Ecological assets of northern Australia to inform water resource assessments A technical report from the CSIRO Roper River Water Resource Assessment for the National Water Grid Australia’s National Science Agency Danial Stratford1, Rob Kenyon1, Jodie Pritchard1, Linda Merrin1, Simon Linke1, Rocio Ponce Reyes1, Laura Blamey1, Rik Buckworth1,2, Pascal Castellazzi1, Bayley Costin1, Roy Aijun Deng1, Ruan Gannon1, Sophie Gilbey1, Darran King1, Keller Kopf2, Stacey Kopf2, Heather McGinness1, Paul McInerney1, Colton Perna2, Eva Plaganyi1, Nathan Waltham3 1 CSIRO 2 Charles Darwin University 3 James Cook University A group of logos with a sun and waves Description automatically generated A black background with purple text Description automatically generated ISBN 978-1-4863-1901-5 (print) ISBN 978-1-4863-1902-2 (online) Citation Stratford D, Kenyon R, Pritchard J, Merrin L, Linke S, Ponce Reyes R, Blamey L, Buckworth R, Castellazzi P, Costin B, Deng R, Gannon R, Gilbey S, King D, Kopf K, Kopf S, McGinness H, McInerney P, Perna C, Plaganyi E and Waltham N (2022) Ecological assets of northern Australia to inform water resource assessments. A technical report from the CSIRO Roper River Water Resource Assessment for the National Water Grid. CSIRO, Australia. Copyright © Commonwealth Scientific and Industrial Research Organisation 2022. 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 Roper River 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 Government. The Assessment was guided by two committees: i. The Assessment’s Governance Committee: CRC for Northern Australia/James Cook University; CSIRO; National Water Grid (Department of Climate Change, Energy, the Environment and Water); NT Department of Environment, Parks and Water Security; NT Department of Industry, Tourism and Trade; Office of Northern Australia; Qld Department of Agriculture and Fisheries; Qld Department of Regional Development, Manufacturing and Water ii. The Assessment’s joint Roper and Victoria River catchments Steering Committee: Amateur Fishermen’s Association of the NT; Austrade; Centrefarm; CSIRO, National Water Grid (Department of Climate Change, Energy, the Environment and Water); Northern Land Council; NT Cattlemen’s Association; NT Department of Environment, Parks Australia; Parks and Water Security; NT Department of Industry, Tourism and Trade; Regional Development Australia; NT Farmers; NT Seafood Council; Office of Northern Australia; Roper Gulf Regional Council Shire 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 its 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 a large number of 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, Jayne Brim Box, Jonathan Vea, Glen Durie, Thor Sanders, Linda Lee and Ian Leiper. Colleagues in other jurisdictions also provided support, including Frances Verrier (Australian Government). People in private industry, universities, local government and other organisations also helped us. They include Lindsay Hutley, Clement Duvert, Jenny Davis, Erica Garcia, Osmar Luiz, Jeremy Russell-Smith, Alison King and David Cook. The Assessment gratefully acknowledges the members of the Indigenous Traditional Owner groups, residents and corporations from the Roper catchment. 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 Bitter Springs. Source: CSIRO Director’s foreword Sustainable regional development is a priority for the Australian and Northern Territory governments. Across northern Australia, however, there is a scarcity of scientific information on land and water resources to complement local information held by Indigenous owners and landholders. Sustainable regional development requires knowledge of the scale, nature, location and distribution of the likely environmental, social and economic opportunities and the risks of any proposed development. Especially where resource use is contested, this knowledge informs the consultation and planning that underpins the resource security required to unlock investment. In 2019 the Australian Government commissioned CSIRO to complete the Roper River Water Resource Assessment. In response, CSIRO accessed expertise and collaborations from across Australia to provide data and insight to support consideration of the use of land and water resources for development in the Roper catchment. While the Assessment focuses mainly on the potential for agriculture, the detailed information provided on land and water resources, their potential uses and the impacts of those uses are relevant to a wider range of regional-scale planning considerations by Indigenous owners, landholders, citizens, investors, local government, the Northern Territory and federal governments. Importantly the Assessment will not recommend one development over another, nor assume any particular development pathway. 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 C:\Users\bru119\AppData\Local\Microsoft\Windows\Temporary Internet Files\Content.Word\C_Chilcott_high.jpg Project Director The Roper River Water Resource Assessment Team Project Director Chris Chilcott Project Leaders Cuan Petheram, Ian Watson Project Support Caroline Bruce Communications Chanel Koeleman/Kate Cranney, Siobhan Duffy, Amy Edwards Activities Agriculture and socio- economics Chris Stokes, Caroline Bruce, Shokhrukh Jalilov, Diane Jarvis1, Adam Liedloff, Yvette Oliver, Alex Peachey2, Allan Peake, Maxine Piggott, Perry Poulton, Di Prestwidge, Thomas Vanderbyl7, Tony Webster, Steve Yeates Climate David McJannet, Lynn Seo Ecology Groundwater hydrology Indigenous water values, rights, interests and development goals Danial Stratford, Laura Blamey, Rik Buckworth, Pascal Castellazzi, Bayley Costin, Roy Aijun Deng, Ruan Gannon, Sophie Gilbey, Rob Kenyon, Darran King, Keller Kopf3, Stacey Kopf3, Simon Linke, Heather McGinness, Linda Merrin, Colton Perna3, Eva Plaganyi, Rocio Ponce Reyes, Jodie Pritchard, Nathan Waltham9 Andrew R. Taylor, Karen Barry, Russell Crosbie, Phil Davies, Alec Deslandes, Katelyn Dooley, Clement Duvert8, Geoff Hodgson, Lindsay Hutley8, Anthony Knapton4, Sebastien Lamontagne, Steven Tickell5, Sarah Marshall, Axel Suckow, Chris Turnadge Pethie Lyons, Marcus Barber, Peta Braedon, Kristina Fisher, Petina Pert Land suitability Ian Watson, Jenet Austin, Elisabeth Bui, Bart Edmeades5, John Gallant, Linda Gregory, Jason Hill5, Seonaid Philip, Ross Searle, Uta Stockmann, Mark Thomas, Francis Wait5, Peter L. Wilson, Peter R. Wilson Surface water hydrology Justin Hughes, Shaun Kim, Steve Marvanek, Catherine Ticehurst, Biao Wang Surface water storage Cuan Petheram, Fred Baynes6, Kevin Devlin7, Arthur Read, Lee Rogers, Ang Yang, Note: Assessment team as at June 15, 2023. All contributors are affiliated with CSIRO unless indicated otherwise. Activity Leaders are underlined. 1James Cook University; 2NT Department of Industry, Tourism and Trade; 3 Research Institute for the Environment and Livelihoods. College of Engineering, IT & Environment. Charles Darwin University; 4CloudGMS; 5NT Department of Environment, Parks and Water Security; 6Baynes Geologic; 7independent consultant; 8Charles Darwin University; 9Centre for Tropical Water and Aquatic Ecosystem Research. James Cook University. ii | Ecological assets of northern Australia Shortened forms For more information on this figure or equation or table, please contact CSIRO on enquiries@csiro.au Units UNIT DESCRIPTION cm centimetre ha hectare km kilometre (1000 metres) m metre mAHD metres above Australian Height Datum mm millimetre mS MilliSiemens ppt parts per thousand yr year Preface Sustainable regional development is a priority for the Australian and Northern Territory governments. For example, in 2023 the Northern Territory Government committed to the implementation of a new Territory Water Plan. One of the priority actions announced by the government was the acceleration of the existing water science program ‘to support best practice water resource management and sustainable development’. The efficient use of Australia’s natural resources by food producers and processors requires a good understanding of soil, water and energy resources so they can be managed sustainably. Finely tuned strategic planning will be required to ensure that investment and government expenditure on development are soundly targeted and designed. Northern Australia presents a globally unique opportunity (a greenfield development opportunity in a first-world country) to strategically consider and plan development. Northern Australia also contains ecological and cultural assets of high value and decisions about development will need to be made within that context. Good information is critical to these decisions. Most of northern Australia’s land and water resources, however, have not been mapped in sufficient detail to provide for reliable resource allocation, mitigate investment or environmental risks, or build policy settings that can support decisions. Better data are required to 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. In consultation with the Northern Territory Government, the Australian Government prioritised the catchment of the Roper River for investigation (Preface Figure 1-1) and establishment of baseline information on soil, water and the environment. Northern Australia is defined as the part of Australia north of the Tropic of Capricorn. The Murray– Darling Basin and major irrigation areas and major dams (greater than 500 GL capacity) in Australia are shown for context. The Roper River Water Resource Assessment (the Assessment) provides a comprehensive and integrated evaluation of the feasibility, economic viability and sustainability of water and agricultural development. While agricultural developments are the primary focus of the Assessment, 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 and urban development and aquaculture, in relevant locations. 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. As policy and regulations can change, this enables the results to be applied to the widest range of uses for the longest possible time frame. Preface Figure 1-1 Map of Australia showing Assessment area It was not the intention – 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 not directly examined in the Assessment are discussed with reference to and in the context of the existing literature. Functionally, the Assessment adopted an activities-based approach (reflected in the content and structure of the outputs and products), comprising eight activity groups; each contributes its part to create a cohesive picture of regional development opportunities, costs and benefits. Preface Figure 1-2 illustrates the high-level links between the eight activities and the general flow of information in the Assessment. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au Preface Figure 1-2 Schematic diagram of the high-level linkages between the eight activities 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 most reliably informs discussion and decisions concerning regional development when read as a whole. The Assessment has produced a series of cascading reports and information products: •Technical reports; that present scientific work at a level of detail sufficient for technical andscientific experts to reproduce the work. Each of the eight activities has one or morecorresponding technical report. •A Catchment report; that for the Roper catchment synthesises key material from the technicalreports, providing well-informed (but not necessarily-scientifically trained) readers with theinformation required to make decisions about the opportunities, costs and benefits associatedwith irrigated agriculture and other development options. •A Summary report; that for the Roper catchment provides a summary and narrative for ageneral public audience in plain English. •A Summary factsheet; that for the Roper catchment provides key findings for a general publicaudience in the shortest possible format. The Assessment has also developed online information products to enable the reader to better access information that is not readily available in a static form. All of these reports, information tools and data products are available online at https://www.csiro.au/roperriver. The website provides readers with a communications suite including factsheets, multimedia content, FAQs, reports and links to other related sites, particularly about other research in northern Australia. A diagram of a diagram Description automatically generated Executive summary The ecology activity seeks to determine the relative risks between different water resource development scenarios in the Assessment catchment 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 catchment include: •What is the main environmental context of the catchment that could influence waterresource development? •What are the key environmental drivers and stressors that are currently occurring or likelyto occur in the catchment (including key supporting and threatening processes such asinvasive 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 developmentsconsidering impacts from potential changes in flow on species and habitats? This report provides a synthesis of the prioritised ecology assets occurring in the Assessment catchment including developing asset knowledge bases, conceptual relationships, and evidence narratives, including the flow–ecology relationships, and considering their context and application in the Assessment catchment. Contents Ecology Technical Report: Asset Descriptions Director’s foreword i The Roper River Water Resource Assessment Team ii Shortened forms iii Units iv Preface v Executive summary viii Contents x Figures ............................................................................................................................. xiii Tables ............................................................................................................................ xvii Part I Ecology of the Assessment catchment 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 Longitudinal and lateral connectivity .................................................................... 4 1.2.3 Habitat modification and loss ................................................................................ 4 1.2.4 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 Asset identification and prioritisation ................................................................... 5 1.3.2 Review, 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 catchment 10 2.1 Ecology of the Roper catchment ...................................................................................... 10 2.1.1 Roper catchment and its environmental values ................................................. 10 2.1.2 Protected, listed and significant areas of the Roper catchment ......................... 12 2.1.3 Important habitat types and values of the Roper catchment ............................. 14 2.1.4 Significant species and ecological communities of the Roper catchment .......... 16 2.1.5 Current condition and potential threats in the Roper catchment ...................... 16 3 Ecological assets from freshwater systems 18 3.1 Freshwater habitats .......................................................................................................... 18 3.1.1 Floodplain wetlands ............................................................................................ 18 3.1.2 Inchannel waterholes .......................................................................................... 23 3.2 Freshwater fish and shark species and groups ................................................................. 28 3.2.1 Barramundi (Lates calcarifer) .............................................................................. 28 3.2.2 Catfish (Order: Siluriformes)................................................................................ 34 3.2.3 Grunters (Family: Terapontidae) ......................................................................... 39 3.3 Freshwater turtles ............................................................................................................ 44 3.3.1 Turtles (Family: Chelidae) .................................................................................... 44 3.4 Freshwater waterbird groups ........................................................................................... 50 3.4.1 Grouping waterbirds............................................................................................ 50 3.4.2 Colonial and semi-colonial nesting wading waterbirds....................................... 53 3.4.3 Cryptic wading waterbirds ................................................................................... 61 3.4.4 Shorebirds ............................................................................................................ 66 3.4.5 Swimming, grazing and diving waterbirds .......................................................... 77 4 Ecological assets from marine systems 88 4.1 Marine habitats................................................................................................................. 88 4.1.1 Mangroves ........................................................................................................... 88 4.1.2 Saltpans and salt flats .......................................................................................... 93 4.1.3 Seagrass habitats ................................................................................................. 98 4.2 Marine crustacean species and species groups ............................................................. 103 4.2.1 Banana prawns (Penaeus merguiensis) ............................................................. 103 4.2.2 Endeavour prawns (Genus: Metapenaeus) ....................................................... 109 4.2.3 Mud crabs (Genus: Scylla) ................................................................................. 116 4.2.4 Tiger prawns (Penaeus esculentus and P. semisulcatus) ................................... 121 4.3 Marine fish and shark species and species groups ......................................................... 127 4.3.1 Mullet (Family: Mugilidae) ................................................................................ 127 4.3.2 Sawfish (Genus: Pristis) ..................................................................................... 132 4.3.3 Threadfin (Polydactylus macrochir) ................................................................... 137 5 Ecological assets from terrestrial systems 142 5.1 Terrestrial habitats ......................................................................................................... 142 5.1.1 Groundwater-dependent ecosystems............................................................... 142 5.1.2 Surface water dependent vegetation................................................................ 159 References 168 Part II Appendices 211 Figures Preface Figure 1-1 Map of Australia showing Assessment area ..................................................... vi Preface Figure 1-2 Schematic diagram of the high-level linkages between the eight activities and the general flow of information in the Assessment. ..................................................................... vii Figure 2-1 Conceptual diagram of selected ecological assets of the Roper catchment. Ecological assets include species of significance, species groups and important habitats ........................... 11 Figure 2-2 Location of protected areas and important wetlands within the Roper catchment assessment area including management areas protected mainly for conservation through management intervention as defined by the IUCN. ..................................................................... 13 Figure 3-1 White-bellied sea-eagle looking over a wetland in northern Australia ...................... 19 Figure 3-2 Land subject to inundation (potential floodplain wetlands) and nationally important wetlands (DIWA) in the Roper catchment .................................................................................... 20 Figure 3-3 Conceptual model showing the relationship between threats, drivers, effects and outcomes for floodplain wetlands in northern Australia ............................................................. 23 Figure 3-4 A large waterhole with fringing vegetation in the Roper catchment .......................... 24 Figure 3-5 Location of persistent inchannel waterholes in the Roper catchment ....................... 25 Figure 3-6 Conceptual model showing the relationship between threats, drivers, effects and outcomes for inchannel waterholes in northern Australia .......................................................... 28 Figure 3-7 Location of barramundi and their habitat in the Roper catchment and the marine region ............................................................................................................................................ 31 Figure 3-8 Conceptual model showing the relationship between threats, drivers, effects and outcomes for barramundi in northern Australia .......................................................................... 34 Figure 3-9 Location of catfish in the Roper catchment ................................................................ 36 Figure 3-10 Conceptual model showing the relationship between threats, drivers, effects and outcomes for catfish in northern Australia ................................................................................... 39 Figure 3-11 Location of grunters in the Roper catchment and marine region ............................. 41 Figure 3-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for grunters in northern Australia ................................................................................ 44 Figure 3-13 Location of freshwater turtles within the Roper catchment ..................................... 46 Figure 3-14 Conceptual model showing the relationship between threats, drivers, effects and outcomes for freshwater turtles in northern Australia ................................................................ 50 Figure 3-15 Royal spoonbill individual at the nest ........................................................................ 54 Figure 3-16 Location of colonial and semi-colonial nesting wading waterbirds in the Roper catchment ..................................................................................................................................... 56 Figure 3-17 Location of colonial and semi-colonial nesting wading waterbirds in the Roper catchment ..................................................................................................................................... 57 Figure 3-18 Egret hunting among water lilies ............................................................................... 60 Figure 3-19 Conceptual model showing the potential relationship between threats, drivers, effects and outcomes for colonial and semi-colonial nesting wading waterbird species ............ 61 Figure 3-20 Dense aquatic and semi-aquatic vegetation used as habitat by cryptic wading waterbirds ..................................................................................................................................... 62 Figure 3-21 Location of cryptic wading waterbirds in the Roper catchment ............................... 63 Figure 3-22 Conceptual model showing the relationship between threats, drivers, effects and outcomes for cryptic wading waterbirds in northern Australia ................................................... 66 Figure 3-23 Red-capped plover walking along a shore ................................................................. 67 Figure 3-24 Location of shorebirds in the Roper catchment ........................................................ 72 Figure 3-25 Location of shorebirds in the Roper catchment ........................................................ 73 Figure 3-26 Location of shorebirds in the Roper catchment ........................................................ 74 Figure 3-27 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the shorebirds group in northern Australia ........................................................... 77 Figure 3-28 Magpie goose perched on a fallen tree branch ......................................................... 78 Figure 3-29 Location of swimmer, grazer and diver waterbirds in the Roper catchment ........... 83 Figure 3-30 Location of swimmer, grazer and diver waterbirds in the Roper catchment ........... 84 Figure 3-31 Location of swimmer, grazer and diver waterbirds in the Roper catchment ........... 85 Figure 3-32 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the waterbird group swimmers, grazers and divers in northern Australia ............ 87 Figure 4-1 Location of mangroves in the Roper catchment marine region .................................. 90 Figure 4-2 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mangroves in northern Australia ........................................................................... 93 Figure 4-3 Saltpan area in northern Australia, which are generally located between mangrove and saltmarsh areas ...................................................................................................................... 94 Figure 4-4 Location of salt flats in the Roper catchment marine region ...................................... 96 Figure 4-5 Conceptual model showing the relationship between threats, drivers, effects and outcomes for saltpans in northern Australia ................................................................................ 98 Figure 4-6 Distribution of seagrass habitats in the Roper catchment marine region ................ 101 Figure 4-7 Conceptual model showing the relationship between threats, drivers, effects and outcomes for seagrass in northern Australia .............................................................................. 103 Figure 4-8 Fisheries catch of banana prawns and their habitat in the Roper catchment marine region .......................................................................................................................................... 106 Figure 4-9 Conceptual model showing the relationship between threats, drivers, effects and outcomes for banana prawns in northern Australia .................................................................. 109 Figure 4-10 Fisheries catch of red endeavour prawns in the Roper catchment marine region . 112 Figure 4-11 Fisheries catch of blue endeavour prawns in the Roper catchment marine region 113 Figure 4-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for endeavour prawns in northern Australia ............................................................. 115 Figure 4-13 Mangrove and intertidal habitat typical of mud crab habitat in northern Australia ..................................................................................................................................................... 117 Figure 4-14 Location of mud crab habitat in the Roper catchment marine region ................... 118 Figure 4-15 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mud crabs in northern Australia .......................................................................... 121 Figure 4-16 Fisheries catch of brown tiger prawns in the Roper catchment marine region ...... 123 Figure 4-17 Fisheries catch of grooved tiger prawns in the Roper catchment marine region ... 124 Figure 4-18 Conceptual model showing the relationship between threats, drivers, effects and outcomes for tiger prawns in northern Australia ....................................................................... 127 Figure 4-19 Records of capture of mullet in the Roper catchment and marine region ............. 129 Figure 4-20 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mullet in northern Australia ................................................................................. 132 Figure 4-21 Records of sawfish capture in the Roper catchment and the marine region ......... 134 Figure 4-22 Conceptual model showing the relationship between threats, drivers, effects and outcomes for sawfish (Pristis pristis) in large rivers in northern Australia ................................. 137 Figure 4-23 Records of threadfin capture in the Roper catchment marine region .................... 139 Figure 4-24 Conceptual model showing the relationship between threats, drivers, effects and outcomes for threadfin in northern Australia ............................................................................ 141 Figure 5-1 Conceptualisation of obligate and facultative groundwater-dependent vegetation 144 Figure 5-2 Conceptualisation of terrestrial GDEs ....................................................................... 145 Figure 5-3 Bitter Springs thermal pools in Elsey National Park .................................................. 146 Figure 5-4 Distribution of known and potential groundwater-dependent aquatic ecosystems in the Roper catchment .................................................................................................................. 148 Figure 5-5 Locations of springs and sinkholes in the Roper catchment ..................................... 149 Figure 5-6 Distribution of potential groundwater-dependent terrestrial ecosystems in the Roper catchment ................................................................................................................................... 151 Figure 5-7 Locations of observed obligate terrestrial GDEs in the Roper catchment ................ 152 Figure 5-8 Locations of facultative and potential GDE vegetation species in the Roper catchment grouped by relevant vegetation type ......................................................................................... 153 Figure 5-9 Locations tested for the presence of subterranean GDEs in the Roper catchment and locations of caves and alluvial and karstic aquifers that may provide habitat for subterranean GDEs ............................................................................................................................................ 155 Figure 5-10 Conceptual model showing the relationship between threats, drivers, effects and outcomes for aquatic GDEs in northern Australia ...................................................................... 157 Figure 5-11 Conceptual model showing the relationship between threats, drivers, effects and outcomes for terrestrial GDEs in northern Australia .................................................................. 158 Figure 5-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for subterranean GDEs in northern Australia ............................................................ 159 Figure 5-13 Locations of observed selected surface water dependent vegetation types in the Roper catchment ......................................................................................................................... 164 Figure 5-14 Conceptual model showing the relationship between threats, drivers, effects and outcomes for surface water dependent vegetation in northern Australia ................................ 167 Tables Table 1-1 Freshwater, marine and terrestrial ecological assets with freshwater dependences ... 6 Table 3-1 Ecological functions for floodplain wetlands and their supporting flow requirements ....................................................................................................................................................... 21 Table 3-2 Ecological functions for inchannel waterholes and their supporting flow requirements ....................................................................................................................................................... 26 Table 3-3 Ecological functions for barramundi and their supporting flow requirements ............ 33 Table 3-4 Ecological functions for catfish and their supporting flow requirements .................... 37 Table 3-5 Ecological functions for grunters and their supporting flow requirements ................. 43 Table 3-6 Ecological functions for freshwater turtles and their supporting flow requirements . 48 Table 3-7 Waterbird species groups and example representative species for northern Australia ....................................................................................................................................................... 52 Table 3-8 Species in the colonial and semi-colonial nesting wading waterbird group, and their national and international conservation status ............................................................................ 58 Table 3-9 Ecological functions for colonial and semi-colonial nesting waders and their supporting flow requirements ...................................................................................................... 59 Table 3-10 Species in the cryptic wading waterbird group, and their national and international conservation status ....................................................................................................................... 62 Table 3-11 Ecological functions for cryptic wading waterbirds and their supporting flow requirements ................................................................................................................................. 64 Table 3-12 Species in the shorebirds group, and their national and international conservation status ............................................................................................................................................. 68 Table 3-13 Ecological functions for shorebirds and their supporting flow requirements............ 75 Table 3-14 Species in the swimmers, grazers and divers waterbird group, and their national and international conservation status ................................................................................................. 79 Table 3-15 Ecological functions for swimmer, grazer and diver waterbirds and their supporting flow requirements ......................................................................................................................... 85 Table 4-1 Ecological functions mangroves and their supporting flow requirements .................. 91 Table 4-2 Ecological functions of saltpans and their supporting flow requirements ................... 97 Table 4-3 Ecological functions for seagrass habitats and their supporting flow requirements . 102 Table 4-4 Ecological functions for banana prawns and their supporting flow requirements .... 107 Table 4-5 Ecological functions for endeavour prawns and their supporting flow requirements ..................................................................................................................................................... 114 Table 4-6 Ecological functions for mud crabs and their supporting flow requirements ............ 120 Table 4-7 Ecological functions for tiger prawns and their supporting flow requirements ........ 125 Table 4-8 Ecological functions for mullet and their supporting flow requirements .................. 130 Table 4-9 Ecological functions for sawfish and their supporting flow requirements ................. 135 Table 4-10 Ecological functions for threadfin and their supporting flow requirements ............ 140 Table 5-1 Ecological functions for GDEs and their supporting flow requirements .................... 156 Table 5-2 Ecological functions for surface water dependent vegetation and their supporting flow requirements ....................................................................................................................... 165 Apx Table A-1 List of GDEs observed in the Roper catchment ................................................... 212 Apx Table B-1 Red gum species (including subspecies) observed in northern Australia based on ALA (Atlas of Living Australia, 2021a; 2021b; 2021c) data within Roper catchment (tick) ........ 215 Apx Table B-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, 2021a; 2021b; 2021c) and bark texture ..................................................................... 215 Apx Table B-3 Monsoon forest species that occur where extra water (in addition to rainfall) is available, for example surface water flows or shallow groundwater ........................................ 216 Part I Ecology of the Assessment catchment Development of water resources can lead to a range of impacts to the environment including changes in flow regimes, land use change and changes to connectivity by building instream structures. The rivers, floodplains and coastal regions of northern Australia are highly diverse and have significant conservation, economic and cultural values. To understand the potential risks to the natural environment associated with water resource development, the ecology activity is using an ecological asset approach by undertaking a prioritisation of assets, developing asset knowledge bases, conceptual relationships and evidence narratives, including the flow–ecology relationships, and considering the context and distribution of the assets in the Assessment catchment. Analysis to understand impacts draws upon the knowledge base of these assets, combined with aspects of ecosystem function to model outcomes of water resource development and climate change scenarios. This report presents this knowledge base for our prioritised freshwater-dependent ecological assets across freshwater, marine and terrestrial habitats. 1 Introduction 1.1 Ecology and water resource development Development of water resources can lead to a range of impacts to the environment including changes in flow regimes, land use change and changes to connectivity by building instream structures. The flow regimes of rivers are regarded as 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 leading to potentially significant changes in important flow attributes such as the magnitude, timing, duration and rate of change of flow events to which flora and fauna of the ecosystem are adapted. Also, the frequency and duration of spells, and any modification of water quality including changes to temperature regimes can also be important for native species. 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 of this is potential ecological changes and consequences for the biota, habitats and ecosystem processes of a catchment (Poff et al., 1997). The freshwater systems in northern Australia contain a high level of diversity, 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 reptiles, 60 amphibian species and 100 macroinvertebrate families (van Dam et al., 2008b). The estuaries of northern Australia similarly support a rich biodiversity, and are critical in supporting productive fisheries, where increased freshwater inflow to estuaries increases fishery production (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 fisheries for barramundi (Lates calcarifer), mud crab (Scylla serrata and Scylla olivacea) and a suite of other species important to commercial, recreational and Indigenous fisheries. Species of conservation significance, including some migratory waterbird species, sea turtles and a variety of sharks and rays, and habitats, such as mangrove forests and seagrass beds, are also dependent on coastal and estuarine systems and the discharge of water and nutrients provided from catchment runoff. The ecology activity seeks to determine the relative risks of water resource development in the Assessment catchment primarily by understanding possible ecological consequences resulting from changes in the flow regime, as well as other impacts such as changes to connectivity and land use change. The key questions that this activity seeks to address include: • What is the main environmental context of the catchment that could influence water resource development? • What are the key environmental drivers and stressors that are currently occurring or likely to occur in the catchment (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? In order to understand the potential risks to the natural environment associated with water resource development, the ecology activity uses an ecological asset approach, in combination 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., 2018). The steps involved for this include undertaking a review and prioritisation of assets in the Assessment catchment, developing asset knowledge bases, conceptual relationships and evidence narratives, including the flow– ecology relationships, and considering the context and distribution of assets and system processes in the Assessment catchment. For the Assessment, the ecology activity utilises a range of modelling methods for the prioritised ecology assets to understand the possible range of changes and impacts that may occur. The ecology assets cover species, groups of species, habitats and their processes that are freshwater dependent, with sufficient understanding of their requirements, and are significant within the Assessment catchment. The broader approach operates at catchment scales and utilises a range of methods that are best suited to individual assets and the questions being asked. The modelling methods range from more qualitative to quantitative as sufficient relationships between flow and ecological outcomes are sufficiently known and can be suitably supported by the methods for the individual assets. The quantitative ecology modelling uses hydrology scenarios developed by the surface water hydrology activity (Hughes et al., 2023) as primary inputs and 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. A summary of the ecology approach is provided in Section 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, inundation and discharges into coastal regions. Globally, water resource development has a range of known impacts on ecological systems. The influence of each of these are dependent upon a range of factors including catchment properties (including physical, geographic and climate characteristics), the type of impacts (for example dams, water harvesting, groundwater extraction), the source location or distribution of the impacts within the catchment, the magnitude and pattern of change, the habitats and species that will be affected and their location, and how any changes may be managed or mitigated. The extent to which impacts may occur are also highly uncertain. Impacts associated with water resource development can include the following: 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., 2019a; Poff et al., 1997). Flow conditions provide the physical habitat in streams and rivers which derives biotic use and composition, to which life-history strategies are evolved, and which enables movement and migration between habitats and exchange of nutrients and materials (Bunn and Arthington, 2002; Jardine et al., 2015). In a river system, the natural periods of both high and low 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). Water resource development through the attenuation of flows can lead to impacts across significant distances downstream of the development including into coastal and near-shore marine habitats (Broadley et al., 2020; Pollino et al., 2018). 1.2.2 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 impact 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., 2019). 1.2.3 Habitat modification and loss Water resource development can result in direct loss of habitat. This can include artificially creating lake habitat behind an impoundment resulting in loss of terrestrial and stream habitat due to inundation by the impoundment. Agricultural development results in the conversion of habitat to more intensive agriculture. Infrastructure including roads and canals can lead to fragmentation of terrestrial habitat or the artificial connection of aquatic habitat that has been historically distinct. 1.2.4 Invasive and non-native species Water resource development often results in homogenisation of flow or habitats. This can be due to the changed patterns of capture and release of flows or creation of impoundments for storage and regulation. Invasive species are recognised to often be at an advantage in such modified habitats (Bunn and Arthington, 2002). Modified landscapes such as lakes or the conversion of ephemeral streams into perennial streams can be a pathway for introduction and support the establishment of non-native species (incidental, accidental or deliberate) including pest plant and fish species (Bunn and Arthington, 2002; Close et al., 2012; Ebner et al., 2020). Increased human activity can lead to increased risk of invasive species being introduced. 1.2.5 Synergistic and co-occurring processes both local and global In addition to 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 stock, 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 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 using a modelling approach which considers a set of ecology assets including species, habitats and catchment ecosystem functions. The ecology assets occur within the catchment and are dependent upon freshwater flows to support their persistence or function. Assets are spread across freshwater, marine and terrestrial regions (including terrestrial habitats with a groundwater or flood flow and inundation dependence). These ecological assets occupy different positions within the catchment, have different watering 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 where and under what conditions different impacts may occur 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 that are described below: • asset identification and prioritisation • review, conceptual modelling and developing evidence narratives • mapping the location and distribution of the assets • understanding flow–ecology relationships and quantitative modelling. 1.3.1 Asset identification and prioritisation For the purpose of the ecology activity, assets are classified as species, species groups or habitats and can be considered as either partially or fully freshwater dependent, or terrestrial or marine dependent upon freshwater flows (or services provided by freshwater flows). To identify assets with which to undertake the ecological analysis, a review and prioritisation of species, species groups or habitats has been undertaken for the Assessment catchments. For the purposes of identifying assets for the ecology assessment, freshwater-dependent assets were considered if they are: • a species or community that is listed as Threatened, Vulnerable or Endangered • a habitat, species or community that is formally recognised in conservation agreements • habitat that provides vital, near-natural, rare or unique habitat for water-dependent flora and fauna • supporting important or notable biodiversity of water-dependent flora and fauna • providing recreational, commercial or cultural value. From the full range of potential assets occurring in the Assessment catchment, the process for selecting priority assets considered if they are: • distinctive – to create an association between flow and outcomes of change, and to consider a broad range of water requirements across different assets • representative – to consider flow requirements for other biota and ecological processes that are not explicitly modelled • describable – with sufficient peer-reviewed evidence available to identify and describe relationships with flow • significant – considering ecological, conservation, cultural and recreational importance, and relevance to the Assessment catchment. The prioritised ecological assets are shown in Table 1-1. These assets are used in the ecology analysis and modelling to quantify the risks associated with different water resource development scenarios considering the type, magnitude and location of change. Asset types described in this report include: • Species, taxonomic groups or functional groups are individual species (such as barramundi in Section 3.2.1) or groups of species that are closely related (such as grunters in Section 3.2.3) or 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. Examples of species groups are the ‘colonial and semi-colonial wading waterbirds’ (Section 3.4.2). • Important habitats include geographical areas identified as sharing similar characteristics such as position on the floodplain or channel, water retention / shedding properties, or other structural features that may make it important for the catchment ecology and support biota within and around the catchment. Examples of habitats include floodplain wetlands (Section 3.1.1), seagrasses (Section 4.1.3) and groundwater-dependent ecosystems (GDEs; Section 5.1.1). Habitats are important for supporting species or communities and may include, but are not limited to, identified or listed locations such as national parks or Directory of Important Wetlands in Australia (DIWA) sites. Table 1-1 Freshwater, marine and terrestrial ecological assets with freshwater dependences 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. 1.3.2 Review, conceptual modelling and developing evidence narratives For each asset, a review of literature and data has been undertaken. This includes catchment- specific data and knowledge, important water and flow requirements, habitat use and distribution, and the collation of information to support an understanding of the asset’s response to potential environmental change. A conceptual model for each asset has been developed to represent this ecological understanding. The asset conceptual models provide a visual summary and an evidence-based hypothesis for the potential impacts that may be possible under water resource development and climate change scenarios considering a range of pathways to impact. Each model is built on a scientific evidence base incorporating a broad range of literature and research, which is documented in the ‘Pathways of change’ section for each asset. The conceptual models provide a structured and generalisable framework to underpin the analyses of the impacts of water resource development by linking possible threats, through drivers and effects, to outcomes. These key terms are defined as follows: • Threat – An action or activity that has the capacity to adversely affect an ecological asset and its value (Hart et al., 2005). • Driver (ecological driver) – The mechanism, process or change via which a threat affects an asset. •Effect–The direct change in,orresponse of, the asset that hasoccurredas a consequenceofthe driver. •Outcome–Theoverall observable or measurable impact(tangibleor otherwise)on theassetorits function within the catchment. These modelsrepresent theavailableknowledge of flow–ecology relationships, consideringaspects such as life history, flow triggers, movement, refuge, foraging and productivity, as relevantto each asset. They explore relationships between key potentialthreats, includingwaterresourcedevelopment, and land use and climate change, and the effects and outcomes of thesethreatingprocesses from the perspective ofeach assetincluding loss of biodiversity orhabitat quality. The conceptual models are box and arrow models that have a standardised structurethat represents links betweenthethreats,drivers,effects and outcomes.The standardised structure of these models allowsforcomparison across assets. Combiningor integratingmodels can alsobe used toexplorethe potential outcomesof threatening processes, including water resourcedevelopment, at systems scalesand support broader synthesis. 1.3.3Mappingthe locationand distribution ofthe assets The locationordistribution of assets has been mapped across theAssessmentcatchmenttounderstand theoccurrencesand/or importantlocationsof the assets within subcatchmentsandthenear-shore coastal region.Water resource development scenarios consider a range ofdevelopment options and pathwayswithin each catchment, with the flowdownstream of eachdevelopmentbeing affected. Assets located downstream ofthese scenariowater resourcedevelopmentsites will beexposed tochanges in flow with different scenarios resultingin differentchanges. Dependingupon the scenario and if thedevelopment is point source or distributed, changes in flow could accumulateor diminishas flowsattenuatethrough thecatchment. The levelof flow change and impact will depend on theasset’slocationrelative tothe development.A range of data sources, includingAtlas of Living Australia, government departments, and fisheriescatch records,wereusedto develop mapsand spatial relationships of theassets in theAssessmentsubcatchments. 1.3.4Understandingflow–ecologyrelationshipsand modelling The assessment approachundertakenin theecology activityusesa combination of qualitative, semi-quantitativeand quantitative modelling methods. The choiceofmethod(s)usedfor eachasset depends uponthe ability tosupport modellinggiventhestrengthof theknowledge base, data and the types of relationshipsfor each asset. Modelling requiresunderstandingrelationshipsbetween flow,ecologicalresponses,andthepotentialoutcomesof changes(including definingmodel endpoints). Different models are used torepresent and incorporatevaryingprocesses that may bemoreimportant for differentassetsincluding changes in flow, inundation and connectivity. The primaryinputs toanalysisaredaily hydrology data generated with river systemmodels(flow dischargeandquantitative propertiesofthe hydrograph)orhydrodynamic models(depth, velocity,inundation) that can be used toquantify the relative differencesbetween scenarios and abaselineoverthesame modelled period(Hughes etal., 2023; Kim et al.,2023).A summary of 8 |Ecological assets of northern Australia some of the ecology modelling used in the ecology assessment are provided below, and further details provided in the companion report (Stratford et al., 2024). Flow requirements A common base to the analysis is the flow requirements method. 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 that are identified for each asset are important for ecological function such as supporting life- history requirements, movement or provision of important habitat. 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. Results indicate sensitivities to the types of change in the hydrograph because of assets having different flow associations and needs, occupying different locations within the catchment and due to water resource development scenarios manifesting different changes in hydrology. 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 methodology of stock assessment approaches and extend this to a representation of multiple co-existing species and stressors in an ecosystem. Hydrodynamic habitat suitability Hydrodynamic habitat suitability modelling uses depth, velocity and inundation extent outputs from hydrodynamic models to quantify and map the occurrence of species suitable and preferred habitat within the catchment across different flow events. The method uses species or species group 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 where depth and velocity are recognised as important determinants of habitat suitability and can be defined to quantify habitat use and preference. Habitat preference can be informed by results of tracking studies that position species use of in channel and floodplain habitat and relate this to the experienced hydraulic properties of the location. Connectivity assessment The connectivity assessment uses hydrodynamic modelling to develop a daily time series of inundation extents 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. wetlands) to the main river channel via connection across the floodplain or via flood runners (latitudinal connectivity). Differences in the extent and duration of connections to wetlands is quantified between the scenarios. 2 Ecology of the catchment 2.1 Ecology of the Roper catchment 2.1.1 Roper catchment and its environmental values The comparatively intact landscapes of the Roper catchment are important for the ecosystem services they provide, including recreational activities, tourism, traditional and commercial fisheries, and areas of agricultural production, notably cattle grazing on native pastures. In addition, they hold important ecological and environmental values. The Roper River is a large perennial flowing river and at 77,400 km2 it drains one of the largest catchment areas flowing into the western Gulf of Carpentaria. Within this catchment, and the surrounding marine environment are rich and important ecological assets including species, ecological communities, habitats, and ecological processes and functions (conceptualised summary in Figure 2-1). The ecology of the Roper catchment is maintained by the river’s flow regime, shaped by the region’s wet-dry climate and the catchment’s complex geomorphology and topography, and driven by seasonal rainfall, evapotranspiration and groundwater discharge. Figure 2-1 Conceptual diagram of selected ecological assets of the Roper catchment. Ecological assets include species of significance, species groups and important habitats See Table 1-1 for a complete list of the freshwater, marine and terrestrial ecological assets considered in the Roper catchment. Biota icons for the Roper catchment adapted from the Integration and Applicaiton Network (2022) Roper catchment conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Much of the natural environment of the Roper catchment is low relief consisting of open woodlands, with escarpments, gorges and plateaux occurring across parts of the catchment. The wet-dry, tropical climate results in highly seasonal river flow with 90% of rainfall occurring between November and March (Faulks, 2001; McJannet et al., 2023). The dynamic occurring between wet and dry seasons provides both challenges and opportunities for biota (Warfe et al., 2011). During the dry season, river flows are reduced and the streams in the catchment recede, many to isolated pools. However, in parts of the Roper catchment the persistence of water during the dry season is supported by discharge from aquifers including the Tindall Limestone Aquifer and the Dook Creek Formation (Faulks, 2001). In the dry season, the streams and waterholes that persist, including the important spring-feed streams between Mataranka Thermal Pools and the Red Lily Lagoon, provide critical refuge habitat for many aquatic species (Barber and Jackson, 2012; Faulks, 2001). In this respect, the Roper catchment is atypical of many of the other catchments of the wet-dry tropics (Kennard, 2010; Pettit et al., 2017) in having as many tributaries supplemented by groundwater discharges (Faulks, 2001; Pettit et al., 2017). Due to the flat topography in parts of the catchment, the Roper River contains sections that braid into smaller channels. These braids and anabranches provide a diverse habitat structure. During the wet season, flooding inundates significant parts of the catchment 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 floodplains, wetlands and intertidal flats of the Limmen Bight (an inlet extending for 135km between Groote Eylandt and the Sir Edward Pellow Group), and delivers extensive discharges into the marine waters of the western Gulf of Carpentaria. 2.1.2 Protected, listed and significant areas of the Roper catchment The protected areas located in the Roper catchment include two national parks, an Indigenous Protected Area and other conservation parks (Figure 2-2). Of the national parks, Elsey National Park covers approximately 140 km² and Limmen National Park approximately 9300 km², although much of Limmen National Park extends beyond the Roper catchment (Department of Agriculture‚ Water and the Environment, 2020b). The South East Arnhem Land Indigenous Protected Area covers an area of approximately 18,000 km2, also extending beyond the Roper catchment. Also within the Roper catchment are Wongalara Sanctuary and the St Vidgeon management area (approximately 2000 km2 and 2800 km2 respectively (Department of Agriculture‚ Water and the Environment, 2020b)). In the Roper catchment marine region are two contiguous marine parks, Limmen Bight in Territory waters and the Limmen Marine Park in Commonwealth waters, covering an area of approximately 870 km2 and 1400 km2, respectively. Further out in the Gulf of Carpentaria is the 7300 km2 Anindilyakwa Indigenous Protected Area consisting of Groote Eylandt and the surrounding waters (Department of Agriculture‚ Water and the Environment, 2020a). Figure 2-2 Location of protected areas and important wetlands within the Roper catchment 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) The Roper catchment includes two DIWA sites (Figure 2-2), namely the groundwater-fed Mataranka Thermal Pools, and the coastal Limmen Bight (Port Roper) Tidal Wetlands System (Environment Australia, 2001; SKM, 2009). These two DIWA wetlands demonstrate a striking Map of Roper catchment. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. contrast between persistent freshwater riparian habitat and marine, coastal and near-shore habitats and reinforce the diversity of aquatic habitats that can be found within the Roper catchment. The Roper catchment contains no Ramsar sites. The Mataranka Thermal Pools DIWA site is located in 4 ha of Elsey National Park in the upper Roper catchment that is maintained by permanent thermal springs (Department of Agriculture, 2019b). The artificially modified pools containing sections of paved and cemented areas are fringed by palm forest and drain into the nearby Waterhouse River (SKM, 2009). The pools provide stable habitat for flyspecked hardyhead (Craterocephalus stercusmuscarum) and chequered rainbowfish (Melanotaenia splendida inornata), while up to 200,000 little red flying-fox (Pteropus scapulatus) roost in the surrounding forest (Department of Agriculture, 2019b). Groundwater- dependent vegetation fringe the pools that are supported by persistent discharges from deep aquifers. The Limmen Bight Tidal Wetlands System DIWA site is part Aboriginal freehold and part private lease and is the second-largest area of saline coastal flats in the NT (1848 km2, excluding subtidal seagrass areas) (Department of Agriculture, 2019a). Limmen Bight forms a highly important habitat system of tidal wetlands (intertidal mud flats, saline coastal flats and estuaries) and while the whole site is tidal, it receives large volumes of freshwater inflows from the contributing catchments. The DIWA site provides one of the most important habitat sites nationally for dugongs (Dugong dugon) (Palmer and Smit, 2019), as well as being an important habitat for several species of marine turtles (Department of Agriculture, 2019a). The seagrass beds of Limmen Bight are a major breeding area for prawns and help support an important fishing industry (Department of Agriculture, 2019a; Palmer and Smit, 2019; Tomkovish and Weston, 2008). Limmen Bight is also an important feeding ground for migratory shorebirds in the NT, with counts in the tens of thousands (Palmer and Smit, 2019). Shorebirds such as the eastern curlew (Numenius madagascariensis; Critically endangered) and great knot (Calidris tenuirostris; Critically endangered) migrate from their breeding grounds in the northern hemisphere and use the intertidal flats for feeding (Department of Agriculture, 2019a; Palmer and Smit, 2019; Tomkovish and Weston, 2008). Due to the provision of important habitat, Limmen Bight is a declared Important Bird Area by BirdLife International (BirdLife International, 2022). 2.1.3 Important habitat types and values of the Roper catchment Within the freshwater sections of the Roper catchment are diverse habitats including persistent and ephemeral rivers, anabranches and braiding channels, wetlands, floodplains and groundwater-dependent ecosystems (GDEs) (Faulks, 2001). The diversity, and complexity of, and connection between, habitats within a catchment are vital for providing a range of habitat needs to support both aquatic and terrestrial biota (Schofield et al., 2018). In the wet season, flooding connects rivers to floodplains. Floodplain habitats, due to their water exchange during floods, support higher levels of primary and secondary productivity in comparison 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 routinely 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 abundance and species richness compared to surrounding habitats (Pettit et al., 2011; Xiang et al., 2016). The riparian habitats that fringe the rivers and streams of the Roper catchment are largely intact and include Eucalyptus camaldulensis overstory with Livistona mariae rigida, Pandanus spp. and Melaleuca communities across many parts of the catchment (Faulks, 2001). Conversely, in the dry season, biodiversity is supported within the inchannel waterholes that persist in the landscape. Waterholes that remain become increasingly important as the dry season progresses and provide important refuge habitat for species and enable recolonisation into surrounding habitats upon the return of larger flows (Hermoso et al., 2013). Waterholes provide direct 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). GDEs occur across many parts of the Roper catchment and come in different forms, including aquatic, terrestrial and subterranean habitats. Aquatic GDEs, including DIWA-listed Mataranka Thermal Pools, 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 often provide critical refuge during periods of the late dry season (James et al., 2013). Vegetation occurring adjacent to the waterways in the Roper catchment rely on water from a range of sources (surface water, soil water, groundwater) which are seasonally dynamic and highly spatially variable across floodplains. The sources of water may be from a combination of direct rainfall, bank recharge from in-stream 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 is only infrequently investigated (e.g. Canham et al. (2021)). In some locations vegetation may be sustained by water available in unsaturated soils and never use groundwater, in other locations vegetation may use groundwater sourced from local alluvial recharge processes or regional groundwater may be critical for maintaining vegetation condition. Sources of water used by vegetation can be patchy across floodplains (e.g. work undertaken by Charles Darwin University at Mataranka) and vary from season to season. Subterranean aquatic ecosystems in the Roper catchment support a diverse and largely hidden stygofauna community within the Tindall Limestone karstic aquifer (Cambrian Limestone Aquifer; Oberprieler et al. (2021)). Some subterranean species are distributed across a broad spatial range, while others have highly restricted ranges, which makes them more vulnerable to local changes where they occur (Oberprieler et al., 2021). For the marine and estuarine environments, the Roper catchment, including the area of Limmen Bight and beyond, has 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 (Bradley, 2018; Department of Agriculture, 2019a; Poiner et al., 1987). The intertidal flats in the Roper catchment marine region are extensive with the mangrove communities containing at least 19 woody plant species fringing the banks of streams and rivers (Palmer and Smit, 2019). Seagrass beds in nearby coastal Gulf of Carpentaria are of high diversity, are vigorous and provide an important food resource for dugongs, 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, while harvest of mud crabs (mainly Scylla serrata) also occurs along the coasts near Port Roper (Bayliss et al., 2014). 2.1.4 Significant species and ecological communities of the Roper catchment The Roper catchment supports some of northern Australia’s most archetypical and important wildlife species, including sawfish (Pristis pristis; Vulnerable), marine turtles and dugong that occur in the coastal waters of the Gulf of Carpentaria. The regionally endemic Gulf snapping turtle (Elseya lavarackorum; Endangered) can be found associated with vegetated freshwater reaches of the catchment. Freshwater crocodiles (Crocodylus johnstoni) are common within the Roper River and its tributaries. While the saltwater crocodile (Crocodylus porosus) frequently occurs in the lower Roper River to around Ngukurr, it can be found upstream as far as near Elsey National Park (Atlas of Living Australia, 2021a). Across the catchment are many lesser-known plants and animals that are also of great importance. Diversity in the Roper catchment is high, containing an estimated 270 vertebrate species (Dasgupta et al., 2019). Among the diversity in the Roper catchment is over 130 species freshwater fishes, sharks and rays. Most of these fish species do not enter the marine environment and remain within the riverine and wetland habitats of the catchment. Owing to healthy floodplain ecosystems and free-flowing rivers (Grill et al., 2019; Pettit et al., 2017), very few freshwater fishes in the study area are threatened with extinction. The Roper catchment is an important stopover habitat for migratory shorebird species that are listed under the Environment Protection and Biodiversity Conservation Act 1999 (EPBC Act) (Australian Government, 1999) including the northern Siberian bar-tailed godwit (Limosa lapponica menzbieri; Critically endangered), eastern curlew (Critically endangered) and the Australian painted snipe (Rostratula australis; Endangered). 2.1.5 Current condition and potential threats in the Roper catchment In conjunction with the diversity of landscapes, habitats, ecological communities and species of the Roper catchment there are also a range of economic enterprises, infrastructure and human impacts. The nature and extent to which human activities have modified the habitats and had an impact on species of the Roper catchment varies. Previous assessments have indicated the riverine habitat in the Roper catchment as being of high quality condition and largely intact and unimpacted by clearing or development (Close et al., 2012; Faulks, 2001), although threatening processes continue to operate in more recent years including as a result of pest species and other disturbances. Intertidal habitats including salt flats and mangroves are recognised as being of good condition and are often of national significance (Department of Agriculture, 2019a). 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 banana prawns, mud crab and barramundi. The study area includes the towns of Ngukurr, Mataranka and Daly Waters, which provide Indigenous homelands, support a vital tourism industry and act as regional hubs for many of the stations across the catchment. While a moderate proportion of the catchment is under conservation reserves the study area does face environmental threats. This includes the potential of tourism-related impacts at sensitive and vulnerable sites. In the Roper catchment the more significant or of higher concern impacts are largely localised and includes areas such as Mataranka Thermal Pools (Department of Agriculture, 2019b). 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). Riparian vegetation characteristics of the Roper are considered to have not been impacted by extensive clearing or development, although impact that occurs is often associated with stock and pest species (Faulks, 2001). One of the most significant environmental threats to remote regions across northern Australia is that of introduced plants and animals. In the Roper catchment, cane toad (Rhinella marina), water buffalo (Bubalus bubalis) and wild pig (Sus scrofa) are among the introduced animals (Atlas of Living Australia, 2021a; Department of Agriculture, 2021). Weeds of national significance in the aquatic systems of northern Australia include giant sensitive tree (Mimosa pigra), olive hymenachne (Hymenachne amplexicaulis), cabomba (Cabomba caroliniana), salvinia (Salvinia molesta) and rubber vine (Cryptostegia grandiflora) (Close et al., 2012). Weed species of interest in and around the Roper catchment include gamba grass (Andropogon gayanus), para grass (Brachiaria mutica), giant sensitive tree and prickly acacia (Vachellia nilotica) (Department of Agriculture, 2021) and some of these, including sensitive tree (M. pigra) and para grass, are recognised to have significantly impacted undeveloped rivers more broadly in northern Australia (Davies et al., 2008). 3 Ecological assets from freshwater systems Northern Australia’s rivers, floodplains and coastal regions contain high diversity including at least 170 fish species, 150 waterbird species, 30 aquatic and semi-aquatic reptiles, 60 amphibian species and 100 macroinvertebrate families (van Dam et al., 2008b). The ecology of the freshwater systems is 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 freshwater sections of the Assessment catchment. Table 1-1 presents the full list of assets used in the ecology activity. 3.1 Freshwater habitats 3.1.1 Floodplain wetlands Description and background to ecology Wetlands in the wet-dry tropics of Australia are considered to have great conservation value (Finlayson et al., 1999), and are considered 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., 2008a), spawning grounds and nurseries for floodplain-dependent fish (Ward and Stanford, 1995) as well as habitat for many other aquatic and riparian species (van Dam et al., 2008a) (Figure 3-1). 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 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 high and low 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 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 Northern Territory 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 (Northern Territory Government, 2020). For the purpose of this Assessment, we are not considering areas within the river channel as wetlands (considered as inchannel waterholes (see Section 3.1.2)). Similarly, marine or saline habitats including mangroves and coastal saltmarshes (salt flats) are also considered as separate assets within this project (Sections 4.1.1 and 4.1.2). Figure 3-1 White-bellied sea-eagle looking over a wetland in northern Australia Photo attribution: CSIRO Floodplain wetlands in the Roper catchment The Roper catchment has two nationally significant wetlands listed under the DIWA: Limmen Bight (Port Roper) Tidal Wetland System and Mataranka Thermal Pools (Figure 3-2) (Department of Agriculture‚ Water and the Environment, 2021). There are no Ramsar-listed wetlands within the Roper catchment. Limmen Bight (Port Roper) Tidal Wetland System is approximately 185,000 ha and is located at the mouth of the Roper River (Department of Agriculture‚ Water and the Environment, 2021). This wetland system includes intertidal mud flats, saline coastal flats and estuaries, and has a high volume of freshwater inflows (McJannet et al., 2009). The area includes the Limmen National Park south of the Roper River, and South East Arnhem Land Indigenous Protected Area to the north of the Roper River. The area is considered a site of conservation significance, and is important for seabirds, waterbirds and migratory shorebirds (Smyth and Turner, 2019). It also supports commercial fisheries for prawns, mud crabs and barramundi (Smyth and Turner, 2019). Traditional practices are still carried out in the South East Arnhem Land Indigenous Protected Area (Gambold, 2015). The Mataranka Thermal Pools in Elsey National Park in the upper reaches of the Roper catchment are a series of permanent, groundwater-connected thermal springs, fed via flows through the Tindall Limestone Aquifer. The thermal pools are inchannel habitat of less than 10 ha in total area (McJannet et al., 2009). For the purpose of this Assessment, the Mataranka Thermal Pools are covered within the waterholes and groundwater-dependent ecosystem (GDE) assets (see Sections 3.1.2 and 5.1.1). Photo of white-bellied sea-eagle in a wetland For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. As well as these two nationally significant wetlands, there are several floodplain areas within the Roper catchment (see land subject to inundation, Figure 3-2). These floodplain areas flood during the wet season, replenishing associated semi-permanent and permanent wetlands. Significant numbers of freshwater floodplains occur in association with the rivers and creeks within the Roper catchment, particularly the Roper, Hodgson, Jalboi and Wilton rivers, and on the Flying Fox, Maiwok, Birdum, Jasper, Horse and Showell creeks (Figure 3-2). Figure 3-2 Land subject to inundation (potential floodplain wetlands) and nationally important wetlands (DIWA) in the Roper catchment DIWA = Directory of Important Wetlands in Australia Datasets: Geoscience Australia (2017); Department of the Environment and Energy (2010) Map For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 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 the major determinant of the level of species diversity, productivity and habitat structure (Close et al., 2015). This, in turn, can impact 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, reducing the viability of populations (Bunn and Arthington, 2002). Table 3-1 outlines these important ecological functions and their corresponding flow component or attribute. Table 3-1 Ecological functions for floodplain wetlands 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 floodplain wetlands There are several threatening processes that can impact on the persistence of wetlands in the Assessment catchment. These threatening processes include river regulation, water extraction, climate change and land use change. River regulation can impact wetlands by the capture of flows by regulating structures and when 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 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 transition into terrestrial habitats. Disruption to the natural flow regime, including the alteration of the magnitude, frequency, duration, timing and rate of change of flows within a system can impact 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, also known as water harvesting, is the pumping of water from a water source, often into an offstream storage. Water can be extracted from both groundwater and surface water sources, with surface water sources being a river or standing water body, such as a lake or wetland. Water extraction of surface water can generally have less impact on the environment than instream storages, as surface water extraction tends to occur during high-flow events such as floods, and not during low-flow periods (Petheram et al., 2008). As a result, water extraction can lower the peak of a flood, providing 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). 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 5.1.1). Water extraction often occurs as a result of an increase in demand from agricultural production, or from urban or industrial development. In this study, water extraction is considered as pumping from a river. Climate change is considered to be a major threat to wetlands (Salimi et al., 2021). A change in the climate may affect rainfall, runoff and evapotranspiration patterns (Grieger et al., 2020; Salimi et al., 2021), impacting on the hydrology of a system, including the baseflow and flood patterns (Erwin, 2009). Changes to the hydrology can also impact 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 impact 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 impact the quantity of inflows to a river. Vulnerability of wetlands to climate change depends on the experienced hydrology conditions and the wetlands position within the landscape (Winter, 2000). Hydrological landscapes are defined by their water source and their flow characteristics. Winter (2000) found that wetlands that were dependent on rainfall were more vulnerable to changes in climate than wetlands dependent on regional groundwater, due to the buffering capacity of these systems. Coastal wetlands in particular may be vulnerable to climate change. Climate change impacts 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 a decrease of freshwater inputs can lead to the saltwater intrusion of wetlands (Close et al., 2015; Close et al., 2012), which in turn can result in the conversion of freshwater floodplains to saline habitats (Finlayson et al., 1999). Land use change can include modification to 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. The interactions of these threats can however compound these impacts. The ecological outcomes of threatening processes on floodplain wetlands in northern Australia are presented below, with their implications for changes to floodplain wetland biodiversity and function (Figure 3-3). Figure 3-3 Conceptual model showing the relationship between threats, drivers, effects and outcomes for floodplain wetlands in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.1.2 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 retract to a series of discrete and disconnected waterholes (Faulks, 2001; 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, as well as providing 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 (Figure 3-4), and can be assisted through having more reliable groundwater (see aquatic GDEs, Section 5.1.1), and further enhancing the habitat value of the area. 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 that have a higher number of persistent waterholes, and often those with a range of different habitat characteristics, are recognised to enhance Conceptual model For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. biodiversity at regional scales (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 is exclusive of large lakes and storages, and pertains to locations of retained water occurring within often disconnected locations within the river channel, rather than on the floodplain or in the estuary (see Floodplain wetlands, Section 3.1.1). 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 otherwise through connection to contributions such as groundwater inflows in locations where this occurs (also see GDEs in Section 5.1.1). Figure 3-4 A large waterhole with fringing vegetation in the Roper catchment Photo attribution: CSIRO Inchannel waterholes in the Roper catchment Larger and more persistent inchannel waterholes occur throughout many parts of the Roper catchment (Figure 3-4). The Roper River is unusual among those of northern Australia in having many of its tributaries fed by groundwater discharge from limestone aquifers. The groundwater discharge maintains an often significant level of baseflow during periods that would otherwise result in highly reduced flow or cease-to-flow conditions (Faulks, 2001; Wallace et al., 2011). Areas of the Roper catchment with groundwater discharge include Mataranka, Flying Fox Creek, Mainoru River and some of the Wilton River (Faulks, 2001; Waugh, 2014) (Figure 3-5). In areas other than these however, many tributaries demonstrate the ephemeral flows that are seasonally characteristic of northern Australian rivers more broadly (Faulks, 2001; Petheram et al., 2008). In these ephemeral reaches, waterholes that persist provide important habitat values. Photo of a large waterhole in the Roper catchment For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. In the Roper catchment, these biodiversity values are highlighted by providing habitat for species listed under the EPBC Act, including the Gulf snapping turtle; Endangered; Section 3.3.1), the freshwater sawfish (Vulnerable; Section 4.3.2) and numerous fish species. Waterhole habitats in the Roper catchment include the Mataranka Thermal Pools listed within the DIWA. Figure 3-5 Location of persistent inchannel waterholes in the Roper catchment 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. A buffer of 1 km is applied to this waterhole mapping to make it visible at the map scale. Data source: Sims et al. (2016) Map. 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 3-2). The habitat conditions within waterholes or the persistence of waterholes within the landscape declines where the duration of low-flow periods is extended, where water is removed from the river during low flows, or where water extraction directly from waterholes occurs. Where this occurs, 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 and their supporting flow requirements for inchannel waterholes are summarised in Table 3-2. Table 3-2 Ecological functions for inchannel waterholes and their supporting 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 3-6). Changes in the flow regime associated with upstream water capture and storage, surface 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 and also 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), as well as 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 migration including longitudinal connectivity along the river channel important to biotic movement, and water quality often deteriorates due to lower exchange along the watercourse. During periods of low flows, 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., 2018). Waterholes persist as a result of the hydrological balance within the system, resulting from the timing and duration of both filling events and drawdown (Close et al., 2012). While loss of waterholes can result in a range of impacts, alternatively any increases in the persistence of waterholes may also result in environmental impacts. This 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 is adapted, with possible impacts to 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 change to provide conditions which may favour 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 result in a shift in the species composition of a region. As 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, alternatively, 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 impact the plants providing this habitat at local and regional scales. Pest species such as buffalo and pigs, as well as unrestricted cattle access to waterholes can cause damage to riparian vegetation, increase sedimentation, and result in higher levels of 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 creation of perennial streams can be a source for introduction (incidental, accidental or deliberate) or create habitat for the establishment of pest plant and fish species (Close et al., 2012; Ebner et al., 2020). The ecological outcomes of threatening processes on inchannel waterholes in northern Australia are illustrated below, with their implications for changes to biodiversity and ecosystem function (Figure 3-6). Figure 3-6 Conceptual model showing the relationship between threats, drivers, effects and outcomes for inchannel waterholes in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.2 Freshwater fish and shark species and groups 3.2.1 Barramundi (Lates calcarifer) Description and background to ecology Barramundi are a large (>1 m standard length) catadromous opportunistic-predatory fish (order Perciformes) that occurs throughout northern Australia. The species occurs in ‘catchment to coast’ Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. habitats throughout the west Indo-Pacific region, including estuaries, rivers, lagoons and wetlands across northern Australia (Crook et al., 2016; Pender and Griffin, 1996; Russell and Garrett, 1983; 1985). The fish is a long-lived (living up to about 32 years) fast-growing, protandrous hermaphrodite species that occupies 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 community (Blaber et al., 1989; Brewer et al., 1995; Milton et al., 2005; Russell and Garrett, 1985). Barramundi are arguably the most important fish species to commercial, recreational and Indigenous subsistence fisheries throughout Australia’s wet-dry tropics. Barramundi make up a substantial component of the total commercial fish catch in northern Australia (Savage and Hobsbawn, 2015). 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 has increased from about 7 kg 100 m net-1 d-1 in the early 1980s to over 30 kg 100 m net-1 d-1 in the 2010s (Northern Territory Government, 2018). The commercial and recreational catches make up then largest proportions of all catches in the NT, though the indigenous artisanal catch is significant in some years. Barramundi are 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 to Indigenous communities across tropical Australia (Green et al., 2010). Their movements are related to habitat requirements during their life cycle and the reliance of barramundi on seasonal variation in river flows to access these habitats. In the NT, the Indigenous catch of barramundi in the study area is less certain than other fisheries. Barramundi life history renders them critically dependent on river flows (Tanimoto et al., 2012). Large females (older fish) and smaller males (younger fish) reside in estuarine and littoral coastal habitats. Mating and spawning occurs in the lower estuary during that later dry season to early wet season and new recruits move into supra-littoral and freshwater habitats; with coastal salt flat, floodplain and palustrine habitats dependent 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. Ontogenetic migration requires palustrine–riverine and riverine–estuarine connectivity; hence migration is dependent on catchment flows. Barramundi transform to females at about 6 years old when they mix with younger males within river estuaries and breed. Over the last 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 males delay downstream spawning migrations for 6 to 10 years until they have undergone the transition to females in freshwater habitats; and (iii) some barramundi can 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 flow regime (Crook et al., 2016), making the species particularly vulnerable to water resource development (Robins et al., 2021). Moreover, the effects of different levels of river flow now are better understood. During high-flow years (a strong wet season), barramundi tend to remain within the estuary (Brien Roberts (World Fisheries Conference 2021), 2021, pers. comm. (WFC 2021)); 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 were likely to emigrate to riverine and palustrine habitats, suggested to be seeking better foraging conditions (Crook et al., 2022). Roberts et al. (2019) demonstrated that approximately 62% of barramundi in tropical Australian rivers were catadromous, while many remained within the estuary. Barramundi in the Roper catchment Barramundi occupy relatively pristine habitats in both freshwater and estuarine reaches of the Roper River (Figure 3-7). Barramundi are abundant in the river. Historically, the Roper River was less-well scientifically studied than other tropical rivers in the NT (e.g. the Daly River and South Alligator River (Larson et al., 2013)). The Roper River estuary is roughly at the centre of a Northern Territory Fisheries catch-reporting grid (#1435) and Territory-wide, the ongoing harvest of the commercial stock is considered sustainable (Northern Territory Government, 2018). Between 1990 and 2009, the annual catch of barramundi from the Roper River estuary and nearby coastal areas fluctuated between 25 t and 102 t (Halliday et al., 2012). The remote location of the Roper catchment is linked to low numbers of reports of barramundi in freshwater reaches of the river and its tributaries (e.g. from recreational fishers). However, given the commercial catch and the known ecology of the species from other catchments it is likely that the Roper catchment represents an important system for barramundi (Crook et al., 2016; Dostine and Crook, 2016). Four barramundi were caught in an upstream freshwater billabong in the vicinity of Elsey Station during a Northern Territory Museum and Art Galley sponsored survey in 2008 (Dally and Larson, 2008). In 2015, barramundi were acoustically tagged in the Roper River to measure movement and migration up to 100 km upstream and downstream within the freshwater and estuarine reaches of the river (Dostine and Crook, 2016). In addition, barramundi were caught in both the freshwater and marine reaches of the Roper River for otolith microchemistry work and showed that fish spend considerable portions of time in freshwater reaches of the river (Crook et al., 2016). Moreover, four-wheel-drive vehicle assisted-access and the desire of recreational fishers for ‘wilderness experience’ over recent decades has supported the establishment of accommodation and guided fishing ventures on established enterprises within the catchment. Moroak Station (a cattle enterprise) in the mid-catchment has established guest accommodation and promotes several wilderness experiences on the station. Fishing targeting barramundi is specifically listed on their website and the promotion states that ‘hundreds of barramundi live in the Roper River’ (Moroak Guided Safari, n.d.). No data from these fishing ventures are available; however, the presence of the activity in the catchment demonstrates that barramundi are abundant in the freshwater reaches of the Roper River and its tributaries. Figure 3-7 Location of barramundi and their habitat in the Roper catchment and the marine region Barramundi juveniles use the freshwater habitats upstream of the estuary as well as brackish shallow estuarine littoral and supra-littoral habitats in the wet season. Barramundi travel upstream to freshwater habitats during the wet season and emigrate downstream during the wet season. Data source: Atlas of Living Australia (2021a) Map. 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, a life history that would be significantly impacted by interruptions to the natural flows of northern Australian rivers (Crook et al., 2022). 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). Historically, individual male fish were thought to move upstream to freshwater riverine and palustrine habitats for about 3 to 4 years before maturing and migrating downstream to the estuary. They then transformed to females at about 6 years old when they mix with younger males and breed. Recent studies using new technologies have proposed a diversity of estuarine and riverine strategies adopted by subsets of young-of-year individuals (Roberts et al., 2019). Different migration types 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. 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 allow fish to 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). Around 1 year of age, individuals move out of the supra-littoral habitats and 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 ephemeral or perennial freshwater habitats from months to years until flood-moderated connectivity allows them to 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 runoff 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. Large instream dams that sever upstream/downstream connectivity may have greater effects on barramundi populations than reduced flows by negating access to riverine and palustrine habitats within a large proportion of a catchment. The ecological functions and their supporting flow requirements for barramundi are summarised in Table 3-3. Table 3-3 Ecological functions for barramundi 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 barramundi In the wet-dry tropics of northern Australia, 90% of rainfall and runoff occurs from January to March (Petheram et al., 2012). 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), but also 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 end (April to June) also promote superior growth compared to 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; 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 ecological outcomes of threatening processes on barramundi in northern Australia catchments are presented below, with their implications for changes to growth, mortality, refuge and habitat, and community structure (Figure 3-8). Figure 3-8 Conceptual model showing the relationship between threats, drivers, effects and outcomes for barramundi in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.2.2 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 consisting of freshwater species, some 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 found in both freshwater and marine habitats. Most catfish are bottom feeders, but generally 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 also notable for its reproductive traits such as the largest eggs of any teleost group (>10 mm) and males exhibiting strong parental care behaviour, Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 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 and thus can make up a lot of biomass in a catchment (Crook et al., 2020). The key Plotosid species are reasonably tolerant to high temperature and low dissolved oxygen (DO) levels, however fish kills at very low DO levels have been reported (Bishop, 1980). The key threat to the two dominant Neosilurus species are potential flow barriers. Plotosidae need high flows to trigger spawning migration, and require a barrier-free passage spawning grounds in the headwater streams. While not as important as barramundi or sooty grunters, Neoarius graeffei (fork-tailed catfish) has considerable importance as a subsistence fish for Indigenous communities (Finn and Jackson, 2011; Jackson et al., 2011). Catfish in the Roper catchment Catfish in the Roper catchment belong to two families: Ariidae (7 species, a family that is split between marine and freshwater species), as well as Plotosidae (4 species, mainly freshwater species in the Roper). Apart from the species in Figure 3-9, these include the Ariids Cinetodus froggatti, Neoarius berneyi, Sciades leptaspis and the marine Nemapteryx armiger, which we excluded from the map to reduce clutter. The larger-bodied Ariid catfish like Neoarious graeffei, Neoarius midgleyi and Sciades paucus are mainly found on the main stems of the Roper River. The usually smaller-bodied Neosilurus species are mainly found in smaller tributaries. Figure 3-9 Location of catfish in the Roper catchment Data source: Atlas of Living Australia (2021a) Map. 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 Siluridae. 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/early dry seasons, Neosilurus hyrtlii was recorded from lowland lagoons, floodplain lagoons and perennial streams of the escarpment. This seems to be context and habitat dependent as another study in a Queensland catchment had 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 show upstream spawning migration – from their adult habitat in the lowland rivers to tributaries where they 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’. Siluridae are a fairly resilient family of 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 leptasis, 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 Neoarius graeffei (Kailola and Pierce, 1988). Neoarius midgley requires connection to the offchannel floodplain as habitat during the dry season. Kailola and Pierce (1988) also report Neoarius midgley 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 and their supporting flow requirements for catfish are summarised in Table 3-4. Table 3-4 Ecological functions for catfish and their supporting 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 The possible ecological outcomes of threatening processes on catfish in northern Australia are discussed below, with 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 and river regulation, with the added threat of climate change. All-in-all the Ariid catfish species present in some northern catchments are fairly pollution tolerant, yet they are all to some degree dependent on a natural flow regime (Pusey, 2004). All species are dependent on connections to the floodplain, often for the purpose of recruitment. River regulation and extraction can reduce overbank flows, leading to a decrease in connection frequency and therefore a loss in recruitment opportunities. Also, as agricultural growing seasons often overlap with fish spawning seasons, water is likely extracted at these important times, while 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) hypothesise 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 that may have similar effects. Some Plotosidae species prefer flowing water in the main channel. This could be impacted by overextraction or even structural changes like dams, which can alter cease-to-flow periods. (Allen, 1982; Bishop et al., 1990). As stated above, the combination of impact on movement, together with missing spawning migration triggers, is highly likely to impact population sizes of Plotosidae, especially Neosilurus ater (Pusey, 2004). Although this could differ under varying circumstances as catfish in Queensland have 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 show upstream spawning migration – from their adult habitat in the lowland rivers to tributaries where they 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’. The ecological outcomes of threatening processes on catfish in northern Australia are illustrated below, with their implications for changes to habitat, community structure and population size (Figure 3-10). Figure 3-10 Conceptual model showing the relationship between threats, drivers, effects and outcomes for catfish in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.2.3 Grunters (Family: Terapontidae) Description and background to ecology Grunters inhabit riverine, estuarine and marine waters and in northern Australia there are a total of 37 species of grunter from 11 genera, with the most species-rich genera being Hephaestus, Scortum, Syncomistes and Terapon. 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, and are a soniferous family (i.e. they can both vocalise and hear well), thus may be sensitive to noise (Smott et al., 2018). One of the most widespread species is the sooty grunters (Hephaestus fuliginosus). Sooty grunters are omnivorous and their diet is diverse in composition, containing 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 are an important recreational species, with environmental flow being 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 overall biomass, they are an important link in the overall food chain. They bridge lower trophic levels with top-level predators, such as long tom (Strongylura krefftii) or crocodiles. Grunters are also important species for Indigenous people 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 Roper catchment and marine region There are six species of grunters in the Roper catchment (Figure 3-11): Spangled grunter (Leiopotherapon unicolor), barred grunter (Amniataba percoides), sooty grunter (Hephaestus fuliginosus), Gulf grunter (Scortum ogilbyi) and estuarine trumpeter (Pelates quadrilineatus). There is a single instance of Barcoo grunter (Scortum barcoo) in the Atlas of Living Australia. It is slightly out of range; however, it is a museum record, hence a reliable observation. Of these, sooty grunters are the key species for recreational and cultural purposes (Chan et al., 2012). In the Roper catchment, grunters are likely widespread with headwaters being spawning and nursery grounds, as well as habitat for adults of the smaller species (spangled grunter for example). Waterholes on the main stem of the Roper River represent habitat for adult grunters. Figure 3-11 Location of grunters in the Roper catchment and marine region Data source: Atlas of Living Australia (2021a) Map. 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. 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, with adults having a preference for flowing water (McDowall, 1996). In some tropical streams in the 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 as the most rheophilic grunter species 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, but also drown shallow refuge habitats. This can reduce riffles and runs, and reduction of 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 impact 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 grunter. It is of little use to assume a charismatic taxon such as barramundi can be a surrogate for the flow requirements of all life stages of sooty grunters Not many studies exist about the western sooty grunter Hephaestus jenkinsii; however, we assume that the very close relative of H. fuliginosus has very similar preferences (Allen et al., 2002). Leiopotherapon unicolor and other smaller-bodied grunters have additional requirements to hydraulic habitat. Leiopotherapon unicolor need a relatively high spawning water temperature of 20–26 °C (Allen et al., 2002). This can be compromised by hypolimnetic releases from impoundments. Movement is also key for L. unicolor’s life cycle, where as a smaller species compared to the two Hephaestus taxa, 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 highly adaptable, and while not much life-history data are available, Pusey et al. (2004) postulate the fish also needs a balanced flow regime, close to the natural regime as they observed several requirements: • Amniataba percoides has been observed to move during its life cycle, hence impoundments are a threatening process (Bishop et al., 1995). • Amniataba percoides show a preference for flowing water; however, high flows are likely to reduce population sizes. • While Amniataba percoides are not dependent on floodplain spawning like some other fish taxa, floodplain connections do increase population fitness. • Amniataba percoides need high spawning temperatures, and desynchronisation of flow and thermal regimes by impoundments can reduce their fitness. The ecological functions and their supporting flow requirements for grunters are summarised in Table 3-5. Table 3-5 Ecological functions for grunters 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 grunters The possible ecological outcomes of threatening processes on grunters in northern Australia are discussed and provided in the conceptual model (Figure 3-12). Four of the key threats in the conceptual model are related to flow modification – water harvesting, dam infrastructure and river regulation, with the added threat of 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, but also juvenile mortality 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 a change in sediment composition, which leads to habitat alteration. Similarly, Hephaestus fuliginos 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 is an effect exacerbated by land use change. Climate change will obviously interact with these threats as it will both enhance high flows that reduce populations, but droughts will also interact including with increased impacts by water extraction. The ecological outcomes of threatening processes on grunters in northern Australia are synthesised in the conceptual model below, considering their implications for changes to habitat, community structure and population size (Figure 3-12). Figure 3-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for grunters in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.3 Freshwater turtles 3.3.1 Turtles (Family: Chelidae) Description and background to ecology Freshwater turtles are one of the world’s more endangered taxonomic groups, with 52% of the global species extinct or threatened (Böhm M et al., 2013; Van Dijk, 2014). Freshwater turtles in Australia can be divided into three families, these are the 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 long periods of time. Chelids retract their necks sideways into their shells and their dietary habits vary between genera. Long-necked species, such as Chelodina are largely carnivorous, feeding on fish, invertebrate and gastropods (Legler, 1982; Thomson, 2000); while short-necked species, such as Elseya, are herbivorous or specialised to eat fruits (Kennett, 1993). Freshwater turtles depend upon flooded wetland systems for breeding, nesting, food provision and refuge. Changes to regional hydrology, habitat loss and climate change are some of their key threatening processes (Stanford et al., 2020). In northern Australia, turtles occupy a range of aquatic habitats, 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 of the turtle species in northern Australia have developed adaptive traits to survive in the interannual variation between the wet and dry seasons, such as the emergence of hatching with the wet season onset (Cann and Sadlier, 2017). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. During the dry season, the movements of the freshwater turtles on and off the floodplain are limited, 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. This includes the consumption of some species by Indigenous peoples as a seasonal source of protein (Jackson et al., 2012). A recent collaboration between Yangbala Rangers and the Atlas of Living Australia provided shared cultural values, and threats, ecological knowledge and distribution of freshwater turtles in the Roper catchment. This is the first time that regionally specific Aboriginal observational occurrence data and Aboriginal historical knowledge are included in the Atlas of Living Australia (Daniels et al., 2022). Indigenous people have widespread 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, 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 oblonga; previously known as Chelodina rugosa).The main threats to the freshwater turtles, identified by the Indigenous peoples, were natural predators (including birds of prey (such as eagles and hawks), crocodiles, goannas and dingoes), feral animals (such as pig, buffalo, horse, donkey, cattle and cane toad) and climate change (e.g. lower rainfall) (Russell et al., 2021). Freshwater turtles in the Roper catchment There are ten species of freshwater turtles described in the NT (NT Government, 2017). In the Roper catchment there are records of five freshwater turtle species: Gulf snapping turtle, northern snapping turtle (Elseya dentata), northern snake-necked turtle, Cann’s snake-necked turtle (Chelodina canni) and red-bellied short-necked turtles (Emydura subglobossa) (Figure 3-13). Until the recent collaboration with Yangbala Rangers, records for this area were sparse compared to many other regions of Australia. These turtles occur in different habitats, from permanently flowing riverine habitats to lakes, billabongs and swamps, from the Roper River mouth to Mataranka, but more surveys are required to assess their current distribution and conservation status in study area. Currently all five species are listed as Least concern by the Northern Territory Government; however, the northern snapping turtle is listed federally as Endangered by the EPBC Act. Due to the richness of natural resources across parts of the Roper catchment, this area was probably seasonally exploited in a fisher, hunter, gatherer economy, allowing large groups of people to gather for ceremony and other purposes. This is represented in rock art sites in the region, with at least one known site showcasing turtles (David, 2017; Earth Sea Heritage Surveys, 2013). Note the recognised Australian distribution of the pig-nosed turtle (Carettochelys insculpta) occurs in the western and northern draining catchments of the Gulf of Carpentaria in the NT and has been reported in the Roper catchment, although not confirmed (Georges et al., 2008). Similarly, the sandstone snake-necked turtle (Chelodina burrungandjii), currently listed as data deficient by the Northern Territory Government, is known to occur in the Wilton River, a tributary of the Roper River (Thomson et al., 2000). Figure 3-13 Location of freshwater turtles within the Roper catchment The freshwater turtle dataset was created from a collaboration between Ngukurr Yangbala Rangers, members of Ngukurr and Numbulwar communities (South East Arnhem Land), Macquarie University ecologists and the ALA through a series of mapping workshops and interviews to record local knowledge of the distribution in the South East Arnhem Land IPA of the freshwater turtles. Elseya lavarackorum distribution modelled through Species of National Environmental Significance (SNES). Data sources: ALA (2022); Department of Agriculture Water and the Environment (2019); Department of Environment Parks and Water Security (2019) Map. 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 breeding and nesting, food provision and refuge (Cann and Sadlier, 2017) (Table 3-6). In Australia, some species have an indirect flow dependency through habitat function driven by the flow regime. While for other species, the flow dependency supports critical phases of their life history, for example, the emergence of northern snapping turtle hatchings with the onset of the wet season (Cann and Sadlier, 2017). The nesting occurs during the dry season, in nests close to water (<4 m) in alluvial soils, sand or soil mix on steep to gently sloping banks. Nesting appears to occur at dispersed localities along the watercourse (Cann and Sadlier, 2017). Many of the turtle species in northern Australia have developed adaptive traits to survive in the highly variable wet-dry environment (Cann and Sadlier, 2017). The northern snake-neck turtle’s nesting behaviour starts in February (wet season) and it is completed by July (in the middle of the dry season). Eggs are laid in the mud, under shallow water, surrounded by flooded waterholes. Embryo development regresses while the eggs are in the water and continues once 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 are more vulnerable to changes in water quality, invasive species and habitat degradation (Cann and Sadlier, 2017). During the dry season, turtles often move to the shallows to aestivate. Weeks immediately before drying are the riskiest in terms of predation on turtles. 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). Also, feral pigs have an undesirable effect on turtle habitat, degrading aquatic ecosystems by upheaval of sediments, the destruction of aquatic vegetation, creation of anaerobic and acidic conditions, and enrichment wetlands with nutrients. Additionally, turbid conditions would limit visibility compromising hunting opportunities. Destruction of vegetation significantly alters production and respiration regimes, causing anoxic conditions and pH imbalances (Doupe et al., 2009). Freshwater turtles use large riparian zones to complete several aspects of their life cycle, such as nesting. Altering or eliminating these riparian habitats could reduce nest survival and, consequently, juvenile recruitment into the breeding population. It would also impact on adult survival through lack of feeding areas and refuge habitat for the dry season, and therefore increase the risk of extinction for freshwater turtle populations (Bodie, 2001). However, more comprehensive data for freshwater turtles are needed, especially on the timing and extent of riparian use. The ecological functions and their supporting flow requirements for freshwater turtles are summarised in Table 3-6. Table 3-6 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 critical for feeding and breeding areas for freshwater turtles (Cann and Sadlier, 2017; Cosentino et al., 2010; Gibbons et al., 2000; Marchand and Litvaitis, 2004). Fragmentation and habitat loss can make freshwater turtles more vulnerable, as they can disrupt their nesting sites and refugia, and also restrict their emigration and dispersal among wetlands (Bodie and Semlitsch, 2000; Bowne et al., 2006). Similarly, changes to hydrological patterns (timing, velocity, persistence and flow extent) through creation of barriers and water extraction could result in changes to the distribution of freshwater species, population growth and reproduction (Hunt et al., 2013). Determining how these drivers affect turtles is critical for improving environmental management and conservation at landscape scales (Bodie and Semlitsch, 2000). As turtle species from northern Australia are not as well studied as those in eastern Australia and elsewhere, mainly due to the remoteness of their habitats (Cann and Sadlier, 2017), much of 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 are discussed below, with their implications for changes to community structure, population viability and biodiversity and ecosystem function (Figure 3-14). Movement is critical for freshwater turtles’ access across breeding, feeding, estivation and refuge habitats (Ocock et al., 2018). Access to water and connectivity between suitable habitats is key by allowing the movement of the turtles within the river channels. Threats that reduce river–wetland connectivity, such as water harvesting, dam infrastructure or climate change, are some of the key threatening processes to the freshwater turtles in northern Australia (Figure 3-14) (Stanford et al., 2020). During wet to dry season transitions, freshwater turtles move on and off the floodplains. In perennial rivers, reduced dry-season baseflows (by extraction) could decrease the availability of suitable habitat that is supported by flows. Such a baseflow reduction 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 a reduced baseflow hamper the freshwater turtles’ movements on and off the floodplain during the transition between wet to dry season (Warfe et al., 2011). Impoundment, regulation and channelisation of riverbanks and beaches can result in the reduction of 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 impoundments (Tucker et al., 2012; Waltham et al., 2013). Similarly, if migratory routes are interrupted through habitat deterioration (especially of nesting sites), the gene flow between populations can be disrupted (Alho, 2011; Lees et al., 2016), which reduces the genetic diversity of populations. Changes to the inundation and flow regime reduce freshwater turtles’ feeding and the suitability of habitats such as waterholes (Warfe et al., 2011), which increases the 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). Channelisation of rivers and shoreline hardening may eliminate nesting and basking areas and alter the hydrodynamic processes that maintain critical nesting habitat (Roosenburg, 2014). Removal of exposed logs and snags to promote recreational boating eliminates critical basking sites and prey habitat (Lindeman, 1999). Changes in flow regimes due to water use and regulation can also impact freshwater turtles by disrupting breeding cues, and by 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 would probably lead female turtles to seek alternate, less-suitable habitat (Bodie, 2001) that can lead to less successful recruitment. The loss of optimal nesting habitat due to fluctuating water levels through water management infrastructure can inundate freshwater turtles’ nests, resulting in egg mortality (Waltham et al., 2013). Therefore, reduced breeding success, survival and population size of freshwater turtles can have an impact on the community and population structure (Georges et al., 1993; Tucker et al., 2012). Similarly, rapid shifts in temperature might preclude successful gradual responses that functioned historically, like active modification of geographic range. Also, early nesting and early egg maturity, due to temperature rises, result in the eggs perishing in the ground, while late nesting risks the eggs being prematurely flooded by rising waters (Jolly, 2008). Refuge habitats during extended drought conditions are critical, as estivation is limited by fat reserves and dehydration, rarely lasting more than 7 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 inundation in floodplains. They can also act as barriers to downstream sediment transmission, resulting in a reduction in in the size and abundance of sandbars (Pusey and Kennard, 2009). Groundwater discharge via springs can be important for maintaining perennial river baseflow in the dry season. Persistent surface water provides important refuge habitat for freshwater turtles (Warfe et al., 2011). Figure 3-14 Conceptual model showing the relationship between threats, drivers, effects and outcomes for freshwater turtles in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.4 Freshwater waterbird groups 3.4.1 Grouping waterbirds The 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, social organisation, 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, migrating 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 (Grus 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, driven by changes including 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 other anthropogenic impacts. Consequently, waterbirds are a focal group for the conservation and Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 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 grouped into four high-level groups. These groups are based on foraging behaviour and habitat dependencies, together with 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 migrate annually to take advantage of foraging opportunities and avoid the northern hemisphere winter. The four waterbird groups are: 1. colonial and semi-colonial nesting waders 2. shorebirds 3. cryptic waders 4. swimmers, grazers and divers. Group 1: ‘Colonial and semi-colonial nesting waders’ (Section 3.4.2). Colonial and semi-colonial 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 3.4.3). 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 3.4.4). Shorebirds have a high level of dependence on end-of- system flows and large inland 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, mostly breed in the northern hemisphere rather than Australia, and are a group of international concern. Group 4: ‘Swimmers, grazers and divers’ (Section 3.4.5). These are species with a relatively high level of dependence on semi-open, open and deeper water environments, who commonly swim when foraging (including diving, filtering, dabbling, grazing) or when taking refuge. These species 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 and are presented in Table 3-7. Their selection is based on whether they 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 synthesizing the pathways to impact associated with potential water resource developments. Table 3-7 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 habitat loss, fragmentation and change; toxins from pollution or contaminants; disturbance from human activities; predation by invasive or feral animals; and changes in disease, or parasite burdens. Habitat loss, fragmentation and change are the most important drivers 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, with nestling and fledgling predation 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, 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 cause changes in bird behaviour and affect temporal and spatial distribution of waterbirds. Human disturbance can be equivalent to habitat loss or degradation because waterbirds may avoid or underuse areas (Fernandez and Lank, 2008). Temporary loss of foraging habitats can occur with the capacity to compensate by foraging for longer periods, varying between species (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 affect population sizes and trajectories. Disease and parasites can affect waterbird nest success, fledging rates, juvenile survival and adult survival, and 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 there are high infection rates of Campylobacteria in waders (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 result in a stronger concentration of 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 (e.g. predation, disease) 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, which all have the potential to influence waterbird populations positively and negatively, and directly and indirectly (Chambers et al., 2005; Sutherland et al., 2012). 3.4.2 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 that have 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 3-8). The species in this group are often easily detectable when breeding and relatively good datasets are available for most, compared to other species or groups. The species in this group are often dependent 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 of birds 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 significant distances to use these sites, ranging up to thousands of kilometres (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 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 3-15). 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 3-15 Royal spoonbill individual at the nest Royal spoonbills are a representative species of the colonial and semi-colonial nesting waders waterbird group. Photo attribution: CSIRO Colonial and semi-colonial nesting waders in the Roper catchment Colonial and semi-colonial nesting waders, including the royal spoonbill, are found widely throughout the Roper catchment (Figure 3-16 and Figure 3-17). The large wetlands and extensive mangroves including areas throughout Limmen National Park support a range of colonial and semi-colonial nesting waders (Delaney, 2012). Aerial surveys by Chatto (2006) found large numbers of egrets (including Egretta spp.) and the red-necked avocet (Recurvirostra novaehollandiae). The mangrove habitats along the north side of the Roper River supports significant colonies of great egret (Ardea alba), intermediate egret (Ardea intermedia), little egret Photo royal spoonbill. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. (Egretta garzetta) and pied heron (Egretta picata), with the nankeen night-heron (Nycticorax caledonicus) found along the banks of the river in the mangrove habitat as far as Ngukurr (Smyth and Turner, 2019). The Roper River is also considered to be a major breeding area for brolgas (Antigone rubicunda) during the wet season (Chatto, 2006). Permanent waterholes and wetlands around Mataranka also support a variety of colonial and semi-colonial nesting waders (Figure 3-16 and Figure 3-17). Figure 3-16 Location of colonial and semi-colonial nesting wading waterbirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-17 Location of colonial and semi-colonial nesting wading waterbirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Table 3-8 Species in the colonial and semi-colonial nesting wading waterbird group, and their national and international conservation status (LC = Least concern) SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME IUCN STATUS Australian white ibis Threskiornis moluccus Threskiornithidae LC Banded stilt Cladorhynchus leucocephalus Recurvirostridae LC Black-winged stilt (pied stilt) Himantopus himantopus (Himantopus leucocephalus) Recurvirostridae LC Cattle egret Bubulcus ibis (Ardea ibis) Ardeidae LC Eastern reef egret Egretta sacra Ardeidae LC Glossy ibis Plegadis falcinellus Threskiornithidae LC Great Egret (eastern great egret) Ardea alba (Ardea modesta, Ardea alba modesta) Ardeidae LC Great-billed heron Ardea sumatrana Ardeidae LC Intermediate egret Ardea intermedia Ardeidae LC Little egret Egretta garzetta Ardeidae LC Nankeen night-heron Nycticorax caledonicus Ardeidae LC Pied heron Egretta picata (Ardea picata) Ardeidae LC Red-necked avocet Recurvirostra novaehollandiae Recurvirostridae LC Royal spoonbill Platalea regia Threskiornithidae LC Sarus crane Grus Antigone Gruidae Vulnerable Straw-necked ibis Threskiornis spinicollis Threskiornithidae LC White-faced heron Egretta novaehollandiae Ardeidae LC White-necked heron Ardea pacifica Ardeidae LC Yellow-billed spoonbill Platalea flavipes Threskiornithidae LC Black-necked stork Ephippiorhynchus asiaticus Ciconiidae LC Brolga Antigone rubicunda Gruidae 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 3-18) (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 and their supporting flow requirements for colonial and semi-colonial nesting waders are summarised in Table 3-9. Table 3-9 Ecological functions for colonial and semi-colonial nesting waders and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-18 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 3-19). 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/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 result in reduced water quality and food availability, and increased 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), or 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; hence when these regimes are changed, life histories are impacted. Photo egret. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-19 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, while black arrows represent non-hydrological changes. 3.4.3 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 3-20). Their habitats require 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 3-10). 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). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 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 3-20 Dense aquatic and semi-aquatic vegetation used as habitat by cryptic wading waterbirds This habitat provides protection from predators and weather. Photo attribution: CSIRO Table 3-10 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 IUCN STATUS Australian little bittern Ixobrychus dubius (Ixobrychus minutus) Ardeidae LC Australian painted snipe Rostratula australis Rostratulidae Endangered Australian spotted crake Porzana fluminea Rallidae LC Photo aquatic habitat. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. SPECIES NAME SPECIES SCIENTIFIC NAME FAMILY SCIENTIFIC NAME IUCN STATUS Baillon's crake Porzana pusilla (Zapornia pusilla) Rallidae LC Black bittern Ixobrychus flavicollis Ardeidae LC Buff-banded rail Hypotaenidia philippensis Rallidae LC Chestnut rail Eulabeornis castaneoventris (Gallirallus castaneoventris) Rallidae LC Latham's snipe Gallinago hardwickii Scolopacidae LC Lewin's rail Lewinia pectoralis Rallidae LC Red-necked crake Rallina tricolor Rallidae LC Spotless crake Zapornia tabuensis (Porzana tabuensis) Rallidae LC Striated heron Butorides striatus (Butorides striata) Ardeidae LC White-browed crake Amaurornis cinerea (Poliolimnas cinereus) Rallidae LC Cryptic wading waterbirds in the Roper catchment Cryptic waders are found throughout the Roper catchment, particularly in the Roper River estuary and the area around Mataranka (Figure 3-21). Only five species of cryptic waders are recorded in the Roper catchment, with the black bittern (Ixobrychus flavicollis) the most common (Atlas of Living Australia, 2021a). Figure 3-21 Location of cryptic wading waterbirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. 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 and nests are therefore at risk when water levels change (Garnett et al., 2015; Marchant and Higgins, 1990). They 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 result in reduced water quality and food availability, and increased rates of competition, predation and disease (McGinness, 2016). Such changes can occur when water extraction directly from these habitats occurs, 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 and their supporting flow requirements for cryptic wading waterbirds are summarised in Table 3-11. Table 3-11 Ecological functions for cryptic wading waterbirds and their supporting 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; however, 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 pathway of potential water resource development impact on cryptic waders is 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 3-22). Changes in hydrology and vegetation also put these species at risk from increased disturbance from human activities and predation, in addition to pure disturbance (Kingsford and Norman, 2002). Human disturbance can be equivalent to habitat loss or degradation because waterbirds may 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 affect 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, competing for habitat or food, or affecting other predators and prey (Cruz et al., 2013; MacDonald and Bolton, 2008; Skorka et al., 2014). Figure 3-22 Conceptual model showing the relationship between threats, drivers, effects and outcomes for cryptic wading waterbirds in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.4.4 Shorebirds Description and background to ecology The ‘shorebirds’ group consists of waterbirds that have 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 have a piscivorous or invertivorous diet and 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 are dependent 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 3-12). Approximately 35 species are common, regular visitors or residents. Several species in this group Conceptual model. 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, curlew sandpiper, eastern curlew, great knot, lesser sand plover and red knot. The eastern curlew is listed as Critically endangered 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 (Figure 3-23) 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 coasts. 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 3-23 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 3-12 Species in the shorebirds group, and their national and international conservation status (LC = Least concern) 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. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Shorebirds in the Roper catchment The Roper catchment is one of the most important areas in the NT for shorebirds, and is a major stopover area for migratory species, especially godwits (Limosa sp.) and knots (Calidris sp.) (Delaney, 2012; Department of Agriculture‚ Water and the Environment, 2021). The intertidal mudflats and coastal flats provide important habitat for shorebirds, as do the large open shallow wetlands (Chatto, 2006). A survey of shorebirds by Chatto (2003) found red knot (Calidris canutus), great knot (Calidris tenuirostris), red-necked stint (Calidris ruficollis) and lesser sand plover (Charadrius mongolus) were common in the coastal area of the Roper catchment. Shorebirds are also commonly found in the area around Mataranka and Daly Waters, as well as throughout the catchment (Figure 3-24, Figure 3-25 and Figure 3-26). Figure 3-24 Location of shorebirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-25 Location of shorebirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-26 Location of shorebirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. 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 and 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 and their supporting flow requirements for shorebirds are summarised in Table 3-13. Table 3-13 Ecological functions for shorebirds and their supporting 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 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, through high food availability, 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 act as fuel before the 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, temporary loss of foraging habitats can occur and 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 Roper catchment are illustrated below, with their implications for changes to biodiversity and ecosystem function (Figure 3-27). Figure 3-27 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the shorebirds group in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 3.4.5 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, who 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 3-14). 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 breeding while usually seasonal can be stimulated by flooding or large rainfall events (Kingsford and Norman, 2002). Species diets may be piscivorous, omnivorous or herbivorous (Barker and Vestjens, 1990). Changes in water depth, water extent or water duration can result in nest exposure to predation, drowning of nests, or reduced food availability, resulting in nest failure (McGinness, 2016; Poiani, 2006). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. The blue-billed duck (Oxyura australis) is one example of the swimmers, grazers and divers group and is nationally listed as Near threatened. The magpie goose (Anseranas semipalmata) is also part of this 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 (Figure 3-28) is an ancient and unique species of particular importance to Indigenous people, 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 enhance the impacts of such change in northern Australia for this species (Poiani, 2006; Traill et al., 2009a). Figure 3-28 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 3-14 Species in the swimmers, grazers and divers waterbird group, and their national and international conservation status (LC = Least concern) 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. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Swimmers, grazers and divers in the Roper catchment Swimmer, grazer and diver waterbirds are found throughout the Roper catchment, and are common in the estuary of the Roper River, continuing upstream, and at Mataranka (Figure 3-29, Figure 3-30 and Figure 3-31). A survey of the Roper River mouth and surrounding wetlands by Chatto (2006) found significant numbers of magpie goose (Anseranas semipalmata), Australian pelican (Pelecanus conspicillatus), whiskered tern (Chlidonias hybrida) and white-winged black tern (Chlidonias leucopterus). The pied cormorant (Phalacrocorax varius) and grey teal (Anas gracilis) were also common (Chatto, 2006). Figure 3-29 Location of swimmer, grazer and diver waterbirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-30 Location of swimmer, grazer and diver waterbirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Maps. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 3-31 Location of swimmer, grazer and diver waterbirds in the Roper catchment Data source: Atlas of Living Australia (2021a) Flow–ecology relationships for swimmer, grazer and diver waterbirds Waterbird species in the swimmer, grazer and diver waterbird 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 3-15) (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 result in reduced water quality and food availability, and increased rates of competition, predation and disease (Douglas et al., 2005; McGinness, 2016). Such changes can occur when water extraction directly from these habitats occurs, 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 and their supporting flow requirements for swimmer, grazer and diver waterbirds are summarised in Table 3-15. Table 3-15 Ecological functions for swimmer, grazer and diver waterbirds and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Maps. 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 swimmer, grazer and diver waterbirds The primary pathways of potential water resource development impact on waterbirds in the swimmers, grazers and divers group include: habitat loss, fragmentation and change, including water depth changes and weed invasion changing habitats; climate change and extremes – including inundation of freshwater habitats by seawater when river flows are reduced, and inundation of nests by extreme flood events; toxins from pollution or contaminants; disturbance and hunting from human activities; predation by invasive or feral animals; and 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 3-32). Reduced extent, depth and duration of inundation of waterhole and other deep- water environments is likely to reduce habitat availability and food availability for this group, increasing competition and predation and also increasing risk of disease and parasite spread. Conversely, species in this group such as magpie geese that nest at water level or just above 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 3-32 Conceptual model showing the relationship between threats, drivers, effects and outcomes for the waterbird group swimmers, grazers and divers in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 4 Ecological assets from marine systems The marine and estuarine habitats of northern Australia include some of the most important, extensive and intact habitats of their type in Australia, many of which are of national significance. Marine habitats in northern Australia are vital for supporting important fisheries including the common banana prawn, mud crab and barramundi, as well as for biodiversity more generally, including waterbirds and marine mammals and turtles. In addition, the natural waterways of the sparsely populated catchments support globally significant stronghold populations of endangered and endemic species (e.g. sharks and rays) that use both marine and freshwater habitats. This section provides a synthesis of the prioritised assets relevant to marine sections of the Assessment catchment. Table 1-1 presents the full list of assets used in the ecology assessment. The Roper River Marine region as considered here is an area south of Groote Eylandt that depends upon recruitment from littoral habitats from the Roper to the mainland coast to the west of Groote Eylandt. 4.1 Marine habitats 4.1.1 Mangroves Description and background to ecology Mangroves are a group of woody plant species, 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, they 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 ranging from molluscs (McClenachan et al., 2021), crustaceans (Guest et al., 2006; Thimdee et al., 2001) and birds (Mohd-Azlan et al., 2012)), to 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, use them 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 4.2.1), barramundi (Section 3.2.1), mud crabs (Section 4.2.3), threadfin (Section 4.3.3) and mullet (Section 4.3.1) (Blaber et al., 1995; Brewer et al., 1995). Mangrove forests provide a diverse array of ecosystem services including stabilising shoreline areas from erosion and sever weather events (Zhang et al., 2012), and 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 ~44 to 1022 g carbon m-2 y-1 from leaves and 912 to 6870 g carbon m-2 y-1 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 storage of mangrove carbon, either through burial in their burrows and uptake directly into production. The decomposition and processing of mangrove material is important also in the cycling of nutrients, which if consumed and released support both a local food web (Abrantes et al., 2015; Guest et al., 2004), but also some of the organic carbon can be transported offshore supporting fisheries production more broadly (Connolly and Waltham, 2015; Dittmar and Lara, 2001; Lee, 1995). Mangroves in the Roper catchment and marine region In a study by Lymburner et al. (2020) the extent of mangroves in Australia was mapped by using 25-m spatial resolution Landsat sensor data that determined that there is 11,142 km2 + 57 km2 (95% confidence interval (CI)) in 2017, which is down slightly from the extent from 2010 (11,388 km2 + 38 km2 (95% CI). Most of the change was found to have occurred along the northern Australian coastline and 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 invoked mangrove loss in northern Australia, the most significant and obvious 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). Mangroves in the Roper River estuary are restricted to a narrow fringe immediately along both sides of connecting tidal channels and main estuary as shown Figure 4-1. Limmen Bight at the mouth of the Roper River forms a highly important habitat system of tidal wetlands and estuaries and while the whole of site is tidal, it receives large volumes of freshwater inflows from the contributing catchments. Limmen Bight has extensive estuarine communities including mangroves that are highly productive, have high cultural value, and are often of national significance (Bradley, 2018; Department of Agriculture, 2019a; Poiner et al., 1987). These habitats at the mouth of the Roper River contain 19 mangrove plant species fringing the banks of streams and rivers (Palmer and Smit, 2019). The spatial distribution illustrates the extent of mangroves in this river, however, it is important to note that the extent of mangroves was altered as part of the wide-scale dieback of mangroves in the Gulf of Carpentaria between late 2015 and early 2016 (Duke et al., 2017). Fish and crustacean communities occupying mangrove forests in the Roper catchment marine region (Larson et al., 2013; Staples, 1979), though most of the same species that occur in this estuary have also been reported in other estuaries in the Gulf of Carpentaria (Blaber et al., 1995; Brewer et al., 1995). Species such as the mud crab can be found occupying mangrove forest areas (Robins et al., 2020), and during periods where tidal connection permits access, a number of fish species would also access the mangroves for shelter and food – which is no different to mangrove forests on 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, particularly when compared to the extent of intertidal saltpans (Section 4.1.2), they still provide important linkage to coastal fisheries production, carbon sequestration and habitat for local wildlife, in addition to providing erosion and sediment accumulation. The near-coastal waters and estuaries of the Roper catchment support a commercial barramundi fishery, while harvest of mud crabs (mainly Scylla serrata) also occurs along the coasts near Port Roper (Bayliss et al., 2014). Figure 4-1 Location of mangroves in the Roper catchment marine region Data source: Geoscience Australia (2017); NVIS Technical Working Group (2017) Map. 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 this to the coastal zone. This can be detrimental to some coastal habitats (e.g. seagrass beds that can be smothered thereby inhibiting 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 the accumulation of carbon in sediments (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 moisture content changes, particularly if soils dry out and the moisture content reduces, mangroves can be impacted greatly. The large mangrove dieback in the Gulf of Carpentaria is a case in point where soil moisture content was low as a consequence of lower sea levels and mangroves were not able to access water (Duke et al., 2017). Given mangroves are connected to 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 moisture 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 4-1. Table 4-1 Ecological functions mangroves and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for mangroves There are several threatening processes that can impact mangroves in northern Australia. These include 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 river regulation driven, but as 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 were not able to 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 impact 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 to the structure, function, sedimentation and biodiversity of mangrove communities. Building dams and other hydrological barriers also affect 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 lead to a reduction in spring flows and a lowering of the groundwater table. Coastal wetlands, including mangroves, are particularly vulnerable to climate change (Feller et al., 2017). Climate change 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) impacting on the hydrology of a system can alter the baseflow and flood patterns (Erwin, 2009). These hydrology changes can also impact 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, but also reports of this same event has been reported elsewhere in northern Australia (Duke et al., 2017; Lovelock et al., 2017). Changes to the hydrology and temperature can impact 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. There are many examples of mangrove loss in developing areas (Xu et al., 2019), which has contributed to loss of mangroves directly or because of changes in hydrology and flow, causing increased erosion. These can include modification to land management practices; changes to the intensity or type of agricultural production; increased vegetation clearing; and increased 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 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 impacted by the loss of mangroves (Sheaves et al., 2016). The ecological outcomes of threatening processes on mangroves in northern Australia are illustrated below, with their implications for changes to biodiversity and ecosystem function (Figure 4-2). Figure 4-2 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mangroves in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.1.2 Saltpans and salt flats Description and background to ecology Saltpans and salt flats are intertidal areas that are devoid of marine plants located between mangrove and saltmarsh meadows. Saltmarshes (Figure 4-3) are the result of intertidal areas that are covered only infrequently with tidal inundation of coastal land in the supra-littoral zones, when 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), as well as metapenaeid shrimps (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). During high-level overbank flood flows, saltpans would be most productive. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. The inundation of saltpans expands the available habitat to hitherto estuarine benthic fish and crustaceans, provided they can tolerate euryhaline conditions. In northern Australia, coastal saltpan extent can be tens to hundreds of square kilometres. They provide habitat for a range of benthic infauna (Dias et al., 2014), which are an important food source for high-order consumers including shorebird species that use saltpans as either 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 5–10 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 Section 3.4.4) 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 4-3 Saltpan area in northern Australia, which are generally located between mangrove and saltmarsh areas Photo attribution: Nathan Waltham Saltpans in the Roper catchment and marine region The Limmen Bight Tidal Wetlands System Directory of Important Wetlands in Australia (DIWA) site is the second-largest area of saline coastal flats in the NT (1848 km2, excluding subtidal seagrass areas) (Department of Agriculture, 2019a). Limmen Bight forms a highly important habitat system of tidal wetlands (intertidal mud flats, saline coastal flats and estuaries) and while the whole of the site is tidal, it receives large volumes of freshwater inflows from the contributing catchments. Saltpans in the Roper River estuary are restricted to a tidal inundation area on the landward side of the mangroves that line the main river channel (Figure 4-4). The spatial data presented Photo of saltpan. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. illustrates the extent of saltpans in this river, however, it is important to note that the extent was altered as part of the wide-scale dieback of mangroves in the Gulf of Carpentaria between late 2015 and early 2016 (Duke et al., 2017). Coastal habitats have high cultural value and are often of national significance (Bradley, 2018; Department of Agriculture, 2019a; Poiner et al., 1987). The intertidal flats in the Roper catchment marine region are extensive (Palmer and Smit, 2019) and provide an important feeding ground for migratory shorebirds in the NT, with counts in the tens of thousands (Palmer and Smit, 2019). Staples (1979) surveyed the banana prawn community in this river system, and Larson (1996) examined estuarine fish, while Thorburn et al. (2003) recorded elasmobranchs in the Roper River. Figure 4-4 Location of salt flats in the Roper catchment marine region Data source: Geoscience Australia (2017) Map. 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 when soil moisture content changes, particularly if soils dry out and the moisture content reduces, which causes these areas to become hypersaline in the surface soils. Given their connection 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 moisture 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 and their supporting flow requirements for saltpans are summarised in Table 4-2. Table 4-2 Ecological functions of saltpans and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for saltpans There are several threatening processes that can impact on 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). Water 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, and not 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., 2017). However, Asbridge et al. (2016) outlines, importantly, the replenishment of mangrove habitats due to natural flows in the southern Gulf – 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 impacted (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; Waltham et al., 2020). Changes in this connectivity could alter soil moisture 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 are illustrated below, with their implications for changes to biodiversity and ecosystem function (Figure 4-5). Figure 4-5 Conceptual model showing the relationship between threats, drivers, effects and outcomes for saltpans in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.1.3 Seagrass habitats Description and background to ecology Seagrasses are marine flowering plants that provide valuable habitat and food resources to a diverse community of animals, such as invertebrates, fish, dugongs, sea turtles and many other Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. marine organisms. In northern Australia, there are 15 species of seagrasses in tidal reaches of rivers, coastal, reef and deep-water habitats. Their distribution is limited by light and suitable substrate availability, with most seagrasses found in shallow inshore and intertidal zone areas to water depths of around 25 m (but seagrasses can occur in depths of 50–60 m) (Carruthers et al., 2002). Seagrass species are not currently listed for conservation. However, they provide crucial habitat for threatened species such as the dugong and marine turtles including the green (Chelonia mydas), flatback (Natator depressus), olive ridley (Lepidochelys olivacea) and hawksbill (Eretmochelys imbricata) turtles. Commercially important species, like prawns, are also highly associated with this habitat. In addition to being a critical food for dugongs and some species of marine turtles (Carruthers et al., 2002; Morgan et al., 2017), seagrasses provide a substrate for epiphytic algae, which provides 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 4.2.1 and 4.2.4), where juvenile prawn production in the north-east Gulf of Carpentaria seems to be mostly based on seagrass-derived organic matter (Loneragan, 1997). Seagrasses provide marine animals with shelter from currents and predation and stabilise bed sediments. They are important in providing nursery habitat for juvenile fish and prawns, including commercially important species (Coles, 2004; Roelofs, 2005; Unsworth, 2019). Additionally, seagrasses influence coastal morphology by reducing the water velocity near the bottom, promoting sediment accretion, and therefore importantly, reducing erosion (Fonseca and Cahalan, 1992). Seagrasses are considered as one of the most valuable ecosystems globally, due to their provision of key ecological services, such as climate regulation (Duarte et al., 2013), carbon sinks (Fourqurean et al., 2012) and fisheries habitat (Unsworth, 2019), nutrient cycling, enhanced biodiversity, and sediment stabilisation (Orth, 2006). Yet, seagrasses globally are declining rapidly (with a loss of 29% of their known areal extent since 1879) (Waycott et al., 2009) due to a range of 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 Roper catchment marine region Australia’s seagrass meadows constitute a large proportion of the planet’s known seagrass species (Green and Short, 2003), with half (37) of the world’s 72 seagrass species found in Australian waters (Kilminster et al., 2015; Short et al., 2011). In the Gulf of Carpentaria, the seagrass distribution is very fragmented, having aggregated seagrass patches with many bare areas between them (Figure 4-6). There relatively few meadows that have continuous seagrass cover, although these tend to also be large meadows (Roelofs, 2005). There have been two main expeditions to map seagrass in the Roper catchment marine region (Poiner, 1987; Roelofs, 2005). Differences between surveys are interpreted as a consequence of using different methodology between surveys and due to potential disturbances from tropical storms and cyclones. While this section of broader coastline is complex, it has few river inputs, making the water turbid than other regions in the Gulf of Carpentaria. The dominant cover type for the Roper catchment marine region was meadows with aggregated seagrass patches. The seagrass meadows from Limmen Bight were found to typically have high numbers of feeding flatback turtles (50+ on some meadows) and several dugongs (Roelofs, 2005). These habitats are recognised as being important for dugong in the western Gulf of Carpentaria including important dreaming (Bradley, 2018). Five seagrass species have been identified in the Roper catchment, all of them listed as ‘Special least concern’ by the Queensland Nature Conservation Act 1992. Open-coastline communities were the major intertidal meadow types mapped and they were dominated by monospecific stands of Halophila ovalis and Halodule uninervis intertidally (Poiner, 1987; Roelofs, 2005). Mixed- species meadows of Cymodocea rotundata and Thalassia hemprichii occurred on reef flats, while some small sheltered embayments were dominated by Enhalus acoroides and Halophila ovalis (Poiner, 1987; Roelofs, 2005). Figure 4-6 Distribution of seagrass habitats in the Roper catchment marine region Data sources: ALA (2022); Department of Agriculture Water and the Environment (2019); Department of Environment Parks and Water Security (2019) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Flow–ecology relationships for seagrass habitats River regulation can change the water quality of flows discharging into coastal regions. Seagrasses are negatively impacted by changes in water quality, in particular by increases in turbidity by larger flood pulses (Turschwell et al., 2021). Turbidity is influenced by runoff and basin management and plume movement among other factors. Such events can smother seagrass or inhibit penetration of sunlight. Large flood events from rivers can create significant sedimentation events, which negatively affect seagrass beds and can result in changes in their distribution and vigor. The ecological functions and their supporting flow requirements for sea grass habitats are summarised in Table 4-3. Table 4-3 Ecological functions for seagrass habitats and their supporting flow requirements For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Pathways of change for seagrass habitats Four key threatening processes have been assessed to negatively impact seagrass species in northern Australia: land use change, climate change, river regulation and demersal fisheries (Turschwell et al., 2021) (Figure 4-7). Although not all seagrass species are affected equally 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., 2021). Agricultural practices that lead 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). But also, increasing nutrient loads influence changes in seagrass community composition, due to the different responses of various species to these changes (Burkholder et al., 2007; Cardoso et al., 2004). Certain herbicides and pesticides commonly used in northern Australia have been found to be 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. More recently, new 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. They identified that climate change induced sea-level rise and temperature changes might have significant impacts on seagrass species and their distribution. If the 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) and showed that benthic irradiance (which is closely correlated with depth) and wave height are sufficient to map seagrass extent with 83% accuracy. Beaman et al. (2016) demonstrated that water depth is a key determinant of benthic habitat in the Coral Sea. Similar work has demonstrated the relationship in marine habitats around the world. In terms of 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)), and under experimental conditions, it has been shown that some tropical seagrasses are more susceptible to toxins when subject to thermal stress (Koch and Erskine, 2001). Changes in temperature can lead to changes in the seagrass community composition, as some species thrive at higher temperatures, while others cannot survive with elevated temperatures (e.g. Campbell et al. (2006); Evans et al. (1986) Mcmillan (1984)). River regulation can change the water quality. Seagrasses are generally intolerant of fresh water for more than brief periods of exposure (though degree of tolerance varies by species) (Adams and Bate, 1994; Collier et al., 2014), and 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., 2021). The ecological outcomes of threatening processes on seagrass habitats in northern Australia are illustrated below, with their implications for changes to population viability and community structure (Figure 4-7). Figure 4-7 Conceptual model showing the relationship between threats, drivers, effects and outcomes for seagrass in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.2 Marine crustacean species and species groups 4.2.1 Banana prawns (Penaeus merguiensis) Description and background ecology Banana prawns are large-bodied decapod crustaceans around 80 g in size 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, Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. the common banana prawn (Penaeus merguiensis) and the redleg banana prawn (Penaeus 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 prolific 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 (recent 10-year mean) of the NPF (worth about $70–80 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 offshore from the Roper River (Laird, 2021). The influence of rainfall and runoff from 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). Postlarval 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 4.1.1) 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, Mollusca, Crustacea, Annelida), 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) and 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 euryhaline 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 >12 mm carapace length (CL) emigrate when salinity is about 30–35 parts per thousand (ppt), while prawns in the size range ≤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 = about 85 g) and about 47 mm CL for males (40 mm CL = about 50 g). Banana prawns in the Roper catchment marine region Adult common banana prawns live and spawn offshore in waters 10–30 m deep, the larvae and postlarvae move by drift 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 from Blue Mud Bay (north of the map extent of Figure 4-8) to the south-east Gulf of Carpentaria support abundant populations of juvenile banana prawns (Staples, 1979). Common banana prawns are found at highest densities offshore from the Roper River in relatively shallow waters, as well as south of Groote Eylandt in deeper water (Figure 4-8). Common banana prawns are abundant elsewhere in the western Gulf of Carpentaria in Blue Mud Bay north of Groote Eylandt and offshore of the McArthur River to the south-east of the Roper catchment marine region. The Roper catchment marine region lies within the southern portion of the ‘Groote’ NPF statistical region of the Gulf of Carpentaria (adjacent to the Blue Mud Bay and Roper River coasts). This statistical region accounts for about 2% (about 95 t – 16-year mean catch) of the total NPF banana prawn catch (Laird, 2021). In all locations, the highest abundances of banana prawns were caught inshore in about 15–20 m depth, in proximity to the river estuaries (Zhou et al., 2015). In the 1970s, the use of mangrove habitats by juvenile banana prawns within the Roper River estuary was documented by Staples (1979) using a float plane to access a series of rivers in the region. However, the remoteness of the river systems in the western Gulf of Carpentaria rendered both the estuarine habitats and their fish and crustacean fauna poorly studied. Knowledge of estuarine banana prawn habitats from other Gulf of Carpentaria rivers showed that the mangrove forest and creek mudbank habitats (indicated as juvenile habitat in Figure 4-8) are critical for juvenile banana prawn survival and growth (Staples, 1979; Vance et al., 1990). These habitats are prolific within the estuaries of many rivers in the western Gulf of Carpentaria, including the Roper River, as well as other rivers along the south-west Gulf of Carpentaria coastline such as the Limmen Bight and McArthur River (Duke et al., 2017). Figure 4-8 Fisheries catch of banana prawns and their habitat in the Roper catchment marine region Banana prawn juveniles use the estuary and adult prawns are caught offshore in water about 10-20 m deep in the marine habitat. Units are kilograms as total catches for the 10-year period 2011 to 2020. Data sources: Kenyon et al. (2022) and Staples (1979) Map. For more information on this figure please contact CSIRO on enquiries@csiro.au 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 postlarvae use currents to move shoreward to river estuaries (Vance and Rothlisberg, 2020). Prior to the annual wet season, postlarvae 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 prawn 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 4-4). 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 in marine habitats is lower (Gwyther, 1982), the overall result is a larger adult population of banana prawns in coastal marine habitats the higher the flood flows from adjacent estuaries (Duggan et al., 2019). Both high-level pulsed flood flows during the monsoon season and low-level ‘early-season’ flows, sustained flows during the wet season and the persistence of wet-season flows 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 and Joseph Bonaparte Gulf ecosystems are stressed, often hypersaline habitats during the latter dry season (Kenyon et al., 2004; Vance et al., 1990). They 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 4-4). 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 4-4. Table 4-4 Ecological functions for banana prawns and their supporting 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 banana prawns The life history of banana prawns would be significantly impacted 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, the maintenance of estuarine brackish habitats, diversity of river flow regimes and high-pulse flood flows enhance the population 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 retainment of, low-level flows removes a large proportion of ‘early-season’ low-level river flows with subsequent impacts on estuarine banana prawns. The interruption of 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 prior to 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 during 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 ecological outcomes of threatening processes on banana prawns in northern Australia are illustrated below, with their implications for changes to growth and mortality, community structure, habitat and population (Figure 4-9). Figure 4-9 Conceptual model showing the relationship between threats, drivers, effects and outcomes for banana prawns in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.2.2 Endeavour prawns (Genus: Metapenaeus) 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 10–40 m deep, the larvae and postlarvae 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 widespread 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 (north of the map extent in Figure 4-10) to Mornington Island and juvenile blue endeavour prawns are abundant within the habitat (Kenyon, 1999; Poiner et al., 1987). Seagrass habitats (see Section 4.1.3) are critical to juvenile endeavour prawn survival. They provide habitat structure for shelter and Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 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 is dependent on seagrass species and morphological type (Haywood et al., 1998; Kenyon et al., 1995). Juvenile endeavour prawns emigrate offshore at about 10–20 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 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 the majority of the carbon source in their diet from seagrass and seston rather than mangroves (Loneragan et al., 1997). Their diet likely is small bivalves, gastropods, ophiuroids, crustaceans and polychaete worms, similar to tiger prawns (Heales et al., 1996; Wassenberg and Hill, 1987). 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). Endeavour prawns support a 500-t component (recent 10-year average) 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 = about 40 g) for females and about 34 mm CL (34 mm CL = about 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 Roper catchment marine region Two species of endeavour prawns are found in the Gulf of Carpentaria: Metapenaeus endeavouri (blue endeavour prawn) and M. ensis (red endeavour prawn). 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 distributions of the two species are relatively allopatric, dependent 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 western Gulf of Carpentaria in coastal waters about 10–45 m deep, offshore from the Roper catchment (Buckworth, 1992) (Figure 4-10 and Figure 4-11). The juvenile phase of both species inhabit the seagrass community (and other shallow coastal habitats), which forms a near continuous swathe of aquatic vegetation in the littoral zone along the coast, both to the north and south of the Roper River (Poiner et al., 1993). Although not studied seasonally or by community type due to the remoteness of the western Gulf of Carpentaria coastline, annual snapshots of the penaeid prawn community within the seagrass communities to the north and south of the Roper River were taken as part of a study of the impact of a category five cyclone (Cyclone Sandy, Gulf of Carpentaria 1985) on the seagrasses (Poiner et al., 1993). Endeavour prawns were found in the littoral seagrass habitats along the full extent of the coastline between the Sir Edward Pellew Islands and Groote Eylandt, including seagrass habitat within the Roper catchment marine region (Poiner et al., 1993). Though not reported, juvenile endeavour prawns were abundant in the same seagrass habitats in the Groote Eylandt and Blue Mud Bay coastal habitats as reported by Loneragan et al. (1998), at various densities depending on microhabitat (Rob Kenyon, CSIRO unpublished data). The Roper catchment marine region lies within the southern portion of the ‘Groote’ Northern Prawn Fishery statistical region of the Gulf of Carpentaria (adjacent to the Blue Mud Bay and Roper River coasts). This statistical region accounts for about 28% (approximately 118 t – 16-year average) of the total NPF endeavour prawn catch (both species grouped) (Laird, 2021). In the Roper catchment marine region, blue endeavour prawns dominate the species distribution on the fishing grounds to the south and south-west of Groote Eylandt (Kenyon, 2021). Figure 4-10 Fisheries catch of red endeavour prawns in the Roper catchment marine region Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022) and Kenyon (1999) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 4-11 Fisheries catch of blue endeavour prawns in the Roper catchment marine region Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022) and Kenyon (1999) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Flow–ecology relationships for endeavour prawns The life-history strategy of endeavour prawns incorporates an inshore dependence on littoral seagrass habitats as juveniles (Coles and Lee Long, 1985; Dall et al., 1990). Endeavour prawn postlarvae use currents to move shoreward to shallow, subtidal habitats, especially seagrass communities (Condie et al., 1999; Jackson and Rothlisberg, 1994). Prior to the annual wet season, blue endeavour prawn postlarvae 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). They shelter, forage and grow within vegetated habitat where leaf structure would reduce predation and promote primary productivity and prawn growth (Haywood et al., 1998; Kenyon et al., 1995). Seagrass habitats are found within some estuaries, though mostly adjacent to estuaries along the Gulf of Carpentaria coast. Seagrass thrives in oligotrophic waters, while a low-level nutrient 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 from the catchment to deposition within the flood plume and littoral zone adjacent to Gulf of Carpentaria rivers (Burford et al., 2012; Burford and Faggotter, 2021). Long-shore transport of a proportion of nutrients deposited adjacent to river mouths would benefit the productivity of coastal seagrasses in the Gulf of Carpentaria (Table 4-5). The hydrology within the inshore Gulf of Carpentaria may retain terrigenous-sourced nutrients in a less-well mixed layer of water inshore (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 4-5). 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 4-5. Table 4-5 Ecological functions for endeavour prawns and their supporting 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 endeavour prawns Coastal waters in Australia’s tropics such as the Gulf of Carpentaria and upstream catchment water bodies form oligotrophic though productive ecosystems, 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 would benefit from high-level flood flows in previous years supporting nutrient dynamics within the littoral community. The ecological outcomes of threatening processes on endeavour prawns in northern Australia are illustrated below, with their implications for changes to growth and mortality, community structure, habitat and population (Figure 4-12). Figure 4-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for endeavour prawns in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 4.2.3 Mud crabs (Genus: Scylla) Description and background to ecology Mud crabs are a large-bodied, large-clawed, short-lived, fast-growing decapod crustacean (>200 mm carapace width) that inhabit the estuarine and shallow subtidal community along tropical and subtropical coastlines, especially mangrove-dominated habitats (Figure 4-13). 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; 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. Mud crabs occupy mangrove forest (see Section 4.1.1) 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 ha-1 yr-1 in the mangrove forest and 2100 kg biomass ha-1 yr-1 in mangrove fringe habitat (Alberts- Hubatsch et al., 2016). Mud crabs dig burrows to rest during the day, reworking mud substrates within mangrove forests and mudbanks. 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 not only migration as crabs grow to the juvenile stage (ontogenetic migration) and settle to their inshore habitats, but 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, while 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). However, recent analyses of Gulf of Carpentaria catches support the notion of river flow enhancing catch, but also show high air temperature over the wet season as a dominant negative influence on mud crab abundance within the Roper River and southern Gulf of Carpentaria estuarine habitats (Robins et al., 2020). Figure 4-13 Mangrove and intertidal habitat typical of mud crab habitat in northern Australia Photo attribution: CSIRO Mud crabs in the Roper catchment and marine region From 2006 to 2018, the average harvest of mud crabs for the Roper catchment marine region was 71 t (an average 35% of the harvest from the Northern Territory Western Gulf of Carpentaria Mud Crab Fishery) (Robins et al., 2020). The Roper catchment marine region had a high variation in catch: a minimum catch of 3.3 t in 2016 and a maximum catch of 123.5 t in 2009 (Robins et al., 2020). Robins et al. (2020) conducted their recent comprehensive analysis of the effect of environmental drivers on Gulf of Carpentaria mud crab catches. Within the western and south- western regions of the Gulf of Carpentaria, heat, evaporation, precipitation, water stress and sea level, as well as hemisphere-wide phenomena create a high-stress environment for mud crabs (i.e. within the Roper and McArthur river estuaries). Catchment rainfall and floodwaters within estuaries support mud crab abundance as defined by commercial catch. Recorded observations of mud crabs in the Roper catchment are shown in Figure 4-14. Photo of mud crab habitat. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 4-14 Location of mud crab habitat in the Roper catchment marine region Mud crab juveniles use the mangrove forest/ mudbank habitats within the estuary and adult crabs are caught within the estuary and in shallow subtidal habitats in the littoral zone. Female mud crabs migrate offshore to spawn. Map. 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 with ontogeny. Mud crabs neither use freshwater riverine or palustrine habitats as juveniles, nor do they emigrate from their estuarine habitats to offshore adult habitats. Optimal estuarine conditions for mud crab growth and survival are a brackish ecotone, between marine habitats and the freshwater riverine habitats (Ruscoe et al., 2004). For mud crabs, no emigration cue exists as such, 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 Gulf of Carpentaria and other northern estuaries have been identified previously (Robins et al., 2005), many other environmental parameters are correlated with catch (Plagányi et al., 2022; Robins et al., 2020). Female mud crabs spawn offshore during 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 15–45 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 and a salinity of 10–20 ppt (growth) and 10–30 ppt (survival) (Meynecke and Lee, 2011; Meynecke et al., 2010; Ruscoe et al., 2004). For short periods, they can tolerate cool conditions (<20 °C), but require temperatures >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 4-6). Adult mud crabs are euryhaline animals, capable of living in freshwater-flooded to hypersaline waters (<5 to 45 ppt) (Alberts-Hubatsch et al., 2016). 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). As well, 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 impact inshore crab habitats in the year of the flood, though being beneficial in subsequent years due to medium-term productivity enhancement (Robins et al., 2020) (Table 4-6). 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), pers. comm.). In contrast, during lower-level floods survival of juvenile crabs in salinities of 5–40 ppt was high (optimally 15–25 ppt), and adult crabs were abundant in brackish estuaries (Robins et al., 2020). The ecological functions and their supporting flow requirements for mud crabs are summarised in Table 4-6. Table 4-6 Ecological functions for mud crabs 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 mud crabs Mud crabs exhibit a life history that would be significantly impacted 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 caused declines in mud crab catch. Analysis of environmental factors and commercial catch by Robins et al. (2020) showed that river flow and water stress (rainfall/evaporation, less stress if rainfall is high) had a positive effect on mud crab catch in the Roper catchment marine region, 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 SOI 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 Roper catchment 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 ecological outcomes of threatening processes on mud crabs in northern Australia are illustrated below, with their implications for changes to growth and mortality, community composition, habitat and population (Figure 4-15). Figure 4-15 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mud crabs in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.2.4 Tiger prawns (Penaeus esculentus and P. semisulcatus) Description and background to ecology Tiger prawns are a relatively large decapod crustacean (90–110 g) from the family Penaeidae, a family that exhibit 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 about 50-m depth. 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). 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. Tiger prawns support a 1749 t approximate component (recent 10-year average) of the commercial NPF (worth about $40–50 million annually) and 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 10–40 m deep, the larvae and postlarvae advect inshore 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; they provide Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 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 prawns do not bury; hence they depend on seagrass leaf structure to avoid predators, especially fish (Kenyon et al., 1995). The degree of refuge within seagrass habitat is dependent on seagrass species and morphology (Haywood et al., 1998; Kenyon et al., 1995; Kenyon et al., 1997). Juvenile tiger prawns emigrate offshore at about 10–20 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, 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 fragmented crustaceans 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 the majority of 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 = about 100 g) for females and about 47 mm CL (43 mm CL = about 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 = 55–65 g). Tiger prawns in the Roper catchment marine region Adult tiger prawns occupy relatively soft-sediment substrates in relatively shallow waters within the north-west and south-west Gulf of Carpentaria. The distributions of the two species is relatively allopatric, dependent 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 during the day and emerge to feed at night (when they are fished). Both grooved and brown tiger prawns are found in the western Gulf of Carpentaria in coastal waters ranging from about 10–45 m deep, offshore from the Roper River (Figure 4-16 and Figure 4-17). The juvenile phase of both species inhabit the seagrass community, which forms a near continuous swathe of aquatic vegetation in the littoral zone along the coast, both to the north and south of the Roper River (Poiner et al., 1987). Although not studied seasonally or by community type due to the remoteness of the western Gulf of Carpentaria coastline, annual snapshots of the penaeid prawn community within the seagrass communities to the north and south of the Roper River were taken as part of a study of the impact of a category five cyclone (Cyclone Sandy, 1985) on the seagrasses (Poiner et al., 1993). Both brown and grooved tiger prawns were found in the littoral seagrass habitats along the full extent of the coastline between the Sir Edward Pellew Islands and Groote Eylandt, including seagrass habitat within the Roper catchment marine region (Poiner et al., 1993). The Roper catchment marine region lies within the southern portion of the ‘Groote’ Northern Prawn Fishery statistical region of the Gulf of Carpentaria (adjacent to the Blue Mud Bay and Roper River coasts). This region accounts for about 29% (about 474 t – 16-year average) of the total NPF tiger prawn catch (both species grouped) (Laird, 2021). In the Roper catchment marine region, brown tiger prawns dominate the species distribution on the fishing grounds to the south and south-west of Groote Eylandt (Kenyon, 2021). Figure 4-16 Fisheries catch of brown tiger prawns in the Roper catchment marine region Units are kilograms as total catches for the 10-year period (2011 to 2020). Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Data source: Kenyon et al. (2022) and Kenyon (1999) Figure 4-17 Fisheries catch of grooved tiger prawns in the Roper catchment marine region Units are kilograms as total catches for the 10-year period (2011 to 2020). Data sources: Kenyon et al. (2022) and Kenyon (1999) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Flow–ecology relationships for tiger prawns The life-history strategy of tiger prawns incorporates an inshore dependence on littoral seagrass habitats as juveniles (Dall et al., 1990; Loneragan et al., 1994). Seagrass habitats are found both within some estuaries, though mostly adjacent to estuaries along the Gulf of Carpentaria coast. Tiger prawn postlarvae use currents to move shoreward to shallow, subtidal habitats, especially seagrass communities (Vance et al., 1996b). Prior to the annual wet season, postlarvae 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 low-level nutrient environment supports seagrass growth. Floodwaters transport terrigenous nutrients from the catchment to deposition 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, carbon-productivity generated by wetted algal crusts are deposited within the plume (Burford et al., 2016). Long-shore transport of a proportion of nutrients deposited adjacent to river mouths would benefit the productivity of coastal seagrasses in the Gulf of Carpentaria. The hydrology within the inshore Gulf of Carpentaria may retain terrigenous-sourced nutrients in a less-well 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 40 years of flow-catch relationship as exists for banana prawns, when wet-season rainfall is high and salinity declines in coastal estuarine or embayment seagrass habitats, 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 4-7. Table 4-7 Ecological functions for tiger prawns and their supporting 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 tiger prawns Coastal waters in Australia’s tropics such as the Gulf of Carpentaria and upstream catchment water bodies form oligotrophic though productive ecosystems, 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). Once an abundant population of juvenile tiger prawns is established within the coastal seagrass community, they benefit from high-level flood flows in previous years supporting 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 are illustrated below, with their implications for changes to growth and mortality, community composition, habitat and population (Figure 4-18). Figure 4-18 Conceptual model showing the relationship between threats, drivers, effects and outcomes for tiger prawns in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.3 Marine fish and shark species and species groups 4.3.1 Mullet (Family: Mugilidae) Description and background to ecology Mullet (a guild including the genera Liza, Mugil and Valamugil) are fish that use marine habitats as adults to spawn and freshwater habitats as juveniles (i.e. catadromous). They have life histories that entail ‘catchment to coast’ habitats (i.e. freshwater, estuarine and marine habitats) (Marin et al., 2003; Whitfield et al., 2012). Mullet are distributed in tropical and temperate coastal waters worldwide. 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 subviridus), sea mullet (Mugil cephalus), roundhead mullet (Valamugil cunnesius), bluespot mullet (Valamugil seheli) and bluetail mullet (Valamugil buchanani) are common species in the Australian tropics and range across the Indo-Pacific (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 and small invertebrates (Górski et al., 2015; Soyinka, Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 2008; Whitfield et al., 2012). Being themselves preyed upon by larger species, they 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, fast growth rates and high fecundity makes 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, 1975; Whitfield et al., 2012). By about 4 years, 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, 1975; Kailola et al., 1993; Robins et al., 2005). Short-lived, fast-growing, and productive, mullet are important as a commercial, recreational, and Indigenous fish resource. Mullet are one of the most important species taken in NT recreational catches and the third most prominent species in (non-Indigenous) recreational catches in the east coast / Gulf of Carpentaria area of the NT (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 among the most numerous species 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 (Northern Territory Government, 2022c). Mullet species are minor bycatch in the NPF (Pender et al., 1993). Mullet in the Roper catchment and marine region Due to their remoteness, river systems in the western Gulf of Carpentaria are poorly studied, and fishing effort by both commercial and recreational sectors can be relatively limited. Nevertheless, bait fishing is a source of commercial mullet catches in the western Gulf of Carpentaria, including the Roper River. Several NT commercial fisheries, including the Fishing Tour Operator sector, are permitted to use the Roper River estuaries and coastal area for bait collection (Northern Territory Government, 2022c). Although mullet are the primary species taken in the Coastal Net Fishery, the Roper River area is not within the zones in which this fishery operates (Northern Territory Government, 2022b). Northern Territory Aboriginal coastal fishing licences are designed to provide for an increased supply of fresh seafood in communities, as well as business opportunities (Northern Territory Government, 2022a) These enterprises have been successfully developed in the Maninigrida area (Boyer, 2018; Wilton et al., 2018). Although several mullet species are caught in the Roper catchment, diamond-scale mullet and bluetail mullet are the most important mullet species for the Indigenous communities of the Roper River area (Mr Steve Barret (Limmen Bight Fishing Camp), 2022, pers. comm.). The Yanyuwa Traditional Owners of the Roper River to Robinson River areas have a vision of development in supportive fishing industries, including Aboriginal coastal fishing licences, to grow local economies (Sinclair et al., 2020). In 2009–10, estimated non-Indigenous recreational fishing effort directed at mullet in the east coast / Gulf of Carpentaria area, which includes the Roper catchment study area, was 10,313 (SE = 1733) fisher days (Appendix 12, West et al., 2012). This produced an estimated annual harvest of 5010 (SE = 3297) fish (Appendix 13, West et al., 2012). For the Roper catchment study area specifically, April-November 2009, the estimated number of mullet kept by NT residents and visitors in the Roper River accommodation survey was 580 (SE = 420) fish (Appendix 29, West et al., 2012). The dominance of non-Northern Territory residents (64.9% of those interviewed) in the survey reflects the importance of fishing tourism in the area. Known locations of mullet are shown in Figure 4-19. Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 4-19 Records of capture of mullet in the Roper catchment 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 (2021a) 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 prolifically 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 impact mullet negatively via reduction in 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 4-8). 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 4-8). 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 all three study areas, 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 and their supporting flow requirements for mullet are summarised in Table 4-8. Table 4-8 Ecological functions for mullet and their supporting 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 mullet With marine and freshwater habitat use similar to barramundi (Section 3.2.1), juvenile and early- adult phase mullet prefer fresh and brackish waters, including palustrine wetlands, supporting 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 impact mullet populations through reduction in the extent and connectivity of the estuarine and freshwater habitats (Faggotter et al., 2013; Jardine et al., 2013; O'Mara et al., 2021), as well as disrupting cues for spawning movements. Disrupted connectivity by built-barriers may limit use of freshwater habitats (Grant and Spain, 1975; 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). Some individuals, nevertheless, 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 are illustrated below, with their implications for changes to growth and mortality, community structure, habitat and population (Figure 4-20). Figure 4-20 Conceptual model showing the relationship between threats, drivers, effects and outcomes for mullet in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 4.3.2 Sawfish (Genus: Pristis) Description and background to ecology Sawfish are a type of ray from the Pristiformes, with a characteristic tooth-lined rostrum (or ‘saw’). As adults, sawfish can attain very large sizes in the order of five to seven meters total length. 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, where inshore waters, including bays and estuaries, are important nursery grounds for neonates and juvenile sawfishes, up until about 4 to 6 years of age (Morgan, 2011; Morgan et al., 2017; Peverell, 2005). Then as adults, they primarily live in tropical and subtropical coastal marine waters (Dulvy, 2016; Last and Stevens, 2008). Worldwide, sawfish are among the most threatened marine taxa (Dulvy, 2016). In Australian waters there are four species of sawfish, all listed of conservation significance at national and international level. The freshwater or large tooth sawfish (Pristis pristis), the green sawfish (P. zijsron) and the dwarf sawfish (P. clavata) are listed as Vulnerable under the EPBC Act. The narrow sawfish (Anoxypristis cuspidata) is listed as Migratory but because it is also listed on Appendix I and II on the Conservation of Migratory Species, it has similar protection status under the EPBC Act. Additionally, sawfishes have a significant cultural and spiritual relevance to Indigenous Australians (Ebner et al., 2016). Indigenous communities on Groote Eylandt, an island located Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. approximately 150 km from the Roper River mouth, believe an ancestral sawfish was the creator of the Angurugu River (McDavitt, 1996). Sawfish are susceptible to multiple threats, in part due to their morphology (the shape of their rostra) and behaviour, and in part due to their life-history characteristics: long lived, slow growth and low reproductive rates, late maturation, relatively low abundance and high specificity of different habitats during their life stages (Peverell, 2005; Phillips, 2017a; Stevens et al., 2009). Due to the overlap of sawfish habitats with coastal fisheries and because the shape of their rostra they are highly disposed to capture in gill-net and trawl fisheries and in recreational fishing. Sawfish rostrums have been collected as trophy items for decades (McDavitt, 1996), and more recently, there is an increasing demand for live sawfish for display in public aquaria (Buckley et al., 2020; Compagno et al., 2006). Fishing mortality over recent decades has been high (Fry, 2021). 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. Sawfish in the Roper catchment and marine region The Roper catchment and marine region represents a potentially important stronghold for the remnant populations of these threatened species (Morgan et al., 2015; Peverell, 2005; Phillips, 2017a; 2017b). Narrow sawfish were caught in the Roper catchment marine region south of Groote Eylandt during both the banana prawn and tiger prawn fishing seasons (Fry, 2021). During the tiger prawn season, the species was caught about 15 nautical miles offshore from the Roper River estuary. From 2003 to 2019 during the tiger prawn fishing season, green sawfish were caught in the Roper catchment marine region. From 2003 to 2019, no largetooth or dwarf sawfish were caught offshore in the Roper catchment marine region (Fry, 2021). However, these catches were taken in relatively deep waters able to be accessed by ~25 m trawler vessels and outside inshore spatial closures. Atlas of Living Australia data show a largetooth sawfish record from the Roper River estuary and two records from freshwater reaches approximately 200 km upstream (Figure 4-21). In addition and not mapped, in 1 year across the extent of the NT barramundi inshore fishery, nine largetooth sawfish, 26 green sawfish, 24 dwarf sawfish and one narrow sawfish were reported caught in littoral habitats (Northern Territory Government, 2018), reinforcing the presence of the four sawfish species in estuarine and shallow inshore habitats in this catchment. Recorded observations of sawfish are shown in Figure 4-21. Figure 4-21 Records of sawfish capture in the Roper catchment and the marine region Data sources: Fry (2021); Kenyon et al. (2022) Atlas of Living Australia (2021a) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Flow–ecology relationships for sawfish Sawfish species have known dependencies on estuarine and riverine habitats. Juvenile largetooth sawfish inhabit both estuarine and freshwater environments and they move 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, they remain in refuge pools (see Section 3.1.2) during the dry season in the Australian tropics (up to about 5 years old). At maturity, they migrate downstream to estuarine habitats and are vulnerable to inshore gill-net fisheries, particularly in the monsoonal wet season (February to April) (Peverell, 2005). Riverine and estuarine connectivity, as well as long-stream connectivity are critical to enable largetooth sawfish to access their juvenile habitats and to return to estuarine breeding habitats (Table 4-9). Dwarf sawfish use estuarine habitats and the lowermost riverine reaches seasonally, cued by 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). They move in response to freshwater cues, seeking higher salinity waters. Also, the green sawfish exhibits site fidelity within the estuarine and coastal habitat matrix in the vicinity of the mouth of tropical rivers. They move to shallow coastal habitats at low tide, and mangrove creek habitats at high tide (Morgan et al., 2017). During large river flow discharge events they emigrate from the river estuary to coastal habitats. The ecological functions and their supporting flow requirements for sawfish are summarised in Table 4-9. Table 4-9 Ecological functions for sawfish 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 sawfish The potential implications of threatening processes for sawfish in northern Australia are summarised in Figure 4-22. Changes in the depth, extent, duration and timing of flows in river reaches where the sawfish inhabit can result in habitat loss and impact significantly on the sawfish populations, for example, by reducing neonate recruitment (Morgan et al., 2016), affecting the potential growth of individuals (Hunt et al., 2012), reducing the abundance of prey species that use floodplain wetlands during their life cycle (Novak et al., 2017), and/or reducing abundance and survivorship (Close et al., 2014; Jellyman et al., 2016; Morgan et al., 2016). Critical flow characteristics in Australia’s tropical rivers that support sawfish populations have been identified by recent research in the Fitzroy River, WA. Recruitment and survival of largetooth sawfish within riverine freshwater habitats was critically dependent on large flood flows. Largetooth sawfish recruitment to riverine habitats was dependent on extended periods of high- level flows to support access to the upstream freshwater river reaches which are their juvenile habitats (14 or more consecutive days in the 98th percentile of recorded water levels) (Lear et al., 2019b). Importantly, remnant riverine pools are critical refugia for sawfish during the dry season. Restricted to the pools, they lose body condition over the dry season; and following low-volume wet-season flows, the loss of body condition is greater than following high-volume wet-season flows (Lear et al., 2021). Moreover, 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 the health of sawfish and the disruption to natural flows due to water impoundment of extraction has the capacity to impact their survival (Figure 4-22) (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. Fishing mortality over recent decades has been high (Fry, 2021). 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 ecological outcomes of threatening processes on sawfish in large rivers in northern Australia are illustrated below, with their implications for changes to growth, population and community structure (Figure 4-22). Figure 4-22 Conceptual model showing the relationship between threats, drivers, effects and outcomes for 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, while black arrows represent non-hydrological changes. 4.3.3 Threadfin (Polydactylus macrochir) Description and background to ecology King threadfin (Polydactylus macrochir, formerly P. sheridani) are a large (>1.5 m) non- diadromous, carnivorous fish (Order Perciformes). Endemic to Australasia, they range 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 are long- lived (22 years) and fast-growing, and begin life as males but change to females as they age (protandrous hermaphrodites). Their body form and quality of flesh makes them a prized table fish. King threadfin is typically the second-most important target species in the commercial, inshore gill-net fisheries that principally target barramundi (Welch et al., 2010). In 2018–19, 235 t of king and blue threadfin 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). They are also a 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 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). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. King threadfin complete their entire life cycle in turbid coastal waters, in estuaries, mangrove creeks and inshore marine waters, but while tolerating a wide salinity range (as low as 2‰), they 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‰) 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. They may also benefit from turbid waters during wet-season flows, protecting young threadfin from large predators (Welch et al., 2014). They particularly inhabit the mid- to upper-estuary, but 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. They 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 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 typify an estuary-dependent fish that hunt 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 impacted 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 Roper catchment marine region There is little knowledge of the stock structure of king threadfin in waters of the NT, although Welch et al. (2010) detected two different stocks in the NT based on the two locations sampled (i.e. Chambers Bay, near Darwin, and Blue Mud Bay). Finer-scale sampling conducted in Queensland 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 4-23). Difficult access to the remote Gulf of Carpentaria coast has also meant that fishing effort by both commercial and recreational sectors can be relatively limited. Historically, commercial fishing occurred from a self- contained, refrigerated fishing vessel and several smaller dories deploying nets or lines. Annual commercial landings of king threadfin in the Roper River in the period 1990 to 2009 were 3–27 t (value less than ~$100,000) (Halliday et al., 2012). Catch estimates for the subsequent years were not available but were similar to this range (Thor Saunders (Department of Industry, Tourism and Trade, Northern Territory), 2022, pers. comm.). King threadfin are important to the Indigenous communities within the catchment and are an important recreational target species for the broader area, including the townships of Katherine, Tindal and Mataranka. Fishing tourism was relatively important with angler effort in April to November 2009 estimated as 5561 days (SE = 1241); this comprised 93.8% non-Northern Territory residents (West et al., 2012). The western Gulf of Carpentaria produced just 5% of the recreational catch of king threadfins (West et al., 2012). Figure 4-23 Records of threadfin capture in the Roper catchment 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. Map. 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 for 1990–2009 was significantly, positively correlated with spring rain lagged by 3 years and was significantly but negatively correlated with autumn rain in the year of catch. King threadfin do not use freshwater habitats, hence 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 4-10). 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 an advantage to fish that do not rely on visual predation alone, such as threadfin. The ecological functions and their supporting flow requirements for threadfin are summarised in Table 4-10. Table 4-10 Ecological functions for threadfin 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 threadfin During a 5-year study, analysis of 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 in a large Queensland subtropical estuary (Halliday et al., 2008). In the Gulf of Carpentaria and the Daly River (NT), 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, 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 on 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 zone 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). The ecological outcomes of threatening processes on threadfin in northern Australia are illustrated below, with their implications for changes to growth and mortality, community structure, habitat and population (Figure 4-24). Figure 4-24 Conceptual model showing the relationship between threats, drivers, effects and outcomes for threadfin in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. 5 Ecological assets from terrestrial systems The terrestrial habitats of northern Australia include a range of varied 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 groundwater dependent ecosystems (GDEs) including aquatic, terrestrial and subterranean habitats. Many of these are highly dependent upon freshwater from rivers and can also be supported by groundwater discharges for their persistence and condition. This section provides a synthesis of the prioritised assets relevant to terrestrial sections of the Assessment catchment. Table 1-1 presents the full list of assets used in the ecology assessment. 5.1 Terrestrial habitats 5.1.1 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 in areas where it would not otherwise persist such as the red cabbage palm (Livistonia mariae; Box et al. (2008)), 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. Aquatic groundwater-dependent ecosystems Aquatic GDEs are surface water habitats that require the presence of near-surface groundwater (e.g. for hyporheic exchange) or groundwater discharge to the surface. These 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., 2009b). 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. Kath et al. (2014) threshold responses). They 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, Figure 5-1) 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 There is a common misconception that 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 following terms are defined: 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 5-1 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 5-1 Conceptualisation of obligate and facultative groundwater-dependent vegetation Source: Pritchard et al. (2010) Figure 5-2 depicts natural terrestrial vegetation vigour associated with changes in water availability. During the wet season (i) vegetation has access to sufficient water resulting in high productivity and diversity. During the dry season (ii) there is reduced water availability and productivity. As soils dry, annual vegetation species die back while deeper rooted species maintain greenness 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 it more susceptible to invasive plants. Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-2 Conceptualisation of terrestrial GDEs Note: (i) represents vigorous ecosystems with seasonally high water availability, (ii) represents ecosystem condition with seasonally low water availability, and (iii) represents 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 and stygofauna; Richardson et al. (2011a)). Subterranean fauna has 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 is 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 potentially provide ecosystem services, such as nutrient cycling and water purification (Glanville et al., 2016; Smith et al., 2016). GDEs in the Roper catchment The National GDE Atlas (Bureau of Meteorology, 2017) contains maps of the distribution of known and potential groundwater-dependent inland aquatic and terrestrial ecosystems. 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 GIS (Doody et al., 2017). There is little known about coastal or submarine groundwater discharge along the northern coast of Australia. The Roper River estuary is considered to have a strong groundwater influence (Sinclair Knight Merz, 2012). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. In Australia, the biodiversity and distribution of subterranean ecosystems remains largely unknown. There are three types of aquifers known to provide subterranean ecosystems that can support stygofauna: karstic, fractured rock and alluvial, typically where the depth to groundwater is less than 30 m (Doody et al., 2019). Karstic aquifers are a prominent feature in the Roper catchment and shallow alluvial aquifers associated with river channels also occur (see section on Karstic aquifers below). Aquatic GDEs in the Roper catchment Regional studies have shown that the Roper River and many of its tributaries contain springs and river sections that hold flowing water throughout most dry seasons due to groundwater discharge, and these support aquatic life and fringing vegetation (Bureau of Meteorology, 2017). These springs include Mataranka Springs thermal pools in the Roper catchment (Figure 5-3). These aquatic ecosystems are mapped as ‘Known GDEs’ in the GDE Atlas (Figure 5-4). There are also hundreds of river sections, lakes 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 5-4). Subsequent environmental tracer studies at Mataranka Springs (Lamontagne et al., 2021) have shown that Rainbow Spring is fed by groundwater originating from the Daly (southerly) flow path within the Cambrian Limestone Aquifer and Bitter Spring, Warloch Pond and Fig Tree Spring are fed by groundwater from the Georgina (northerly) flow path within the Cambrian Limestone Aquifer. However, further work is required to attribute sources of groundwater to most other aquatic GDEs in the catchment and whether they are connected to groundwater flow systems that are of interest for water resource development (Taylor et al., 2023). Figure 5-3 Bitter Springs thermal pools in Elsey National Park Photo attribution: CSIRO Photo of Bitter Springs. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. There are 132 springs mapped in the Roper catchment (Figure 5-5), but only a handful of these are currently incorporated in the GDE Atlas. It is assumed that because these surface water features are named springs, they are groundwater discharge features and should be considered aquatic GDEs. Sinkholes associated with the Dook Creek Formation, Tindall Limestone, Gum Ridge Formation and Jinduckin Formation have been mapped (Figure 5-5). It is unconfirmed whether any of these contain groundwater and support aquatic ecosystems through the dry season. However, it is assumed here that sinkholes that are classified as broad depressions associated with the Tindall Limestone near Mataranka, are recharge features that drain into the aquifer and do not contain water during the dry season. It is also assumed here that sinkholes classified as caves potentially provide habitat for subterranean fauna that depend on the presence of groundwater (Subterranean GDEs). Such caves are mapped in the karstic Dook Creek Formation in the north and Tindall Limestone in the south-west. There is little known about coastal or submarine groundwater discharge in the Roper catchment. Stories shared by local Indigenous people report that the coastal area had very poor fresh surface water resources, and they had to travel many days carrying water through the coastal area of the Roper catchment (Zaar, 2009). However, the Roper River estuary is considered to have a strong groundwater influence (Sinclair Knight Merz, 2012). Global-scale modelling suggests that there is potential for submarine groundwater discharge off the coast of the Roper River (Luijendijk et al., 2020), but there have been no local-scale studies to substantiate this. Figure 5-4 Distribution of known and potential groundwater-dependent aquatic ecosystems in the Roper catchment Note: A buffer of 1 km has been applied to GDE mapping so that they are visible on the map scale. Dataset: Bureau of Meteorology (2017) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-5 Locations of springs and sinkholes in the Roper catchment Dataset: Department of Environment‚ Parks and Water Security (NT) (2013; 2014) Terrestrial GDEs in the Roper catchment Regional studies have shown that terrestrial vegetation adjacent to many of the waterways in the Roper catchment use groundwater. These terrestrial ecosystems are mapped as ‘Known GDEs’ in the GDE Atlas (Bureau of Meteorology (2017); Figure 5-6). Groundwater-dependent terrestrial ecosystems are known to exist in four key areas including the Limmen Bight (Port Roper) Tidal Wetlands System (Melaleuca cajuputi, Melaleuca viridiflora open forest), surrounding Red Lilly Lagoon (M. viridiflora, M. cajuputi low open-woodland), fringing Mataranka Hot Springs (M. cajuputi, Melaleuca acaciodes open forest) and riparian vegetation along many of the rivers and creeks (Melaleuca leucadendra, Melaleuca argentea open forest and M. leucadendra, M. viridiflora woodland). There is also vegetation adjacent to waterways in the south-west and the northern catchment that are believed to be groundwater dependent based on remote sensing work and expert opinion and are mapped as ‘Potential GDEs’ (Bureau of Meteorology (2017); Figure 5-6). Terrestrial GDEs mapped in the GDE Atlas within the Roper catchment include monsoon vine forests and Melaleuca forests and woodlands (including M. cajuputi, M. acaciodes, Melaleuca citrolens, Melaleuca minutifolia, M. viridiflora, M. argentea, M. leucadendra, Melaleuca nervosa), however, there are a number of other terrestrial vegetation species that exist in the Roper catchment that are likely to use groundwater (e.g. river red gum; Eucalyptus camaldulensis). A preliminary indication of where other potential terrestrial GDE species exist within the Roper catchment is shown in Figure 5-7 and Figure 5-8. Figure 5-7 maps observed occurrence of three tree species (E. camandulensis, M. argentea, Barringtonia acutangular) that are 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 are riparian, along the major rivers, however comprehensive species distribution mapping does not exist and B. acutangular is also known to occur in freshwater mangroves and M. argentea is also known to occur in floodplain and swamp habitats. Therefore, the distribution of obligate GDE vegetation species is expected to be more extensive than mapped in Figure 5-7. Figure 5-8 shows the observed occurrence of many other known GDE species grouped by vegetation type ‘GDEs’ that have been identified 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 are suspected to use groundwater, but groundwater use remains unconfirmed. A complete list of species included in Figure 5-8 is provided Appendix A. 150 | Ecological assets of northern Australia Figure 5-6 Distribution of potential groundwater-dependent terrestrial ecosystems in the Roper catchment Note: A buffer of 1 km has been applied to GDE mapping so that they are visible on the map scale. Dataset: Bureau of Meteorology (2017) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-7 Locations of observed obligate terrestrial GDEs in the Roper catchment Note: A buffer of 1 km has been applied to NT Melaleuca mapping so that they are visible on the map scale. Datasets: Atlas of Living Australia (2021a); Department of Environment Parks and Water Security (2000a) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-8 Locations of facultative and potential GDE vegetation species in the Roper catchment grouped by relevant vegetation type Note: A buffer of 1 km has been applied to NT Melaleuca and Monsoon vine forest mapping so that they are visible on the map scale. Datasets: Atlas of Living Australia (2021a); Department of Environment Parks and Water Security (2000a; 2000b) Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Subterranean GDEs in the Roper catchment Subterranean aquatic ecosystem sampling was limited to 16 boreholes within the Roper catchment, however it revealed that a diverse stygofauna community exists within the Tindall Limestone karstic aquifer (Cambrian Limestone Aquifer; Oberprieler et al. (2021)). Some of the subterranean species are distributed across a broad spatial range (~300 km), but smaller species may exist that are less mobile and are at risk if aquifer conditions change due to their short-range endemism (Oberprieler et al., 2021). Figure 5-9 shows where stygofauna have been tested and detected in the Roper catchment and based on the presence of favourable habitats for subterranean GDE species (caves and alluvial and karstic aquifers), where additional subterranean GDEs may also be found. Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-9 Locations tested for the presence of subterranean GDEs in the Roper catchment and locations of caves and alluvial and karstic aquifers that may provide habitat for subterranean GDEs Datasets: Subterranean GDEs Rees et al. (2020); Department of Environment‚ Parks and Water Security (NT) (2014). 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/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 3.1.1 and 3.1.2 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 (Table 5-1). However, there is some crossover between groundwater-dependent and surface water dependent terrestrial vegetation (Section 5.1.2), for which flow–ecology relationships are reported in Table 5-2. 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 5-1 specifies broad flow–ecology relationships that need to be considered when assessing the impact of changes in flow on subterranean GDEs. Table 5-1 Ecological functions for GDEs and their supporting 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. Here we discuss 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 5-10), terrestrial (Figure 5-11) and subterranean GDEs (Figure 5-12). 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 there is reduced groundwater recharge, the impacts on GDEs over the long term are similar to groundwater drawdown. In areas where there is enhanced groundwater recharge, there could be: (i) 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 (potentially favouring non-native aquatic species, e.g. (Yarnell et al., 2015)); (ii) shallower groundwater levels, potentially leading to soil salinisation due to evapotranspiration from shallow watertables (e.g. Smith and Price, 2009) and/or may cause a shift in the type of terrestrial vegetation supported (e.g. from Melaleuca swamp to grassland, (Department of Environment and Science Queensland, 2013); and (ii) 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 impact the volume, timing, frequency and quality of surface water flows or inundation are likely to impact aquatic GDEs and fringing vegetation. The ecological outcomes of threatening processes on GDEs in northern Australia are illustrated below with their implications for changes to biodiversity and ecosystem function (Figure 5-10, Figure 5-11, Figure 5-12). Conceptual model. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-10 Conceptual model showing the relationship between threats, drivers, effects and outcomes for aquatic GDEs in northern Australia Blue arrows represent hydrological changes, while 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 5-11 Conceptual model showing the relationship between threats, drivers, effects and outcomes for terrestrial GDEs in northern Australia Blue arrows represent hydrological changes, while 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 5-12 Conceptual model showing the relationship between threats, drivers, effects and outcomes for subterranean GDEs in northern Australia Blue arrows represent hydrological changes, while black arrows represent non-hydrological changes. 5.1.2 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, 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. Critically endangered Zyzomys palatalis, unique to monsoon forest, Crowley (2010)) and provide critical habitat for fauna, for example 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 high and low flows) • groundwater depth • timing and flood frequency (return interval) • 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. They have some inbuilt resilience to natural changes in water availability, but prolonged change is likely to result in dieback after some lag period and 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, the water regimes that support three terrestrial vegetation types are reviewed here: • paperbark swamps • river red gum • monsoon vine forest. In northern Australia, these ecosystems provide food and habitat for high levels of biodiversity (e.g. for migratory waterbirds, flying foxes, crocodiles and honeyeaters), playing 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’ is considered 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 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 floodprone 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 moisture 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. There is not much 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 abundance of saplings, most often in recently burnt areas. 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 ~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. (2009a); 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 ranging from every 1 to 3 years for 2 to 8 months (Rogers and Ralph, 2010) to every 3 to 5 years for up to 2 months (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: These flooding relationships exist for trees found in the MDB where there has been extensive research completed on maintaining this ecosystem type. However, these relationships cannot be directly extrapolated to the different hydrologic-soil- climate conditions of northern Australia. Specific water requirements for E. camaldulensis and subspecies found in northern Australia are unknown. Monsoon vine forest The 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 600–2000 mm mean annual rainfall (Bowman and Bowman, 2000). Most monsoon vine forests seem limited to areas with permanent soil moisture such as creek lines, springs and seeps and are thought to be remnants of a wetter period during Australia’s geological history where changes in climate, fire regime and water availability has restricted their distribution to small pockets across northern Australia of less than several hectares (Bowman and Bowman, 2000). However, the hydrological and geomorphic environments of these ecosystem communities are poorly understood, and while the monsoon forest can typically be found in areas that offer fire protection such as boulder outcrops and areas of high soil moisture, a change in water availability may make the 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 thickets or a 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 will therefore be used throughout this report. Furthermore, Russell‐Smith (1991) categorised monsoon vine forests into 16 different floristic assemblages or rainforest types; he defined these by what area they grew (coastal vs inland), water regime (wet vs dry), rainforest type (forest vs vine thicket) and canopy type / height. This report uses the 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’ and having near constant waterlogging of soils with very little soil drying out, and ‘dry’, which may occur on floodplains or are seasonally flooded with regular drying out of soils. See Appendix B 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 are considered Endangered under the Act (Fisher et al., 2014). Surface water dependent vegetation in the Roper catchment The distribution of red gum is not comprehensively mapped, but available data indicate that it occurs along the banks of major rivers in the Roper catchment (based on ALA data, Atlas of Living Australia (2021a); Figure 5-13). In the Roper catchment, paperbark is found in five key areas including the floodplains around the Limmen Bight Tidal Wetland system (Melaleuca cajuputi, M. viridiflora, M. leucadendra), in the topographic depression around Mataranka Springs (M. cajuputi, M. acacioides), on the floodplains and drainage features around Red Lilly Lagoon (M. viridiflora, M. cajuputi, M. dealbata), and adjacent most of the major rivers and creek lines and the plains and drainage areas surrounding creek lines particularly in the southern area of the catchment (M. citrolens, M. minutifolia, M. nervosa, M. viridiflora, M. dealbata). Whilst Figure 5-13 shows known observed occurrences of paperbark species in the Roper catchment, it is unconfirmed whether all of these occurrences actually fall within swamp habitats. ‘Wet’ monsoon forest species diagnostic of springs, seasonal flooding or groundwater use (obligate GDEs) occur adjacent to river channels predominantly in the northern reaches of the Roper catchment, with those species diagnostic of seasonal flooding extending further south (based on ALA data, Atlas of Living Australia (2021a)). Significant areas of monsoon forest diagnostic of springs and groundwater use occur along much of the Roper River. Map. For more information on this figure, table or equation please contact CSIRO on enquiries@csiro.au. Figure 5-13 Locations of observed selected surface water dependent vegetation types in the Roper catchment Species within each vegetation type are listed in Appendix B Datasets: Atlas of Living Australia (2021a); Department of Environment Parks and Water Security (2000a) Flow–ecology relationships for surface water dependent vegetation Red gum, paperbark and ‘wet’ monsoon forest vegetation are sensitive to changes in water availability because they need more water than is available from local rainfall alone to sustain them. Some require periodic inundation by floodwaters and/or access to groundwater to survive, flower, fruit and/or reproduce, as summarised in Table 5-2. The amount, source, timing and frequency of extra water needed by vegetation will vary depending on climate, local soils and vegetation type. The water needs for all vegetation types are not well defined, particularly in northern Australia. Table 5-2 Ecological functions for surface water dependent vegetation and their supporting flow requirements 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 The changes of water availability to each of these vegetation types (both groundwater and surface water dependent) could affect ecosystem function and how they may persist 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 a groundwater source of water (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 affected the surface water and groundwater availability for a wet monsoon forest; this in turn 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). The 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 they 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, 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/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 is 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 velocity of waterways seem to influence what ecosystem types are present in northern Australia (Figure 5-14), if a location is waterlogged / spring fed, with 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, if the same environmental conditions are present with high levels of disturbances (as mentioned above), then this may support certain types of paperbark swamps, or the ecosystem may change as the result of this disturbance (Franklin et al., 2007). Indeed, if a paperbark swamp has a high frequency of continual inundation, with disturbances, then the environment could shift to a grassland (Department of Environment and Science Queensland, 2013); however, Bren (1992) have also shown that flooded grasslands are at risk of encroaching river red gum forests (E. camaldulensis) 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 5-14). 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Zhou S, Buckworth RC, Ellis N, Deng RA and Pascoe S (2015) Getting all information out of logbooks: estimating banana prawn fishable biomass, catchability, and fishing power increase, with a focus on natural mortality. ICES journal of marine science 72(1), 54-61. 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 Terrestrial GDE observations in the catchments Roper catchment terrestrial GDE observations in ALA Apx Table A-1 List of GDEs observed in the Roper catchment Notes: This is based on a search of the literature and species mapped in ALA (Atlas of Living Australia, 2021a) and is not a fully comprehensive list. Any subspecies of these varieties present in the ALA database are included in the mapping (Figure 5-8). OBLIGATE GDE FACULTATIVE GDE OR TYPE OF DEPENDENCY UNCONFIRMED POTENTIAL GDE Riparian Eucalyptus camaldulensis  Melaleuca argentea  Acacia auriculiformis  Cathormion umbellatum  Corymbia bella  Nauclea orientalis  Lophostemon lactifluus  Pandanus spiralis  Leptocarpus tenax  Typha orientalis  Pandanus aquaticus  Carallia brachiata  Paperbark swamp Melaleuca leucadendra  Melaleuca viridiflora  Melaleuca dealbata  Melaleuca alsophila  Melaleuca citrolens  Melaleuca cajuputi  Monsoon vine forest Tylophora cinerascens  Atalaya variifolia  Bauhinia cunninghamii or Lysiphyllum cunninghamii  Celtis philippensis  Clerodendrum floribundum var. ovatum  OBLIGATE GDE FACULTATIVE GDE OR TYPE OF DEPENDENCY UNCONFIRMED POTENTIAL GDE Diospyros humilis  Exocarpos latifolius  Grewia breviflora  Gyrocarpus americanus subsp. pachyphyllus  Sersalisia sericea  Terminalia ferdinandiana  Syzygium nervosum  Carpentaria acuminata  Croton habrophyllus  Dodonaea platyptera  Flueggea virosa subsp. melanthesoides  Hypoestes floribunda var. varia  Planchonia careya  Flagellaria indica  Abrus precatorius  Capparis lasiantha  Jasminum didymum  Operculina aequisepala  Opilia amentacea  Tinospora smilacina  Acacia aulacocarpa  Antidesma parvifolium  Calophyllum sil  Canarium australianum  Denhamia obscura  Fagraea racemose  Ficus benjamina  Ficus coronulata  Ficus racemose  Ficus virens  Glochidion perakense  Gmelina schlechteri  Helicia australasica  Homalanthus novo- guineensis  Ilex arnhemensis  OBLIGATE GDE FACULTATIVE GDE OR TYPE OF DEPENDENCY UNCONFIRMED POTENTIAL GDE Lindsaea ensifolia  Litsea glutinosa  Lygodium flexuosum  Lygodium microphyllum  Macaranga tanarius  Maranthus corymbose  Melastoma affine  Nephrolepis biserrate  Planchonella DNA 47005  Schefflera actinophylla  Sterculia holtzei  Syzygium angophoroides  Syzygium forte  Syzygium minutuliflorum  Terminalia microcarpa  Vitex glabrata  Xanthostemon eucalyptoides  Other habitats Barringtonia acutangular  Eucalyptus miniata  Eucalyptus tetrodonta  Melaleuca nervosa  Corymbia opaca  Atalaya hemiglauca  Cyperus conicus  Surface water dependent vegetation observations in the Roper catchment Apx Table B-1 Red gum species (including subspecies) observed in northern Australia based on ALA (Atlas of Living Australia, 2021a; 2021b; 2021c) data within Roper catchment (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 ROPER Eucalyptus camaldulensis   Eucalyptus camaldulensis subsp. acuta P  Eucalyptus camaldulensis subsp. camaldulensis P  Eucalyptus camaldulensis subsp. obtusa P  Apx Table B-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, 2021a; 2021b; 2021c) and bark texture 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. * Denotes species for which subspecies exist in the ALA datasets and are included in mapping (Figure 5-13). PAPERBARK SWAMP SPECIES GDE ROPER Melaleuca acacioides  *Melaleuca alsophila   Melaleuca argentea   Melaleuca cajuputi P  Melaleuca clarksonii Melaleuca citrolens P  Melaleuca dealbata   *Melaleuca ferruginea  Melaleuca foliolosa Melaleuca fluviatilis  *Melaleuca lanceolata  Melaleuca leucadendra   Melaleuca minutifolia  Melaleuca nervosa   Melaleuca saligna Melaleuca stenostachya  PAPERBARK SWAMP SPECIES GDE ROPER Melaleuca tamariscina  *Melaleuca trichostachya P Melaleuca viridiflora   Apx Table B-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 Roper catchment based on ALA data (Atlas of Living Australia, 2021a; 2021b; 2021c). 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 MVF species present in ALA datasets are included in mapping. MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE ROPER 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  Calophyllum soulattri typical Canarium australianum typical typical  Capparis lasiantha typical P  Capparis sepiaria typical  Carpentaria acuminata typical P  MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE ROPER Cayratia maritima typical  Celtis philippensis typical P  Celtis strychnoides Clerodendrum floribundum P  Cordyline terminalis typical Croton habrophyllus P  Cupaniopsis anacardiodes 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  Glycosmis trifoliata typical  Gmelina schlechteri typical  Grewia breviflora P  Gymnanthera nitida typical MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE ROPER 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  Melhania oblongifolia typical  Micromelum minutum typical  Mimusops elengi  Nauclea orientalis typical   MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE ROPER Nephrolepis biserrata typical  Operculina aequisepala P  Opilia amentacea P  Passiflora foetida typical  Piper novae-hollandiae diagnostic Planchonella DNA 47005 typical ? Planchonia careya P  Pleomele angustifolius 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  Stychnos 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  Terminalia subacroptera typical Tinospora smilacina P  Tylophora cinerascens   MONSOON FOREST SPECIES DRAINAGE LINE SEASONALLY FLOODED SPRING GDE ROPER Vincetoxicum cinerascens P  Vitex glabrata typical  Xanthostemon eucalyptoides diagnostic  As Australia’s national science agency and innovation catalyst, CSIRO is solving the greatest challenges through innovative science and technology. 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