) ................................................................................. 32 3.1.4 Grunters (family Terapontidae) ........................................................................... 38 3.1.5 Mullet (family Mugilidae) .................................................................................... 44 3.1.6 Sawfishes (Pristis spp.) ........................................................................................ 50 3.1.7 Threadfin (Polydactylus macrochir) ..................................................................... 57 3.2 Waterbirds ........................................................................................................................ 63 3.2.1 Colonial and semi-colonial nesting wading waterbirds....................................... 66 3.2.2 Cryptic wading waterbirds ................................................................................... 75 3.2.3 Shorebirds ............................................................................................................ 82 3.2.4 Swimming, grazing and diving waterbirds .......................................................... 94 3.3 Turtles, prawns and other species .................................................................................. 105 3.3.1 Banana prawns (Penaeus merguiensis) ............................................................. 105 3.3.2 Endeavour prawns (Metapenaeus spp.) ........................................................... 111 3.3.3 Freshwater turtles (family Chelidae) ................................................................. 117 3.3.4 Mud crabs (Scylla serrata) ................................................................................. 124 3.3.5 Tiger prawns (Penaeus esculentus and P. semisulcatus) ................................... 131 3.4 Flow-dependent habitats ............................................................................................... 137 3.4.1 Floodplain wetlands .......................................................................................... 137 3.4.2 Groundwater-dependent ecosystems............................................................... 145 3.4.3 Inchannel waterholes ........................................................................................ 161 3.4.4 Mangroves ......................................................................................................... 166 3.4.5 Saltpans and salt flats ........................................................................................ 172 3.4.6 Seagrass habitats ............................................................................................... 177 3.4.7 Surface-water-dependent vegetation ............................................................... 181 References 190 Part II Appendices 235 Figures Preface Figure 1-1 Map of Australia showing Assessment area (Southern Gulf catchments) and other recent CSIRO Assessments .................................................................................................... vi Preface Figure 1-2 Schematic of the high-level linkages between the eight activity groups and the general flow of information in the Assessment ...................................................................... vii Figure 2-1 Conceptual diagram of selected ecology assets and threatening processes of the Southern Gulf catchments. Ecology assets include species of significance, species groups and important habitats ........................................................................................................................ 12 Figure 2-2 Location of protected areas and important wetlands within the Southern Gulf catchments Assessment area, including management areas protected mainly for conservation through management intervention as defined by the IUCN ........................................................ 15 Figure 2-3 Distribution of threatened fauna species listed under the EPBC Act (Cth) and by the Northern Territory and Queensland governments in the Southern Gulf catchments ................. 18 Figure 3-1 Observed locations of barramundi (Lates calcarifer) and their modelled probability of occurrence in the Southern Gulf catchments ............................................................................... 24 Figure 3-2 Conceptual model showing the relationship between threats, drivers, effects and outcomes for barramundi in northern Australia .......................................................................... 28 Figure 3-3 Records of bull shark capture in the Southern Gulf catchments ................................ 30 Figure 3-4 Conceptual model showing the relationship between threats, drivers, effects and outcomes for bull sharks in northern Australia ............................................................................ 32 Figure 3-5 Location of catfish in the Southern Gulf catchments .................................................. 34 Figure 3-6 Modelled potential species distribution for fork-tailed catfish (Neoarius graeffei) in the Southern Gulf catchments ...................................................................................................... 35 Figure 3-7 Conceptual model showing the relationship between threats, drivers, effects and outcomes for catfish in northern Australia ................................................................................... 38 Figure 3-8 Location of grunters in the Southern Gulf catchments and marine region ................ 40 Figure 3-9 Modelled potential species distribution for sooty grunter (Hephaestus fuliginosus) in the Southern Gulf catchments ...................................................................................................... 41 Figure 3-10 Conceptual model showing the relationship between threats, drivers, effects and outcomes for grunters in northern Australia ................................................................................ 44 Figure 3-11 Records of capture of mullet in the Southern Gulf catchments and marine region . 47 Figure 3-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mullet in northern Australia ................................................................................... 50 Figure 3-13 Records of sawfish capture in the Southern Gulf catchments and the marine region ....................................................................................................................................................... 53 Figure 3-14 Modelled potential species distribution for freshwater sawfish (Pristis pristis) in the Southern Gulf catchments ............................................................................................................ 54 Figure 3-15 Conceptual model showing the relationship between threats, drivers, effects and outcomes for freshwater sawfish (Pristis pristis) in large rivers in northern Australia ................ 57 Figure 3-16 Records of threadfin capture in the Southern Gulf catchments marine region ....... 60 Figure 3-17 Conceptual model showing the relationship between threats, drivers, effects and outcomes for threadfin in northern Australia .............................................................................. 63 Figure 3-18 Royal spoonbill (Platalea regia) at the nest .............................................................. 67 Figure 3-19 Location of colonial and semi-colonial nesting wading waterbirds in the Southern Gulf catchments in alphabetic order of species name: Antigone rubicunda (brolga) to Egretta novaehollandiae (white-faced heron) .......................................................................................... 68 Figure 3-20 Location of colonial and semi-colonial nesting wading waterbirds in the Southern Gulf catchments in alphabetic order of species name: Egretta picata (pied heron) to Recurvirostra novaehollandiae (red-necked avocet) .................................................................... 69 Figure 3-21 Location of colonial and semi-colonial nesting wading waterbirds in the Southern Gulf catchments in alphabetic order of species name: Threskiornis molucca (Australian white ibis) to Threskiornis spinicollis (straw-necked ibis) ....................................................................... 70 Figure 3-22 Modelled potential species distribution for royal spoonbill (Platalea regia) in the Southern Gulf catchments ............................................................................................................ 71 Figure 3-23 Egret hunting among water lilies ............................................................................... 74 Figure 3-24 Conceptual model showing the potential relationship between threats, drivers, effects and outcomes for colonial and semi-colonial nesting wading waterbird species ............ 75 Figure 3-25 Dense aquatic and semi-aquatic vegetation used as habitat by cryptic wading waterbirds ..................................................................................................................................... 76 Figure 3-26 Location of selected cryptic wading waterbirds in the Southern Gulf catchments .. 78 Figure 3-27 Modelled potential species distribution for Australian painted snipe (Rostratula australis) in the Southern Gulf catchments .................................................................................. 79 Figure 3-28 Conceptual model showing the relationship between threats, drivers, effects and outcomes for cryptic wading waterbirds in northern Australia ................................................... 82 Figure 3-29 Red-capped plover walking along a shore ................................................................. 83 Figure 3-30 Observed locations of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Actitis hypoleucos (common sandpiper) to Calidris ruficollis (red-necked stint) .............................................................................................................................................. 86 Figure 3-31 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Calidris subminuta (long-toed stint) to Esacus magnirostris (beach stone-curlew) ....................................................................................................................................................... 87 Figure 3-32 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Gallinago megala (Swinhoe’s snipe) to Peltohyas australis (inland dotterel) ..... 88 Figure 3-33 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Phalaropus lobatus (red-necked phalarope) to Vanellus miles (masked lapwing) ....................................................................................................................................................... 89 Figure 3-34 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Vanellus tricolor (banded lapwing) to Xenus cinereus (Terek sandpiper) ............ 90 Figure 3-35 Modelled potential species distribution for eastern curlew (Numenius madagascariensis) in the Southern Gulf catchments ................................................................... 91 Figure 3-36 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the shorebirds group in northern Australia ........................................................... 94 Figure 3-37 Magpie goose perched on a fallen tree branch ......................................................... 95 Figure 3-38 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Amaurornis moluccana (pale-vented bush- hen) to Chenonetta jubata (Australian wood duck) ..................................................................... 98 Figure 3-39 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Cygnus atratus (black swan) to Nettapus coromandelianus (cotton pygmy-goose) ...................................................................................... 99 Figure 3-40 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Nettapus pulchellus (green pygmy-goose) to Porphyrio porphyrio (purple swamphen) .................................................................................... 100 Figure 3-41 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Spatula rhynchotis (Australian shoveler) to Tribonyx ventralis bBlack-tailed native-hen) .............................................................................. 101 Figure 3-42 Modelled potential species distribution for magpie goose (Anseranas semipalmata) in the Southern Gulf catchments ................................................................................................ 102 Figure 3-43 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the swimming, grazing and diving waterbirds group in northern Australia ........ 104 Figure 3-44 Fisheries catch of banana prawns the Southern Gulf catchments marine region .. 107 Figure 3-45 Conceptual model showing the relationship between threats, drivers, effects and outcomes for banana prawns in northern Australia .................................................................. 110 Figure 3-46 Fisheries catch of red endeavour prawns in the Southern Gulf catchments marine region .......................................................................................................................................... 113 Figure 3-47 Fisheries catch of blue endeavour prawns in the Southern Gulf catchments marine region .......................................................................................................................................... 114 Figure 3-48 Conceptual model showing the relationship between threats, drivers, effects and outcomes for endeavour prawns in northern Australia ............................................................. 117 Figure 3-49 Location of freshwater turtles within the Southern Gulf catchments .................... 119 Figure 3-50 Modelled potential species distribution for northern snake-necked turtle (Chelodina oblonga) in the Southern Gulf catchments................................................................................. 120 Figure 3-51 Conceptual model showing the relationship between threats, drivers, effects and outcomes for freshwater turtles in northern Australia .............................................................. 124 Figure 3-52 Mangrove and intertidal habitat typical of mud crab habitat in northern Australia ..................................................................................................................................................... 126 Figure 3-53 Location of mud crab habitat in the Southern Gulf catchments marine region ..... 127 Figure 3-54 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mud crabs in northern Australia .......................................................................... 131 Figure 3-55 Fisheries catch of brown tiger prawns in the Southern Gulf catchments marine region .......................................................................................................................................... 133 Figure 3-56 Fisheries catch of grooved tiger prawns in the Southern Gulf catchments marine region .......................................................................................................................................... 134 Figure 3-57 Conceptual model showing the relationship between threats, drivers, effects and outcomes for tiger prawns in northern Australia ....................................................................... 137 Figure 3-58 Brolgas flying into the sunset at Lake Moondarra ................................................... 141 Figure 3-59 Land subject to inundation (potential floodplain wetlands) and important wetlands in the Southern Gulf catchments ................................................................................................ 142 Figure 3-60 Conceptual model showing the relationship between threats, drivers, effects and outcomes for floodplain wetlands in northern Australia ........................................................... 145 Figure 3-61 Conceptualisation of obligate and facultative groundwater-dependent vegetation ..................................................................................................................................................... 147 Figure 3-62 Conceptualisation of terrestrial GDEs: (I) vigorous ecosystems with seasonally high water availability, (II) ecosystem condition with seasonally low water availability, and (III) seasonal low after groundwater development .......................................................................... 148 Figure 3-63 Distribution of potential groundwater-dependent aquatic ecosystems in the Southern Gulf catchments .......................................................................................................... 150 Figure 3-64 Distribution of potential groundwater-dependent terrestrial ecosystems in the Southern Gulf catchments .......................................................................................................... 152 Figure 3-65 Locations of observed obligate terrestrial GDEs in the Southern Gulf catchments 153 Figure 3-66 Locations of facultative and potential GDE vegetation species in the Southern Gulf catchments grouped by relevant vegetation type ..................................................................... 154 Figure 3-67 Distribution of potential groundwater-dependent vegetation in Southern Gulf catchments .................................................................................................................................. 155 Figure 3-68 Distribution of known and potential subterranean GDEs, alluvial and karstic aquifers and caves that may provide habitat for subterranean GDEs in the Southern Gulf catchments 157 Figure 3-69 Conceptual model showing the relationship between threats, drivers, effects and outcomes for aquatic GDEs in northern Australia ...................................................................... 159 Figure 3-70 Conceptual model showing the relationship between threats, drivers, effects and outcomes for terrestrial GDEs in northern Australia .................................................................. 160 Figure 3-71 Conceptual model showing the relationship between threats, drivers, effects and outcomes for subterranean GDEs in northern Australia ............................................................ 161 Figure 3-72 Location of persistent inchannel waterholes in the Southern Gulf catchments ..... 163 Figure 3-73 Conceptual model showing the relationship between threats, drivers, effects and outcomes for inchannel waterholes in northern Australia ........................................................ 166 Figure 3-74 Location of mangroves in the Southern Gulf catchments marine region ............... 168 Figure 3-75 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mangroves in northern Australia ......................................................................... 171 Figure 3-76 Saltpan area in northern Australia, which are generally located between mangrove and saltmarsh areas .................................................................................................................... 173 Figure 3-77 Location of saltpans in the Southern Gulf catchments marine region .................... 174 Figure 3-78 Conceptual model showing the relationship between threats, drivers, effects and outcomes for saltpans in northern Australia .............................................................................. 176 Figure 3-79 Distribution of seagrass habitats in the Southern Gulf catchments marine region 178 Figure 3-80 Conceptual model showing the relationship between threats, drivers, effects and outcomes for seagrass in northern Australia .............................................................................. 181 Figure 3-81 Locations of observed selected surface-water-dependent vegetation types in the Southern Gulf catchments .......................................................................................................... 186 Figure 3-82 Conceptual model showing the relationship between threats, drivers, effects and outcomes for surface-water-dependent vegetation in northern Australia ............................... 189 Tables Table 1-1 Freshwater, marine and terrestrial ecology assets with freshwater dependences ....... 7 Table 3-1 Ecological functions supporting barramundi and their associated flow requirements 26 Table 3-2 Ecological functions supporting river sharks and their supporting flow requirements 31 Table 3-3 Ecological functions supporting catfish and their associated flow requirements ........ 36 Table 3-4 Ecological functions supporting grunters and their associated flow requirements ..... 43 Table 3-5 Ecological functions supporting mullet and their associated flow requirements ........ 49 Table 3-6 Ecological functions supporting sawfishes and their associated flow requirements ... 55 Table 3-7 Ecological functions supporting threadfin and their associated flow requirements ... 61 Table 3-8 Waterbird species groups and example representative species for northern Australia ....................................................................................................................................................... 65 Table 3-9 Species in the colonial and semi-colonial nesting wading waterbird group, and their international conservation status ................................................................................................. 72 Table 3-10 Ecological functions supporting colonial and semi-colonial nesting waders and their associated flow requirements....................................................................................................... 73 Table 3-11 Species in the cryptic wading waterbird group, and their national and international conservation status ....................................................................................................................... 77 Table 3-12 Ecological functions supporting cryptic wading waterbirds and their associated flow requirements ................................................................................................................................. 80 Table 3-13 Species in the shorebirds group, and their national and international conservation status ............................................................................................................................................. 84 Table 3-14 Ecological functions supporting shorebirds and their associated flow requirements 92 Table 3-15 Species in the swimming, grazing and diving waterbirds group, and their national and international conservation status .......................................................................................... 96 Table 3-16 Ecological functions supporting swimming, grazing and diving waterbirds and their associated flow requirements..................................................................................................... 103 Table 3-17 Ecological functions supporting banana prawns and their associated flow requirements ............................................................................................................................... 109 Table 3-18 Ecological functions supporting endeavour prawns and their associated flow requirements ............................................................................................................................... 116 Table 3-19 Ecological functions for freshwater turtles and their supporting flow requirements ..................................................................................................................................................... 122 Table 3-20 Ecological functions supporting mud crabs and their associated flow requirements ..................................................................................................................................................... 129 Table 3-21 Ecological functions supporting tiger prawns and their associated flow requirements ..................................................................................................................................................... 136 Table 3-22 Nationally important wetlands in the Southern Gulf catchments ........................... 141 Table 3-23 Ecological functions supporting floodplain wetlands and their associated flow requirements ............................................................................................................................... 143 Table 3-24 Ecological functions supporting GDEs and their associated flow requirements ...... 158 Table 3-25 Ecological functions supporting inchannel waterholes and their associated flow requirements ............................................................................................................................... 164 Table 3-26 Ecological functions supporting mangroves and their associated flow requirements ..................................................................................................................................................... 169 Table 3-27 Ecological functions supporting saltpans and their associated flow requirements . 175 Table 3-28 Ecological functions for seagrass habitats and their supporting flow requirements 179 Table 3-29 Ecological functions supporting surface-water-dependent vegetation and their associated flow requirements..................................................................................................... 187 Part I Ecology of the Southern Gulf catchments Development of water resources can lead to a range of impacts to the environment, including changes in flow regimes, land use impacts and changes to connectivity by building instream structures. The rivers, floodplains and coastal regions of northern Australia are highly diverse and have significant conservation, cultural and economic values. To understand the potential risks to the natural environment associated with water resource development, this investigation (the ecology activity) takes an ecology asset approach. This involves prioritising assets; developing asset knowledge bases, conceptual relationships and evidence narratives, including flow–ecology relationships; and considering the context and distribution of the assets in the catchments of the Southern Gulf. Analysis to understand impacts draws upon the knowledge base of these assets and aspects of ecosystem function to model outcomes of water resource development and climate change scenarios. This technical report presents this knowledge base for the prioritised freshwater-dependent ecology assets across freshwater, marine and terrestrial habitats in the Southern Gulf catchments. 1 Introduction 1.1 Ecology and water resource development Development of water resources to support agriculture or aquaculture including water harvest from river flows, instream engineered dams or groundwater development can lead to a range of impacts on the environment, including changes in flow regimes, land use related habitat loss and changes to connectivity caused by building instream structures. The flow regimes of rivers are a primary driver of riverine, wetland, floodplain and near-shore coastal ecology (Bunn and Arthington, 2002; Junk et al., 1989; Poff and Zimmerman, 2010a). Water resource development alters flow regimes and leads to potentially significant changes in important flow attributes, such as the magnitude, timing, duration and rate of change of flow events upon which flora and fauna of the ecosystem are adapted. Also, changes in the frequency and duration of wet or dry spells and any modification of water quality (including changes to temperature regimes or sediment discharges) can affect species and their habitat. Additional impacts can result from instream barriers, land use change and a range of other threatening processes, either directly, indirectly, or in synergy with development. The result is potential ecological changes and consequences for the biota, habitats and ecosystem processes of a catchment (Poff et al., 1997). Freshwater systems in northern Australia contain a high level of biodiversity, with many unique and significant species and habitats. The catchments of northern Australia support at least 170 fish species, 150 waterbird species, 30 aquatic and semi-aquatic reptile species, 60 amphibian species and 100 macroinvertebrate families (van Dam et al., 2008). The freshwater and estuary habitats are critical in supporting productive fisheries, which increase production as freshwater inflow to estuaries increases (Aquatic Ecosystems Task Group, 2012). Catchment flows also support high- value commercial and recreational marine fisheries, such as the Northern Prawn Fishery, as well as barramundi (Lates calcarifer) and mud crab (Scylla serrata). Species and habitats of conservation significance also depend on the discharge of water and nutrients provided from catchment runoff. These include migratory waterbirds, sea turtles and a variety of sharks and rays, and habitats such as mangrove forests and seagrass beds. This activity (the ecology activity) seeks to determine the relative risks of alternative water resource development scenarios in the Southern Gulf catchments. The key questions that this activity seeks to address include: •What is the main environmental context of the catchment that could influence water resourcedevelopment? •What are the key environmental drivers and stressors that are currently occurring or likely tooccur in the catchment (including key supporting and threatening processes such as invasivespecies, water quality and habitat changes)? •What are the known linkages between flow and ecology? •What are the key ecological trade-offs between different water resource developmentsconsidering impacts from potential changes in flow on species and habitats? To understand the potential risks to the natural environment associated with water resource development, the ecology activity combines an ecology asset approach with other approaches that explore system processes such as lateral and longitudinal connectivity, inundation, habitat provision and end-of-system discharge. The ecology work builds upon and adapts the methods used in the ecology synthesis and assessment components of the Northern Australia Water Resource Assessment (NAWRA; Pollino et al. (2018a) and Pollino et al. (2018b)). This involves undertaking a review and prioritising assets; developing asset knowledge bases, conceptual relationships and evidence narratives, including flow–ecology relationships; and considering the context and distribution of the assets in the catchments of the Southern Gulf. The analysis approach undertaken in the ecology activity uses a combination of modelling methods with the choice of method(s) used for each asset depending upon the ability to support modelling given the strength of the knowledge base, the asset data and the types of relationships important for each asset (methods and results provided in the Asset Analysis reporting (Ponce Reyes et al., 2024)). The ecology assets include species, groups of species and habitats and their processes that are freshwater dependent and significant within the Southern Gulf catchments and for which there is sufficient understanding of their requirements. The use of modelling methods for specific assets depends on the relationships between flow and ecological outcomes and if these relationships are sufficiently known and can be suitably supported by the knowledge base of the asset. Quantitative ecology modelling uses hydrology scenarios developed by the surface water hydrology activity as primary inputs. The ecology modelling compares outcomes as relative differences between the scenarios and a baseline to identify where change occurs and by how much. This analysis enables the identification of assets that may be most sensitive to the type of changes in the different scenarios, and the scenarios that lead to the greatest ecological change. A summary of the ecology approach is provided in 1.3. 1.2 Water resource development and ecological changes The importance of the natural flow regime for supporting environmental function has become increasingly well understood, as has the importance of rivers operating as systems, including the connection of floodplains via inundation, the distribution of refuges, and discharges into coastal regions. Globally, water resource development has a range of known impacts on ecological systems. The influence of each of these impacts depends upon a range of factors, including catchment properties (e.g. physical, geographic and climate characteristics), the type of developments (e.g. dams, water harvesting, groundwater development), the source location or distribution of the developments within the catchment, the magnitude and pattern of change, how any changes may be managed or mitigated, and the habitats and species that will be affected and their distribution. The extent to which impacts may occur are also highly uncertain. Impacts associated with water resource development include the following, which are described below: •flow regime change (Section 1.2.1) •altered longitudinal and lateral connectivity (Section 1.2.2) •habitat modification and loss (Section 1.2.3) •increased invasive and non-native species (Section 1.2.4) •synergistic and co-occurring processes both local and global (Section 1.2.5). 1.2.1 Flow regime change Water resource development, including water harvesting and creating instream structures for water retention, can influence the timing, quality and quantity of water that is provided by catchment runoff into the river system. The natural flow regime (including the magnitude, duration, timing, frequency and pattern of flow events) is important in supporting a broad range of environmental processes upon which species and habitat condition depend (Lear et al., 2019; Poff et al., 1997). Flow conditions provide the physical habitat in streams and rivers which determines biotic use and composition and to which life-history strategies are adapted, and enables movement and migration between habitats and the exchange of nutrients and materials (Bunn and Arthington, 2002; Jardine et al., 2015). In a river system, the natural periods of both low and high flow (including no-flow events) are important to support the natural function of habitats, their ecological processes and the shaping of biotic communities (King et al., 2015). Through the attenuation of flows, water resource development can lead to impacts significant distances downstream of the development, including into coastal and near-shore marine habitats (Broadley et al., 2020; Pollino et al., 2018a). 1.2.2 Altered longitudinal and lateral connectivity River flow facilitates the exchange of biota, materials, nutrients and carbon along the river and into the coastal areas (longitudinal connectivity), as well as between the river and the floodplain (lateral connectivity) (Pettit et al., 2017; Warfe et al., 2011). Physical barriers such as weirs and dams, or a reduction in the magnitude of flows (and the duration or frequency) can affect longitudinal and lateral connectivity, changing the rate or timing of exchanges (Crook et al., 2015). These impacts can include changes in species’ migration and movement patterns as well as altered erosion processes and discharges of nutrients into rivers and coastal waters (Brodie and Mitchell, 2005). Seasonal patterns and rates of connection and disconnection caused by flood pulses are important for providing seasonal habitat, enabling movement of biota into new habitats and their return to refuge habitats during drier conditions (Crook et al., 2020). 1.2.3 Habitat modification and loss Water resource development can cause direct loss of habitat. For example, artificially creating a lake inundates habitat behind an impoundment resulting in loss of terrestrial and stream habitat. Habitat can be lost due to agricultural development and infrastructure, including roads and canals, which can fragment terrestrial habitat or artificially connect aquatic habitats that had been historically distinct. 1.2.4 Increased invasive and non-native species Water resource development often homogenises flow or habitats, for example, through changed patterns of capture and release of flows or creation of impoundments for storage and regulation. It is recognised that invasive species are often at an advantage in such modified habitats (Bunn and Arthington, 2002). Modified landscapes, such as lakes or homogenised perennial streams that were previously ephemeral, can be a pathway for introduction and support the incidental, accidental or deliberate establishment of non-native species, including pest plants and fish (Bunn and Arthington, 2002; Close et al., 2012; Ebner et al., 2020). Increased human activity can increase the risk of invasive species being introduced. 1.2.5 Synergistic and co-occurring processes both local and global Along with water resource development comes a range of other pressures and threats, including increases in fishing, vehicles, habitat fragmentation, pesticides, fertilisers and other chemicals, erosion, degradation due to increased stock pressure, changed fire regimes, climate change and other human disturbances, both direct and indirect. Some of these pressures are the direct result of changes in land use associated with water resource development, others may occur locally, regionally or globally, and act synergistically with water resource development and agricultural development to increase the risk to species and their habitats (Craig et al., 2017; Pettit et al., 2012). 1.3 Ecology asset-based approach to modelling and assessment The goal of the ecology activity is to understand the potential impacts of water resource development on ecological systems. This is achieved by modelling a set of ecology assets, including species, habitats and catchment ecosystem functions. Ecology assets in northern Australia depend on freshwater flows to support their persistence or function. Assets are spread across freshwater, marine and terrestrial habitats (including terrestrial habitats dependent on groundwater or flood flow and inundation). The ecology assets have different distributions within the catchment, have different flow associations and needs, and are likely to have different trajectories of change when exposed to a potential range of threatening processes. The ecology activity uses these assets in a range of models to infer what impacts may occur, and where within the catchment, as a result of different water resource development and climate change scenarios. The ecology activity is built upon four main components of work, which are described below: • identifying and prioritising assets (Section 1.3.1) • reviewing, conceptual modelling and developing evidence narratives (Section 1.3.2) • mapping the location and distribution of the assets (Section 1.3.3) • understanding flow–ecology relationships and quantitative modelling (Section 1.3.4). 1.3.1 Identifying and prioritising assets For the purpose of the ecology activity, assets can be considered either partially or fully freshwater dependent, including terrestrial or marine assets dependent upon freshwater flows (or services provided by freshwater flows). Assets can include: • species − individual species (such as barramundi in Section 3.1.1) • taxonomic groups − groups of species that are closely related (such as grunters in Section 3.1.4) • functional groups − groups of often unrelated species that may occupy similar niches, use similar habitat, have other attributes or requirements, and that are likely to respond to change in a similar way (such as colonial and semi-colonial nesting wading waterbirds in Section 3.2.1) •habitats−importanthabitatsincludegeographical areasidentifiedassharingsimilar characteristics (such as position on the floodplainor channel, water retention or sheddingproperties)or other structural features that maymake it importantfor the catchment ecology and support biota within and around the catchment. Habitats include floodplain wetlands(Section3.4.1) and groundwater-dependent ecosystems (GDEs; Section 3.4.2). Habitats areimportant for supporting speciesor communitiesand may include, but are not limited to, identified or listed locations such asnational parks orDirectory of Important Wetlandsin Australia (DIWA) sites. To identify assets, species, species groups and habitats have been reviewed and prioritised fortheSouthern Gulfcatchments. Freshwater-dependent assetswere considered if theymeetany of the following: •a speciesor communitythat is listed asThreatened,Vulnerableor Endangered(EPBC orState/Territory listing) •a habitat, species or communitythat isformally recognisedin governmentconservationagreements •habitat that providesvital, near-natural,rareor uniquehabitat for water-dependent floraand fauna •supportingimportantor notablebiodiversityofwater-dependent flora and fauna •providing recreational,commercialor cultural value. From the full range ofpotential assetsidentifiedin theSouthern Gulfcatchments,the process forselecting priorityassetsfor considerationisifthey are: •distinctive–havingan association between flow and outcomes of change,and a broad range of water requirementsare representedacross different assets •representative–can provideinsights fortheflow requirementsof other biota and ecological processesthat are not explicitly modelled •describable–having sufficient available peer-reviewed evidence to identify and describe relationships with flow •significant–havingecological, conservation, cultural or recreational importance,and relevancetothe Assessment catchment. The prioritisedecologyassets described in this report are listed inTable1-1. These assets are usedin theecology analysis toassesstherelativerisksassociated withdifferent water resourcedevelopment scenarios consideringthe type, magnitude and location of change. 6| Ecological asset descriptionsof the Southern Gulf catchments Table 1-1 Freshwater, marine and terrestrial ecology assets with freshwater dependences For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 1.3.2 Reviewing, conceptual modelling and developing evidence narratives Literature and data were reviewed for each asset. Information collated includes catchment- specific data and knowledge, important water and flow requirements, habitat use and distribution, and information required to support an understanding of the asset’s response to potential environmental change, considering aspects such as life history, flow triggers, movement, refuge, foraging and productivity. A conceptual model for each asset was developed to represent and summarise this ecological understanding and an evidence narrative used to support and communicate the flow–ecology relationships and pathways to change for each asset based upon available data and literature. The asset conceptual models provide a simplified visual summary and outline an evidence-based hypothesis for the potential impacts under water resource development and climate change scenarios. The conceptual models are box and arrow models with a standardised structure that represent links between the threats, drivers, effects and outcomes. These key terms are defined as follows: • Threat is an action or activity that has the capacity to adversely affect an ecology asset and its value (Hart et al., 2005). • Driver (ecological driver) is the mechanism, process or change by which a threat affects an asset. • Effect is the direct change in, or response of, the asset that has occurred as a consequence of the driver. • Outcome is the overall observable or measurable impact on the asset or its function within the catchment (tangible or otherwise). The conceptual models explore relationships between key potential threats (including water resource development, land use and climate change) and the effects and outcomes of these threating processes from the perspective of each asset, including loss of biodiversity or habitat quality. The standardised structure of these models allows for comparison across assets and understanding commonalities and differences between assets. Each model is built on a scientific evidence base incorporating a broad range of literature and research, which is documented in the ‘Pathways to change’ section for each asset. 1.3.3 Mapping the location and distribution of the assets The location or distribution of assets has been mapped across the Southern Gulf catchments to understand the occurrences and/or important locations of the assets across the catchment and the near-shore coastal region. For species with suitable data, species distribution models have been created to extrapolate potential suitable habitat within the catchment based upon environmental relationships established across northern Australia, thereby drawing upon a larger number of data points. For some of the habitat assets, remote sensing is used to identify distribution across the catchment. Water resource development scenarios consider a range of development options and pathways within each catchment, with the flow downstream of each development being affected. Assets located downstream of these scenario water resource development sites will be exposed to changes in flow with different scenarios resulting in different changes. The level of flow change and impact will depend on the asset’s location relative to the development. A range of data sources, including Atlas of Living Australia (https://www.ala.org.au/ ), government department databases, and fisheries catch records, were used to develop maps and spatial relationships of the assets in the Southern Gulf catchments. Species distribution modelling uses the observed locations of species across northern Australia to create correlations with environmental covariates that are then used to predict suitable habitat associations within the Southern Gulf catchments. 1.3.4 Understanding flow–ecology relationships and modelling The analysis approach undertaken in the ecology activity uses a combination of modelling methods with the choice of method(s) used for each asset depending upon the ability to support modelling given the strength of the knowledge base, the asset data and the types of relationships important for each asset. Modelling requires understanding relationships between flow, ecological responses and the potential outcomes of changes. Different models are used to represent and incorporate varying processes (including changes in flow, inundation and connectivity) that range in importance for different assets. The primary inputs to analysis are daily hydrology data generated with river system models (flow discharge and quantitative properties of the hydrograph) or hydrodynamic models (depth, velocity, inundation) that can be used to quantify the relative differences between scenarios and a baseline over the same modelled period (River system modelling (Gibbs et al., 2024)). A summary of some of the modelling used in the ecology assessment is provided below, and further details and the result of modelling are provided in the companion report (Asset Analysis reporting (Ponce Reyes et al., 2024)). Flow requirements A common base to the analysis is the flow requirements method which is used for all assets. This analysis identifies the specific components of the hydrograph that are important to each asset (quantified using hydrometrics: statistical properties of the long-term flow regime). The specific set of flow metrics identified for each of the assets are important for ecological function such as supporting life-history requirements, movement, or provision of important habitat (for example, high flows may be required to support migration upstream or onto the floodplain for fish species thereby enhancing habitat quality or availability resulting to improved condition and higher abundances). The flow requirements assessment calculates the change in these asset-specific hydrometrics occurring between the model scenarios and the baseline to enable relative comparison between locations, assets and scenarios. Depending upon the scenario, and whether the development is point source or distributed, changes in flow could accumulate or diminish as flows attenuate through the catchment. Results indicate sensitivities to the types of change in the hydrograph due to assets having different flow associations and needs, occupying different locations within the catchment, and because water resource development scenarios manifest different changes in hydrology. Hydrodynamic habitat suitability Hydrodynamic habitat suitability modelling uses depth, velocity and inundation extent outputs from hydrodynamic models to map and quantify the occurrence of flow habitat that would be suitable or preferred for species across different flood events. The method uses species’ (or species groups’) specific habitat preferences informed by literature and/or data to provide mechanistic links between hydraulic variables from hydrodynamics modelling. The form of these relationships can be used for a range of biota, such as fish and waterbirds, for which depth and velocity are important determinants of habitat preferences or associations. For example, habitat preference can be informed by the results of tracking studies that position species’ use of inchannel and floodplain habitat and relate this to the experienced hydraulic properties of the location. Changes in the hydrodynamics between flood events and between scenarios may reduce or enhance the availability of suitable or preferred habitat within the catchment with an effect on the availability of resources for the assets. Connectivity assessment The connectivity assessment uses hydrodynamic modelling to develop a daily time series of inundation extents or depths for a range of scenarios. For these scenarios, across a sample of flood events, the pattern and extent of inundation can quantify the connectivity of assets (e.g. floodplain wetlands) to the main river channel via connection across the floodplain or via flood runners (latitudinal connectivity) or across instream barriers that may limit movement along the river channel during periods of low flow (longitudinal connectivity). Differences in the extent and and/or duration of connections to these assets is quantified between the scenarios. Models of Intermediate Complexity for Ecosystem assessment Models of Intermediate Complexity for Ecosystem assessment (MICE) are methods for simultaneously assessing the status of both fisheries and other non-targeted species, including those of high conservation concern, and evaluating the trade-offs among management plans aimed at addressing conflicting objectives. They are dynamic, spatially resolved models of intermediate complexity that draw on quantitative and statistical methods of stock assessment approaches and extend this to a representation of stressors and outcomes in an ecosystem. 2 Ecology of the Southern Gulf catchments 2.1 Ecology of the Southern Gulf catchments 2.1.1 Southern Gulf catchments and its environmental values The Southern Gulf catchments span an area of 108,200 km2 across the Northern Territory (NT) and Queensland and are comprised of Settlement (17,600 km2), Nicholson (52,200 km2), Leichhardt (33,400 km2) and Mornington Inlet (3,700 km2) catchments, as well as the Mornington Islands (1,200 km2). Agricultural production is the largest land use in the catchments, mostly cattle grazing on native pastures (85%) (ABARES, 2022). Other land uses include recreational activities, tourism, traditional and commercial fisheries, mining and traditional Indigenous uses. In addition, these catchments have important ecological and environmental values. Within this catchment and the surrounding marine environment are rich and important ecology assets, including species, ecological communities, habitats, and ecological processes and functions (conceptualised summary in Figure 2-1). The ecology of the Southern Gulf catchments is maintained by the flow regime in each catchment, shaped by the region’s wet-dry climate and complex geomorphology and topography, and driven by seasonal rainfall, evapotranspiration and groundwater discharge. Figure 2-1 Conceptual diagram of selected ecology assets and threatening processes of the Southern Gulf catchments. Ecology assets include species of significance, species groups and important habitats See Table 1-1 for a complete list of the freshwater, marine and terrestrial ecology assets considered in the Southern Gulf catchments. Biota icons for the Southern Gulf catchments adapted from the Integration and Applicaton Network (2023). The Southern Gulf catchments have a highly seasonal climate with an extended dry season. Rainfall averages 602 mm/year with 94% of the rainfall falling in the wet season (McJannet et al., 2023). The dominant vegetation types in the catchments are open eucalypt open woodlands, Melaleuca forests and woodlands and tussock grasslands (Department of Climate Change‚ Energy‚ the Environment and Water, 2020). There are two major water storages within the Southern Gulf catchments: Lake Julius and Lake Moondarra. Both storages are located on the Leichhardt River and are listed in the Directory of Important Wetlands in Australia (DIWA). While the water in these two storages predominantly supplies urban, mining and industrial demand around Mount Isa and Cloncurry, the permanent water provides important dry-season refuge for waterbirds and supports a variety of freshwater fish species (Department of Agriculture‚ Water and the Environment, 2021a). During the wet season, flooding inundates significant parts of the lower reaches of the catchments connecting wetlands to the river channel, inundating floodplains and driving a productivity boom. This flooding is particularly evident in the lower parts of the catchment, including the floodplain wetlands, and extensive intertidal flats on the mainland coastline south of Bentinck and Sweers islands, where it delivers extensive discharges into the marine waters of the south-western Gulf of Carpentaria. While most rivers in the Southern Gulf catchments are ephemeral, the Gregory and O'Shanassy rivers and Lawn Hill Creek are permanent, being fed by groundwater from the Thorntonia Limestone formation. These perennial waterways provide critical refuge habitat for many aquatic species in this semi-arid environment. During the dry season, river flows are reduced and the streams in the catchment contract, resulting in series of instream waterholes, which also provide critical habitat in the dry season. 2.1.2 Protected, listed and significant areas of the Southern Gulf catchments The protected areas located in the Southern Gulf catchments include the UNESCO World Heritage listed Australian Fossil Mammal Sites (Riversleigh), three Indigenous Protected Areas, namely Ganalanga-Mindibirrina, Nijinda Durlga and Thuwathu/Bujimulla, and Boodjamulla (Lawn Hill) and Finucane Island national parks and other conservation parks (Figure 2-2). The Australian Fossil Mammal Sites (Riversleigh) is largely situated within the Boodjamulla (Lawn Hill) National Park. It is considered one of the richest and most extensive fossil deposits in the world. The Ganalanga-Mindibirrina Indigenous Protected Area is located in the upper reach of the Nicholson River and covers over 1 million ha. The Nijinda Durlga Indigenous Protected Area covers over 186,850 ha and includes habitat for marine turtles, dugongs, shorebirds and seabirds. The Thuwathu/Bujimulla Indigenous Protected Area spans across the Wellesley Islands and includes over 1.6 million ha of marine area and over 120,000 ha of land. The area contains significant habitat for sea turtles, shorebirds and seabirds. It is also culturally significant, including having the largest collection of stone fish traps in the southern hemisphere (Australian Indigenous Australians Agency, 2023). Boodjamulla (Lawn Hill) National Park includes sandstone ranges, Lawn Hill Gorge and the Australian Fossil Mammal Sites (Riversleigh) UNESCO World Heritage site. Lawn Hill Gorge is formed by Lawn Hill Creek, which is groundwater fed, as are the Gregory and O'Shanassy rivers. Lawn Hill Creek and Gregory River contain wet riverine forest and support a variety of plant species. Lawn Hill Creek supports freshwater turtles, including the Gulf snapping turtle (Elseya lavarackorum; listed as Endangered under the Commonwealth Environment Protection and Biodiversity Conservation Act 1999 (EPBC Act)) and more than 20 species of fish. Across the catchment, permanent water sources provide important refuge habitat for a broad variety of species (Department of Environment and Science (Qld), 2023a). Finucane Island National Park is 7610 ha and is an island within a river network, located within the Southern Gulf Aggregation DIWA site. It includes estuarine wetlands, salt flats, mangroves and grasslands. The park provides important habitat for fish and waterbirds (Department of Environment and Science (Qld), 2023b). The Southern Gulf catchments contain 13 nationally significant wetlands listed in the DIWA: Bluebush Swamp, Buffalo Lake Aggregation, Forsyth Island Wetlands, Gregory River, Lake Julius, Lake Moondarra, Lawn Hill Gorge, Marless Lagoon Aggregation, Musselbrook Creek Aggregation, Nicholson Delta Aggregation, Southern Gulf Aggregation, Thorntonia Aggregation and Wentworth Aggregation (Figure 2-2) (Department of Agriculture‚ Water and the Environment, 2021a). These DIWA-listed wetlands include a variety of wetland types, ranging from estuarine wetlands with salt flats and saltmarshes to man-made lakes and spring-fed creeks and rivers. For more information on the DIWA-listed wetlands within the Southern Gulf catchments see Section 3.4.1. No wetlands listed under the Ramsar Convention on Wetlands of International Importance occur within the Southern Gulf catchments. Figure 2-2 Location of protected areas and important wetlands within the Southern Gulf catchments Assessment area, including management areas protected mainly for conservation through management intervention as defined by the IUCN Data sources: Department of Agriculture‚ Water and the Environment (2020a; 2020b) Department of the Environment and Energy (2010) 2.1.3 Important habitat types and values of the Southern Gulf catchments The freshwater sections of the Southern Gulf catchments include diverse habitats such as persistent and ephemeral rivers, wetlands, floodplains and groundwater-dependent ecosystems (GDEs). The diversity and complexity of habitats, and the connections between habitats within a catchment, are vital for providing the range of habitats needed to support both aquatic and terrestrial biota (Schofield et al., 2018). In the wet season, flooding connects rivers to floodplains. The exchange of water from the river across the floodplain supports higher levels of primary and secondary productivity than compared to surrounding areas with less frequent inundation (Pettit et al., 2011). Infiltration of water into the soil during the wet season and along persistent streams often enables riparian habitats to form an important interface between the aquatic and terrestrial environment. While riparian habitats often occupy a relatively small proportion of the catchment, they frequently have a higher species richness and abundance of individuals than surrounding habitats (Pettit et al., 2011; Xiang et al., 2016). In the dry season, biodiversity is supported by the perennial rivers and creeks, permanent lakes, and inchannel waterholes. These water sources that persist through the dry season become increasingly important as the season progresses, as they provide important refuge habitat for species and enable recolonisation into surrounding habitats upon the return of larger flows (Hermoso et al., 2013). These water sources provide habitat for water-dependent species including fish, sawfish and turtles, as well as providing a source of water for other species more broadly within the landscape (McJannet et al., 2014; Waltham et al., 2013). The terrestrial habitats of northern Australia include a range of diverse and significant habitat types. While much of the tropics of northern Australia is savanna, eucalypt forest and grasslands, other habitats include riparian and floodplain communities and GDEs. Many of these are highly dependent upon freshwater supplied from rivers, and their persistence and condition can also be supported by groundwater discharges. GDEs occur across many parts of the Southern Gulf catchments and come in different forms, including aquatic, terrestrial and subterranean habitats. Aquatic GDEs, contain springs and river sections that hold water throughout most dry seasons due to groundwater discharge. Aquatic GDEs are important for supporting aquatic life and fringing vegetation and in the wet-dry tropics that can provide critical refuge during periods of the late dry season (James et al., 2013). Vegetation occurring adjacent to the waterways in the Southern Gulf catchments relies on water from a range of sources (surface water, soil water, groundwater) which are seasonally dynamic and highly spatially variable. Water may be sourced from a combination of direct rainfall, bank recharge from instream flows, local floodplain recharge from surface water inundation during overbank flows, and/or shallow groundwater connected to intermediate or regional aquifer systems. Perennial floodplain vegetation often uses groundwater when it is within reach of the root network, particularly during the dry season or drought, but the origin of the groundwater used has only been infrequently investigated (e.g. Canham et al. (2021)). In some locations vegetation may be sustained by water available in unsaturated soils and so never use groundwater. In other locations vegetation may use groundwater sourced from local alluvial recharge processes; alternatively, regional groundwater may be critical for maintaining vegetation condition. Subterranean aquatic ecosystems in the limestone that occur in the upper south- western region of the Southern Gulf catchments support subterranean fauna that depend on the presence of groundwater (e.g. troglofauna, which live in caves, and stygofauna, which live in groundwater systems). The marine and estuarine environments of the Southern Gulf catchments, including the mainland area adjacent to Mornington and Sweers islands, have extensive intertidal flats and estuarine communities including mangroves, salt flats and seagrass habitats. These habitats are highly productive, have high cultural value, and are often of national significance (Poiner et al., 1987). Seagrass beds in the nearby coastal Gulf of Carpentaria have high diversity, has vigorous stands and provide an important food and habitat for dugongs (Dugong dugon), green turtles (Chelonia mydas) and prawns (Loneragan et al., 1997; Poiner et al., 1987). These near-coastal and estuary habitats support a major commercial barramundi fishery (Bayliss et al., 2014). Mud crabs (mainly Scylla serrata) are also harvested (Bayliss et al., 2014). 2.1.4 Significant species and ecological communities of the Southern Gulf catchments There are a number of aquatic and terrestrial species in the Southern Gulf catchments currently listed as Critically Endangered, Endangered and Vulnerable under the EPBC Act and by the NT Government’s wildlife classification system and the Queensland Government’s threatened species conservation system (Figure 2-3). These include freshwater or largetooth sawfish (Pristis pristis; Vulnerable, EPBC Act) and the Gulf snapping turtle (Elseya lavarackorum; Endangered, EPBC Act). The Southern Gulf catchments are important stopover habitat for migratory shorebird species listed under the EPBC Act, including the eastern curlew (Numenius madagascariensis; Critically Endangered) and the Australian painted snipe (Rostratula australis; Endangered) (Atlas of Living Australia, 2021; Department of Agriculture‚ Water and the Environment, 2021b). Figure 2-3 Distribution of threatened fauna species listed under the EPBC Act (Cth) and by the Northern Territory and Queensland governments in the Southern Gulf catchments Data source: Department of Environment and Science (Qld) (2023c); Department of Environment Parks and Water Security (2019a) 2.1.5 Current condition and potential threats in the Southern Gulf catchments Northern Australia more broadly encompasses some of the last relatively undisturbed tropical riverine landscapes in the world with low levels of flow regulation and low development intensity (Pettit et al., 2017; Vörösmarty et al., 2010). Fishing in northern Australia is a valuable industry, and the waters of the Gulf of Carpentaria contribute significantly to the national catch of important species including prawns, mud crab and barramundi. There is a range of economic enterprises, infrastructure and human impacts in the Southern Gulf catchments and the nature and extent to which these have modified the habitats and affected species of the Southern Gulf catchments varies. Intensive industries such as mining affect portions of the catchments, and there is potential for further mining and tourism-related impacts at sensitive and vulnerable sites. The study area includes the towns of Burketown, Doomadgee and Mount Isa which support tourism and mining and act as regional hubs for many of the stations across the catchment. While a proportion of the catchment is under conservation reserves, the study area does face environmental threats, including the potential for mining and tourism-related impacts at sensitive and vulnerable sites. One of the most significant environmental threats to remote regions across northern Australia is that of introduced plants and animals. In the Southern Gulf catchments, feral horses (Equus caballus), wild pig (Sus scrofa) and cane toads (Rhinella marina) are among the introduced animals (Department of Agriculture‚ Water and the Environment, 2021a; 2021b). Weeds of national significance in the aquatic systems of northern Australia include salvinia (Salvinia molesta) and rubber vine (Cryptostegia grandiflora) (Close et al., 2012). Weed species of interest in and around the Southern Gulf catchments include prickly acacia (Vachellia nilotica), buffel-grass (Cenchrus ciliaris), rubber vine (Cryptostegia grandiflora) and water hyacinth (Eichhornia crassipes) (Department of Agriculture‚ Water and the Environment, 2021b). 3 Ecological assets from the Southern Gulf catchments and marine region Northern Australia’s rivers, floodplains and coastal regions contain high biodiversity, including at least 170 fish species, 150 waterbird species, 30 aquatic and semi-aquatic reptile species, 60 amphibian species and 100 macroinvertebrate families (van Dam et al., 2008). The ecologies of the freshwater systems are supported by, and adapted to, the highly seasonal flow regimes of the wet-dry tropics. This section provides a synthesis of the prioritised assets relevant to the Southern Gulf catchments for the purpose of understanding the ecology outcomes of flow regime change. Table 1-1 presents the full list of assets used in the ecology activity. 3.1 Fish, sharks and rays 3.1.1 Barramundi (Lates calcarifer) Description and background to ecology Barramundi are a large (>1 m standard length) opportunistic-predatory fish (order Perciformes) that occurs throughout northern Australia. The species is non-obligatory catadromous (i.e. it migrates down rivers to spawn in the sea) and occurs in ‘catchment to coast’ habitats throughout the west Indo-Pacific region, including estuaries, rivers, lagoons and wetlands across northern Australia (Crook et al., 2016; Pender and Griffin, 1996; Roberts et al., 2023; Russell and Garrett, 1983; 1985). The fish is long-lived (living up to about 32 years) and fast-growing, and individuals begin life as a male but change to female as they age (protandrous hermaphrodite): they occupy freshwater habitats as males in the first years of life and saltwater habitats as older females. The species is of ecological importance, capable of modifying the estuarine and riverine fish and crustacean communities (Blaber et al., 1989; Brewer et al., 1995; Milton et al., 2005; Russell and Garrett, 1985). Barramundi is arguably the most important fish species for commercial, recreational and Indigenous subsistence fisheries throughout Australia’s wet-dry tropics. It makes up a substantial component of the total commercial fish catch in northern Australia (Savage and Hobsbawn, 2015). Queensland’s ‘southern Gulf of Carpentaria' barramundi stock encompasses fish from the Southern Gulf catchments as well as coastal rivers and estuaries further east in the gulf (Filar and Streipert, 2022; Streipert et al., 2019). Grubert et al. (2020) summarised commercial stocks across northern Australia. They suggested that commercial catch in the southern Gulf of Carpentaria was high; from 1989 to 2011, the commercial catch of barramundi increased from 520 t to 960 t, before decreasing through 2013–2015. The average harvest from 2013 to 2017 was 468 t (Streipert et al., 2019). Grubert et al. (2020) suggest the 2019 commercial catch of 496 t was lower than average catches from earlier decades, though the catch rate was high (catch per unit effort of 28 kg 100 m of net per day. The catch per unit effort was similar to recent barramundi catch rates in NT tropical coastal waters. In 2013–14, barramundi comprised 28% of the $31 million wild- caught fishery production in the NT. Commercial catch per unit effort in the NT increased from about 7 kg 100 m of net per day in the early 1980s to over 30 kg 100 m of net per day in the 2010s (NT Government, 2018). Barramundi is also a fish of cultural significance for the Indigenous community as well as being an important food source (Jackson et al., 2012). The movements of barramundi between habitats are indicators of the change in season for Indigenous communities across tropical Australia (Green et al., 2010). The movements relate to the barramundi’s habitat requirements during its life cycle, which rely on seasonal variation in river flows to access habitats. Barramundi life history renders the species critically dependent on river flows (Plagányi et al., 2023; Tanimoto et al., 2012). Large females (older fish) and smaller males (younger fish) reside in estuarine and littoral coastal habitats. Mating and spawning occur in the lower estuary during the late dry season to early wet season, and new recruits move into supra-littoral and freshwater habitats. Coastal salt flat, floodplain and palustrine (i.e. non-tidal wetland) habitats depend on overbank flows for maintenance and connectivity (Crook et al., 2016; Russell and Garrett, 1983; 1985). Young fish, as males, may move large distances upstream and reside in palustrine billabongs for 3 to 4 years before maturing and migrating downstream. This ontogenetic migration (i.e. migration between habitats at different life stages) requires palustrine–riverine and riverine– estuarine connectivity; hence migration depends on catchment flows. Barramundi transform to females at about 6 years old when they mix with younger males within river estuaries and breed. Juvenile barramundi growth rates are consistent over a range of ‘tropical’ temperatures commonly encountered in northern Australia, including 29°C and 35°C (Scheuffele et al., 2021). Over the past decade, studies using otolith microchemistry and fish-tag telemetry have provided greater understanding of barramundi use of freshwater, estuarine and marine habitats (Crook et al., 2016; Roberts et al., 2019) than initial life-history studies in the 1980s (Pender and Griffin, 1996). Crook et al. (2016) proposed three primary life-history strategies employed by barramundi: (i) some male adults return to the estuary to spawn after spending several years in freshwater habitats, (ii) some individuals delay downstream spawning migrations for 6 to 10 years until they have undergone the transition to females in freshwater habitats, and (iii) some barramundi remain in estuarine waters and complete their life cycle without entering freshwater systems (Crook et al., 2016; Roberts et al., 2019; Robins et al., 2021). The variation in migration strategy is thought to be triggered by variation in the strength of the flow regime (Crook et al., 2016; Roberts et al., 2023), making the species particularly vulnerable to water resource development (Robins et al., 2021). Moreover, the effects of different levels of river flow are now better understood. During high-flow years (a strong wet season), barramundi tend to remain within the estuary (Roberts et al., 2023); the estuary becomes a brackish habitat, and terrestrial and palustrine productive inputs probably render the estuary prime habitat (Burford and Faggotter, 2021). During low-flow years (a drier wet season), barramundi are twice as likely to immigrate to riverine and palustrine habitats, probably seeking better foraging conditions among freshwater habitats (~80% move to riverine habitats in dry years) (Crook et al., 2022; Roberts et al., 2023). In years with strong monsoon rainfall, about 60% of barramundi remain within the estuarine habitats that are strongly linked and receiving inputs from the catchment (Roberts et al., 2023). Overall, approximately 62% of barramundi in tropical Australian rivers were catadromous, that is, migrating to freshwater habitats and returning to saline waters to spawn (Roberts et al., 2019). In Gulf of Carpentaria rivers, quantitative modelling of the relationship between flow and catch for barramundi has confirmed the dependence of barramundi on flow regime, predicting declines in barramundi biomass directly related to water extraction or impoundment that modifies the seasonal historical flows (Plagányi et al., 2023). Barramundi in the Southern Gulf catchments Barramundi are abundant within freshwater, estuarine and marine reaches of the southern Gulf of Carpentaria rivers (Figure 3-1). In estuarine and coastal locations of these rivers, barramundi are the dominant target species for the inshore net fishery (Queensland N3 fishery, Department of Agriculture and Fisheries (2015); Campbell et al. (2017); Streipert et al. (2019)). In addition, barramundi were identified within the freshwater extents of the Southern Gulf catchments during two surveys of the freshwater fish of the major rivers in the study area. Barramundi were caught at five sites in the Leichhardt River, the site furthest upstream was also the highest at 209 mAHD (metres above Australian Height Datum), and four sites in the Nicholson River and its tributary, the Gregory River (furthest upstream and highest site at 148 mAHD) (Hogan and Vallance, 2005). In the Nicholson River catchment two barramundi were caught at Kingfisher Billabong (instream billabong, 83 mAHD) about 170 km upstream; five at Lake Corinda (instream pool, 31 mAHD) about 80 km upstream on the Gregory River; three at Pear Tree (instream pool, 84 mAHD) about 170 km upstream on the Gregory River; and two at Riversleigh Station (instream pool, 148 mAHD), about 180 km upstream on the O’Shanassy River (a tributary of Gregory River) (Hogan and Vallance, 2005). In addition, Hagedoorn and Smallwood (2007) sampled seven locations spanning the freshwater reaches of the Gregory River using electrofishing techniques during 2000 to 2005. They caught barramundi during replicate samples within at least one of the seven sites each year. In the Leichhardt River catchment five barramundi were caught below the Leichhardt Falls (instream billabong and pool, 6 mAHD) about 70 km upstream on the Leichhardt River; three at The Washpool (instream billabong, 24 mAHD) about 75 km upstream on the Alexandra River (tributary of the Leichhardt River); six at Abdy’s Waterhole (instream billabong, 26 mAHD) about 75 km upstream on the Alexandra River (a tributary of the Leichhardt River); three at Rocky Bar (instream billabong, 162 mAHD) about 300 km upstream on the Leichhardt River; and one below the Lake Julius Dam (waterfall pool, 209 mAHD) about 300 km upstream on the Leichhardt River (Hogan and Vallance, 2005). One survey of the freshwater reaches of the rivers provided a snapshot of barramundi distribution in freshwater habitats. However, it provided neither temporal trends of abundance nor spatial data on all stages of the life history of barramundi across the catchments. The survey conducted by Hagedoorn and Smallwood (2007) did provide catch per unit effort data for electrofishing sampling over six years, ranging from 0.29 fish to 3.02 fish per year, averaged over seven sites. Commercial catch data from Queensland’s N3 Net Fishery ‘southern gulf’ barramundi stock shows tonnages caught range from approximately 300 t to 900 t between 1981 to 2015 (Campbell et al., 2017). The ‘southern gulf’ stock extends beyond the bounds of the rivers of the Southern Gulf catchments considered in this report. Hence, barramundi abundance data from the estuarine and littoral habitats within and adjacent to the rivers of the Southern Gulf catchments are not specific. However, grid square ‘AF18’ of the Queensland fishing logbook program encompasses the ‘rivers of the Southern Gulf catchments’ group. The grid square is fished at a moderate level and barramundi is a target species. Using recreational fishing survey data to provide information on fish weights, Streipert et al. (2019) modelled recreational catch as a sensitivity test during assessment of the southern gulf barramundi stock. They estimated the recreational catch of barramundi from the southern gulf stock to be 143 t in 2000 and 79 t in both 2011 and 2014. In addition, to support scientific growth, ageing and movement studies, barramundi otoliths and fish frames have been collected from the commercial harvest in the southern gulf for decades (Leahy et al., 2022; Wright et al., 2021). The freshwater survey data and commercial catch data demonstrate that over recent decades barramundi have been abundant within the rivers of the Southern Gulf catchments as their life history necessitates use of estuarine and freshwater habitats within these river systems to sustain the adult population. Figure 3-1 Observed locations of barramundi (Lates calcarifer) and their modelled probability of occurrence in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for barramundi The barramundi life-history strategy is critically dependent on the natural flow regime in the wet- dry tropics and would be significantly affected by interruptions to the natural flows of northern Australian rivers (Crook et al., 2022; Roberts et al., 2023). Spawning occurs in the lower estuary and young male fish move into floodplain and freshwater habitats when suitable flows provide access (Crook et al., 2016; Roberts et al., 2019; Russell and Garrett, 1985). It had been thought that individual male barramundi moved upstream to freshwater riverine and palustrine habitats at about 3 to 4 years old before maturing and migrating downstream to the estuary at about 6 years old; there they would transform into females and mix with younger males and breed. Recent studies using new technologies have proposed that subsets of young-of-year individuals adopt a range of estuarine and riverine strategies (Roberts et al., 2019). Different migration strategies across freshwater and marine zones are triggered by variation in the wet- season flow regime and connectivity (Crook et al., 2016), making the species particularly vulnerable to water resource development (Doupé et al., 2005). The recruitment of barramundi to nursery habitats is moderated by floodwater access to supra- littoral, lagoon and riverine habitats (Russell and Garrett, 1983). Both longitudinal and floodplain connectivity require significant flood heights that let fish travel upstream or out of the river channel in search of habitats that increase their survival and growth during their juvenile stage. Peak spring tides also may facilitate access to supra-littoral habitats, supplemented by small early- season floods (Russell and Garrett, 1983); however, individuals also recruit to the population after spending larval and juvenile stages completely in estuarine water (Milton et al., 2008). Individuals around 1 year old move out of the supra-littoral habitats − they may move upstream into freshwater reaches (Russell and Garrett, 1985) or return to the estuary (Blaber et al., 2008; Milton et al., 2008) where they may reside for several years. Adolescents and adults remain in perennial freshwater habitats for periods of months to years until ephemeral flood-moderated connectivity lets them return to the river before emigrating downstream to the estuary and near-shore zones as adults to spawn (Blaber et al., 2008). Consequently, the annual wet season, and subsequent increase in flows, is a major determinant of their access to juvenile habitat and connectivity back to the coastal zone. There is a correlation between seasonal flood flow and juvenile recruitment strength and subsequent adult stocks, possibly lagged by 1 to 5 years (Halliday et al., 2012; Leahy and Robins, 2021; Robins et al., 2005). Historically, most northern rivers are unregulated with no large dams as barriers to migration, which supports life-history diversity in response to variability in monsoon-driven river flows (Roberts et al., 2023). Large instream dams that sever upstream−downstream connectivity may have greater effects on barramundi populations than reduced flows by blocking access to riverine and palustrine habitats within a large proportion of a catchment (Doupé et al., 2005). The ecological functions that support barramundi, and their associated flow requirements, are summarised in Table 3-1. Table 3-1 Ecological functions supporting barramundi and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for barramundi In the wet-dry tropics of northern Australia, the wet season stimulates primary productivity and connectivity within and between isolated palustrine and riverine habitats that are stressed by the end of the dry season (Ndehedehe et al., 2020a; Ndehedehe et al., 2021). Barramundi juvenile recruitment to freshwater habitats and fish growth rates are enhanced by large wet-season flows during the ‘peak flows’ wet-season months of January to March (Crook et al., 2022; Leahy and Robins, 2021) and also by flows higher than baseflows that precede and follow the wet-season peak-flow months. Higher flows during an early start to the wet season (October to December) or late rainfall (April to June) also promote superior growth than do low-level flows over the same months (Leahy and Robins, 2021). The research demonstrates that both the level and the seasonality (timing) of flood flows affects barramundi growth. High river flows expand the extent of palustrine and estuarine-margin habitats, increase connectivity, deliver nutrients from terrestrial landscapes, create hot spots of high primary productivity and food webs, increase prey productivity and availability, and increase migration within the river catchment (Burford et al., 2016; Burford and Faggotter, 2021; Crook et al., 2019; Leahy and Robins, 2021; Ndehedehe et al., 2021). These factors promote the successful recruitment of juvenile barramundi to freshwater habitats and the growth and survival of those that inhabit both freshwater and estuarine habitats within the river catchments. In years of poor wet seasons and low rainfall that result in naturally low-level flows, or flows that are reduced by anthropogenic activity such as water extraction, the range of facultative habitat and ecosystem processes available to barramundi is reduced; hence, growth and survival are reduced (Leahy and Robins, 2021; Robins et al., 2006; Robins et al., 2005). The impact of water resource development in three Gulf of Carpentaria rivers on coastal barramundi populations from the construction of dams and water harvest at several levels of extraction has been modelled (Plagányi et al., 2023). An array of water harvest and impoundment scenarios on the Mitchell, Flinders and Gilbert rivers in the eastern and southern Gulf of Carpentaria reduced both the biomass and commercial catch of barramundi by 4 to 61% depending on the water resource development scenario. The risk to the barramundi population was assessed as minor for one of four water resource development scenarios, the remaining three were assessed as negligible. However, risk to the commercial harvest of barramundi was assessed as moderate for two scenarios, minor for one and negligible for one (Plagányi et al., 2023). The model outputs included an explicit representation of the dependence of barramundi on lower estuarine species as prey, the abundance of which was also modified by change in river flows, that influenced the decline in the barramundi population. Recent research on monsoon-driven habitat use by barramundi has shown that, during drier years with lower river flows, a large proportion of the juvenile barramundi immigrate upstream from estuarine spawning habitat to freshwater habitats (Roberts et al., 2023). The modelling of the relationship between flow and catch undertaken for Gulf of Carpentaria rivers showed that maintaining low-level flows was important to support the barramundi population (Plagányi et al., 2023). Low-level flows maintained by pump-initiation thresholds that protected river flows from extraction until relatively high river flow levels were reached had less impact on barramundi biomass than the same allocation of water pumped at low-level river flows. The results of Roberts et al. (2023) assist to interpret the benefit of water-extraction pump thresholds for barramundi: during years of limited monsoon-driven flood flows, a higher proportion of juvenile barramundi immigrate upstream than during years of high-level flood flows that deliver abundant resources to estuaries. During drier years, barramundi use those unregulated lower-level flows to access riverine and palustrine habitats, to the benefit of the population in subsequent years. During years of high-level flood flows driven by a strong monsoon, most barramundi remain in the estuary, benefiting from riverine inputs transported to the estuary and not necessarily migrating upstream. Plagányi et al. (2023) showed that both constructing dams and harvesting river flows via pumped water extraction affect aspects of the barramundi life history that limit the resilience of its population. Anthropogenic reduction in the volume and duration of high-level flows and induced variability in the seasonality and volume of low-level flows affect habitat connectivity, migration, predation, growth and survival of barramundi (Leahy and Robins, 2021; Plagányi et al., 2023; Roberts et al., 2023). The ecological outcomes of threatening processes on barramundi in northern Australia catchments, and their implications for changes to growth, mortality, refuge and habitat, and community structure, are presented in Figure 3-2. Figure 3-2 Conceptual model showing the relationship between threats, drivers, effects and outcomes for barramundi in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.1.2 Bull sharks (Carcharhinus leucas) Description and background ecology Bull sharks (Carcharhinus leucas) are abundant in estuarine and freshwater reaches of tropical Australian rivers. Juvenile bull sharks occupy the freshwater reaches of most tropical Australian rivers (Constance et al., 2023a; Dwyer et al., 2019). Following access during floods, bull sharks can survive in isolated freshwater pools for over a decade awaiting future floods to re-establish connectivity with the riverine habitats (Gausmann, 2024). Adult bull sharks are ubiquitous in estuarine and marine habitats (Tillett et al., 2012; Werry et al., 2018). They can migrate >500 km along coastlines (Lubitz et al., 2023). There are no historical or recent records of speartooth sharks (Glyphis glyphis) and northern river sharks (Glyphis garricki) in the Southern Gulf catchments. Less well studied in the tropics, within a sub-tropical New South Wales coastal estuary, the primary carbon source of juvenile bull sharks within estuarine habitats varied with age. Young sharks (<2.5 years of age) relied on prey that consumed particulate organic matter, while older juveniles (4-6.5 years of age) preyed on species that relied on saltmarsh habitats to forage (Niella et al., 2022). Sharks that inhabit the freshwater reaches of rivers (especially freshwater-dwelling bull sharks) are of significance to the Indigenous Peoples using the freshwater and estuarine resources of the tropical rivers. Comments from elders may reflect a seasonal appearance of sharks in the freshwater reaches of the tropical rivers, which along with other species were food sources. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Bull sharks in the Southern Gulf catchments and marine region The Atlas of Living Australia records two instances of bull sharks in Southern Gulf catchments (Figure 3-3). A project investigating the reproductive philopatry of bull sharks in the Gulf of Carpentaria obtained 27 shark tissue samples (from commercial fishers) from the Roper River to the Robinson River, west of the Southern Gulf catchments, and 17 shark tissue samples from the Mitchell to Wenlock rivers to the east (Tillett et al., 2012), suggesting that bull sharks are common in all Gulf of Carpentaria rivers. The shark samples were collected in the river estuaries during commercial fishing activities. Though the rivers of the Southern Gulf catchments were not sampled, they also have extensive estuaries and freshwater reaches that were bull shark habitats in the sampled rivers to the east and west. Figure 3-3 Records of bull shark capture in the Southern Gulf catchments Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for bull sharks Long-stream riverine connectivity sustained by high-level flood flows is critical for bull sharks to access their riverine juvenile habitats and to return to estuarine breeding habitats and marine adult habitats (Table 3-2). Reduced high-level flows due to water extraction would reduce their access to habitats due to interrupted riverine−estuarine connectivity (Constance et al., 2023a; Tillett et al., 2012). On the Queensland and New South Wales coasts, bull shark abundance in coastal marine habitats increases following substantial rainfall (> 100 mm) in adjacent catchments (Smoothey et al., 2023; Werry et al., 2018), suggesting the sharks migrate from freshwater and estuarine habitats during flood flows. The ecological functions that support river sharks, and their associated flow requirements, are summarised in Table 3-2. Table 3-2 Ecological functions supporting river sharks and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for bull sharks The interruption of wet-season low flows and the reduction of both low and high flood flows by water diversion or impoundment will have a reduce the ability of bull sharks to access riverine habitats (Constance et al., 2023a). Juvenile bull sharks access freshwater habitats in lower and upper river reaches, where they forage for freshwater prey. Reduced river flows would reduce the spatial and temporal extent of riverine connectivity for juvenile sharks pupped within an estuary and immigrating upstream. Similarly, reduced flood flows would restrict the ability of sub-adult sharks to move downstream from palustrine habitats to estuarine and marine habitats (Gausmann, 2024). Physical barriers to low-level, dry-season flows (e.g. instream dams and barrages) may render the estuary hypersaline, prohibit connectivity with freshwater riverine habitats and expose rare species to high levels of predation (Morgan et al., 2017b). Bull sharks are vulnerable to both recreational and commercial fishing, mostly as non-target catch (Kyne and Feutry, 2017). Barriers and flow reduction would affect the habitat extent and movement of sharks within the estuary. The reduction of both low- and high flood flows would reduce prey availability for bull sharks. For example, sharks consume river prawns (Macrobrachium equidens), Johnius novaeguineae, an estuarine fish species, and barramundi (freshwater resident individuals described by Every et al. (2019)). The ecological outcomes of threatening processes on bull sharks in northern Australia, with their implications for changes to growth and mortality, community structure, habitat and population, are illustrated in Figure 3-4. Figure 3-4 Conceptual model showing the relationship between threats, drivers, effects and outcomes for bull sharks in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.1.3 Catfish (order Siluriformes) Description and background to ecology Catfish are a highly diverse group that inhabit both inland and coastal waters globally. In northern Australia, catfish are found both in freshwater and marine habitats. The group includes freshwater species, marine species, and some that move between the river and the estuary (Pusey et al., 2020). Catfish in northern Australia belong to two families: Plotosidae (19 species in total) and Ariidae (17 species). Plotosidae are found in the Eastern Pacific and Indian Ocean whereas Ariidae are a global family. Both are found in both freshwater and marine habitats. Most catfish are bottom-feeding omnivores. They feed on algae, submerged macrophytes, invertebrates and smaller fish. Species within the Ariidae family are slow-growing and generally large bodied. The family is notable for its reproductive traits: it has the largest eggs of any teleost For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au group (>10 mm) and males exhibit strong parental care behaviour, incubating the eggs and developing the young in the mouth for up to 5 weeks (Pusey et al., 2004). Because of the tendency to feed opportunistically, ariid catfish can be very competitive, consuming a variety and large volumes of food. Thus, they can make up a lot of biomass in a catchment (Crook et al., 2020). The key plotosid species are reasonably tolerant to high temperatures and low dissolved oxygen levels, but fish kills at very low dissolved oxygen levels have been reported (Bishop, 1980). The key threat to the two dominant Neosilurus species is potential flow barriers. Plotosidae need high flows to trigger spawning migration, and they require a barrier-free passage to spawning grounds in the headwater streams. While not as culturally and commercially important as barramundi or sooty grunters (Hephaestus fuliginosus), the fork-tailed catfish (Neoarius graeffei) has considerable importance as a subsistence fish for Indigenous communities (Finn and Jackson, 2011; Jackson et al., 2011). Catfish in the Southern Gulf catchments Catfish in the Southern Gulf catchments belong to two families: Ariidae (7 species, a family that is split between marine and freshwater species) and Plotosidae (4 species, mainly freshwater species in the Southern Gulf catchments). Figure 3-5 shows the observed locations of six key species: Neoarius graeffei (Figure 3-6), Neoarius midgleyi, Neosilurus ater, Neosilurus hyrtlii, Neosilurus pseudospinosus and Porochilus rendahli. Other catfish species include the ariid species Sciades leptaspis, Amissidens hainesi, Hexanematichthys mastersi, Nemapteryx armiger and Plicofollis argyropleuron. The larger-bodied ariid catfish like Neoarius graeffei, Neoarius berneyi and Sciades paucus are mainly found on the main stems of the Leichhardt, Gregory and Nicholson rivers. The usually smaller-bodied Neosilurus species are mainly found in smaller tributaries. The modelled distribution of N. graeffei is shown in Figure 3-6. Figure 3-5 Location of catfish in the Southern Gulf catchments Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-6 Modelled potential species distribution for fork-tailed catfish (Neoarius graeffei) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for catfish Flow–ecology relationships for catfish depend upon life histories, and therefore differ between the two families Plotosidae and Ariidae. The smaller-bodied Plotosidae can be found in many types of hydraulic habitat, including dune lakes (Arthington, 1984), but not the very large estuarine reaches (Pusey et al., 2004). Habitat use can change seasonally – a study in the Alligator River found Neosilurus hyrtlii in sandy creeks only during the late dry season. In the late wet and early dry seasons, N. hyrtlii was recorded from lowland lagoons, floodplain lagoons and perennial streams of the escarpment. Distribution seems to depend on context and habitat, as another study in a Queensland catchment found upstream migration in the wet season (Orr and Milward, 1984). Neosilurus ater prefers faster-flowing habitats in the main channel (Allen, 1982; Bishop et al., 1990) but has also been found to migrate upstream from its adult habitat in the lowland rivers to tributaries to spawn (Orr and Milward, 1984). Based on these observations, Pusey et al. (2004) conclude that ‘the development of water infrastructure that inhibits upstream movement, or which captures high-flow events and therefore removes the probable stimulus for spawning migrations, is highly likely to negatively impact on this species’. Ariidae is a family of fairly resilient catfish that, unlike Plotosidae, often prefer larger river channels or estuaries (Bishop et al., 2001). This is especially the case for Neoarius graeffei and Sciades leptaspis, which can tolerate slow-flowing or stagnant water; however, barriers can hinder dispersal for smaller size classes, even if barrier mitigation is provided (Stuart and Berghuis, 1999). Also, while not directly flow related, cold water from stratified impoundments can hinder spawning cues for N. graeffei (Kailola and Pierce, 1988). Neoarius midgleyi requires connection to the offchannel floodplain as habitat during the dry season. Kailola and Pierce (1988) also report N. midgleyi in a variety of habitats, including fast- flowing rivers, billabongs, creeks, deep pools and desiccating waterholes. This species is found less in the main channel and estuary. The ecological functions that support catfish, and their associated flow requirements, are summarised in Table 3-3. Table 3-3 Ecological functions supporting catfish and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for catfish This section discusses the possible ecological outcomes of threatening processes on catfish in northern Australia and their implications for changes in habitat shifts, community structure and population sizes presented in a conceptual model. Four of the key threats in the conceptual model are related to flow modification: water harvesting, dam infrastructure, river regulation and climate change. Overall, the ariid catfish species present in some northern catchments are fairly pollution tolerant, yet they all depend to some degree on a natural flow regime (Pusey, 2004). All species depend on connections to the floodplain, often for the purpose of recruitment. River regulation and extraction can reduce overbank flows, reducing connection frequency and therefore recruitment opportunities. As agricultural growing seasons often overlap with fish spawning seasons, water is likely extracted at these important times. Furthermore, environmental flows can be released at the wrong time (Linke et al., 2011), again leading to a possible reduction in recruitment and thus population size. Apart from being barriers to movement, dams can contribute to cold water pollution as released, stratified water can be significantly colder. While there are no data on catfish in tropical streams, Pusey (2004) hypothesises that in upland areas winter thermal tolerances of Neoarius graeffei are close to the thermal limit, indicating potential vulnerability to cold water releases from a dam. Some Plotosidae species prefer flowing water in the main channel. This could be affected by overextraction or even structural changes like dams, which can alter cease-to-flow periods. (Allen, 1982; Bishop et al., 1990). As described above, the combination of impact on movement and missing spawning migration triggers is highly likely to affect population sizes of Plotosidae, especially Neosilurus ater (Pusey, 2004). However, this could differ under varying circumstances; for example, some catfish in Queensland have been found to migrate upstream in the wet season (Orr and Milward, 1984). The ecological outcomes of threatening processes on catfish in northern Australia, with their implications for changes to habitat, community structure and population size, are illustrated in Figure 3-7. Figure 3-7 Conceptual model showing the relationship between threats, drivers, effects and outcomes for catfish in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.1.4 Grunters (family Terapontidae) Description and background to ecology In northern Australia there are a 37 species of grunter from 11 genera, with the most species-rich genera being Hephaestus, Scortum, Syncomistes and Terapon. Grunters inhabit riverine, estuarine and marine waters. Many grunter species spend their entire lives in fresh water, while other species inhabit marine or estuarine waters, only sometimes venturing into fresh water (Pusey et al., 2004). The Terapontidae are a perciform (perch like) family of fishes of medium diversity, restricted to the Indo-Pacific region. They are characterised by a single long-based dorsal fin, which has a notch marking the boundary between the spiny and soft-rayed portions. Terapontidae are soniferous (i.e. they can both vocalise and hear well) so may be sensitive to noise (Smott et al., 2018). One of the most widespread species is the sooty grunter (Hephaestus fuliginosus). Sooty grunters are omnivorous and eat a diverse diet, including terrestrial insects and vegetation, fish, aquatic insect larvae, macrocrustacea (shrimps and prawns) and aquatic vegetation. Sooty grunters switch diet from being insectivorous while juvenile to being top-level predators as adults, often feeding on smaller fish as well as juvenile grunters. Juvenile grunters are often associated with flowing water, suggesting that water harvesting that reduces or ceases flow could pose a threat. Tree root masses and undercut banks are also important microhabitat, especially for adult fish (Pusey et al., 2004). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Grunters prefer medium to high oxygen levels as well as medium to low salinity (Hogan and Nicholson, 1987). Grunters will move out of the dry-season refugial habitats and into ephemeral wet-season habitats for spawning (Bishop et al., 1990), with juveniles known to swim up to 7 km. The sooty grunter is an important recreational species, with environmental flow managed to maintain suitable habitat conditions (Chan et al., 2012). Because grunters are omnivorous and able to integrate many sources of food, as well as having a high total biomass, they are an important link in the overall food chain. They link lower trophic levels with top-level predators, such as long tom (Strongylura krefftii) and crocodiles. Grunters are also important species for Indigenous Peoples in northern Australia, both culturally (Finn and Jackson, 2011; Jackson et al., 2011) and as a food source (Naughton et al., 1986). Grunters in the Southern Gulf catchments and marine region Six species of grunters occur in the Southern Gulf catchments (Figure 3-8): spangled grunter (Leiopotherapon unicolor), barred grunter (Amniataba percoides), sooty grunter (Hephaestus fuliginosus), Gulf grunter (Scortum ogilbyi / S. hillii) and the estuarine terapon theraps (Pelates quadrilineatus). Waterholes on the main stem represent habitat for adult grunters, including the Gulf grunter, which was listed as Near Threatened by the IUCN in 2019. Figure 3-8 Location of grunters in the Southern Gulf catchments and marine region Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-9 Modelled potential species distribution for sooty grunter (Hephaestus fuliginosus) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for grunters Terapontidae have varying flow requirements. Not many studies have investigated the western sooty grunter Hephaestus jenkinsi; however, the very close relative H. fuliginosus is likely to have very similar preferences (Allen et al., 2002). It cannot be assumed that knowledge of the flow requirements of a charismatic taxon such as barramundi can be a surrogate for understanding the flow requirements of all life stages of sooty grunters. The most important species for recreational and cultural reasons, Hephaestus fuliginosus, is found in a variety of habitats, between headwater streams and the river mouths of the larger northern Australian streams; adults are rheophilic (i.e. have a preference for flowing water) (McDowall, 1996). In some tropical streams in Cape York Peninsula, passage to spawning habitat has been reported as a requirement (Herbert et al., 1995). There is scientific consensus that altered flow regimes are of concern to H. fuliginosus populations. It is the most rheophilic grunter species and is highly adapted to flowing water conditions (Pusey et al., 2004). Impoundments can inundate upstream riffles and fast-flowing sections that provide critical spawning areas. In general, regulation can both dry out critical habitat and connections and drown shallow refuge habitats. This can reduce riffles and runs, and reducing the diversity of flow environments within reaches is likely to reduce spawning success (Harris and Gehrke, 1994). While too much water can reduce population fitness, the loss of medium-magnitude flushing flows in the wet season would affect spawning sites (Hogan, 1994). Medium-size flow events also stimulate secondary production, and their loss could lead to a lack of food for grunter populations. Movement in both directions must be possible in order to accommodate the needs of both adult and juvenile grunters. Leiopotherapon unicolor and other smaller-bodied grunters have additional requirements to hydraulic habitat. L. unicolor needs a relatively high spawning water temperature of 20 to 26 °C (Allen et al., 2002). This can be compromised by hypolimnetic releases from impoundments (i.e. releases of the cold bottom layer of water). Movement is also key for L. unicolor’s life cycle. As a smaller species than the two Hephaestus species, L. unicolor often prefers to use floodplain wetlands as nursery habitat, making intermittent flooding important for recruitment success (Merrick and Schmida, 1984). • Amniataba percoides is a highly adaptable species, but data on its life-history are scarce. Pusey et al. (2004) postulated the fish also needs a balanced flow regime that is close to the natural regime on the basis of the following observations: • Amniataba percoides moves during its life cycle, hence impoundments are a threatening process (Bishop et al., 1995). • Amniataba percoides shows a preference for flowing water, but high flows are likely to reduce population sizes. • While Amniataba percoides is not dependent on floodplain spawning like some other fish taxa, floodplain connections increase population fitness. • Amniataba percoides needs high spawning temperatures, and desynchronisation of flow and thermal regimes by impoundments can reduce their fitness. The ecological functions that support grunters, and their associated flow requirements, are summarised in Table 3-4. Table 3-4 Ecological functions supporting grunters and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for grunters The possible ecological outcomes of threatening processes on grunters in northern Australia are discussed in this section and provided presented in the conceptual model (Figure 3-10). Four of the key threats in the conceptual model are related to flow modification: water harvesting, dam infrastructure, river regulation and climate change. For Amniataba percoides, changes in flow regimes that lead to faster-flowing environments can lead to decreased population viability – for example, a dam structure that first holds back water then releases it at higher velocity (Pusey et al., 2004). The key mechanisms for this are desynchronisation of thermal regimes and juvenile mortality caused by out-of-season high flows. The impact of regulation on Leiopotherapon unicolor has been documented by Gehrke (1997), who found that abundance was greatly reduced in regulated reaches. This is partly attributable to barriers to mobility, but also to a change in sediment composition, which leads to habitat alteration. Similarly, Hephaestus fuliginosus relies on flowing water, especially for spawning runs (Hogan, 1994). Barriers can interrupt these runs, leading to lower population viability. The loss of flushing flows can also lead to sediment build-up in key pool habitats – this effect is exacerbated by land use change. Climate change will interact with these threats in two ways: (i) it will enhance high flows that reduce populations, and (ii) droughts will interact with flow regimes, including with increased impacts of water extraction. The ecological outcomes of threatening processes on grunters in northern Australia, and their implications for changes to habitat, community structure and population size, are synthesised in the conceptual model shown in Figure 3-10. Figure 3-10 Conceptual model showing the relationship between threats, drivers, effects and outcomes for grunters in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.1.5 Mullet (family Mugilidae) Description and background to ecology Mullet (a guild including the genera Liza, Mugil and Moolgarda) are fish that use marine habitats as adults to spawn and freshwater habitats as juveniles (i.e. catadromous), although many species are more abundant in estuaries than freshwater habitat (Whitfield and Durand, 2023). They have life histories that entail ‘catchment to coast’ habitats (i.e. freshwater, estuarine and marine habitats), a feature which has supported speciation over evolutionary time (Marin et al., 2003; Whitfield et al., 2012). Mullet are distributed in tropical and temperate coastal waters worldwide (Whitfield and Durand, 2023); many species occur within any coastal ecosystem and mullet species often are dominant groups by biomass in coastal habitats. About 20 tropical mullet species occur in northern Australian waters from Townsville on the east coast to Broome in the west (Blaber et al., 2010). Diamond-scale mullet (Liza vaigiensis), largescale mullet (Liza macrolepis), greenback mullet (Liza subviridis), sea mullet (Mugil cephalus), roundhead mullet (Moolgarda cunnesius), bluespot mullet (Moolgarda seheli) and bluetail mullet (Moolgarda buchanani) are common species in the Australian tropics and range across the Indo-Pacific region (Blaber et al., 1995; Larson et al., 2013; Whitfield et al., 2012). These catadromous species are an abundant component of the fish community, being both forager and prey in the coastal ecosystem. Larvae are planktivorous, and juveniles feed on benthic invertebrates as well as prey in the water column. Adult mullet feed on organic detritus, benthic microalgae, filamentous algae, meiofauna (i.e. tiny sediment-dwelling invertebrates) and small Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. invertebrates (Górski et al., 2015; Soyinka, 2008; Whitfield et al., 2012). Being themselves preyed upon by larger species, mullet transfer energy from low to high trophic levels in the estuarine fish community; mullet are thus ecological link species (Górski et al., 2015). Their position as detritivores in the food chain and their fast growth rates and high fecundity make them a species group with high harvest potential. Mullet tend to grow fastest during the summer or tropical wet season, suggesting the influence of a seasonal increase in productivity of coastal waters (Grant and Spain, 1975b; Whitfield et al., 2012). By about 4 years of age, they leave nursery habitats for lower estuaries and the ocean. In general, mullets in Australia aggregate and spawn in marine waters in the lower reaches of estuaries or adjacent coastal waters in autumn to mid-winter before moving into coastal open- water habitats (De Silva, 1980; Grant and Spain, 1975b; Kailola et al., 1993; Robins et al., 2005). In the tropics Grant and Spain (1975a) suggest sea mullet live to about 7 years of age (~600 mm fork length), however in temperate regions, sea mullet may live to 16 years of age and grow to 640 mm fork length. In both climate regions they mature by 3-4 years of age at ~300-350 mm fork length (Lovett et al., 2022). Short-lived, fast-growing and maturing, and productive, mullet are important as a commercial, recreational and Indigenous fish resource. Mullet are one of the most important species groups taken in Queensland and NT recreational catches and the third most prominent finfish species taken in (non-Indigenous) recreational catches(Henry and Lyle, 2003; West et al., 2012). Most of the NT recreational mullet catches (92.4%) are targeted (West et al., 2012) rather than bycatch. Mullet are of cultural significance for Indigenous communities throughout Australia and are the most numerous species group in their catch (Henry and Lyle, 2003). In NT fisheries, they are a target for Aboriginal coastal fishing licences (Boyer, 2018; Wilton et al., 2018) and a target or bycatch in several fisheries (NT Government, 2022). Mullet species are minor bycatch in the Northern Prawn Fishery (NPF) (Pender et al., 1993). Mullet in the Southern Gulf catchments and marine region River diamond (or yellowfin) mullet (Planiiiza ordensis), greenback mullet (Liza subviridis), diamond-scale mullet (Liza vaigiensis), popeye mullet (Rhinomugil nasutus) and fantail mullet (Paramugil georgii) were found within the freshwater−brackish ecotone above the estuary, the estuarine reaches of the southern Gulf of Carpentaria rivers and the coastal marine environment (Figure 3-11). During two surveys of the freshwater fish of the major rivers in the study area, few mullet were identified within the freshwater extents of the Southern Gulf catchments. River diamond mullet (three individuals) were caught at one site in the lower Leichhardt River below the Leichhardt Falls and above the estuary (~6 mAHD) (Hogan and Vallance, 2005). They found no other species of mullet in the Leichhardt River. No mullet were found in the Nicholson or Gregory rivers or any other water body sampled within the Southern Gulf catchments (Hogan and Vallance, 2005). The major riverine and billabong habitats (e.g. Kingfisher Billabong, Pear Tree and Lake Corinda) within the Southern Gulf catchments (41 locations) were sampled by Hogan and Vallance (2005), and Hagedoorn and Smallwood (2007) sampled seven locations spanning much of the Gregory River. Hagedoorn and Smallwood targeted several mullet species (goldspot mullet (Liza argentea), sea mullet (Mugil cephalus), pinkeye mullet (Trachystoma petardi)), so their absence from abundance counts represented actual mullet absence. Immediately east of the Southern Gulf catchments, four river diamond mullet were caught in the Bynoe River at the Burke and Wills monument and one was caught in the Flinders River at the Burketown Crossing. Both of these catches were taken relatively close to the coast at 5 mAHD. Their presences reinforces the understanding that diamond mullet use the ecotone between riverine and estuarine habitats in rivers similar to those in the Southern Gulf catchments. One-off surveys of the freshwater reaches of the rivers of the Southern Gulf catchments provide a snapshot of mullet distribution in freshwater habitats. However, they provide neither temporal trends of abundance nor spatial data on all stages of the life history of mullet across the catchments. Recreational fishing surveys show that fishers target freshwater and enclosed estuaries along the Gulf of Carpentaria coast (Webley et al., 2015). However, mullet species are not a high proportion of the target catch and no data specific to them were reported from the Gulf of Carpentaria region (or Southern Gulf catchments) (Webley et al., 2015). Data from south-east Queensland show that mullet are not a high proportion of target catch for recreational fishers fishing from small vessels (Webley et al., 2009). However, mullet do comprise a significant portion of the Australia-wide recreational catch, and they represent the third most numerous portion of the finfish catch (8 and 14%, respectively) in Queensland and the NT. They often are targeted as bait and caught by bait net (Henry and Lyle, 2003). Mullet are a key target fish for Indigenous fishers, comprising 38% in of the Indigenous finfish catch in Queensland and 46% in the NT (Henry and Lyle, 2003). The indigenous take is mainly caught using nets, with a small proportion taken by line and spear in estuarine and coastal habitats. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-11 Records of capture of mullet in the Southern Gulf catchments and marine region Mullet juveniles use the mangrove and mudbank habitats within the estuary, and adult fish are caught within the estuary and in shallow subtidal habitats in the littoral zone. Mullet may use brackish and freshwater habitats during their juvenile phase. Data source: Atlas of Living Australia (2023a; 2023b) Flow–ecology relationships for mullet Mullet spawn in coastal marine areas where the larvae inhabit marine-salinity waters. As they grow, juvenile mullet migrate into estuaries and upstream to freshwater habitats (including palustrine wetlands) (Blaber et al., 1995; Gillson et al., 2009; Rolls et al., 2014). The frequency and duration of high flood events supports the inundation and availability of river floodplain and estuarine supra-littoral habitats used extensively by juvenile mullet during the wet season (O’Mara et al., 2021). Flooded palustrine habitats are hot spots for primary productivity (Burford et al., 2016; Ndehedehe et al., 2020a; Ndehedehe et al., 2020b) and refugia during the subsequent dry season (O’Mara et al., 2021). Reduced river flow volume and disrupted seasonality of flows affect mullet negatively by reducing the extent and connectivity of estuarine and freshwater habitats, affecting growth and survival via lower seasonal food accessibility and non-optimal environmental conditions (Jardine et al., 2013; Ndehedehe et al., 2021; Ndehedehe et al., 2020b) (Table 3-5). In extreme situations, a cessation of flow that closes an estuary to the marine environment for an extended period can cause mass mortality of mullet (Krispyn et al., 2021). Dry-season baseflows facilitate connectivity between estuarine and riverine reaches of Gulf of Carpentaria rivers. Brackish water and freshwater habitats are optimal for the growth and survival of mullet (Cardona, 2000; Whitfield et al., 2012) and lost connectivity reduces the population. Monsoon-season flood flows support the upstream and downstream migration of juvenile and adult mullet, respectively (Table 3-5). High-level flows allow access to inundated floodplain habitats for juvenile mullet, and cue emigration of sub-adults and adults to the marine environment. Constructed barriers such a weirs or dams block access to up-river habitats for juvenile mullet. Mullet are found in the riverine reaches of rivers in the Southern Gulf catchments so their catadromous life history remains critical to ontogenetic habitat selection over the full extent of catchment to coast (Larson et al., 2013; Waltham et al., 2013). The ecological functions that support mullet, and their associated flow requirements, are summarised in Table 3-5. Table 3-5 Ecological functions supporting mullet and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for mullet With marine and freshwater habitat use similar to barramundi, juvenile and early-adult phase mullet prefer fresh and brackish waters, including palustrine wetlands, that support optimal growth and survival (Cardona, 2000; Whitfield et al., 2012). Seasonal rainfall and flow likely influence downstream movements (Cardona, 2000; Gillson et al., 2009). A reduction in flow volume and seasonality may negatively affect mullet populations by reducing the extent and connectivity of the estuarine and freshwater habitats (Faggotter et al., 2013; Jardine et al., 2013; O'Mara et al., 2021) and disrupting cues for spawning movements. Disrupted connectivity by built barriers may limit use of freshwater habitats or block access to marine habitats if low-level flows cease (Grant and Spain, 1975b; O'Mara et al., 2021; Robins and Ye, 2007; Stuart and Mallen‐ Cooper, 1999). Wetland ‘perimeter to area ratio’ and wetland ‘number of patches’ can be strongly related to mullet catch, suggesting the extent and connectivity of estuarine habitats, intertidal and supra-littoral areas, and creeks and channels are important to mullet production (Meynecke et al., 2008). However, some individuals occupy wholly marine habitats despite available access to nearby estuaries (Górski et al., 2015). The ecological outcomes of threatening processes on mullet in northern Australia, and their implications for changes to growth and mortality, community structure, habitat and population, are illustrated in Figure 3-12. Figure 3-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mullet in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.1.6 Sawfishes (Pristis spp.) Description and background to ecology Sawfish belong to the order Pristiformes. They are characterised by a distinctive tooth-lined rostrum or ‘saw’. As adults, sawfish can attain very large sizes, ranging from 5 to 7 m in total length (TL) (Constance et al., 2023b). They are widely distributed in northern Australian marine waters, although they are not necessarily abundant (Last and Stevens, 2008; Morgan, 2011; Stevens et al., 2009). These species can migrate at landscape and oceanic scales through their life cycle, using inshore waters, including bays and estuaries, as important nursery grounds for neonates and juvenile sawfish, until about 4 to 6 years of age (Morgan, 2011; Morgan et al., 2017a; Peverell, 2005). As adults, they primarily inhabit tropical and subtropical coastal marine waters (Dulvy, 2016; Last and Stevens, 2008). Globally, sawfishes are considered one of the most threatened marine taxa (Dulvy, 2016). In Australian waters, there are four species of sawfishes, all listed of conservation significance at both national and international levels. The freshwater or largetooth sawfish (Pristis pristis), the green sawfish (P. zijsron) and the dwarf sawfish (P. clavata) are all listed as Vulnerable under the EPBC Act and Critically Endangered by the IUCN. The narrow sawfish (Anoxypristis cuspidata) is listed as Migratory under the EPBC Act, but because it is also listed in Appendix I and II under the Convention on the Conservation of Migratory Species of Wild Animals (Bonn Convention), it has For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au similar protection status under the EPBC Act. Additionally, sawfish hold significant cultural and spiritual relevance to Indigenous Australians (Ebner et al., 2016). Sawfish are vulnerable to multiple threats, partly due to their morphology (the shape of their rostra) and behaviour, and partly due to their life-history characteristics: long lives, slow growth and low reproductive rates, late maturation, relatively low abundance and high specificity of different habitats in different stages (Peverell, 2005; Phillips, 2017; Stevens et al., 2009). Their habitats often overlap with coastal fisheries, making them highly susceptible to capture in gill-net and trawl fisheries and recreational fishing (because the shape of their rostra). Marine and coastal protected areas have been established as refugia for sawfishes, though protection of their terrestrial range is deficient (Devitt et al., 2015). Sawfish rostra have been collected as trophies for decades (McDavitt, 1996), and there is a growing demand for live sawfish for display in public aquaria (Buckley et al., 2020; Compagno et al., 2006). Recent decades have seen high fishing mortality (Fry, 2021). Other pressures include cumulative impacts from climate change, habitat loss, artificial passage barriers and declining water quality that may have a significant impact on the movements of sawfish between freshwater and estuarine environments. Sawfishes in the Southern Gulf catchments and marine region The Gulf of Carpentaria is a haven ecosystem for the four recognised species of sawfishes found in Australia, with observations occurring across the Southern Gulf marine region (Figure 3-13). Worldwide, sawfish are among the most threatened marine taxa (Dulvy, 2016). Despite regional population declines due to fishing mortality over the past 50 years, northern Australia has viable populations of sawfish in contrast to sawfish populations elsewhere in Indo-Pacific region waters. Historically remote and minimally exploited, fishing activity in the Gulf of Carpentaria accelerated in the 1960s and 1970s, both inshore in littoral habitats and offshore in deeper waters. The most common sawfish in the Gulf of Carpentaria is the narrow sawfish, which is caught by the NPF as bycatch and is widespread over the western part of the Gulf of Carpentaria (caught in 546 trawls across the NPF from 2003 to 2019). Its abundance was an order of magnitude greater than the green sawfish (caught in 23 trawls) and the freshwater sawfish (caught in 14 trawls). Only a few dwarf sawfish were caught in the NPF over the same time period. For each species, the proportion of trawls in which no sawfish were caught was greater than 98%. Inshore, sawfishes are caught in gillnet fisheries and a study of sawfish rostra sourced from commercial gillnet fishers (known capture location over the past 100 years) recorded fish from estuaries and coastlines within the Southern Gulf catchments marine region, including Mornington Island (Wueringer et al., 2023). Within the littoral habitats, all four sawfish species were caught, and the abundances of species changed pre-2000 verses post-2000. Post-2000 the catches of green sawfish, in particular, declined (Wueringer et al., 2023). Historical anecdotal records and observations indicate the presence of large freshwater sawfish (P. pristis) and green sawfish (P. zijsron) in Southern Gulf catchments and estuaries. Photographs at the Gregory Downs Hotel show a 4.12 m freshwater sawfish caught nearby in the Gregory River at Gregory Downs Station in 1982. A rostrum of a 5.3 m long green sawfish was attached to the wall of the hotel – it was said to have been caught in the Gregory River even though green sawfish use estuarine and marine habitats. Similarly, rostra of green sawfish were attached to the walls of the Hells Gate Roadhouse (Nicholson River catchment) and the Kalkadoon Hotel, Kajabbi (Leichhardt River catchment), though in both cases the fish were likely caught in downstream estuaries. Residents of Floraville Station mentioned that 15 years ago researchers ran a sawfish acoustic tagging project in the area. They tagged two approximately 2 to 2.5 m long fish in the billabong below the station house. Recently, Floraville residents caught two small sawfish while cast netting for cherabin around Leichhardt Falls. One sawfish died and one was disentangled and released alive. Also, the residents mentioned that during a large flood a vehicle ran over a juvenile sawfish on a station roadway. Anecdotes by the publican of the Gregory Downs Hotel suggest that she had seen about four sawfish caught locally during her greater-than-40-year tenure at the hotel, and Terry Vallance (Hogan and Vallance, 2005) told of a small sawfish (~70 cm) caught at the ‘two mile’ on the Gregory River at Easter 2022 (T Vallance (Tropical River Consulting), 2022, pers. comm.). The physical and photographic evidence of sawfishes occurring in the Southern Gulf catchments is conclusive, though the locations of capture of the green sawfish cannot be confirmed. The verbal reports of sawfish interaction are anecdotal but common in the catchments. Figure 3-13 Records of sawfish capture in the Southern Gulf catchments and the marine region The location of sawfish species marked with a black square in the background, are estimated from Figure 7 of Wueringer et al. (2023). The locations may not be precise but would have an error <5 km. The approximate location of capture is representative of estuarine and coastal habitat used by sawfish in the Southern Gulf Catchments Marine Region. Data sources: Atlas of Living Australia (2023a; 2023b); Department of Environment and Science (2023); Department of Environment Parks and Water Security (2019a); Fry G (2021); Kenyon R (2022); OBIS (2023) Figure 3-14 Modelled potential species distribution for freshwater sawfish (Pristis pristis) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for sawfishes Sawfish species depend on estuarine and riverine habitats (Constance et al., 2023a). Juvenile freshwater sawfish (P. pristis) inhabit both estuarine and freshwater environments and migrate upstream to riverine reaches using high flood flows to access freshwater habitats (Morgan et al., 2016; Thorburn, 2007; Whitty, 2017; 2009). They can be found over 300 km upstream in freshwater riverine reaches. As juveniles (up to about 5 years old), they remain in refuge pools (see Section 3.4.2) during the dry season in the Australian tropics. At maturity, sawfish migrate downstream to estuarine habitats and become vulnerable to inshore gill-net fisheries, particularly in the monsoonal wet season (February to April) (Peverell, 2005). Riverine–estuarine connectivity and long-stream connectivity are critical for freshwater sawfish to access their juvenile habitats and to return to estuarine breeding habitats (Table 3-6). Dwarf sawfish use estuarine habitats and the lowermost riverine reaches seasonally, responding to salinity changes. In the Fitzroy River of WA, dwarf sawfish were found in a single large high- salinity pool at the uppermost tidal limit in the late dry season (August–November), before migrating downstream to close proximity to the river mouth or in King Sound during the wet and early dry seasons (December–July) (Morgan et al., 2021) in response to freshwater cues and seeking higher salinity waters. The green sawfish exhibits site fidelity within the estuarine and coastal habitat matrix in the vicinity of the mouth of tropical rivers. It moves to shallow coastal habitats at low tide and mangrove creek habitats at high tide (Lear et al., 2023; Morgan et al., 2017a). During large river-flow-discharge events they emigrate from the river estuary to coastal habitats. The ecological functions supporting sawfishes and their flow requirements are summarised in Table 3-6. Table 3-6 Ecological functions supporting sawfishes and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for sawfishes The potential implications of threatening processes for sawfishes in northern Australia are summarised in Figure 3-15. Changes in the depth, extent, duration and timing of flows in river reaches inhabited by sawfish can lead to habitat loss and significantly affect sawfish populations. These changes may reduce neonate recruitment (Morgan et al., 2016), affect the potential growth of individuals (Hunt et al., 2012), reduce the abundance of prey species that use floodplain wetlands during their life cycle (Novak et al., 2017), and/or reduce abundance and survivorship (Close et al., 2014; Jellyman et al., 2016; Morgan et al., 2016). Recent research in the Fitzroy River, WA, has identified critical flow characteristics in Australia's tropical rivers that support sawfish populations. Recruitment and survival of freshwater sawfish within riverine freshwater habitats was critically dependent on large flood flows. Freshwater sawfish recruitment to riverine habitats depended on extended periods of high-level flows (14 or more consecutive days in the 98th percentile of recorded water levels) to support access to the upstream freshwater river reaches that are their juvenile habitats (Lear et al., 2019). Remnant riverine pools are crucial refugia for sawfish during the dry season, within and during which they lose body condition – they lose greater body condition following low-volume wet-season flows than high-volume wet-season flows (Lear et al., 2021). It is likely that certain rivers are stronghold nursery habitats for freshwater sawfish as some rivers support consistent and high numbers of recruits (Lear et al., 2021). The maintenance of depth and stability of river pools during the dry season is critical to sawfish health, and disruptions to natural flows due to water impoundment or extraction can affect their survival (Figure 3-15) (Lear et al., 2020). During the early wet season, re-established connectivity downstream to estuarine habitats also is crucial, and modification of early-season flows or low- level flows during a poor wet season may delay or reduce riverine connectivity. Other pressures include the cumulative impacts from climate change, habitat loss, artificial passage barriers and declining water quality that may have a significant impact on the movements of sawfish between freshwater and estuarine environments. The impact of water resource development in three Gulf of Carpentaria rivers on coastal sawfish populations has been documented for the construction of dams and water harvest at several levels of extraction (Plagányi et al., 2023). Different water resource development scenarios on the Mitchell, Flinders and Gilbert rivers in the eastern and southern Gulf of Carpentaria reduced both the biomass of sawfishes by approximately 50 to 80% depending on the scenario. The risk to the sawfish population was assessed as severe and/or extreme for three of four water resource development scenarios and moderate for the remaining one (Plagányi et al., 2023). The model outputs included an explicit representation of the dependence of sawfish on lower estuarine prey species abundance that also were flow-dependent, hence their decline influenced the decline in the sawfish population. Plagányi et al. (2023) showed that both the construction of dams and the harvest of river flows via pumped water extraction affect aspects of sawfish life history that reduce their population resilience. Anthropogenic reduction in the volume and duration of high-level flows, as well as variability in the seasonality and volume of low-level flows affects riverine and estuarine connectivity, habitat suitability, body condition, growth, survival and especially upstream migration of sawfish (Lear et al., 2019; Lear et al., 2021; Plagányi et al., 2023). The ecological outcomes of threatening processes on sawfish in large rivers in northern Australia, and their implications for changes to growth, population and community structure, are illustrated in Figure 3-15. Figure 3-15 Conceptual model showing the relationship between threats, drivers, effects and outcomes for freshwater sawfish (Pristis pristis) in large rivers in northern Australia The conceptual model has only been developed for P. pristis, owing to the lack of information on the other three relevant sawfishes in relation to hydrological change. Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.1.7 Threadfin (Polydactylus macrochir) Description and background to ecology King threadfin (Polydactylus macrochir, formerly P. sheridani) is a large (>1.5 m) non-diadromous, carnivorous fish (Order Perciformes). Endemic to Australasia, it is found from the Ashburton River / Exmouth Gulf, WA, across northern Australia, southern Papua New Guinea and Irian Jaya to the Brisbane River in Queensland (Motomura et al., 2000). King threadfin is long-lived (22 years) and fast-growing. Individuals begin life as a male but change to female as they age (protandrous hermaphrodites). They reach sexual maturity at around 60–80 cm TL and 2–4 years of age, and change from male to female around 70–100 cm TL at 4–8 years of age (Leigh et al., 2021). Their body form and quality of flesh makes threadfin a prized table fish. It is typically the second- most important target species in the commercial inshore gill-net fisheries that principally target barramundi (Welch et al., 2010). From 2015 to 2019, the Gulf of Carpentaria (Queensland) harvest of king threadfin averaged 203 tonnes per year (Leigh et al., 2021). In 2018–19, 235 t of king threadfin and blue threadfin (Eleutheronema tetradactylus) worth $923,000 were taken in the NT, while 218 t of king threadfin worth $946,000 were taken in Queensland waters (Steven et al., 2021). Threadfin are also target species for recreational and Indigenous fisheries throughout wet-dry tropical Australia (Moore et al., 2011). King threadfin are of cultural significance for the Indigenous Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. community, and in key localities in the vicinity of Indigenous townships in the NT they are subject to management plans specifying season and bag limits (Malak Malak: Land and Water Management, 2016). King threadfin complete their entire life cycle in turbid coastal waters, in estuaries, mangrove creeks and inshore marine waters. They tolerate a wide salinity range from marine salinity down to as low as 2 ppt, but are not found in freshwater habitats (Blaber et al., 1995; Moore et al., 2012). Adults probably spawn in inshore coastal waters and lower parts of estuaries (Halliday and Robins, 2005; Welch et al., 2014). High salinity (>32 ppt) is important for survival of the pelagic eggs, and spawning occurs in marine waters away from the outflows of river mouths, avoiding lower salinity levels (Halliday et al., 2008; Robins and Ye, 2007; Welch et al., 2014). Young fish likely enter estuaries during the wet season when prawns and other prey species are seasonally abundant. Turbid waters during wet-season flows may protect young threadfin from large predators (Welch et al., 2014). King threadfin particularly inhabit the mid-to-upper estuary, but they are thought to restrict their use of estuarine habitats to permanent water areas in the main channels and tributaries of creeks and rivers (Halliday et al., 2008). Older fish inhabit estuarine and marine systems. King threadfin are a top predator capable of modifying the estuarine fish and crustacean community in which they live (Salini et al., 1990; Salini et al., 1998). Although king threadfin are restricted to brackish estuarine and marine conditions, the extent and patchiness of wetland and salt flat habitats are likely to be important to king threadfin production (Meynecke et al., 2008), perhaps via productivity and availability of prey. Preying on a range of fish and crustaceans in the coastal ecosystem (Blaber et al., 1995; Salini et al., 1990), king threadfin exemplify an estuary-dependent fish that hunts successfully in turbid waters (Salini et al., 1998). Threadfin are not obligate visual predators; they also use tactile sensors (pectoral filaments) to detect their dominant crustacean prey (prawns) (Pember, 2006; Salini et al., 1998). As adults, their success as a predator may be significantly affected by interruptions to the high-level natural river flows that maintain the turbid, brackish ecotone of tropical rivers within which they successfully hunt. Threadfin in the Southern Gulf catchments marine region King threadfin are numerous in coastal habitats of the Southern Gulf catchments marine region, though due to the remoteness of the rivers, commercial fish catch is the key source of most information (Leigh et al., 2021); this is reflected in the recorded observations of this group in the Southern Gulf catchments (Figure 3-16). In 2009, 289 t of king threadfin (approximate value $1,200,000) were harvested from Queensland Gulf of Carpentaria waters, while catch from 2015–19 averaged 203 tonnes per year, with significant fishing based around towns such as Burketown and Karumba, within and adjacent to the Southern Gulf catchments marine region (Leigh et al., 2021; Welch et al., 2014; Welch et al., 2010). A 2020 stock assessment used fishery logbook data collected from the ‘Mornington’ region within the Gulf of Carpentaria management region (Leigh et al., 2021). Within the region, Moore et al. (2012) collected fish frames of king threadfin from Arthurs Creek (42 fish), Albert River (35 fish) and Morning Inlet (50 fish) to study stock population connectivity across tropical Australia. Using parasites as indicators of stock similarity and limited movement, fish from these rivers were identified as an overlapping stock, though with connectivity to rivers such as the Roper River to the north and the Flinders River to the east (Moore et al., 2012). A 2017 study Moore et al. (2017) collected king threadfin from the Albert River and Morning Inlet (within the Southern Gulf catchments marine region) as well as four other southern gulf rivers and demonstrated that overfishing in the post-2000 decade has had major impacts on the demography of the population compared to the 1980s (Moore et al., 2017). The authors suggest that the removal from the population of large female fish due to fishery harvest has reduced the maximum and modal age of fish from 14 years and 5 years (respectively) to 8 years and 3 years (respectively) between the two periods. In addition, the sex ratio of the population changed, and individuals changed from male to female younger at a smaller size (Moore et al., 2017). Across tropical Australia, sampling conducted within Queensland, the NT and WA revealed stocks separated by tens to hundreds of kilometres or by large coastal geographical features (Moore et al., 2011; Welch et al., 2010). Figure 3-16 Records of threadfin capture in the Southern Gulf catchments marine region Threadfin juveniles use the mangrove and mudbank habitats within the estuary and fish are caught within the estuary and in shallow subtidal habitats in the littoral zone. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for threadfin River flow effects are evident in commercial catch data. Halliday et al. (2012) recorded that, after adjusting for fishing effort, the annual king threadfin commercial catch from 1990 to 2009 was significantly positively correlated with spring rain lagged by 3 years. It was also significantly but negatively correlated with autumn rain in the year of catch. King threadfin do not use freshwater habitats, so the effect of flood flows on their abundance is less-well defined (Halliday et al., 2012). However, flood flows are key environmental drivers for king threadfin prey, so flow effects on threadfin populations are moderated by food webs, tide regimes, and catchment and estuarine productivity (Jinks et al., 2020). In some tropical and subtropical rivers, the year-class strength of king threadfin was positively correlated with spring and summer flood flows (Halliday et al., 2008; Halliday et al., 2012). Baseflow in the spring and early-season low flows are used by threadfin larvae in marine habitats as cues to access estuaries. Monsoon flows create a brackish ecotone within estuaries that is prime habitat for threadfin and their prey (Cardona, 2000; Russell and Garrett, 1983; Vance et al., 1998) (Table 3-7). In addition, flood flows deliver nutrients and increase turbidity in estuaries, supporting the food chain and minimising predation; both aspects enhance the survival of juvenile threadfin. Small fish and crustaceans (including penaeid prawns, the prime prey of king threadfin), are abundant in tropical estuaries in the pre-wet and wet seasons (Jinks et al., 2020; Salini et al., 1990; Salini et al., 1998; Vance et al., 1998). In the Australian tropics, the prey community is supported by turbid wet-season flows, though turbidity is also advantageous for fish that do not rely on visual predation alone, such as threadfin. The ecological functions that support threadfin, and their associated flow requirements, are summarised in Table 3-7. Table 3-7 Ecological functions supporting threadfin and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for threadfin During a 5-year study in a large Queensland subtropical estuary, king threadfin year-class strength (indicating recruitment and survival of juvenile king threadfin) was positively related to the annual levels of freshwater flow during spring and summer (Halliday et al., 2008). In the Gulf of Carpentaria and the Daly River, both commercial catch (as a measure of abundance) and year-class strength were positively related to monsoon rainfall (often year lagged) in some rivers, but not for all river flows (Halliday et al., 2012; Welch et al., 2014). The survival and growth of king threadfin is likely supported by higher estuarine productivity and abundant prey in years of high flood flow, though these relationships are not robustly studied in the Gulf of Carpentaria (Halliday et al., 2012; Moore et al., 2012). The frequency and duration of high flood events supports the annual inundation and enhanced primary productivity of floodplain and estuarine supra-littoral habitats (Burford et al., 2016; Ndehedehe et al., 2020a). Carbon and nutrients that are exported to the estuarine and near-shore habitats are used by king threadfin and their prey. Reduced natural flow volumes and interrupted seasonality of monsoon floods would reduce the growth and abundance of king threadfin, as has been found for other large predatory fish that use Gulf of Carpentaria estuaries as prime habitat (Leahy and Robins, 2021). In the southern Gulf of Carpentaria, overfishing has stressed the population of king threadfin, reducing individual size and age within the population, and causing individuals to mature younger at a smaller size (Moore et al., 2017); hence further stress on the population due to the interruption of the historical flow regime may confound population recovery. The ecological outcomes of threatening processes on threadfin in northern Australia, and their implications for changes to growth and mortality, community structure, habitat and population, are illustrated in Figure 3-17. Figure 3-17 Conceptual model showing the relationship between threats, drivers, effects and outcomes for threadfin in northern Australia Blue arrows represent hydrological and black arrows represent non-hydrological changes. 3.2 Waterbirds Freshwater and saltwater habitats throughout northern Australia are home to a diverse range of waterbird species. Waterbirds are highly dependent on the resources provided by these habitats, including food, shelter and nesting opportunities, all of which are critical for species survival and population maintenance. In most of these habitats, waterbird behaviour, movement and distribution and reproductive ecology are largely dependent on natural flooding and rainfall events (Kingsford and Johnson, 1998). Waterbirds respond to flooding and rainfall and subsequent primary and secondary productivity by building condition, moving to important wetland sites and breeding (Brandis et al., 2009). Consequently, waterbirds are recognised as important indicators of aquatic ecosystem quality and environmental variability (Garnett et al., 2015 ; Rahman and Ismail, 2018). Worldwide, populations of waterbirds are in decline, with many species listed as Threatened, Endangered, or Critically Endangered. In Australia, species such as the eastern curlew (Numenius madagascariensis), brolga (Antigone rubicunda) and Australian painted snipe (Rostratula australis) are listed as priority species through state, federal or international agreements and legislation (Kingsford, 2013). Waterbird population declines are primarily driven by changes in habitat and food availability and quality, usually driven in turn by changes in river flow and flood regimes through construction of dams and weirs, water extraction from rivers, water harvesting from floodplains, draining of wetlands, loss of intertidal habitat, over-fishing, water quality changes and For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au other anthropogenic impacts. Consequently, waterbirds are a focal group and indicator for the conservation and management of aquatic and semi-aquatic habitats across northern Australia (Bellio et al., 2004; Butchart et al., 2010). Their unique characteristics, visual appeal and social behaviours have historically influenced human culture and continue to engage people and communities with their environments, for example through cultural activities, traditional stories, symbology, hunting and birdwatching (Kushlan et al., 2002). To provide a simple basis for understanding and communicating the associated risks and opportunities for waterbirds related to potential water resource development in northern Australia, waterbird species have been divided here into four high-level groups. These groups are based on foraging behaviour and habitat dependencies, and nesting behaviour and habitat dependencies. Both foraging and nesting dependencies need to be taken into account, because while some species both forage and nest in northern Australia, others are more mobile, using environments outside of northern Australia for either nesting or foraging. The four waterbird groups are: •colonial and semi-colonial nesting waders •cryptic waders •shorebirds •swimmers, grazers and divers. Group 1: Colonial and semi-colonial nesting waders (Section 3.2.1). Colonial and semi-colonial nesting wading species have a high level of dependence on flood timing, extent, duration, depth, vegetation type and condition for breeding. They are also often dependent on specific important breeding sites in Australia. They are usually easily detectable when breeding, and good datasets are available for most species. These species are typically nomadic or partially migratory. Group 2: Cryptic waders (Section 3.2.2). Cryptic wading species have a high level of dependence on shallow temporary and permanent wetland habitats with relatively dense emergent aquatic vegetation that requires regular or ongoing inundation to survive (e.g. reeds, rushes, sedges, wet grasses and lignum). These species breed in Australia and usually nest as independent pairs, though some may occasionally nest semi-colonially. They may be sedentary, nomadic, migratory or partially migratory. Few data are available; however, habitat requirements can be used as surrogates to assess vulnerability. Group 3: Shorebirds (Section 3.2.3). Shorebirds have a high level of dependence on end-of-system flows and large flood events that provide broad areas of very shallow water and mudflat type environments. They occur across freshwater and marine habitats and are largely migratory or nomadic. Many breed in the northern hemisphere rather than Australia but depend on Australian environments to survive the rest of the year. Shorebirds are a group of international concern. Group 4: Swimmers, grazers and divers (Section 3.2.4). These are species with a relatively high level of dependence on semi-open, open and deeper-water environments. These species commonly swim when foraging (including diving, filtering, dabbling, grazing) or when taking refuge. They breed in Australia and may be sedentary, nomadic, migratory or partially migratory. To support the ecology assessment, example species from each group have been selected (see Table 3-8). Species selected are good representatives of the group as a whole, of conservation or cultural importance, and likely to be affected by water resource development. These species provide examples for synthesising the pathways to impact associated with potential water resource developments. Table 3-8 Waterbird species groups and example representative species for northern Australia For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. The primary pathways of potential water resource development impact on waterbirds include: (i)habitat loss, fragmentation and change, (ii) toxins from pollution or contaminants, (iii)disturbance from human activities, (v) predation by invasive or feral animals, and (v) changesin disease, or parasite burdens. Habitat loss, fragmentation and change are the most importantdrivers of changes in waterbird abundance, population size and diversity worldwide (McGinness, 2016). The toxic effects of pollution or contaminants such as pesticides, heavy metals, nutrients and other chemicals are known to have caused declines in many populations of waterbirds worldwide (De Luca-Abbott et al., 2001; Howarth et al., 1981; Kim and Oh, 2015). Besides their direct toxic effects, pesticides and herbicides can reduce food availability for waterbirds, depending on their diet. Changes in the extent or intensity of water resource development and subsequent agricultural developments are often associated with increases in the amounts of pollution or contaminants such as pesticides, heavy metals, nutrients and other chemicals in catchments, and therefore present risks to waterbird populations. Predation is a natural component of waterbird population biology. However the nature and importance of its impact can be changed by anthropogenic changes, in particular, the introduction of feral predators such as pigs and habitat alteration via introduced plants and herbivores (Sovada et al., 2001). Changes in water levels during nesting periods can make nests more accessible and vulnerable to predators (McGinness, 2016). Many studies have shown that predation on waterbirds occurs mainly during nesting and is dominated by egg predation; nestling and fledgling predation are also reported. Predation on adult waterbirds is relatively rare, but is probably additive to mortality due to other factors (e.g. hunting, pollution (Sovada et al., 2001)). Predators such as pigs can reduce the survival of waterbirds, and consequently population size, either through direct predation or indirectly by causing adults to desert their nests or foraging sites. They can also affect population size by competing for habitat or food, or affecting other predators and prey (Cruz et al., 2013; MacDonald and Bolton, 2008; Skorka et al., 2014). Disturbance from human activities can affect bird behaviour and temporal and spatial distribution of waterbirds. Human disturbance can be equivalent to habitat loss or degradation because it may lead waterbirds to avoid or underuse areas (Fernandez and Lank, 2008). Temporary loss of foraging habitats can occur, and species vary in their capacity to compensate by foraging for longer periods (Sutherland et al., 2012). During the breeding season, human disturbance may also influence nest incubation and chick rearing, affecting overall nest success and eventual recruitment, which then affects population sizes and trajectories. Disease and parasites can affect waterbird nest success, fledging rates, juvenile survival and adult survival. They are more likely to be a problem where there is insufficient habitat and birds are crowded, which can occur following changes in flood regimes and habitats due to water resource development or land development (McGinness, 2016). Infectious diseases are an important and dominant mortality factor in waterbird populations. Bacteria such as Clostridium botulinum and viruses such as avian influenza, West Nile virus, Newcastle disease virus, avian poxvirus, duck plague, avian bornavirus, reoviruses and adenoviruses may contribute to population declines of both wild and domestic waterbirds. The infection rate by Plasmodium parasites (avian malaria) is rapidly increasing in many birds, and infection rates of campylobacteria in waders are high (Sutherland et al., 2012). Ticks parasitising nestlings can reduce survival and nest success, and potentially also transmit viruses. Changes in land use and global climate may concentrate waterbirds on remaining high-quality sites, making them potentially more vulnerable to infections (Sutherland et al., 2012). Where impacts on waterbird populations are natural processes (e.g. predation, disease), anthropogenic influences have almost always altered those processes, as described above. Consequently, such processes can become management problems, even though they are fundamentally natural. Interactions are also likely with climate change. Climate change is affecting seasonal and extreme temperatures; the timing, intensity, amount and duration of rain; and the frequency and severity of extreme weather events, all of which have the potential to influence waterbird populations positively and negatively, and directly and indirectly (Chambers et al., 2005; Sutherland et al., 2012). 3.2.1 Colonial and semi-colonial nesting wading waterbirds Description and background to ecology The colonial and semi-colonial nesting wading waterbirds (colonial waders) group comprises wading waterbird species with a high level of dependence on water for breeding, including requirements for flood timing, extent, duration, depth, vegetation type and vegetation condition. In northern Australia, this group comprises 21 species from 5 families, including ibis, spoonbills, herons, egrets, avocets, stilts, storks and cranes (Table 3-9). The species in this group are often easily detectable when breeding, and relatively good datasets are available for most, unlike for other species or groups. The species in this group often depend on specific important breeding sites (Arthur et al., 2012). Ibis, spoonbills, herons, egrets, avocets and stilts nest in loose groups or dense colonies of hundreds to tens of thousands of birds in specific vegetation types and locations, over or adjacent to water (Bino et al., 2014). Storks (such as the black-necked stork (Ephippiorhynchus asiaticus)) and cranes including the brolga (Antigone rubicunda) and sarus crane (Antigone antigone) usually nest independently, but loose widely spaced groups of nests may occur in suitable habitat. Species in this group may travel up to thousands of kilometres to use these sites (McGinness et al., 2019), and nesting events can last several months, depending on inundation conditions (Kingsford et al., 2012). Species in this group usually have a mixed diet including fish, frogs, crustaceans and insects, and use foraging methods such as walking, stalking and striking to catch their prey. Colonial and semi-colonial nesting waders generally prefer shallow water or damp sediment with medium- to low-density vegetation for foraging (Garnett et al., 2015). These species are typically nomadic or partially migratory but may spend long periods in particular locations when conditions are suitable. For the assessment, the species selected as representative of the colonial and semi-colonial nesting waders group is the royal spoonbill (Platalea regia; Figure 3-18). The royal spoonbill is a large wading species highly adapted to foraging in shallow wetlands (Marchant and Higgins, 1990). This species requires water and water-dependent vegetation for feeding, nesting, refuge, roosting and movement habitat (e.g. ‘stopover’ habitat for longer distance trips) (Marchant and Higgins, 1990). Spoonbills nest in loose colonies, usually in vegetation surrounded by water, including reedbeds, semi-aquatic shrubs and trees. They often nest adjacent to colonies of other species in the group. Figure 3-18 Royal spoonbill (Platalea regia) at the nest Royal spoonbills are a representative species of the colonial and semi-colonial nesting wading waterbird group. Photo attribution: CSIRO Colonial and semi-colonial nesting waders in the Southern Gulf catchments Colonial and semi-colonial nesting wading waterbirds are found throughout the Southern Gulf catchments (see Figure 3-19, Figure 3-20, Figure 3-21 and a modelled potential species distribution for royal spoonbill Figure 3-22). Common species include the black-necked stork, brolgas, herons, egrets, ibises and spoonbills (Atlas of Living Australia, 2023a; 2023b). Figure 3-19 Location of colonial and semi-colonial nesting wading waterbirds in the Southern Gulf catchments in alphabetic order of species name: Antigone rubicunda (brolga) to Egretta novaehollandiae (white-faced heron) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-20 Location of colonial and semi-colonial nesting wading waterbirds in the Southern Gulf catchments in alphabetic order of species name: Egretta picata (pied heron) to Recurvirostra novaehollandiae (red-necked avocet) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-21 Location of colonial and semi-colonial nesting wading waterbirds in the Southern Gulf catchments in alphabetic order of species name: Threskiornis molucca (Australian white ibis) to Threskiornis spinicollis (straw- necked ibis) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-22 Modelled potential species distribution for royal spoonbill (Platalea regia) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Table 3-9 Species in the colonial and semi-colonial nesting wading waterbird group, and their international conservation status (LC = Least concern) SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME POPULATION TYPE IUCN STATUS EPBC ACT STATUS Australian white ibis Threskiornis moluccus Threskiornithidae Australian LC _ Banded stilt Cladorhynchus leucocephalus Recurvirostridae Endemic LC _ Pied stilt (black- winged stilt) Himantopus leucocephalus Recurvirostridae Australian LC _ Cattle egret Bubulcus ibis Ardeidae Australian LC _ Eastern reef egret Egretta sacra Ardeidae Australian LC _ Glossy ibis Plegadis falcinellus Threskiornithidae Australian LC _ Great Egret (eastern great egret) Ardea alba (Ardea alba modesta) Ardeidae Australian LC _ Great-billed heron Ardea sumatrana Ardeidae Australian LC _ Plumed egret (recently split from intermediate egret) Ardea plumifera (recently split from Ardea intermedia) Ardeidae Australian LC _ Little egret Egretta garzetta Ardeidae Australian LC _ Nankeen night-heron Nycticorax caledonicus Ardeidae Australian LC _ Pied heron Egretta picata Ardeidae Australian LC _ Red-necked avocet Recurvirostra novaehollandiae Recurvirostridae Endemic LC _ Royal spoonbill Platalea regia Threskiornithidae Australian LC _ Sarus crane (Australian sarus crane) Grus antigone (Grus antigone gillae) Gruidae Australian (Endemic) Vulnerable _ Straw-necked ibis Threskiornis spinicollis Threskiornithidae Endemic (breeding) LC _ White-faced heron Egretta novaehollandiae Ardeidae Australian LC _ White-necked heron Ardea pacifica Ardeidae Endemic LC _ Yellow-billed spoonbill Platalea flavipes Threskiornithidae Endemic LC _ Black-necked stork Ephippiorhynchus asiaticus Ciconiidae Australian LC _ Brolga Antigone rubicunda Gruidae Australian LC _ Flow–ecology relationships for colonial and semi-colonial nesting wading waterbirds Waterbird species in the colonial and semi-colonial nesting waders group are sensitive to changes in the depth, extent and duration of shallow wetland environments, particularly during nesting events. Colonial nesting waders nest when and where weather, water and vegetation provide optimal conditions, including suitable vegetation structure and water around nests for protection from predation and weather (Kingsford and Norman, 2002) and sufficient food resources (Figure 3-23) (O’Brien and McGinness, 2019). Completion of a full nesting cycle can take several months. During this time, changes in water depth, water extent, water duration or food availability can force adults to abandon their nests or expose nests to predation, resulting in nest failure, and in the long term can result in abandonment of regular breeding sites (Brandis, 2010; Brandis et al., 2011). Adults of these species may not breed every year, and recruitment rates post-breeding are frequently low because of this dependence on suitable hydrological and weather conditions to support food resources and habitats. Nesting failures may have a serious impact on population sizes and trajectories (Kingsford and Norman, 2002). The ecological functions that support colonial and semi-colonial nesting waders, and their associated flow requirements, are summarised in Table 3-10. Table 3-10 Ecological functions supporting colonial and semi-colonial nesting waders and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-23 Egret hunting among water lilies Egrets are species of the colonial and semi-colonial nesting waders group. Photo attribution: CSIRO Pathways to change for colonial and semi-colonial nesting wading waterbirds The primary pathways of potential water resource development impact on colonial waders are habitat loss, fragmentation and change (Figure 3-24). Because of the specific needs of colonial waders regarding water regimes in suitable nesting habitats, colony sites in areas subject to changes in flood regimes due to water resource developments (e.g. river regulation through dams or weirs, water extraction from rivers, floodplain water harvesting) are at high risk of damage or loss, with implications for population maintenance (Brandis et al., 2011). Unnatural or unexpected changes in the depth, extent, frequency and duration of inundation in wetland habitats used by colonial and semi-colonial nesting waders for nesting and foraging can have significant impacts on nesting, nest success, juvenile recruitment and adult survival (Bino et al., 2014; Brandis et al., 2018; Brandis et al., 2011; Kingsford et al., 2011). Changes can also reduce water quality and food availability, and increase rates of competition, predation and disease (McGinness, 2016). Changes can occur when flood peaks are reduced by water extraction or dams (e.g. by reducing flood extent, frequency, duration or depth), when floodwater is captured on floodplains (e.g. by dams, levees or roads), or when the time between inundation events that create these habitats is extended (Kingsford and Thomas, 2004). The life histories of many of these species have evolved to expect natural flooding regimes, so they are affected when these regimes are changed. Photo egret. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-24 Conceptual model showing the potential relationship between threats, drivers, effects and outcomes for colonial and semi-colonial nesting wading waterbird species Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.2.2 Cryptic wading waterbirds Description and background to ecology The cryptic waders group comprises wading waterbird species that are relatively difficult to detect and have a high level of dependence on shallow temporary and permanent wetland habitats with relatively dense emergent aquatic vegetation (Figure 3-25) that requires regular or ongoing inundation to survive (e.g. reeds, rushes, sedges, wet grasses). In northern Australia, this group comprises 13 species from four families, including bitterns, crakes, rails and snipe (Table 3-11). Species from this group are often present in low numbers and are difficult to detect even when breeding; consequently, datasets are generally sparse, and a lack of incidental records does not necessarily mean the species is absent. Cryptic wader species usually nest as independent pairs, though some may nest semi-colonially (Marchant and Higgins, 1990). Nesting generally occurs seasonally. They may be sedentary, nomadic, migratory or partially migratory (Garnett et al., 2015; Marchant and Higgins, 1990). Movements between sites are likely to be partly dependent on the availability of suitable wetland habitats between origin and destination sites for shelter and feeding. Species in this group usually have an invertivorous or omnivorous diet and use foraging methods such as walking, stalking, striking and probing to catch their prey (Barker and Vestjens, 1989). Cryptic waders generally prefer shallow water or damp sediment with medium to high-density vegetation (Garnett et al., 2015). For nesting, some species require deeper water environments with dense vegetation, while others require very shallow water or recently dried wetland environments (Marchant and Higgins, 1990). Changes in water depth, water extent, water duration or food availability may result in nest exposure to predation or reduced food availability, resulting in nest failure (McGinness, 2016). For the purpose of this Assessment, the Endangered Australian painted snipe (Rostratula australis) is a representative species for the cryptic waders group and is rarely seen throughout its range (Rogers et al., 2004). It is a shy species that spends most of its time hidden in vegetation or woody debris in shallow-water areas. The population is small and has declined significantly across much of its range, most likely due to loss and degradation of inland floodplain wetland habitats and in particularly breeding habitats (Rogers et al., 2004). Figure 3-25 Dense aquatic and semi-aquatic vegetation used as habitat by cryptic wading waterbirds This habitat provides protection from predators and weather. Photo attribution: CSIRO Photo aquatic habitat. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Table 3-11 Species in the cryptic wading waterbird group, and their national and international conservation status (LC = Least concern) SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME POPULATION TYPE IUCN STATUS EPBC ACT STATUS Australian little bittern Ixobrychus dubius Ardeidae Australian LC _ Australian painted snipe Rostratula australis Rostratulidae Endemic Endangered Endangered Australian spotted crake Porzana fluminea Rallidae Endemic LC _ Baillon’s crake Zapornia pusilla Rallidae Australian LC _ Black bittern Ixobrychus flavicollis Ardeidae Australian LC _ Buff-banded rail Hypotaenidia philippensis Rallidae Australian LC _ Chestnut rail Eulabeornis castaneoventris Rallidae Australian LC _ Latham’s snipe Gallinago hardwickii Scolopacidae Non-breeding Near Threatened Vulnerable Lewin’s rail Lewinia pectoralis Rallidae Australian LC _ Red-necked crake Rallina tricolor Rallidae Australian LC _ Spotless crake Zapornia tabuensis Rallidae Australian LC _ Striated heron Butorides striata Ardeidae Australian LC _ White-browed crake Amaurornis cinerea Rallidae Australian LC _ Cryptic wading waterbirds in the Southern Gulf catchments Cryptic waders are found throughout the Southern Gulf catchments (see Figure 3-26 and a modelled potential species distribution for the Australian painted snipe Figure 3-27). Many species have been recorded at Lake Moondarra near Mount Isa, a permanent lake with fringing wetlands. Australian painted snipe, Australian spotted crake (Porzana fluminea), Latham's snipe (Gallinago hardwickii), and to a lesser extent, the black bittern (Ixobrychus flavicollis) have been recorded in this area (Atlas of Living Australia, 2023a; 2023b). Black bittern has also been recorded at Lawn Hill Gorge (Department of Agriculture‚ Water and the Environment, 2021a). Striated heron (Butorides striatus), chestnut rail (Eulabeornis castaneoventris) and black bittern have also been recorded at the coastal wetland aggregations (Atlas of Living Australia, 2023a; 2023b). Figure 3-26 Location of selected cryptic wading waterbirds in the Southern Gulf catchments Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-27 Modelled potential species distribution for Australian painted snipe (Rostratula australis) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for cryptic wading waterbirds Waterbird species in the cryptic waders group are sensitive to changes in the depth, extent and duration of shallow wetland environments and the fringes of deeper-water habitats such as waterholes (Kingsford and Norman, 2002; Marchant and Higgins, 1990; McGinness, 2016). Most species nest on the ground or in low vegetation, so nests are at risk when water levels change (Garnett et al., 2015; Marchant and Higgins, 1990). Cryptic waders are particularly sensitive to changes in the type, density or extent of emergent aquatic and semi-aquatic vegetation in and around these habitats. Besides changing foraging, nesting and refuge habitat, such changes can also reduce water quality and food availability and increase rates of competition, predation and disease (McGinness, 2016). Such changes can occur when water is directly extracted from these habitats or when the time between inundation events that create these habitats is extended (Brandis et al., 2009; Kingsford and Norman, 2002). Climate change and climate change−driven extremes are likely to interact with changes induced by water resource development, including inundation of freshwater habitats by seawater and inundation of nests by extreme flood events or seawater intrusion. The ecological functions that support cryptic wading waterbirds, and their associated flow requirements, are summarised in Table 3-12. Table 3-12 Ecological functions supporting cryptic wading waterbirds and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for cryptic wading waterbirds Few data are available for cryptic waders, but habitat requirements can be used as surrogates to assess vulnerability and pathways to change. The cryptic wader group’s need for appropriate vegetation and shallow-water environments makes them sensitive to changes in both water regimes and vegetation throughout their life cycles (Marchant and Higgins, 1990). Thus, the primary pathways of potential water resource development impacts on cryptic waders are habitat loss, fragmentation and change through changes in the timing, extent, depth and duration of inundation, which in turn change vegetation (Kingsford and Norman, 2002; McGinness, 2016; McKilligan, 2005) (Figure 3-28). In addition to direct disturbance from changes in hydrology and vegetation, species are also at risk from increased disturbance from human activities and predation (Kingsford and Norman, 2002). Human disturbance can be equivalent to habitat loss or degradation because it may lead waterbirds to avoid or underuse areas. During the breeding season, disturbance and predation may influence nest incubation and chick rearing, affecting overall nest success and eventual recruitment, which then affects population sizes and trajectories (McGinness, 2016). Changes in water regimes and vegetation can change predation pressure through increased exposure of cryptic waders and their nests (Sovada et al., 2001). Increased predation due to such changes can reduce the survival of cryptic waders and consequently population size either directly or indirectly by causing adults to desert their nests or foraging sites. Predators can also affect population size by competing for habitat or food, or affecting other predators and prey (Cruz et al., 2013; MacDonald and Bolton, 2008; Skorka et al., 2014). Figure 3-28 Conceptual model showing the relationship between threats, drivers, effects and outcomes for cryptic wading waterbirds in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.2.3 Shorebirds Description and background to ecology The shorebirds group consists of waterbirds with a high level of dependence on end-of-system flows and large inland flood events that provide broad areas of shallow water and mudflat environments. Flood events trigger production of significant food resources for these species – resources that are critical for fuelling long-distance migrations. Shorebirds generally eat fish or invertebrates. Most species walk and wade when foraging, probing sediment, rocks or vegetation for prey (Garnett et al., 2015; Marchant and Higgins, 1990). Shorebirds are largely migratory, mostly breeding in the northern hemisphere. They are in significant decline and are of international concern. Shorebirds depend on specific shallow-water habitats in distinct geographic areas, including northern hemisphere breeding grounds, southern hemisphere non-breeding grounds and stopover sites along migration routes such as the East Asian-Australasian Flyway (Bamford, 1992; Hansen et al., 2016). As the group is of international concern, various management and conservation strategies have been implemented (DAWE, 2021), including bilateral migratory bird agreements with China (CAMBA), Japan (JAMBA), and Korea (ROKAMBA), the Bonn Convention on the Conservation of Migratory Species of Wild Animals (Bonn), and the Ramsar Convention on Wetlands of International Importance. In northern Australia, this group comprises approximately 55 species from four families, including sandpipers, godwits, curlew, stints, plovers, dotterel, lapwings and pratincoles (Table 3-13). Approximately 35 species are common, regular visitors or residents. Several species in this group For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au are endangered globally and nationally, including the bar-tailed godwit (Limosa lapponica), curlew sandpiper (Calidris ferruginea), eastern curlew, great knot (Calidris tenuirostris), lesser sand plover (Charadrius mongolus) and red knot (Calidris canutus). The eastern curlew is listed as Critically Endangered under the EPBC Act and recognised through multiple international agreements as requiring habitat protection in Australia. Eastern curlews rely on food sources along shorelines, mudflats and rocky inlets, as well as roosting vegetation. Developments and disturbances, such as recreational, residential and industrial use of these habitats, have restricted habitat and food availability for the eastern curlew, contributing to population declines. The red-capped plover (Charadrius ruficapillus; Figure 3-29) is a shorebird that breeds in Australia rather than in the northern hemisphere. It is a small species that is widespread and common both inland and along the coast. It prefers open flat sediment areas such as mudflats and beaches for foraging and eats a range of small invertebrates including crustaceans. It breeds in response to flooding or rain inland, and seasonally on the coasts. Figure 3-29 Red-capped plover walking along a shore The red-capped plover is a member of the shorebirds group. Photo attribution: CSIRO Photo red-caped plover. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Table 3-13 Species in the shorebirds group, and their national and international conservation status (LC = Least concern) SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME POPULATION TYPE IUCN STATUS EPBC ACT STATUS Australian pratincole Stiltia isabella Glareolidae Australian LC _ Beach stone-curlew Esacus magnirostris Burhinidae Australian Near Threatened _ Masked lapwing Vanellus miles Charadriidae Australian LC _ Red-capped plover Charadrius ruficapillus Charadriidae Australian LC _ Black-fronted dotterel Elseyornis melanops Charadriidae Endemic LC _ Inland dotterel Charadrius australis Charadriidae Endemic LC _ Red-kneed dotterel Erythrogonys cinctus Charadriidae Endemic LC _ Banded lapwing Vanellus tricolor Charadriidae Endemic LC _ Bar-tailed godwit Limosa lapponica Scolopacidae Non-breeding migrant LC Two sub- species Endangered Black-tailed godwit Limosa limosa Scolopacidae Non-breeding migrant Near Threatened Endangered Broad-billed sandpiper Limicola falcinellus Scolopacidae Non-breeding migrant LC _ Common greenshank Tringa nebularia Scolopacidae Non-breeding migrant LC Endangered Common sandpiper Actitis hypoleucos Scolopacidae Non-breeding migrant LC _ Curlew sandpiper Calidris ferruginea Scolopacidae Non-breeding migrant Near Threatened Critically Endangered Eastern curlew (far eastern curlew) Numenius madagascariensis Scolopacidae Non-breeding migrant Endangered Critically Endangered Great knot Calidris tenuirostris Scolopacidae Non-breeding migrant Endangered Vulnerable Greater sand plover, Large sand plover Charadrius leschenaultii Charadriidae Non-breeding migrant LC _ Grey plover Pluvialis squatarola Charadriidae Non-breeding migrant LC Vulnerable Grey-tailed tattler Tringa brevipes Scolopacidae Non-breeding migrant Near Threatened _ Lesser sand plover Charadrius mongolus Charadriidae Non-breeding migrant Endangered Endangered Little curlew Numenius minutus Scolopacidae Non-breeding migrant LC _ Long-toed stint Calidris subminuta Scolopacidae Non-breeding migrant LC _ Marsh sandpiper Tringa stagnatilis Scolopacidae Non-breeding migrant LC _ Oriental plover, Oriental dotterel Charadrius veredus Charadriidae Non-breeding migrant LC _ Oriental pratincole Glareola maldivarum Glareolidae Non-breeding migrant LC _ Pacific golden plover Pluvialis fulva Charadriidae Non-breeding migrant LC _ Red knot Calidris canutus Scolopacidae Non-breeding migrant Near Threatened Vulnerable Red-necked stint Calidris ruficollis Scolopacidae Non-breeding migrant Near Threatened _ Ruddy turnstone Arenaria interpres Scolopacidae Non-breeding migrant LC Vulnerable Sanderling Calidris alba Scolopacidae Non-breeding migrant LC _ SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME POPULATION TYPE IUCN STATUS EPBC ACT STATUS Sharp-tailed sandpiper Calidris acuminata Scolopacidae Non-breeding migrant Vulnerable Vulnerable Swinhoe’s snipe Gallinago megala Scolopacidae Non-breeding migrant LC _ Terek sandpiper Xenus cinereus Scolopacidae Non-breeding migrant LC Vulnerable Whimbrel Numenius phaeopus Scolopacidae Non-breeding migrant LC _ Wood sandpiper Tringa glareola Scolopacidae Non-breeding migrant LC _ Asian dowitcher Limnodromus semipalmatus Scolopacidae Non-breeding migrant Near Threatened Vulnerable Common redshank, Redshank Tringa totanus Scolopacidae Non-breeding migrant LC _ Double-banded plover Charadrius bicinctus Charadriidae Non-breeding migrant LC _ Pectoral sandpiper Calidris melanotos Scolopacidae Non-breeding migrant LC _ Wandering tattler Tringa incana Scolopacidae Non-breeding migrant LC _ Little ringed plover Charadrius dubius Charadriidae Non-breeding migrant LC _ Pin-tailed snipe Gallinago stenura Scolopacidae Non-breeding migrant LC _ Red-necked phalarope Phalaropus lobatus Scolopacidae Non-breeding migrant LC _ Ruff (reeve) Calidris pugnax Scolopacidae Non-breeding migrant LC _ Baird’s sandpiper Calidris bairdii Scolopacidae Vagrant LC _ Caspian plover Charadrius asiaticus Charadriidae Vagrant LC _ Green sandpiper Tringa ochropus Scolopacidae Vagrant LC _ Buff-breasted sandpiper Tryngites subruficollis Scolopacidae Vagrant Near Threatened _ Dunlin Calidris alpina Scolopacidae Vagrant LC _ Grey (red) phalarope Phalaropus fulicaria Scolopacidae Vagrant LC _ Lesser yellowlegs Tringa flavipes Scolopacidae Vagrant LC _ Little stint Calidris minuta Scolopacidae Vagrant LC _ Common ringed plover Charadrius hiaticula Charadriidae Vagrant LC _ Spotted redshank Tringa erythropus Scolopacidae Vagrant LC _ White-rumped sandpiper Calidris fuscicollis Scolopacidae Vagrant LC _ Shorebirds in the Southern Gulf catchments Shorebirds are found throughout the Southern Gulf catchments (Figure 3-30, Figure 3-31, Figure 3-32, Figure 3-33, Figure 3-34, Figure 3-35). The most common species are Australian pratincole (Stiltia isabella), black-fronted dotterel (Elseyornis melanops), masked lapwing (Vanellus miles), red-kneed dotterel (Erythrogonys cinctus) and sharp-tailed sandpiper (Calidris acuminata) (Atlas of Living Australia, 2023a; 2023b). Figure 3-30 Observed locations of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Actitis hypoleucos (common sandpiper) to Calidris ruficollis (red-necked stint) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-31 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Calidris subminuta (long-toed stint) to Esacus magnirostris (beach stone-curlew) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-32 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Gallinago megala (Swinhoe’s snipe) to Peltohyas australis (inland dotterel) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-33 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Phalaropus lobatus (red-necked phalarope) to Vanellus miles (masked lapwing) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-34 Location of shorebirds in the Southern Gulf catchments in alphabetic order of species name: Vanellus tricolor (banded lapwing) to Xenus cinereus (Terek sandpiper) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-35 Modelled potential species distribution for eastern curlew (Numenius madagascariensis) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for shorebirds Waterbird species in the shorebirds group are sensitive to changes in the depth, extent and duration of inundation of open, very shallow water environments, including the edges of inland floodplains and lakes and estuarine and coastal mudflats and sandflats (Albanese and Davis, 2015; Donnelly et al.; Fernandez and Lank, 2008; Ge et al., 2009; Jackson et al., 2019; Schaffer-Smith et al., 2017). Their preference for open flat areas and good visibility when foraging means that encroachment of dense vegetation or human activity can prevent their use of a site (Baudains and Lloyd, 2007; Ge et al., 2009; Tarr et al., 2010). These species require abundant and spatially dense food, the latter being dependent on good water quality, high productivity of freshly inundated floodplain areas, and end-of-system flows to estuaries and coasts (Saint-Beat et al., 2013; Taft and Haig, 2005; 2006; Tjorve et al., 2008). The ecological functions that support shorebirds, and their associated flow requirements, are summarised in Table 3-14. Table 3-14 Ecological functions supporting shorebirds and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for shorebirds Shorebirds use habitats such as mudflats, sandflats, coastal bays or inlets to recover from migration flights (Atkinson, 2003; Jackson et al., 2019). Quality sites are able to support large numbers of shorebirds by providing abundant food, minimal human disturbance and shelter to rest (Goodenough et al., 2017; Pfister et al., 1992). Throughout the non-breeding season, shorebirds must increase their food intake to fuel their migration back to northern breeding sites (Goodenough et al., 2017). They require undisturbed and productive feeding areas to ensure minimal energy expenditure (Anderson et al., 2019). They rely on the inundation of shallow flat areas such as mudflats and sandflats to provide invertebrates and other food sources (Aharon- Rotman et al., 2017; Galbraith et al., 2002). Without inundation events, these habitats cannot support high densities of shorebird species, and lack of food can increase mortality rates both on- site and during and after migrations (Aharon-Rotman et al., 2017; Goss-Custard, 1977; Rushing et al., 2016). The primary pathways of potential water resource development impact on shorebirds include: habitat loss, fragmentation and change; toxins from pollution or contaminants; and disturbance from human activities (Aharon-Rotman et al., 2016). Habitat loss and disturbance from human activities is of particular concern for shorebird species worldwide. Shorebirds may waste time and energy responding to human disturbance, which may cause temporary loss of foraging habitats. The capacity to compensate by foraging for longer periods may vary between individuals and species (Glover et al., 2011; Pfister et al., 1992; Rogers et al., 2006; St Clair et al., 2010; Tarr et al., 2010; West et al., 2002). During the breeding season, human disturbance may also influence nest incubation and chick rearing, affecting overall nest success and eventual recruitment, which then affect population sizes and trajectories (McGinness, 2016). Climate change is also affecting habitat availability and quality among other factors for shorebirds, including changing freshwater inflows and the availability of mudflats and similar environments (Bellisario et al., 2014; Iwamura et al., 2013). The ecological outcomes of threatening processes on wetlands in the Southern Gulf catchments, and their implications for changes to biodiversity and ecosystem function, are illustrated in Figure 3-36. Figure 3-36 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the shorebirds group in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.2.4 Swimming, grazing and diving waterbirds Description and background to ecology The swimmers, grazers and divers group comprises species with a relatively high level of dependence on semi-open, open and deeper-water environments. These species commonly swim when foraging (including diving, filtering, dabbling, grazing) or when taking refuge. In northern Australia, this group comprises 49 species from 11 families, including ducks, geese, swans, grebes, pelicans, darters, cormorants, shags, swamphens, gulls, terns, noddies and jacanas (Table 3-15). This group can be further broken down into the subgroups: • diving swimmers – e.g. cormorants, pelicans, grebes • aerial divers – e.g. terns, gulls, noddies • grazing swimmers – e.g. swans, coots, swamphens, ducks, geese. These species breed in Australia and may be sedentary, nomadic, migratory or partially migratory. Nesting generally occurs seasonally, usually in dense vegetation such as emergent macrophytes, trees and shrubs (Garnett et al., 2015). Nests are usually independent or semi-colonial, and while breeding while is usually seasonal, it can be stimulated by flooding or large rainfall events (Kingsford and Norman, 2002). Species diets may be piscivorous, omnivorous or herbivorous (Barker and Vestjens, 1989). Changes in water depth, water extent or water duration can expose nests to predation or drowning, or reduce food availability, resulting in nest failure (McGinness, 2016; Poiani, 2006). For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au The magpie goose (Anseranas semipalmata; Figure 3-37) is one example of the swimmers, grazers and divers group, and while it is an iconic species in northern Australia, it is also the source of some conflict with humans when resources are limited (Corriveau et al., 2022; Frith and Davies, 1961; Traill et al., 2010). The magpie goose is an ancient and unique species of particular importance to Indigenous Peoples, providing eggs, meat and feathers. This species feeds on aquatic vegetation and often nests colonially (Marchant and Higgins, 1990). While currently abundant in northern Australia, wild magpie goose populations have largely disappeared from southern Australia due to human-driven change such as habitat destruction and hunting (Nye et al., 2007), and climate change is likely to exacerbate the impacts of such changes on magpie gees in northern Australia (Poiani, 2006; Traill et al., 2009a). Figure 3-37 Magpie goose perched on a fallen tree branch Magpie geese are a representative species of the swimmers, grazers and divers group. Photo attribution: CSIRO Photo of magpie goose. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Table 3-15 Species in the swimming, grazing and diving waterbirds group, and their national and international conservation status (LC = Least concern) SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME POPULATION TYPE IUCN STATUS EPBC ACT STATUS Australian (Australasian) shoveler Spatula rhynchotis Anatidae Australian LC _ Australian wood duck (maned duck) Chenonetta jubata Anatidae Endemic LC _ Spotted whistling-duck Dendrocygna guttata Anatidae Australian LC _ Garganey, Garganey teal Spatula querquedula Anatidae Non-breeding migrant LC _ Freckled duck Stictonetta naevosa Anatidae Endemic LC _ Chestnut teal Anas castanea Anatidae Endemic LC _ Grey teal Anas gracilis Anatidae Australian LC _ Pacific black duck Anas superciliosa Anatidae Australian LC _ Hardhead Aythya australis Anatidae Australian LC _ Black swan Cygnus atratus Anatidae Endemic LC _ Wandering whistling-duck Dendrocygna arcuata Anatidae Australian LC _ Plumed whistling-duck Dendrocygna eytoni Anatidae Australian LC _ Pink-eared duck Malacorhynchus membranaceus Anatidae Endemic LC _ Cotton pygmy-goose Nettapus coromandelianus Anatidae Australian LC _ Green pygmy-goose Nettapus pulchellus Anatidae Australian LC _ Blue-billed duck Oxyura australis Anatidae Endemic LC Radjah shelduck Radjah radjah Anatidae Australian LC _ Australian shelduck Tadorna tadornoides Anatidae Endemic LC _ Australasian darter Anhinga novaehollandiae Anhingidae Australian LC _ Magpie goose Anseranas semipalmata Anseranatidae Australian LC _ Comb-crested jacana Irediparra gallinacean Jacanidae Australian LC _ Common noddy Anous stolidus Laridae Australian LC _ Whiskered tern Chlidonias hybrida Laridae Australian LC _ White-winged black tern Chlidonias leucopterus Laridae Non-breeding migrant LC _ Silver gull Chroicocephalus novaehollandiae Laridae Australian LC _ SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME POPULATION TYPE IUCN STATUS EPBC ACT STATUS Australian gull-billed tern Gelochelidon macrotarsa Laridae Endemic (breeding only) LC _ Common gull-billed tern Gelochelidon nilotica Laridae Non-breeding migrant LC _ Caspian tern Hydroprogne caspia Laridae Australian LC _ Bridled tern Onychoprion anaethetus Laridae Australian LC _ Sooty tern Onychoprion fuscatus Laridae Australian LC _ Roseate tern Sterna dougallii Laridae Australian LC _ Common tern Sterna hirundo Laridae Non-breeding migrant LC _ Black-naped tern Sterna sumatrana Laridae Australian LC _ Little tern Sternula albifrons Laridae Australian LC _ Lesser crested tern Thalasseus bengalensis Laridae Australian LC _ Crested tern Thalasseus bergii Laridae Australian LC _ Australian pelican Pelecanus conspicillatus Pelecanidae Endemic (breeding only) LC _ Little pied cormorant Microcarbo melanoleucos Phalacrocoracidae Australian LC _ Great cormorant Phalacrocorax carbo Phalacrocoracidae Australian LC _ Little black cormorant Phalacrocorax sulcirostris Phalacrocoracidae Australian LC _ Pied cormorant Phalacrocorax varius Phalacrocoracidae Australian LC _ Great crested grebe Podiceps cristatus Podicipedidae Australian LC _ Hoary-headed grebe Poliocephalus poliocephalus Podicipedidae Endemic LC _ Australasian grebe Tachybaptus novaehollandiae Podicipedidae Australian LC _ Pale-vented bush-hen, Bush-hen Amaurornis moluccana Rallidae Australian LC _ Eurasian coot Fulica atra Rallidae Australian LC _ Dusky moorhen Gallinula tenebrosa Rallidae Australian LC _ Purple swamphen Porphyrio porphyrio Rallidae Australian LC _ Black-tailed native-hen Tribonyx ventralis Rallidae Endemic LC _ Swimmers, grazers and divers in the Southern Gulf catchments Swimmers, divers and grazers are found throughout the Southern Gulf catchments (see Figure 3-38, Figure 3-39, Figure 3-40, Figure 3-41 and a modelled potential species distribution for magpie geese Figure 3-42). The most common species are Australasian darter (Anhinga novaehollandiae), Australasian grebe (Tachybaptus novaehollandiae), Australian pelican (Pelecanus conspicillatus), grey duck, grey teal (Anas gracilis), hardhead (Aythya australis), little black cormorant (Phalacrocorax sulcirostris) and little pied cormorant (Microcarbo melanoleucos) (Atlas of Living Australia, 2023a; 2023b). Figure 3-38 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Amaurornis moluccana (pale-vented bush-hen) to Chenonetta jubata (Australian wood duck) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-39 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Cygnus atratus (black swan) to Nettapus coromandelianus (cotton pygmy-goose) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-40 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Nettapus pulchellus (green pygmy-goose) to Porphyrio porphyrio (purple swamphen) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-41 Locations of swimming, grazing and diving waterbirds in the Southern Gulf catchments in alphabetic order of species name: Spatula rhynchotis (Australian shoveler) to Tribonyx ventralis bBlack-tailed native-hen) Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-42 Modelled potential species distribution for magpie goose (Anseranas semipalmata) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for swimmer, grazer and diver waterbirds Species in the swimming, grazing and diving waterbirds group are sensitive to changes in the depth, extent and duration of perennial semi-open and open deeper-water environments such as waterholes and wetlands (Table 3-16) (Marchant and Higgins, 1990; McGinness, 2016). They can also be sensitive to changes in the type, density or extent of the fringing aquatic or semi-aquatic vegetation in and around these habitats. Besides changing foraging, nesting and refuge habitat, such changes can also reduce water quality and food availability and increase rates of competition, predation and disease (Douglas et al., 2005; McGinness, 2016). Such changes can occur when water is extracted directly from these habitats or when the time between connecting flows or rainfall events that fill these habitats is extended (Kingsford and Norman, 2002). Climate change and extremes are likely to interact with changes induced by water resource development, including inundation of freshwater habitats by seawater and inundation of nests by extreme flood events or seawater intrusion (Nye et al., 2007; Poiani, 2006; Traill et al., 2009a; Traill et al., 2009b). The ecological functions that support swimming, grazing and diving waterbirds, and their associated flow requirements, are summarised in Table 3-16. Table 3-16 Ecological functions supporting swimming, grazing and diving waterbirds and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for swimming, grazing and diving waterbirds The primary pathways of potential water resource development impact on waterbirds in the swimmers, grazers and divers group include: (i) habitat loss, fragmentation and change, including water depth changes and weed invasion changing habitats, (ii) climate change and extremes – including inundation of freshwater habitats by seawater when river flows are reduced and inundation of nests by extreme flood events, (iii) toxins from pollution or contaminants, (iv) disturbance and hunting from human activities, (v) predation by invasive or feral animals, and (vi) changes in disease or parasite risk or burdens (Bayliss, 1989; Corbett and Hertog, 1996; Douglas et al., 2005; Morton, 1990; Nye et al., 2007; Poiani, 2006; Traill et al., 2010; Traill et al., 2009a; Traill et al., 2009b) (Figure 3-43). Reduced extent, depth and duration of inundation of waterhole and other deep-water environments are likely to reduce habitat and food availability for this group, increasing competition and predation and also increasing risk of disease and parasite spread. Conversely, species in this group that nest at water level or just above, such as magpie geese, are particularly at risk of nests drowning when water depths increase unexpectedly (Douglas et al., 2005; Poiani, 2006; Traill et al., 2010; Traill et al., 2009a; Traill et al., 2009b). Figure 3-43 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the swimming, grazing and diving waterbirds group in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au 3.3 Turtles, prawns and other species 3.3.1 Banana prawns (Penaeus merguiensis) Description and background ecology Banana prawns are large-bodied decapod crustaceans around 80 g in weight of the family Penaeidae that are found throughout the Indo-West Pacific. They are a prized fishery target species throughout their geographic distribution. Two species of banana prawns are found in Australia: the common banana prawn (Penaeus merguiensis) and the redleg banana prawn (P. indicus). Both banana prawn species are globally widespread throughout the Indian Ocean and south-east Asian and west Pacific coastal habitats. In Australia, common banana prawns inhabit tropical and subtropical coastal waters (Grey et al., 1983). In contrast, the Joseph Bonaparte Gulf and western Tiwi Island region in north-west Australia are the south-eastern limit of the worldwide distribution of redleg banana prawns (Grey et al., 1983). Common banana prawns are abundant in the western Gulf of Carpentaria, with significant commercial catches taken adjacent to their inshore estuarine habitats (Staples et al., 1985). Banana prawns support an approximate 4942 t ‘sub-fishery’ component (based on mean take over a recent 10-year period of the NPF that is worth about $70 to $80 million annually (Laird, 2021). The major portion of the common banana prawn catch is taken in the eastern Gulf of Carpentaria; however, significant catches are taken east and west of Mornington Island within the Southern Gulf catchments marine region (Laird, 2021). The influence of rainfall and runoff from south- western Gulf of Carpentaria catchments on banana prawn catches is less clear than for eastern catchments and requires further investigation, though seasonal rainfall and prevailing winds are positively correlated with catch (Vance et al., 2003; Vance et al., 1985). Post-larval banana prawns settle in the mudbank and mangrove forest matrix in the upper reaches of estuarine tributaries (Kenyon et al., 2004; Vance et al., 1996a; 2002). They occupy mangrove forest habitats (see Section 3.4.4) and are forced from the mangroves on each ebb tide, to return on the next flood tide (Vance et al., 2002). Mangrove prop roots and trunks are critical to juvenile banana prawn survival; they provide shelter and refuge from predation (Meager et al., 2005). The substrates within the forest and on the intertidal banks support microflora and meiofauna (algae, molluscs, crustaceans and annelid worms), which they consume on each tide (Burford et al., 2012; Duggan et al., 2014; Vance et al., 2002; Wassenberg and Hill, 1993). Juvenile common banana prawns emigrate from river estuaries cued by wet-season freshwater river flows (Vance et al., 1998). The larger the flow volume, the greater the emigration pulse and the smaller size prawns that emigrate. The tolerance of juvenile common banana prawns to brackish water declines as prawns grow (Dall, 1981). As the salinity of the estuary declines due to flood flows, fewer large juvenile prawns can tolerate the low salinity waters to reside there. Juvenile prawns with a carapace length (CL) greater than 12 mm emigrate when salinity is about 30 to 35 ppt, while prawns less than or equal to 8 mm CL emigrate when salinity drops to about 5 ppt or lower, particularly when the decline was abrupt (Staples and Vance, 1986). Emigrants move offshore to reside on muddy sediments in deeper waters (Staples, 1980b; Staples and Vance, 1986; Vance et al., 1998). Using commercial catch as a measure of population abundance, large flood flows cue the prolific population of juvenile banana prawns to emigrate en masse to the near-shore and offshore zones where they rely on marine habitats for enhanced growth and survival (Broadley et al., 2020; Duggan et al., 2019; Lucas et al., 1979). Adult banana prawn distribution is adjacent to their juvenile estuarine mangrove habitats (Staples et al., 1985; Zhou et al., 2015). Adult common banana prawns occupy soft-sediment substrates in relatively shallow waters within the south-west, south-east and eastern Gulf of Carpentaria, and along the Top End / Arnhem Land coastline. Banana prawns are managed by limited effort (licence to fish) and by spatial and temporal closures. The fishing season opens on 1 April annually and continues until catch rates decline to a trigger level defined in the Northern Prawn Fishery Harvest Strategy (AFMA, 2022). Common banana prawns grow to about 55 mm CL for females (50 mm CL ≈ 85 g) and about 47 mm CL for males (40 mm CL ≈ 50 g). Banana prawns in the Southern Gulf catchments marine region Adult common banana prawns live and spawn offshore in waters 10 to 30 m deep, the larvae and post-larvae advect (i.e. are transported by water movement) inshore to settle in the mangrove forest and mudbank matrix in estuarine mangrove habitats (Crocos and Kerr, 1983; Staples, 1980a; Vance et al., 1998). Each of the major rivers along the south-west Gulf of Carpentaria coastline, including those within the Southern Gulf catchment marine region, supports abundant populations of juvenile banana prawns (Staples, 1979). Commercial catches of common banana prawns are taken to the east and west of Mornington Island, offshore from the rivers of the Southern Gulf catchments (Figure 3-44). The Southern Gulf catchments marine region spans the ‘Mornington’ (16-year mean catch: 265 t; 5.8% of total common banana prawn catch) and ‘Sweers’ (16-year mean catch: 273 t; 6%) reporting regions for the NPF (Laird, 2021). The highest catch rates are taken inshore in waters about 15 to 20 m deep, close to the coastal estuaries of the Leichhardt, Albert, Nicholson and other rivers of the region. The mangrove forest and creek mudbank habitats of juvenile common banana prawns are located within the estuaries of the adjacent rivers. In the 1970s, the use of mangrove habitats by juvenile banana prawns within the Albert River, Massacre Inlet and the Calvert River estuaries was documented as part of a survey across river estuaries in the Gulf of Carpentaria (Staples, 1979). From 1970 to 1973, these samples were taken by float plane in main river channels; the banana prawn catches were seasonal and varied in abundance, but showed the estuaries were inhabited by juvenile banana prawns. However, the remoteness of the river systems in the south-western Gulf of Carpentaria renders them poorly studied. More comprehensively, the Albert River estuarine habitats were sampled for banana prawns over the ‘juvenile recruitment season’ from September 1978 to March 1979 (Staples and Vance, 1987). They confirmed the strong recruitment of immigrating planktonic post-larvae, subsequent benthic settlement, juvenile growth and survival, and prawn emigration (to the near shore) within the mangrove-lined estuary. These samples were taken by vessel either fortnightly or monthly over the 7-month period and represent a comprehensive quantification of several life-history stages of banana prawns within the Albert River estuary (Staples and Vance, 1987). Figure 3-44 Fisheries catch of banana prawns the Southern Gulf catchments marine region Juvenile common banana prawns inhabit tropical river estuaries in the marine / brackish zone, particularly the mangrove forest / tributary creek matrix within river estuaries. Adult common banana prawns are caught offshore in water about 10 to 30 m deep in the marine habitat, adjacent to their juvenile habitats. Units are kilograms as total logbook catches for the period 2011 to 2020. Data sources: Kenyon et al. (2022) and Staples and Vance (1987) Flow–ecology relationships for banana prawns The life-history strategy of banana prawns renders them critically dependent on the natural flow regime in the Australian wet-dry tropics. Adult prawns spawn at sea, and pelagic eggs and larvae occupy the marine habitat, before post-larvae use currents to move shoreward to river estuaries (Vance and Rothlisberg, 2020). Prior to the annual wet season, post-larvae settle to benthic habitats in the estuarine mangrove forest and mudbank matrix, particularly upper tributary mangrove forests (at high tide) and creeks (Vance et al., 2002; Vance et al., 1990). They shelter and grow within the estuary, and a brackish ecotone supports lower mortality and faster growth (Staples and Heales, 1991; Vance et al., 1998; Wang and Haywood, 1999). Predation by fish within the estuary is high, and a significant proportion of the estuarine population is lost (Wang and Haywood, 1999). Floodwaters cue juvenile banana prawns to emigrate. The larger the flood the greater the emigration event, and the lower the estuarine salinity the smaller the prawns that emigrate (Staples and Vance, 1986) (Table 3-17). Emigrant juveniles and sub-adults move to the near-shore zone (Staples, 1980b) and probably benefit from nutrient deposition within the flood plume (Burford et al., 2012; Burford and Faggotter, 2021). In addition, mortality is lower in marine habitats than in the estuary (Gwyther, 1982). The overall result is a larger adult population of banana prawns in coastal marine habitats when the flood flows from adjacent estuaries are higher (Duggan et al., 2019). High-level pulsed flood flows during the monsoon season, low-level early-season flows, sustained flows during the wet season and persistent wet-season flows all have important effects on the estuarine population of both species of banana prawns. During the September to December recruitment window for juvenile prawns, estuaries within the Gulf of Carpentaria ecosystems are stressed, and habitats are often hypersaline during the latter part of the dry season (Kenyon et al., 2004; Vance et al., 1990). The estuaries are a refuge habitat for many fish and crustaceans living under severe environmental conditions prior to the onset of the wet season, usually January to March (Babcock et al., 2019; Robins et al., 2020). Early low-level flows that might occur during November to December condition tropical estuaries to brackish, cooler habitats, more favourable to growth and survival of crustaceans and fish within them, including juvenile banana prawns (Leahy and Robins, 2021; Ruscoe et al., 2004; Staples and Heales, 1991) (Table 3-17). Once an abundant estuarine population of juvenile banana prawn is established, high-level flood flows cue emigration and result in a large prawn population offshore. Persistent flows in the latter portion of the wet season continue to facilitate both a brackish estuary to support the growth of small juveniles and emigration of the larger juvenile population (Duggan et al., 2014; Staples and Vance, 1986). The ecological functions and their supporting flow requirements for banana prawns are summarised in Table 3-17. Table 3-17 Ecological functions supporting banana prawns and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for banana prawns The life history of banana prawns would be significantly affected by interruptions to the natural flows of northern Australian rivers. Large flows during the wet season cue emigration of banana prawns from estuarine habitats to the near-shore zone and further offshore. The emigration cues triggered by the annual monsoon-season flow regime renders banana prawns particularly vulnerable to water resource development. During high-flow years (strong wet season), banana prawns emigrate en masse from the estuary and commercial catches of prawns (as an indicator of abundance) are high (Broadley et al., 2020; Plagányi et al., 2022). During low-flow years (drier wet season), a proportion of banana prawns remains within the estuary and is subject to predation and mortality. Therefore, maintaining estuarine brackish habitats, diversity of river flow regimes and high-pulse flood flows enhances the populations of banana prawns and inshore and offshore habitat connectivity. Water resource development has the capacity to reduce the population of banana prawns (Broadley et al., 2020; Plagányi et al., 2022). Extraction from, or impounding, low-level flows removes a large proportion of early-season low- level river flows, with subsequent impacts on estuarine banana prawns. Interrupting early-season low-level flows reduces the capacity of freshwater inputs to the estuary to create brackish habitats and may render an estuary continuously hypersaline. A hypersaline estuary is a stressful habitat for juvenile banana prawns during the annual recruitment window from September to January. Threshold levels of river flow as a trigger to water extraction can sustain the provision of flow, and hence the ecosystem services to the estuary, during this window of possible low-level flows before the onset of the bulk of wet-season precipitation during January to March (Plagányi et al., 2022). Significant extraction or impoundment of pulsed high-level flood flows from January to March reduces the emigration cue for juvenile banana prawns, reducing the proportion of the population reaching offshore habitats (Broadley et al., 2020; Plagányi et al., 2022). The impact of water resource development in three Gulf of Carpentaria rivers on coastal banana prawn populations has been modelled for the construction of dams and water harvest at several levels of extraction (Plagányi et al., 2023). The biomass and commercial catches of the common banana prawn were predicted to decrease by 4 to 40% depending on the extent of water extraction or impoundment from the Mitchell, Flinders and Gilbert rivers in the eastern and southern Gulf of Carpentaria. The risk to the banana prawn population was assessed as negligible for one of four water resource development scenarios, moderate for two and major for the remaining scenario. The risk to the commercial fishery for banana prawns mimicked that for their population (Plagányi et al., 2023). The model outputs included an explicit representation that the decline in the banana prawn population flowed on to detrimental effects on their predators. Plagányi et al. (2023) showed that both the construction of dams and the harvest of river flows via pumped water extraction affect aspects of banana prawn life history, reducing the resilience of their populations. Reduction in the volume and duration of high-level flows, and variability in the seasonality and volume of low-level flows, affect estuarine habitat suitability (brackish conditions preferred), growth, survival and emigration of banana prawns (Plagányi et al., 2023; Vance and Rothlisberg, 2020). The ecological outcomes of threatening processes on banana prawns in northern Australia, with their implications for changes to growth and mortality, community structure, habitat and population, are illustrated in Figure 3-45. Figure 3-45 Conceptual model showing the relationship between threats, drivers, effects and outcomes for banana prawns in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au 3.3.2 Endeavour prawns (Metapenaeus spp.) Description and background ecology Endeavour prawns are a medium-sized commercial decapod crustacean (40–60 g) from the family Penaeidae, a family that exhibits a larval life-history strategy (Dall et al., 1990) with inshore and offshore phases. Two species of endeavour prawns, blue endeavour prawns (Metapenaeus endeavouri) and red endeavour prawns (M. ensis), inhabit littoral and coastal ecosystems in tropical Australia. Blue endeavour prawns are endemic to Australia (Grey et al., 1983), while red endeavour prawns are found throughout the Indo-West Pacific, from Sri Lanka to Japan (Grey et al., 1983). In Australia, red endeavour prawns are restricted to tropical coastlines, while blue endeavour prawns are found in both tropical and subtropical latitudes. Adult endeavour prawns live and spawn offshore in waters 10 to 40 m deep; the larvae and post- larvae are transported inshore to settle in the littoral zone (Crocos et al., 2001; Gribble et al., 2007; Jackson and Rothlisberg, 1994; Staples et al., 1985). Blue endeavour prawns use seagrass habitats as juveniles, while red endeavour prawns are ubiquitous across a range of shallow habitats including bare substrates and mangrove-lined mudbanks (Dall et al., 1990; Staples et al., 1985). A continuous seagrass habitat thrives in the south-west Gulf of Carpentaria ecosystem from Blue Mud Bay to Mornington Island, and juvenile blue endeavour prawns are abundant within the habitat (Kenyon, 1999; Poiner et al., 1987). Seagrass habitats (see Section 3.4.6) are critical to juvenile endeavour prawn survival. They provide habitat structure for shelter and stability; they provide food directly to penaeid prawns and via micro- and meio-benthos (Loneragan et al., 1997; Wassenberg and Hill, 1987). Seagrasses provide a refuge from predation for penaeid prawns (Haywood et al., 1998). The degree of refuge within seagrass habitat depends on seagrass species and morphological type (Haywood et al., 1998; Kenyon et al., 1995). Juvenile endeavour prawns emigrate offshore at about 10 to 20 mm CL to an epibenthic existence in deeper waters (Coles and Lee Long, 1985; Watson and Turnbull, 1993). Adult endeavour prawns occupy soft-sediment substrates in relatively shallow waters (Buckworth, 1992; Kenyon, 2021). As large juveniles and adults, both species of endeavour prawns bury themselves in the sediment during the day and emerge to feed at night (when they are fished) (Park and Loneragan, 1999; Wassenberg and Hill, 1994). The diet of juvenile and adult endeavour prawns in the Gulf of Carpentaria has not been well studied; however, estuarine juveniles derive most the carbon in their diet from seagrass and seston (i.e. suspended particles, including living organisms) rather than from mangroves (Loneragan et al., 1997). Their diet likely is small bivalves, gastropods, ophiuroids, crustaceans and polychaete worms, which is similar to that of tiger prawns (Heales et al., 1996; Wassenberg and Hill, 1987). Within their seagrass habitats and offshore, both species of endeavour prawn are preyed upon by fish (sharks and teleosts), squid and cuttlefish (Brewer et al., 1991; Brewer et al., 1995). Endeavour prawns support a 500 t component (recent 10-year mean) of the commercial NPF (worth about $5 million annually), and the major portion of the endeavour prawn catch is taken in the western Gulf of Carpentaria (Laird, 2021; Savage and Hobsbawn, 2015). The endeavour prawn species group is managed via a stock assessment (Deng et al., 2021; Dichmont et al., 2010; Punt et al., 2011). Over the decades, management provisions for the NPF have been continually revised and improved. Currently, the fishery is managed by effort regulation (Deng et al., 2021; Deng et al., 2015). Blue endeavour prawns grow to about 45 mm CL (38 mm CL ≈ 40 g) for females and about 34 mm CL (34 mm CL ≈ 30 g) for males. Red endeavour prawns are larger: females grow to about 48 mm CL (43 mm CL ≈ 50 g) and males to about 36 mm CL (34 mm CL ≈ 30 g) (Kenyon, 2021). Endeavour prawns in the Southern Gulf catchments Two species of endeavour prawns are found in the Gulf of Carpentaria: Metapenaeus endeavouri (blue endeavour prawn) and M. ensis (red endeavour prawn). Both species of endeavour prawn are found in the marine region offshore from the Southern Gulf catchments (Laird, 2021). Adult endeavour prawns occupy soft-sediment substrates in relatively shallow waters within the north- west and south-west Gulf of Carpentaria (Buckworth, 1992; Kenyon, 2021). The two species have relatively allopatric distributions (i.e. geographically separated), which depend on sediment texture (Somers, 1987; Somers, 1994). The proportion of red endeavour prawns in the commercial catch is higher in the northern Gulf of Carpentaria, particularly north of Groote Eylandt (Kenyon, 2021). Both red and blue endeavour prawns are found in the south-western Gulf of Carpentaria in coastal waters about 10 to 45 m deep offshore from the rivers of the Southern Gulf catchments (Somers, 1994) (Figure 3-46 and Figure 3-47). Endeavour prawns are abundant to the west and east of Mornington Island, which lies within the Southern Gulf catchments marine region (Kenyon, 2021; Robertson et al., 1985; Somers, 1994). The Southern Gulf catchments marine region spans the ‘Mornington’ (16-year mean catch: 43 t; 10% of total endeavour prawn catch) and ‘Sweers’ (16-year mean catch: 29 t; 6.8%) reporting regions for the NPF (Laird, 2021). In the Southern Gulf catchments marine region, blue endeavour prawns dominate the species distribution on the fishing grounds to the east and west of Mornington Island (Kenyon, 2021). The seagrass habitats of the juvenile phase of both species of endeavour prawns form a near-continuous swathe of vegetation in the littoral zone along the coast to the west of Mornington Island, and along the east coast of Mornington Island in the coastal bays such as Charlie Bush Bay (Coles and Lee Long, 1985; Poiner et al., 1987). The juvenile endeavour prawn population within littoral habitats along the extensive Gulf of Carpentaria coastline to the north-west of Mornington Island has not been sampled. However, Coles and Lee Long (1985) verified the presence of a population of juvenile blue endeavour prawns in seagrass habitats on the east coast of Mornington Island. Figure 3-46 Fisheries catch of red endeavour prawns in the Southern Gulf catchments marine region Juvenile habitat for red endeavour prawns includes shallow littoral seagrass habitats in reasonably clear waters along coasts. Adult red endeavour prawns are caught offshore in water about 20 to 45 m deep in the marine habitat adjacent to their juvenile habitats. Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022) and Coles and Lee Long (1985) Figure 3-47 Fisheries catch of blue endeavour prawns in the Southern Gulf catchments marine region Juvenile habitat for blue endeavour prawns includes shallow littoral seagrass habitats in reasonably clear waters along coasts. Adult blue endeavour prawns are caught offshore in water about 20 to 40 m deep in the marine habitat adjacent to their juvenile habitats. Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022) and Coles and Lee Long (1985) Flow–ecology relationships for endeavour prawns The life-history strategy of endeavour prawns includes a juvenile stage dependent on littoral seagrass habitats (Coles and Lee Long, 1985; Dall et al., 1990). Endeavour prawn post-larvae use currents to move shoreward to shallow, subtidal habitats, especially seagrass communities (Condie et al., 1999; Jackson and Rothlisberg, 1994). Before the annual wet season, blue endeavour prawn post-larvae settle within the shallow, shoreward extent of the seagrass community (Coles and Lee Long, 1985; Staples et al., 1985). Red endeavour prawns settle in a range of estuarine and littoral habitats (Staples et al., 1985). Both species shelter, forage and grow within vegetated habitat where leaf structure reduces predation and promotes primary productivity and prawn growth (Haywood et al., 1998; Kenyon et al., 1995). Seagrass habitats are found within some estuaries, but along the Gulf of Carpentaria coast, seagrass habitats are mostly offset from estuaries and away from the direct influence of river flows (Poiner et al., 1987). Seagrass thrives in oligotrophic (i.e. nutrient poor) waters, while a sufficient level of nutrients in the environment supports seagrass growth. The primary productivity of the river and coastal shallow habitats of the Gulf of Carpentaria is nutrient limited (Burford et al., 2012; Burford and Faggotter, 2021). Floodwaters transport terrigenous nutrients (i.e. eroded from the land) from the catchment that are deposited within the flood plume and littoral zone adjacent to Gulf of Carpentaria rivers (Burford et al., 2012; Burford and Faggotter, 2021). Prolonged turbidity or direct smothering is a risk to seagrasses close to river mouths (Longstaff et al., 1999; Longstaff and Dennison, 1999). Longshore transport of a proportion of nutrients deposited adjacent to river mouths maintains a nutrient balance and would benefit the productivity of coastal seagrasses in the Gulf of Carpentaria, while not adversely affecting water quality within the seagrass habitat (Table 3-18). The hydrology within the inshore Gulf of Carpentaria may ensure that terrigenous-sourced nutrients are retained inshore in a poorly mixed layer of water (Burford and Faggotter, 2021). As a consequence, river flows are crucial in maintaining the balance of coastal productivity in the Gulf of Carpentaria littoral zone via low-level source nutrients from adjacent catchments. The growth and stability of the littoral seagrass communities in the Gulf of Carpentaria are enhanced by the riverine-source nutrient inputs to the coastal waters. Under current flow regimes, low sediment loads and limited effects of turbidity during floods do not smother coastal seagrasses (Table 3-18). Hence, endeavour prawn juveniles benefit from the natural flow regime of Gulf of Carpentaria rivers. The ecological functions and their supporting flow requirements for endeavour prawns are summarised in Table 3-18. Table 3-18 Ecological functions supporting endeavour prawns and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for endeavour prawns Coastal waters in Australia’s tropics, such as the Gulf of Carpentaria and upstream catchment water bodies, form oligotrophic though productive ecosystems, which are stressed by heat, high evaporation, hypersaline estuaries and lack of precipitation for 9 months of the year (Blondeau- Patissier et al., 2014; Robins et al., 2020). The monsoon season delivers environmental flux that stimulates the ecosystem and estuarine and marine communities (Blondeau-Patissier et al., 2014). Biota benefit from the annual dynamic freshwater pulse flows. Primary productivity (Ndehedehe et al., 2020a), fish growth (Leahy and Robins, 2021) and crustacean and fish populations (Plagányi et al., 2022) benefit from the maintenance of trend in historical flows within the ecosystem. Though not yet well understood, littoral seagrass communities within the Gulf of Carpentaria and their dependent fauna benefit from the dynamic provision of monsoon-driven inputs to the system on an annual basis (Plagányi et al., 2022). An established population of juvenile endeavour prawns within the coastal seagrass community benefits from high-level flood flows in previous years that supported nutrient dynamics within the littoral community. The ecological outcomes of threatening processes on endeavour prawns in northern Australia, with their implications for changes to growth and mortality, community structure, habitat and population, are illustrated in Figure 3-48. Figure 3-48 Conceptual model showing the relationship between threats, drivers, effects and outcomes for endeavour prawns in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.3.3 Freshwater turtles (family Chelidae) Description and background to ecology Freshwater turtles represent one of the world’s more imperilled taxonomic groups, with 52% of the global species facing extinction or being under threat (Böhm et al., 2013; Van Dijk, 2014). In Australia, freshwater turtles can be classified into three families: Chelidae (32 species), Trionychidae (two species) and Carettochelyidae (one species) (Georges and Thomson, 2010). Chelids, members of the Chelidae family, are highly aquatic species. They have webbed feet and can stayed submerged in water for extended periods. Chelids retract their necks sideways into their shells, and their dietary habits vary between genera. Long-necked species such as Chelodina spp. are primarily carnivorous, consuming fish, invertebrates and gastropods (Legler, 1982; Thomson, 2000). In contrast, short-necked species such as Elseya spp. are herbivorous or specialised fruits eaters (Kennett, 1993). Freshwater turtles depend upon flooded wetland systems for breeding, nesting, food provision and refuge. Some of their key threatening processes are changes to regional hydrology, habitat loss, degradation, and fragmentation; predation; and, in some cases, competition by invasive species; disease (Petrov et al., 2023) and alteration of environments predicted by climate change (Stanford et al., 2020). In northern Australia, turtles inhabit a diverse range of aquatic environments, including both river and floodplain wetland habitats such as the main channel, waterholes, floodplain wetlands and oxbow lakes (Cann and Sadlier, 2017; Thomson, 2000). Many turtle species in this region have developed adaptive traits to survive in the inter-annual variation between the wet and dry For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au seasons, such as the emergence of hatchlings with the wet season onset (Cann and Sadlier, 2017). During the dry season, freshwater turtle movements on and off the floodplain become restricted, making them more vulnerable to changes in water quality, invasive species and habitat degradation (Cann and Sadlier, 2017; Doupe et al., 2009). Australian freshwater turtles are of both ecological and cultural significance in Australia. For example, some Indigenous Peoples consume some species as a seasonal protein source (Jackson et al., 2012). Freshwater turtles are also represented in Indigenous rock art (Ferronato and Georges, 2023). Indigenous Peoples maintain profound connections to freshwater turtles through songlines and ceremonies, and certain people have roles as custodians and caretakers according to the kinship system. Knowledge holders described seasonal knowledge and indicators that related to freshwater turtle hunting, behaviour, diet and physiology, including aestivation (i.e. dry- season torpor), fatness and breeding cycles. For example, knowledge holders said the dry (cold) season is the time to hunt for northern snake-necked turtle (Chelodina oblonga; formerly Chelodina rugosa). Indigenous Peoples identify natural predators (including birds of prey e.g. eagles and hawks, crocodiles, goannas and dingoes), feral animals (e.g. pig, buffalo, horse, donkey, cattle and cane toad) and climate change (e.g. lower rainfall) as the main threats to the freshwater turtles (Russell et al., 2021). Freshwater turtles in the Southern Gulf catchments There are ten species of freshwater turtles described in the NT (Department of Environment Parks and Water Security, 2019b) of which four have been recorded in the Southern Gulf catchments: Cann’s snake-necked turtle (Chelodina canni), northern snake-necked turtle (Chelodina oblonga), Gulf snapping turtle (Elseya lavarackorum) and red-bellied short-necked turtle (Emydura subglobosa) (Atlas of Living Australia, 2021; Department of Environment Parks and Water Security, 2019a) (Figure 3-49). Records for the Southern Gulf catchments are sparse compared to many other regions of Australia. Currently all four species are listed as Least concern by the NT Government (Department of Environment Parks and Water Security, 2019a); however, the Gulf snapping turtle (Elseya lavarackorum) is listed nationally as Endangered under the EPBC Act. Freshwater turtle species native to northern Australia inhabit diverse environments, depending on the species. The northern snapping turtle is found in deep-water pools in upper catchments of permanently flowing spring-fed rivers, reaching their highest density in areas with good-quality riparian vegetation. Riverside pandanus (Pandanus aquaticus) provide important habitat. Submerged pandanus branches serve as resting sites for the northern snapping turtle, while above the water, dense stands of pandanus offer protection against predators. The northern snake-neck turtle is found in lakes, billabongs and swamps with substantial beds of water plants, and Cann’s snake-necked turtle is found in similar habitats. Invasive animal and plant species can have severe impacts on freshwater turtle populations as resource competitors, predators and parasites (Stanford et al., 2020). Invasive plants and climate change can exacerbate the frequency and intensity of fires (Stanford et al., 2020). For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-49 Location of freshwater turtles within the Southern Gulf catchments Data sources: Atlas of Living Australia (2023a; 2023b); Department of Environment and Science (2023); Department of Environment Parks and Water Security (2019a); OBIS (2023) Figure 3-50 Modelled potential species distribution for northern snake-necked turtle (Chelodina oblonga) in the Southern Gulf catchments Probability of occurrence is based upon a general linear model with model predictors provided in Appendix A. For the SDMs, only records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5 km were used. Red points show locations from Atlas of Living Australia. Data source: Atlas of Living Australia (2023a; 2023b) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for freshwater turtles Freshwater turtles depend upon water and wetland systems for various crucial aspects of their life cycle, including breeding and nesting, food supply and seeking refuge (Cann and Sadlier, 2017) (Table 3-19). In Australia, some species have an indirect dependence on water flow through habitat function driven by the flow regime, for example, food availability for turtles can be affected by changes in water flow. While for other species, the flow dependency supports critical phases of their life history, such as the emergence of northern snapping turtle hatchlings coinciding with the onset of the wet season (Cann and Sadlier, 2017). This species nests during the dry season in nests located within 4 m of the water, in alluvial soils, sand or soil mix on steep to gently sloping banks. Nests seem to be dispersed along the watercourse (Cann and Sadlier, 2017). Numerous turtle species in northern Australia have developed adaptive traits to survive in the highly variable wet-dry environment (Cann and Sadlier, 2017). The nesting behaviour of the northern snake-neck turtle starts in February (wet season) and concludes by July (middle of the dry season). Eggs are laid in the mud, under shallow water, surrounded by flooded waterholes. Embryo development halts while the eggs are submerged and resumes once the water recedes. Hatchling emergence coincides with the onset of the wet season (Cann and Sadlier, 2017). During dry periods, dispersal by freshwater turtles is reduced and they become more vulnerable to changes in water quality, invasive species and habitat degradation (Cann and Sadlier, 2017). In the dry season, turtles often move to the shallows for aestivation. The weeks preceding drying are the riskiest in terms of predation on turtles. The presence of introduced feral pigs represents a high risk of predation on turtles and eggs (Approved Conservation Advice for Elseya lavarackorum (Gulf snapping turtle), 2008; Fordham, 2006; Pusey and Kennard, 2009). Feral pigs also negatively affect turtle habitats, degrading aquatic ecosystems through upheaval of sediments, destruction of aquatic vegetation, creation of anaerobic and acidic conditions, and enrichment of wetlands with nutrients. Additionally, turbid conditions would limit visibility compromising the turtles’ hunting opportunities. Vegetation destruction significantly alters production and respiration regimes, leading to anoxic conditions and pH imbalances (Doupe et al., 2009). Freshwater turtles use large riparian zones to complete various aspects of their life cycle, including nesting. Altering or eliminating these riparian habitats could reduce nest survival and, consequently, juvenile recruitment into the breeding population. It would also affect adult survival through lack of feeding areas and refuge habitat for the dry season, thereby increasing the risk of extinction for freshwater turtle populations (Bodie, 2001). However, more comprehensive data on freshwater turtles are needed, particularly regarding the timing and extent of riparian use. The ecological functions supporting freshwater turtles and their flow requirements are summarised in Table 3-19. Table 3-19 Ecological functions for freshwater turtles and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for freshwater turtles Aquatic and riparian habitats are crucial for freshwater turtles, serving as vital feeding and breeding areas (Cann and Sadlier, 2017; Cosentino et al., 2010; Gibbons et al., 2000; Marchand and Litvaitis, 2004). Fragmentation and habitat loss can increase the vulnerability of freshwater turtles by disrupting their nesting sites and refugia and also restricting their emigration and dispersal among wetlands (Bodie and Semlitsch, 2000; Bowne et al., 2006). Similarly, changes to hydrological patterns (such as timing, velocity, persistence and flow extent) due to barriers and water extraction can lead to changes to the distribution of freshwater species, population growth and reproduction (Hunt et al., 2013). Understanding the impact of these drivers on turtles is critical for improving environmental management and conservation at landscape scales (Bodie and Semlitsch, 2000). As turtle species in northern Australia are less studied than those in eastern Australia and elsewhere, mainly due to the remoteness of their habitats (Cann and Sadlier, 2017), much of the knowledge about their flow requirements and responses to flow are inferred from research on eastern turtle species. The ecological outcomes of threatening processes on freshwater turtles in northern Australia, and their implications for changes to community structure, population viability and biodiversity and ecosystem function, are discussed below and shown in Figure 3-51. Movement is critical for freshwater turtles’ access across breeding, feeding, aestivation and refuge habitats (Ocock et al., 2018). Access to water and connectivity between suitable habitats are key because they let turtles move within the river channels. Threats that reduce river–wetland connectivity, such as water harvesting, dam infrastructure or climate change, are key factors threatening freshwater turtles in northern Australia (Figure 3-51). During transitions from wet to dry seasons, freshwater turtles move on and off the floodplains. In perennial rivers, a reduction in dry-season baseflows (due to extraction) could decrease the availability of suitable habitat supported by flows. Such a reduction in baseflow could even shift the rivers from perennial to intermittent status, which can lessen the turtles’ chances of reaching a freshwater shelter for the dry season (Hunt et al., 2013). Disconnections caused by reduced baseflow hinder the freshwater turtles’ movements on and off the floodplain during the wet-to-dry-season transition (Warfe et al., 2011). Impoundment, regulation and channelisation of riverbanks and beaches can reduce nesting and feeding habitat for most turtle species. Long-lived freshwater turtles are likely to respond slowly to changes in their environment. Due to delays in recruitment, impacts may not be evident until many years after the creation of an impoundment (Tucker et al., 2012; Waltham et al., 2013). In addition, if migratory routes are interrupted through habitat deterioration (especially of nesting sites), gene flow between populations can be disrupted (Alho, 2011; Lees et al., 2016), reducing the genetic diversity of populations. Changes to the inundation and flow regime reduce the food available to freshwater turtles and the suitability of habitats such as waterholes (Warfe et al., 2011), increasing competition for resources (Chessman, 1988; DSITIA, 2014). High abundance of turtles can reduce hatchling survival due to direct predation and resource competition between adults and juveniles (Trembath, 2005). Channelising rivers and shoreline hardening may eliminate nesting and basking areas and alter the hydrodynamic processes that maintain critical nesting habitat (Roosenburg, 2014). Removing exposed logs and snags for recreational boating eliminates critical basking sites and prey habitat (Lindeman, 1999). Changes in flow regimes due to water use and regulation can also affect freshwater turtles by disrupting breeding cues and reducing feeding and/or nesting grounds. For example, a high dry-season flow reduces or eliminates emergent sandbars (Tracy-Smith, 2006), affecting the availability of favoured nesting habitat. This may lead female turtles to seek alternative, less-suitable habitat (Bodie, 2001), resulting in less successful recruitment. Optimal nesting habitat can be lost due to fluctuating water levels caused by water management infrastructure; for example, freshwater turtles’ nests can be inundated, resulting in egg mortality (Waltham et al., 2013). The resulting reduced breeding success, survival and population size of freshwater turtles can affect the community and population structure (Georges et al., 1993; Tucker et al., 2012). Similarly, rapid temperature shifts might prevent historically successful responses, such as the active modification of geographic range. Also, early nesting and early egg maturity resulting from temperature rises may lead to the eggs perishing in the ground, while late nesting risks the eggs being prematurely flooded by rising waters (Jolly, 2008). Refuge habitats are critical during extended drought conditions, as aestivation is limited by fat reserves and dehydration and rarely lasts more than seven months (Roe et al., 2008). Extracting water may reduce the size, number and persistence of waterholes, and may delay their reconnection between seasons (Warfe et al., 2011). Reservoirs, weirs and barrages can reduce floodplains inundation and act as barriers to downstream sediment transmission, resulting in smaller and fewer sandbars (Pusey and Kennard, 2009). Groundwater discharge via springs is essential for maintaining perennial river baseflow in the dry season. Persistent surface water provides important refuge habitat for freshwater turtles (Warfe et al., 2011). Figure 3-51 Conceptual model showing the relationship between threats, drivers, effects and outcomes for freshwater turtles in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.3.4 Mud crabs (Scylla serrata) Description and background to ecology Mud crabs are large-bodied, large-clawed, short-lived, fast-growing decapod crustaceans (>200 mm carapace width) that inhabit the estuarine and shallow subtidal community along tropical and subtropical coastlines, especially mangrove-dominated habitats (Figure 3-52). They are targeted throughout their range as a commercial, recreational and Indigenous fishery resource and a prized table species (commercial catch 40,000 t worldwide in 2012) (Alberts-Hubatsch et al., 2016). Two species of mud crab are found in tropical Australia: Scylla serrata and S. olivacea (Alberts-Hubatsch et al., 2016; Robins et al., 2020). Mud crabs are distributed across the Indo- Pacific region; though in Australia, S. serrata is the dominant commercial species by abundance (Robins et al., 2020). Scylla olivacea is found only in the north-east Gulf of Carpentaria in the Weipa region (Alberts-Hubatsch et al., 2016; Robins et al., 2020). The combined NT and Queensland Gulf of Carpentaria mud crab catch contributed about 25% of the reported mud crab commercial harvest in Australia between 2008 and 2017. The NT crab catch in 2018–19 was 270 t (valued at $7,881,000), while the Queensland crab catch was 1949 t (valued at $19,825,000) (all crab species, Steven et al. (2021)). At the Sydney Fish Market, the price for mud crabs averaged about $34/kg in 2018–19, making them a high-value regional resource (Robins et al., 2020). The mud crab’s high fecundity, high natural mortality and relatively short life span suggest that they are a moderately resilient species suitable for sustainable harvest. The high market price commanded by mud crabs supports their fishery within, and transport from, remote coastal locations in tropical Australia, including the Gulf of Carpentaria regions. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Mud crabs occupy mangrove forest (see Section 3.4.4) and nearby shallow subtidal habitats within estuarine and coastal ecosystems (Alberts-Hubatsch et al., 2016); hence they use the estuary and shallow-water coasts in the Gulf of Carpentaria as habitat. Mud crabs are an important ecological species, being both predator and prey in the coastal ecosystem. As small juveniles, mud crabs are detritivores; as large juveniles and as adults they are benthic predators feeding on crustaceans, molluscs and fish. Estimates suggest that the mud crab population consumes 650 kg biomass per hectare per year in the mangrove forest and 2100 kg biomass per hectare per year in mangrove fringe habitat (Alberts-Hubatsch et al., 2016). Mud crabs dig burrows to rest during the day and during receding high tides (79% resting behaviour), reworking and aerating mud substrates within mangrove forests and mudbanks (Hewitt et al., 2023). Mud crabs forage on the low-rising tide (Hewitt et al., 2023). They play a significant trophic role in mangrove ecosystems. Mud crabs demonstrate a larval life-history strategy (see Robins et al. (2020) for recent comprehensive review): adult crabs mate in the estuary and the females migrate offshore to spawn (September to November; larvae require marine salinity) (Hill, 1994; Hill, 1975; Meynecke et al., 2010; Welch et al., 2014). Their larvae transform to megalopae (the final larval stage) that move by drift inshore where they settle as benthic juveniles in estuarine mangrove and mudflat habitats (Alberts-Hubatsch et al., 2016; Meynecke et al., 2010; Robins et al., 2020). The larval form facilitates ontogenetic migration as crabs grow to the juvenile stage and settle to their inshore habitats and also long-distance dispersal and genetic mixing (Gopurenko and Hughes, 2002; Gopurenko et al., 2003; Robins et al., 2020). Initial recruitment to inshore habitats occurs at the mangrove forest fringe, and as crabs grow their dependence on estuarine mangroves declines (Alberts-Hubatsch et al., 2014). Mud crabs remain in the estuary for several years as sub-adults and adults before the females alone emigrate to spawn (Hill, 1994). Regionally, the annual wet season and subsequent runoff is a significant determinant of their recruitment strength and total catch (possibly lagged by 1 to 2 years) in the estuary and near-shore zone (Meynecke and Lee, 2011; Meynecke et al., 2010). Recent analyses of Gulf of Carpentaria catches support the notion that river flow enhances catch, but also show high air temperature over the wet season as a dominant negative influence on mud crab abundance within the southern Gulf of Carpentaria estuarine habitats (Robins et al., 2020). Figure 3-52 Mangrove and intertidal habitat typical of mud crab habitat in northern Australia Photo attribution: CSIRO Mud crabs in the Southern Gulf catchments and marine region Mud crabs are common in littoral habitats in the Southern Gulf catchments marine region (Figure 3-53). Robins et al. (2020) analysed mud crab catch and environment relationships for an area encompassing Settlement Creek, Nicholson River, Albert River and Leichhardt River as their ‘South West’ region. The region is fished intermittently and contributes about 4% of the harvest of mud crabs from the Queensland Gulf of Carpentaria region. Between 2000 and 2018, the mean harvest for the South West region was 6 t (Robins et al., 2020); it ranged from a minimum catch of less than 1 t (multiple years) to a maximum catch of 15 t in 2003. Robins et al. (2020) concluded that mud crab stocks in the southern Gulf of Carpentaria are often exposed to ‘extreme’ environmental conditions and are close to their thermal tolerance. Photo of mud crab habitat. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-53 Location of mud crab habitat in the Southern Gulf catchments marine region Mud crab juveniles use the mangrove forest and mudbank habitats within the estuary, while adult crabs are caught within the estuary and in shallow subtidal habitats in the littoral zone. Female mud crabs migrate offshore to spawn. Data source: Department of Climate Change‚ Energy‚ the Environment and Water (2020), Geoscience Australia (2017) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for mud crabs The mud crab life-history strategy renders the species critically dependent on the natural flow regime in the wet-dry tropics. Juvenile and adults mud crabs are estuarine and littoral-coast residents, and both of these habitats are influenced by freshwater inflows. The effect of freshwater flows on mud crabs is difficult to define compared to species that emigrate as they grow into a new life-history stage. Mud crabs do not use freshwater riverine or palustrine habitats as juveniles, nor do they emigrate from their estuarine habitats to marine adult habitats. Adult female mud crabs emigrate from inshore to marine habitats to spawn, but they reside in estuarine and coastal habitats as adults prior to their reproductive response (Alberts-Hubatsch et al., 2016). Optimal estuarine conditions for both male and female mud crab growth and survival are found in a brackish ecotone between marine habitats and the freshwater riverine habitats (Pati et al., 2023; Ruscoe et al., 2004). Mud crabs are not subject to emigration cues, though freshwater inflows may cause movement down the estuary as upper reaches become too fresh to tolerate (Robins et al., 2020). Although positive relationships between flow and mud crab catches across the Gulf of Carpentaria and other northern estuaries have been identified previously (Robins et al., 2005), many other environmental parameters are also correlated with catch (Plagányi et al., 2022; Robins et al., 2020). Female mud crabs spawn offshore from September to November. Their larvae require marine salinities (25–30 ppt) and warm waters (26–30 °C) for optimal growth (Alberts-Hubatsch et al., 2016; Welch et al., 2014). Megalopae are tolerant of 15 to 45 ppt salinity, facilitating their occupation of diverse inshore habitats where physical parameters can be variable. Though larvae survive best in marine waters, the growth and mortality of juvenile mud crabs is optimal in brackish waters characteristic of the tropics: about 25 to 30 °C with a salinity of 10 to 20 ppt (for growth) and 10 to 30 ppt (for survival). Mud crabs can tolerate cool conditions (<20 °C) for short periods, but require temperatures higher than 20 °C to grow and function (~25–30 °C is optimal). Juvenile mud crabs resident in estuaries can tolerate a broader salinity range (5–45 ppt); they benefit from perennial baseflows and low flood flows that create brackish conditions in the estuary (Alberts-Hubatsch et al., 2016; Welch et al., 2014). Estuaries in the Australian tropics often become hypersaline in the lead up to the wet season and in years of very low rainfall. Under hypersaline conditions, growth and survival of the crabs may be inhibited until first rains and low- level river flows reduce estuarine salinity to brackish levels (Table 3-20). Adult mud crabs are euryhaline animals, capable of living in freshwater-flooded to hypersaline waters (<5−45 ppt) (Alberts-Hubatsch et al., 2016), with optimal salinity ~20 ppt (Pati et al., 2023). High-level flows benefit the estuarine mud crab population via increased productivity due to nutrient loads delivered to estuarine and near-shore littoral habitats (Burford et al., 2016; Burford et al., 2012; Burford and Faggotter, 2021). Also, mangroves rely on the depositional environment sustained by sediment loads on large floods to maintain their intertidal habitat (Asbridge et al., 2016). However, very large floods cause the loss of marine influence and may negatively affect inshore crab habitats in the year of the flood; though they may be beneficial in subsequent years due to medium-term productivity enhancement (Robins et al., 2020) (Table 3-20). Large floods that create a freshwater estuary cause mortality and movement from estuaries: juvenile crabs in fresh water suffer 100% mortality (Ruscoe et al., 2004) and during a one-in-fifty-year flood in the south-east Gulf of Carpentaria in 2009, adult crabs in freshwater estuaries emigrated elsewhere (Gary Ward (Gulf of Carpentaria fisher), 2010, pers. comm.). In contrast, during lower-level floods survival of juvenile crabs in salinities of 5 to 40 ppt was high (optimally 15–25 ppt), and adult crabs were abundant in brackish estuaries (Robins et al., 2020). The ecological functions that support mud crabs, and their associated flow requirements, are summarised in Table 3-20. Table 3-20 Ecological functions supporting mud crabs and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for mud crabs Mud crabs exhibit a life history that would be significantly affected by interruptions to the natural flows of northern Australian rivers. In the western Gulf of Carpentaria, the short wet season (3 months at most) and unreliability of annual rainfall (including consecutive years of low rainfall) render mud crabs highly vulnerable to climate events, especially cumulative heat from November to March (Robins et al., 2020). While river flow and rainfall have been shown to be positively related to mud crab catch in the eastern, southern and western Gulf of Carpentaria, environmental stressors in the ecosystem, such as evaporation and heat stress, can be extreme and have major negative impacts on mud crab populations (Robins et al., 2020). Particularly in the western and south-western Gulf of Carpentaria, evaporation during the dry season and heat stress during the wet season decreased mud crab catch. Analysis of environmental factors and commercial catch by Robins et al. (2020) showed that river flow and water stress (rainfall offset by evaporation; less stress if rainfall is high) had a positive effect on mud crab catch in some Gulf of Carpentaria catchments, while other stressors such as evaporation during the dry season and heat stress during the wet season had negative effects on catch. Mean sea-level anomaly during the wet season and the Southern Oscillation Index were positive for catches in this region. Hence, reduction in river flows due to water resource development would be expected to have detrimental effects on mud crab catches in the Southern Gulf catchments and marine region. In particular, reduced low-level flows − those flows that condition estuaries to brackish habitats after the extended dry season − would reduce the growth and survival of mud crabs in a hypersaline estuary. The impact of water resource development in three Gulf of Carpentaria rivers on coastal mud crab populations has been modelled for the construction of dams and water harvest at several levels of extraction (Blamey et al., 2023). With the exception of perennial rivers such as the Mitchell River, an array of water harvest and impoundment scenarios predicted a reduction in both the biomass and commercial catch of mud crab by 36 to 46% on average in the ephemeral and temporally variable Flinders and Gilbert rivers in the eastern and southern Gulf of Carpentaria. The risk to mud crab population was assessed as negligible, to major, to severe (two scenarios) for the four water resource development scenarios. The risk to mud crab population was assessed as negligible (one scenario), major (one scenario) and severe (two scenarios) for the four water resource development scenarios. The risk to the commercial fishery for mud crabs was assessed as major for two of the scenarios, severe for one, and negligible for the remaining scenario (Plagányi et al., 2023). Blamey et al. (2023) showed that both constructing dams and harvesting river flows via pumped water extraction affect aspects of the mud crab life history and reduce the resilience of the populations. Reduced volume and duration of high-level flows due to water resource development, and variability in the seasonality and volume of low-level flows, affect estuarine habitat suitability (brackish conditions preferred), growth and survival of mud crabs (Blamey et al., 2023; Robins et al., 2020). The ecological outcomes of threatening processes on mud crabs in northern Australia, and their implications for changes to growth and mortality, community composition, habitat and population, are illustrated in Figure 3-54. Figure 3-54 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mud crabs in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.3.5 Tiger prawns (Penaeus esculentus and P. semisulcatus) Description and background to ecology Tiger prawns are relatively large decapod crustaceans (90–110 g) from the family Penaeidae, a family that exhibits a larval life-history strategy (Dall et al., 1990) with inshore and offshore phases. They inhabit littoral and coastal ecosystems in tropical Australia, from the intertidal to a depth of about 50 m. Two species of tiger prawns are found in the Gulf of Carpentaria: the grooved tiger prawn (Penaeus semisulcatus) and the brown tiger prawns (P. esculentus). Both species are found in the marine regions offshore from the Southern Gulf catchments. Brown tiger prawns are endemic to Australia (Grey et al., 1983), while grooved tiger prawns are found throughout the Indo-West Pacific, from southern Africa to Japan (Grey et al., 1983). In Australia, grooved tiger prawns are restricted to tropical coastlines, while brown tiger prawns are found in both tropical and subtropical latitudes. The annual tiger prawn harvest in the commercial NPF is approximately 1749 t (recent 10-year mean), which is worth about $40 to $50 million annually). The major portion of the tiger prawn catch is taken in the western Gulf of Carpentaria (Laird, 2021). Adult tiger prawns live and spawn offshore in waters 10 to 40 m deep, the larvae and post-larvae advect inshore on currents to settle in littoral seagrass habitats (Condie et al., 1999; Crocos, 1987a; 1987b; Dall et al., 1990; Loneragan et al., 1994; Loneragan et al., 1998; Somers and Kirkwood, 1991). Seagrass habitats are critical to juvenile tiger prawn survival. They provide habitat structure for shelter and stability and provide food directly and via micro- and meio-benthos (Loneragan et al., 1997; Wassenberg and Hill, 1987). Seagrasses provide a refuge from predation (Haywood et al., 1998). Small juvenile tiger For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au prawns do not bury themselves in the sediment, so they depend on seagrass leaf structure to avoid predators, especially fish (Kenyon et al., 1995). The degree of refuge within seagrass habitat depends on seagrass species and morphology (Haywood et al., 1998; Kenyon et al., 1995; Kenyon et al., 1997). Juvenile tiger prawns emigrate offshore at about 10 to 20 mm CL to an epibenthic existence in deeper waters (Loneragan et al., 1994; Vance et al., 1996b). The diet of juvenile and adult tiger prawns consists of small bivalves, gastropods, ophiuroids (brittle stars), crustaceans and polychaete worms (Heales et al., 1996; O'brien, 1994; Wassenberg and Hill, 1987). Bivalves and gastropods are the most common food of juvenile and adult brown and grooved tiger prawns, while crustacean fragments are also common in the diet of grooved tiger prawns. In turn, within their seagrass habitats and offshore, both species are preyed upon by fish (sharks and teleosts), squid and cuttlefish (Brewer et al., 1991; Brewer et al., 1995). Tiger prawns are an important fishery resource, with most fishing effort occurring from August to November annually. The tiger prawn species group is managed via a stock assessment (Deng et al., 2021; Dichmont et al., 2010; Punt et al., 2011; Zhou et al., 2009). Over the decades, management provisions for the NPF have been continually revised and improved. Currently, the fishery is managed by effort regulation (Deng et al., 2021; Deng et al., 2015). Brown tiger prawns grow to about 55 mm CL (53 mm CL ≈ 100 g) for females and about 47 mm CL (43 mm CL ≈ 65 g) for males. Grooved tiger prawns are larger; females grow to about 58 mm CL (53 mm CL ≈ 110 g) and males to about 47 mm CL (43 mm CL ≈ 65 g). Tiger prawns in the Southern Gulf catchments Adult tiger prawns occupy relatively soft-sediment substrates in relatively shallow waters within the north-west and south-west Gulf of Carpentaria. The two species have relatively allopatric distributions, which depend on sediment texture (Somers, 1987; Somers, 1994) and latitude (Venables and Dichmont, 2004). Importantly, adult tiger prawns of both species are found adjacent to littoral seagrass communities, their critical juvenile habitat (Staples et al., 1985). As large juveniles and adults, both species of tiger prawns bury into the sediment during the day and emerge to feed at night (when they are fished). Both grooved and brown tiger prawns are found in the south-western Gulf of Carpentaria in coastal waters ranging from 10 to 45 m deep, offshore from the rivers of the Southern Gulf catchments (Somers, 1994) (Figure 3-55 and Figure 3-56). Tiger prawns are abundant to the west and east of Mornington Island, within the Southern Gulf catchments marine region (Kenyon, 2021; Robertson et al., 1985; Somers, 1994). The rivers that feed into the Southern Gulf catchments marine region span the ‘Mornington’ (16-year mean catch: 177 t; 11% of total tiger prawn catch) and ‘Sweers’ (16-year mean catch: 60 t; 3.7%) reporting regions for the NPF (Laird, 2021). In the Southern Gulf catchments marine region, brown tiger prawns dominate the species distribution on the fishing grounds to the east and west of Mornington Island (Kenyon, 2021). The seagrass habitats of the juvenile phase of both species of tiger prawns form a near-continuous swathe of vegetation in the littoral zone along the coast to the west of Mornington Island and along the east coast of Mornington Island in the coastal bays such as Charlie Bush Bay (Coles and Lee Long, 1985; Poiner et al., 1987). The juvenile tiger prawn population within littoral habitats along the extensive Gulf of Carpentaria coastline has not been sampled. However, Coles and Lee Long (1985) verified the presence of a population of juvenile brown tiger prawns in seagrass habitats on the east coast of Mornington Island. Figure 3-55 Fisheries catch of brown tiger prawns in the Southern Gulf catchments marine region Juvenile habitat for brown tiger prawns includes shallow littoral seagrass habitats in reasonably clear waters along coasts. Adult brown tiger prawns are caught offshore in water about 20 to 45 m deep in the marine habitat adjacent to their juvenile habitats. Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022); Coles and Lee Long (1985) Figure 3-56 Fisheries catch of grooved tiger prawns in the Southern Gulf catchments marine region Juvenile habitat for grooved tiger prawns includes shallow littoral seagrass habitats in reasonably clear waters along coasts. Adult grooved tiger prawns are caught offshore in water about 20 to 45 m deep in the marine habitat adjacent to their juvenile habitats. Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022); Coles and Lee Long (1985) Flow–ecology relationships for tiger prawns The life-history strategy of tiger prawns includes a dependence on littoral seagrass habitats as juveniles (Dall et al., 1990; Loneragan et al., 1994). Seagrass habitats are found both within some estuaries, but along the Gulf of Carpentaria coast, seagrass habitats are mostly offset from estuaries away from the direct influence of river flows (Poiner et al., 1987). Tiger prawn post- larvae use currents to move shoreward to shallow, subtidal habitats, especially seagrass communities (Vance et al., 1996b). Before the annual wet season, post-larvae settle within the shallow, shoreward extent of the seagrass community in relatively clear waters favourable to seagrass growth (Loneragan et al., 1994; Loneragan et al., 1998; Vance et al., 1996b). They shelter, forage and grow within the seagrass habitat where leaf structure reduces predation and promotes primary productivity and prawn growth (Haywood et al., 1998; Kenyon et al., 1995). Predation by fish within the seagrass habitats is high, and a significant proportion of the population is lost (Brewer et al., 1995). The primary productivity of the river and coastal shallow habitats of the Gulf of Carpentaria is nutrient limited (Burford et al., 2012; Burford and Faggotter, 2021). Seagrass thrives in oligotrophic waters, while a sufficient level of nutrients in the environment supports seagrass growth. Floodwaters transport terrigenous nutrients from the catchment that are deposited within the flood plume and littoral zone adjacent to Gulf of Carpentaria rivers (Burford et al., 2012; Burford and Faggotter, 2021). In addition, during large floods that inundate salt flats adjacent to the lower estuary, organic matter produced by wetted algal crusts are deposited within the plume (Burford et al., 2016). Prolonged turbidity of direct smothering is a risk to seagrass habitats close to river mouths (Longstaff et al., 1999; Longstaff and Dennison, 1999). However, longshore transport of a proportion of nutrients deposited adjacent to river mouths maintains a nutrient balance and would benefit the productivity of coastal seagrasses in the Gulf of Carpentaria, while not adversely affecting water quality within the seagrass habitat. The hydrology within the inshore Gulf of Carpentaria may ensure that terrigenous-sourced nutrients are retained in poorly mixed layer of water inshore. Therefore, river flows are crucial in maintaining the balance of coastal productivity in the Gulf of Carpentaria littoral zone via low-level source nutrients from adjacent catchments. The growth and stability of the littoral seagrass communities in the Gulf of Carpentaria are enhanced by the riverine-source nutrient inputs to the coastal waters. Hence, tiger prawn juveniles benefit from the natural flow regime of Gulf of Carpentaria rivers. In the Australian tropics, the levels of nutrient inputs are comparatively low, maintaining oligotrophic coastal waters. In addition, though not verified by data similar to the 40 years of flow−catch relationship recorded for banana prawns, when wet-season rainfall is high and salinity declines in coastal estuarine or embayment seagrass habitats, the grooved tiger prawn fishery catch is enhanced (Bishop et al., 2016; Vance et al., 1996b). The ecological functions and their supporting flow requirements for tiger prawns are summarised in Table 3-21. Table 3-21 Ecological functions supporting tiger prawns and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for tiger prawns Coastal waters in Australia’s tropics, such as the southern Gulf of Carpentaria and upstream catchment water bodies, form oligotrophic though productive ecosystems, which are stressed by heat, high evaporation, hypersaline estuaries and lack of precipitation for 9 months of the year (Blondeau-Patissier et al., 2014; Robins et al., 2020). The monsoon season delivers environmental flux that stimulates the ecosystem and estuarine and marine communities (Blondeau-Patissier et al., 2014). Biota benefit from the annual dynamic freshwater pulse flows. Primary productivity (Ndehedehe et al., 2020a), fish growth (Leahy and Robins, 2021) and crustacean and fish populations (Plagányi et al., 2022) benefit from the maintenance of trend in historical flows within the ecosystem. Though not yet well understood, littoral seagrass communities within the southern Gulf of Carpentaria and their dependent fauna benefit from the dynamic provision of monsoon-driven inputs to the system on an annual basis (Plagányi et al., 2022; Plagányi et al., 2023). Once an abundant population of juvenile tiger prawns is established within the coastal seagrass community, it benefits from high-level flood flows in previous years that supported nutrient dynamics within the community. In addition, when catchment flows are large enough to cause salinity declines in outer estuarine or embayment coastal seagrass habitats, evidence suggests that juvenile grooved tiger prawns leave the shallow habitats and move seaward to adult habitats offshore and enhance fishery catch (Bishop et al., 2016; Vance et al., 1996b). Hence reduced catchment runoff would mimic drier years when catch declines earlier in the season (Bishop et al., 2016). The ecological outcomes of threatening processes on tiger prawns in northern Australia, and their implications for changes to growth and mortality, community composition, habitat and population, are illustrated in Figure 3-57. Figure 3-57 Conceptual model showing the relationship between threats, drivers, effects and outcomes for tiger prawns in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.4 Flow-dependent habitats 3.4.1 Floodplain wetlands Description and background to ecology Wetlands in the wet-dry tropics of Australia have great conservation value (Finlayson et al., 1999), and are one of the most diverse aquatic ecosystems in Australia (Douglas et al., 2005). Wetlands provide permanent, temporary or refugia habitat for both local and migratory waterbirds (van Dam et al., 2008), spawning grounds and nurseries for floodplain-dependent fish (Ward and Stanford, 1995), and habitat for many other aquatic and riparian species (van Dam et al., 2008). Floodplain wetlands are an important source of nutrients and organic carbon, driving primary and secondary productivity (Junk et al., 1989; Nielsen et al., 2015). Wetlands also provide a range of additional ecosystem services, including water quality improvement, carbon sequestration and flood mitigation (Mitsch et al., 2015). Hydrological regimes are fundamental to sustaining the ecological characteristics of rivers and their associated floodplains (Pettit et al., 2017). In the wet-dry tropics of northern Australia, the ecology of wetlands is highly dependent on the seasonal rainfall−runoff pattern and the associated low and high flows (Pidgeon and Humphrey, 1999; Warfe et al., 2011). These flows are important drivers of floodplain wetland ecosystem structure and processes (Close et al., 2012; Warfe et al., 2011). Changes to these flow characteristics are likely to have a significant impact on the aquatic For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au biota (Close et al., 2012). The timing, duration, extent and magnitude of wetland inundation has the greatest impact on the ecological values, including species diversity, productivity and habitat structure (Close et al., 2015). Under the Ramsar convention a wetland is defined as (Ramsar Convention Secretariat, 2004): ‘areas of marsh, fen, peatland or water, whether natural or artificial, permanent or temporary, with water that is static or flowing, fresh, brackish or salt, including areas of marine water the depth of which at low tide does not exceed six metres.’ The NT Government defines wetlands as including coastal saltmarshes, mangrove swamps, freshwater lakes and swamps, floodplains, freshwater ponds, springs and saline lakes, that can be permanent, seasonal or intermittent, and can be natural or artificial (NT Government, 2020). This Assessment does not consider areas within the river channel to be wetlands (they are considered to be inchannel waterholes (see Section 3.4.2)). Similarly, marine or saline habitats including mangroves and coastal saltmarshes (salt flats) are also treated as separate assets in this project (sections 3.4.4 and 3.4.5, respectively). Floodplain wetlands in the Southern Gulf catchments The Southern Gulf catchments contain 13 nationally significant wetlands listed in the DIWA: Bluebush Swamp, Buffalo Lake Aggregation, Forsyth Island Wetlands, Gregory River, Lake Julius, Lake Moondarra, Lawn Hill Gorge, Marless Lagoon Aggregation, Musselbrook Creek Aggregation, Nicholson Delta Aggregation, Southern Gulf Aggregation, Thorntonia Aggregation and Wentworth Aggregation (Table 3-22; Figure 3-59) (Department of Agriculture‚ Water and the Environment, 2021a). There are no Ramsar-listed wetlands within the Southern Gulf catchments. The Wentworth Aggregation is found in the Settlement Creek catchment (Figure 3-59) and covers an area of 82,306 ha (Department of Agriculture‚ Water and the Environment, 2021a). It contains the full range of wetland types (estuarine, lacustrine (lake), palustrine and riverine) and is considered to have high wilderness value due to its remoteness. The main habitat types are: estuarine salt flats and saltmarshes (55.6% of the wetland area); coastal and sub-coastal non- floodplain tree swamp – Melaleuca spp. and Eucalyptus spp. (13.5% of the wetland area); coastal and sub-coastal floodplain grass, sedge, herb swamp (7.2%); and estuarine – mangroves and related tree communities (6.7% of the wetland area) (Department of Environment and Science (Qld), 2022). It is considered an important wetland for waterbirds (Department of Agriculture‚ Water and the Environment, 2021a). The Marless Lagoon Aggregation spans the Settlement Creek and Nicholson River Basin catchments (Figure 3-59) and has an area of 166,948 ha (Department of Agriculture‚ Water and the Environment, 2021a). It has extensive palustrine wetlands (97.6% of the wetland area) in which the dominant habitat type is coastal and sub-coastal non-floodplain tree swamp – Melaleuca spp. and Eucalyptus spp. (97.3% of the wetland area) (Department of Environment and Science (Qld), 2022). Bluebush Swamp is located in the Nicholson River Basin catchment (Figure 3-59) and is 879 ha (Department of Agriculture‚ Water and the Environment, 2021a). It is a scrub−shrub wetland with Acacia stenophylla as the dominant species, and it has areas of shallow, open water. Bluebush Swamp provides habitat for waterbirds in the late wet season and in autumn and spring (Department of Agriculture‚ Water and the Environment, 2021a). Lawn Hill Gorge is located within Boodjamulla (Lawn Hill) National Park in the Nicholson River Basin catchment (Figure 3-59) and has an area of 1133 ha (Department of Agriculture‚ Water and the Environment, 2021a). Prior to being gazetted as a national park, it was extensively grazed, with feral pigs and other invasive animals continuing to affect the area. A variety of recreational activities are available within the national park, although visitor numbers are controlled. Indigenous usage of the area dates back to between 17,000 and 30,000 years (Department of Agriculture‚ Water and the Environment, 2021a). The gorge itself is located on Lawn Hill Creek, which is spring-fed from limestone but also receives wet-season flushes (Department of Agriculture‚ Water and the Environment, 2021a; Department of Environment and Science (Qld), 2021). Musselbrook Creek Aggregation is located in the Nicholson River Basin catchment (Figure 3-59) and has an area of 45,114 ha (Department of Agriculture‚ Water and the Environment, 2021a). The most dominant habitat types are coastal and sub-coastal floodplain tree swamp – Melaleuca spp. and Eucalyptus spp. (34.6% of the wetland area), arid and semi-arid tree swamp (floodplain) (32.8% of the wetland area) and riverine (27.6% of the wetland area) (Department of Environment and Science (Qld), 2022). The area has been extensively grazed by cattle and feral horses (Department of Agriculture‚ Water and the Environment, 2021a). Nicholson Delta Aggregation is in the Nicholson River Basin catchment (Figure 3-59) and has an area of 63,646 ha (Department of Agriculture‚ Water and the Environment, 2021a). This aggregation has a freshwater–saltwater gradient, as estuarine waters flood the salt flats and tidal channels, particularly during the dry season. There is also a series of freshwater wetlands (Department of Agriculture‚ Water and the Environment, 2021a). The main habitat is estuarine – salt flats and saltmarshes (40.8% of the wetland area) and riverine (31.1% of the wetland area) (Department of Environment and Science (Qld), 2022). The wetlands in this aggregation are a mix of permanent, semi-permanent and seasonal wetlands, and they provide refugia habitat for waterbirds in the dry season (Department of Agriculture‚ Water and the Environment, 2021a). Thorntonia Aggregation is located in the Nicholson River Basin catchment (Figure 3-59) and has an area of 298,629 ha (Department of Agriculture‚ Water and the Environment, 2021a). It is partly within the Boodjamulla (Lawn Hill) National Park and contains the internationally significant Riversleigh fossil field. It has deep permanent channels which are spring-fed and shallower seasonal channels. The perennial channels provide refugia habitat in the dry season (Department of Agriculture‚ Water and the Environment, 2021a). Gregory River crosses the Nicholson River Basin and Leichhardt River Basin (Figure 3-59) and is 26,630 ha (Department of Agriculture‚ Water and the Environment, 2021a). It is the largest perennial river in semi-arid and arid Queensland, and is spring-fed, with additional flow from runoff. The area is used for recreational purposes, including camping, canoeing and fishing, and has had extensive cattle grazing. Freshwater crocodiles are very common (Department of Agriculture‚ Water and the Environment, 2021a). Lake Julius is an artificial lake located in the Leichhardt River Basin catchment (Figure 3-59). Lake Julius is 1936 ha and was created when the Leichhardt River was damned to secure water for irrigation and town water supply (Department of Agriculture‚ Water and the Environment, 2021a). The lake is used for recreational purposes, including boating and fishing, with the land surrounding the lake used to graze cattle. As a permanent lake, it provides important habitat for waterbirds during the dry season. It also has a variety of freshwater fish, the red-clawed crayfish (Cherax quadricarinatus) and the freshwater crocodile (Department of Agriculture‚ Water and the Environment, 2021a). Lake Moondarra (Figure 3-58) is an artificial lake located just outside the town of Mount Isa in the Leichhardt River Basin catchment (Figure 3-59). Situated on the Leichhardt River, it is 1742 ha and is upstream of Lake Julius (Department of Agriculture‚ Water and the Environment, 2021a). As a permanent water body, it provides refugia habitat for waterbirds, and is considered an important recreations area, allowing for both boating and fishing. Cattle grazing is extensive, including at the water’s edge, causing damage. Seasonal turbidity is an issue (Department of Agriculture‚ Water and the Environment, 2021a). Buffalo Lake Aggregation has an area of 1911 ha and is located in the Morning Inlet Basin catchment (Figure 3-59) (Department of Agriculture‚ Water and the Environment, 2021a). The lake is shallow (<1 m depth) and ephemeral, flooding in extreme wet seasons and drying out completely most dry seasons. It is also occasionally flooded by tidal surges (Department of Agriculture‚ Water and the Environment, 2021a). The dominant habitats are coastal and sub- coastal floodplain lake (69.0% of the wetland area) and coastal and sub-coastal floodplain grass, sedge, herb swamp (24.3% of the wetland area) (Department of Environment and Science (Qld), 2022). The lake provides important habitat for waterbirds, including migratory species. The area has been extensively grazed (Department of Agriculture‚ Water and the Environment, 2021a). Southern Gulf Aggregation is the largest continuous estuarine wetland in Australia at 545,577 ha, and it crosses into all four catchments (Figure 3-59) (Department of Agriculture‚ Water and the Environment, 2021a). Dominated by estuarine – salt flats and saltmarshes (77.3% of the wetland area) and estuarine – mangroves and related tree communities (20.4% of the wetland area) (Department of Environment and Science (Qld), 2022), the area is governed by estuarine tides, and during the wet season, freshwater flooding (Department of Agriculture‚ Water and the Environment, 2021a). The wetlands located along the inland edges of the aggregation are all seasonal and are brackish. The area is considered one of the most important shorebirds sites in Australia (Department of Agriculture‚ Water and the Environment, 2021a). Forsyth Island Wetlands are located on Forsyth Island, about 10 km off the coast of Bayley Point. It is an estuarine wetland, with important seagrass habitats in Government Bay (Department of Agriculture‚ Water and the Environment, 2021a). The habitat consists of estuarine – salt flats and saltmarshes (60.9% of the wetland area) and estuarine – mangroves and related tree communities (39.1% of the wetland area) (Department of Environment and Science (Qld), 2022). The island itself is an Indigenous reserve, with the area, including the surrounding waters, used for fishing and hunting (Department of Agriculture‚ Water and the Environment, 2021a). As this is a marine wetland, it is considered out of scope for this study. Significant parts of the floodplain areas within the Southern Gulf catchments are already incorporated into the existing 13 nationally significant wetlands (see land subject to inundation Figure 3-59). There is additional floodplain on Musselbrook Creek, extending beyond the Musselbrook Creek Aggregation. There is also additional floodplain near the Wentworth Aggregation (Figure 3-59). Table 3-22 Nationally important wetlands in the Southern Gulf catchments For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-58 Brolgas flying into the sunset at Lake Moondarra Photo attribution: CSIRO For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-59 Land subject to inundation (potential floodplain wetlands) and important wetlands in the Southern Gulf catchments Dataset: Geoscience Australia (2017); Department of the Environment and Energy (2010) Flow–ecology relationships for floodplain wetlands The inundation pattern, including the extent, duration, depth, rate of inundation and timing are important factors for maintaining the ecological function of wetlands (Bunn and Arthington, 2002; Pettit et al., 2017). The pattern of connectivity is important for the movement of nutrients and biota on and off the floodplain (Junk et al., 1989). Changing the pattern of connectivity can change primary production on the floodplain, which is thought to be a major determinant of the level of species diversity, productivity and habitat structure (Close et al., 2015). This, in turn, can affect the productivity of the overall system (Brodie and Mitchell, 2005; Hamilton, 2010). The timing and duration of flooding events can be important factors to determine the success of a breeding event (e.g. bird nesting, fish spawning) (Close et al., 2012). The extent of the flood influences the extent to which habitat is provided for biota. A reduction in the flood extent will lead to a reduction in suitable habitat available to biota, potentially reducing the viability of populations (Bunn and Arthington, 2002). Table 3-23 outlines these important ecological functions and their corresponding flow component or attribute. Table 3-23 Ecological functions supporting floodplain wetlands and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways to change for floodplain wetlands Several threatening processes could affect the persistence of wetlands in the Southern Gulf catchments, including river regulation, water extraction, climate change and land use change. River regulation can affect wetlands when regulating structures capture flows and there is a downstream demand for water. Instream dams can have a significant impact on the immediate upstream and downstream environments, but may have less impact lower down in the catchment if the storage is located in the upper catchment and if other tributaries are unregulated (Petheram et al., 2008). The ecological impacts of dams on wetlands can be numerous. Dams can prevent water from flowing onto floodplain wetlands by capturing water from moderate to large rainfall events, preventing flood pulses from moving down the channel (Kingsford, 2000). This loss of connectivity to the floodplain can result in the reduction of wetland area, and even loss of wetlands, as they may transition into terrestrial habitats. Disruption to the natural flow regime, including altered magnitude, frequency, duration, timing and rate of change of flows within a system, can affect all aspects of a riverine ecosystem (Poff and Zimmerman, 2010b). These aspects include the structure, function and biodiversity of wetland ecosystems (Poff and Zimmerman, 2010b; Richter et al., 1996). Water extraction can be from both groundwater and surface water sources, the latter being rivers or standing water bodies, such as lakes or wetlands. For a range of reasons, extraction of surface water generally has less impact on the environment than instream storages. One reason is that surface water extraction can occur during high-flow events, such as floods, and not during low- flow periods (Petheram et al., 2008). Water extraction can lower the quantity of water in the river, providing less water for the environment. Reduced flooding extent and duration is also likely to reduce local groundwater recharge and thus reduce groundwater flows back into wetlands once floodwaters recede, placing pressure on floodplain wetland ecosystems that depend on groundwater discharge to sustain them during dry periods (Froend and Horwitz (2018); refer to aquatic GDEs section 3.4.2). Climate change is a major threat to wetlands (Salimi et al., 2021). Future changes in the climate may affect rainfall, runoff and evapotranspiration patterns (Grieger et al., 2020; Salimi et al., 2021), affecting the hydrology of a system, including the baseflow and flood patterns (Erwin, 2009). Changes to the hydrology can also affect the water quality through, for example, increased erosion and changes to water temperature (Erwin, 2009). Changes to the hydrology and water temperature of wetlands can affect their biogeochemistry and function, and therefore the ecosystem services that they provide (Salimi et al., 2021). Climate change, including changes in precipitation and rates of evaporation, can affect the quantity of inflows to a river. Vulnerability of wetlands to climate change depends on the hydrology conditions experienced and the wetlands’ positions within the landscape (Winter, 2000). Hydrological landscapes are defined by their water source and their flow characteristics. Winter (2000) found that wetlands that depended on rainfall were more vulnerable to changes in climate than wetlands that depended on regional groundwater, due to the buffering capacity of groundwater systems. Coastal wetlands in particular may be vulnerable to climate change. Climate change impacts on coastal wetlands may include accelerated sea-level rise, a change in freshwater inputs, and changes to the frequency and intensity of storms and storm surges (Day et al., 2008; Nicholls et al., 1999). Sea-level rise and reduced freshwater inputs can lead to saltwater intrusion of wetlands (Close et al., 2015; Close et al., 2012), which in turn can convert freshwater floodplains to saline habitats (Finlayson et al., 1999). Land use change can include modification of land management practices, changes to the intensity or type of agricultural production, increased vegetation clearing, or increased mining or urbanisation. These changes can affect water quality by increasing nutrient loads, sediment and turbidity levels (Finlayson et al., 1999). Changes to land use can also increase the likelihood of invasive species, due to the increased level of disturbance (Finlayson et al., 1999). Taken individually, these threats can each have significant impacts on wetlands and their ability to provide ecosystem services and habitat, and the interactions of these threats can compound these impacts. The ecological outcomes of threatening processes on floodplain wetlands in northern Australia, with their implications for changes to floodplain wetland biodiversity and function, are presented in Figure 3-60. Figure 3-60 Conceptual model showing the relationship between threats, drivers, effects and outcomes for floodplain wetlands in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.4.2 Groundwater-dependent ecosystems Description and background to ecology GDEs are defined as habitats that require groundwater at critical times (continuously, seasonally or only sporadically) to continue their existence and support the plants and animals that inhabit them and other ecosystem functions and services they provide (modified from Richardson et al. (2011b)). For example, in these habitats, groundwater may support vegetation such as the red cabbage palm (Livistona mariae; Box et al. (2008)) in areas where it would not otherwise persist, fish may persist in groundwater-fed waterholes during dry seasons (e.g. McNeil et al., 2013), and stygofauna may live in underground water cavities maintained by groundwater (e.g. karst aquifers; Oberprieler et al. (2021)). In the wet-dry tropics typical of northern Australia, groundwater is important, being recharged over wet periods and supporting ecological function of water- dependent habitats and species during dry periods. GDEs are typically categorised into three functional types: • aquatic groundwater-dependent ecosystems • terrestrial groundwater-dependent ecosystems • subterranean groundwater-dependent ecosystems. These functional types are described in the following sections. Conceptual model For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Aquatic groundwater-dependent ecosystems Aquatic GDEs are surface water habitats that require groundwater discharge to the surface or the presence of near-surface groundwater (e.g. for hyporheic exchange, which is mixing of the surface water with shallow subsurface water in the sediment surrounding rivers and wetlands). They include groundwater-fed spring, wetland, river, estuary and coastal (submarine groundwater discharge) ecosystems. The loss of groundwater can have extreme consequences, such as the complete drying out of mound springs and loss of all dependent species (e.g. Fairfax and Fensham, 2002). Habitats largely supported by surface water flow can still rely on groundwater at specific times or to maintain processes, such as maintaining the quality or temperature of water available (e.g. for fish spawning (Geist et al., 2002)) or nutrients for animal and plant growth (Moore (2010)). The impacts of reduced groundwater can appear over long periods and may lead to lower recruitment, loss of species diversity and abundance, proliferation of invasive species, and changes in the structure and function of the ecosystem (e.g. Nevill et al., 2010). Terrestrial groundwater-dependent ecosystems Terrestrial GDEs are vegetated habitats supported by subsurface groundwater, for example, trees that use groundwater and the various plants and animals supported by the habitat the trees provide. Groundwater-dependent terrestrial vegetation requires access to groundwater at critical times for survival (varies depending on species, climate, environment and soil water−holding properties), flowering and successful recruitment (e.g. Horner et al. (2009)). Some terrestrial vegetation species only occur where groundwater is available (obligate GDEs), while other species use groundwater in some habitats (facultative GDEs) but can also exist in habitats where sufficient water within unsaturated soils (driven by climate and plant-available water capacity of soils) removes the need for groundwater (e.g. Pritchard et al., 2010). Regardless of the species, mature vegetation is unlikely to be able to adapt to changes in water availability outside natural variation (e.g. threshold responses; Kath et al. (2014)). Terrestrial GDEs have some inbuilt resilience to changes in water availability and quality, but long-term change in groundwater regime (driven by water resource development or climate change) is likely to result in dieback of groundwater- dependent vegetation (whether obligate or facultative) after some lag period. Dieback of groundwater-dependent vegetation may have broad environmental implications, causing shifts in ecosystem composition and structure (change in the density and diversity of species) and function (e.g. change in the ecosystem’s ability to provide suitable food or habitat for animal species, e.g. Betts et al. (2010), Fleming et al. (2021)). Obligate versus facultative GDEs – challenging definitions A common misconception has broadly propagated though GDE literature that the term ‘obligate GDE’ refers to ecosystems that require a permanent source of groundwater, and the term ‘facultative GDE’ refers to ecosystems that only use groundwater opportunistically, implying that groundwater is not critical to the survival of the ecosystem and that facultative GDEs will survive if groundwater availability is permanently removed. This definition is misleading. Facultative GDEs will become degraded if groundwater is not available at critical times. Therefore, within this project, the terms are defined as follows: Obligate GDE: an ecosystem that will only naturally occur where groundwater is available at critical times (this may be continuous, seasonally or sporadically). Facultative GDE: an ecosystem that naturally occurs in some environments (under specific climate and site conditions) in which it must receive groundwater at critical times (this may be continuous, seasonally or sporadically), but it can also occur in other environments in which it naturally receives enough water from other sources (e.g. rainfall, surface water flows, unsaturated soil stores) that it never uses groundwater. In the case of facultative GDEs, groundwater dependence cannot be proven based on species composition alone. Further studies will be required to determine sources of water used. For example, Melaleuca leucadendra uses groundwater in some environments (Canham et al., 2021) but not in others (O’Grady et al., 2006). Figure 3-61 demonstrates that obligate groundwater-dependent vegetation only occurs in parts of the landscape where there is a reliable source of groundwater. In contrast, facultative GDEs grow and depend on groundwater in some areas but can also establish and thrive in areas where there is sufficient soil water to sustain them without ever having access to groundwater. Obligate GDEs are always vulnerable to unprecedented declines in groundwater availability. Facultative GDEs are vulnerable to groundwater declines in some parts of the landscape, but in other parts they may not require groundwater. Further site assessment is required to establish water dependence of facultative GDE species. Figure 3-61 Conceptualisation of obligate and facultative groundwater-dependent vegetation Phreatophytes are vegetation that draw their water from near the water table. Source: Pritchard et al. (2010) Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-62 depicts natural terrestrial vegetation vigour associated with changes in water availability. During the wet season (I), vegetation has access to sufficient water and productivity and diversity are high. During the dry season (II), there is reduced water availability and productivity. As soils dry, annual vegetation species die back while deeper-rooted species stay green through access to deeper soil water or groundwater. If water availability is reduced beyond natural dry-season variation (III), deeper-rooted species also die back once deeper soil water and groundwater sources become inaccessible. This is likely to result in a shift in ecosystem type (e.g. forest to savanna) and makes the ecosystem more susceptible to invasive plants. Figure 3-62 Conceptualisation of terrestrial GDEs: (I) vigorous ecosystems with seasonally high water availability, (II) ecosystem condition with seasonally low water availability, and (III) seasonal low after groundwater development Source: Rohde et al. (2017) Subterranean groundwater-dependent ecosystems Subterranean GDEs are cave and aquifer systems that provide habitat for subterranean fauna that depend on the presence of groundwater (e.g. troglofauna (cave dwelling) and stygofauna; Richardson et al. (2011a)). Subterranean fauna have limited mobility, and changes in groundwater beyond natural fluctuation in watertable elevation or groundwater quality risks loss of the local communities (Hose et al., 2015). Some subterranean fauna are only known to exist in discrete localities (e.g. Hancock and Boulton, 2008), so loss of local communities can result in species extinction. Apart from their intrinsic biodiversity value, subterranean ecosystems are indicators of groundwater health, and they potentially provide ecosystem services such as nutrient cycling and water purification (Glanville et al., 2016; Smith et al., 2016). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. GDEs in the Southern Gulf catchments The National GDE Atlas (Bureau of Meteorology, 2017) contains maps of the distribution of known and potential groundwater-dependent inland aquatic and terrestrial ecosystems and also some subterranean ecosystems in Queensland. Mapping of potential GDEs within the GDE Atlas was based on the location of known GDEs and their extrapolation to regional scales using a process that relied on the integration of expert opinion, remote sensing data (2000 to 2010) and geographic information systems (Doody et al., 2017). There is little known about coastal or submarine groundwater discharge along the northern coast of Australia. In Australia, the biodiversity and distribution of subterranean ecosystems remain largely unknown. Three types of aquifers are known to provide subterranean ecosystems that can support stygofauna: karstic, fractured rock and alluvial. Typically these occur where the depth to groundwater is less than 30 m (Doody et al., 2019), but some have recently been found to depths of 70 m (Oberprieler et al. (2021)). Karstic aquifers are a prominent feature in the Southern Gulf catchment area (see section on subterranean GDEs below). Aquatic GDEs in the Southern Gulf catchments Regional studies in the NT have shown there are many permanent or near-permanent flowing river sections and springs in the Settlement Creek catchment (Bureau of Meteorology, 2017), and these support aquatic life and fringing vegetation. In Queensland, numerous freshwater springs originating from the limestone plateau feeding into Lawn Hill Creek have been mapped west of Gregory in Boodjamulla (Lawn Hill) National Park. These aquatic ecosystems are mapped as ‘known GDEs’ in the GDE Atlas (Figure 3-63). There are also hundreds of river sections, lakes, springs and wetlands that are believed to be supported by groundwater discharge based on remote sensing work and expert opinion, and these are mapped as ‘potential GDEs’ (Bureau of Meteorology (2017); Figure 3-63). Additional mapped springs are assumed to arise from groundwater discharge (Department of Environment and Science, 2020; Department of Environment‚ Parks and Water Security (NT), 2013) and therefore to be aquatic GDEs. Little is known about coastal or submarine groundwater discharge in the Southern Gulf catchments. Global-scale modelling suggests there is potential for minor submarine groundwater discharge off the coast from the Southern Gulf catchments (Luijendijk et al., 2020), but there have been no local-scale studies to substantiate this. Figure 3-63 Distribution of potential groundwater-dependent aquatic ecosystems in the Southern Gulf catchments Note: A buffer of 1 km has been applied to GDE mapping so that they are visible on the map scale. Only the distribution for the mainland is shown. Dataset: Bureau of Meteorology (2017) Terrestrial GDEs in the Southern Gulf catchments GDE Atlas analysis (based on remote sensing work and expert opinion) suggests that groundwater- dependent vegetation potentially occurs across broad areas of the Southern Gulf catchments, and these are mapped as ‘potential GDEs’ (Bureau of Meteorology (2017); Figure 3-64). However, no on-ground studies have verified groundwater dependence of vegetation within the Southern Gulf catchments. Terrestrial GDEs mapped in the GDE Atlas within the Southern Gulf catchments include various acacia, eucalypt, monsoon vine forest, mangrove and Melaleuca species. However, a number of other terrestrial vegetation species that occur in the Southern Gulf catchments are likely to use groundwater (e.g. river red gum (Eucalyptus camaldulensis)). Preliminary indications of where specific potential terrestrial GDE species exist within the Southern Gulf catchments are shown in Figure 3-65 and Figure 3-66. Figure 3-65 maps observed occurrence of three tree species (E. camaldulensis, Melaleuca argentea, Barringtonia acutangular) thought to only occur naturally where they have access to groundwater at critical times (i.e. obligate GDE species; Lamontagne et al. (2005); Mensforth et al. (1994)). Most of the observed occurrences of E. camaldulensis and M. argentea are riparian, along the major rivers, although M. argentea is also known to occur in floodplain and swamp habitats. However, comprehensive species distribution mapping does not exist. B. acutangular (obligate GDE) is known to occur in freshwater mangrove forests. Therefore, the distribution of obligate GDE vegetation species is expected to be more extensive than mapped in Figure 3-65. Figure 3-66 shows the observed occurrence of many other known GDE species grouped by vegetation types (GDEs) that are known to use groundwater in some locations, but under some climate and/or site conditions may not be groundwater dependent (i.e. facultative GDE species). It also shows the occurrence of potential GDE species grouped by vegetation type (potential GDEs), that is, species that are suspected to use groundwater but this remains unconfirmed. A complete list of species included in Figure 3-66 is provided in Appendix B. Recent work by Castellazzi et al. (2023) (Figure 3-67) proposes a more detailed, finer-resolution map of the distribution of drought-resilient vegetation across the Southern Gulf catchments than previously available. The map is formed by combining radar and optical satellite imagery products and Atlas of Living Australia (ALA) data of known locations of obligate groundwater-dependent vegetation types (Figure 3-65; Appendix B). It is more detailed and up to date than the groundwater-dependent vegetation mapping included in the GDE Atlas (Figure 3-67). Figure 3-67 shows that potential terrestrial GDEs occupy smaller areas of the landscape in the Queensland portion of the Southern Gulf catchments than indicated by the GDE Atlas. The source of groundwater used by vegetation and the timing and frequency of groundwater use remain challenges for further investigation. Groundwater in Southern Gulf catchments exists at various groundwater flow scales, ranging from local-scale flow systems (Lawn Hill Platform, South Nicholson Basin and Karumba Basin) to intermediate to regional-scale (Georgina Basin carbonate rocks, Carpentaria Sub-basin of the Great Artesian Basin sandstones; Section 2.4.1 (Raiber et al., 2023)). Figure 3-64 Distribution of potential groundwater-dependent terrestrial ecosystems in the Southern Gulf catchments Note: A buffer of 1 km has been applied to GDE mapping so that they are visible on the map scale. Only the distribution for the mainland is shown. Dataset: Bureau of Meteorology (2017) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Figure 3-65 Locations of observed obligate terrestrial GDEs in the Southern Gulf catchments Note: Only the distribution for the mainland is shown. Datasets: Atlas of Living Australia (2023a; 2023b); Department of Environment Parks and Water Security (2000); NVIS Technical Working Group (2017) Figure 3-66 Locations of facultative and potential GDE vegetation species in the Southern Gulf catchments grouped by relevant vegetation type Note: A buffer of 1 km has been applied to NT Melaleuca mapping and NVIS so that they are visible on the map scale. Only the distribution for the mainland is shown. Datasets: Atlas of Living Australia (2023a; 2023b); Department of Environment Parks and Water Security (2000); NVIS Technical Working Group (2017) Figure 3-67 Distribution of potential groundwater-dependent vegetation in Southern Gulf catchments Dataset: Castellazzi et al. (2023) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Subterranean GDEs in the Southern Gulf catchments Regional studies in Queensland have shown the existence of subterranean GDEs in limestones that occur in the upper south-western region of the Southern Gulf catchments (Bureau of Meteorology, 2017). These subterranean ecosystems are mapped as ‘known GDEs’ in the GDE Atlas (Figure 3-68). Most of the limestone system in the Southern Gulf catchments is considered potential habitat for subterranean GDEs (Bureau of Meteorology, 2017). Other aquifers also provide potential habitat for subterranean GDEs associated with fractured rock and alluvial systems in the Southern Gulf catchments (Figure 3-68). Figure 3-68 Distribution of known and potential subterranean GDEs, alluvial and karstic aquifers and caves that may provide habitat for subterranean GDEs in the Southern Gulf catchments Only the distribution for the mainland is shown. Datasets: Bureau of Meteorology (2017); Geoscience Australia (2012); Lau et al. (1987) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships and water requirements for GDEs GDEs are sensitive to changes in water quality and availability. Aquatic GDEs may be sustained by surface flows for much of the year, but when surface flows become low, they are often sustained by groundwater discharge. For some aquatic GDEs, there may be recruitment or breeding events that are exclusively triggered by groundwater discharges (this could be caused by timing and/or quality of groundwater inputs). Relationships between groundwater discharge and aquatic GDEs in northern Australia remain unknown. ‘Floodplain wetland’ and ‘inchannel waterhole’ flow–ecology relationships are reported in Sections 3.4.1 and 3.4.2 respectively. Terrestrial GDEs are often sustained by a mixture of soil water and groundwater, however some may also require periodic flooding to induce flowering and seed fall (e.g. river red gum, George (2004)) and recruitment. Groundwater requirements of terrestrial GDEs are highly variable depending on the species present, and soil and climate conditions. Surface water inundation requirements for maintaining terrestrial GDE function and services are largely unknown. However, there is some crossover between groundwater-dependent and surface-water-dependent terrestrial vegetation (Section 3.4.6), for which flow–ecology relationships are reported in Table 3-29. Most subterranean fauna have limited mobility and become stranded and die in unsaturated soils when groundwater levels drop rapidly (Hose et al., 2015). Conversely, when groundwater levels rise, subterranean fauna may not be able to rise with groundwater and become stranded in waters with insufficient oxygen to sustain them (Hose et al., 2015). The water level and quality changes that subterranean GDEs can withstand probably varies broadly with species and aquifer type, but is largely unknown. Table 3-27 specifies broad flow–ecology relationships that need to be considered when assessing the impact of changes in flow on subterranean GDEs. Table 3-24 Ecological functions supporting GDEs and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for GDEs Changes to GDEs from water resource development can occur due to a range of different processes, depending on the type of water resource development and how it is managed. This section discusses the impacts of water harvesting from groundwater and directly from rivers, dam infrastructure and regulation of river flows, climate change and land use change on aquatic (Figure 3-69), terrestrial (Figure 3-70) and subterranean (Figure 3-71) GDEs. GDEs are inherently sensitive to changes in the availability and quality of groundwater at critical times, but most are highly dependent on surface water as well. Groundwater drawdown near GDEs may result in reduced discharge to aquatic GDEs (e.g. wetlands, rivers), reduced connection between groundwater and dependent vegetation (terrestrial GDEs) or loss of subterranean GDEs altogether. Surface water harvesting, river regulation, dam infrastructure, climate change and land use change can all disturb the natural groundwater recharge regime, altering the depth to water, the seasonal cyclicity of groundwater levels, and groundwater quality. In areas where groundwater recharge is reduced, the impacts on GDEs over the long term are similar to those of groundwater drawdown. In areas where groundwater recharge is enhanced, there could be: • local increases in groundwater discharge to aquatic GDEs. In some areas this can be a source of high salt loads to surface water systems (e.g. Jolly et al. (1993)) that potentially increase the longitudinal connectivity along rivers during the dry season, putting pressure on some aquatic GDE species and potentially favouring non-native aquatic species (e.g. Yarnell et al. (2015)) • shallower groundwater levels, potentially leading to soil salinisation due to evapotranspiration from shallow watertables (e.g. Smith and Price, 2009) and/or a shift in the type of terrestrial vegetation supported (e.g. from Melaleuca swamp to grassland (Department of Environment and Science Queensland, 2013)) • potential mortality of subterranean GDEs in anoxic waters if stygofauna lose connection with relatively aerated water at the top of the watertable (Hose et al., 2015). Most aquatic and terrestrial GDEs require surface water in addition to groundwater to sustain their water requirements. Activities that affect the volume, timing, frequency and quality of surface water flows or inundation are likely to affect aquatic GDEs and fringing vegetation. The ecological outcomes of threatening processes on aquatic, terrestrial and subterranean GDEs in northern Australia, and their implications for changes to biodiversity and ecosystem function, are illustrated in Figure 3-69, Figure 3-70, and Figure 3-71. Figure 3-69 Conceptual model showing the relationship between threats, drivers, effects and outcomes for aquatic GDEs in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-70 Conceptual model showing the relationship between threats, drivers, effects and outcomes for terrestrial GDEs in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-71 Conceptual model showing the relationship between threats, drivers, effects and outcomes for subterranean GDEs in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.4.3 Inchannel waterholes Description and background to ecology Rivers located in northern Australia’s wet-dry tropics are subject to highly seasonal rainfall, often resulting in high wet-season flows, and low dry-season flows (Close et al., 2012; Petheram et al., 2008). During the dry season, many rivers cease to flow and can retract to a series of discrete and disconnected waterholes (McJannet et al., 2014; Waltham et al., 2013). In ephemeral river systems, the waterholes that retain water for periods sufficient to outlast dry spells provide vital refuge habitat and resources for both flora and fauna (Sheldon, 2017). Waterholes are also an important social resource, particularly during the dry season, by providing places for recreation as well as providing cultural functions (Centre of Excellence in Natural Resource Management, 2010; McJannet et al., 2014). Waterholes provide direct habitat for water-dependent species, including fish, sawfish and turtles, and a source of water for other species more broadly within the landscape (McJannet et al., 2014; Waltham et al., 2013). Larger more-stable waterholes that retain water during extended dry periods also often support a vibrant riparian vegetation community that often further enhances Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. the habitat value of waterholes, with vegetation often assisted through having more reliable groundwater (see aquatic GDEs, Section 3.4.2). Once river flows recommence and reconnect aquatic habitats in the early wet season, waterholes act as habitat sources for recolonisation of other parts of the catchment (Garcia et al., 2015; Lymburner and Burrows, 2008). Areas with a higher number of persistent waterholes, and often those with a range of different habitat characteristics, are recognised as enhancing biodiversity at regional scales and are considered as important refuge habitat during the dry season (Arthington et al., 2010; DERM, 2011). Despite their comparatively small contribution to the total area of the catchment, waterholes often provide high habitat value with often disproportionately high biodiversity values. For the purpose of this Assessment, waterholes are defined as locations within river channels or watercourses that retain water during periods of low or no flow. This definition excludes large lakes and storages, and it pertains to areas of retained water occurring within often-disconnected locations within the river channel, rather than on the floodplain or in the estuary (alternatively, see Floodplain wetlands, Section 3.4.1). Waterholes can include bodies of water occurring in main channels, braided channels or oxbows, with persistence maintained due to the size or position of the waterhole or in some locations through connection to contributions such as groundwater inflows (also see GDEs in Section 3.4.2). Inchannel waterholes in the Southern Gulf catchments Larger and more persistent inchannel waterholes occur throughout many parts of the Southern Gulf catchments, including sections of river around Lawn Hill, Gregory and Mount Isa (Figure 3-72). The highly anabranching channels lower in the catchment (towards the Gulf of Carpentaria) limit the size of waterholes that persist and that can be identified by remote sensing to all but the larger channels. In many locations, persistent waterholes are supported by groundwater discharge that maintains an often-significant level of baseflow during periods that would otherwise result in highly reduced flow or cease-to-flow conditions. Permanent or near-permanent flowing river sections and springs that originate from the limestone plateau feeding into Lawn Hill Creek include Settlement Creek (Bureau of Meteorology, 2017) and west of Gregory in Boodjamulla (Lawn Hill) National Park. In areas other than these however, many tributaries demonstrate the seasonal ephemeral flows that are broadly characteristic of northern Australian rivers (Petheram et al., 2008). In these ephemeral reaches, waterholes that persist provide important habitat values. In the Southern Gulf catchments, these biodiversity values are highlighted by the waterholes providing habitat for species listed under the EPBC Act, including the freshwater sawfish (Vulnerable; Section 3.1.6). Figure 3-72 Location of persistent inchannel waterholes in the Southern Gulf catchments Waterholes are mapped at the end of each dry season using Landsat imagery as described in Sims et al. (2016). For this, an inchannel mask containing a 500 m buffer from the watercourse is divided into 200 m segments along each watercourse. The percentage of dry seasons containing at least one pixel of water within each 200 m segment is calculated to allow for the fact that a waterhole can vary in shape and location through time. Data source: Sims et al. (2016) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for inchannel waterholes Waterholes are sensitive to changes in low-flow magnitudes, low-flow duration, periods of cease to flow, and timing of first wet-season inflows (Table 3-25). The habitat conditions within waterholes and the persistence of waterholes within the landscape decline where the duration of low-flow periods is extended, where water is removed from the river during low flows, or where water is extracted directly from waterholes. Where water is extracted, waterholes are increasingly prone to drying out, resulting in a loss of habitat quality and extent, reduced water quality, and changes in competition and food web structure for biota. The timing of a first-flow pulse is important for breaking the dry period, improving water quality and reconnecting habitats. Similarly, conversion of ephemeral systems to perennial systems due to dam or barrier construction will alter the cycle of ephemeral systems and change the natural habitat conditions as low flows and cease-to-flow conditions are important for maintaining ecosystem function, including habitat partitioning and limiting habitat suitability and persistence of non-native species (Yarnell et al., 2015). Infrequent large flows are likely important for maintaining structure within the waterhole. The ecological functions that support inchannel waterholes, and their associated flow requirements, are summarised in Table 3-25. Table 3-25 Ecological functions supporting inchannel waterholes and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for inchannel waterholes Changes to waterholes resulting from water resource development can occur due to a range of different processes, depending upon the type of water resource development and how it is managed. This assessment considers upstream water capture and storage, water harvesting direct from flows within the channel and climate change (Figure 3-73). Changes in the flow regime associated with upstream water capture and storage, surface water and groundwater extraction, and rainfall and higher evaporation due to climate change have the potential to reduce inflows and influence the natural filling and drying cycles of waterholes (Arthington et al., 2010; McJannet et al., 2014; Waltham et al., 2013). Waterholes are likely to be particularly sensitive to changes in the duration and severity of dry periods and changes in the timing of first flushes and inflows. Other drivers to waterhole persistence and quality can include use of groundwater that results in reduced inflows or faster drawdown of waterholes. Maintaining the quality of waterhole habitat during periods of low flow is crucial for the local persistence of many of aquatic species (Department of Environment and Resource Management, 2010). Lower dry-season flows resulting in longer periods of low flows due to water resource development threaten to reduce the habitat value of waterholes. This can occur due to loss of waterholes within the landscape or decreases in the condition of the waterholes that remain (Department of Environment and Resource Management, 2010). Capturing or harvesting water upstream, or extracting water directly from the waterhole, can lead to drawdown or early loss of the waterhole from within the landscape (McJannet et al., 2013). This may result in a localised loss of dependent biota (both aquatic and terrestrial) and the loss or degradation of habitat (McJannet et al., 2014). Where loss of waterholes occurs more frequently within the landscape, it has the potential to result in biodiversity impacts from local to more regional scales across the catchment (James et al., 2013). The number, size and heterogeneity of waterholes is considered important for sustaining biodiversity at larger spatial scales. Modification of the current duration or timing of low-flow or cease-to-flow periods threatens to change the ecological character of waterholes. During cease-to-flow events, when no surface water enters waterholes, species lose pathways for movement, including longitudinal connectivity along the river channel important to biotic movement. In addition, water quality often deteriorates due to lower exchange along the watercourse. During periods of low inflows, waterhole area is reduced resulting in the loss of important ‘slide’ and riffle habitat, or potential loss of entire waterholes. The location of individual waterholes within the catchment is an important contributing factor to the duration of the cease-to-flow period, with waterholes in upper catchments more likely to undergo prolonged periods of disconnection under current conditions (Pollino et al., 2018a). Waterholes persist owing to the hydrological balance within the system that results from the timing and duration of both filling events and drawdown (Close et al., 2012). While loss of waterholes can cause a range of impacts, the alternative of an increase in the persistence of waterholes may also have environmental impacts. Increased water persistence could occur due to the construction of instream barriers, and as a result of persistent or unseasonal releases from upstream storages. Increases in waterhole persistence can alter the natural system to which the flora and fauna are adapted, with possible impacts on habitat structure, water quality, productivity and food web complexity. For example, shifts in the characteristics of waterholes may change predator–prey balances, reduce predator-free habitat for communities of smaller fish species due to loss of many smaller waterholes, or cause conditions to change to those favouring non-native species (McJannet et al., 2013; Yarnell et al., 2015). The species in each catchment have adapted to the range of conditions that result from the climate and geomorphology of the system. Changes in the range of conditions experienced during the dry or wet seasons, or the transitions between seasons, can alter the species composition of a region. Decreases in flow during the wet season result in loss of connectivity and decreases in flow during the dry season result in loss of critical refuge habitat. Also, homogenisation and loss of the extent of seasonal variation changes the environment to which species have adapted. Waterholes are typically surrounded by riparian vegetation, which offers shade and structural diversity and acts as an interface between aquatic and terrestrial ecosystems. Changes in waterhole permanence could affect the plants providing this habitat at local and regional scales. Pest species such as buffalo and pigs, and unrestricted cattle access to waterholes, can damage riparian vegetation and increase sedimentation, turbidity and nutrients within waterholes. Changes in the condition or persistence of waterholes could also provide a competitive advantage to non-native fish species. Invasive species are recognised to often be at an advantage in modified habitats (Bunn and Arthington, 2002). Modified landscapes such as lakes or newly created perennial streams can create habitat for the establishment of pest plant and fish species or be a source for their introduction, whether incidental, accidental or deliberate (Close et al., 2012; Ebner et al., 2020). The ecological outcomes of threatening processes on inchannel waterholes in northern Australia, and their implications for changes to biodiversity and ecosystem function are illustrated in Figure 3-73. Figure 3-73 Conceptual model showing the relationship between threats, drivers, effects and outcomes for inchannel waterholes in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.4.4 Mangroves Description and background to ecology Mangroves are a group of woody plant species, ranging from shrub to large tree to forest, that are highly specialised to deal with daily variation in their niche within the intertidal and near-supra- littoral zones along tidal creeks, estuaries and coastlines (Duke et al., 2019; Friess et al., 2020; Layman, 2007). Their occurrence is a result of changes across temporal scales from twice-daily tides to seasonal and annual cycles; mangroves have acclimatised to variable inundation, changing salinity, anoxic sediments, drought and floods, and sea-level change. Mangrove forests provide a complex habitat that offers a home to many marine species including molluscs (McClenachan et Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. al., 2021), crustaceans (Guest et al., 2006; Thimdee et al., 2001), birds (Mohd-Azlan et al., 2012), reptiles (Fukuda and Cuff, 2013) and numerous fish species. During periods of inundation at high tide, fish and crustaceans access mangrove forests for shelter against predation. Fish and crustaceans use mangroves as refugia during larval phases and settle there as benthic juveniles (Meynecke et al., 2010) or access them for food (Layman, 2007; Skilleter et al., 2005). Mangrove forests support many of the species and groups reported as biota assets in this report, particularly commercial species such as banana prawns (Section 3.3.1), (Section 3.3.1), barramundi (Section 3.1.1), mud crabs (Section 3.3.4), threadfin (Section 3.1.7) and mullet (Section 3.1.5) (Blaber et al., 1995; Brewer et al., 1995). Mangrove forests provide a diverse array of ecosystem services, including stabilising shoreline areas from erosion and severe weather events (Zhang et al., 2012), and they play an important role in greenhouse gas emission and carbon sequestration (Lovelock and Reef, 2020; Owers et al., 2022; Rogers et al., 2019). Mangroves continually shed leaves, branches and roots, contributing from approximately 44 to 1022 g carbon per square metre per year from leaves and 912 to 6870 g carbon per square metre per year from roots, though these rates continue to be explored (Robertson, 1986; Robertson and Alongi, 2016). Intertidal crabs living in mangrove forests play an important role in processing and storing mangrove carbon, either through burial in their burrows or uptake directly into production. The decomposition and processing of mangrove material is important also in the cycling of nutrients. If consumed and released, these nutrients support a local food web (Abrantes et al., 2015; Guest et al., 2004),and some of the organic carbon can be transported offshore where it supports fisheries production more broadly (Connolly and Waltham, 2015; Dittmar and Lara, 2001; Lee, 1995). Mangroves in the Southern Gulf Catchments and marine region Lymburner et al. (2020) mapped the extent of mangroves in Australia using 25 m spatial resolution Landsat 5 (TM, ETM, OLI) sensor data, finding an area of 11,142 ± 57 km2 (95% confidence interval (CI)) in 2017, which is down slightly from the 2011 extent of 11,388 ± 38 km2 (95% CI). Most of the change was found to have occurred along the northern Australian coastline and be concentrated in major gulfs and sounds. While coastal urban and industrial development can result in direct loss of coastal wetland ecosystems, including mangroves (Firth et al., 2020; Murray et al., 2022), climate change has also notably caused mangrove loss in northern Australia. The most significant and obvious example was the dieback event between late 2015 and early 2016 along more than 1000 km of coastline in the Gulf of Carpentaria (Duke et al., 2017). Mangrove communities in the Southern Gulf catchments marine region were affected by the dieback. Mangroves in the Southern Gulf catchments are restricted to along the coastline and a narrow fringe lining both sides of connecting tidal channels and main estuaries in the Southern Gulf catchments marine region, and as shown in Figure 3-74 (Short, 2020). The distribution of mangroves in the Southern Gulf catchments marine region was reduced following extensive dieback of mangroves between late 2015 and early 2016, particularly along the coastline (Duke et al., 2017). Fish and crustacean communities occupy mangrove forests in the Southern Gulf catchments marine region (Staples and Vance, 1987), though community species’ composition is poorly studied. It is likely that most of the same species occur in these estuaries as elsewhere in the Gulf of Carpentaria (Blaber et al., 1995; Brewer et al., 1995). Species such as mud crab (Scylla serrata) are highly associated with the mangrove community (Robins et al., 2020), as are other fish species that access the mangrove forests during periods where tidal connection permits access, presumably for shelter and food – which is similar to the east coast of Queensland (Sheaves and Johnston, 2009; Sheaves et al., 2016). While the extent of mangrove forests in this catchment area is relatively small compared to the extent of intertidal saltpans (Section 3.4.5), they still provide important linkages to coastal fisheries production when connected to the estuary, in addition to providing erosion protection, sediment accumulation and carbon sequestration services. Figure 3-74 Location of mangroves in the Southern Gulf catchments marine region Data source: Department of Climate Change‚ Energy‚ the Environment and Water (2020), Geoscience Australia (2017) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for mangroves Large flood events from rivers mobilise catchment sediments and deliver them to the coastal zone. This can be detrimental to some coastal habitats (e.g. seagrass beds can be smothered when sediments inhibit sunlight penetration through the water column). In mangrove forests, while sediment delivered to the coast can also smother mangrove root systems, sediment accumulation is generally considered beneficial as it assists with habitat substrate stability and the accumulation of carbon in sediments for the benefit of benthic foragers (Owers et al., 2022). Asbridge et al. (2016) suggest that Gulf of Carpentaria mangroves have expanded seaward in recent years and that without sediment replenishment these mangrove forests would erode. The hydrology of mangroves is complex; it is influenced by tidal inundation, rainfall, soil water moisture content, groundwater seepage and evaporation, all of which influence soil salinity that can have profound effects on mangrove growth and survival. Mangroves require access to fresh water, though many species are found at the upper salinity threshold (Robertson and Duke, 1990). A challenge for mangroves is when soil water content changes, and they can be greatly affected if soils dry out and the water content reduces. The large mangrove dieback in the Gulf of Carpentaria is an example of when soil water content was low as a consequence of lower sea levels and mangroves were not able to access water (Duke et al., 2017). Mangroves are connected to the sea and estuaries via tidal inundation, which rehydrates soils; the only other time soils become waterlogged is during rainfall or wet-season flow, which recharge soil water and groundwater in mangrove forests (Duke et al., 2019). Altered freshwater flow in catchments that previously caused rivers to overtop their banks and spread across coastal floodplains could therefore contribute to mangrove stress and potentially die back. The ecological functions and their supporting flow requirements for mangroves are summarised in Table 3-26. Table 3-26 Ecological functions supporting mangroves and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for mangroves Several threatening processes can affect mangroves in northern Australia, including river regulation, water extraction, climate change and land use change. The ecological impacts of dams and river regulation can be numerous; most obviously they can prevent water from flowing onto floodplains by capturing large rainfall events, preventing flood pulses from moving down catchment and reaching dynamic estuaries and near-shore coastal areas. This loss of connectivity to the coastal floodplain areas, including mangrove forests, can result in the reduction or loss of coastal wetland vegetation areas. This was the case in the Gulf of Carpentaria mangrove dieback − while not driven by river regulation, it was a consequence of an unusually lengthy period of severe drought conditions, unprecedented high temperatures and a temporary drop in sea level (Duke et al., 2017). In this extreme event, high temperatures resulted in mangrove dehydration and death – they could not access freshwater sources during critical periods of high summer temperature (Duke et al., 2017). River regulation can disrupt the natural flow regime. The alteration of the magnitude, frequency, duration, timing and rate of change of flows within a system can affect all aspects of a riverine and floodplain ecosystem (Abrial et al., 2019; Chemagin, 2019; Poff and Zimmerman, 2010b). For mangroves these changes can include impacts on the structure, function, sedimentation and biodiversity of mangrove communities. Building dams and other hydrological barriers also affects mangrove forests by choking off sediment loading while increasing nutrient pollution (Godoy et al., 2018). Sedimentation, for example, is critical for the protection of mangrove forests; without sediment supply from river catchments, they would erode (Asbridge et al., 2016). Groundwater extraction may reduce spring flows and lower the watertable. Coastal wetlands, including mangroves, are particularly vulnerable to climate change (Feller et al., 2017). Climate change impacts may include accelerated sea-level rise, changes in freshwater inputs, and changes to the frequency and intensity of storms and storm surges (Day et al., 2008; Nicholls et al., 1999). Sea-level rise and a decrease of freshwater inputs can lead to the saltwater intrusion of wetlands (Close et al., 2015; White and Kaplan, 2017), which in turn can result in the loss or retreat of mangroves and the conversion of freshwater floodplains to estuarine ecosystems (Duke et al., 2019; Finlayson et al., 1999). Changes in rainfall, runoff and evapotranspiration patterns (Grieger et al., 2020; Salimi et al., 2021) affecting the hydrology of a system can alter the baseflow and flood patterns (Erwin, 2009). These hydrology changes can also affect the water quality through, for example, increased erosion and changes to temperature (Erwin, 2009). Drought and a lower sea level have been shown to be the cause of mangrove loss in the Gulf of Carpentaria in 2015, and the same event has been reported elsewhere in northern Australia (Duke et al., 2017; Lovelock et al., 2017). Changes to hydrology and temperature can affect the biodiversity ecosystem services that mangroves provide (Dudgeon et al., 2006; Finlayson et al., 2006; Mitsch et al., 2015). Land use change is a major threat to the extent and fragmentation of mangroves, and there are many examples of mangrove loss in developing areas (Xu et al., 2019). Land use change has contributed to loss of mangroves directly or because of changes in hydrology and flow, causing increased erosion. Changes includes modifying land management practices; changing the intensity or type of agricultural production; increasing vegetation clearing; and increasing mining, urbanisation and industrial development. Evidence of landward expansion of mangroves has been documented (Armitage et al., 2015), but this expansion can only occur where there is sufficient space, and it will be restricted by hard engineering structures or urbanisation that prevents this expansion (Doody, 2004; Leo et al., 2019). The loss of extent or fragmentation of mangroves as a direct result of land use changes or deforestation can reduce carbon sequestration stock (Atwood et al., 2017). In addition, mangroves, when inundated with tidal water, provide critical nursery habitat for local species, including commercial fishery species that would also be affected by the loss of mangroves (Sheaves et al., 2016). The ecological outcomes of threatening processes on mangroves in northern Australia, and their implications for changes to biodiversity and ecosystem function, are discussed below and illustrated in Figure 3-75. The impact of water resource development, such as dam construction and several levels of pumped water extraction, on Gulf of Carpentaria mangrove communities has been modelled using predicted streamflow data generated under water resource development scenarios (Plagányi et al., 2023). Mangrove biomass declined by up to approximately 40% in some river estuaries (predicted average declines of 26, 28 and 44% for Mitchell, Flinders and Gilbert river systems, respectively, under a ‘high extraction’ scenario). The assessed risk to the mangrove community ranged from negligible to severe across the four water resource development scenarios (Plagányi et al., 2023). Both the construction of dams and the harvest of river flows via pumped water extraction affect mangrove community stability and replenishment, reducing mangrove community resilience. Anthropogenic reduction in the volume and duration of high-level flows, as well as induced variability in the seasonality and volume of low-level flows, affect freshwater delivery to the mangrove community and its survival, particularly during the latter dry season when water stress is high (Duke et al., 2017; Plagányi et al., 2023). Figure 3-75 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mangroves in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 3.4.5 Saltpans and salt flats Description and background to ecology Saltpans and salt flats are intertidal areas that are devoid of marine plants and are located between mangroves and saltmarsh meadows. Saltmarshes (Figure 3-76) occur in the supra-littoral zones that is inundated infrequently by the tide and where subsequent water evaporation leaves behind expanses of minerals and salts (Cotin et al., 2011). Despite their infrequent inundation, saltpans provide habitat for some estuarine fish, such as barramundi (Russell and Garrett, 1983), and shrimps of the genus Metapenaeus (Bayliss et al., 2014) during periods when the tide covers these habitats. Inundation of saltpans mostly occurs during the annual wet season when large tides and rainfall surface runoff ponds as shallow wetted areas within the saltpans and shallow tidal-cut gutters that intersect them. In northern Australia, saltpan sediments are infused with dormant algae that remain inactive in a desiccated state during the dry season (most of the year). However, during overbank inundation from flooded rivers or extensive rainfall, the saltpan soil algae become active and photosynthesise and increase nutrient contribution to the ecosystem (Burford et al., 2016). After several days, active algal growth occurs and carbon, nitrogen and phosphorous compounds are produced. Estimates suggest that saltpans can contribute an extra 0 to 13% of ecosystem primary productivity depending on the extend of saltpan inundation during the wet season (Burford et al., 2016). Saltpans would be most productive during high-level overbank flood flows. The inundation of saltpans expands the available habitat to estuarine benthic fish and crustaceans, provided they can tolerate brackish conditions. In northern Australia, coastal saltpans can extend tens to hundreds of square kilometres. They provide habitat for a range of benthic fauna (Dias et al., 2014), which are an important food source for high-order consumers including shorebird species that use saltpans as resting and/or feeding areas during their migration, which can include long flights to Asia (Cotin et al., 2011; Lei et al., 2018; Rocha et al., 2017). The extent of saltpans in Australia is unknown, though they are common and extensive in more arid coastal areas, most notably in northern Australia (Duke et al., 2019). The northern Australian coastline extends for thousands of kilometres and is relatively pristine; low beach profiles backed by extensive saltpans, possibly 5 to 10 km inland, are characteristic of hundreds of kilometres of coastline (Short, 2022). Despite limited tidal exchange, saltpans provide important habitat resources for migratory birds (see shorebirds in Section 3.2.3) that access these areas for feeding and shelter (Lei et al., 2018). In addition, these habitat features also provide erosion and sediment accumulation opportunities in estuaries as well as carbon sequestration services. Figure 3-76 Saltpan area in northern Australia, which are generally located between mangrove and saltmarsh areas Photo attribution: Nathan Waltham Saltpans in the Southern Gulf catchments and marine region Saltpans in the Southern Gulf estuaries are extensive behind tide-dominated beaches as shown in Figure 3-77 (Short, 2020). They are mostly restricted to a tidal inundation area on the landward side of the mangroves that line the main river channel but also occur adjacent to Buffalo and Sweet swamps. The spatial data presented illustrates the extent of saltpans in the rivers, but it is important to note that the extent presumably increased as part of the wide-scale dieback of mangroves in the Gulf of Carpentaria between late 2015 and early 2016 (Duke et al., 2019; Duke et al., 2017). There has been no targeted scientific survey of fish and crustacean communities over the saltpans of the Southern Gulf catchments marine region, presumably because they are located so high in the intertidal zone and are only infrequently covered with tidal water. Also, the catchment is remote, which leads to difficulties with access and sampling. Similar to saltpans elsewhere in northern Australia, these saltpans provide important habitat opportunities for many species, including fish and crustaceans during inundation from tides or floods, and migratory birds also use many of them for resting, feeding and shelter. Photo of saltpan. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-77 Location of saltpans in the Southern Gulf catchments marine region Data source: Geoscience Australia (2017) For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Flow–ecology relationships for saltpans in northern Australia The hydrology of saltpans is complex. Tidal inundation, rainfall, soil water, groundwater seepage and evaporation all influence soil salinity, which can have profound effects on the services saltpans provide in the seascape. A great challenge to the flora and fauna found on saltpans is change in soil water content, particularly if soils dry out and the moisture content reduces, which causes these areas to become hypersaline in the surface soils. Saltpans are connected to sea and estuaries via infrequent tidal inundation, which rehydrates soils. The only other time soils become waterlogged is during rainfall or wet-season flow, which recharges soil water and groundwater (Duke et al., 2019). Altered freshwater flow in catchments that would otherwise have caused rivers to overtop their banks and spread across coastal floodplains could contribute to wide-scale impacts on the services provided by these habitat resources. The ecological functions that support saltpans, and their associated flow requirements, are summarised in Table 3-27. Table 3-27 Ecological functions supporting saltpans and their associated flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for saltpans Several threatening processes can affect the saltpans in northern Australia, including river regulation, water extraction, climate change and land use change. The ecological impacts of dams and river regulation can be numerous; most obviously they can prevent water from flowing onto floodplains by capturing large rainfall events, preventing flood pulses from moving down catchment and reaching dynamic estuaries and near-shore coastal areas. This loss of connectivity to coastal floodplain areas, including saltpans, can result in the reduction or loss of coastal wetland areas (Lei et al., 2018; Velasquez, 1992). Extraction of surface water generally has less impact on the environment than instream storages, as surface water extraction tends to occur during high-flow events such as floods rather than during low-flow periods (Petheram et al., 2008). As a result, water extraction can lower the peak of a flood, allowing less water for the environment. The reduction in peak flow can decrease the duration and extent of a flood event, and can also prevent overbank flooding altogether (Kingsford, 2000). Coastal wetlands are particularly vulnerable to climate change (Feller et al., 2017). Climate change impacts include accelerated sea-level rise, a change in freshwater inputs, and changes to the frequency and intensity of storms and storm surges (Day et al., 2008; Nicholls et al., 1999). Sea- level rise and a decrease of freshwater inputs can lead to the saltwater intrusion of wetlands (Close et al., 2015; White and Kaplan, 2017), which in turn can result in the conversion of freshwater floodplains to salt flats (Duke et al., 2019; Finlayson et al., 1999). In the Gulf of Carpentaria, a dieback of mangroves occurred along a large stretch of the coast. This dieback was a response to low rainfall and freshwater runoff from catchments, warmer temperature conditions and a lower sea level than typical during the summer wet season in the Gulf of Carpentaria (Duke et al., 2019; Duke et al., 2017). Asbridge et al. (2016) described the replenishment of mangrove habitats due to natural flows in the southern Gulf of Carpentaria, supporting the idea that reduction in flow may reduce sediment loads and set up conditions for erosion of mangrove foreshores and possibly the saltpan habitats behind them. The loss of saltpan extent or fragmentation as a direct result of land use changes or sea-level rise can reduce carbon sequestration stocks (Atwood et al., 2017). In addition, saltpans when inundated with tidal water provide critical nursery habitat for local species, including commercial fishery species that would also be affected (Sheaves et al., 2016). Invasive species, such as feral pigs, and vehicles driving across saltpans can also change the habitat quality directly through trampling or digging and tyre tracks left behind, which has the potential to alter hydrological connectivity of saltpans with river channels (Trave and Sheaves, 2014; Vulliet et al., 2023; Waltham et al., 2020). Changes in this connectivity could alter soil water and leave saltpans degraded and of low-quality habitat for migratory birds (Duke et al., 2019). The ecological outcomes of threatening processes on saltpans in northern Australia, with their implications for changes to biodiversity and ecosystem function, are illustrated in Figure 3-78. Figure 3-78 Conceptual model showing the relationship between threats, drivers, effects and outcomes for saltpans in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 3.4.6 Seagrass habitats Description and background to ecology Seagrasses are marine flowering plants that play a vital role providing valuable habitat and food resources to a diverse community of marine animals, including invertebrates, fish, dugongs, sea turtles. In the tidal reaches of rivers and coastal, reef and deep-water habitats in northern Australia, there are 15 species of seagrasses. Their distribution is limited by light availability and suitable substrate; most seagrasses live in shallow inshore and intertidal zones extending to water depths of approximately 25 m (though they can occur in depths of 50–60 m) (Carruthers et al., 2002). Although seagrass species are not currently listed for conservation, they serve as crucial habitat for threatened species such as dugongs and various marine turtle species, including the green (Chelonia mydas), flatback (Natator depressus), olive ridley (Lepidochelys olivacea) and hawksbill (Eretmochelys imbricata) turtles. Some commercially important species, such as prawns, are also highly associated with seagrass habitats. In addition to being a critical food for dugongs and some species of marine turtles (Carruthers et al., 2002; Morgan et al., 2017a), seagrasses provide a substrate for epiphytic algae, which provide a basal resource for marine food webs (Moriarty, 1990). Detrital seagrass biomass is an important food source for species in the Northern Prawn Fishery (NPF), including banana and tiger prawns (sections 3.3.1 and 3.3.5, respectively). Juvenile prawn production in the north-east Gulf of Carpentaria seems to be mostly based on seagrass-derived organic matter (Loneragan, 1997). Seagrasses offer shelter from currents and predation while stabilising bed sediments, crucial for providing nursery habitat for juvenile fish and prawns, including commercially important species (Coles, 2004; Roelofs, 2005; Unsworth, 2019). Additionally, seagrasses contribute to coastal morphology by reducing water velocity near the bottom, promoting sediment accretion and therefore, importantly, reducing erosion (Fonseca and Cahalan, 1992). Globally, seagrasses are recognised as one of the most valuable ecosystems, providing key ecological services such as climate regulation (Duarte et al., 2013), carbon sinks (Fourqurean et al., 2012) fisheries habitat (Unsworth, 2019), nutrient cycling, enhanced biodiversity and sediment stabilisation (Orth, 2006). Despite their ecological importance, seagrasses worldwide are experiencing rapid declines, with a loss of 29% of their known areal extent since 1879 (Waycott et al., 2009). This loss is attributed to various threats including rising sea-surface temperatures, extreme temperature events, coastal development, coastal urban and agricultural runoffs, and untreated sewage and industrial waste outfalls (Arias-Ortiz et al., 2018; Freeman et al., 2008; Grech et al., 2012). Seagrass habitats in the Southern Gulf catchments Australia’s seagrass meadows constitute a large proportion of the planet’s known seagrass species (Green and Short, 2003): just over half (37) of the world’s 72 seagrass species are found in Australian waters (Kilminster et al., 2015; Short et al., 2011). The distribution of seagrasses in the Gulf of Carpentaria is fragmented, characterised by aggregated seagrass patches with many bare areas between them (Figure 3-79). Relatively few meadows have continuous seagrass cover, although these tended to also be large meadows (Roelofs et al., 2005). Figure 3-79 Distribution of seagrass habitats in the Southern Gulf catchments marine region Data sources: Atlas of Living Australia (2023a; 2023b); Department of Agriculture Water and the Environment (2019); Department of Environment and Science (2018; 2021; 2023); Department of Environment Parks and Water Security (2019c); OBIS (2023) Seven seagrass specieshave been recorded inthe Southern Gulf catchments, all listed as ‘Special least concern’underQueensland’sNature Conservation Act1992. There have been someexpeditions to this catchment to identify and map seagrass species.Poiner et al. (1987)showed that 74% (661.2km2) of the seagrass occurred along open coastlines in depth-zoned distributions, and each zonewas dominated by one or two species (includingSyringodium isoetifolium, Halophila ovalis, Halodule uninervis, Cymodocea serrulata andHalophila spinulosa).The subtidalzone was dominated byS. isoetifoliumandCymodocea serrulata.Otherhabitats hadmixed- species meadows of Cymodocea rotundataand Thalassia hemprichiion reefflats, while somesmall sheltered embaymentswere dominated byEnhalus acoroidesand Halophila ovalis(Poiner etal., 1987;Roelofset al., 2005). Monospecificcommunities of H. uninerviswere found in the shelterofthe large islands, principally the Sir EdwardPellew Group. During an expedition in 2005, Roelofs mapped intertidal seagrassmeadows in coastal areasbetween eastern Van Diemen Gulf in theNT andHorn Island in north Queensland(Roelofs et al., 2005). Roelofsdescribed the coastline fromthe NT border to Aurukun asdominated by open barecoasts, turbid water and a few areas of extensiveintertidal seagrass. Those meadows containedallthe species listed aboveexcept Cymodocea rotundata. Most meadows across the Southern Gulf catchments marine region were aggregatedpatches; there wererelatively fewcontinuousmeadows. Previous studies have also identified significant areas of seagrassnear BamagaandatMornington and WellesleyislandsandPort Musgrave (Coles et al.,2001). Flow–ecology relationships for seagrass habitats River regulation can changethe waterquality of flows discharging into coastal regions, adverselyaffectingseagrasses,particularlywith elevatedturbidity resulting from larger flood pulses(Turschwell et al.,2021b). Turbiditylevels areinfluenced by runoff, basinmanagement and plumemovement, among otherfactors.Large turbidflood pulsescanpotentiallysmother seagrass orinhibit penetration of sunlight.Thedetrimental effects on seagrassbedsof substantialsedimentation eventstriggeredby large floodsfrom riverscan lead to shifts in theirdistribution and vigour. The ecological functions and theirsupporting flow requirements forseagrasshabitatsaresummarised inTable3-28. Table3-28Ecological functions for seagrass habitats and their supporting flow requirements ECOLOGICAL FUNCTIONREQUIREMENTFLOWCOMPONENTORATTRIBUTE Clear waters to support Negative associationwithInfrequent large discharges from end- photosynthesissedimentation from rivers and of-systemturbid watersRisk of sedimentation covering Negative associationwithFor more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. seagrass habitatslarge sedimentation events of-systemfrom rivers Chapter3 Ecological assets from the Southern Gulf catchments and marine region|179 Pathways of change for seagrass habitats Four key threatening processes have been evaluated as having negative impacts on seagrass species in northern Australia: land use change, climate change, river regulation and demersal fisheries (Turschwell et al., 2021b) (Figure 3-80). While not all seagrass species are equally affected by the same pressures, destructive demersal fishing and poor water quality have shown a strong association with a rapidly declining extent of seagrass meadows (Turschwell et al., 2021b). Agricultural practices leading to sediment and nutrient runoff and reduced light available for photosynthesis are the principal contributors to declines in seagrass extent (Grech et al., 2012; Quiros et al., 2017). Increasing nutrient loads also influence changes in seagrass community composition, due to the varying responses of different species to these changes (Burkholder et al., 2007; Cardoso et al., 2004). Specific herbicides and pesticides commonly used in northern Australia are toxic to seagrasses (e.g. Flores et al. (2013); Negri et al. (2015); Wilkinson et al. (2015)). These herbicides are frequently detected in waters of the Great Barrier Reef (e.g. Flores et al. (2013); King et al. (2013); Negri et al. (2015); Waterhouse et al. (2012)), but there is limited knowledge of the herbicide use in the study catchments. Recent studies on the impact of microplastic pollution are emerging (Seng et al., 2020). Short and Neckles (1999) and Duarte (2002) reviewed the impacts of climate change on seagrass and found that climate-change-induced sea-level rise and temperature changes might significantly affect seagrass species and their distribution. If marine depth increases, seagrasses and other organisms like corals and some macroalgae might then receive too little light. However, since seagrasses can migrate to new areas, including to newly inundated areas, their distributions can potentially change. Saunders et al. (2013) modelled the potential impacts of sea-level rise on seagrasses in Moreton Bay (south-east Queensland), demonstrating that benthic irradiance (which is closely correlated with depth) and wave height can map seagrass extent with 83% accuracy. Beaman et al. (2016) showed that water depth is a key determinant of benthic habitat in the Coral Sea. Similar relationships have been demonstrated in marine habitats worldwide. Regarding temperature changes, the relationship between seagrass metabolism (photosynthesis and respiration) and temperature is well established in the literature (e.g. Lee et al. (2007); Masini et al. (1995)). Under experimental conditions, it has been observed that some tropical seagrasses are more susceptible to toxins when subject to thermal stress (Koch and Erskine, 2001). Temperature changes can lead to modifications in the seagrass community composition; some species thrive at higher temperatures while others cannot survive elevated temperatures (e.g. Campbell et al. (2006); Evans et al. (1986); (McMillan, 1984)). River regulation can affect water quality. Seagrasses are generally intolerant of fresh water for extended periods of exposure (although tolerance varies by species) (Adams and Bate, 1994; Collier et al., 2014). They can be harmed by direct exposure to flood plumes (Collier et al., 2014) or by sedimentation and high levels of turbidity associated with large flood events and their discharges (Turschwell et al., 2021b). The ecological outcomes of threatening processes on seagrass habitats in northern Australia, and their implications for changes to population viability and community structure, are discussed below and shown in Figure 3-80. The potential impact of water resource development, including dam construction and several levels of pumped water extraction, on seagrass communities in the Gulf of Carpentaria has been modelled using predicted streamflow data generated under different water resource development scenarios (Plagányi et al., 2023). The models predict a potential increase in seagrass biomass by 2 to 7% inshore, particularly adjacent to river estuaries. The risk to the seagrass community was assessed as negligible for each of the four water resource development scenarios (Plagányi et al., 2023). Both the construction of dams and the harvest of river flows via pumped water extraction have limited impacts on aspects of the seagrass community stability and growth. Reducing the volume and duration of high-level flows mitigates the impact of freshwater flood plumes and turbidity on the Gulf of Carpentaria seagrass community, preventing smothering, light reduction or freshwater depredations to the benefit of seagrasses (Plagányi et al., 2023; Turschwell et al., 2021a). Figure 3-80 Conceptual model showing the relationship between threats, drivers, effects and outcomes for seagrass in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. 3.4.7 Surface-water-dependent vegetation Description and background to ecology Across much of the northern Australia, terrestrial vegetation survives on water derived from local rainfall that recharges soils during the wet season and can be accessed by the root systems within unsaturated soils throughout the year. Terrestrial vegetation that receives extra water (i.e. in addition to local rainfall), for example, recharge from flood waters (or by accessing shallow groundwater, see Section 3.4.2), often provide a lush green and productive forest ecosystem (high diversity, dense tree cover) within an otherwise drier or more sparsely vegetated savanna environment (e.g. Pettit et al., 2016). This is referred to as surface-water-dependent vegetation. While water availability influences the distribution of savanna versus forest ecosystems across the northern Australia landscape, their distributions are also linked to fire regime, nutrient availability, soil type and herbivory (Murphy and Bowman, 2012). Terrestrial vegetation that receives extra water may contain unique species (e.g. the Carpentarian Rock-rat (Zyzomys palatalis; Endangered, EPBC Act; Critically Endangered, IUCN) which is unique to monsoon forest, Crowley (2010)) and For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. provide critical habitat for fauna (e.g. Melaleuca forests in the NT support many nationally significant rookeries for waterbirds (Woinarski, 2004)). Such habitats often occur along rivers and floodplains, fringing wetlands and springs or where the depth to groundwater is within reach of the roots. Vegetation naturally inhabits and thrives in niches in the environment that provide the right combination of water conditions, including surface water depths (during low and high flows), groundwater depth, timing and flood frequency (return interval), and flood duration. The optimal water regime will vary for different climate conditions (rainfall regime), site conditions (soil type and water availability) and vegetation types. The water regime supports vegetation survival, growth, flowering and fruiting, germination and successful establishment of new saplings for the diversity of ecosystem species, and maintains their functions and services. Vegetation is unlikely to be able to adapt to changes in water availability outside natural variation. Vegetation has some inbuilt resilience to natural changes in water availability, but prolonged change is likely to result in dieback after some lag period and a shift in ecosystem structure and function (e.g. Mitchell et al., 2016). Terrestrial vegetation that requires surface water inundation and/or access to groundwater is at risk from water resource development if the natural surface water and groundwater regimes are modified beyond some limit. To anticipate potential impacts of any future water resource development in northern Australia, this section reviews the water regimes that support three terrestrial vegetation types: • paperbark swamps • river red gum • monsoon vine forest. In northern Australia, these ecosystems provide food and habitat for many species (e.g. for migratory waterbirds, flying foxes, crocodiles and honeyeaters) and play a role in nutrient cycling and providing buffering against erosion. Paperbark swamps Paperbark is a term commonly used to describe a range of Melaleuca species that have a distinctive papery bark texture. Some paperbark species occur in low-lying areas that are seasonally inundated with fresh water (Department of Environment and Science Queensland, 2013). Many paperbark species co-occur with eucalypt species in riparian and floodplain tree swamps (Department of Environment and Science Queensland, 2013), but here a ‘paperbark swamp’ means the non-tidal coastal and sub-coastal swamp (tree swamp) occurring in the equatorial tropical and subtropical areas of the NT and Queensland (Department of Environment and Science Queensland, 2013) that are dominated by Melaleuca species with papery textured bark. The dominant paperbark swamp species of northern Australia include broad-leaved paperbark (Melaleuca viridiflora), weeping paperbark (M. leucadendra), silver paperbark (M. argentea), blue paperbark (M. dealbata) and yellow-barked paperbark (M. nervosa) (Department of Environment and Energy, 2017), but may also include M. acacioides, M. cajuputi, M. citrolens, M. minutifolia and M. stenostachya in the NT and M. arcana, M. citrolens, M. clarksonii, M. fluviatilis, M. foliolosa, M. saligna, M. stenostachya and M. tamariscina in Queensland (based on Department of Environment and Energy, 2017). The combination of the dominant paperbark swamp species (M. viridiflora, M. leucadendra, M. argentea, M. nervosa and M. dealbata) can flower all year round (Brock, 2022), providing an almost constant source of nectar and pollen for insects, birds and bats (Department of Environment and Science Queensland, 2013). Paperbark swamps provide nesting sites for native birds and flying foxes and are a critical food source for migratory birds (Williams, 2011) and honeyeaters, especially when part of an ecotone (a transition between two ecological communities (Franklin and Noske, 1998)). Fukuda and Cuff (2013) found that about 10% of crocodile nests in the northern coastal and sub-coastal regions of the NT occurred in Melaleuca forests and woodlands. It is unknown whether crocodile nesting would continue in Melaleuca forests if surface water inundation regimes were altered due to water resource development. Coastal paperbark swamps are hypothesised to provide spawning habitat for gudgeon that move between rivers and floodplains during floods (Department of Environment and Science Queensland, 2013). Paperbark swamps can be inundated for 3 to 6 months of the year. If they are inundated for longer periods, they may shift towards more grass, sedge and herb-type wetlands (Department of Environment and Science Queensland, 2013). Some species are more tolerant of extended flooding than others, with M. leucadendra and M. cajuputi occurring in the most flood-prone areas of swamps in northern Australia (Franklin et al., 2007). Investigations at Howard Springs, NT, showed that paperbark swamps were generally inundated between December and June, and water levels fluctuated between 1 m above ground during the wet season and down to 2.5 m below ground level during the dry season (Cook et al., 1998). There appeared to be sufficient water available to M. viridiflora without the need to access shallow groundwater during the monitoring period (based on a water balance study incorporating investigations of evapotranspiration using eddy correlation and sap flow, groundwater dating, soil water properties, runoff (Cook et al., 1998)). However, Melaleuca species were shown to use groundwater in other parts of northern Australia (e.g. M. dealbata; Department of Water and Environmental Regulation (2017), based on water potentials and depth to groundwater data; M. leucadendra; Canham et al. (2021), based on stable isotopes of water analyses), indicating that some paperbark swamps are GDEs. Not much is known about the conditions required for regeneration of paperbark swamps. Major Melaleuca germination may be triggered by the timing and extent of wet-season rains (Woinarski, 2004). In general, Franklin et al. (2007) observed very few paperbark seedlings but occasional areas, most often recently burnt, with abundant saplings. River red gum River red gum (Eucalyptus camaldulensis) commonly line permanent or seasonal rivers and sometimes form forests over floodplains (Costermans, 1981) that are subject to frequent or periodic flooding. The water requirements of E. camaldulensis have not been investigated in northern Australia. However, in the Murray–Darling Basin (MDB), E. camaldulensis experiences episodic flooding and drought and it uses more water than is available from rainfall alone (Doody et al., 2015). It can use groundwater with salinities up to a maximum of approximately 30 mS/cm (Overton and Jolly, 2004). Falling groundwater levels have resulted in E. camaldulensis dieback when groundwater levels dropped below critical levels or thresholds (12 to 22.6 m below ground surface; Horner et al. (2009); Kath et al. (2014); Reardon-Smith et al. (2011)). The threshold groundwater levels are variable and depend on climate conditions and soil characteristics. Flooding requirements for maintaining healthy river red gum have been estimated for various floodplain forests and riparian woodlands in the MDB; they range from a flood duration of 2 to 8 months every 1 to 3 years (Rogers and Ralph, 2010) to up to 2 months duration every 3 to 5 years (Wen et al., 2009). E. camaldulensis may require flood to induce seed fall (George, 2004), but excessive flooding can destroy seeds (Rogers and Ralph, 2010). Note that these flooding relationships exist for trees found in the MDB where extensive research has focused on maintaining this ecosystem type. However, these relationships cannot be directly extrapolated to the different hydrology, soil and climate conditions of northern Australia. Specific water requirements for E. camaldulensis and subspecies found in northern Australia are unknown. Monsoon vine forest Monsoon vine forest can be found in tropical and subtropical regions of northern Australia, with patches spanning the NT, Queensland and WA. While generally falling under the umbrella term ‘rainforest’, with its closed canopy and high leaf cover exceeding 70% (Stork et al., 2008), it can be further characterised by canopy height, leaf size, proximity to permanent moist soils and species composition. This forest type is typically found in areas of 600 to 2000 mm mean annual rainfall (Bowman, 2000). Most monsoon vine forests seem limited to areas with permanent soil water, such as creek lines, springs and seeps. They are thought to be remnants of a wetter period during Australia’s geological history, and changes in climate, fire regime and water availability is thought to have restricted their distribution to small pockets (of less than several hectares) across northern Australia (Bowman, 2000). However, the hydrological and geomorphic environments of these ecosystem communities are poorly understood, and while monsoon forest can typically be found in areas that offer fire protection, such as boulder outcrops and areas of high soil water, a change in water availability may make monsoon vine forests more prone to fire (Larsen et al., 2016; Russell‐Smith, 1991). While a set definition of what constitutes a monsoon vine forest, vine thicket or rainforest is not wholly agreed upon, the definitions provided by (Webb, 1968; Webb, 1959; 1978) and Russell‐ Smith (1991) seem to be widely used and are therefore used throughout this report. Furthermore, Russell‐Smith (1991) categorised monsoon vine forests into 16 different floristic assemblages or rainforest types; he defined these by where they grew (coastal vs inland), water regime (wet vs dry), rainforest type (forest vs vine thicket) and canopy type and height. This report uses water regime as a focus for selecting monsoon forest types. It focuses on forests that require annual inundation, regular watering through streamflow or are groundwater dependent. These are roughly defined as ‘wet’, having near-constant waterlogging of soils with very little soil drying out, or ‘dry’, occurring on floodplains or being seasonally flooded and experiencing regular drying out of soils. See Appendix C for a further breakdown of monsoon forest types. Under the EPBC Act, the semi-deciduous vine thickets of WA are considered a Threatened Ecological Community and Endangered (Fisher et al., 2014). Surface-water-dependent vegetation in the Southern Gulf catchments The distribution of red gum has not been comprehensively mapped, but available data indicate that it is dotted along the banks of major rivers in the Southern Gulf catchments (based on ALA data, Atlas of Living Australia (2021); Figure 3-81). Extensive areas of paperbark (Melaleuca citrolens, M. viridiflora, M. leucadendra) occur within about 100 km of the coast in the Southern Gulf catchments (based on vegetation mapping, Atlas of Living Australia (2021); Department of Environment Parks and Water Security (2000); NVIS Technical Working Group (2017)). While Figure 3-81 shows known observed occurrences of paperbark species in Southern Gulf catchments, it is unconfirmed whether all these occur within swamp habitats. ‘Wet’ monsoon forest species diagnostic of springs and seasonal flooding seem to occur where there are geological transitions that cause groundwater to discharge to the land surface. This may be an artefact of the limited data available (based on ALA data, Atlas of Living Australia (2021)) rather than actual species distribution across the catchment. Figure 3-81 Locations of observed selected surface-water-dependent vegetation types in the Southern Gulf catchments Species within each vegetation type are listed in Appendix C. Only the distribution for the mainland is shown. Datasets: Atlas of Living Australia (2023a; 2023b); Department of Environment Parks and Water Security (2000); NVIS Technical Working Group (2017) Flow–ecology relationships forsurface-water-dependentvegetation Red gum,paperbark and ‘wet’ monsoon forest vegetation are sensitive tochanges in wateravailabilitybecause theyneed more water than isavailablefrom local rainfall aloneto sustainthem.Somerequire periodic inundation byfloodwaters and/or access togroundwaterto survive, flower,fruit and/or reproduce, assummarisedin Table3-29.The amount, source,timingand frequencyof extrawater needed by vegetation will vary depending on climate, localsoilsand vegetation type. The water needsfor all vegetation types are not well defined, particularly innorthern Australia. Table3-29Ecological functionssupportingsurface-water-dependent vegetationand their associatedflowrequirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Chapter3 Ecological assets from the Southern Gulf catchments and marine region|187 For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. References: 1. Franklin and Bowman (2003) includes secondary references, 2. Casanova (2015), 3. Bell (1999), 4. Russell‐Smith (1991), 5. Larsen et al. (2016), 6. Franklin et al. (2007), 7. Wilson et al. (1996), 8. Finlayson and Woodroffe (1996) Pathways of change for surface-water-dependent vegetation Changes in the water available to both groundwater-dependent and surface-water-dependent vegetation could affect ecosystem function and the persistence of each vegetation type into the future. Some paperbark swamps and most wet monsoon forests require near-constant waterlogging or high levels of inundation to maintain health; these may also use groundwater (Franklin et al., 2007; Larsen et al., 2016). These wet environments create conditions that are essentially fireproof in the near-annually burnt fire regime practice of northern Australia (Fisher et al., 2014). Indeed, reductions in water availability can adversely affect these systems through effectively ‘drying them out’, thus making them more fire prone; this in turn could affect recruitment, community structure and the overall biodiversity of the area. In Litchfield National Park, NT, a study was conducted where a naturally occurring, but retreating alluvial knickpoint (see box below) affected the surface water and groundwater availability for a wet monsoon forest. This retreat dried out the histosol soils (peat-like soils) and caused the wet monsoon forest to retreat, becoming more fire prone and suffering fire damage (Larsen et al., 2016). Wet monsoon forests seem particularly sensitive to disturbances such as erosion, flooding, changes to water regimes and fire. If these disturbances increase in frequency in the wet monsoon forest areas, there is the potential of an ecosystem shift from wet monsoon forest to possibly a paperbark/Melaleuca forest over time, as paperbarks are more resilient to these pressures and have a similar watering requirement (Franklin et al., 2007). Other threats include grazing pressure through introduced species such as cattle, damage from wallowing species such as the water buffalo and feral pigs, and weeds. All can cause degradation to the environment and can affect community structure, loss of biodiversity and ecosystem function (Russell-Smith and Bowman, 1992). Alluvial knickpoint explained An alluvial knickpoint is a geomorphological feature of a river or stream where there is a sudden change in elevation or a sudden step or drop in the river or longitudinal profile, like a waterfall (Fryirs and Brierley, 2012). This can be caused by volcanic uplift, an earthquake, landslide, or in the case of Litchfield National Park, bedrock that is resistant to erosional pressures. A retreating alluvial knickpoint occurs when erosion of the bedrock has sped up and the river is retreating or migrating upstream; this in turn changes the topography of the river, and may influence how groundwater interacts with the vegetation downstream of the knickpoint, as is seen in Larsen et al. (2016). Water levels, inundation time and the velocities of waterways seem to influence what ecosystem types are present in northern Australia (Figure 3-82). If a location is waterlogged or spring-fed, and has little disturbance from fire or floods, then the conditions may better support the wet monsoon vine forest type (Franklin et al., 2007; Larsen et al., 2016). However, in the same environmental conditions but with high levels of disturbances (such as those mentioned above), the location may support certain types of paperbark swamp, or the ecosystem may change as the result of this disturbance (Franklin et al., 2007). Indeed, a paperbark swamp with a high frequency of continual inundation and disturbances could shift to a grassland ecosystem (Department of Environment and Science Queensland, 2013). However, Bren (1992) showed that flooded grasslands are at risk of encroaching river red gum (E. camaldulensis) forests if inundation patterns change from yearly, to every couple of years with periods of drying out. Therefore, these ecosystem types are very sensitive to changes in water availability, and a change in watering patterns through dam infrastructure, climate change or water harvesting has the potential to change the current ecosystem and generate an ecosystem shift (Figure 3-82). This conclusion is by no means extensive or will it follow this exact pattern. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-82 Conceptual model showing the relationship between threats, drivers, effects and outcomes for surface- water-dependent vegetation in northern Australia Blue arrows represent hydrological changes and black arrows represent non-hydrological changes. References ABARES (2022) Land use of Australia 2010–11 to 2015–16, 250 m, Australian Bureau of Agricultural and Resource Economics and Sciences. Canberra, Viewed 20 December 2023, (https://www.agriculture.gov.au/abares/aclump/land-use/data-download). 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Zhou S, Punt AE, Deng R, Dichmont CM, Ye Y and Bishop J (2009) Modified hierarchical Bayesian biomass dynamics models for assessment of short-lived invertebrates: a comparison for tropical tiger prawns. Marine and freshwater research 60(12), 1298-1308. Part II Appendices Species Distribution Model parameters To estimate modelled distributions for asset species, we first attributed five major classes of predictors to the 954457 of the Australian Hydrological Geofabric (AHGF, Bureau of Meteorology, 2020)) 3.2 polygons across Northern Australia that intersect with the AHGF streamlines. Predictor variables Predictor variables were attributed to the polygons using the package terra (Hijmans et al., 2022) in R. The predictor classes are: • Land use The six highest level classes from the Catchment Scale Land Use map (CLUM) (ABARES, 2022) are: – Conservation and natural environments – Production from Relatively Natural Environments – Production from Dryland Agriculture and Plantations – Production from Irrigated Agriculture and Plantations – Intensive uses – Water • Soils The fourteen classes from the Australian Soil Classification Map (Searle et al., 2021) are: – Vertosol – Sodosol – Dermosol – Chromosol – Ferrosol – Kurosol – Tenosol – Kandosol – Hydrosol – Podosol – Rudosol – Calcarasol – Organosol. – Anthroposol • Geology Surface Geology of Australia (Raymond, 2012), summarised into classes that mirror the AHGF NCB 2.1: – Igneous – Metamorphic – Sedimentary – Mixed – Other. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Apx. Figure A-1 Percent sedimentary rocks in subcatchment • Vegetation: NVIS 6.0 (Department of Climate Change‚ Energy‚ the Environment and Water, 2020) summarised into classes that mirror the AHGF NCB 2.1: – Bare – Forest – Grasslands – Woodlands – Other. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Apx. Figure A-2 Percent woodland in subcatchment • Climate: Variables from the CSIRO 9-second climate surfaces for Australia (Harwood, 2018) are: – Aridity (x3): Mean, min, max – Evaporation (x5): Mean, min, max, as well as actual scaled by MODIS and modelled using water-holding capacity – Precipitation (x5): Annual, min, max, as well as seasonality (equinox and solstice factor ratios) – Minimum Temperature (x3): Annual mean, min, max – Maximum Temperature (x3): Annual mean, min, max – Temperature range (x3): Annual range, diurnal min and max. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Apx. Figure A-3 Average annual evapotranspiration Ecological data preparation and modelling Using the package ‘galah’ (Westgate M, 2023) in R, we built a pipeline to automatically download ALA datapoints for the entire area of Northern Australia and (see above) and attributed them to the AHGF3.2 polygons using terra. To reduce autocorrelation and tone down artificial weighting inflation, we reduced the 64 predictor variables to 10 principal components which we then used in the modelling (Linke et al., 2008) We then ran three different modelling algorithms: • Random Forests (Breiman, 2001) – an ensemble learning method used for classification and regression. It operates by constructing a multitude of decision trees at training time and outputting the class that is the mode of the classes (classification) or mean prediction (regression) of the individual trees. Random decision forests correct for decision trees' habit of overfitting to their training set. The algorithm combines the simplicity of decision trees with flexibility, making it a robust and accurate model. It handles large data sets efficiently and can manage thousands of input variables without variable deletion. • Generalized Linear Models (Nelder and Wedderburn, 1972) extend linear regression by allowing the linear model to be related to the response variable via a link function and by allowing the magnitude of the variance of each measurement to be a function of its predicted value. GLMs are flexible in handling different types of response variables, like binary, count, or continuous outcomes. They usually avoid overfitting, but result in slightly lower evaluation metrics. • Maxent (Phillips et al., 2006), short for Maximum Entropy Modelling, is a widely used algorithm in ecology for species distribution modelling. It estimates the probability distribution of a species' occurrence based on environmental constraints, using the principle of maximum entropy. This approach assumes that without additional information, the best distribution is the one that maximizes entropy (i.e., is most uniform) while remaining consistent with the given constraints. Maxent is especially popular for its effectiveness with incomplete data sets and its ability to handle presence-only data, making it ideal for predicting species distributions under various environmental conditions. We thinned occurrences that were present in the same subcatchment to reduce observation bias and created 1000 pseudo-absences (or if >1000 observations matched the presences with pseudo- absences). We only used records later than 1960 that intersected with polygons that contain waterways and that had a stated coordinate uncertainty <5km. While Maxent and RF have a built-in variable weighting algorithm, we ran a best subsets selection procedure for GLMs. Terrestrial GDE observations in the catchment Southern Gulf catchments terrestrial GDE observationsin ALA ApxTableB-3 Terrestrial GDEs observed in the Southern Gulf catchments Notes: This is based on a search of the literature and species mapped in ALA(Atlas of Living Australia, 2021)and is nota fully comprehensive list. Any subspecies of these varieties present in the ALA database are included in the mapping(Figure3-66).Manysynonymsexistfor these species. All synonyms recorded in ALA(Atlas of Living Australia, 2024) and Kew(https://www.kew.org/science/collections-and-resources/data-and-digital/names-and-taxonomy(RoyalBotanic Gardens, 2024))were searched(Jan 2024)and included in mapping.Synonyms are notlistedinthe appendices because there arehundreds, butare available upon request. OBLIGATE GDEFACULTATIVE GDE OR TYPE OF DEPENDENCY UNCONFIRMEDPOTENTIAL GDERiparian Eucalyptus camaldulensis Melaleucaargentea Acacia auriculiformis Casuarina cunninghamiana  Cathormion umbellatum  Eucalyptus coolabah  Lophostemon grandiflorus  Lophostemon lactifluus  Melaleuca fluviatilis  Melaleuca lanceolata  Nauclea orientalis  Pandanus spiralis  Corymbia bella  Melaleuca trichostachya  Pandanus aquaticus  Paperbark swamp Melaleuca dealbata  Melaleuca leucadendra  Melaleuca viridiflora  Melaleuca bracteata  240| Ecological asset descriptionsof the Southern Gulf catchments OBLIGATE GDE FACULTATIVE GDE OR TYPE OF DEPENDENCY UNCONFIRMED POTENTIAL GDE Melaleuca cajuputi  Melaleuca citrolens  Melaleuca saligna  Melaleuca stenostachya  Monsoon vine forest Tylophora cinerascens  Abrus precatorius  Bauhinia cunninghamii or Lysiphyllum cunninghamii  Capparis lasiantha  Celtis philippensis  Clerodendrum floribundum var. ovatum  Croton habrophyllus  Diospyros humilis  Dodonaea platyptera  Exocarpos latifolius  Flueggea virosa subsp. Melanthesoides  Grewia breviflora  Gyrocarpus americanus subsp. pachyphyllus  Hypoestes floribunda var. varia  Jasminum didymum  Operculina aequisepala  Opilia amentacea  Planchonia careya  Sersalisia sericea  Terminalia petiolaris  Tinospora smilacina  Vincetoxicum cinerascens  Acacia aulacocarpa  OBLIGATE GDE FACULTATIVE GDE OR TYPE OF DEPENDENCY UNCONFIRMED POTENTIAL GDE Antidesma parvifolium Canarium australianum  Ficus coronulata  Ficus racemosa  Ficus virens  Lindsaea ensifolia  Lygodium microphyllum  Melastoma affine  Nephrolepis biserrata  Syzygium angophoroides  Terminalia microcarpa  Vitex glabrata  Xanthostemon eucalyptoides  Other habitats Acacia stenophylla  Callistemon viminalis  Corymbia opaca  Corymbia tessellaris  Eucalyptus melanophloia  Eucalyptus miniata  Eucalyptus tetrodonta  Livistona lanuginosa  Melaleuca nervosa  Atalaya hemiglauca  Cyperus conicus  Surface-water-dependent vegetation observations in the Southern Gulf catchment Many synonyms exist for the species included in the tables below. All synonyms recorded in ALA (Atlas of Living Australia, 2024) and Kew (https://www.kew.org/science/collections-and- resources/data-and-digital/names-and-taxonomy (Royal Botanic Gardens, 2024)) were searched (Jan 2024) and included in mapping. Synonyms are not listed in the appendices because there are hundreds. Apx Table C-1 Red gum species (including subspecies) observed in northern Australia based on ALA (Atlas of Living Australia, 2021) data within the Southern Gulf catchments (tick) The groundwater-dependent ecosystem (GDE) column denotes whether species are known (tick), or assumed (potential but not specifically investigated, P) to use groundwater. RED GUM SPECIES GDE SOUTHERN GULF Eucalyptus camaldulensis   Eucalyptus camaldulensis subsp. acuta P  Eucalyptus camaldulensis subsp. camaldulensis P Eucalyptus camaldulensis subsp. obtusa P  Apx Table C-2 Paperbark species of northern Australia that occur in seasonally waterlogged habitats based on Melaleuca swamp species and Melaleuca species habitats (Atlas of Living Australia, 2021) and bark texture The groundwater-dependent ecosystem (GDE, See Section 3.4.2) column denotes whether species are known (tick), or assumed (potential but not specifically investigated, P), not considered (blank) to use groundwater. * Denotes species for which subspecies exist in the ALA datasets and are included in mapping (Figure 3-81). PAPERBARK SWAMP SPECIES GDE SOUTHERN GULF Melaleuca acacioides  *Melaleuca alsophila  Melaleuca argentea   Melaleuca cajuputi P  Melaleuca clarksonii Melaleuca citrolens P  Melaleuca dealbata   *Melaleuca ferruginea Melaleuca foliolosa Melaleuca fluviatilis   *Melaleuca lanceolata   PAPERBARK SWAMP SPECIES GDE SOUTHERN GULF Melaleuca leucadendra   Melaleuca minutifolia Melaleuca nervosa   Melaleuca saligna  Melaleuca stenostachya  Melaleuca tamariscina  *Melaleuca trichostachya P  Melaleuca viridiflora   Apx Table C-3 Monsoon forest species that occur where extra water (in addition to rainfall) is available, for example surface water flows or shallow groundwater Some species typically occur in wet habitats (drainage lines, seasonally flooded areas or around springs) but may also occur in drier areas and these are termed ‘typical’. Some species only occur in wet habitats, and these are termed ‘diagnostic’. Lists are based on interpretation of data from Russell‐Smith (1991), identification of presence in the Southern Gulf catchment based on ALA data (Atlas of Living Australia, 2021). The groundwater-dependent ecosystem (GDE, See Section 5.1.1) column denotes whether species are known (tick), or assumed (potential but not specifically investigated, P), not considered (blank) to use groundwater. Note: Subspecies of monsoon vine forest species present in ALA datasets are included in mapping. MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE SOUTHERN GULF Abrus precatorius typical P  Abutilon andrewsianum typical  Acacia aulacocarpa typical  Acacia auriculiformis typical   Acmena hemilampra typical typical Acmenosperma claviflorum diagnostic diagnostic Aglaia sapindina diagnostic Allosyncarpia ternata diagnostic Antidesma parvifolium typical  Atalaya variifolia P Barringtonia acutangular typical  Bauhinia cunninghamii typical typical P  Blechnum indicum typical typical Caesalpinia major P Calophyllum sil diagnostic MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE SOUTHERN GULF Calophyllum soulattri typical Canarium australianum typical typical  Capparis lasiantha typical P  Capparis sepiaria typical  Carpentaria acuminata typical P Cayratia maritima typical  Celtis philippensis typical P  Celtis strychnoides Clerodendrum floribundum P Cordyline terminalis typical Croton habrophyllus P  Cupaniopsis anacardioides typical  Denhamia obscura typical Diospyros cordifolia typical Diospyros humilis P  Dodonaea platyptera P  Drypetes lasiogyna typical Dysoxylum acutangulum typical Dysoxylum latifolium typical typical Ehretia saligna typical  Elaeocarpus culminicola diagnostic diagnostic Erycibe coccinea typical Euodia elleryana typical Exocarpos latifolius P  Fagraea racemosa typical Ficus apodogynum typical Ficus benjamina typical Ficus coronulata typical  Ficus leucotricha typical Ficus opposita typical  Ficus racemosa typical typical  Ficus virens typical typical  Flagellaria indica typical P Flueggea virosa subsp. melanthesoides P  Glochidion perakense typical MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE SOUTHERN GULF Glycosmis trifoliata typical  Gmelina schlechteri typical Grewia breviflora P  Gymnanthera nitida typical Gyrocarpus americanus subsp. pachyphyllus P  Helicia australasica typical Helicteres rhynchocarpa P  Homalanthus novo- guineensis typical Horsfieldia australiana typical Hydriastele wendlandiana typical Hypoestes floribunda var. varia P Ilex arnhemensis typical Jasminum didymum P  Jasminum molle typical  Leea indica typical Lindsaea ensifolia diagnostic  Litsea breviumbellata diagnostic typical Litsea glutinosa typical typical Livistona benthamii typical Lophopetalum arnhemicum diagnostic Lophostemon grandiflorus typical   Lycopodium cernuum typical typical Lygodium flexuosum typical Lygodium microphyllum typical  Macaranga involucrata diagnostic typical Macaranga tanarius diagnostic Maranthus corymbosa typical Melaleuca cajuputi P Melaleuca leucadendra typical   Melastoma affine typical  MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE SOUTHERN GULF Melhania oblongifolia typical  Micromelum minutum typical  Mimusops elengi  Nauclea orientalis typical   Nephrolepis biserrata typical  Operculina aequisepala P  Opilia amentacea P  Passiflora foetida typical  Piper novae-hollandiae diagnostic Planchonella DNA 47005 typical Planchonia careya P  Pleomele angustifolia diagnostic Polyalthia australis typical Polyscias australianum typical typical Rapanea benthamiana diagnostic Rhus taitensis diagnostic Schefflera actinophylla diagnostic Secamone elliptica typical  Sersalisia sericea P  Smilax australis typical Sterculia holtzei typical Sterculia quadrifida typical  Strychnos lucida typical  Syzygium angophoroides diagnostic diagnostic  Syzygium fibrosum typical Syzygium forte typical Syzygium minutuliflorum typical Syzygium nervosum typical P Terminalia ferdinandiana P Terminalia microcarpa typical  Terminalia petiolaris P Terminalia platyphylla diagnostic  MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE SOUTHERN GULF Terminalia subacroptera typical Tinospora smilacina P  Tylophora cinerascens  Vincetoxicum cinerascens P  Vitex glabrata typical  Xanthostemon eucalyptoides diagnostic  As Australia’snational scienceagency and innovation catalyst, CSIRO is solving the greatestchallenges through innovativescience and technology. 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