Impacts of climate change on Australian marine life-Part A: Executive Summary

Editors: Alistair J. Hobday, Thomas A. Okey, Elvira S. Poloczanska, Thomas J. Kunz, Anthony J. Richardson CSIRO Marine and Atmospheric Research report to the Australian Greenhouse Office , Department of the Environment and Heritage September 2006

Impacts of Climate Change on Australian Marine Life Part C: Literature Review

Published by the Australian Greenhouse Office, in the Department of the Environment and Heritage ISBN: 978-1-921297-07-6 © Commonwealth of Australia 2006 This work is copyright. Apart from any use as permitted under the Copyright Act 1968, no part may be reproduced by any process without prior written permission from the Commonwealth, available from the Department of the Environment and Heritage. Requests and inquiries concerning reproduction and rights should be addressed to: Assistant Secretary Land Management and Science Branch Department of the Environment and Heritage GPO Box 787 CANBERRA ACT 2601 This report is in 3 parts: Part A. Executive Summary Part B. Technical Report Part C. Literature Review Please cite this report section as: Hobday, A.J., Okey, T.A., Poloczanska, E.S., Kunz, T.J. & Richardson, A.J. (eds) 2006. Impacts of climate change on Australian marine life: Part C. Literature Review. Report to the Australian Greenhouse Office, Canberra, Australia. September 2006.

Disclaimer The views and opinions expressed in this publication are those of the authors and do not necessarily reflect those of the Australian Government or the Minister for the Environment and Heritage.

While reasonable efforts have been made to ensure that the contents of this publication are factually correct, the Commonwealth does not accept responsibility for the accuracy or completeness of the contents, and shall not be liable for any loss or damage that may be occasioned directly or indirectly through the use of, or reliance on, the contents of this publication.

Editors: Alistair J. Hobday, Thomas A. Okey, Elvira S. Poloczanska, Thomas J. Kunz, Anthony J. Richardson CSIRO Marine and Atmospheric Research report to the Australian Greenhouse Office , Department of the Environment and Heritage

This report is in 3 parts: Part A. Executive Summary Part B. Technical Report Part C. Literature Review September 2006

Impacts of Climate Change on Australian Marine Life Part C: Literature Review

Table of Contents

List of Figures vi

List of Tables vi

Acknowledgements vii

1. Impacts of Climate Change on Phytoplankton 8

1.1 Overview 8

1.2 Introduction 9

1.3 Expected impacts of climate change 11

1.4 Observed impacts of climate change 15

1.5 Impacts and stressors other than climate change 17

1.6 Conclusion 18

2. Impacts of Climate Change on Zooplankton 19

2.1 Overview 19

2.2 Introduction 20


2.4 Observed impacts of climate change (or relationships) 25


2.6 Conclusion 25

3. Impacts of Climate Change on Seagrasses 27

3.1 Overview 27

3.2 Introduction 28

3.3 Expected impacts of changes in climate 30


3.5 Impacts and stressors other then climate change 34

4. Impacts of Climate Change on Mangroves 36

4.1 Overview 36

4.2 Introduction 37



4.5 Impacts and stressors other then climate change 42

5. Impacts of Climate Change on Kelp 44

5.1 Overview 44

5.2 Introduction 44

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6. Impacts of Climate Change on Rocky Shores 53

6.1 Overview 53

6.2 Introduction 53



6.5 Impacts and Stressors Other than Climate Change 57

7. Impacts of Climate Change on Coral Reefs 59

7.1 Overview 59

7.2 Introduction 59

7.3 Climate change in Australia’s tropical and subtropical areas 62

7.4 Potential changes to coral reefs as a result of climate change 65

7.5 Conclusion 67

8. Impacts of Climate Change on Deep-sea and Cold-water Corals 69

8.1 Overview 69

8.2 Introduction 69



8.5 Impacts other than climate change 75

9. Impacts of Climate Change on Soft Sediment Fauna 77

9.1 Overview 77

9.2 Introduction 78




9.6 Conclusion 86

10. Impacts of Climate Change on Benthic and Demersal Fishes 88

10.1 Overview 88

10.2 Introduction 88




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11. Impacts of Climate Change on Pelagic Fishes 94

11.1 Overview 94





12. Impacts of Climate Change on Marine Turtles 102

12.1 Overview 102



12.4 Observed impacts 106

12.5 Threats other than climate change 108

13. Impacts of Climate Change on Seabirds 110

13.1 Overview 110



13.4 Observed impacts of climate change (or relationships) 112


14. References 114

14.1 Phytoplankton 114

14.2 Zooplankton 119

14.3 Seagrasses 121

14.4 Mangroves 127

14.5 Kelps 130

14.6 Rocky shores 135

14.7 Corals 138

14.8 Deep-sea and cold-water corals 142

14.9 Soft sediment fauna 146

14.10 Benthic and demersal fishes 153

14.11 Pelagic fishes 154

14.12 Marine turtles 157

14.13 Seabirds 163

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LIST OF FIGURES

Figure 1-1: Phytoplankton provinces in Australia’s marine environment. .....................10 Figure 2-1: Distribution and abundance of some zooplankton around Australia. ..........20 Figure 4-1: Mangrove species richness around the Australian coastline ........................37 Figure 5-1: The distribution of the kelp species along the Australian coastline ............45 Figure 7-1. Map of shallow coral reefs in the Australian mainland bioregions.............60 Figure 8-1: Known distribution of cold water corals around Australia ..........................70 Figure 11-1: Diversity patterns in large pelagic fish predators around Australia. ..........96

LIST OF TABLES

Table 2-1: Size classes of zooplankton in marine systems. ............................................20

ACKNOWLEDGEMENTS

This review was supported by the Australian Greenhouse Office and the CSIRO Wealth from Oceans National Research Flagship. We thank and acknowledge the additional authors of the literature reviews in Part C of this report: Russ Babcock, Alan Butler, David Milton, Nadia Engstrom, Ove Hoegh-Guldberg, and Richard Matear. We thank Gina Newton and Anna van Dugteren from the Australian Greenhouse Office for their helpful comments and advice on improving the report. Our appreciation to Lisa Albinsson, Sue Blackburn, Gustaaf Hallegraeff, Tony Koslow and Jock Young for providing valuable information regarding plankton; Julie Phillips for comments and advice on Australian marine macroalgae and related research recommendations; Simon Thrush and John S. Oliver for advice on impacts to soft sediment communities; Andrew Baird, Terry Done, Terry Hughes and Iain Suthers for information on time series; Wenju Cai for information on atmospheric and oceanic connections; and Sarah Metcalf for advice on Tasmanian coastal fisheries. Data extraction for a range of the vulnerability indicators was undertaken by Mike Fuller. Access to marine pest data for the vulnerability index was granted by National Introduced Marine Pest Coordinating Group and is gratefully acknowledged. We thank Vivienne Mawson for her editing work, and Toni Cannard for skilful assistance with compiling and formatting the report sections. Editing assistance was also provided by Toni Cracknell. We thank Peter Rothlisberg and Alan Butler for providing guidance and advice. We also thank Tim Skewes and Tom Taranto for producing the map of Australian coral reefs and for the spatial analysis of those reefs that contributed to the vulnerability assessment. Images supplied by Mick Haywood, Frank Coman, Sue Blackburn and CSIRO Marine and Atmospheric Research. Thanks to Peter Parks at ImageQuest3D for the jellyfish image. Phytoplankton and zooplankton maps are from Hallegraeff et al. (in press) and CSIRO, respectively. Karen Miller provided the map of the distribution of deep sea corals. Cover design by Louise Bell.

IMPACTS OF CLIMATE CHANGE ON AUSTRALIAN MARINE LIFE: PART C. LITERATURE REVIEWS

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1. IMPACTS OF CLIMATE CHANGE ON PHYTOPLANKTON

Thomas J. Kunz & Anthony J. Richardson, CSIRO Marine and Atmospheric Research, [email protected]

1.1 Overview

Phytoplankton are diverse microscopic plants inhabiting the light-illuminated surface waters of the world’s the oceans and freshwater ecosystems. Phytoplankton provide most of the primary production in the oceans, thus supporting higher trophic levels as well as playing a central role in the global carbon, oxygen, and nutrient cycles. They are the major food source for young fish, crabs, prawns, and other zooplankton and shoreline filter-feeders such as mussels and oysters. On occasion, toxic phytoplankton species form harmful algal blooms (HABs) that can cause severe health problems and fish kills and may require the temporary closure of mariculture farms. Temperature and stratification of the surface ocean are key determinants of phytoplankton community composition and production. Warmer temperatures, and enhanced southward flow of the East Australian Current should move tropical phytoplankton further south, and HABs may become more frequent. The relatively productive south-eastern temperate phytoplankton province is likely to shrink considerably in area and may constrict to an area west of Tasmania by the 2070s. The change will likely to have considerable repercussions for higher trophic levels (fish, turtles, seabirds and whales), but precise estimations of these repercussions will be very challenging. Although there are few documented changes in Australian phytoplankton, perhaps reflecting limited sampling, the range of one tropical phytoplankton strain is extending southward in eastern Australia. Satellite observations of the sea surface also suggest that phytoplankton abundance may already be changing regionally.

IMPACTS OF CLIMATE CHANGE ON PHYTOPLANKTON

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1.2 Introduction

Definition and distribution

The illuminated zone of coastal and oceanic waters is inhabited by a great diversity of tiny, usually single-celled, microscopic algae known as phytoplankton. Like all plants, phytoplankton utilise sun light, water, and dissolved CO2 and nutrients to grow and reproduce. Phytoplankton usually occur at densities of a few hundred to hundreds of thousands of cells per litre of seawater. In favourable conditions the concentration can increase to several million cells per litre which may become apparent as algal blooms. Most phytoplankton species in Australian waters belong to these major groups of planktonic algae: diatoms, dinoflagellates, small flagellates, coccolithophores and cyanophytes. Broadly speaking, these taxa differ in morphology, biochemistry, motility, sinking characteristics, competitiveness for resources, and growth rate. The surprising diversity of phytoplankton life forms in the relatively homogeneous pelagic zone reflects both the multitude of physicochemical factors they are exposed to and the various mechanisms that increase chances of survival. With current research recognising that phytoplankton species in marine Australian waters represent ecophenotypes of species known from elsewhere in the world, i.e. varieties that have morphological and/or physiological adaptations to local environmental conditions. The estimated number of species is on the order of several 10,000 (S. Blackburn, personal communication). General phytoplankton provinces in marine Australian waters are shown in Figure 1-1.

Roles and ecosystem services (ecological, social, economic)

Phytoplankton are the primary source of organic material and energy in oceanic food webs and they also make up a considerable portion of the primary production in coastal food webs. Zooplankton, including the larvae of many marine benthic invertebrates, fish, and other commercial species such as mussels and oysters feed directly on phytoplankton. Pelagic fish species such as pilchard, jack mackerel, and tuna often feed on zooplankton and so depend on phytoplankton indirectly. Phytoplankton have a central role in the global biogeochemical cycles of key elements such as carbon, nitrogen, and phosphorus, and provide about 50 % of all oxygen on the planet. Because phytoplankton consume mineral nutrients including those from human and industrial waste water, they reduce the amount of nutrients available for the growth of nuisance microorganisms. The ability of the oceans to act as a sink for excessive production of anthropogenic CO2 relies largely on its conversion into biomass by algal photosynthesis and subsequent sinking of phytoplankton to greater ocean depths (the ‘Biological Pump’; Jahnke 1996, Lampitt & Antia 1997). Some phytoplankton species are producers of dimethylsulfonium propionate, a precursor of dimethylsulfoxide (DMS) (Blesinger et al. 1994). DMS evaporates from the ocean and, in the atmosphere, is oxidised into sulfate, which forms condensation nuclei for clouds. Because of their effect on cloud formation phytoplankton indirectly affect precipitation and coastal runoff, and eventually, salinity, stratification, and nutrient supply in the ocean. Clouds are also important for reflecting sunlight (the albedo effect). Phytoplankton therefore have important feedback effects on the climate.


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Figure 1-1: Phytoplankton provinces around Australia. In northern shelf waters westwards from Torres Strait tropical diatom species dominate, with slight regional differences in relative abundances and absolute biomass (1a-c). The shallow waters of the Great Barrier Reef region (3) are dominated by fast-growing nano-sized diatoms. The deeper waters of the Indian Ocean and the Coral Sea are characterised by a tropical oceanic flora (2a and 2c, respectively) that is dominated by dinoflagellates and follows the Leeuwin Current (2b) and the East Australia Current and its eddies (2d). South-eastern coastal waters harbour a temperate phytoplankton flora (4) with seasonal succession of different diatom and dinoflagellate communities. Waters south of the tropical and temperate phytoplankton provinces are characterised by an oceanic transition flora (5a,b) that communicates to the subantarctic phytoplankton province (6) and is highly variable in extent (Figure prepared by G.M. Hallegraeff for CSIRO and National Oceans Office). Some phytoplankton species produce toxins and may become a problem for natural ecosystems and humans if they grow to large numbers, resulting in so-called red tides or harmful algal blooms (HABs). Most toxic species of phytoplankton are dinoflagellates (Alexandrium and Gymnodinium spp.) or cyanobacteria (Nodularia spumigena, Trichodesmium erythraeum; see Falconer 2001 and references therein), but some are diatoms. Although known for forming nuisance blooms rather than harmful algal blooms, the most common red-tide forming species in south-eastern Australian waters nowadays is the dinoflagellate Noctiluca scintillans which started to regularly move into Tasmania with the EAC from 1994, and became an overwintering resident in Tasmania around 2003 (Ajani et al. 2001, G. Hallegraeff, personal communication). Many species of zooplankton and shellfish feed by filtering phytoplankton and so incorporate toxins when HAB species are present. Fish, sea birds, and whales that consume affected zooplankton and shellfish may exhibit a variety of responses detrimental to individual survival (Landsberg 2002, Pearson 2003, Shumway et al. 2003). If concentrated in mussels or oysters, the toxins can cause amnesic, diarrhetic or paralytic shellfish poisoning in humans and so may require the closure of aquaculture farms. For example, blooms of Gymnodinium catenatum in southern Tasmanian waters led to the closure of shellfish farms in the 1980s and 1990s (Hallegraeff & Sumner 1986, Hallegraeff et al. 1988, Hallegraeff 1995). A bloom by Karenia cf. mikimotoi related to the incursion of warm EAC water killed $4 M worth of salmonids in south-


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eastern Tasmanian aquaculture farms in 2003 (G. Hallegraeff, personal communication). Recreation and tourism may also be impacted where toxic phytoplankton occur at high densities, as humans swimming in affected waters can experience skin irritations and respiratory problems. General impacts of climate change on phytoplankton

Phytoplankton have short generation times and are non-sedentary, being largely moved passively with water masses and currents. They may therefore respond to climate variability and change, which alters the physical and chemical environment of plankton, with only a comparatively small time lag. One of the consequences of this is that changes in the plankton manifest faster than changes in most other marine groups.

1.3 Expected impacts of climate change

The ability to respond quickly to environmental change with changes in abundance is an important factor in the observed seasonal cycles in phytoplankton communities. Below we describe how phytoplankton is affected by a suite of environmental parameters and discuss how phytoplankton abundance (biomass) and community composition can be expected to respond under a climate change scenario.

Sea surface temperature Sea surface temperature is an important control of the distribution and abundance of phytoplankton species. Because increasing temperature enhances the metabolic rates of algal cells the growth rate of phytoplankton species is faster at higher temperature but drops considerably beyond an optimal temperature (Eppley 1972, Schoemann et al. 2005). Phytoplankton grow over a range of temperatures that are usually characteristic of the habitat (Suzuki & Takahashi 1995). Natural populations of phytoplankton are, however, often found at temperatures suboptimal for species-specific photosynthesis, and it is believed this is to avoid risking abrupt declines in growth rate associated with sudden increases in temperature (Li 1980, Smayda 1980). A recent modelling study found that temperature variability of the habitat selects against species with lower optimal growth temperature and that phytoplankton communities should be dominated by phytoplankton that grow maximally at temperatures higher than the mean temperature of that habitat (Moisan et al. 2002). Sea temperature is also an important trigger of the germination of resting stages of diatoms and dinoflagellates and it therefore affects the timing of phytoplankton blooms, particularly in coastal regions (McQuoid & Hobson 1995). Increasing sea surface temperature should generally result in a shift in phytoplankton community composition towards warm-water species and reductions in the abundance of the existing phytoplankton flora. Cyanophytes and dinoflagellates (including HAB species) may increase in regions where water warms and stratification intensifies, particularly in the southwest and from northern to southeastern Australian waters (Boyd and Doney 2002). Such changes have been observed in extensive areas of the northeast Atlantic since the 1950s (Edwards & Richardson 2004, Hays et al. 2004, Richardson & Schoemann 2004). The increase of harmful algal blooms in Norwegian coastal waters


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over the same time period has been related to climate variability and change (Sellner et al. 2003, Edwards et al. 2006). Rainfall, coastal runoff, and salinity

Changes in the amount or timing of rainfall and the associated river runoff affect the salinity of estuaries and adjacent coastal waters. Salinity is relatively constant throughout the year in most oceanic waters but may vary considerably in estuarine and coastal waters. While estuarine phytoplankton have been found to grow best at low salinities typical for estuaries, oceanic species seem to grow best at higher seawater salinities which are characteristic for most oceanic regions; coastal phytoplankton appear to prefer an intermediate salinity range but to tolerate lower salinity (Brand 1984, McQuoid 2005). However, some phytoplankton appear to thrive over a wide range of salinities (e.g. a Skeletonema costatum isolate from a Swedish fjord grew well from 15 -35 psu in experimental conditions) or exhibit seasonal differences in growth response to salinity (McQuoid 2005). In coastal regions where diatoms and dinoflagellates produce resting stages to outlive unfavorable seasons, salinity may be an important cue for germination time (McQuoid 2005). Freshwater influx to coastal waters not only affects salinity but also modifies the stratification of the water column thereby affecting nutrient resupply from below. While diatoms seem to be negatively affected by the decrease in nutrient concentrations associated with river discharge in relatively nutrient-poor river catchments, dinoflagellates may profit from freshwater input as this usually increases stratification and the availability of humic substances (Carlson et al. 1995, Edwards et al. 2006, Goffart et al. 2002). Irrespective of the direction of change, changes in rainwater runoff and associated changes in salinity and resource supply can therefore affect the composition and, potentially, the productivity of the phytoplankton community in coastal waters. Wind Stress and Currents

Wind can have a direct effect on phytoplankton by generating water column turbulence that can physically kill sensitive phytoplankton by disrupting internal structures (microtubules). In particular, dinoflagellates such as Gymnodinium catenatum are vulnerable (G. Hallegraeff, personal communication). Most effects of wind on phytoplankton are, however, indirect and involve a range of processes. Wind at the ocean basin scale determines the surface ocean regions of upwelling and downwelling, which strongly affects the supply of macro-nutrients that are essential for phytoplankton growth at the ocean surface. At the region scale, winds can interact with the surface ocean currents to enhance or suppress local upwelling. Wind-driven currents may also transport phytoplankton into other regions, and affect the size and frequency of formation of mesoscale oceanic features such as fronts and eddies (Martin 2005). Locally, wind intensity strongly influences depth and intensity of vertical mixing in the surface layer thereby affecting phytoplankton access to nutrients stored in deeper water layers, light availability for phytoplankton photosynthesis and phytoplankton exposure to potentially harmful UV-B radiation (Mann & Lazier 1996, Garrett 2003). Finally, winds can influence the supply of iron to the surface ocean through aeolian transport of dust from land to sea. In regions south of Tasmania, where


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macro-nutrient concentrations are always high, iron availability influences growth, biomass, and composition of phytoplankton (Sedwick et al. 1999, Boyd et al. 2000). In the macro-nutrient limited regions more typical of the waters around continental Australia, the atmospheric supply of iron may stimulate nitrogen-fixing cyanobacteria which have a higher iron requirement than other phytoplankton (Jickells et al. 2005, Mahaffey et al. 2005). Changes in currents such as the East Australia Current are likely to suppress existing areas of upwelled nutrient-rich waters, or create new ones, depending on local underwater topography and oceanography. As a result, phytoplankton production may experience reductions in some regions but increases in others. The projected greater southward penetration of the East Australia Current should transport tropical phytoplankton and nutrient-poor water from the Coral Sea further poleward in the future replacing temperate phytoplankton. Changes in existing wind and current patterns may therefore considerably affect phytoplankton community composition regionally, and changes in primary production will have repercussions for higher trophic levels and ecosystem services. UV radiation Like visible, photosynthetically active radiation, the highly energetic ultraviolet radiation (UVR) penetrates the water surface and may affect phytoplankton in various ways. It has been shown that UVR can negatively affect several physiological processes and cellular structures of phytoplankton. These include photosynthesis, nutrient-uptake, cell motility and orientation, algal life-span, and the DNA (Häder et al. 1991, Jeffrey et al. 1996, Hessen et al. 1997, Wilhelm et al. 1997, Hogue et al. 2005, Litchman & Neale 2005). While shorter wavelengths generally cause greater damage per dose, inhibition of photosynthesis by ambient UVR increases linearly with increasing total dose (Lorenzen 1979, Smith & Baker 1980). The extent to which UVR affects phytoplankton depends on both environmental conditions and on biological characteristics of phytoplankton cells. While cloudiness affects the absolute amount of both incident visible light and UVR, dissolved and particulate matter reduce the penetration of UVR in water (Kirk 1994, Neale et al. 1998). In clear oceanic waters UV-B radiation can reach depths of at least 30 metres (Smith & Baker 1979, Behrenfeld et al. 1993). By contrast, UVR is largely absorbed within the top few metres of the water column where concentrations of inorganic and organic matter are high (see Keller et al. 1997 for evidence from a field experiment conducted in Narragansett Bay, Rhode Island, USA). The degree of stratification of the water column also crucially affects phytoplankton exposure to UVR. Decreasing depth of the mixed surface layer, and increasing turbulence both increase exposure of phytoplankton to UVR and thus the chance of algal cells receiving harmful doses (Keller et al. 1997, Neale et al. 1998, Hernando & Ferreyra 2005). Phytoplankton damage by UVR may largely be avoided when the depth of the surface mixed layer is close to or deeper than the euphotic depth, i.e. the light-illuminated part of the water column (Barbieri et al. 2002). Although some phytoplankton may partly acclimate to, compensate for or repair damage by UVR, this involves metabolic costs thereby reducing energy available for cell growth and division. Effects of UV radiation on phytoplankton physiology and


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morphology, and protective mechanisms involve profound changes in cellular elemental stoichiometry including increased cellular carbon content, decreased content of chlorophyll a (Chl-a; the main photosynthetic pigment in algal cells), and less frequent cell division resulting in increased cell size (see Hessen et al. 1997 for a review). Studies by Sugawara et al. (2003) and Raven et al. (2005) suggest that UVR intensity affects the size-ratio in phytoplankton communities because small cells are more prone to UVR and have comparatively high metabolic costs to screen out damaging UVR. Many of these changes decrease the nutritional value of phytoplankton. Negative effects of altered food quality have indeed been observed in zooplankton, but not in higher trophic levels (Keller et al. 1997). UV radiation enhances the photochemical decomposition of dissolved organic carbon and colloids and thereby makes essential phytoplankton macronutrients (e.g. nitrate) and micronutrients (e.g. iron and copper) more readily available (Sunda et al. 1981, Rich & Morel 1990, Palenik et al. 1991, Wängberg et al. 1999). Depending on the type of phytoplankton and UV intensity this may translate into increased phytoplankton growth. In Australia, UVR effects on phytoplankton have not been investigated. Internationally, a number of studies have addressed UVR effects at the species level (e.g. Garde & Cailliau 2000, Rech et al. 2005) but only few studies have looked at long-term effects on phytoplankton communities in the field (but see Keller et al. 1997, Wängberg et al. 1999). pH and seawater chemistry The pH of open ocean seawater varies little both spatially and temporally (Skirrow 1975), with a global present-day average of 8.2 units (8.3 units in pre-industrial times). By contrast, the pH of coastal waters may vary by up to 1 unit in response to precipitation and biological activity in the plankton and sediment. Seawater pH may affect phytoplankton via several processes. pH is an important determinant of the growth rate of phytoplankton species, some of which grow well over a range of pH values, while the growth rate of others varies considerably between pH 7.5 to 8.5 (reviewed in Hinga 2002). The calcification rate and quality of calcitic shells (liths) that cover coccolithophore cells has been shown to decrease as waters become more acidic (Riebesell et al. 2000, Engel et al. 2005). Decreasing pH has also been observed to increase the availability of potentially toxic trace elements such as copper (Kester 1986) and to negatively affect nitrification in marine bacteria, which is an important pathway of nitrate supply to phytoplankton, at pH < 8.0 (Huesemann et al. 2002). All these effects should have repercussions for phytoplankton community composition and productivity.

Phytoplankton rely mostly on the carbon from dissolved CO2 for the production of organic material. They also may utilise HCO3

-, but this comes at higher energetic costs. Because the relative consumption of HCO3

- and CO2 differs between phytoplankton species, changes in their availability may thus affect phytoplankton on the cellular, population and community levels (reviewed in Riebesell 2004). Increased CO2 concentration appears to have relatively small effects on phytoplankton productivity (e.g. Tortell et al. 2002, Schippers et al. 2004, Engel et al. 2005). It is difficult to draw general conclusions from the few studies that investigated effects of CO2 concentration


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on phytoplankton community composition, but it is likely that changes in carbon speciation will affect the relative abundances of different phytoplankton groups with repercussions for the entire food web (Riebesell 2004).

Mixed layer depth and stratification The depth of the mixed surface layer affects sea surface temperature (which roughly corresponds to the temperature across the mixed layer), the supply of light and nutrients to phytoplankton (from above and below, respectively), and phytoplankton sinking losses. Mixing depth therefore strongly affects phytoplankton production and abundance in surface waters and deeper (Reynolds 1984, Mitchell & Holm-Hansen 1991, Huisman & Weissing 1995, Diehl 2002, Kunz 2005). Changes in mixing depth may therefore affect phytoplankton production and community structure in the entire water column. Climate models predict decreases in mixed layer depth and increasing stratification in response to warming for large regions of the global ocean, including most Australian waters. This will result in reduced nutrient transport to the surface layer over vast areas of the pelagic zone with repercussions for phytoplankton production. Cyanobacteria, flagellates and dinoflagellates (including nuisance and harmful algal bloom species) may increase in abundance where vertical mixing decreases. The ‘microbial loop’ may replace the relatively more productive ‘classic’ food web in affected areas. As noted previously, the productive temperate phytoplankton province may shrink considerably in area and potentially become restricted to west of Tasmania by 2100. Available evidence of biological effects of changes in stratification comes from the subtropical waters surrounding the North-west Hawaiian Islands, where Chl-a concentration and primary production were low during periods when stratification was supposedly relatively strong (Polovina et al. 1994). In the North Atlantic, increasing water column stratification has been linked to large-scale northward shifts in the distribution of warm-water phyto- and zooplankton between 1958 and 2002, as well as changes plankton abundance during that same period (Richardson & Schoeman 2004, Hays et al. 2005, Edwards et al. 2006). Likewise, the earlier timing of dinoflagellate blooms in spring has been partly attributed to earlier and enhanced stratification (Edwards & Richardson 2004, Richardson & Schoeman 2004).

1.4 Observed impacts of climate change

Because of the near absence of plankton time-series from Australian waters, observations of climate change impacts on marine phytoplankton are mainly anecdotal. Below we summarise observed changes and discuss potential impacts on phytoplankton and community composition. We also address flow-on effects to higher trophic levels and discuss potential consequences for species diversity and community composition. Abundance/distribution Despite the paucity of plankton records in marine Australian waters, a small number of range phytoplankton expansions and associated changes in phytoplankton community


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composition have been documented and related to warming and more southward penetration of the warm East Australia Current into cooler, temperate waters: A tropical strain of the coccolithophore Gephyrocapsa oceanica has been observed to form extensive blooms in temperate south-eastern Australian waters on several occasions since 1992 (Blackburn & Cresswell 1993, Blackburn 2005). The dinoflagellate Noctiluca scintillans, which may form nuisance blooms, has regularly moved into Tasmanian waters since the 1990s and has now become a permanent member of the plankton there (G. Hallegraeff, personal communication). Considerable changes in the concentration of surface ocean chlorophyll-a (Chl-a, a surrogate parameter for phytoplankton biomass) have occurred in southern Australian waters between the time periods 1979-1986 and 1997-2000 (Gregg & Conkright 2002). The Chl-a concentration increased by about 50 % in southeastern shelf and adjacent offshore waters in the Oct-Dec period, but decreased by up to 50 % south of Perth and Brisbane in the Jul-Sept period and by up to 30 % in the southern Coral and the northern Tasman Sea in the Oct-Dec period. While the Chl-a decreases in winter may be attributed to mixed-layer deepening, the Chl-a-decrease in the Tasman Sea in spring-summer may have been related to increased stratification (Gregg & Conkright 2002). Analyses of existing time series of remotely sensed sea surface Chl-a are needed to determine whether the observed trends are persistent. Changes in incident UVR may have contributed to the observed decreases in summer surface Chl-a concentration. This is because of stronger seasonal variation in water column stratification and increasing incident UV-B radiation in mid-latitudes of the Southern Hemisphere since the 1970s (Madronich et al. 1998), and because Chl-a decreases were higher in clear waters (northern Tasman Sea, southern Coral Sea and shelf waters) than in more turbid south-eastern Australian coastal and shelf waters waters. The only extensive study on UVR effects on phytoplankton conducted in the South Pacific region showed that photosynthesis was affected comparatively little by UV-B radiation at latitudes < 30 °S but demonstrated a high variability in dose-rate response at higher latitudes (Behrenfeld et al. 1993). The authors attribute their observations to partial photoadaptation of phytoplankton at lower latitudes where the vertical structure of the water column typically is more stable and ambient UVR has typically been higher than in temperate regions. However, until a detailed analysis of remotely sensed surface Chl-a concentrations and incident UVR is conducted, any attempts to explain the observed changes in phytoplankton biomass remain speculative. Phenology

Increasing water temperature and stratification lead to an elongation of the growing season, which may allow subtropical and temperate phytoplankton species to bloom earlier. To date no observations of changes in phenology exist from Australian waters. Changes in the timing of phytoplankton may be most pronounced in temperate south-eastern waters where the greatest warming is predicted (cf. Figure 1-1). In the North Atlantic, the bloom peaks of different phytoplankton have advanced by up to several weeks as compared with the late 1950s (Beaugrand 2004, Edwards & Richardson 2004, Richardson & Schoemann 2004).


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Physiology, morphology, and behavior

Changes in water column stratification and incident UVR may considerably affect the ecological stoichiometry of phytoplankton (i.e. carbon : chlorophyll- a : nutrient ratios) and the diel vertical migration of some phytoplankton with flow-on effects for higher tropic levels. Observations of changes in phytoplankton stoichiometry require a comprehensive sampling approach and do not currently exist in Australian waters. However, the relatively small changes of mixed layer depth predicted for waters of the Australian continental shelf are unlikely to strongly affect phytoplankton physiology. Although surface waters around Australia may not become undersaturated with respect to calcite before 2100, the ability of coccolithophores to grow coccolith shells may already suffer at the lower end of the supersaturated scale (Feely et al. 2004, Orr et al. 2005, Tyrell 2006). Due to a lack of experimental evidence it is currently impossible to predict any thresholds. Likely consequences for biological communities and biodiversity With the southward flow of the warm East Australia Current expected to increase, and warming of the Leeuwin Current, the productive temperate phytoplankton province of south-eastern shelf waters may contract considerably, potentially to the region off western Tasmania. By contrast, Australian tropical phytoplankton provinces may expand with warming. As a considerable part of phytoplankton productivity at the base of marine food webs propagates to higher trophic levels, climate change effects on phytoplankton can be expected to affect the abundances and distribution of a range of marine organisms including, in particular, pelagic fish species. Decreasing phytoplankton production in the southern Coral Sea and northern Tasman Sea may, for example, reduce the productivity of the Eastern tuna and billfish fishery. However, very little is known currently about trophic relationships in the oceans around Australia and so accurate predictions are difficult to make. Some studies in eastern and south-western Australian waters are currently underway (Keesing & Heine 2005, Young et al. 2004).

1.5 Impacts and stressors other than climate change

Fishing Knowledge is currently limited, although it has been suggested that phytoplankton abundance can be affected by the cascading effects of overfishing (Frank et al. 2005). However, the most conclusive examples were from relatively small, semi-enclosed water bodies and not from the open ocean. Other anthropogenic Eutrophication is usually considered a major threat only in enclosed coastal waters, for example Port Phillip Bay, and in shallow marginal seas. A key stressor of phytoplankton communities is, however, the introduction of phytoplankton species from


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elsewhere around the world. The information available on introduced phytoplankton is largely restricted to species of immediate economic and health concern. For example, several non-native, toxic dinoflagellate species were observed to form episodic nuisance or harmful algal blooms in Australian waters in the past few decades: Alexandrium catenella in Port Phillip Bay, Victoria, A. minutum in Port River, South Australia, and Gymnodinium catenatum in southern Tasmania (Hallegraeff et al. 1988, McMinn et al. 1997). All are considered to have been introduced via the ballast water of commercial vessels, although shellfish imported for aquaculture are another potential source (Hallegraeff & Bolch 1992). However, with both the relatively good record of ballast water control that Australia has and the ‘International Convention for the Control and Management of Ships’ Ballast Water and Sediments’ coming into effect from February 2005, it may be possible to limit the number of future phytoplankton introductions to Australian waters.

1.6 Conclusion

Increasing sea surface temperature and stratification, and enhanced southward flow of warm water are the main physical oceanographic changes that can be expected to have a comparatively strong impact on phytoplankton with flow-on effects on higher trophic levels. While the direction of response of temperate marine phytoplankton can be predicted with some confidence based on the knowledge gained from North Atlantic time-series data, a considerable degree of uncertainty remains regarding the response of tropical and subtropical marine phytoplankton in Australia. Monitoring, modelling, and comparative studies are therefore needed.

Main Findings: Phytoplankton • Expected changes

o Increasing SST and southward flow of the East Australian Current will drive phytoplankton species southwards

o Changes in abundance of phytoplankton including HAB species o Earlier timing of the production peak of some phytoplankton o The productive south-eastern temperate phytoplankton province may retreat to

west of Tasmania by the 2070s o Changing seawater chemistry may shift the ratio of phytoplankton groups o Changes in phytoplankton distribution, abundance, and timing of production

may have drastic effects for most marine life o Decreasing pH may have detrimental effects on coccolithophores

• Observed changes within the Australian region o Phytoplankton abundance may already be changing regionally o A tropical phytoplankton has expanded southward o Lack of confirmed observations of change is most likely due to lack of

sampling.

IMPACTS OF CLIMATE CHANGE ON ZOOPLANKTON

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2. IMPACTS OF CLIMATE CHANGE ON ZOOPLANKTON

Anthony J. Richardson, CSIRO Marine and Atmospheric Research, [email protected] Thomas J. Kunz, CSIRO Marine and Atmospheric Research, Hobart

2.1 Overview

Zooplankton are microscopic- to jellyfish-size animals drifting in ocean currents and distributed throughout all marine environments. The zooplankton community consists of species that are permanent residents together with temporary members such as early and larval life history stages of benthic and pelagic species such as fish, crabs, molluscs, worms and seastars. Zooplankton may be the most numerous multicellular animals on earth; they are also the main secondary producers in the oceans, transferring energy from primary producers such as phytoplankton to higher trophic levels including fish, seabirds, and whales. They thus support directly and indirectly all the world’s fisheries and play a central role in the global carbon and nutrient cycles, removing large quantities of carbon from the surface layers of the ocean and distributing it to the deep. Over years to decades, ocean warming will cause large southward shifts in the distribution of many tropical and sub-tropical zooplankton species displacing many local species, as well as resulting in an earlier annual appearance of many groups. As zooplankton are such a critical food source for higher trophic levels, both these impacts will alter trophic and competitive relationships among species and disrupt foodwebs. There are also implications for larval health and transport, and therefore recruitment to adult benthic and pelagic populations. Evidence suggests that enhanced stratification will lead to lower abundance of zooplankton but increased incidence of jellyfish blooms, with potentially dramatic effects on higher trophic levels and people. By the end of this century, acidification of the world’s oceans will affect the processes of production and maintenance of external calcified shells and plates in molluscs, echinoderms, and some crustaceans. To date, there have been no observed impacts of climate change on this group in Australia, but this is due to the paucity of zooplankton sampling. Climate change will alter the distribution, the timing, and composition of zooplankton communities in the future, with repercussions for all marine life including those groups such as fish, marine reptiles and mammals higher up the foodweb.


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2.2 Introduction

Definition and distribution

Zooplankton is a generic term describing organisms that have limited locomotive ability relative to the water bodies in which they live. Zooplankton are highly diverse, containing organisms from almost all animal phyla and families. These range in size from small stages of copepods only 20-200 µm in size and known as microzooplankton to large jellyfish up to 2 m in size called megazooplankton (Table 2-1). Planktonic organisms can be classified or structured in many ways, but the most common is on the basis of size, which affects mobility and trophic status (prey generally smaller than predators). Table 2-1: Size classes of zooplankton in marine systems. Size class Size range Representative organisms Microzooplankton 20 – 200 µm Small stages of copepods,

foraminifera Mesozooplankton 0.2 – 20 mm Amphipods, appendicularians,

chaetognaths, copepods, thaliaceans (doliolids and salps)

Macrozooplankton 20 – 200 mm Euphausiids, heteropods, jellyfish, larval fish, mysids, pteropods, solitary salps

Megazooplankton > 200 mm Jellyfish, Colonial salps Australian zooplankton communities range from those with tropical affinities in the north to those with cold temperature alliances in the south (Figure 2-1).

Figure 2-1: Distribution and abundance of some zooplankton around Australia. Note that eastern Australia is poorly sampled.


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Roles and ecosystem services (ecological, social, economic)

Zooplankton are the most numerous multicellular animals on earth. They are the main secondary producers in the oceans, transferring energy from the primary producers (phytoplankton) to higher trophic levels including fish, seabirds, and whales. They thus support directly and indirectly all the world’s fisheries and play a central role in the global carbon and nutrient cycles, removing large quantities of carbon from the surface layers of the ocean and distributing it to the deep. In Australia, zooplankton and suspended non-living organic particles directly support a wide variety of suspension feeding organisms and planktivorous fish on coral reefs. Most benthic marine invertebrates and fish also have a planktonic life stage during which they are dispersed by currents.

General impacts of climate change on zooplankton

Zooplankton are highly sensitive to their physical and chemical environment, and are thus ideal, early-warning sentinels of climate change, for a number of reasons (Hays et al. 2005). First, unlike other marine species, such as fish and many intertidal organisms, few species of zooplankton are commercially exploited; therefore, any long-term changes can be more clearly attributed to climate change. Second, most species are short lived and so population size is less influenced by the persistence of individuals from previous years. This leads to tight coupling between environmental change and population dynamics. Third, plankton can show dramatic changes in distribution because they are free floating and can respond easily to changes in temperature and oceanic current systems by expanding and contracting their ranges. Finally, recent evidence suggests that zooplankton are more sensitive indicators of change than are even environmental variables themselves, because the nonlinear responses of biological communities can amplify subtle environmental perturbations.


Sea surface temperature

All zooplankton are poikilothermic and thus physiological rate processes and rates of overall growth are highly sensitive to temperature (e.g. Huntley & Lopez 1992), with many zooplankton having a Q10 close to 3 (a tripling in the speed of rate processes for a 10°C temperature rise). All species have a thermal optimum where growth and reproduction is maximal, and thermal limits beyond which net growth ceases. Little information is available on these ranges for Australian zooplankton, and in most cases experiments have been carried out with temperate species. As individual zooplankton species have their own thermal optimum and limits for growth, thus warming will have differential effects on the growth of individual species. As emphasized earlier, changes in zooplankton community composition will influence higher trophic levels. Although direct effects of temperature changes are fundamentally important to plankton rate processes, indirect effects are also critical shapers of plankton growth rates. Zooplankton grow at temperature-dependent maximal rates only when they are food-saturated (Kleppel et al. 1996, Hirst & Lampitt 1998, Richardson & Verheye 1998).


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Indeed, available evidence from tropical Australia indicates that copepod growth and egg production rates are regulated primarily by food availability rather than temperature (McKinnon & Thorrold 1993, McKinnon 1996, McKinnon et al. 2005). For example, generation times of the common coastal tropical copepod Acrocalanus gibber decreased by 25% with a 5˚C rise in temperature because of food limitation (McKinnon 1996). Therefore, zooplankton growth rates appear to be severely food limited in the warm, oligotrophic waters of tropical Australia. However, this relationship may not apply so strongly in cooler temperate waters. Climate impacts on nutrient enrichment processes are thus likely to be at least as important in Australia as temperature effects. Temperature also has an effect on the body size of individual species of zooplankton. Copepod body length typically decreases with increasing temperature (McKinnon 1996). Effects of temperature on upper trophic levels may be strongly mediated by zooplankton size, which is a key determinant of food quality for planktivorous fish. Zooplankton have exhibited some of the largest range shifts of any marine group (Hays et al. 2005). Members of the warm temperate copepod communities in the Northeast Atlantic have moved more than 1000 km poleward over the last 50 years (Beaugrand et al. 2002, Bonnet et al. 2005), although this may be more associated with changing currents than in situ insolation (heating from the sun). Concurrently, cooler water copepod assemblages have retracted further toward the Pole. Overseas studies show that timing of peak abundance of zooplankton is sensitive to climate warming and this can have effects that resonate to higher trophic levels. In the plankton ecosystem of the North Sea, the timing of all zooplankton taxa associated with low turbulent conditions in summer advanced with warming of 0.9°C from 1958 to 2002, with meroplankton (ephemeral members of the plankton) moving forward by 27 days, copepods by 10 days and non-copepod holozooplankton by 10 days (Edwards & Richardson 2004). Magnitudes of these changes in phenology were greater than those observed on terrestrial communities (Root et al. 2003). Meroplankton are extremely temperature sensitive because they are dependent on temperature to stimulate physiological developments and larval release (Kirby et al. in press). Although many plankton species are responding to climate warming, the magnitude of the response differs throughout the community, having profound implications for the assembly, structure and functioning of the pelagic communities and the entire pelagic ecosystem (Edwards & Richardson 2004). The different extent to which functional groups are moving forward in time in response to warming (e.g. phytoplankton are responding more than zooplankton) may lead to a mismatch between successive trophic levels and a change in the synchrony of timing between primary, secondary and tertiary production. Efficient transfer of marine primary and secondary production to higher trophic levels such as commercially-important fish species is largely dependent on the temporal synchrony between successive trophic production peaks in temperate marine systems, including estuaries (Newton 1996). Thus, marine trophodynamics may have already been radically altered by ocean warming, although it is unknown what effect this is having in Australian waters. The distribution and timing of appearance of jellyfish that are dangerous to bathers may change in the future. The box jellyfish (Chironex fleckerii) is currently at the southern limit of its range in North Queensland and causes problems for bathers on North Queensland beaches there during summer. The box jellyfish is likely to expand its range


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further south as oceanic waters warm. The medusa stages of box jellyfish and the small highly-poisonous Irukandji jellyfish whose stings can be fatal to bathers may also occur earlier in the year as waters warm. Only one species, Carukia barnesi, has been demonstrated to cause Irukandji syndrome but at least six other, mostly undescribed, species may also be responsible. pH and seawater chemistry

Many plankton organisms have calcium carbonate shells, making them susceptible to dissolution in acidic waters. Impacts of acidification will be greatest for zooplankton species with calcified (containing calcium carbonate) shells, plates or scales. For organisms to make these structures, seawater has to be supersaturated in calcium carbonate to ensure that once formed, the calcium carbonate does not re-dissolve. Acidification reduces the carbonate saturation of the seawater, making calcification more difficult and dissolving structures already formed. Although many plankton groups do not have calcium carbonate structures, there are a variety of important groups that do possess such structures such as molluscs, echinoderms and some crustaceans. Even marine organisms with calcium carbonate shells differ in their susceptibility to acidification depending on whether the crystalline form of their calcium carbonate is calcite or aragonite. Calcite has a higher stability (is less soluble) than aragonite, making it less susceptible to dissolution. Pteropods (a type of permanently planktic molluscs) are the most sensitive planktonic group because their shell is composed of aragonite, which will be subject to increased dissolution under more acidic conditions (Orr et al. 2005). Foraminifera (protist plankton) and non-pteropod molluscs produce calcite, the more stable form of calcium carbonate. However, these groups can be sensitive to acidification. Undersaturation of aragonite and calcite in seawater is likely to be more acute and happen earlier further south in the Southern Hemisphere and then move northward. This means that calcifying plankton will not be able to move further south into cooler waters in response to warming seas because they are likely to be undersaturated in calcium carbonate. No quantitative work on thresholds of deleterious effects of acidification has been conducted, but experiments on the pteropod Clio pyrimidata at 788 ppm CO2 for 48 hrs (Orr et al. 2005) led to shell deterioration. This experiment was conducted at CO2 levels approximating those that are likely around 2100 under a business-as-usual scenario of greenhouse gas emissions. Pteropods with their aragonite shells are particularly vulnerable to ocean acidification (Orr et al. 2005). In the Southern Ocean and subarctic Pacific Ocean, shelled pteropods are prominent components of the foodweb contributing to the diet of carnivorous zooplankton, myctophid and other fishes, baleen whales, as well as forming the entire diet of gymnosome molluscs. Pteropods can also account for the majority of the annual export flux of both carbonate and organic carbon in the Southern Ocean. Shells from live pteropods dissolve rapidly when placed in water undersaturated with aragonite, similar to the levels that are likely to exist in 2100. If pteropods cannot grow their protective shell, their populations are likely to decline and their range will contract towards lower-latitude surface waters that remain supersaturated in aragonite.


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In Australian waters, zooplankton with calcium carbonate shells will be affected. Pteropods can be locally common in some areas. For example, the pteropod Cavolinia longirostris can form dense aggregations on the Great Barrier Reef during summer (Russell 1935) occurring in such large numbers that their shells wash up on beaches (D. McKinnon, pers. comm.). It is likely that such pteropods will decline in numbers or disappear. Mussel spat will also be detrimentally affected. UV radiation

UVR is known to damage various life stages of holozooplankton (permanent members of the zoooplankton) such as copepods and shrimp, as well as meroplanktonic eggs and larvae of crabs and fish that are temporary members of the plankton (Hunter et al. 1982, El-Sayed et al. 1996). In copepods, UVR has been found to lower fecundity, increase mortality and affect the sex ratio (Karanas et al. 1981, Bollens & Frost 1990). Propagules of zooplankton such as eggs and larvae, which often reside close to the sea surface, along with the neuston community, are most likely to receive high doses of UVR and thus suffer the most adverse effects. Because UV-B radiation can have deleterious consequences for those organisms that lack photoprotective mechanisms, many of the permanent members of the neustonic copepod community which are common in Australia’s warmer waters have pigments to reduce such damage. Photoenzymatic DNA-repair mechanisms have been observed in copepod nauplii and Antarctic zooplankton (Malloy et al. 1997). With respect to fish eggs and larvae (considered part of the meroplankton), effects of UVR radiation have rarely been investigated. The few existing studies found deleterious effects of UV-B radiation in clear waters and confirm the importance of dissolved organic carbon in ameliorating those effects. For example, Hunter et al. (1982) observed the death of 8 % of anchovy larvae in response to a 20 % increase in UV-B radiation. By contrast, the growth rate and mortality of anchovy larvae raised in turbid estuarine water was not significantly negatively affected even by a 50 % increase in UV-B radiation (Keller et al. 1997). UV radiation also has known adverse effects on eggs and larvae of animals on the shore (Lesser et al. 2003, Przeslawski et al. 2004, 2005, Bonaventura et al. 2006), but more research is needed on specific impacts on both adult and larval zooplankton components. Stratification and mixed layer depth

In the North Atlantic, large-scale changes in the distribution and abundance of zooplankton over the last 50 years have been related to changes in water column stratification (Richardson & Schoeman 2004, Hays et al. 2005, Edwards et al. 2006). In the Central and Subarctic North Pacific Ocean, large changes in plankton secondary production have been linked to a decadal-scale climate change event between the mid-1970s and the late 1980s associated changes in winter and spring mixed-layer depth (Hayward 1997). Impacts of these changes in ocean stratification on lower trophic levels appear to have propagated to higher trophic levels with pelagic larvae, including spiny lobster, squid, salmon, reef and flying fishes, sea birds and monk seals. In the northern California Current, macrozooplankton biomass has decreased by 80 percent since 1951 in response to reduced nutrient transport across the thermocline due to warmer sea surface temperatures and enhance stratification (Roemmich & McGowan 1995). Recent modelling work has suggested that increased stratification in the world’s oceans will lead to an overall decline in primary and secondary productivity.


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2.4 Observed impacts of climate change (or relationships)

There is a virtual absence of zooplankton time series from Australian waters, so there are no documented impacts of climate change on zooplankton in Australian waters. This lack of knowledge of potential climate impacts is a direct consequence of the lack of zooplankton monitoring. The longest extant zooplankton time series in Australia is two years and consists of a single cross-shelf transect for all of Australia. Given the diversity of marine habitats in Australia and the economic and social importance of fishing, this does not compare favourably with the rest of the world. The UK has at least four plankton datasets that span more than 40 years. Globally there are zooplankton time series spanning more than 15 years in no fewer than 30 countries, including relatively poor developing nations such as Bulgaria, Chile, Estonia, Greece, Kazakhstan, Latvia, Faroe Islands, Namibia, Peru, Turkey and the Ukraine. Australia is clearly impoverished in the long-term zooplankton datasets urgently required to assess climate change impacts. Priority should be given to commensing such time-series of data collection.


Although it has been suggested that zooplankton abundance can be affected by the cascading effects of overfishing (Frank et al. 2005), there are few robust examples from oceanic waters but more from semi-enclosed water bodies. Thus overfishing in estuaries and semi-enclosed embayments may have local impacts on zooplankton communities. Zooplankton can be introduced via the ballast water of commercial vessels. However, with both the relatively good record of ballast water control that Australia has and the ‘International Convention for the Control and Management of Ships’ Ballast Water and Sediments’ coming into effect from February 2005, the incidence of introducing exotic zooplankton species to Australian waters will be reduced. However, illegal and unrerulated vessels commonly make deep forays into Australian waters potentially introducing exotic species.

2.6 Conclusion

Over years to decades, ocean warming will cause large southward shifts in the distribution of many tropical and sub-tropical zooplankton, displacing many local species. There will also be the earlier annual appearance of many groups. As zooplankton are such a critical food source for higher trophic levels, both these impacts will alter trophic and competitive relationships among species and disrupt foodwebs. Climate change impacts may considerably influence larval health and transport, and therefore recruitment to adult populations, as a major component of zooplankton communities are the early life history stages of benthic and pelagic species. Evidence suggests that enhanced stratification will lead to lower abundance of zooplankton but increased incidence of jellyfish blooms, with potentially dramatic effects on higher trophic levels and people. By the end of this century, acidification of the world’s oceans will affect the processes of production and maintenance of external calcified shells and plates in molluscs, echinoderms, and some crustaceans. To date, there have been no observed impacts of climate change on this group in Australia because of the paucity of sampling. Climate change will alter the distribution, the timing, and composition of


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zooplankton communities in the future, with repercussions for fish, marine reptiles and mammals higher up the foodweb.

Main Findings: Zooplankton • Expected changes

o Large southward movements of tropical and temperate species as the water warms

o Changes in abundance of particular species o Earlier timing of secondary productivity with warming. o Changes in timing could lead to decreased coupling with lower trophic levels

(phytoplankton) and lower fish yields o Changes to larval health and transport leading to recruitment impacts o Species with calcareous shells such as echinoderms, crustraceans and molluscs

(especially pteropods) may not be able to maintain shell integrity in a more acidic ocean

o Increased incidence of jellyfish with enhanced stratification • Observed Changes

o None yet, but this is likely to be due to the paucity of sampling

IMPACTS OF CLIMATE CHANGE ON SEAGRASSES

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3. IMPACTS OF CLIMATE CHANGE ON SEAGRASSES

Elvira S Poloczanska, CSIRO Marine and Atmospheric Research, Cleveland [email protected]

3.1 Overview

Seagrasses are found in shallow coastal waters, generally on soft sediments. Australia has the highest diversity of seagrasses and most extensive seagrass beds globally. Seagrasses are considered ecosystem engineers; they influence coastal nutrient and carbon cycles, baffle water flow, trap and bind sediments and filter coastal waters. Seagrass systems act as a buffer between terrestrial and oceanic systems, so play a vital role in nutrient cycling and are an important CO2 sink. Seagrass beds provide important nursery habitat for many economically valuable species of fish and crustaceans, so declines in seagrass are often detrimental to the economy of local fisheries. They also support internationally important populations of marine turtles and dugongs. Globally, seagrass meadows are in decline. Australia has lost extensive areas of seagrass habitat in the last 50 years mainly through increased anthropogenic inputs to coastal waters reducing water quality. Temperature and water clarity are major factors controlling the biogeographic distributions of seagrasses in Australian waters. The maximum depth that seagrasses are found is limited by light attenuation in the water column and thus by water clarity. As climate warms, it is expected tropical species will extend or shift their ranges southwards and temperate species will retreat. Changes in phenology will occur with alterations in peak flowering times or frequency of flowering episodes, which appear to be triggered by water temperature. Sea level rise will lead to erosion of coastlines or more frequent incidences of flooding, which will increase turbidity of coastal waters and impact on survival of seagrasses. Sea level rise may also lead to the loss of seagrass habitat in deeper waters as water depth increases. As atmospheric CO2 increases, a higher proportion of dissolved CO2 in the ocean should favour seagrass productivity. Consequences could be an increase in seagrass biomass, in seagrass depth limits and an enhancement of the role of seagrass beds in carbon and nutrient cycles. However, as sea temperatures, UV radiation and frequency natural disturbance regimes such as storms and cyclones are also predicted to alter, their combined impacts may negate the benefits of increased CO2 availability. At present, many seagrass beds are under threat from


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other anthropogenic pressures which, coupled with the extra stressors of a rapidly changing climate, will threaten the persistence and resilience of seagrass beds.

3.2 Introduction

Seagrasses are the only flowering plants found in the sea, usually growing completely submerged and pollinating via water movement. Seagrasses have a typical plant structure, with leaves, roots and underground stems called rhizomes that can represent a considerable portion of the plant biomass (King et al. 1990, Duarte et al. 1998). They reproduce by flowering and fruiting but can also spread through branching and growth of rhizomes (King et al. 1990, Duarte et al. 1998, Marbà & Walker 1999). They inhabit shallow coastal waters, generally on soft sediments and of the 60-plus species found worldwide, over half occur in Australian waters (King et al. 1990, Jernakoff et al. 1996, Walker et al. 1999). Australia has the highest diversity of seagrasses and most extensive seagrass beds globally, ranging from tropical species in the north to temperate species in the south (Walker & Prince 1987, Jernakoff et al. 1996, Kirkman 1997, Walker et al.1999). Overlap zones between temperate and tropical biogeographic regions are centred at Shark Bay on the west coast and Moreton Bay on the east (Kirkman 1997, Walker et al. 1999). Seagrass meadows are typically patchy and most consist of a mixture of two or three species (Walker & Prince 1987, Poiner et al. 1987, Lee Long et al. 1993, Nakaoka 2005). The highest biomasses and highest regional species diversity of seagrasses is found in south-west Australia (Walker et al. 1999, Kirkman 1997). Seagrasses occur over a depth range from the intertidal to 30m or more in very clear water, but the mean depth of occurrence for most species is in the top few metres (Walker & Prince 1987, Coles et al. 1993, Lee Long et al. 1993, Jernakoff et al. 1996). Maximum depth at which seagrasses can occur is limited by light attenuation in the water column thus by water clarity (Backman & Barilotti 1976, Bulthuis 1983, Dennison & Alberte 1985, Best et al. 2001). Due to limitations of remote sensing techniques and restrictions of surveys, our knowledge of Australian seagrass distributions in deeper waters (>20 m) and also in areas of high water movement is poor (Walker et al. 1999). Seagrasses are considered ecosystem engineers, structurally altering the environment in which they occur (Walker et al. 2001). They play many roles: they influence coastal nutrient and carbon dynamics (Pergent et al. 1997, Miyajima et al. 1998, Hemminga et al. 1999); baffle water flow (Peterson et al. 2004); trap and bind sediments (Gacia & Duarte 2001, Gacia et al. 2003), and filter coastal waters (Lemmens et al. 1996). Seagrass productivity is among the highest in coastal environments, it is estimated seagrasses contribute about 0.6 to 0.9 Pg C yr-1 (Pg, or petagram, is a metric unit of mass equal to 1015 grams or 1 gigatonne – one billion metric tons) to marine net primary production (Duarte et al. 1998, Duarte & Cebrián 1996). Seagrass and algal ecosystems are particularly valuable on a global scale for their role in nutrient recycling (Costanza et al. 1997). Australian marine waters are generally nutrient poor, especially in nitrate and phosphate (see Jeffrey et al. 1990, Cambridge & Hocking 1997, Lourey et al. 2006). Input from terrestrial sources is generally low (in pristine regions) as Australian soils have low nutrient status and low run-off (Jeffrey et al. 1990) so seagrasses in Australian coastal waters tend to be primarily nitrogen limited (Udy & Dennison 1997, Udy et al. 1999, Alcoverro et al. 2001, Carruthers et al. 2002). Seagrasses absorb nutrients from


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sediments through their extensive root/rhizome system as well as through stems and leaves which, together with mechanisms such as mobilisation and re-allocation of internal nutrient resources enable seagrasses to survive and flourish in nutrient poor waters (Pedersen et al. 1997, Penhale & Thayer 1980, Hemminga et al. 1999, Walker et al. 2004). Seagrasses systems act as a buffer between terrestrial and oceanic systems, receiving inputs of terrestrial nutrients and organic matter so play a vital role in global biogeochemical cycling and are an important CO2 sink (Gattuso et al. 1998, Duarte et al. 2005). Seagrass meadows and other vegetated marine coastal habitats are considered ‘hotspots’ for carbon burial in the ocean and excess organic carbon is exported to the open ocean (Duarte et al. 2005). The macro and micro-algal epiphyte community growing on seagrasses beds are an important component of the seagrass ecosystems, contributing to productivity and nutrient recycling (van Montfrans et al. 1984, Jernakoff et al. 1996, Moncreiff & Sullivan 2001, Sohma et al. 2004, Smit et al. 2005). For example, it is estimated that epiphyte photosynthesis can double the primary production of Zostera marina plants (Mazella & Alberte 1986). The high numbers of benthic invertebrates associated with seagrass beds also contribute significantly to coastal nutrient dynamics (Edgar 1990, Lemmens et al. 1996). For example, suspension feeders living on and in Posidonia meadows are estimated to filter the water column above the seagrass meadow daily, (~5000 litres seawater m-2 seagrass habitat) representing a filtering capacity 25 times higher then suspension-feeding communities in non-vegetated areas (Lemmens et al. 1996). The decline of epifaunal suspension-feeding community as seagrass beds are lost in areas of coastal eutrophication may magnify negative effects of coastal eutrophication and so reduce capacity for successful revegetation by seagrasses (Lemmens et al. 1996). Seagrass beds function as nursery habitats for a wide range of fauna, increasing survivorship through protection from predation and a high abundance of food resources (Heck et al. 2003). They form important nursery habitat for many economically valuable species of fish and crustaceans (Coles et al. 1993, Gray et al. 1998, Rotherham & West 2002, Smith & Sinerchia 2004). Studies in New South Wales suggest that 70% of fish caught commercially and recreationally are associated with seagrasses or mangroves at some stage of their life cycle (see Zann 2000). Seagrass beds are also important foraging habitats for larger piscivorus (fish-eating) fish, the strength of trophic links between seagrass beds and commercially important fish suggests that a decline in seagrass could be detrimental for economy of local fisheries (Hindell 2006). Globally, a number of commercially exploited species from fish to crustaceans and molluscs, have been linked to seagrass beds (see Jackson et al. 2001 for review). Seagrass beds are important nursery areas for prawn species in Australia such as those targeted by the extremely valuable Northern Prawn Fishery (Coles et al. 1987, Watson et al. 1993, Loneragan et al. 1994, 1998, Haywood et al. 1995). Loss of seagrass habitat, for example by severe cyclones, has been associated with a decline in subsequent prawn catch by fisheries (Penn & Caputi 1986, Rothlisberg et al. 1988). Australian seagrass beds sustain internationally important populations of dugongs and marine turtles which are listed as threatened species in the IUCN Red List (see chapter on marine turtles for further information). Dugongs and turtles are harvested by indigenous people and are highly valued for their cultural significance as well as their meat (Marsh 1996). Tropical seagrass beds also act as important nursery habitats for coral reef fish species; the presence of seagrass beds positively influences densities of


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adult fish species on nearby coral reefs (Nagelkerken et al. 2002, Dorenbosch et al. 2005). Seagrass beds are extremely important for biodiversity and drive primary production in coastal waters (Jernakoff et al. 1996, Pergent et al. 1997, Nakaoka 2005, Bloomfield & Gillanders 2005). Seagrass beds support a large number of associated species from algae living on seagrass blades, to large herbivores such as dugongs to small invertebrate grazers and temporal visitors such as crustaceans and fish (Orth et al. 1984). As seagrasses tend to occur in areas of soft sediment, they represent a hard surface for attachment for benthic plants and animals (Jernakoff & Nielsen 1997, Walker et al. 2001, van Elven et al. 2004, Nakaoka 2005). They represent foraging grounds (Jackson et al. 2001) either through direct grazing on seagrasses themselves (Jernakoff & Nielsen 1997, Lanyon 2003, André et al. 2005, Nagelkerken et al. 2006) or feeding on epibionts (an organim living on another organism) and associated fauna (Jernakoff et al. 1996, Jernakoff & Nielsen 1997, Nagelkerken et al. 2006), provide refugia from predation (Edgar 1999a, 1999b, Watson et al. 1993, Jackson et al. 2001), and provide spawning and breeding grounds (Jackson et al. 2001). Globally, seagrass meadows are in decline (Short & Wyllie-Echeverria 1996). Australia has lost extensive areas of seagrass habitat in the last 50 years (Walker & McComb 1992, Walker et al. 1999, Seddon et al. 2000). This decline has mainly been attributed to anthropogenic effects such as dredging and coastal development although natural events such as cyclones, also destroy seagrass beds (Walker & McComb 1992, Short & Wyllie-Echeverria 1996). However, global climate change is expected to have a great impact on seagrasses in the near future, altering community structure and dynamics (see Short & Neckles 1999 for review).

3.3 Expected impacts of changes in climate

Seagrass beds are under threat from both anthropogenic and natural causes. Natural events, such as coastal erosion, abnormally high temperatures, cyclones, storms and heavy and prolonged rain can all damage or destroy seagrass beds (Preen et al. 1995, Short & Wyllie-Echeverria 1996, Kirkman 1997). Pulsed turbidity events, such as river flooding, can deprive seagrass beds of light availability for a short time as well as increasing stress through a drop in salinity. Small seagrasses can withstand light derivation for short time scales (weeks) which are extended for larger species, but after a prolonged period, rapid die-off of seagrass tends to occur (Longstaff & Dennison 1999, Longstaff et al. 1999). Recovery of seagrass communities from disturbance may take many years depending on the extent of impact. Plants can regrow from small tufts or rhizomes or from seeds in the sediments if these are not damaged (Preen et al. 1995, Preen 1995). When disturbance is severe, the ensuing total destruction of substantial parts or all of seagrass beds may occur (Williams 1988, Preen et al. 1995). Predicted increases in temperature, CO2, UV, sea level and storm intensity will all impact on seagrass communities and are discussed below. Temperature

The waters around Australia are expected to increase by 2 – 3°C by the 2070s (see Climate Change section). Most biological processes such as photosynthesis and respiration occur fastest in some optimal temperature range and decline outside this range. For example, optimum photosynthetic temperature range of Halophila ovalis is


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25-30°C, but a response in photosynthetic yield occurs with temperature changes as little as ± 2.5°C (Ralph 1998). A similar optimal range has been reported for a number of seagrass species (Drew 1978, Marsh et al. 1986). However, the favourable temperature range for growth and survival may be lower than the optimal photosynthetic range as respiration costs increase more rapidly with increasing temperature (Marsh et al. 1986). At extreme temperatures (high and low) photosynthetic inhibition occurs and if severe or prolonged could lead to irreparable damage and death (Ralph 1998, Campbell et al. 2006). The frequency of extreme temperature days may increase as climate warms. Extreme high temperatures coinciding with low tides may be partly responsible for major dieback of seagrass beds in southern Australia in early 1993 (Seddon et al. 2000). Water temperature is also a major trigger for the timing of biological events such as flowering and seed germination in seagrass plants (de Cock 1981, Phillips et al. 1983, Inglis & Smith 1998). Rainfall & coastal salinity

Runoff from Australian rivers tends to be highly pulsed, particularly in the tropical north, and carry high sediment loads. Heavy rainfall can lead to flooding and large-scale plumes of turbid water discharging from rivers (Gillanders & Kingsford 2002) which can cause rapid seagrass die-back due to a drastic reduction in light levels and lowering of salinity (Preen et al. 1995, Carruthers et al. 2002). Seagrass plants can tolerate a wide range of salinities, depending on species, but the germination of seeds may be less tolerant, so population resilience to prolonged disturbance may be affected and shifts in community composition will occur (Gillanders & Kinsford 2002, Fernández-Torquemada & Sánchez-Lizaso 2005). Currents and wind stress

Wind stress will determine the wave exposure at coastal sites which controls attributes of seagrass beds and community composition. Increased wave exposure and increased local current speed will enhance mechanical disturbance which may damage seagrass and resuspend sediments, making it difficult for seedlings to establish or persist (Carruthers et al. 2002). It has been found that the percentage of area covered by seagrass plants declines with increasing wave exposure and/or current speed (Fonseca & Bell 1998). Pollination of seagrasses and seed dispersal are dependant on water movement (Ackerman 1986, Verduin et al. 2002). Cyclones and storms

Cyclones and storms can lead to increased turbidity from increased sediment suspension, mechanical damage or destruction of seagrass plants and smothering by sediment deposition (Williams 1988, Preen et al. 1995). In severe cases, complete devastation of seagrass beds occurs. For example, in 1992 two major river flood events and a cyclone in short succession destroyed approximately 1200 km2 of seagrass beds in south west Hervey Bay (Queensland) primarily through reduced water clarity and uprooting of plants (Preen et al. 1995). However, there is no relationship between cyclone intensity and impact on seagrass beds, the degree of destruction will depend on the cyclone track, local topography and coastal hydrodynamics (Rothlisberg et al. 1988). As the intensity of cyclones and storms may increase with global warming, the likelihood of severe perturbations of seagrass beds is expected to increase.


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Sea level rise

Sea level rise may lead to erosion of coastlines or increased incidences of flooding, which will increase turbidity of coastal waters and adversely impact on seagrass communities. Sea level rise may also lead to the loss of seagrass habitat in deep waters as water depth increases but may create new habitat in shallow waters as the sea encroaches inland. Sea level rise should result in a shift in seagrass distributions landward, tracking changes in depth but these will depend on dispersal rates and habitat availability. Shifts will be impeded by coastal development along many shorelines so overall a loss of seagrass beds is expected (Short & Neckles 1999). UV radiation

Excessive UV radiation causes a decline in photosynthetic efficiency of plants and, over a long period, a reduction in plant biomass (Dawson & Dennison 1996, El-Sayed et al. 1996, Häder et al. 1998). The decline in photosynthetic efficiency will vary between species, depending on factors such as the ability to produce UV blocking pigments. For example, the seagrass Zostera capricorni produces more UV blocking pigment when grown at higher light levels (Abal et al. 1994). Seagrasses most sensitive to UV radiation may become restricted at their upper depth limits if UV RADIATIONlevels increase which will result in a shift in species dominance within seagrass beds (Dawson & Dennison 1996). Seawater chemistry

Dissolved inorganic carbon occurs in high concentrations in the sea in different forms and the proportion of each form present is determined largely by pH and the CaCO3 saturation state of seawater (see Climate Change section). Dissolved CO2 is the preferred form used by seagrasses but most is in the form of bicarbonate HCO3

- , which is used with poor efficiency by seagrasses (see Invers et al. 1997; 2002, Short & Neckles 1999). As atmospheric CO2 increases, an increase in dissolved organic carbon and decrease in surface water pH will result in a higher proportion of dissolved CO2 which should favour seagrass productivity (Invers et al. 2002). Consequences could be an increase in seagrass biomass, in seagrass depth limits and an enhancement of the role of seagrass beds in carbon and nutrient cycles (Invers et al. 2002). Differences in ability to extract dissolved CO2 among species would lead to shifts in species distributions within seagrass beds as atmospheric CO2 concentrations rise (Short & Neckles 1999).


There appears to be no published observations explicitly on the observed impacts of climate change on Australian seagrass communities. A discussion of potential impacts of climate change on abundance, distribution, phenology and physiology is given below together with consideration of consequences for associated biodiversity.

Abundance/distribution

An increase in temperature regimes is predicted to lead to a major shift in species distributions towards higher latitudes and this is already being observed globally in many flora and fauna (Parmesan & Yohe 2003, Walther et al. 2005). Temperature is a


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major factor controlling the biogeographic distributions of seagrasses in Australian waters (Walker & Prince 1987, Jernakoff 1996). As climate warms, tropical species will extend ranges southwards and, depending on resource availability and dispersal abilities, temperate species will retreat. The result will be shifts in community composition as a consequence of variability in the rates of species responses. The possibility to shift ranges southwards is confined by the southern coast of Australia which, coupled with high human populations living along the coast in temperate Australia, particularly in the south-east, and the fact that greatest warming of Australian waters is predicted off south-east Australia (see Climate Change section) suggests some cold-temperate seagrass beds may highly threatened by warming seas. Plants growing at sites near northern edges of geographical distributions (in southern hemisphere) may be already near upper thermal limits so will be particularly susceptible to small increases in temperature. Phenology

Seed germination, seedling growth rate and timing of flowering in seagrasses are driven by daylength and water temperatures (Inglis & Smith 1998, Campbell & McKenzie 2004). Therefore, all these biological processes will be affected by climate change with implications for the persistence of populations. For example, regular flowering of the seagrass Posidonia australis occurs between April and June in south-western Australia, probably induced by a seasonal decline in water temperatures (West & Larkum 1979, Cambridge & Hocking 1997). Further north, in Shark Bay, P. australis meadows do not flower every year (Larkum 1976). Similarly, off Central New South Wales on the east coast wide-spread flowering P. australis is also rare (Walker et al. 1988). Shark Bay and central New South Wales are near the northern limits for this temperate seagrass species so the threshold decline in water temperature required to trigger flowering does not happen every year. As a warming of coastal waters is predicted over the next hundred years (see Climate Change chapter) particularly off south-east Australia, episodes of flowering of Posidonia australis may become even rarer in northern meadows. Physiology, morphology

In general, climate change could favour seagrass beds, enhancing productivity as the availability of dissolved CO2 increases. However, sea temperatures, extreme climatic events and climate variability are also predicted to change, and their combined impacts on seagrass communities are more difficult to predict and may negate the benefits of increased CO2 availability. At present, many seagrass beds are under great threat from other anthropogenic pressures (Short & Wyllie-Echeverria 1996) which, coupled with the extra stressors of a rapidly changing climate, greatly threaten the persistence and resilience of seagrass beds. Although average rainfall may decrease in parts of Australia, rainfall extremes may increase and air temperatures are also predicted to increase, resulting in increased droughts with periods of heavy rain. This can lead to episodic periods of major flooding and high turbidity in coastal waters which cause dieback events of seagrass beds. Consequences for biological communities and diversity

Decline of seagrass beds will have dire consequences for the economic value of commercial and recreational fisheries. Seagrass beds support a large number of


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exploited species such as prawns in northern Australia (Hayward et al. 1995, Loneragan et al. 1998, Jackson et al. 2001), so loss of seagrass beds would have an immediate impact on the economic value of associated fisheries. Population declines of globally threathened dugongs have been attributed to loss of seagrass beds (Preen & Marsh 1995). For example, the loss of over 1000 km2 of seagrass beds from Hervey Bay in 1992 caused mass mortality of dugongs in the area, apparently due to starvation, with large numbers of surviving dugongs abandoning the area (Preen & Marsh 1995). As conditions alter, species dominance within seagrass beds will shift which will have profound implications for associated biodiversity and for nutrient recycling within seagrass beds. Shifts in seasonality will lead to a mis-match between species phenology. Ecological interactions are complex and it is difficult to anticipate consequences for ecosystems of changes at population level. Population response rates to changing climate will depend on the life-history characteristics and growth rates of individual species. It is predicted communities will become less resilient as climate changes so more prone to extinction (Emmerson et al. 2005). As seagrasses are considered ecological engineers, increasing local biodiversity and re-cycling nutrients, it can be assumed that loss of seagrass beds will be detrimental for coastal systems.

3.5 Impacts and stressors other then climate change

Fishing

As seagrasses typically occur in shallow coastal waters, they are at high risk from the direct impacts of human activities such as boat use, trampling, and bait harvesting (Walker et al. 1989, Creed & Amado Filho 1999, Skilleter et al. 2006). Seagrass beds may be scarred by boat propellers, where propeller use creates narrow trenches in seagrass beds removing the plants and reducing the abundance of macrofauna in and around scarred areas (Zieman 1976, Dawes et al. 1997, Uhrin & Holmquist 2003). Trawling for prawns may also damage seagrass beds (Kirkman 1997).

Other anthropogenic factors

Declines of seagrass beds in Australia (Kirkman 1997, Abal & Dennison 1996, Walker et al. 1999) and world-wide (Cambridge et al. 1986, Ruiz & Romero 2003) have been attributed to a loss of water quality from increased land use and coastal development. Rivers transport suspended sediments, organic material, nutrients and pollutants into coastal waters. Modification of river catchments and catchment use, for example dams or mining, afforestation and deforestation, or pollution discharge will all determine river characteristics and flood events (Carruthers et al. 2002; Gillanders and Kingsford 2002). River discharge into coastal waters result in turbid plumes which will influence light availability, as well causing reductions in salinity and increased nutrient loading, which may smother seagrass beds (Kirkman 1997, Gillanders & Kingsford 2002). Distribution and abundance of seagrasses in coastal waters around Australia is constrained by limited nutrient availability so increased input from coastal development and agriculture should favour seagrasses to a point (Udy et al. 1999). Generally, increased nutrient availability will favour the development of plankton communities or harmful algal blooms potentially leading to a loss of water clarity, smothering of seagrass beds and anoxia (Kirkman, 1997, Best et al. 2001, Albert et al. 2005). Dredging to clear passages for ships and construction of harbour walls and groynes will


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change local hydrology and may lead to loss of seagrass beds (Kirkman 1997). Dredging, mining and seismic testing have all caused seagrass loss (Kirkman 1997).

Main Findings: Seagrasses • Expected changes

o Productivity of seagrass beds may increase o Warming will drive species distributions southwards o Timing of flowering may shift or flowering episodes decrease/increase o Damage or destruction by flood events may increase o Likelihood of destruction of tropical seagrass beds by severe storms or

cyclones may increase o Frequency of diebacks will increase as number of extreme hot days

increases o Increased likelihood of algal blooms will increase frequency of diebacks

• Observed changes within the Australian region o None as yet. Decline in seagrass beds generally attributed to anthropogenic

impacts although some diebacks may be related to changes in climate

IMPACTS OF CLIMATE CHANGE ON AUSTRALIAN MARINE LIFE: PART C

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4. IMPACTS OF CLIMATE CHANGE ON MANGROVES

Elvira S Poloczanska, CSIRO Marine and Atmospheric Research, Cleveland [email protected]

4.1 Overview

Mangrove communities are diverse assemblages of trees and shrubs that are found fringing most of the coastline of mainland Australia, with the most extensive communities found in the tropics. The mangrove flora of Australia is one of the most diverse globally. Mangroves act as a buffer between land and sea, filtering terrestrial discharge and decreasing sediment loading. They are valuable for their role in nutrient and carbon recycling and export of organic material to coastal foodwebs. Mangroves provide a host of services for coastal communities – they support fish and fisheries, maintain the integrity of the coastline, and stabilise sediments and provide resources such as wood and food. Mangroves frequently occur in conjunction with seagrass beds and coral reefs and there are strong interactions between them. Mangroves support a diversity of species that are ecologically and commercially important for example they act as a nursery and breeding habitat for marine species such as fish, crabs and prawns and they support a diverse bird population. Sea level rise is considered the major climate change threat to mangroves, as mangroves grow in calm areas on shorelines with a low profile so a small rise in sea level may inundate large areas of mangroves. The response of mangroves to sea level rise will depend on coastal dynamics; for example if the sedimentation rate exceeds the rate of sea level rise then mangrove area may not alter but if sea level rise is greater, mangroves may retreat landwards. Mangroves also require freshwater input, which will further increase susceptibility to rising sea levels. Many mangroves live close to their salinity tolerance levels, so increases in storm regimes or reduction of rainfall may also threaten mangrove persistence. Increases in atmospheric CO2 may enhance growth of mangroves but this will depend on other factors such as nutrient availability and salinity levels. Presence of mangroves will act to mitigate climate change effects on coastal communities, providing protection from storms and coastal erosion but are themselves vulnerable to climate change. Further, given the valuable services provided by mangrove systems to local communities, mangrove conservation should be viewed as a useful adaptive management strategy in response to climate change.

IMPACTS OF CLIMATE CHANGE ON MANGROVES

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4.2 Introduction

Mangroves occur chiefly in tropical and subtropical latitudes where they grow at the interface between land and sea (the term mangrove can either refer to an individual plant within a community or the community itself). They occur round most of the Australian coastline, often forming dense intertidal forests where suitable habitat exists, with most diverse and extensive mangroves found in the tropics (Figure 4-1). Over 30% of the tropical Australian coastline is covered with mangroves. The most southerly mangroves in Australia, and indeed globally, occur at Corner Inlet in Victoria – there are no mangroves in Tasmania. Mangrove distribution in Australia is strongly influenced by latitude and rainfall.

Figure 4-1: Mangrove species richness around the Australian coastline (source Environment Australia 2000)

Mangroves grow in sediments that are generally low in oxygen and are highly saline. Different taxa exhibit diverse morphological and physiological mechanisms to deal with such environmental stresses. For example, the main adaptations to deal with excessive salt, depending on species, include salt-excreting glands on leaves, accumulation of salt in the leaves (which is lost when leaves are shed) and filtration of water during uptake by the roots. Some species, such as Rhizophora have aerial prop roots that extend down from branches the trunk while others, such as Avicennia have pneumatophores, aerial roots for gas exchange that grow vertically up from a shallow extensive network of anchoring roots extending above the mud surface. The location of mangrove species on a coastline can be generally described by two factors: estuarine location and intertidal position (Duke 1992). Location on the open coast or in an estuary will determine fresh water input while the intertidal position, that is low, mid or high shore, will determine the length of time the area is inundated by the sea. These factors also influence soil salinity, soil waterlogging and water chemistry in any particular area. A ‘zonation’ of mangrove species can generally be observed going from the sea to the land or upriver from the mouth of estuaries reflecting the ecophysiological response of the plants along these and other environmental gradients and disturbance regimes (Ellison & Farnsworth 1993).


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Mangrove trees reproduce by flowering and seed production. Seedling development is highly specialised, with most species producing viviparous propagules - the seeds germinate while still attached to the parent tree. When shed, the propagules are able to root and anchor rapidly thus reducing the likelihood of being washed out of mangrove areas. The degree of vivipary varies between species but most propagules are buoyant and viable for a short period at least, thus allowing some potential for dispersal along coastlines (Duke et al. 1998). Mangroves support diverse communities of flora and fauna, both marine and terrestrial animals are found in mangroves. For example, in the tropics birds such as the mangrove robin and mangrove fantail are found almost exclusively within mangroves while other terrestrial birds visit the mangroves at various times of the seasonal or tidal cycle. Waterbirds are also found feeding, roosting or nesting within mangroves. The marine fauna associated with mangroves includes sessile benthic species such as barnacles that grow on the trunks, roots and even leaves. Crabs, particularly fiddler crabs, tend to be abundant in mangroves where they are important recyclers of fallen leaf litter. Fish move in and out of the mangroves with the tide or will remain in water-filled creeks during low tide. Mangroves are considered important nursery areas for fish and crustacean species, including many commercial species (Kathirisan & Bingham 2001, Manson et al. 2005, Loneragan et al. 2005). Mangroves provide a range of ecosystem services such as recycling of carbon and nutrients, shoreline protection, and enrich coastal productivity (Constanza et al. 1997, Ewel et al. 1998, Kathirisan & Bingham 2001, Duarte et al. 2005, Bloomfield & Gillanders 2005). Mangroves produce large amounts of litter that contribute to the recycling of nutrients in these coastal systems and potentially enrich coastal waters. Nutrient enrichment by mangroves may be particularly important in waters where nutrient concentrations are typically low (Kathirisan & Bingham 2001). Mangroves trap sediments resulting in a rise of the substrate (if sedimentation is greater than erosion) and presumably reducing turbidity in coastal waters (Furukawa et al. 1997). Coastal fringing mangroves are important for shoreline protection from storms and erosion (Ewel et al. 1998, Badola & Hussain 2005, Kathiresan & Rajendran 2005).


Mangroves are adapted to deal with a naturally stressful environment but as they live close to their tolerance levels, they may be particularly sensitive to anthropogenic disturbances (Kathiresan & Bingham 2001). Over a third of mangrove forests have been lost globally over the last 50 years, primarily through coastal development and clearing for aquaculture and agriculture (Alongi 2002). The rate of loss in Australia is low compared to many other countries but mangrove habitat is being cleared here for land development while changes in land use and catchment modification may alter hydrodynamic regimes thus threatening mangroves. Coastal morphology and geophysical processes are important environmental drivers of the structure and ecology of mangrove ecosystems (Thom 1982, Lugo et al. 1998). Even minor changes in hydrological regimes can lead to massive mortality of mangroves (Blasco et al. 1996). Climate change may negatively impact these important ecosystems, the major threats being changing sea-level and alteration of rainfall patterns especially a reduction in rainfall in conjunction with increasing temperatures.


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Temperature

Globally, mangrove distribution is generally constrained by the 20°C winter sea isotherm; there are a few exceptions, the more southerly distribution of mangroves in eastern Australia being one (Duke 1992). It has been suggested that this distribution is the result of small-scale extensions of warmer currents, such as the East Australian Current, or the southern populations are a relict representing refuges of more polewards distributions in the past (Duke 1992). As mangrove species show considerable variation in their sensitivity to temperature, species composition of mangrove forests will alter as temperatures rise and species distributions are expected to shift polewards (Field 1995). Temperature is also linked to coastal water and soil salinity through evaporation. Higher temperatures will increase evaporation thus increasing soil salinities and stress levels on mangroves (Twilley & Chen 1998; see Rainfall & Coastal Salinity below). Hence a reduction in rainfall or groundwater seepage coupled with increasing temperatures could alter the community composition of mangroves and lead to disappearance of some species.

Rainfall & Coastal Salinity

Salinity stress is a major factor determining the distribution of mangrove species along coastlines and estuaries. Rainfall directly influences the salinity of the intertidal waters and sediments but also influences salinity through freshwater runoff from the land and freshwater seepage into the soil (see Twilley & Chen 1998). Hydrology of mangroves is complex; tidal inundation, rainfall, groundwater seepage and evaporation all influence soil salinity and have a profound effect on mangrove growth. As salinity levels increase, mangroves are faced with increasing salt levels in the tissues (see Field 1995). Mangroves are considered highly susceptible to alteration in rainfall abundance or frequency. In Senegal, decreasing rainfall has led to marked changes in mangrove populations (Diop et al. 1997). Hydrology model simulations of mangrove systems in south-west Florida have demonstrated that mangroves in the upper intertidal are particularly sensitive to reductions in rainfall, even though these are areas with minor freshwater input (Twilley & Chen 1998). Animals and plants living in the upper intertidal are generally near the upper limits of environmental tolerance limits so small alterations in climate may impact this area more then the lower intertidal. In India, Rhizophora apiculate seedlings in the lower intertidal grew more rapidly than those in the upper intertidal presumably reflecting the additional stresses in the upper intertidal (Kathiresan et al. 1996). Precipitation patterns also control growth, flowering and fruiting patterns of mangroves; growth in some species may cease completely during dry periods (Imbert & Menard 1997, Coupland et al. 2005). For example, in southeast India, low salinity and high nutrients from land runoff induces rapid growth in mangroves after the monsoon season (Kathirisan et al. 1996).


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Currents and wind stress

Mangrove distributions may be constrained by current patterns, even if environmental conditions are favourable for growth and survival (Duke et al. 1998). Most mangroves are typically dispersed by water-buoyant propagules, which are viable for a number of days (Duke et al. 1998). Dispersal distance will reflect on the length of time that propagules remain viable and buoyant for, and this is expected to differ between species (Duke et al. 1998). For example, propagules of the mangrove Avicennia marina may only be able to establish successfully within the first 4-5 days of dropping (de Lange & de Lange 1994). It has been suggested the southern limit of this species in New Zealand may be controlled by limited transport by coastal drift and lack of suitable habitat within the dispersal range of existing populations, rather than by climatic factors (de Lange & de Lange 1994). It is hypothesised the southerly limit of mangroves in Australia is set by water movement through Bass Strait, which is primarily driven by east-west tidal currents (de Lange & de Lange 1994). Southward water movement by wind-induced drift is slow and is insufficient to transport the propagules to Tasmania within the 5 day period for viable establishment. Therefore even if temperatures in Tasmania become warm enough to support Avicennia populations (conventional wisdom is latitudinal range edges of mangroves are mainly determined by temperatures) they are unlikely to become established there.

Cyclones and storms

Cyclones in Australia and elsewhere have caused large scale devastation of mangroves (Woodroffe & Grime 1999). Loss of mangroves can result in coastal erosion, as occurred after Hurricane Andrew hit south Florida in 1992 (Davis 1995, Doyle et al. 1995). Storms surges will cause salinity intrusion into mangrove systems and lead to excess sedimentation in river flood events which can kill mangroves if disturbances are severe or prolonged. Differential mortality after storms can cause shifts in mangrove community composition (see Kathirisan & Bingham 2001). Sea level rise

Different mangrove species are adapted for different tidal inundation regimes as apparent in the zonation patterns of coastal mangroves. Rising sea-level will alter the tidal inundation regime experienced by mangroves and presumably increase environmental stress on individual plants. Laboratory experiments have shown increased tidal inundation reduced growth and photosynthetic rates in Rhizophora mangle seedlings (Ellison & Farnsworth 1997). Mangroves typically occur on low profile, low energy coastlines and are ecologically restricted to saline intertidal environments so are considered particularly susceptible to rapid changes in sea-level (Woodroffe 1992, Ellison 1993, Parkinson et al. 1994, Field 1995). When sea levels rise, as projected over the next century, shorelines will move landward but if sedimentation is more rapid then sea-level rise, shorelines may actually move seaward. Mangroves trap suspended sediments for example, a field study in a mangrove swamp in Cairns found 80% of suspended sediment bought in from coastal waters at spring flood tide was trapped in the mangroves resulting in a rise of the substrate of 1mm per year and presumably reducing turbidity in coastal waters (Furukawa et al. 1997). If sediment accretion rates in mangroves are equal to or exceed


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sea-level rise then mangroves will persist. The ability to accrete sediments will depend on the availability of suspended sediments in coastal waters so in areas where suspended sediment load is low, mangroves may not be able to track rising sea-levels (Ellison & Stoddart 1991, Parkinson et al. 1994). Mangroves growing in such areas, such as on carbonate settings or low islands may be strongly threatened by sea-level rise over the next century (Ellison & Stoddart 1991, Ellison 1993). Over the last few thousand years sea level has fluctuated continually, probably with dramatic consequences for many mangrove ecosystems depending on coastal characteristics and processes (Woodroffe 1992). 8000-6000 years ago, great mangrove swamps formed in northern Australia following a dramatic rise in sea-level which flooded low-lying land and created habitat for mangroves (Crowley 1996). These swamps subsequently filled in so reducing mangrove habitat but a gradual rise in sea level may again create conditions for the expansion of mangroves in northern Australia (Crowley 1996). Mangroves communities have responded to substantial fluctuations in sea-level in the past, however additional anthropogenic-induced pressures such coastal development and pollution may reduce the ability of mangroves to respond to present day climatic pressures. For example, coastal development will constrain shoreward progression of mangroves, resulting in a substantial loss of mangrove area, even if sediment accretion rates are equal to sea-level rise.

UV radiation

Increasing UV exposure will alter photosynthetic efficiency in mangroves depending on species, even minor increases in UV radiation may have significant effects (Detres et al. 2001). Moorthy & Kathiresan (1997) grew Rhizophora apiculate seedlings under increased UV-B regimes. They found net photosynthesis increased by 45% under a 10% increase in UV radiation but a 59% decrease in net photosynthesis occurred with a 40% increase in UV radiation.

Seawater chemistry and atmospheric CO2

Terrestrial vegetation is expected to flourish as atmospheric CO2 levels rise depending on the availability of nutrients and water. Evidence suggests that mangrove growth may also be stimulated as CO2 levels increase, seedlings of Rhizophora mangle grown under double ambient CO2 for a year exhibited increases in growth and photosynthetic rate (Farnsworth et al. 1996). The young plants also became reproductive an year earlier than in the field so elevated CO2 may accelerate maturation as well as growth (Farnsworth et al. 1996). However, the long term response of mature mangrove forests to elevated CO2 is unknown.


There appears to be little or no published observations explicitly on the observed impacts of climate change on Australian mangroves. Temperature, tidal inundation and rainfall are major factors controlling biogeographic distributions of mangroves globally. Mangroves are found mainly in the tropics and sub-tropics so as climate warms, mangroves in the southern hemisphere may penetrate further south. Changes to


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hydrological regimes will have profound impacts in mangroves. Reductions in rainfall will lead to the loss of species near upper limits of salt tolerance while changes in rainfall patterns will impact on the growth rates and reproduction. A general decrease in rainfall is projected for Australia over the next century. Sea-level rise is projected to be greatest on the east coast of Australia, which is also the most populated coast, so mangroves here may be under greatest threat. As responses to changing climate will be species specific, shifts in community composition are expected to occur. In south-east Australia mangroves are expanding as they migrate into saltmarsh areas (Saintilan & Williams 1999, Harty & Cheng 2003, Harty 2004, Rogers et al. 2006). For example, at Botany Bay, New South Wales, mangrove area increased by 32.8% overall between 1956 and 1996 while saltmarsh coverage decreased by 78.7% (Evans & Williams 2001). Although no single factor is responsible, it has been suggested increased rainfall in this area associated with climate change has reduced salinity levels within saltmarshes thereby allowing mangroves to migrate and outcompete saltmarsh plants (Harty & Cheng 2003, Harty 2004, Rogers et al. 2006). However, hydrodynamic modification related to urban and rural development is likely to be the overriding factor driving this mangrove expansion. Loss of mangroves will have immediate impact on the economic value of associated fisheries (Loneragan et al. 2005, Manson et al. 2005) as well as associated biodiversity. Further, mangroves commonly occur in conjunction with seagrass beds and tropical coral reefs and there are strong interactions between them. Coral reefs shelter the coastline from wave action while mangrove and seagrass beds filter coastal waters. Reductions in mangrove or seagrass cover may reduce resilience of neighbouring coral reefs to climate change.

4.5 Impacts and stressors other then climate change

The diverse fauna supported by mangroves is of direct and indirect economic and social value to human societies (Field 1995). Mangroves are under threat from coastal development because they occur in coastal and estuarine areas near centres of human settlement. Over 90% of Australia’s population lives within 120km of the coast. Deterioration of mangroves have been attributed to changes in hydrology bought about by human activities such as building of seawalls, foreshore development or dredging activities (Kathirisan & Bingham 2001) although in some areas mangroves may be favoured may such changes (Harty 2003, Harty & Cheng 2003). Mangrove abundance in New Zealand has increased over the past century attributed to changes in catchment land use increasing sedimentation rates (Ellis et al. 2004), however excessive sedimentation resulting from coastal building activities can smother mangroves leading to death (Ellison 1998). Coastal water enrichment through nutrient runoff may lead to expansion of mangroves. Many are nutrient limited so changes in nutrient levels will alter growth and production and shift community composition but severe enrichment can increase anoxia of coastal waters and pollute mangroves. A severe and widespread dieback of mangroves in the Mackay region, Queensland, since the 1990s has been linked to herbicides from agricultural runoff (Duke et al. 2005). Reclamation of coastal land for development and diversion of freshwater are major threat to mangroves (Twilley et al. 1998). Further, human activity hundreds of kilometres inland can change ground water flow resulting in


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the loss of vast mangrove areas (Tack & Polk 1997 in Kathiresan & Bingham 2001). Such activities will also exacerbate climatic change impacts on mangroves.

Main Findings: Mangroves • Expected changes

o Productivity of mangroves may increase o Sea level rise will reduce mangrove area depending on locality o Warming will drive species distributions southwards o Timing of flowering may shift o Damage or destruction by storm events may increase

• Observed changes within the Australian region o None as yet. Recent declines attributed to anthropogenic impacts

although some changes may be related to climate


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5. IMPACTS OF CLIMATE CHANGE ON KELP

Thomas A. Okey, Nadia Engstrom, & Russ Babcock CSIRO Marine and Atmospheric Research, [email protected]

5.1 Overview

Kelp is the common name for large brown subtidal macroalgae that form surface or mid-water canopies over hard substrates in temperate regions. Kelp species comprise a small proportion of marine macroalgae, but they are foundation organisms in their habitats, and are thought of as bioindicators of the integrity of these temperate reef ecosystems. They can form large floating canopies up to 30 metres above the sea floor in the form of ‘kelp forests’. This habitat architecture is inhabited by diverse assemblages of animals and plants, and the primary production of kelp is also utilised by a broad community of organisms. Kelp species are distributed throughout the southern half of Australia wherever persistent hard substrata receives enough light. The distributions of some individual kelp species are limited to Tasmania or other particular areas such as the South Eastern and South Western Large Marine Domains. The distributions of kelp species in Australia and elsewhere are shaped by temperature, nutrients, turbidity, light penetration and quality, wave action and sand scour, and the relative prevalence of herbivores and their predators. Large declines of giant kelp and other macroalgae along the east coast of Tasmania in the last 50 years have been attributed to rising sea temperatures, but deforestation, agriculture, urban development, and other land-use changes might have changed sedimentation and runoff characteristics decreasing the quality of kelp habitat in some settings. The removal of predators through fishing or the invasion of new herbivores, or both, might lead to kelp forest degradation as it has elsewhere throughout the world. One known example of herbivore invasion of Tasmanian Kelp forests is likely the result of climate change.

5.2 Introduction

The hard substrata of Australia’s temperate rocky intertidal and subtidal habitats are distributed unevenly across the coastline of Australia’s southern latitudes and are inhabited by a diverse and highly endemic macroalgal flora. Macroalgae are multicellular, photosynthetic marine algae of macroscopic size, which have complex life cycles (e.g. Abbott & Hollenberg 1976). Kelp is the common name for large ‘corticated’ macroalgal species of generally temperate distributions and which have quite differentiated and specialized structures for attachment, light collection, materials transport, food storage, and reproduction. Kelps are classified as ‘brown’ algae (Phylum: Phaeophyta) of the order Laminariales, whereas a large proportion of marine macroalgae are less conspicuous species classed as ‘red’ (Phylum: Rhodophyta) or

IMPACTS OF CLIMATE CHANGE ON KELP

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‘green’ (Phylum: Chlorophyta) with less differentiated structures. We focus on kelp here for four reasons: (1) they are well studied relative to most other macroalgae; (2) they are ‘foundation’ species in subtidal temperate rocky reef ecosystems; (3) they are sensitive to temperature and other associated variables; and (4) they are conspicuous. Australia’s surface-canopy-forming kelp species include giant kelp (Macrocystis pyrifera) and other large laminarian species (e.g. Macrocystis angustifolia and Ecklonia radiata) making these antepodean kelp forests generally resemble kelp forests found in other parts of the world. Floating canopies of kelp plants can form at the water’s surface up to 30 metres above the sea floor and over intermediate canopies of other kelp and macroalgal species and sessile invertebrate communities (as described by Dayton 1985). The Australian distributions of five canopy forming kelp species (Figure 5-1) illustrate that some kelp species require specialised conditions and that certain species might be vulnerable to climate change due to their limited ranges and the ‘poleward boundedness’ of their habitats.

Figure 5-1: The distribution of the kelp species Macrocystis pyrifera (light blue), Macrocystis angustifolia (blue green), Ecklonia radiate (purple), and two other conspicuous macroalgae species Durvillaea potatorum (pink) and Acrocarpia robusta (light green) along the Australian coastline, from Sanderson (1997).

The ecology and dynamics of Australia’s kelp and macroalgal communities have been relatively poorly known (see Underwood & Kennelly 1990, Kennelly & Underwood 1992), though kelp forest research has been gaining momentum in Australia (e.g. Hatcher et al. 1987, Kennelly 1987a, b, Cheshire & Hallam 1988a, b, 1989, Andrew 1993, Kendrick et al. 1999, O'Hara 2001, Fowler-Walker & Connell 2002, Wernberg et al. 2003, Valentine & Johnson 2004). As in other parts of the world, Australian kelp facilitates and continually produces large amounts of coastal fish and invertebrates because it provides extensive three dimensional structures with a variety of micro-habitats, refugia, attachment


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opportunities and food. Kelp forests produce very large amounts of phytal detritus (Gerard 1976) that fuels high secondary production of invertebrates which in turn support higher trophic levels (e.g. Duggins et al. 1989). This fuelling of marine food webs occurs both within kelp forests and at remote distances and depths (e.g. Bustamante et al. 1995, Vetter & Dayton 1999, Okey 2003) through the oceanographic transport of this phytal detritus. For these and other reasons, Australian kelp supports a multitude of recreational, commercial, and scientific activities both directly and indirectly (Kingsford et al. 1991). Australia’s temperate macroalgal flora exhibits unusually high endemism and diversity due to isolation from other continents and limited connectivity amongst the suitable habitats of Australia’s southern latitudes (Womersley 1990, Phillips 2001). This diversity explanation is illustrated by Womersley’s (1990) finding that only 13% of Australia’s temperate macroalgal species were common to all of Australia’s temperate (southern) regions. Southern Australia contains by far the highest diversity of brown algae (140 species per 100 km section) of all the rich kelp areas of the world—California, southwestern Africa, and north-central Chile (Bolton 1996), but this number vastly underestimates the important relationship of these macroalgae to Australia’s marine biodiversity. The important point about these kelp is that their presence facilitates a very broad biodiversity of invertebrates, fishes, and higher vertebrates (Keough & Butler 1995, Steneck et al. in press). This has been noted of kelp forests throughout the world for decades by our colleagues (Bernstein & Jung 1980, Duggins 1980, Reed & Foster 1984, Dunton & Schell 1987), but it has likely been recognized for many millennia (Simenstad et al. 1978). The degradation or disappearance of particular macroalgal species or assemblages might cause a dis-proportionately large simplification of Australia’s reef ecosystems and patterns of coastal production, thus causing a cascading simplification of these systems as these animal species disappear, as observed elsewhere (Bodkin 1988, Dayton et al. 1998). Degradation, or even extirpation/extinction, of macroalgae is plausible in Australia because many of Australia’s temperate macroalgal species are geographically ‘bounded’ in the sense that these species exist along the extreme southern reaches of Australia’s hard substrata (Figure 5-1), precluding adaptive migration to the south. Kelp forest ecosystems are generally vulnerable to potential climate changes because they are sensitive to (1) increased temperatures and associated decreases in nutrients, (2) increased turbidity and resulting decreases in light penetration, and (3) outbreaks of herbivores due to depletions of predators (e.g. resulting from the combined effects of fishing and climate change), or due to invasions of herbivores (e.g. resulting from climate-related range shifts or human-facilitated transport). The current understanding of the sensitivities of macroalgae, and particularly kelp, to these factors is reviewed in the following sections with particular reference to Australian kelp communities.


Sea temperature / Nutrients It is well known in other regions of the world that many marine algae, e.g. kelp and kelp communities, are sensitive to sea temperature increases and associated nutrient decreases (e.g. Gerard 1982, Foster & Schiel 1985, Deysher & Dean 1986, Gerard 1997, Dayton et al. 1998, 1999, Adey & Steneck 2001), and this is also recognised in


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Australia (Bolton & Anderson 1987, Hatcher et al. 1987). Declines of giant kelp forest communities in Tasmania have been reported to be associated with changing oceanographic conditions (Edyvane 2003, Edgar et al. 2005) as in California (e.g. Dayton & Tegner 1984). Along coastal California, the El Niño phase of ENSO (El Niño-Southern Oscillation) leads to elevated sea surface temperatures, reduced upwelling, a stratified water column, and low nutrient concentrations causing considerable net decreases in forest canopy cover. Increases in the frequency and magnitude of El Nino events can thus lead to shifts in the kelp forest community there. In Australia, the La Niña phase of the ENSO weakens the zonal westerly winds, thus weakening the coastal upwelling and impingement of nutrient-rich, subantarctic waters along Tasmania’s east coast (Harris et al. 1987, Harris et al. 1988). These changes combine with the strengthening of the warmer and nutrient poorer East Australia Current (also related to changes in wind patterns), which apparently results from the effects of stratospheric ozone depletion (Cai et al. 2005, Cai 2006), to produce an even more severe impact on kelp and associated species. Ozone depletion causes a north-south temperature gradient change in the stratosphere, with a greater cooling toward the pole than at higher latitudes. This causes a change in stratospheric circulation whereby a decreasing mean sea level atmospheric pressure at high latitudes causes increased rising air masses, with the opposite occurring in mid-latitudes. This decreases westerly winds in the mid-latitudes while increasing them in the high latitudes, and this increases the strength of the East Australia Current, bringing it and warmer conditions further southwards into the Tasman sea (W. Cai, CSIRO Marine and Atmospheric Research, personal communication, 16 September 2006). The strength of the East Australia Current has increased some 20% since the late 1970s (Cai 2006). Given what is known about kelp forests, the resulting habitat changes—increased temperatures and decreased nutrients—would be expected to stress kelp species physiologically where they currently exist, initiate changes in the phenology of kelp life cycles, and lead to declines in Australia’s southern bounded kelp forests. The survival of kelp plants during the 1983 El Niño event along the coast of California was greater when positioned in the path of nutrient pulses, indicating that nutrients play an important role in thermally-induced mortality (Zimmerman & Robertson 1985). Indeed, nutrients rather than temperature have been long suspected as being the key variable to kelp productivity (Jackson 1977, Dayton 1985). But temperature and nutrients are connected because colder water can hold more nutrients, and also because nutrients are depleted when summer conditions or El Niño events stratify the water column; i.e. they prevent mixing of the warm surface waters with the cooler nutrient-rich waters below. Although some areas of the Australian coastline obtain nutrients from impingement of sub-Antarctic waters onto shelves and along coasts, Australian coastal marine ecosystems are nutrient poor in general due to limited upwelling, ancient geologies, and the tropical origins of the predominant coastal currents. Furthermore, temperate Australia receives relatively little rain, and evaporation often exceeds precipitation, thereby limiting runoff. El Niño events tend to reduce rainfall thereby further limiting nutrients in coastal waters. California kelp forests have experienced considerable impacts on different temporal and spatial scales in connection with El Niño events, but the mechanisms of impact were multiple, involving catastrophic storms, overfishing of predators, trophic cascades, and


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disease, making it difficult to attribute impacts to just temperature and associated decreases in nutrients, and making the story of degradation more complex and interesting (Ebeling et al. 1985, Tegner & Dayton 1991, Tegner et al. 1996, 1997, Tegner & Dayton 2000). Observed declines of Australia’s kelp forests have likewise resulted from multiple factors, and considerably more research is needed to distinguish the causes and tell the story. The dispersal potential of kelp rafts and hence connectivity and maintenance of kelp forest meta-populations does appear to be related to sea surface temperature. Hobday (2000) found that rafts float well and last long during cool months, and they degrade quickly when temperatures exceed 20ºC. Theil & Gutow (2005) showed that kelp rafts were indeed less common in warmer water. These findings imply that increasing sea surface temperatures might markedly decrease dispersal distances, thus leading to greater population fragmentation and reduced subsidy of remote systems. The resistance of macroalgae to environmental change is highly variable among species. During the strong El Niño event in 1983 on the coast of California, both the adults and new recruits of the giant kelp Macrocystis pyrifera were seriously affected by high sea temperatures and low nutrient levels, while associated understory kelp species experienced very little impact (Dayton 1985). The giant kelp Macrocystis is particularly sensitive to rising temperatures and reduced nutrients, partly due to its position in the water column, and partly due to its limited nitrate storage capacity (Gerard 1982). If current temperature trends continue in Australia’s shallow marine ecosystems, macroalgal assemblages might fundamentally shift from canopy dominated assemblages to understory dominated assemblages, or in some other fundamental way through influencing competitive interactions, as implied by the related work of Kennelly (1987a, 1987b) in Australia and Dayton et al. (1999) in California. Mixed layer depth In conditions of temperature and nutrient stress, kelp forest mortality often occurs in the surface mixed layer above the thermocline, whereas the flora and fauna beneath the thermocline can escape mortality, even if on the same plant (Dayton et al. 1992). Any deepening of the mixed layer in these critical shallow areas might have a large effect on the macroalgal communities of temperate Australia.

Rainfall/Coastal runoff/Salinity/Irradiance Light penetration is a main factor limiting the distribution of marine macroalgae, both in Australia (Kennelly 1989) and elsewhere (Reed & Foster 1984, Deysher & Dean 1986, Dayton et al. 1999, Spalding et al. 2003). Increases in turbidity associated with increases in rainfall or coastal development, or other human activities, would thus be expected to degrade kelp communities by generally decreasing the light penetration and depth ranges of macroalgal species (e.g. Vadas & Steneck 1988, also see review in Okey et al. 2004). Keough & Butler (1995) suggested, based on their experience in Australia, that increases in turbidity can give competitive advantage to shade-tolerant flora and non-photosynthetic organisms and that it can have other disproportionate impacts on kelp forest organisms, though they also pointed out the lack of studies that have distinguished this factor from others. Edyvane (2003) suggested that the observed decline in Tasmanian kelp forests could be partly explained by increased sedimentation and subsequent reduction in available substrate.


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Wind and currents Southward moving currents such as the East Australia Current and Zeehan Current (a continuation of the western Leeuwin current) interact with southern waters to create creating complex ocean structure along Australia’s east and west coasts and further offshore thus creating a variety of coastal and oceanic habitats, micro-habitats, and interfaces of temperature and production (e.g. Harris et al. 1987). Changes in the oceanographic conditions that create these features would undoubtedly lead to changes in sea surface temperatures, nutrient levels, and productivity regimes, thereby influencing kelp forest communities. Changes in these features should indeed be expected in the near future because stratospheric ozone depletion appears to be causing a change in oceanic wind patterns that is already strengthening the East Australian Current measurably and driving more warm water into the southern cool waters of Australia (Cai et al. 2005, Cai 2006). These changes might well degrade the remaining refuge habitat of species such as giant kelp and its associated diverse assemblage (e.g. Edyvane 2003).

Disturbance regimes Storm regimes help shape kelp forest ecosystems (e.g. Reed & Foster 1984, Dayton et al. 1992, Graham 1997) and their high diversity is undoubtedly promoted by storm disturbances, but unusually intense storms associated with El Niño are known to have severely impacted California’s kelp forest ecosystems (Dayton & Tegner 1984, Ebeling et al. 1985, Tegner & Dayton 1987, 1991, Tegner et al. 1997). Recovery after these storms can be relatively quick, but there is concern that prolonged increases of the frequency or magnitude of these storms and conditions would cause a net shift in community structure. In Australia, Kay & Butler (1983) found disturbance associated with storm waves to be important in structuring hard substrate habitats. Kennelly (1987a, b) found that recolonisation or growth of Australian sub-canopy kelp forest algae increased after disturbances removed canopy algae. Some of these understory algae might inhibit the recovery of the canopy algae, as found by Dayton & Tegner (1984). Disturbances do indeed shift kelp forest assemblages in Australia, and they might well maintain diversity as noted in southern Australia by G. Jones (see Steneck et al. in press), but a shift from a ‘natural’ disturbance regime (to which the kelp and other organisms are adapted) to one with more frequent or intense storms would affect Australia’s kelp forests adversely.

Ultraviolet radiation Ultraviolet (UV) radiation has been shown to inhibit photosynthesis in Antarctic marine macroalgae (Bischof et al. 1998) and in temperate California kelp forests (Graham 1996) indicating that current increases in UV-B radiation in the northern and southern hemispheres is likely to affect the fitness and available depth distribution of marine macroalgae. The former authors found that the photosynthesis of brown algae (Phaeophyta) was inhibited to an intermediate extent, implying that surface canopy-forming kelp might be impacted strongly, and the latter author found that gametophytes and embryonic sporophytes of Macrocystis pyrifera did not survive in the high light levels found in shallow water. UV-B radiation has increased in the southern oceans and Australia (e.g. Cai 2006), and so such effects should be expected to have manifested in Australia’s kelp forests, though recent evidence suggests that the ozone hole has now begun to decrease (de Jager et al. 2005) perhaps leading to amelioration of these effects.


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Acidification Decreasing ocean pH resulting from increased concentrations of atmospheric CO2 is expected to have profound effects on corals and all other organisms that rely on CaCO2 structures (Feely et al. 2004, Orr et al. 2005, Guinotte et al. 2006). Some green and many red marine macroalgae incorporate calcium carbonate in their tissues or as skeletal structures. These crustose or articulated forms of ‘coralline’ algae are common and play a prominent ecological role in both temperate reef ecosystems and tropical coral reef ecosystems (see the coral reefs section of this report). The projected acidification is expected to adversely affect this component by decreasing its fitness, competitive abilities, grazing resistance, and abundance. This will certainly have cascading effects in kelp forests and temperate rocky reef ecosystems.

Sea level Changes in sea level may considerably change the amount and distribution of available substrata for kelp and other macroalgae species to attach because they inhabit a very narrow vertical zone where light reaches appropriate hard substrata. Such major changes have occurred during the last 18,500 years causing large-scale ecological shifts in southern California from a highly productive rocky-reef coastline to one dominated by soft sediment (Graham et al. 2003). Steneck et al. (in press) suggest that an expected 20 cm sea level rise by 2050 would decrease kelp depth distribution by increasing coastal erosion and turbidity. Given current suspicions about the role of turbidity in the demise of Australian kelp forests, and given the pace of changes in coastal watershed function, Australian kelp forests would be vulnerable to such impacts.

Species interactions A general pattern that has emerged from studies of kelp de-forestation around the world is that the reduction or removal of predators causes population explosions of herbivores, usually echinoids (sea urchins), that de-foliate kelp forest ecosystems, often causing them to collapse or shift to a degraded state (e.g. Simenstad et al. 1978, Tegner & Dayton 1991, Estes & Duggins 1995, Tegner & Dayton 2000, Steneck et al. 2002, Estes et al. 2004, Steneck et al. in press). Sea urchins are also important in structuring Australian kelp forest communities, at least along eastern Australian coasts and where urchins can find refuge from predators (Andrew 1993, Fowler-Walker & Connell 2002), or where predators are reduced (e.g. Valentine and Johnson 2005). Indeed, Edgar and Barrett (1999) suggested that Tasmanian coastal reefs have been heavily overfished by a combination of commercial and recreational fisheries. A variety of gear types such as commercial and recreational gillnetting, commercial fish traps, hook and line, and recreational angling and spearing have lead to declines in a variety of Tasmanian kelp forest fish species. Abundances of some have declined (e.g. Bastard Trumpeter, Blue Warehou, Purple Wrasse, Blue Throated wrasse, Banded Morwong, and some leatherjackets, and lobsters), but much of the conspicuous effects has been shifts in size and age at maturity (Ziegler et al. 2006). The presence of a predator fauna is of primary importance in kelp forest ecosystems, as illustrated by findings by Halpern et al. (2006) that top down forces explained up to ten times more of the variance in abundance of bottom and mid-trophic levels than does bottom-up control in Southern California kelp forest. This is consistent with the findings of Shears & Babcock (2002) in New Zealand kelp forests. Indeed, Edgar and Barrett (1999) found not only that Tasmanian reef ecosystems are overfished but that marine protected areas in these habitats enable predator assemblages to recover, thereby controlling herbivore populations and enabling recovery of foundational kelp


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assemblages which in turn provides habitat for a broader biodiversity (see also Edgar et al. 2005). This evidence all bolsters the notion that effective management of fisheries (commercial and otherwise) would increase the resiliency (reduce the vulnerability) of these systems to climate change. In addition to predator depletions through overexploitation (e.g. Dayton et al. 1998), species introductions or invasions can also cause kelp forest degradation. Given that the ranges of shallow temperate marine species are shifting polewards along coastlines (Barry et al. 1995, Sagarin et al. 1999), major changes in species interactions and compositions are expected in Australia’s kelp forests, exacerbating the problem of poleward boundedness discussed previously.


Physiological stress, abundance, and distribution Giant kelp canopy losses have likely been greater than 50% in Tasmanian waters over the last 50 years, and it is more than interesting that the annual (Jun-Nov) mean sea temperatures at Maria Island in coastal Tasmania increased by approximately 1.0-1.5 ºC between1944 and 1999 (Edyvane 2003). It seems likely that this dramatic decrease in abundance and cover reflects direct physiological stress caused by changing oceanographic conditions associated with increased temperatures and decreased nutrients (Edyvane 2003, Edgar et al. 2005), but it is also plausible that other factors such as changes in rainfall, coastal runoff, upland erosion (associated with land changes), and sedimentation might have contributed to this observed decline in Tasmania (Edyvane 2003). Still another plausible explanation for the demise of Tasmanian kelp forests is invasive herbivores and algae. The recent invasions of the sea urchin sea urchin Centrostephanus rodgersii in Tasmanian kelp forests has caused extensive deforestations (e.g. Johnson et al. 2005), and some of this opened space is now occupied by an invasive Japanese kelp Undaria pinnatifida (Valentine & Johnson 2004). These southward range expansions might well have been initiated by the strengthening of the East Australian Current (Edyvane 2003), and this might represent a quasi-permanent and fundamental shift in these Tasmanian kelp forests, with more changes to come. Phenology It is likely that changes in local oceanographic conditions will cause shifts in phenology—the timing of reproductive stages and life cycles. However, very little work has been conducted on the phenology of Australian kelp, except for on invasive kelp species (Schaffelke et al. 2005). This is a promising area of climate impacts research for Australia.

Predicted consequences for biological communities and biodiversity Kelp forests are foundation species in marine ecosystems and there is compelling evidence that their degradation or removal leads to the collapse of whole functioning and highly productive communities (e.g. Bodkin 1988, Dayton et al. 1998, Pinnegar et al. 2000), which subsidize adjacent ecosystems and communities in a variety of ways. The degradation of kelp forests thus has far-reaching ecological effects. The fact that they are sensitive to oceanographic fluctuations and changes (Dayton et al. 1998), relatively easy to measure, and of high ecological importance makes Australian kelp forests an excellent candidate for climate impacts studies as an indicator system.


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Effects of fishing It is well established that overfishing can have profound impacts on kelp forests due to the cascading effects of removing predators that leads to outbreaks of herbivores (Dayton et al. 1998, Tegner & Dayton 2000, Jackson 2001, Jackson et al. 2001, Steneck et al. 2002). Whole suites of organisms can vanish once deforestation occurs (Bodkin, 1988), and this can lead to alternate states that are persistent or stable. For example, reductions of predators cause outbreaks of herbivores that consume all the existing kelp and prevent new kelp from recruiting and establishing, thereby producing ‘barrens’ where kelp forests used to stand. Because fishing is a strong shaper of marine ecosystems, including kelp forests, accounting for fishing effects is necessary when undertaking studies of climate impacts. Distinguishing these effects is challenging, but it would be a very profitable field of study not just for understanding climate impacts, but also for understanding how these marine ecosystems are shaped, structured, and regulated in general.

Coastal pollution Kelp forests are exposed, and vulnerable, to various types of pollution due to their position along coastal fringes. Runoff, sedimentation, and turbidity, were discussed previously, but other types include thermal and chemical. There are a few published studies on pollution and Australian kelp or kelp forest communities (Smith & Simpson 1993, Roberts 1996, Roberts et al. 1998, Burridge & Bidwell 2002). Since pollution is also an important shaper of kelp forest communities, climate impacts studies should attempt to account for these effects as well.

Main Findings: Kelp • Expected changes

o Pronounced increases in SST will drive species distributions southwards increasing the risk of extirpation and extinction

o Reductions of kelp forests can lead to degradation and collapse of a diverse biodiversity that depend on kelp species

o Non-climate related stressors such as fisheries and pollution and coastal watershed degradation are also important in structuring kelp forest ecosystems and so successful management of these stressors can increase the resiliency of Australian kelp forest ecosystems to climate change impacts

• Observed changes within the Australian region

o Recent declines in macroalgae along east coast of Tasmania are largely attributed to ocean warming, though this effect likely combines with other coastal stressors

o Kelp forests in Tasmania and elsewhere are overfished and impacted by coastal modifications to water quality related to watershed modification and human activities

o Marine protected areas in Tasmania have been shown to be effective in restoring the ecological integrity of kelp forest communities thereby increasing their resilience to climate change

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6. IMPACTS OF CLIMATE CHANGE ON ROCKY SHORES

Elvira S. Poloczanska & Russ Babcock CSIRO Marine and Atmospheric Research, [email protected]

6.1 Overview

Rocky shores are one of the most easily accessible marine habitats and are a transition zone between the land and sea. The rocky shorelines of southern Australia contain very high numbers of endemic species; it has been estimated that over 90% of mollusc and echinoderm species and 60% of seaweeds may be endemic. Many animals and plants in the rocky intertidal such as barnacles, limpets and seaweeds, are sessile as adults and can be quickly and easily surveyed using non-destructive methods. The rocky intertidal has a long history as a test area for important concepts in ecology and a good knowledge of processes and species biology exists for many Northern Hemisphere shores and for limited parts of the Australian coastline. Some of the best and most numerous examples globally of marine range shifts with warming temperatures have come from rocky shore surveys. Further, monitoring programmes for climate change impacts have been initiated in the UK, USA and other countries. The rocky intertidal provides an ideal sentinel system to monitor climate change in a cost effective manner.

6.2 Introduction

Australia crosses temperate and tropical regions, and the intertidal areas are very different within them. The tropical coasts tend to be sand or mangrove dominated with rocky outcrops while temperate regions have more extensive rocky shores. Much of rocky shore fauna and flora of southern Australia is endemic (Bolton 1996, Poore 2001, Phillips 2001, Griffiths 2003). The high endemism in this area is partly the result of low dispersal abilities of species and the presence of ecological barriers to dispersal along the southern coastal waters such as a sharp temperature gradient near the cessation of the Leeuwin Current and the absence of nearshore rocky reefs in the centre of the Great Australian Bight and at other locations along the southern Australian coastline. This review will mainly focus on Australia’s temperate rocky shores. Rocky shore organisms are strongly influenced two processes - tidal regime and exposure or wave action. As intertidal habitats are transitional between the terrestrial and marine realms, sessile animals and plants growing on the shore are periodically immersed in seawater and then exposed to air as the tides rise and fall. Position on the shore (low, mid or high) will determine the length of time a sessile organism is covered by seawater, organisms living on the very top of the shore may only be immersed for short periods and even then only during the largest tides. Eastern and western Australia have two tides a day, with tidal range ranging from less then 1 m in south-western


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Australia to around 1-2 m in south-eastern Australia with the largest tides of over 7 m occurring in north-western Australia. Tides along the southern coast of Australia are more complex, in some parts there is only one tide per day and extended periods do occur where tidal level does not change. Air temperatures fluctuate more widely then sea temperatures on a daily basis and there may be considerable difference between sea temperature and air temperature. As intertidal organisms are marine in origin, there is therefore a gradient in environmental stresses going from the bottom of the shore to the top which is evident in the banding of plants and animals on the shore. For example, animals with a high tolerance to desiccation stress, such as the barnacle Chthamalus antennatus, are found on the high shore. Tidal influence determines the position of animals and macroalgae on the shore but tidal effects are modified by factors such as wave force, humidity, substrate topography and biotic interactions. Wave action exerts strong mechanical stresses on intertidal organisms and is an important force shaping of rocky shore communities. Wave action varies greatly along the shore depending on factors such as topography of the coastline, fetch (length of open water over which the wind blows) and bathymetry. Large swell waves, generated in the Southern Ocean, pound much of the southern coast of Australia. Wave action can extend the vertical distribution of organisms on the shore. Macroalgae are an obvious component of rocky shores, these range from encrusting species which are found on the high shore, to foliose algae which are most abundant on the lower shore, to the canopy forming species which generally occur at the very bottom of the shore. Many rocky shore animals are sessile with the most conspicuous being barnacles, mussels, oysters, chitons, sea anemones and tubeworms. Shores which have strong wave action, that is very exposed shores, tend to be dominated by the surf barnacles. Sheltered shores tend to have extensive cover of bivalves such as oysters, especially on tropical shores. Most of these sessile species have a larval dispersive stage. Mobile macrofauna such as starfish, sea urchins, crabs and marine snails range over the intertidal when either exposed or submerged while fish will move in with the tide to feed on the animals and plants in the intertidal zone. Some fish are intertidal, taking refuge in rockpools when the tide retreats. There is a long history of study of the intertidal, due to the ease of access, the conspicuousness and ease of identification of fauna and flora and pronounced physical gradients found on shores. Early seminal ecological experiments have been conducted in the intertidal and intertidal populations and they have been used to test and explore ecological hypotheses (e.g. Connell 1961, Roughgarden et al. 1985; Burrows & Hawkins 1998, Johnson 2000). Impacts of climate change have already been recorded on rocky shores in the UK, Ireland, USA and Chile where historical records were available (Barry et al. 1995, Sagarin et al. 1999, Rivadeneira & Fernandez 2005, Mieszkowska et al. 2005). Changes in intertidal populations may provide a valuable warning system for comparable changes in demersal and pelagic species in coastal waters (Southward 1991). Monitoring programmes are being established overseas using rocky shore organisms as sentinels of climate change (Mieszkowska et al. 2005). Studies in south-east Australia have already identified molluscs as reliable indicator species of rocky shore biodiversity (Gladstone 2002, Smith 2005).


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Temperature As most rocky shore species are restricted by thermal tolerances, both to air and water temperature, temperature is a major factor influencing large scale distributions of rocky shore organisms (e.g. Hiscock et al. 2004). Large die-backs of flora and fauna are frequently observed in the intertidal when low tides coincide with the middle of the day during summer (e.g. Tsuchiya 1983, Perez et al. 2000). Extreme temperatures, both high and low, can lead to irreparable damage and death of coastal organisms if severe or prolonged (Wethey 1985, Davenport & Davenport 2005). An unusual dieback of the shallow sublittoral brown alga Phyllospora comosa along the east coast of Tasmania in 2001 was attributed to above-average sea water temperatures coupled with nutrient stress (Valentine & Johnson 2004). At sustained extreme temperatures, photosynthetic inhibition occurs in marine plants which may lead to irreparable damage and death (Bruhn & Gerard 1996, Campbell et al. 2006). Much of southern Australia rocky shore biodiversity is characterised by warm-temperate species with cold-temperate species such as the giant brown algae Macrocyctis and Durvillaea found around Tasmania and Victoria. As temperatures increase, a shift in species distributions towards the poles is expected, such shifts over the past 50 years have already been documented from rocky shores outside Australia (Barry et al. 1995, Sagarin et al. 1999, Rivadeneira & Fernandez 2005, Mieszkowska et al. 2005). For example, recent annual monitoring in the UK has found evidence of on-going range extensions in warm water species and bioclimatic modelling predicts these range shifts will continue in the future (Mieszkowska et al. 2005). Even a small change in temperature for some species may lead to major shifts in distributions. Similar responses are expected or may be occurring on Australian rocky shores. The high levels of endemism along Australia’s southern coastline could increase vulnerability to temperature increases compared to temperate rocky shore elsewhere, many endemic species may have more stringent temperature limits and so may be particularly susceptible to warming (Beardall et al. 1998). Rainfall & Costal Salinity Changes in rainfall patterns, and associated changes in watershed geomorphological dynamics, will affect the dynamics of coastal marine ecosystems through the physiological effects of large-scale flooding, fluctuations in salinity, and increases in turbidity and nutrients on resident organisms such as intertidal fauna. Currents and Wind Stress Hydrodynamic stress will affect growth forms and morphological adaptations of plants and animals (e.g. Denny & Gaylord 1996) as well as community diversity and structure (e.g. Coates 1998, Benkendorff & Davis 2004). An increase or decrease in the size of waves and the frequency of storms with climate change may result in significant changes in population abundances of rocky shore organisms and a shift in community composition in areas that are affected. For example, the large fucoid alga Carpophyllum flexuosum is characteristic of calm conditions and is now common in areas of north eastern New Zealand where it was once virtually absent (Cole et al. 2001). This range


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expansion coincides with a significant decrease in storm frequency and intensity in this part of New Zealand over the past 30 years related to decadal scale climate variation (de Lange & Gibb 2000). Cyclones and Storms In coastal systems, extreme wind events such as storms and cyclones are important natural disturbance regimes (see Currents and Wind Stress above). Severe storms can remove large areas of intertidal and subtidal fauna and macroalgae (e.g. Thomsen et al. 2004). Intense natural disturbance regimes also act as an important selective pressure on coastal organisms as seen in a rapid shift in the shell morphology of intertidal gastropod populations after a severe storm (Trussell 1997). Sea Level Rise Sea-level rise will also alter the magnitude of local tidal ranges, depending on interactions with coastal topography. An increase or decrease in tidal range will directly impact algae and fauna in the intertidal. UV radiation UV-B radiation is known to have inhibitory effects on photosynthetic performance and can damage DNA. Intertidal plants and animals are exposed to high levels of solar radiation so have developed a range of photoprotection mechanisms, depending on species. However, species less tolerant to UV radiation are expected to decrease in abundance or disappear from intertidal if UV radiation levels rise. Early lifestages may be may be more susceptible to UV radiation than mature plants, thus regulating distributions of adults (Graham 1996, Rijstenbil et al. 2000, Swanson & Druel 2000, Cordi et al. 2001). Similarly, egg and larval stages of many marine animals are susceptible to UV damage, particularly when stressed by warmer temperatures or increased salinity (Lesser et al. 2003, Przeslawski et al. 2004, 2005, Bonaventura et al. 2006). pH and seawater chemistry

Macroalgae take up primarily bicarbonate ions for photosynthesis, as only a small proportional change in bicarbonate concentration will occur as atmospheric CO2 levels rise, little enhancement of growth is expected (Beardall et al. 1998). However, increased acidification of the oceans may have severe consequences for coralline algae (Gao et al. 1993) therefore enhancing competitive advantages of non-calcifying species over calcifying species (Gao et al. 1993, Beardall et al. 1998). Animals which produce carbonate shells will be at risk, acidification may be expected to increase physiological stress on marine animals. Metabolic efficiency and growth rates of bivalves and other molluscs will be impaired (Michaelidis et al. 2005, Berge et al. 2006). Experiments have also shown that under lowered pH conditions the fertilization rate of eggs of intertidal echinoderms declined and larvae were severely malformed (Kurihara et al. 2004).


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Abundance/Distribution There appears to be no published observations explicitly on the observed impacts of climate change on Australian rocky shores. Given the mounting evidence from outside Australia that distributions of rocky shore species are shifting, it is expected that similar shifts will occur, or are occurring, on Australian coastlines. Tropical species may extend further south, particularly given the projected strengthening of the East Australian Current on the east coast. Temperate species will retreat towards southwards. Cold-temperate species around Tasmania and Victoria may be at highest risk given the lack of suitable habitat further south. The southern coastline of Australia also contains many endemic species with restricted ranges and these may similarly be threatened as climate changes. However, local microclimate will temper large-scale climate impacts on organisms. The climate experienced by an sessile intertidal organism when exposed at low tide is the result of multiple, interacting climate factors such as temperature, wind speed, cloud cover, etc. – creating a unique ‘thermal signal’ at each site (Helmuth 2002). The complexity of coastlines and the influence of tides and weather create a complex mosaic of thermal environments along a coastline, so providing refuges from warming temperatures (Helmuth et al. 2002). Given this, climate change may not lead to a ‘smooth’ poleward shift in intertidal organisms but instead result in localised extinctions at a series of ‘hotspots’ (Helmuth et al. 2002). Consequences for biological communities and diversity Several key functional or structural species, such as the canopy forming macroalgae, are likely to be affected by climate change. In some cases, loss of these key species may lead to the disappearance of the associated community. Shifts in distributions will alter biotic interactions within communities leading to alteration in species dominance. Alteration of rocky shore communities will impact coastal ecosystems. For example, many temperate rocky shore fauna, such as barnacles, release large numbers of larvae which are important components of coastal plankton. Loss or alteration of such species will also have consequences for larval and juvenile stages of fish that feed on the planktonic stages.

6.5 Impacts and Stressors Other than Climate Change

As rocky shores are among the most accessible of the marine environments, they can be visited by large numbers of people, particularly shores near major cities, for aesthetic reasons, to collect food or bait or for fishing (Kingsford et al. 1991, Underwood 1993). Harvesting for food and bait can disrupt rocky shore communities, and also deplete some species while allowing for increase in dominance of other species (Keough et al. 1993, Keough & Quinn 1998, Keough & Quinn 2000, Airoldi et al. 2005). Trampling can damage and kill intertidal flora and fauna (Keough & Quinn 1998, Keough & Quinn 2000, Airoldi et al 2005). Coastal development will change hydrology and topography of areas and thus the characteristics of rocky shore communities. Artificial structures such as seawalls and pontoons offer new substrate for colonisation and may facilitate spreading of species over long distances (Chapman & Bulleri 2003, Bulleri et al. 2004). Alteration of water quality and input of pollutants will also impact on rocky intertidal communities.


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Main Findings: Rocky Shores • Expected changes

o Warming will drive species distributions southwards, evidence of such shifts is already apparent from rocky shores outside Australia

o Changes in population abundance and community composition will occur o Alteration of reproductive output o Frequency of diebacks will increase as number of extreme hot days

increases • Observed changes within the Australian region

o None investigated as yet.

IMPACTS OF CLIMATE CHANGE ON CORAL REEFS

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7. IMPACTS OF CLIMATE CHANGE ON CORAL REEFS

Ove Hoegh-Guldberg, Centre for Marine Studies, The University of Queensland, St Lucia, 4072, Queensland, [email protected]

7.1 Overview

Tropical coral reefs are composed mostly of hermatypic corals (Phylum Cnidaria, Class Scleractinia) containing microscopic symbiotic algae (Symbiodinium) that together produce the carbonate substrate on which the reef ecosystems are based. Much of Australia’s northern shores along both east and west coasts are lined by extensive coral reef ecosystems. Globally, coral reefs are estimated to house as many as one million species, of which less than 10% are known to science. In Australia, coral reefs provide critical habitat for a diversity of fauna and flora that includes over 400 species of corals, 4000 species of molluscs and over 1500 species of fish. The pristine nature of Australia’s coral reefs provide critical coastal protection against potential wave impacts, and generates over $5 billion in tourism alone per annum. Based on current coral physiology, a 1-2°C warming in tropical and subtropical Australia will lead to annual bleaching and regular large-scale mortality events where the symbiosis between the coral and its brown dinoflagellate symbionts disintegrates. Ocean acidification will tip the balance from coral calcification to erosion, with atmospheric CO2 levels above 500 ppm severely compromising coral viability. As temperatures warm, tropical reefs will only be able to extend several hundred kilometres southward; beyond this carbonate concentrations will decrease below their critical limit for calcification. Changes in intensity and frequency of storms, increased aridity leading to greater sediment load on reefs, and sea level rise are likely to act synergistically with temperature and acidification to reduce the coral populations. Although not seen before 1979, recent warming has led to repeated coral bleaching events in 1983, 1987, 1991, 1998 and 2002, with reefs in northwest Australia most vulnerable probably because northwest corals were more exposed to longer durations of heated water during these episodes than Great Barrier Reef corals. There is currently some debate over whether the impacts of ocean acidification are clearly evident Climate change will cause increased coral bleaching, reduced rates of calcification, increased sediment loads, higher sea levels, and shifts in storm intensity and frequency, reducing growth and survival of reef-building corals. This is likely to have major dramatic consequences for Australia’s coastal biodiversity, people and industry.


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7.2 Introduction

Much of Australia’s northern shores are lined by extensive coral reef ecosystems (Figure 7-1). They range from rocky reefs on which coral populations grow (non-carbonate reefs) to extensive coral communities living up the coastline carbonate reefs. These ecosystems are estimated to house perhaps as many as one million species (Reaka-Kudla 1996), the majority of which are unknown to science. Coral reefs are critical to Australia’s economic well-being: as well as providing resources and habitat for commercially important fish and crustacean species, they are the basis for one of Australia’s largest industries — tourism. They also play a critical role in coastal protection, moderating the impacts of waves on our coasts. However, both here in Australia and internationally, coral reefs are experiencing unprecedented stress from a range of disturbances, including overfishing, declining water quality and climate change.

Figure 7-1. Map of shallow coral reefs in the Australian mainland bioregions (Produced by CSIRO Marine and Atmospheric Research, 2006).

Two aspects of climate change are of greatest concern for the future of coral reefs. The first is the impact of rising sea temperatures, which drives mass bleaching and mortalities. The second is associated with the increasing acidity of the world’s oceans arising from the increasing amounts of carbon dioxide entering the ocean. Both of these aspects of climate change, when coupled with other factors such as the desertification of the Australian landmass and the shifts in storm intensity and frequency, are likely to cause major changes to coral populations around Australia. The central role that corals play in providing habitat for many species suggests that changes in the growth and survival rates of reef-building corals are likely to have major implications for Australia’s coastal biodiversity and industry.


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Coral reefs stretch across 284,300 square kilometres of the shallow, low-turbidity waters of the world. The major coral reef ecosystems around Australia stretch in the east from Frazer Island to Torres Strait, while on the west, they stretch from the Houtman Abrolhos reefs along the north-west coast of Australia to the western coast of the Gulf of Carpentaria. Throughout this range there is tremendous variability in the structure of the coral reefs, from poorly developed reefs that fringe continental islands and the Australian coastline to extensive carbonate barrier reefs. Scleractinian corals are a fundamental element of coral reefs, providing the framework in which the many thousands of associated plants, animals and protists live. The calcareous remains of corals and other organisms are cemented together by organisms such as calcareous red and green algae. Overall, coral reefs provide critical habitat for a diversity of fauna and flora, including over 400 species of corals, 4000 species of molluscs and over 1500 species of fish. The role of coral reefs in underpinning coastal economies in Australia is becoming increasingly recognised. The pristine nature of Australia’s coral reefs is the main reason international tourists visit them. The reef-associated tourist economy currently exceeds $5 billion per annum (Hoegh-Guldberg and Hoegh-Guldberg 2004). Despite the value of coral reefs to the economy, the reefs are among the most poorly studied of Australia’s ecosystems, with a large amount of their associated biodiversity as yet unidentified. International studies put the biodiversity of coral reefs at somewhere between 1 and 9 million species, with as few as 10% of those species being known to science (Reaka-Kudla 1996). This biodiversity is important potentially as a storehouse of genetic material, but it is also an attractive feature of such reefs as the Great Barrier Reef and Ningaloo Reef. The rich ecosystems of coral reefs depend on the myriad of interactions and functional groups to provide an inherent resilience to the effects of disturbance events. The importance of ecological complexity was illustrated in the Caribbean when intense fishing activity removed the grazing fish, leaving a single major herbivore—a sea urchin—on Jamaican coral reefs (Hughes 1994). The structure of the benthic community was again changed massively when a disease subsequently decimated the sea urchins. This example, and many others, suggests the important role of biodiversity in making coral reefs resilient to the effects of change. Australian coral reefs are in fairly good shape, compared to many elsewhere in the world (Pandolfi et al. 2005). The relatively healthy coral reefs that stretch along the vast, largely unpopulated, coastline have lower rates of reef degradation than elsewhere. This said, however, Australian coral reefs are facing significant pressures from changes to coastal water quality and overexploitation of key ecological functional groups, particularly in areas close to urban and agricultural developments. Historic studies such as that completed by C. M. Yonge in 1928 provide important data on how the Great Barrier Reef may have changed over the past 150 years (Yonge and Nichols 1931). For example, the extensive coral reefs in the tidal areas of the Low Isles have shown marked decreases in areal extent over the past 78 years. These changes, while documented only in a few studies, appear to be typical of many inshore reefs (Wachenfeld 1997). Water quality is thought to be the primary near-term threat to Australian coral reefs. Underlying the changes in water quality in the Great Barrier Reef catchment and coastal waters are large-scale changes to land-use along the Australian coastline. Since the first European settlements, cattle farming and intensive agriculture have dramatically


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increased the amount of sediment and inorganic nutrients flowing down the rivers and into the Great Barrier Reef lagoon. These apparently substantial increases seem to have occurred almost immediately after European agriculture started in the catchments. In the case of the Burdekin River, sedimentation increased 20-fold within a few decades of the arrival of cattle (McCulloch et al. 2004). Although other catchments have not been studied in this respect, there appears to be growing evidence that European farming methods have made major changes to sediment and nutrient flows into Great Barrier Reef waters, with corresponding impacts on water quality and the benthic ecology of coral reefs in coastal Queensland. Other impacts over the past 50 years are suspected of being linked to human influences. The crown-of-thorns starfish Acanthaster planci is a case in point. In outbreaks over the past 50 years, which starfish have reduced the abundance of corals to very low levels (Moran 1988, Moran et al. 1988, Pratchett 2005). Several hypotheses have been put forward to explain why crown-of-thorns starfish populations have had large fluctuations in abundance. Two major hypotheses are that crown-of-thorns starfish survivorship has been enhanced by either (a) increased concentrations of nutrients and hence phytoplankton food for crown-of-thorns starfish larvae or (b) the removal of key predators such as the triton (a predatory mollusc) and red emperor Lutjanus sabae increases survivorship. Recent evidence supports the first hypothesis (Brodie et al. 2005), although the effect of reducing predation through fishing pressure cannot be ruled out. This example highlights the complex nature of human influences and the response of coral reef communities. 7.3 Climate change in Australia’s tropical and subtropical

areas Problems of water quality and over-harvesting of Australia’s coral reefs are now occurring in the context of an environment that is also changing rapidly. It is also apparent that reef-building corals are particularly sensitive to changes being driven by climate change; in particular, changing sea temperatures and acidity from increased atmospheric CO2 in the water column are having strong influences on the survivorship and abundance of reef-building corals. Other factors such as changes in sea level, rainfall and storms may also impact coral reefs, although their effects are usually of lesser magnitude. Sea temperature

Tropical waters in Australia have warmed significantly over the past 150 years, with much of the increase occurring over the last 50 years (Lough 1999). The rates of warming are close to that derived for global tropical waters: +0.73oC between 1951 and 1990 (Lough 2004). These changes have brought corals closer to their thermal maxima, with the result that warm years arising from natural variability push corals beyond this point, which induces coral bleaching (a stress response in which the symbiosis between the coral and its brown dinoflagellate symbionts disintegrates). Corals bleach in response to a variety of stresses, including reduced salinity (Kerswell and Jones 2003); high or low irradiance (Vaughan 1914, Yonge & Nichols 1931, Hoegh-Guldberg & Smith 1989, Gleason & Wellington 1993, Lesser et al. 1990); the presence of toxins such as cyanide (Jones & Hoegh-Guldberg 1999) and copper ions (Jones 1997); microbial infection such as Vibrio (Kushmaro et al. 1996); and elevated or reduced


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temperatures (Jokiel & Coles 1977, 1990, Coles & Jokiel 1978, Hoegh-Guldberg & Smith 1989, Glynn & D’Croz 1990, Saxby et al. 2003, Hoegh-Guldberg & Fine 2004). Widespread coral-bleaching events have affected many parts of the world over the last 30 years. Mass coral-bleaching events stretching over hundreds to thousands of square kilometres of coral reefs have been reported in the scientific literature only since 1979. These events, which are triggered by warmer than normal conditions, can be predicted from sea-surface temperature anomalies measured by satellites (Hoegh-Guldberg 1999). Light is a co-factor: corals that are shaded do not bleach as severely as those in normal irradiances (Jones et al. 1998). Coral bleaching is also affected by water motion: corals that are in still, warm and sunlit conditions experience the greatest impact, which is consistent with the first observations that coral bleaching is associated with the doldrums conditions that are typical of El Niño years (Glynn 1993). While anecdotal reports suggest mass coral-bleaching events may have occurred before 1979, the lack of scientific reports before this date suggests that they have not been as frequent or severe as in the past four decades. Over this time, mass bleaching events have occurred about every 3-5 years and have affected almost every coral reef worldwide. Bleached corals may recover their dinoflagellate symbiotic populations in the months after a bleaching event if the conditions triggering the event are mild and short-lived, but mortality up to 100% has occurred when conditions have been more stressful for longer. In the severest global cycle of mass coral-bleaching, 16% of corals that were surveyed before the 1998 event had died by the end of 1998 (Hoegh-Guldberg 1999, Knowlton 2001). However, this figure is an average and hides the fact that in some oceans, for example the Western Indian Ocean, up to 46% of corals had died (GCRMN 2000). Mass bleaching events have occurred in Australia repeatedly over the past 30 years: in 1983, 1987, 1991, 1998, 2002 and 2006, when large sections of the Great Barrier Reef were bleached. Mortality rates have been relatively low, however, primarily because the conditions were less severe than in other parts of Australia and the world. Scott Reef in the north-west waters of Australia has not been so lucky. In 1998, a very warm core of water sat above these oceanic coral reefs for several months, resulting in an almost total bleaching and mortality of corals down to 30 m depth. The recovery of these reefs has been very slow (GCRMN 2004). Projections of temperature changes in Australian seas are that they may soon exceed the thresholds for coral bleaching every year. Based on the current responses of corals to rises in sea temperature, it is estimated that an increase of 2°C in the average sea temperature in tropical and subtropical Australia would result in annual bleaching and quite possibly regular, large-scale mortalities (Hoegh-Guldberg 1999, 2004). The risk associated with these changes in sea temperature has been elegantly analysed by Done et al. (2003). In this study, the succession of devastating mass coral-bleaching events would be such that the ability of reefs to recover would be severely compromised. Deterioration of coral populations is likely in most of the scenarios examined by Done et al. (2003). These conclusions reinforce those of an earlier study (Hoegh-Guldberg 1999) and more recent studies such as that by Donner et al. (2005). Ocean pH and seawater chemistry

The changes to the atmospheric concentration of CO2 are a significant threat to calcifying organisms such as corals when CO2 enters the ocean and disturbs the pH


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balance of seawater (Raven et al. 2005). With the rising concentration of CO2 in the atmosphere, CO2 has increasingly entered the oceans, where it reacts with water to form carbonic acid. The rising CO2 levels have acidified the ocean and are shifting the ion balance from carbonate to bicarbonate ions. As calcification is linearly related to the carbonate ion concentration (Langdon et al. 2000), this shift is a major threat to calcifying organisms. Doubling atmospheric CO2 effectively decreases the carbonate concentration to 200 µmol·kg-1 (although there is a small dependency on temperature). At carbonate ion concentrations, calcification of corals and many other calcifying organisms effectively becomes zero. Given that coral reefs represent a balance between calcification and erosion, it would appear that atmospheric concentrations that exceed 500 ppm (pre-industrial CO2 levels were 280 ppm, in 2006 levels were 380 ppm, and levels of 600ppm are predicted for 2100) would severely compromise the ability of corals to maintain themselves against the forces of physical and biological erosion. There has been some debate how serious ocean acidification is for coral reefs (see McNeil et al. 2004 versus Kleypas et al. 2005). Cores from long-lived corals such as Porites allow one to estimate how calcification has changed over the past century. When Lough and Barnes (2000) examined coral cores, they saw no evidence of a decline in the calcification of Great Barrier Reef corals over the past 50 years; in fact they saw an increase. They also found that calcification was highly correlated with average sea temperature, with an annual average increase in calcification of 0.3 g·cm-

2·y-1 for each degree of increase in temperature. Lough and Barnes (2000) suggested that the increase in calcification was probably due to the 0.25°C increase in sea temperature on the Great Barrier Reef over the same period. Critically, these measurements were undertaken on coral cores that ended in mid 1980s, which leaves in question what has happened during the past 20 years. The physiology also does not support the continued increased in calcification with temperature. Although calcification does increase with temperature, it does not increase indefinitely with temperature; several studies have shown that it increases up to the summer sea temperature maximum but declines rapidly at higher temperatures (Kleypas et al. 2005). The link between increasing calcification and higher sea temperatures has misled some authors as to how corals will respond to future increases in ocean acidity. McNeil et al. (2004), for example, conclude that the effect of acidification on coral calcification will be compensated by the increase in calcification rate due to warmer conditions. This conclusion is invalid, for corals would have problems with bleaching (see above for a critique of McNeil et al. 2004, and see Kleypas et al. 2005). The physiological literature tells us that calcification will increase up to the summer sea-temperature maxima and decrease thereafter. On the balance of evidence, coral reefs will not be able to able to maintain themselves or grow if carbon dioxide concentrations rise above 500 ppm (Hoegh-Guldberg 2005). Changes to sea level, storm intensity and other factors

Sea level has risen about 25 cm over the past 100 years (Pittock 1999) and is currently rising at 1–2 mm a year, an order of magnitude larger than the average rate over the previous several millennia (Church et al. 2001). The rise is slow compared to the rates of coral growth (20 cm per year; Done 1999) and hence is not a major challenge to healthy coral populations. The problem comes with a combination of sea-level rise and slower growth due to the effects of increased ocean temperatures and acidity: coral populations might not be able to keep up with the sea-level rise.


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Recent events and modelling studies have highlighted the possible influence of climate change on storm intensity and frequency (Webster et al. 2005). Storms impact coral reefs through the direct physical impact and also their influence on river flows and hence the amount of sediment flowing down to the reefs. The impact of storms such as Hurricane Andrew has been documented in examples, which hit Florida and other parts of the Caribbean in 1992 has been documented. Porter and Meier (1992) recorded that the coral reefs were affected by wave impacts. However, Meier and Porter (1994) commented that the physical damage from this category 4 storm is much less than that caused over longer periods by other factors that lower water quality and create coastal pollution. The immediate impact of storms on Australian coral reefs is likely to be similar to those that have impacted the Caribbean coral reefs. Greater intensities and frequencies of storms will increase these impacts, however, and could act in conjunction with other factors, such as increased ocean temperature and acidity, to further reduce coral populations, tipping the balance of reef growth towards a net dissolution. On the Australian mainland there has been a drying trend along the eastern seaboard over the past 70 years (DEH 2005). The trend is reducing vegetation and destabilising sediments in catchments. With more intense storms, more destabilised sediments are likely to be washed down river catchments and into coastal areas. The impacts would probably be similar to those seen when hard-hoofed cattle were introduced into river catchments on the east coast of Australia: within a few decades of cattle farmers arriving, sediment flows increased by 20-fold (McCulloch et al. 2004). Increased sediment load from such modification to watersheds or from increased rainfall or storm severity is well known to impact corals through direct smothering and changes in water quality, light penetration, and nutrient regimes (e.g. Rogers 1990, Dubinsky & Stambler 1996, Neil et al. 2002, Fabricius 2005). 7.4 Potential changes to coral reefs as a result of climate

change The combined impact of local and global stressors appears to be rapidly changing the distribution and abundance of corals on reef systems worldwide (GCRMN 2004). Coral cover—one of the primary measures of coral abundance—has declined on many reef systems, with subsequent losses of fish and other associated organisms. For example, on the Seychelles coral reefs Graham et al. (2006) demonstrated that coral bleaching resulted in local extinctions, reduced taxonomic distinctiveness and species richness, and a loss of species within key functional groups of reef fish. The authors considered that these changes raised doubts that reefs in this state had the potential to recover in the near term. Interestingly, with a note in reference to the Great Barrier Reef, the authors suggest that oceanic reefs, such as those in the Seychelles, may be more vulnerable to stresses resulting from climate change than those associated with continental islands and land masses. Whether being remote from other impacting factors (such as river run-off and increased sedimentation) is advantageous is debatable. Inspection of historic photos and surveys like those of the Great Barrier Reef Expedition in 1928-29 clearly demonstrates that coastal populations of corals have declined over the past century (Yonge & Nichols 1931). Large areas of intertidal reef surrounding in the Low Isles in north-eastern Australia, for example, do not have any scleractinian corals growing on them, although they were named after these corals e.g.


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Porites pond, and Madrepore Moat (personal observation). Many inshore reefs (usually within 10 kilometres of the Australian land mass) have gone from having rich scleractinian coral populations to having little or no coral left. In more offshore regions, the changes are less clear and a matter of active debate. Recent studies, such as those by Bellwood et al. (2004) contrast to some extent with the results of the Long-Term Monitoring Project LTMP (Miller & Sweatman 2005). In the former case, the authors report data that show an overall decline in coral cover over the past 40 years, while the latter suggest that coral reefs with the Great Barrier Reef have deteriorated over the same period. One of the key questions is whether coral reefs in Australia will move in a poleward direction. The current determinants of coral reefs correlate with temperature, light and carbonate ion concentrations of the seas in which they grow (Kleypas et al. 1999); the distribution of reef-building corals is not determined by temperature alone hence, a poleward shift in the temperature regime does not mean that coral populations will shift. Given that the carbonate ion concentration decreases in a poleward direction and will further decrease as atmospheric CO2 increases, carbonate concentrations would decrease below the critical level of 200 µ mol.kg-1 not too far south of the present day limit of coral reefs. Kleypas et al. (2001) therefore suggest that the poleward creep of coral reefs would be restricted to a few hundred km at most. It is interesting to speculate, however, as to how the more mobile fauna of coral reefs would move. Some fishes associated with coral reefs appear able to populate reefs that do not have corals, as shown by the appearance of coral reef fishes in southern New South Wales and Victoria in the summer months. These fishes recruit into coastal areas and grow for several months, disappearing when cold conditions return. While species that are corallivores would clearly be missing from these temporary migrants, it appears that many coral reef fish might be able to move southwards as the global ocean warms. This has already been observed in other parts of the world; for example, Mexican fish arriving in Californian waters (Holbrook et al. 2000, Hoegh-Guldberg 2004). The loss of calcifying organisms on coral reefs would have implications for limestone structures provided by coral reefs off Australian coastal areas. Currently, reef systems such as the Great Barrier Reef and Ningaloo Reef are buffers between the power of the Pacific and Indian Ocean waves and the extensive ecosystems and human infrastructure that line Australia’s coastal areas. The limestone structures of coral reefs are a balance between calcification by corals and other organisms and biological and physical erosion. Rates of calcification are low relative to erosion: corals grow at 1–10 cm per year (Gladfelter et al. 1978, Gladfelter 1984, Pittock 1999), whereas coral reefs grow at 0.5–1.0 m per 1000 years (Odum & Odum 1955, Smith & Kinsey 1976, Davies & Hopley 1983). Given that erosion is likely to remain unchanged or possibly even increase if the pH changes and more carbonate is dissolved, the net rate of reef growth is likely under many climate-change scenarios to become negative. Such changes have as yet been documented on only a few coral reefs, such as Faaa in French Polynesia (Pari et al. 1998). Coupled with sea-level rise, the loss of reef structures from coastal areas could have major implications for not only reef habitats but for the functioning of entire coastal areas in Australia. The degree of resiliency of corals and coral reefs to predicted changes in climate and ocean conditions is under some debate (Hoegh-Guldberg et al. 2002). Some coral species are more vulnerable than others as the result of various adaptations to different conditions and environmental variability. After losing their Symbiodinium in a severe


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bleaching event, for example, some corals show the ability to shift their symbiont populations to more thermally tolerant varieties (e.g. Berkelmans & van Oppen 2006). These changes come from corals having evolved to be symbiotic with several strains of Symbiodinium. The degree of resilience and adaptability this affords the host corals, given geographic and taxonomic variability in such adaptive capacities, is still poorly known (Rowan et al. 1997, Baker 2003) but it is unlikely to have a big impact on the trajectories of coral populations under climate change as few corals show multiple Symbiodinium in their tissues. Also relating to resiliency is the observation that coral reefs throughout the world have become severely degraded—many more so than the Great Barrier Reef—and the possibility that there are reinforcing feedbacks, such as cyanobacteria and marine algae, that might prevent their recovery by holding the ecosystems at a degraded ‘alternate stable state’ (e.g. Pandolfi et al. 2005, Fong et al. 2006). Resolving these dilemmas is a critical field of research for the management and protection of the world’s coral reef ecosystems. Socio-economic consequences

How changes to coral reefs would affect the humans that depend on them is clearly a question that requires considerably more space than a concluding paragraph. Australia’s coral reefs are currently an important economic engine for its communities and coastal economies. The best-developed example of the economic importance of coral reefs in Australia is the Great Barrier Reef. Two industries — tourism and fishing — dominate the productivity of coastal economies associated with this reef. Tourism is the largest industry in this region, generating over $5 billion per annum, but it is largely dependent on the pristine reputation of the Great Barrier Reef (Pandolfi et al. 2003). This reputation enables the Australian tourism sector to compete successfully for international tourists interested in seeing untouched coral reefs. Hoegh-Guldberg & Hoegh-Guldberg (2004) analysed the potential effect of losing that competitive edge if the coral reefs on the Great Barrier Reef deteriorated as a result of climate change. The effects vary, depending on domestic and international trends in such aspects as climate, markets and politics. However, if one uses the reef-interested component of these economies as a guide to how things might change if reefs continue to degrade, it becomes clear that degradation would reduce international tourist income by as much as $8 billion over 19 years (Hoegh-Guldberg & Hoegh-Guldberg 2004). Although similar problems might be expected to occur in other areas of the world, Australia has the disadvantage that in-coming tourists have to travel two or three times as far to see Australia’s reefs. Economic losses of this magnitude would have serious consequences for coastal communities where, in some cases, over 80% of external earnings are dependent on international tourism. Understanding the economic and social ramifications of climate change on coral reefs will be critical in preparing people, industries and governments for the changes that seem almost certain to occur over the next few 50 years.

7.5 Conclusion

Much of Australia’s northern shores are lined by extensive coral reef ecosystems, and coral reefs are critical to Australia’s economic well-being. They are indisputably a resource of global significance, and the social and economic consequences of coral reef degradation in Australia have been estimated to be huge. Australia’s coral reefs harbour tremendous biodiversity and provide considerable ecosystem services, and although they are in relatively good shape compared to the degraded coral reef systems


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elsewhere, they are threatened by impacts and stressors that are increasing and combining in worrisome ways. The climate changes of greatest concern include rising sea temperatures, which drive mass bleaching and mortalities, and increasing acidity of the world’s oceans, which will directly impact corals and other organisms with calcium carbonate structures. Increasing storm severity is another looming climate change impact that is a big worry for the future of Australia’s coral reefs, especially when combined with other stressors such as water quality degradation and over-harvesting. The biological and ecological impacts of these changes will be varied, but they will undoubtedly continue to degrade coral reef resources, and probably severely. The degree of resiliency of corals and coral reefs to predicted changes in climate and ocean conditions is thus critical to developing an effective strategy for adaptation and mitigation.

Main Findings: Corals • Expected changes

• Based on current coral physiology, a 2°C warming in tropical and subtropical Australia will lead to annual bleaching and large-scale mortality events (at least once per decade but probably more frequently)

• Ocean acidification will tip the balance from coral calcification to erosion; atmospheric CO2 that exceeds 500 ppm will severely compromise the ability of corals to build and maintain limestone reefs.

• Changes in intensity and frequency of storms, increased aridity leading to greater sediment load, and sea level rise are likely to act synergistically with temperature and acidification to reduce coral viability

• Tropical reefs will only be able to extend ranges several hundred km south with warming because carbonate concentrations will decrease below their critical limit

• Observed changes within the Australian region • Warming over the last 30 years has led to repeated coral bleaching events,

with the reefs in NW Australia most vulnerable

IMPACTS OF CLIMATE CHANGE ON DEEP-SEA AND COLD-WATER CORALS

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8. IMPACTS OF CLIMATE CHANGE ON DEEP-SEA AND COLD-WATER CORALS

Elvira Poloczanska & Alan Butler, CSIRO Marine and Atmospheric Research [email protected]

8.1 Overview

Deep-sea cold-water corals occur globally, but research is relatively recent and limited. In Australian waters, these corals are presently known only from sites around south and southeast Australia, particularly within the Tasmanian Seamounts Marine Reserve and on seamounts in Australian territorial waters around Norfolk Island (Figure 8-1). The reefs represent essential fish habitat, with large aggregations of fish such as the commercially exploited orange roughy, Hoplostethus atlanticus found on some seamounts. They are considered ‘hotspots’ for biodiversity, comparable to shallow-water tropical coral reefs, despite a paucity of knowledge of their ecology, population dynamics, distribution and ecosystem functioning. Importantly, it is acknowledged that the reefs are very slow growing, taking many thousands of years to develop. The three major threats to cold water corals are bottom trawling by fisheries, ocean acidification and seabed mining. Global distribution of cold water corals appears to be influenced by the depth of the aragonite saturation horizon, below which corals will have difficulty laying down skeletal matrix. As atmospheric CO2 levels increase, the depth of the aragonite saturation horizon shallows and it is estimated around 70% of known cold water coral reefs globally will be affected by 2100. Seamount habitats off south Australia may become inhospitable for cold water corals below a few hundred metres. As Australian seamount cold water corals are areas of high endemism and speciation, the resulting loss to Australian marine biodiversity will be high.

8.2 Introduction

Cold water corals have been recorded from all the world’s oceans and comprise a diverse taxonomic group including colonial stony corals (Scleractinia), soft corals (Octocorallia), black corals (Antipatharia) and lace corals (Hydrozoa) (Freiwald et al. 2004, Morgan 2005, Guinotte et al. 2006). These corals differ from shallow tropical reef-forming species in that they lack symbiotic algae and are found down to depths of over 1000 meters below sea level. The extent and profusion of cold water corals in the deep sea is only now being appreciated with recent advances in technology for underwater exploration (Roberts et al. 2006). Only some species are reef forming but most form 3-dimensional complex structures on normally flat, featureless, muddy deep sea floor (Freiwald et al. 2004). Scleractinian species can accumulate to form large reefs and carbonate mounds which can be tens of kilometres across and several metres high (Reed 2002, Roberts et al. 2003, Freiwald et al. 2004). Cold water corals may rival shallow tropical coral reef structures in area of sea floor covered (Freiwald et al. 2004).


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Lophelia pertusa, Madrepora oculata and Oculina varicosa are the most prolific reef builders in northern hemisphere waters (Friewald 2004). Madrepora oculata is also found in waters off northwest Australia and New Zealand while Goniocorella dunosa and Solenosnilia variabilis are pronounced reef builders on seamounts and ocean banks in Tasmanian and New Zealand waters (Friewald 2004) (Figure 8-1). Oculina varicosa is known from sites on the continental rise off south and east Australia.

Figure 8-1: Sampling locations illustrating known distribution of cold water corals around eastern Australia (courtesy Karen Miller, University of Tasmania)

Seamounts tend to deflect currents, inducing circular currents and enhancing local upwelling. A seamount is defined as an underwater mountain, often volcanic in origin, rising more than 1000m above the surrounding flat silty seafloor - with their topographically enhanced currents and unique environmental conditions, seamounts tend to be areas of enhanced biological productivity and can support huge numbers of cold water corals and other filter feeding animals (Rogers 1994, Koslow et al. 2001, Genin 2004, White & Mohn 2004). Smaller pinnacles (20m or more), often loosely referred to as seamounts, also support these distinct biological communities (Rogers 1994). The Tasmanian Seamounts Marine Reserve, between 50 and 170 km off southern Tasmania, contains about 70 pinnacles between 20 and 500m high and several kilometres across at their base. These are the only known examples of these structures in Australian continental waters and those peaking at a depth of less than 1400m are dominated by dense stands of the coral Solenosmilia variabilis (Figure 8-1) (Bulman et al. 2002, see www.deh.gov.au for more details on the Tasmanian Seamounts Marine Reserve). Seamounts with cold water coral communities are also found in Australian territorial waters to the east of Norfolk Island and along the Norfolk Ridge, adjacent to Norfolk Island (Williams et al. 2006).


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Cold water corals represent essential fish habitat (Husebo et al. 2002, Keiger & Wing 2002) providing shelter, feeding grounds, spawning grounds, nursery area and other uses (Bull et al. 2001, Dower & Perry 2001, Reed 2002b). Foraging seabirds, marine mammals and pelagic fish concentrate above seamounts as the shallower water depth, localised circulation features plus often elevated productivity provide good feeding opportunities (Blaber 1986, Haney et al. 1995, Holland et al. 1999, Freon & Dagorn 2000, Genin 2004, Yen et al. 2004). Fish aggregations have been found to be highly localised on the sides on seamounts (Bulman et al. 2002), and commercial deep water fisheries may target these fish (Koslow 1997, Clark 1999, Koslow et al. 2000, Bull et al. 2001, Anderson & Clark 2003, McClatchie & Coombs 2005, O’Driscoll & Clark 2005, Shepherd & Rogan 2006). For example, orange roughy (Hoplostethus atlanticus) seasonally migrate to seamount areas in Australian and New Zealand waters to form large spawning aggregations that are targeted by fisheries (Clark 1991, McClatchie & Coombs 2005). Cold water coral and sponge communities can be considered ‘foundation species’ (Jensen & Frederiksen 1992, Rogers 1999, Reed 2002b, Jonsson et al. 2004). Seamounts and coral reefs are often referred to as ‘hotspots’ for biodiversity. Lophelia pertusa reefs in the northeast Atlantic were found to have over 1,300 species of invertebrates suggesting a biodiversity comparable to shallow-water tropical coral reefs (Jensen & Frederiksen 1992, Freiwald et al. 2004, Roberts & Freiwald 2005). Even dead Lophelia pertusa colonies support a high biodiversity of epifauna (Jensen & Frederiksen 1992). Similarly, the biodiversity of benthic fauna on Oculina reefs in southwest Atlantic was found to be very rich and equivalent to the biodiversity of shallow tropical coral reefs (Reed 2002b). Non reef forming species, such as the deepwater gorgonian corals Primnoa spp. also sustain a high diversity of associated macrofauna (Krieger & Wing 2002). In Australian waters, over 850 macro- and megafaunal species were recently found on seamounts in the Tasman and southeast Coral Seas of which 29 - 34% are new to science and potential endemics (Richer de Forges et al. 2000, Williams et al. 2006). The unique hydrological characteristics of seamounts, coupled with their biological isolation, means many may have high levels of species endemism and could be important centres of speciation (Richer de Forges et al. 2000, Koslow et al. 2001). Such speciation could be the result of reproductive and genetic isolation due to geographic separation as well as unique localised hydrographic conditions and the large variety of substrata (Rogers 1994). Low species overlap between seamounts in Tasman and southeast Coral seas, and even at separate ridges at the same latitude, indicates the seamounts are acting as ‘island groups’ with highly localised species distribution (Richer de Forges et al. 2000, Williams et al. 2006). It is hypothesized that cold water coral structures may function as areas of spreading for associated species, acting as a ‘stepping stone’ for long range dispersal between regions (Moore et al. 2004, Williams et al. 2006). The reefs appear to contain core populations of species that may become unsustainable or have difficulties spreading if the reefs disappear (Fossa et al. 2002). Study of samples from north Atlantic Lophelia pertusa reefs found almost 40% of recorded species had not been previously recorded from the area (Jensen & Frederiksen 1992) while a third of species of fish and invertebrate recorded from seamounts in Tasmanian waters were restricted to the seamount environments (Koslow et al. 2000, Koslow et al. 2001).


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Very little is known about their ecology, population dynamics, distribution and ecosystem functioning but it is becoming increasingly clear that cold water corals inhabit places where natural disturbance is rare (Morgan 2005). Cold water corals and sponges are long-lived, slow-growing and susceptible to physical disturbance, making them extremely slow to recover from disturbance (Morgan 2005). These ecosystems have evolved in highly stable conditions they are therefore likely to be impacted by climate change (Guinotte 2005).


Temperature

Reef building cold water corals are largely restricted to temperatures between 4°C and 12°C (Roberts et al. 2003, Roberts et al. 2006). It is assumed that as the corals have evolved over many thousands of years for this narrow stable temperature range, any increase or decrease in temperatures will negatively impact on coral physiology. Rising temperatures will influence calcification rates, physiology and biochemistry of cold water corals. Laboratory experiments showed a that 3°C (sub-lethal) increase in mean temperature increased calcification rates of shallow water hermatypic corals by 24-39% (Edmunds 2005). Calcification rates are fastest within an optimal temperature range and decline with temperatures higher or lower than this (Edmunds 2005, Schneider & Erez 2006). Preliminary findings suggest Lophelia pertusa respiration rates increase at higher temperatures but synergistic effects of increased temperature and respiration on calcification rates are unknown (Guinotte 2005). Chemical analysis of deepwater corals off southern Australia has indicated a long-term cooling of ambient waters which commenced in the mid 18th century (Thresher et al. 2004). This cooling may be accounted for by an enhanced flow of the warm EAC increasing coastal sea surface temperatures and resulting in a shift the depth distribution of temperature isotherms along the coastal margin—not by a wide-spread cooling of the deep ocean. Chemical analysis of the deep-water corals indicated a cooling of about 1°C of ambient deep water over the past 250 years, though spacing of the coral growth rings gave no indication of changes in growth rates so far in this limited sample (Thresher et al. 2004). The strengthening of the East Australian Current is predicted to continue as global climate warms, and so these resulting changes may eventually impact cold-water coral physiology. Mixed layer depth

In the Northern hemisphere cold water corals occur in waters that have higher than average surface primary productivity (Heikoop et al. 2002, Roberts et al. 2003, Duineveld et al. 2004). Cold water corals are suspension feeders surviving in deep waters below the photic zone so may be dependant on surface productivity filtering down. If climate change alters surface productivity then this could impact dramatically on deep sea corals. Further, there is evidence of seasonality in North Atlantic cold water corals with the spring phytoplankton blooms influencing reproductive periodicity (Waller & Tyler 2005). Changes in the phenology of plankton species may either result in a corresponding alteration in the phenology of deep sea coral reproduction or lead to a mismatch between coral reproduction and food availability thus resulting in lower reproductive potential for the corals and a possible lack of suitable food for larval stages. Seamounts peaking at depths less than 1000m may continuously intersect deep


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scattering layers of organisms (Fock et al. 2002, Williams et al. 2006) and intercept diurnally migrating zooplankton. Additionally, diurnally migrating zooplankton may become trapped above seamounts or swept onto seamounts by prevailing currents (see Rogers 1994, Butler et al. 2002, Genin 2004). Links with surface productivity may be weak for seamount cold water corals ecosystems (Bulman et al. 2002) but studies are very limited. It must be assumed alteration to the mixed layer depth will have implications for cold water coral communities. Currents

Cold water corals are sessile animals that depend on water flow to supply food and remove waste and excess sediment, so any changes in local current regime could have a significant impact on their distribution (Roberts et al. 2003, Guinotte 2005). It could affect nutrition, hence growth, reproduction and longevity in a local population, and it could affect dispersal, hence altering the ‘stepping stone’ function of seamount chains. Cold water corals tend to occur in areas of high current activity and this is evident on Australian seamounts where corals tend to occur in distinct depth zones (Koslow et al. 2001). Alteration of currents may make areas no longer favourable for coral growth and, given the low growth rate, colonisation of newly available areas with optimal environmental conditions may be slow and may take many decades or longer before a viable population size is reached. High water flow may also be necessary for the supply or retention of larvae to maintain populations (Genin et al. 1986). Sea level

Changing sea level can affect cold water corals by altering bathymetry and currents hence altering the environment in areas presently favourable for coral growth (Lindberg & Mienert 2005) – see section on ‘Currents’. pH and seawater chemistry

Global distribution of cold water corals appears to be influenced by seawater carbonate chemistry with a clear relationship between cold-water scleractinian occurrences and the depth of the aragonite saturation horizon (Guinotte et al. 2006). Aragonite is the form of calcium carbonate deposited by corals and the aragonite saturation horizon represents the limit between the upper saturated and the deeper undersaturated waters; above the horizon calcium carbonate can form but below it dissolves (Raven et al. 2005). Corals will have difficulty laying down skeletal matrix in undersaturated waters - research has already shown that tropical hermatypic corals show reduced calcification rates when exposed to elevated CO2 (Langdon et al. 2000, Langdon & Atkinson 2005). As atmospheric CO2 levels increase, the depth of the aragonite saturation horizon will decrease and the entire Southern Ocean could become undersaturated by 2100 (Orr et al. 2005, Caldeira & Wickett 2005). The saturation of high latitude waters is already low due to colder temperatures; saturation also decreases with depth so cold water corals are inhabiting areas generally less favourable for calcium carbonate deposition (Raven et al. 2005). It is suggested that cold water corals are much more vulnerable to changes in ocean chemistry than shallow tropical hermatypic corals (Raven et al. 2005).


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There appear to be no published observations explicitly on the observed impacts of climate change on Australian cold water coral communities. A discussion of potential impacts of climate change on abundance, distribution, phenology and physiology is given below together with consideration of potential consequences for associated biodiversity. Distribution

Cold water coral reefs and mounds are slow growing and have developed over long periods of time in extremely stable environments so are unlikely to be able to shift distributions with rapid climate change. The growth rate of a Lophelia pertusa reef is estimated to be 1.1mm per year, so it could take thousands of years for a reef 10-30 m thick to develop (Hall-Spencer et al. 2002). Samples of Lophelia pertusa for the surface of reefs from waters west of Scotland were dated as 3800 years old suggesting the reefs themselves may have formed up to 10,000 years ago (Roberts et al. 2005). Deep water corals (Keratoisis spp.) collected off southern Australia were aged as between 300 - 400 years old (Thresher et al. 2004). As the oceans acidify decreased carbonate (aragonite and calcite) saturation will have disastrous effects for calcium building animals. Corals will build weaker skeletons leaving them more susceptible to physical disturbance and erosion and growth rates may slow (Odhe & Hossain 2004). There is some debate as to whether increased temperatures, providing these are sub-lethal, may lead to an increase in growth rates which may counter-balance the reduction caused by increased atmospheric CO2 but, considering all the evidence, this is unlikely (McNeil et al. 2004, Kleypas et al. 2005, Hoegh-Guldberg 2005, see Coral Reef section for details). The predicted decrease aragonite saturation horizon in global oceans mean, within a hundred years, only 30% of known deep-sea corals reefs and mounds will remain in supersaturated waters compared to the present day figure of >95% (Guinotte et al. 2006). Seamount habitats off south Australia may become inhospitable for cold water corals below a few hundred meters (Guinotte et al. 2006) and they, with their multitude of associated fauna, could simply disappear . Cold water corals may be significant contributors to global calcium carbonate budget, and may account for at least 1% of global carbonate production (Lindberg & Mienert 2005). This has implications for the global carbon cycle; the biological deposition of calcium carbonate acts as a significant buffer of ocean chemistry and atmospheric CO2 (Ridgwell & Zeebe 2005). As the oceans become increasingly acidic, biological carbonate deposition will reduce but it is uncertain how this will impact on the transfer of CO2 from the atmosphere to the deep ocean carbon reservoirs (Ridgwell & Zeebe 2005) but the disappearance of the cold water coral reefs could lead to further perturbation of the global carbonate cycle. Phenology

Evidence suggests diurnally migrating zooplankton and overwintering calanoid copepods may be ecologically significant prey species of cold water corals (Duineveld et al. 20004). Changes in these populations through alteration in trophic web, phenology


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and distributions of plankton in the photic zone may have severe knock-on consequences for deeper benthic communities. Where it has been studied in the North Atlantic, reproduction in cold water corals appears to be triggered by deposition of organic material (marine snow) from the surface water spring plankton bloom (Waller & Tyler 2005), so changes in this may either result in a mismatch of timing events or a shift in cold water coral phenology. In Australian waters known occurrences of cold water corals are generally on seamounts which have enhanced currents, so trophic links with surface productivity may be weaker (Bulman et al. 2002, see Mixed Layer Depth section). Consequences for biological communities and biodiversity

As three dimensional complex structures in the often featureless deep sea, the cold water corals provide habitats for many species. Our knowledge of these communities is limited compared to other systems but ecological effects of degraded or destroyed reefs may be substantial. These corals function as foundation species so if they disappear, the entire ecosystem based on them will also disappear. Cold water coral reefs and seamount communities are also areas of high endemism and speciation, so the resulting loss to Australian marine biodiversity will be very high if these are adversely impacted by climate change. Further, it is likely that a high proportion of species on coral communities are unknown to science, and may never be known from living records if these coral communities are lost.

8.5 Impacts other than climate change

The three major threats for these cold water coral ecosystems have been identified as bottom trawling by fisheries, ocean acidification, and hydrocarbon drilling and seabed mining (Roberts et al. 2006). Fishing

Cold water coral reefs and sea mount communities are fragile and at high risk from bottom trawling and other disturbances (Clark 1999, Hall-Spencer et al. 2001, Andrews et al. 2002, Fossa et al. 2002). The fisheries over seamounts and reefs typically use heavy duty trawl gear designed to cope with the rough, hard seabed (Fossa et al. 2002, Hall-Spencer et al. 2002, Anderson & Clark 2003, Freiwald et al. 2004). The trawls are extremely damaging for the fragile sessile corals, smashing their structures, as well as removing large bycatches of corals, gorgonians and other fish and invertebrate species (Koslow et al. 2001, Hall-Spencer et al. 2002, Anderson & Clark 2003, Freiwald et al. 2004). Between 30 -50% of estimated total area of cold water corals in Norwegian waters have found to be damaged by trawling (Fossa et al. 2002) and substantial dead coral was found on heavily fished Australian seamounts (Koslow et al. 2001). Additionally, intense trawling may also keep coral colonies below a minimum reproductive size causing sterility of the coral community (Waller & Tyler 2005). In the north Atlantic, fishermen expressed concerns in the early 1990s about the effect of trawling on Lophelia pertusa reefs; it was claimed that as corals disappeared from trawling grounds, catches of fish in these areas were also lowered (Fossa et al. 2002). A decline in the bycatch of benthic fauna concurrent with a decline in target species was noted in fisheries on seamounts in New Zealand waters (Clark 1999).


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Fish fauna associated with cold water coral reefs tend to be slow growing, with low natural mortality and low fecundity rendering them highly vulnerable to over fishing and unlikely to recover from population crashes (Mace et al. 1990, Koslow et al. 2000, Morato et al. 2006). In addition, some species form large spawning aggregations on seamounts that are targeted by the fisheries (Clark 1991, McClatchie & Coombs 2005). Seamount fisheries tend to follow a pattern of an initial fisheries development phase with large catches followed by rapid declines in catches, such as occurred with orange roughy fisheries in New Zealand and Australian waters (Clark 1999, Clark 2001, Koslow et al. 2000). A progressive serial depletion of seamount stocks was recorded as the fishery moved from seamount to seamount. Other anthropogenic factors

Some seamounts are rich in minerals and it is likely that commercial mining of these will eventually be proposed as mining technology develops and other reserves become depleted (Butler et al. 2002). If the sea bed is mined for mineral resources, impacts may result directly on benthic fauna through operation of machinery resulting in high localised mortality and indirectly though excessive sedimentation, effects of which may be more widespread (Johnston & Santillo 2004, Freiwald et al. 2004). Exploitation of methane gas hydrates could impact on sediment motility and the potential for seafloor collapse and sediment resuspension could adversely affect benthic fauna such as cold water corals (Johnston & Santillo 2004).

Main Findings: Cold water corals • Expected changes

o Reduction in distribution of deep sea corals as the aragonite saturation horizon shallows

o Estimated that 70% of deep corals will be impacted • Observed changes within the Australian region

o None investigated as yet

IMPACTS OF CLIMATE CHANGE ON SOFT SEDIMENT FAUNA

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9. IMPACTS OF CLIMATE CHANGE ON SOFT SEDIMENT FAUNA

Thomas A. Okey, CSIRO Marine and Atmospheric Research, [email protected]

9.1 Overview

Soft sediment invertebrate fauna such as crabs, bivalves, prawns, and worms cover a broad spectrum of sizes and life habits. These invertebrate assemblages are extremely diverse in Australia overall; this is due, in part, to the subtle diversity and vast extent of soft sediment habitats in Australia’s marine environment. Extremely high diversities of soft sediment fauna are also found at very small spatial scales (e.g., 0.1 m2) on some Australian continental shelves. Very high densities, biomass, and diversity of these fauna occur in some areas in Australia due, in part, to relatively high supplies of detrital food at ‘productivity hotspots’ in an otherwise oligotrophic (nutrient poor) system. The diversity of soft sediment fauna is apparently also shaped by energy in the form ocean temperature, in addition to the chemical energy in food. Soft sediment fauna is of key importance in utilizing and supporting the detrital food web in the ocean, as opposed to the planktonic food web. It recycles a considerable portion of the overall energy in a system that would otherwise be lost. They are indeed a main link coupling the production originating from primary producers with the production that relies on detritus. To a very large extent, they support and mediate the production and maintenance of Australia’s marine ecosystems thereby supporting the Australian economy and society through ecological stabilisation, fisheries production, pollution mitigation, and other services. Australia’s soft sediment communities are thus sensitive to changes in production in the overlying water column and to changes in adjacent coastal systems that affect production or sedimentation, or cause organic enrichment or other pollution. Anthropogenic (human generated) disturbances such as seafloor trawling considerably modify the structure and functions of soft sediment ecosystems that are heavily fished, though non-climate related human impacts such as pollution are often local. Because of the very high proportion of biological flows (e.g. food energy that supports ecosystems) mediated by soft sediment invertebrate fauna, both destructive fishing and climate change will impact commercially valuable invertebrates (e.g. prawns, crabs, pearl oysters), fish, and other valued vertebrates throughout Australian marine ecosystems, but what will the relative impacts of these stressors be, and how will they interact?. What are the effects of climate change on benthic and demersal fisheries? Can


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‘manageable’ stressors such as fisheries be adjusted to bolster the resilience of these systems, thus alleviating the potential impacts of climate change? Australia’s soft sediment communities are an ideal model system with which to explore the interactions and trade-offs among such factors because this fauna is shaped strongly by (productivity) food, temperature, and disturbances and stressors such as fishing. The challenge is to design comparative and experimental management studies (e.g. closed areas) that control for, and can clearly distinguish, the various factors that shape the structure of these biological communities. Such short term empirical studies should be designed in concert with long-term monitoring and ecological modelling studies

9.2 Introduction

The soft sediment habitats of the sea floor cover most of the earth’s surface and the vast majority of Australia’s marine domains (Bunt 1987, Poore 1995). The benthic (sea floor) biota associated with these habitats includes a broad range of taxa, sizes, habitat preferences, life habits, and distributions. Most of these organisms are invertebrates, and they range in size from the microfauna (< 0.062 mm) living in the interstitial spaces between sediment grains through meiofauna (0.062 mm - 0.5 mm) and macrofauna (> 0.5 mm) (see Nybakken 2001), and finally to megafauna; e.g. larger than ~20 or 50 mm. Infauna live in the three-dimensional top layer of sediment (mostly in the top 5 cm, but to a metre or more deep), while epifauna live at or on the sediment surface. The fish fauna inhabiting Australia’s soft sediment habitats are discussed separately. Some benthic invertebrates are mobile, while others are sessile, though many species switch modes at different life stages, or diurnally, or according to lunar or oceanographic cycles. Many emerge from the sediment regularly or occasionally, for example, to swim or migrate vertically in the water column for feeding, reproduction, dispersal, or migration. Soft sediment habitats

Soft sediment benthic habitats extend from the high water mark of muddy estuaries and sandy beaches across the continental shelves and down the continental slopes to the abyssal plain of the deep sea (i.e. deeper than 2000 m), which alone covers about 40% of the world’s sea floor. Although a single sediment type can typically cover “many thousands of km2” in soft sediment environments (Gray 2002), the composition of Australia’s soft sediment types and habitats varies at different spatial scales and according to depth, latitude, and sediment source (e.g. Bunt 1987, Poore 1995). Much of the seafloor of Australia’s shelf is covered with relict biogenic carbonate sands that grade from course near shore to finer offshore. To a lesser extent, terrigenous mud dominates some nearshore northern tropical areas near rivers, while quartz sand is found only anomalously around Australia (Bunt 1987). The northeast region exemplifies continual production of carbonate sands from the Great Barrier Reef seaward of a zone of terrigenous muds inshore, which has expanded due to watershed modifications. In Australia’s southern regions the relict carbonate sands interface with rocky habitats. Production and diversity

The sediment-water interface is the site of highly concentrated primary production by films of microphytobenthic organisms (see Bunt et al. 1972, Sournia 1976), and this


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primary production often rivals or surpasses the production of phytoplankton in overlying water columns (Cahoon & Cooke 1992, MacIntyre et al. 1996), especially in shallow zones of continental shelves, but also across the continental shelves of clear water systems (Okey et al. 2004). Furthermore, any detritus produced by the organisms in the system that is not consumed within the system or exported laterally is deposited at the sediment-water interface, as is excess imported detritus. The combination of high in situ primary production and deposition of detritus at the sediment-water interface can support high secondary production of consumers manifesting as high abundances, biomass, turnover (energy throughput), and biological diversity. The feeding modes of benthic fauna include suspension feeding, filter feeding, deposit feeding, herbivory, detritivory, omnivory, and predation, and thus almost the full gamut of trophic levels is represented within soft sediment ecosystems—even without considering vertebrate faunas. Soft sediment benthic organisms interact with the organisms of adjacent systems in addition to with each other, and they are the primary food of many commercial species, such as prawns, lobsters, crabs, and fishes (Hall 2002), in addition to some seabirds and marine mammals. They support indirectly a large proportion of the overall biota in adjacent water columns through trophic connections (Okey et al. 2004). Benthic fauna thus often dominate and mediate the biomasses, energy flows, and dynamics of continental shelf and slope ecosystems, such as the relatively low nutrient systems surrounding Australia. Benthic communities are often coupled strongly with neritic (nearshore water column) and pelagic (open water column) systems through biological, physical, and chemical pathways. The biomasses and secondary production rates of marine soft sediment fauna can be extremely high on continental shelves and slopes (Vetter 1994, Okey 2003), and the biodiversity of these systems can be very high in the deep sea (e.g. Hessler & Sanders 1967, Sanders 1968) and in shallow water (Gray et al. 1997, Gray 2002). Likewise, very high diversity is found in some of Australia’s soft sediment habitats (Poore et al. 1975, Coleman et al. 1978, Parry et al. 1990, Rainer 1991, Poore & Wilson 1993, Poore et al. 1994, Coleman et al. 1997, Gray et al. 1997) but especially along the Victorian coast of Bass Strait were Coleman et al. (1997) reported up to 198 species in a single 0.1 m2 grab sample! Poore (1995) suggests that the high infaunal diversity observed in Australia’s soft sediments is partly explained by the predominance of course calcareous biogenic sands (see also Gray 2002), and this is supported by findings that muddy terrigenous sediments in Australian estuaries harbour a less diverse fauna (Jones & Candy 1981). Little is known yet about the soft sediment communities enhabitingn the abyssal plains surrounding Australia’s continental shelves, but the little we do know suggests that they are far more diverse than previously thought. There are internationally-recognised hotspots of tropical marine biodiversity just to the north of Australia (i.e. coral reef species and fish), but it is unclear whether this is reflected in soft-sediment invertebrate communities. Infaunal diversity is generally lower in nearshore muddy habitats of the north, but muddy habitats do give way to carbonate sands further offshore with various regimes of disturbance and food input. The latitudinal gradient of species richness in the northern hemisphere is apparently not reflected in the southern hemisphere (Clarke 1992) possibly due to hot-spots of available productivity in southern high latitudes (Gray 2002) as well as influential differences in sediment type and habitat. The patterns of Australian infaunal abundance and diversity are still somewhat unexplained, but a well designed empirical study on


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climate change impacts could complement and synthesize existing information to distinguish climate impacts from other factors. Climate change impacts

In general, oceanographic and other changes associated with global climate change can affect soft sediment communities strongly, thereby affecting entire marine ecosystems, especially in Australia’s oligotrophic marine domains. The main drivers of climate change impacts on Australia’s soft bottoms covered here are sea temperature, rainfall / coastal runoff, wind and currents, ultraviolet radiation, and pH. These variables drive biological changes mainly because they: • Influence the production or provision of food for soft sediment organisms; • Influence sediment and substrate types, distributions, and habitat chemistries; and • Shape the natural disturbance regimes that influence organisms directly and

indirectly through modifications to substrate, food, and biological interactions. A conspicuous problem is that research and information on marine soft sediment habitats and communities has been extremely limited and highly variable, considering the extent, magnitude, and fundamental ecological role of this environment and fauna to marine ecosystems (Peterson 1991, Constable 1999). This disproportionately small amount of research on soft sediment environments is particularly true for Australia (Fairweather 1990, Fairweather & Quinn 1995). Given this paucity of information, the present discussion is limited mostly to macro-infaunal communities. Epibenthic mega-invertebrate assemblages of soft bottoms are discussed only briefly. This macro-infaunal scope is particularly useful here for two reasons: • Certain metrics of macroinfaunal community structure might prove to be excellent

indicators of climate change for assessment and long term monitoring programs—especially if combined with metrics describing the overlying plankton communities that they largely depend on; and

• Soft sediment fauna encompasses a large portion of Australia’s marine biodiversity, despite the general perception that coral reefs harbour the vast majority of Australian marine species.

The potential for using macroinfaunal assemblages in the development of climate change indicators relates to: • The sensitivity of this assemblage to the oceanographic manifestations of climate

change; • The potential to distinguish the impacts of climate change from the impacts of other

oceanographic variability and other anthropogenic disturbances; and • The relative ease of sampling and quantifying these assemblages. Gray (2002), for example, points out that a single 0.1 m2 grab sample can contain over 50 species of macrofauna and that most species are non-motile making detection of change relatively easy. The challenges include expensive ship time and difficulties in sorting and identifying large numbers of infauna (Hall et al. 1994), attributing causality in a multi-variate system, and choosing useful community metrics.


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The distributions, abundances, and persistence of species and the assembly and structure of biological communities are shaped by six general aspects: 1. Dimensions of physical, chemical, and biogenic habitat structure; 2. The provision, availability, and distribution of food resources; 3. Existing disturbance regimes; 4. The spatial and temporal variability and patterns of these habitats, food resources,

and disturbance regimes; 5. Biological interactions in the context of these dimensions; and 6. The particular histories of a given assemblage in terms of biological abundances,

assembly rates, extinctions, or invasions (see Okey 2003). Modification of any of these aspects can modify soft sediment community structure. General themes of relationships between environmental variables and patterns of biota can be extracted from the limited number of Australian studies. Food, habitat, and disturbance

Soft sediment infaunal communities are shaped by a combination of temperature, food input, and disturbance regimes. This is highlighted in several studies in northeastern Australia along a gradient from the Great Barrier Reef lagoon to bathyal and abyssal depths of the Solomon and Coral Seas (Alongi 1987, Alongi & Pichon 1988, Alongi 1989, 1992, Alongi & Christoffersen 1992). The high diversity of infauna on the continent’s northwest shelf also appears to have resulted from the combined effects of disturbance, predation, and productivity (Rainer 1991). A combination of similar factors (sediment types, disturbance regimes, temperature, productivity, and biological interactions) has also been invoked to explain the high diversity of nematode assemblages distributed across the Great Barrier Reef Lagoon (Tietjen 1991). All these results are consistent with the broad theme that soft sediment communities are structured by a combination of temperatures, productivity, and natural disturbance regimes, in the context of biological interactions. Sediment type has also long been thought to shape infaunal assemblages. The distributions of infaunal assemblages in Port Phillip Bay, Victoria, have been found to be related at least partly to sediment type (Poore and Rainer 1979); assemblages and distributions of the infauna of the Gulf of Carpentaria were related most strongly to sediment grain size and depth (Long and Poiner 1994); and similar relationships have been found in Bass Strait along the Victorian coast of southeast Australia (Coleman et al. 1997). Similarly, the distributions of epibenthic mega-invertebrates in the Gulf of Carpentaria were associated strongly with sediment type with clear distinctions between sandy- and muddy-bottom communities (Long et al. 1995). Depth is often associated with benthic infaunal patterns because depth can be correlated with the factors that shape these communities. Depth, for example, has been found to be strongly associated with the diverse isopod fauna along the southeastern Australian continental slope from 200 to 300 metres depth (Poore et al. 1994). This study did not measure the roles of food, productivity, biological interactions, and disturbance explicitly, but the results are again consistent with the overall pattern described previously; grain size distribution and sediment type are shaped by depth-associated wave and storm disturbance regimes, current regimes, and trawling regimes in the


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context of sediment sources (see Snelgrove & Butman 1994). Also, sediment type and depth are associated with productivity and temperature. All of the Australian studies reviewed here are consistent with the more generally held notion that soft sediment assemblages of continental shelf systems are shaped by combinations of factors (Hall 2002), whether alternatively (Menge & Sutherland 1987) or simultaneously (Okey 2003). Factors that regulate soft-sediment communities thus fit into three dominant themes: food (productivity), habitat, and disturbance. It follows that empirical studies that can account for variability in habitat characteristics and disturbance regimes could reveal the effects of climate changes such as temperature and other factors related to productivity. Comparative snapshot studies could provide this insight immediately without waiting for the results of long-term monitoring. Such empirical studies should, however, be designed in concert with such monitoring programs and integrated with ecological modelling approaches that incorporate predicted changes in environmental variables that will be affected by climate change such as sea temperature, rainfall/runoff/salinity, wind and currents, acidification and UV radiation. But first, theoretical frameworks emerging from studies of marine soft sediment fauna will help us identify how climate change impacts will manifest in these communities, in the context of other disturbances. Species-energy relationships and productivity hotspots

Gray (2002) presents some compelling evidence that species richness in marine soft sediments is a function of the amount of energy (e.g. food and temperature) that an area receives. This is consistent with Wright’s (1983) species–energy hypothesis, which had previously gained favour in explaining patterns of species richness in terrestrial biota (Kerr & Packer 1997, Kerr 2001). Productivity, i.e. food, is of fundamental importance in marine ecosystems. Energy can also be measured as temperature, and some authors have found sea temperature to be the best indicator of marine taxonomic richness in shallow subtidal settings (Fraser & Currie 1996, Roy et al. 1996, Roy et al. 1998). Recognizing the combined influence of temperature and productivity can resolve longstanding dilemmas in marine ecosystem understanding. For instance, there appears to be no latitudinal gradient of species richness in the southern hemisphere, as there is in the northern hemisphere (Clarke 1992, Gray 2002). This could be because of hot-spots of available productivity in the high latitudes of the south (Gray 2002). Productivity

The availability of food has been demonstrated to be a principal factor shaping at least some marine soft sediment communities in Western Australia (Peterson & Black 1986, 1987, Black & Peterson 1988, Peterson & Black 1988b, a, Peterson 1991, Peterson & Black 1991). Food is a key factor for soft-bottom fauna in many settings (Levin et al. 2001), and this might be true generally for much of Australia’s soft sediment communities because of the generally oligotrophic (nutrient poor) status of Australia’s marine realm. Seafloor secondary production depends on the downward flux of organic material produced in the water column (Alongi 1998), and the degree to which these two systems are coupled depends on the degree of recycling of detritus in the water column and other factors that influence the access of benthic organisms to water-column detritus (Grebmeier & Barry 1991) The deposition of detritus from water column productivity results in enhanced sea floor production, which in turn makes nutrients available to support pelagic production (McLusky & McIntyre 1998).


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Food is a principal shaper of marine soft sediment communities, even in cases where other factors, such as disturbance, limits populations. In the case of Australian soft sediment communities, any factors that would lead to decreases in the productivity of Australia’s marine plankton (see sections in this report on phytoplankton and zooplankton) and the downward flux of detritus from those systems, or the advection of food from adjacent regions, will likely impact Australia’s impressive soft sediment biodiversity adversely. This could have important implications for important ecosystem services and commercial resources. In similar ways, these factors would likely lead to decreases in the in situ benthic primary production at the sediment-water interface. Sea temperature

Temperature is related to diversity and community structure because it mediates a whole suite of physical, chemical, and biological mechanisms. For instance, temperature influences metabolism and thus productivity rates, but it has also been shown to mediate interactions such as predator avoidance thus influencing diversity patterns (Crame 2002). There is some recent evidence that the distributions of soft sediment fauna are sensitive to temperature or climate changes (Godinez-Dominguez & Gonzalez-Sanson 1998, Kroncke et al. 1998, Figueroa et al. 2003, Moreno et al. 2006) complementing broader evidence of the impacts of climate-related temperature change on marine hard bottoms (e.g. Barry et al. 1995, Sagarin et al. 1999). Depth is associated with temperature and productivity (and distance from productive shores and irradiance), and thus depth can explain faunal patterns that are really associated with temperature and productivity (Levin et al. 2001, Gray 2002). Temperature and productivity are indeed the main factors that underpin Longhurst’s (1998) useful definitions of biogeographical provinces in the world’s marine systems. Oxygen concentrations in seawater are temperature dependent, and oxygen concentrations at the sediment-water interface shape soft sediment communities very strongly (Gray et al. 2002, Okey 2003). Increased temperatures can decrease oxygen concentrations thereby expanding zones of hypoxia and anoxia caused by watershed modifications and nutrient enrichment (Hall 2002), such as the dead zone in the Gulf of Mexico (Rabalais et al. 2002). The net deposition of organic detritus from the water column (or from land-based sources) to the sea floor implies that marine soft sediment is a global sink of carbon that was once in the atmosphere in the form of CO2. Increases in ocean temperatures can potentially release this stored carbon into seawater and the atmosphere, in the form of methane, thus causing a feedback that can potentially speed global warming (Hall 2002). Rainfall / coastal runoff / salinity

Rainfall and riverine output influences the primary production and benthic communities in coastal marine waters through the supply of organic matter (e.g. Creutzberg 1985), nutrients, and pollutants (Boesch et al. 2000), but it also imposes strong influences through changes in salinity and sedimentation. For instance, Jones et al. (1986) found that macrobenthic community patterns in the Hawkesbury Estuary, New South Wales were more strongly associated with salinity than with sediment characteristics, and that most species were found at the most seaward transects while least species were found at the transects furthest upstream. In this case, salinity appears to be a principal shaper of marine soft sediment communities in estuaries, as opposed to further offshore where


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sediment type, food, and other factors are likely to be more important. All of these characteristics are influenced by river discharge which is a primary shaper of soft bottom communities on most continental shelves, but particularly in tropical areas where smaller watersheds produce more sediment (Milliman 1991, Alongi & Robertson 1995, Hall 2002). Rhoads et al. (1985) highlighted these strong influences in a generalized model of the influence of river discharge on soft sediment benthic fauna, and this is relevant to contrasts between Australia’s tropical north and temperate south. In addition to recent attention to riverine impacts in Northern Australia and adjacent areas (e.g. Alongi & Robertson 1995, Alongi & McKinnon 2005), and a well developed literature on the effects of sedimentation on coral reef communities (Rogers 1990, Wilkinson 1999, Fabricius 2005), a prominent New Zealand team has examined the impacts of modified coastal sedimentation regimes on soft sediment communities (Norkko et al. 2002, Thrush et al. 2003a, Thrush et al. 2003b, Lohrer et al. 2004, Thrush et al. 2004, Thrush et al. 2005). This growing body of work suggests that increases in rainfall, or extreme rainfall events, can increase upland erosion and sediment transport considerably. This results in significant impacts to estuarine and nearshore coastal ecosystems, especially along coastlines where development and other human land uses have degraded the integrity of watersheds. Australia’s estuaries and coastal watersheds are highly prone to modification and changes in both salinity and sedimentation regimes due to the concentration of human population along coasts, the growing need for coastal water for human uses, rapid changes in land use and development, and the political reluctance to allow for ecological allocations. The fauna associated with these estuaries and dependent coastlines are thus extremely vulnerable to changes in salinity and sedimentation regimes. Natural disturbance regimes

Regimes of sediment transport from coastal watersheds is one example of a natural disturbance regime that can be modified by changes in climate, but others include disturbances associated with winds, waves, storms, and currents. An extreme, but illustrative example is the one presented by Rothlisberg et al. (1988) where severe impacts on the tiger prawn stocks of northern Australia would be expected from increases in the frequency or intensity of cyclones, since the seagrass nursery areas of tiger prawns would be vulnerable to such changes (discussed further in the section in this report on seagrasses by Poloczanska). The frequency of intense cyclones has indeed increased during the last 50 years (Webster et al. 2005) and the destructiveness of cyclones has increased since 1970 (Emanuel 2005). There are many more subtle changes to natural disturbance regimes that can influence benthic fauna aside from changes in rainfall/runoff discussed previously. Changes in winds, weather, and hydrographic regimes, for example, might cause major changes to much of Australia’s soft sediment environments and habitats, as waves and currents re-work and re-suspend bottom sediments (Hall 1994). There are several Australian examples of disturbance regimes and sedimentation causing shifts and changes in benthic biota (e.g. Jones & Candy 1981).


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pH and seawater chemistry

Decreases in pH and aragonite saturation will directly impact many important soft sediment organisms that rely on calcium carbonate structures (e.g. mollusc groups and benthic foraminifera), as well as many planktonic organisms that the benthic fauna relies on, thus impacting primary and secondary production both on and above the bottom (see sections on phytoplankton and zooplankton in this volume for a discussion of the mechanism of acidification impacts). Any adverse impacts to water-column components cause impacts to soft sediment biota because the two are often tightly coupled in terms of food transfer. In addition to direct impacts on the biota, changes in pH will also change the composition and nature of the sediment itself, thereby modifying the habitat and causing indirect impacts to the soft sediment biota. This is because a large fraction of the sediment is calcium carbonate in origin.


There appear to be no published observations explicitly on the observed impacts of climate change on Australian soft sediment environments and communities


The main non-climate driven anthropogenic impacts and stressors affecting Australia’s soft bottoms and their communities are sea floor trawling, modification of sediment transport from coastal watersheds, and pollution. These impacts include the effects of mine tailings Sea-floor trawling

There is considerable interest and research currently on the impacts of trawling on sea-floor ecosystems and communities (e.g. Jones 1992, Tuck et al. 1998, Watling & Norse 1998, Collie et al. 2000, Gray 2000, Jennings et al. 2001a, Jennings et al. 2001b, NRC 2002, Thrush & Dayton 2002), and some such research has been conducted in Australia (Sainsbury 1987, Moran & Stephenson 2000, Pitcher et al. 2000, Koslow et al. 2001, Burridge et al. 2003, Tanner 2003, Pitcher et al. 2004, Tanner 2005). This body of knowledge has established that trawling often does considerably degrade and modify sea floor marine ecosystems, particularly in settings from which very slow-growing biogenic structures such as gorgonians and other architectural organisms are removed. However, there are cases where impacts are not detected (e.g. Drabsch et al. 2001). Indeed the impacts of trawling are likely to be often underestimated because there are no documented examples where pre-trawling communities are described and quantified for comparisons to post-trawl communities at the same location. There is a smaller body of research indicating that seafloor trawling also impacts soft bottom communities adversely by removing many fine habitat structures on, and in, the soft substratum (Sparks-McConkey & Watling 2001, Watling et al. 2001, Rosenberg et al. 2003). It is not surprising that some opportunistic organisms would benefit from seafloor trawling in some settings (Engel & Kvitek 1998, Pitcher et al. 2004). Prawns are one of the groups that are apparently facilitated by moderate levels of trawling (e.g. Cushing 1984).


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Pollution

The effects of some aspects of pollution on marine soft bottoms are relatively well known (Pearson & Rosenberg 1978, Gray et al. 2002). Some of Australia’s marine pollution problems are considerable, but many have only local effects. Published research and overviews of the impacts of pollution on marine soft sediment fauna or habitats include those focusing on eutrophication (Peterson et al. 1994), associated toxic cyanobacterial blooms (Garcia & Johnstone 2006), oil spills (Burns & Codi 1999), sedimentary heavy metals in estuaries and continental margins (Birch 1996, 2000, Matthai & Birch 2000, Liaghati et al. 2004, Burton et al. 2005), plastics pollution (Gregory 1999), hazardous wastes such as organochlorine pollutants (Thompson et al. 1992, Matthai & Birch 2000), sewage (Scanes & Philip 1995, Morris & Keough 2002), general industrial pollution (He & Morrison 2001), and pollutants from land-based activities in general (Pearce 1991, Williams 1996). Australia is developing some plans to deal with marine and upland pollution problems (Pearce 1991, Thompson et al. 1992, Nelson 2000). It is expected that some of these problems might be alleviated in time, but increases in intense rainfall episodes in agricultural and urban areas, including increasing flood frequencies, could exacerbate other marine pollution problems in Australia thereby reducing resiliency to climate change impacts. Areas that experience net decreases in rainfall or changes in the frequency of severity of extreme events would also influence coastal pollution regimes and the resilience of biological communities that are adapted to particular hydrodynamic regimes. Geosequestration might become a common tool for the mitigation of excess CO2. Some counteries have proposed burying CO2 under the seafloor, while others have proposed disposing of it in the water column where it sinks and forms pools on seabed. This is an important future potential threat to soft sediment communities, particularly in deeper waters.

9.6 Conclusion

Australia’s soft sediment communities directly support a number of important fisheries, but the indirect services they provide are much more widespread and pervasive because they are the conduit for the detritus-based portion of marine food webs. Diversity, abundance, and biomass of infaunal communities has been shown to be extremely high in some parts of Australia, but the factors explaining these patterns are complex and can be organised into three dominant themes food (productivity), habitat, and disturbance. Future changes in sea temperature and other factors that affect productivity such as rainfall and coastal runoff are likely to influence soft sediment communities strongly. Acidification may drive strong changes as well, but this is less certain. Soft-sediment habitats are an ideal model system with which to distinguish climate change impacts from other factors such as fishing impacts, natural disturbances, and habitat characteristics. The challenge is to design comparative and experimental management studies (e.g. closed areas) that control for, and can clearly distinguish, these various factors that shape the structure of these biological communities. Comparative snapshot studies could provide this insight immediately without waiting for the results of long-term monitoring, but such short term empirical studies should be designed in concert with long-term monitoring and ecological modelling studies designed to estimated the community impacts of predicted changes in environmental variables that will be affected by climate change such as sea temperature and rainfall/runoff/salinity, but also wind and currents, acidification and UV radiation.


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Main Findings: Soft sediment fauna • Expected Changes

o Changes in water column primary production will impact abundance and composition of soft sediment fauna by reducing food sources

o Increases in water temperatures are likely to shift many species southwards

o Ocean acidification will erode the shells of mussels and other groups with carbonate shells

o A range of factors including temperature rise, changes in freshwater input, and altered storm activity will impact soft sediment communities through ecological and physical disturbances

o Changes in biological communities will change the physical characteristics of soft sediment, possibly influencing stability

• Observed Changes o There are no observed changes to soft sediment communities as yet

due to the limited research on this key habitat


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10. IMPACTS OF CLIMATE CHANGE ON BENTHIC AND DEMERSAL FISHES

Thomas A. Okey, CSIRO Marine and Atmospheric Research, Cleveland Alistair J. Hobday, CSIRO Marine and Atmospheric Research, Hobart [email protected]

10.1 Overview

Australia’s benthic and demersal fish fauna is expected to continue shifting southwards, where geographically possible, with continued strengthening of the East Australian Current. Their populations and biomasses, and thus their central functional roles in ecosystems, are expected to decline where their ranges are bounded geographically, and because of predicted changes in factors such as zonal westerly winds that are expected to decrease recruitment, though some species presumably might increase as functional niche space opens. Declines in several commercially-important benthic and demersal fish species have been observed in response the climate-related changes that have already been recorded; most of these observations have occurred in eastern or south-eastern Australian marine domains, though some such changes have been observed in the west. Projected climate changes such as warming, increases in the strength of East Australian Current, and decreases in productivity are likely to decrease the overall abundance, biomass, productivity, and diversity of benthic and demersal fish species in Australian marine waters, thus substantially modifying overall marine communities and services. It is possible, however, that some species or assemblages may benefit in some areas.

10.2 Introduction

The benthic and demersal fishes of Australia include all teleost fishes (a subclass of fishes that are bony and with rayed fins) and elasmobranchs (a subclass of fishes that are cartilaginous, comprising the sharks, the rays, and the Chimaera) that inhabit the ocean floor (whether geological, biogenic, or anthropogenic in origin) throughout Australia’s marine realm. Much of the energy and nutrients in Australia’s marine domain flows from water-column and sea floor production sources to fish species with benthic and demersal lifestyles. These fishes play a key role in shaping most aspects of Australia’s marine biological communities through predation on secondary producers,

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as forage for other fishes, vertebrates, and scavengers; and sometimes through bioturbation. These species provide tremendous social and economic services as targets for valuable commercial and recreational fisheries in Australia. Australia’s benthic and demersal fish fauna is extremely diverse and endemic by world standards due to a high diversity of tropical and temperate habitats, and due to the geographic isolation of its temperate regions. This fauna is expected to continue shifting southwards in response to projected climate changes, but only where geographically possible, as the East Australian Current and the Leeuwin current continue to strengthen. Their populations and biomasses, and thus their central functional roles in ecosystems, are expected to decline where their ranges are bounded geographically, but other more tropical benthic and demersal fish species will likely expand their ranges and increase their overall biomass in certain areas of Australia’s marine realm due to the spatial and temporal expansions of their preferred conditions and due to decreases in competition and the opening of functional niche space.


Expected declines in benthic and demersal fish biomasses and distributions due to climate changes will result from projected increases in temperature throughout Australia’s marine domains and changes in other environmental variables such as decreases in zonal westerly winds over southern Australia. The effects of changes in other environmental variables might be less important for most, but not all, benthic and demersal fishes. Increased temperatures are expected to change the phenology of fish life cycles and change (shift) the forage resources of fishes differentially thus likely leading to very-difficult-to-predict community changes relating to mis-matches of co-evolved predator-prey relationships, in addition to shifting ranges southward. Furthermore, decreases in zonal westerly winds will likely change local current patterns and decrease primary and secondary production, thereby decreasing recruitment. Because of the poleward boundedness of southern Australia, projected climate changes will likely decrease the biodiversity of benthic and demersal fish species, thus potentially modifying overall marine communities and services substantially. These effects will occur primarily through increases in sea surface temperature, increases in the strength of East Australia and Leeuwin currents, and decreases in productivity. These general predictions of change assume, for convenience and optimistically, that fishing pressure will decrease to sustainable levels and remain stable. A number of fished stocks in the south-east region and elsewhere are over-exploited currently, and many others are strongly influenced by fisheries, making it necessary to account for fishing impacts when examining climate impacts. From a fisheries perspective, the impacts of climate changes on target stocks and their communities are of particular concern for the future sustainability of these fisheries, whether or not they will be managed successfully. Lyne and his colleagues (2003) conservatively projected that the expected increases in Australian ocean temperatures would cause a 35% overall economic decline of Australian fisheries by 2070, and that temperate Australian fisheries will be more vulnerable than tropical ones. For example, they projected the economic impacts on


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Tasmanian, Victorian, and Western Australian fisheries to be 64%, 40%, and 38% declines, respectively, by 2070, assuming that fishing effort is managed well. Sea temperature The large predicted changes in the sea-surface temperatures of Australia’s south-eastern region to climate change will likely have a strong effect on the ranges of many fish species and assemblages in that region, and such changes have already been observed there (discussed later). Other Australian marine domains will be vulnerable to sea temperature changes as well, though it appears that the most dramatic impacts will be in the southeast. At the core of this biological impact is a tendency for shifts in species distributions poleward and towards newly appropriate thermo-habitats and shifted food resources. It is well recognised that sea temperature is a principal determinant of fish species abundance and distribution (e.g. Lehodey et al. 1997, Roessig et al. 2004, Perry et al. 2005), biomass (e.g. Ware 1995, O’Brien et al. 2000, Drinkwater 2005), and other critical life history and physiological processes (e.g. Burkett et al. 2001). Much of the relationship between temperature and benthic and demersal fish populations is likely due to temperature-related productivity in the ocean, in addition to physiological dependencies. This relationship between temperature and fish production and distribution began to become apparent through examination of the relationship between decadal-scale oceanographic (temperature and productivity) regime shifts and fisheries catches (Beamish et al. 1997, Mantua et al. 1997, Beamish et al. 1999) and zooplankton and seabirds (McGowan et al. 1998). The CSIRO climate models and a suite of other coupled climate models predict that Australia’s southeast region will experience the largest increase in sea surface temperature (Figure 2.1) due to the likely strengthening effect of a poleward shift in zonal winds on the warm East Australia Current. These predictions are consistent with the Maria Island (Tasmania) time series, which has shown summer sea-surface temperatures to rise by over 1°C over the last 60 years, and the deep-water corals proxy temperature record developed by Thresher et al. (2004). These temperature changes are expected to have large effects on the distribution of many benthic and demersal fish species, the composition of their communities, and ecosystem function in general. The east-west orientation of the temperate Australian coastline adds a critical boundary to the southward movement of many species, thus increasing the risk of extirpation and extinction with increases in water temperature. In addition to fish species, the poleward shift in distribution of the European shore crab and the barren-forming urchin indicates that marine invertebrates are also reshaping marine ecological communities by shifting their ranges polewards (e.g. Johnson et al 2005).

Winds The climate models also consistently predict a poleward shift in the zonal winds that normally cross the southern part of Australia, and these predictions are consistent with recent changes in the Antarctic Oscillation Index (Gillett and Thompson 2003; though see also Jones and Widmann 2004). This predicted subsequent general weakening of those winds is expected to hinder recruitment for both fished and non-fished marine populations because strong relationships between wind strength and fish recruitment have been shown, including in a coastal rocky reef fish (Heteroclinus sp.; Thresher et


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al. 1989) and in a commercially fished gadoid of the outer shelf (Macronurus novaezelandiae, blue grenadier; Thresher et al. 1992). In southeastern Australia, Harris et al. (1992) found evidence that the weak jack mackerel (Trachurus declivis) production off Tasmania resulted from decreased wind stress and subsequent decreases in large zooplankton. Such decreases in secondary production can also considerably influence the production of benthic and demersal fishes, as discussed in the soft sediment biota and plankton sections of the present report. Some years earlier, Harris et al. (1988) related the variability in the annual frequency of strong zonal west winds to catch rates and recruitment variability in several southeastern demersal fisheries. The collapse of the gemfish fishery in that region was likely due to the combination of weak recruitment due to declining winds and over-fishing (Thresher 1996). A variety of south-eastern shelf teleosts exhibit a decadal-scale recruitment cycle, in several cases directly linked to regional wind (Thresher 2002; Jenkins et al. 2005). Rainfall/Coastal runoff/Salinity Close to tropical Australia in New Caledonia, Wantièz et al. (1996) found the biomass of fishes to be negatively associated with both temperature and rainfall, indicating that a warmer and wetter climate would decrease fish biomass. The relationship of temperature to fish biomass was discussed previously. Increased rainfall could adversely impact fishes by increasing turbidity, sedimentation, and nutrient input, and decreasing light penetration to the seafloor, thereby changing several attributes of fish habitat and food resources. The degradation of coral reefs, or the initiation of phase shifts from corals to fleshy algae has profound effects on fish assemblages. Important fish nursery habitat such as estuaries, mangroves, and seagrass meadows via changes in rainfall, runoff, and salinity might strongly affect early life history stages and recruitment, as well as transport of young fishes to offshore shelf and slope habitats. UV Radiation As discussed in the pelagic fishes section of this report, increases in UV radiation might affect survival of planktonic egg and larval stages of many of Australia’s benthic and demersal fishes. Such effects are not well known in Australia or elsewhere.


Abundance/distribution Declines in several commercially important benthic and demersal fish species have been observed in response the climate-related changes that have been recorded already. Most of these observations have occurred in eastern or southeastern Australian marine domains, though some such changes have been observed in the west. The distributions of at least 36 species of Tasmanian marine fish species (in 22 fish families) have changed considerably during the previous decade, almost always in the form of poleward shifts. Many of these are warm temperate reef species normally distributed adjacent to the coast of New South Wales. These have become established south of Bass Strait, and others have shifting their ranges south along the Tasmania coast, and some are totally new records for Tasmania (P. Last, CSIRO, in prep.).


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The gemfish fishery, already under pressure from apparent over-fishing (like many of the south-east stocks), collapsed altogether when the zonal winds declined to their predicted low point in the 10-year cycle (1989). In fact, the zonal winds in this year were at the lowest levels ever reported. Since then, the winds have fallen further, perhaps reflecting the predicted onset of overall weak zonal winds in the south-east region due to climate change, and also perhaps explaining why, despite the fishery being closed, there has been little or no sign of recovery of the eastern gemfish stock (Caton and McLoughlin 2004).

Phenology Changes in the phenology, or the timing of life stages or migrations, of benthic and demersal fish specie have been observed in Australia by commercial and recreational fishermen and by other local and indigenous people. Rigorous studies are needed to provide quantitative descriptions or evaluations of these observed changes in phenology. Predicted consequences for biological communities and biodiversity Benthic and demersal fishes are key structuring forces of Australian marine ecosystems, and this is indicated by some dynamic models of Australian marine ecosystems (e.g. Okey, in press), but such predictions are just now emerging.


Fishing The fishing pressure on benthic and demersal fishes can be substantial, particularly in some regions. For instance, eight of the 17 stocks listed as overfished in the Australian Bureau of Resource Sciences 2004 annual report on the state of Australia’s fisheries are in the south-east demersal fishery group (Caton and McLoughlin 2004). Fishing is recognized as having a huge impact on fish assemblages in the world’s oceans (Pauly et al. 1998), but particularly along coastlines and on continental shelves. The effect of fishing is presumably so important that assessments and projections of climate change impacts should not be conducted without considering the impacts of fishing in the area of interest. Pollution Benthic and demersal fishes are known to be affected in a variety of ways by pollution, and some marine areas of Australia, especially localized areas have severe pollution problems. Fishes absorb contaminants directly from the water and sediment and indirectly through food web transport. Contaminant concentrations in fish tissues indicate exposure and accumulation, but effects are often sublethal and much more difficult to discern. Other indirect effects of pollution on benthic and demersal fish assemblages include ecosystem level effects of nutrient enrichment and sedimentation, which changes habitats, food, and predator assemblages. Pollution should be considered


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a stressor that reduces the resilience of benthic and demersal fish assemblages to climate changes.

Main Findings: Benthic and demersal fishes • Expected changes

o Increases in sea temperature will drive benthic and demersal fish species southwards

o The benthic and demersal fishes of Australia’s southern temperate regions are more vulnerable than the tropical fauna, partly because of the poleward-bounded geography of essential fish habitat

o Decreases in zonal westerly winds will adversely affect recruitment and production of a number of benthic and demersal fish species in southeastern Australia due to changes in ocean currents and productivity

• Observed changes within the Australian region o Many benthic and demersal fish species, and some important mega-invertebrate

species, are expanding their ranges southwards, and this is particularly evident in eastern and southeastern Australian marine domains.


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11. IMPACTS OF CLIMATE CHANGE ON PELAGIC FISHES

Alistair J. Hobday, CSIRO Marine and Atmospheric Research, Hobart [email protected]

11.1 Overview

Pelagic fishes occupy the largest ecosystem on the planet, the surface waters and water column of the 70% of the planet that comprises the open ocean. There are approximately 200 pelagic species around Australia, and while the most obvious members are the apex predators such as tunas, billfish and sharks, the mid-trophic level pelagic species, such as sardines, anchovies, and squids, are also important as food for larger marine animals. The majority of species are widely distributed: for example swordfish are found worldwide. These are valuable species economically and socially, representing high value to both commercial and recreational fishers. The expected impacts of climate change will be seen first on the distribution and abundance of pelagic species. Water temperature is a key variable with regard to distribution, and so around Australia, the range of many species will expand to the south. The impact of changes in other environmental variables, such as salinity, UV, pH and sea level are expected to be minor. For pelagic species living near regions of coastal upwelling, such as sardines, mackerel and redbait, increases in upwelling favourable winds may lead to increased local productivity, and benefit these species and their predators. Consequences for biological communities are speculative, but might include increased predation pressure due to increased presence of tropical apex predators in more southerly regions. The observed impacts of climate change are few; one of the few examples is the replacement in eastern Tasmania of cold-water jack mackerel with warm-water redbait from the East Australian Current. This is consistent with a warming trend on the east coast of Australia and Tasmania, observed in temperature records. The main impact of climate change on pelagic species will be on the distribution of the species. The impact on overall population size is less certain for many species.

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11.2 Introduction

a. Definition and distribution

The open ocean is the largest ecosystem on earth, comprising 70% of the surface of the planet. Australia is surrounded by pelagic systems, including the boundary currents (East Australian and Leeuwin Currents) and the waters extending to the outer edge of the EEZ and beyond. This vast system is often beyond the comprehension of land-based humans, yet anthropogenic impacts have been detected in declines of exploited large pelagic species (e.g. Meyers and Worm 2003, Meyers and Worm 2005, Polacheck 2006).

Pelagic species occupy the open ocean; the iconic pelagic fishes include species of tuna and billfish (swordfish and marlin) and sharks. While these high trophic level species may be the most well known, species from intermediate trophic levels are also crucial to maintenance of biodiversity in the pelagic realm. Species at the intermediate levels (e.g. sardines and anchovies, and squid) are particularly sensitive to climate impacts based on studies elsewhere in the world (e.g. Jacobsen et al. 2001, Chavez et al. 2003).

The pelagic fishes are not particularly diverse; there are approximately 260 pelagic species, out of about 12,000 marine species worldwide. Distribution is worldwide for many species. For example, there is a single species of swordfish found worldwide, while others are worldwide at the genera level (e.g. tunas, Thunnus, sardines, Sardinops). A variety of pelagic sharks (e.g. blue shark Prionace glauca) are also distributed worldwide. There is also limited diversity (few species) at intermediate trophic levels, the so-called wasp-waist level (Cury et al. 2000). At these mid-trophic levels, it is sardine and anchovy species that dominate the biomass, particularly in near-shore pelagic systems in Australia and elsewhere (e.g. Cole and McGlade 1998, Ward et al. 2006).

In general, most information on the distribution and abundance of pelagic species has come from catch records where the species are exploited (e.g. Worm et al. 2003, Zainuddin et al. 2006). Fishery-independent data on the distribution and environmental associations of the larger pelagic species has been gathered by electronic tags that record the location of the fish and limited environmental information such as temperature and depth (e.g. Arnold and Dewar 2001, Gunn and Block 2001, Block et al. 2003, Schaefer and Fuller 2003). Data from these electronic tags are being used to understand the movement and behaviour of fishes in response to climate forcing at a variety of temporal and spatial scales (e.g. Prince and Goodyear 2006, Hobday and Hartmann in press).

Biodiversity hotspots have been identified for exploited large pelagic fishes (Worm et al. 2003). In most of the hotspot regions identified there are relatively low catch rates, which led the authors to suggest that fish productivity in hotspot regions is not high. Worm et al. (2003) found instead that high catch rates occurred in areas of low diversity. The east coast of Australia was been identified as a biodiversity hotspot for large pelagic fishes, in the area between southern Queensland and Lord Howe Island (Figure 11-1). In contrast to the general pattern, however, the Australian pelagic fish hotspot is located in an area of high catch rates and fishing effort (Campbell and Hobday 2003).


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Figure 11-1: Diversity patterns in large pelagic fish predators (from Worm et al. 2003). Left hand panel: Predator diversity in the ocean predicted from Australian observer data. Color codes indicate levels of species diversity calculated by rarefaction and expressed as the expected number of species per 50 individuals. Red cells indicate areas of maximum diversity, or hotspots. The dotted lines represent 1,000-m isobaths, identifying the outer margins of continental slopes. Right hand panel: Predator diversity (species per 50 individuals) as a function of latitude for Australia.

b. Roles and ecosystem services (ecological, social, economic)

Pelagic tuna and billfish are captured in longline and purse-seine fisheries that are among the five most valuable Australian Government-managed fisheries, although all have suffered recent declines in value (ABARE 2005). For example, the gross value of production in the eastern tuna and billfish fishery fell from $68 million in 2002-03 to under $47 million in 2003-04, while in the southern and western tuna and billfish fishery it fell from $20 million to $8.3 million, and in the southern bluefin tuna wild fishery it fell from $78 million to $38 million in 2003-04 (ABARE 2005). These fisheries are regionally important as employment industries (e.g Port Lincoln South Australia and towns in coastal New South Wales). Small pelagic species, sardines, jack mackerel and redbait, and squid, are captured in lower value, but high-volume, coastal fisheries operating from a number of Australian ports. The small pelagic fishery previously harvesting jack mackerel and now dominated by redbait, off eastern Tasmania is one such high volume fishery (Welsford and Lyle 2003), the other is the South Australian pilchard fishery (http://www.afma.gov.au/fisheries/small_pelagic/default.htm). With regard to ecosystem services, top predators that are migratory or disperse widely have the potential to transfer energy across wide areas (Zainuddin et al. 2006). Thus, they may be important connections between mostly separate foodwebs. Small pelagic species also play a crucial ecosystem role; whole food chains depend on sufficient abundance of these taxa (e.g. Cole and MacGlade 1998, Cury et al. 2000).


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c. General impacts of climate change on pelagic fishes

The main impact of climate change on large pelagic species will be on species distribution. For smaller pelagic species, such as sardines and squids, the impacts may be on local productivity and abundance. In general, however, the impact on overall population size is highly uncertain. Exploring climate relationships in the pelagic realm, particularly the offshore region, will be a challenge due to the short and patchy time-series of pelagic species abundance.


The timing of migration events may be impacted if the timing of expansion or contraction of seasonal currents, such as the Leeuwin or East Australian Current, changes. For example, southern bluefin tuna have a seasonal presence along the east coast of Australia, which may be restricted further if Tasman Sea warming continues. Preliminary analyses indicate that changes may have already occurred, with fewer fish moving to the east coast in the Austral winter (Polacheck et al. 2006). Too little is known to even speculate on changes in physiology or morphology that might be driven by climate change. A single example from the tropical Atlantic and Pacific regions indicates how physiology, changes in climate, and fishing pressure may stress fish populations (Prince and Goodyear 2006). Prince and Goodyear (2006) demonstrated that large areas of cold hypoxic water in the eastern tropical Pacific (ETP) and Atlantic oceans, occurring as a result of high productivity initiated by intense nutrient upwelling, restricts the depth distribution of tropical pelagic marlins, sailfish, and tunas. The acceptable physical habitat is compressed into a narrow surface layer. This in turn makes them more vulnerable to over-exploitation by surface fishing gears. They showed that the long-term landings of tropical pelagic tunas from areas of habitat compression have been far greater than in surrounding areas. If climate change further compresses the habitat, or the area of habitat compression increases, many tropical pelagic species could be quite sensitive to increased fishing pressures. In Australia habitat compression or expansion may occur as a result of climate-forced ocean changes. In the following sections, the impact of change in particular environmental variables is considered. Sea Surface Temperature (ocean temperature) Change in ocean temperature, especially in the surface layers, will have an impact on the distribution of pelagic species. Southern bluefin tuna (T. maccoyii, SBT) are restricted to the cooler waters south of the East Australian Current and range further north when the current contracts up the New South Wales coast (Majkowski et al. 1981). This response to climate variation has allowed real-time spatial management to be used to restrict catches of SBT by non-quota holders in the east coast fishery by restricting access to ocean regions believed to contain SBT habitat (Hobday and Hartmann in press).

Rainfall/Coastal runoff/Salinity Changes in these variables are not likely to influence the pelagic species around Australia, as the change in salinity in the open ocean is expected to be negligible.


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Wind Wind indirectly impacts pelagic species through mixing of the surface waters (Cury and Roy 1989), although the direction of productivity changes is hard to predict. Pelagic regions may become more or less productive, depending on the relative balance of nutrients and light and stability for phytoplankton production (Cury and Roy 1989, Bakun and Weeks 2004). In the case of large increases in wind-forcing, upwelling could intensify and result in catastrophic consequences for coastal pelagic systems (Bakun and Weeks 2004). In such cases, as occurs in south-west Africa, intense wind-forced upwelling introduces nutrients at high levels and offshore advection is increased. Planktonic grazers have difficulty maintaining substantial populations within and adjacent to the strongly divergent upwelling zone. A massive buildup of phytoplankton biomass results, much of which sinks to the sea floor, forming thick accumulations of unoxidized organic matter and extensive areas where dissolved oxygen concentrations are very low or entirely lacking. Methane and poisonous hydrogen sulphide gas are generated within a metres-deep layer of anoxic diatom sludge and bubble back to the surface, lowering oxygen levels and resulting in large fish kills (Bakun and Weeks 2004). While this does not currently occur in Australia, it remains a possibility with future increased upwelling intensity. UV There are no known studies on the impacts of UV radiation on pelagic fishes; however, some life stages such as eggs and larvae which are close to the surface may be impacted. pH The magnitude of predicted pH change in the open ocean is small compared to existing variation and direct impacts on adult pelagic fishes are not expected. Differences in pH across the range of pelagic fishes span approximately 0.3 units, from a low of 7.9 in the intense upwelling regions to 8.1 in the oligotrophic ocean gyres (Raven et al 2005). Additional research with appropriate pH levels are needed to understand potential future impacts. Impacts may, however, occur via disruption to the lower trophic levels (see zooplankton and phytoplankton sections), or to sensitive larval stages. Again, more research is required. Mixed layer depth Some species are constrained by physiology to remain in the warmer mixed layer (e.g. skipjack tuna). The anticipated changes according the CSIRO Mk3.5 model are minor in the regions around Australia (Tony Hirst, CSIRO pers. comm), and impacts on open ocean pelagic fishes and sharks may not be detectable, or significant. If the changes in mixed layer depth do result in changes in available habitat, such as mixed layer compression, then impacts such as changes in fish catch are expected (e.g. Prince and Goodyear 2006). Sea level Sea level rise is less relevant to species of the pelagic ocean. The ocean will just get a little deeper. If the sea level rise is accompanied by fresher, less dense water, then ocean mixing may be reduced, which may have impacts on productivity of lower trophic levels.


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Of the four categories of biological impact (abundance, phenology, physiology, and community impacts), there are no studies that have considered the impact of climate variability or change on pelagic species’ phenology or physiology or community membership. In Australian pelagic systems, limited evidence of links to climate variability has been gathered for abundance and distribution changes and changes in ecosystem productivity; these are discussed further below.

The known relationships between the distribution and abundance patterns of some pelagic species on the east coast suggest that changes in the strength of the East Australian Current would have dramatic effects on the distribution of a variety of pelagic species. The seasonal expansion and contraction of the East Australian Current is linked to the local abundance of species such as yellowfin (Thunnus albacares) and bigeye tuna (T. obesus) captured in the east coast longline fishery (Campbell 1999). An increased southward penetration of the EAC may increase the suitable habitat for these species. At a finer scale the distribution of yellowfin tuna has been linked to the distribution of mesoscale environmental features such as eddies generated by the East Australian Current (Young et al. 2001). Again, an increase in the strength of the EAC may result in increased eddy formation; these are productive feeding grounds for a suite of species.

Changes in productivity can also affect the pelagic ecosystem and propagate upwards to species at the top of the food chain; research in this area is in its infancy (Young and Hobday 2004). Changes in the relative dominance of the East Australian Current and the sub-Antarctic water masses, and consequently the regional prey communities have also been implicated in changes in local productivity off the east coast of Tasmania (Young et al. 1993, 1996). For example, the disappearance of krill, Nyctiphanes australis, from the shelf ecosystem of eastern Tasmania during a warm (La Nina) event in 1989 was linked to the simultaneous disappearance of their main predator, jack mackerel (Trachurus declivis) (Young et al. 1993). Given that this krill species is at the base of most Tasmanian shelf foodwebs, and that it is a cool-water species, any persistent warming of the regional oceanography would have a profound effect on krill-dependent food chains. These food chains include cephalopods (Pecl and Jackson 2005), seabirds (Bunce 2004), small pelagic fish and tunas (Young et al. 1993).

Off the coast of Tasmania, declining growth rates of jack mackerel and a change in the age structure of the fishery catch through the 1990s may have both an environmental and an anthropogenic component (Lyle et al. 2000, Browne 2005). Changes in the fishing method (purse-seine to midwater trawl) have confounded the detection of environmental relationships (Lyle et al. 2000) that have been so clearly documented elsewhere in the world (e.g. Chavez et al. 2001, Jacobson et al. 2003).

In southern Australia, juvenile southern bluefin tuna (SBT) have been the subject of several studies investigating distribution and abundance relationships to mesoscale environmental variability (Hobday 2001, Cowling et al. 2003) or to prey (Young et al. 1996). In general, the environmental linkages to abundance are not strong at the mesoscale, although problems with the spatial resolution of some biological data have confounded analyses (Hobday et al. 2004). Seasonal changes in the abundance of juvenile SBT (ages 1-5) in southern Australia are well documented. SBT are resident along the shelf during the austral summer (Cowling et al. 2003) and then migrate south during the winter. Interannual variation in SBT abundance within the main fishing grounds in the Great Australia Bight has not been linked to the environment, although


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variation in the arrival time of schools has been attributed to unspecified environmental factors (Cowling et al. 2003). Variation in the availability of SBT prey (sardines and anchovies) as a result of changes in wind-driven upwelling (Dimmlich et al. 2004) are also likely if climate change affects the strength of upwelling favourable winds (Hertzfeld and Tomczak 1997), which might ultimately affect the pelagic predators. Finally, the impact of climate change on the winter SBT feeding grounds in the southern ocean may be more dramatic (e.g. Sarmiento et al. 2004).

Squid (e.g. Nototodarus gouldi) abundance fluctuates in southern Australia, and growth rates can have regional variation linked to ocean chlorophyll levels (e.g. Jackson et al. 2003). Regional differences in growth and age/size at maturity may be a result of local environmental differences; the populations may be adapted to local conditions (Jackson et al. 2003) which makes them vulnerable to climate change (Pecl and Jackson 2005). Due to the recent development of the squid fishery in Australia, available catch time-series are too short for demonstrating larger-scale links to processes such as ENSO, which have been shown elsewhere in the world (e.g. Jackson and Domier 2003).


Commercial fishing is the main stressor on pelagic ecosystems around Australia. Several offshore fisheries capture species that are classified as fully exploited; e.g. southern bluefin tuna, Indian and Pacific Ocean bigeye tuna (Caton and McLoughlin 2004). The status of other pelagic species, such as swordfish, is uncertain, while others are classified as underfished (e.g. yellowfin tuna). As part of these legal commercial fisheries, bycatch of other pelagic species such as sharks and sunfish remains a concern, with too little known of impacts and stresses. Illegal fishing for sharks in northern Australia remains a threat (http://www.daffa.gov.au/fisheries/iuu/overview_illegal,_unreported_and_unregulated_iuu_fishing). Offshore waters are relatively safe from other anthropogenic impacts such as pollution and habitat modification, although open ocean oil spills may occur and threaten pelagic species. Finally, the impacts of chronic oil spillage into the surface of the ocean as part of “everyday” boat operation are unknown. Overall, the impacts of climate change on pelagic species exploited by Australian fisheries remain uncertain. The limited evidence of climate variability impacts in Australian waters coupled with the greater body of evidence elsewhere in the world on the same or related species, does suggest that the impacts will be expressed first in the distribution and abundance of these generally widely distributed and mobile species, such as tuna, billfish and sharks. Impacts that result in changes in the other three biological categories - phenology and physiology, composition and interactions within communities, and structure and dynamics of communities - are even more unclear. Preliminary analyses have found spatial differences in productivity and pelagic ecosystem structure; these regional differences could mimic the temporal changes that might occur as a result of climate change (Jock Young CSIRO, unpublished data). Without a greater research effort on all four categories, an understanding of potential climate impacts on pelagic species will be limited.


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Main Findings: Pelagic Fishes • Expected changes

o Warming is likely to result in species ranges expanding southwards o Increases in wind that affect upwelling intensity may benefit some of the

mid-trophic level pelagic species o Strong seasonal and interannual changes in distribution suggest that

climate change will also have a strong impact • Observed changes within the Australian region

o Limited observations for most species o A change in species composition from jack mackerel to redbait along the

east coast of Tasmania has been attributed to warming


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12. IMPACTS OF CLIMATE CHANGE ON MARINE TURTLES

Elvira S Poloczanska & David Milton, CSIRO Marine and Atmospheric Research [email protected]

12.1 Overview

Marine mammals, turtles and seabirds are found throughout Australian waters. Some species, such as dugongs and little penguins are permanent residents in Australian waters whereas others are occasional visitors, migrating to important breeding or feeding grounds around Australia. Of the turtles, flatbacks are endemic to Australia nesting only on tropical Australian beaches. Large cetaceans such as the southern right whale and humpback whales move from summer feeding grounds off Antarctica to areas such as the Great Australian Bight and Hervey Bay to mate and calve. Of the seabirds, almost a quarter of albatross species have nesting sites on islands around Tasmania and on Macquarie Island. Some species are highly valued by coastal indigenous communities for their cultural and spiritual significance and may also be hunted for food. They also benefit local economies through ecotourism. There is some controversy regarding the extent that seabirds, cetaceans, and pinnipeds remove exploited fish species and compete with fishers. Warming is likely to be a major climate change threat to marine turtles as all stages of life history are strongly influenced by temperature. Gender of embryos is determined by ambient nest temperature and a small increase in temperature may bias the sex ratio of hatchlings towards females, potentially eliminating the production of males in certain regions. Climate change is also likely to influence phenology and distributions of all species; globally there is evidence that turtles are nesting earlier in response to warming and ranges are extending or shifting poleward. Increases in sea level will lead to substantial loss of nesting sites and inshore foraging habitats for marine turtles, particularly where coastal development constrains evolution of shorelines. The greatest threat to turtles is likely to come from climate-induced changes of food resources in their critical habitats. Green turtle populations in Australia are increasing, which have been attributed partly to the recent high frequencies of strong ENSO anomalies favouring food availability.

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12.2 Introduction

There are seven living species of marine turtles and six of these are found in Australian waters: flatback Natator depressus, green turtle Chelonia mydas, loggerhead Caretta caretta, olive ridley turtle Lepidochelys olivacea, hawksbill Eretmochelys imbricata and leatherback Dermochelys coriacea. All of these, with the exception of the flatback turtle, are listed in the IUCN 2004 Red List of threatened species. The flatback is catagorised as data deficient in the IUCN 2004 Red List but is listed as vulnerable by the Australian Environment Protection and Biodiversity Conservation Act 1999. Flatbacks are endemic to Australian tropical and subtropical coastal waters but are also found on feeding grounds in waters around Papua New Guinea and South Indonesia. The remaining species found in Australian waters have global distributions in tropical and subtropical waters. Loggerheads and leatherbacks are also found in temperate waters with large leatherbacks sighted at very high latitudes (Dutton et al. 1999, Eckhert 2002, Hays et al. 2004). Turtles use different habitats at different stages of their life cycle: natal beach, breeding grounds, internesting habitat, feeding grounds and open water pelagic systems (Environment Australia 1998). All six species found in Australian waters nest on Australian tropical and subtropical mainland and island beaches with flatbacks nesting exclusively on Australian beaches. Major loggerhead, green turtle and flatback rookeries of international significance occur in Queensland and the Torres Strait (Gyuris & Limpus 1988, Limpus et al. 1989, Parmenter & Limpus 1995). Turtles return to their natal regions to breed but not all females will nest each year (Carr et al. 1978, Limpus et al. 1983c, 1992, Miller et al. 1998, Bowen et al. 2004, James et al. 2005). After hatching, turtles undertake often extensive migrations from natal beaches to open ocean pelagic habitat where juveniles stay for to 15 years but information on this life stage is scarce (Carr 1987, Bjorndal et al. 2000b, Bowen et al. 2005), with the exception flatbacks which have no open water pelagic stage (Walker & Parmenter 1990). Young turtles then move to, generally coastal, adult feeding grounds. Turtles also exhibit site fidelty to feeding grounds and turtles from widely seperated rookeries may be found on one feeding ground (Limpus 1992, 1993, Kennett et al. 2004). Adult leatherbacks, however, tend not to move to coastal waters but remain in the open ocean foraging almost exclusively on gelatinous plankton concentrated at ocean frontal systems (Holland et al. 1990, Bels et al. 1998, Ferraroli et al. 2004). Turtles and eggs, particularly the green turtle which is prized for its meat, are harvested by coast-dwelling indigenous communities in subsistence fisheries (Limpus et al. 1983a,b, 1989, Marsh 1996, Collins et al. 1996). Marine turtles are also highly valued by indigenous people for their cultural and spiritual significance (Marsh 1996). Marine turtles may provide economic benfit to local economies through tourism and with the rise of ecotourism and establishment of educational visitor centres, such as in the Mon Repos Conservation Park, Queensland and the Jurabi Turtle Centre in Western Australia, contribute to raising public awareness of marine conservation issues (see Tisdell & Wilson 2002). Green turtles and dugongs are the largest herbivores in Australian coastal waters and through selective grazing, play an important role in structuring sea grass communities (Aragones 2000, Aragones & Marsh 2000, André et al. 2005, Moran & Bjorndal 2005). Hawksbills, which prey on sponges and other invertebrates around coral reefs, play a


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similar role in structuring coral reef communities (Meylan 1988, Hill 1998, León & Bjorndal 2002). By feeding on select sponges that are competitively superior to corals, hawksbill grazing helps maintains coral diversity (Hill 1998). Green turtles enhance seagrass biodiversity and productivity through selective cropping (Brand-Gardner et al. 1999, Aragones 2000, André et al. 2005, Moran & Bjorndal 2005). Green turtles may also play an important role in nutrient cycling within seagrass systems, short-circuiting the detrital cycle through regular grazing and faecal production so nutrient turnover rate is increased substantially (Thayer et al. 1982, Aragones 2000). Green turtles transport nutrients from seagrass beds to adjacent ecosystems such as coral reefs through faecal production and other processes (Thayer et al. 1982). Marine turtles are predicted to be very sensitive to rapid global warming because of their life history characteristics: namely slow growth rate, temperature dependant sex determination in the nest, and natal homing (Davenport 1997). Marine turtles are very old species that have survived for millions of years and endured past shifts in climate (see Dutton et al. 1999, Reece et al. 2005) but over-exploitation, human disturbance and habitat loss have greatly reduced populations comprising population resilience to a changing environment (Jackson et al. 2001, Lewison et al. 2004, Ferraroli et al. 2004, Kaplan 2005).


As marine turtles move long distances and utilise different habitats at different stages during their life cycle they are subject to a wide range of environmental influences so may be seriously impacted by climate change (see Robinson et al. 2005 for a review). Temperature

Temperature is likely to be one of the biggest climate change threat to sea turtles, all stages of marine turtle life history are influenced strongly by temperature (Davenport 1997). All marine turtles rely on ambient nest temperatures for gender determination. In a nesting year, females will lay multiple clutches digging nests above high water line on sandy beaches, the critical period for sex differentiation is during the middle third of the incubation period when thermosensitive development of gonads occur (Yntema & Mrosovsky 1982, Hewavisenthi & Parmenter 2002). Small changes in temperature close to the pivotal temperature (~29°C) at which a 50:50 sex ratio is produced, can skew the sex ratio of hatchlings with higher temperatures producing more females (Yntema & Mrosovsky 1982, Godfrey et al. 1999, Booth & Astill 2001b). Many nesting beaches already have a strong female bias (Godfrey et al. 1996, Binckley & Spotila 1998, Godley et al. 2002b, Hays et al. 2003a, Glen & Mrosovsky 2004) so if temperatures rise, the production of males may quickly be eliminated. Inter-beach thermal variations at rookeries ensure some beaches are producing males (Hays et al. 2003a), the occurrence of light-coloured (cooler) beaches within nesting regions is important for population persistence. In southern Queensland, the lighter-coloured sand of beaches on offshore coral cays and islands is important for producing males while the warmer mainland beaches tend to be heavily female-biased (Environment Australia 1998). Within Australia, rookeries on northern beaches generally produce more females and cooler southern beaches more males. For loggerheads, beaches south of Fraser Island probably only produce males (Environment Australia 1988). If temperatures rise on currently male-producing beaches then the gross skewing in sex bias may lead to


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extinction of local nesting populations. Further, extreme high temperatures will lead to heat stress of embryos and a decline on hatching success (Broderick et al. 2001). The seasonal pattern of nesting in turtles may be driven by environmental temperature (Godley et al. 2002a). There is already evidence that marine turtles in Florida, USA, are nesting earlier in response to increasing sea surface temperatures (Weishampel et al. 2004). As well as an earlier onset of nesting, increased sea temperatures also reduce the interval length between the multiple clutches laid within a nesting season (Sato et al. 1998, Hays et al. 2002). Incubation duration, hatchling survival and hatchling weight are also influenced by incubation temperatures (Booth & Astill 2001a, Hewavisenthi & Parmenter 2001, Godley et al. 2002b). Adult distribution is limited by minimum temperatures around 15-20°C for most species (Davenport 1997) although leatherbacks and loggerheads penetrate far colder waters. Large leatherbacks are reported from waters as cool as 8°C but juvenile leatherbacks (<100cm carapace length) are rarely found in waters less than 26°C (Eckert 2002). Increased temperatures may extend ranges of turtles restricted to tropical and subtropical waters further south, although green turtles and hawksbills, which forage in tropical/subtropical seagrass beds and coral reefs respectively, are likely to be restricted by food availability. Range shifts in marine turtles are already being noted in the northern hemisphere with an increasing number of turtles being recorded from UK shores. Currents

Transport by large ocean currents play a major role in the movements of hatchlings and early juveniles to ocean pelagic nursery habitats where young turtles exploit biologically rich environments linked to current systems and convergence zones (Carr 1987, Witherington 2002, Ferraroli et al. 2004). However, larger juveniles and adults have been tracked swimming against prevailing currents so may use current flows opportunistically to facilitate transport (Luschi et al. 2003, Polovina et al. 2004). Juvenile loggerheads originating from Australian populations have been identified from feeding grounds off Baja California, representing a journey that crosses the entire Pacific Ocean most likely aided by north Pacific current (Bowen et al. 1995). Adult loggerheads and leatherbacks forage at fronts and eddies and are associated with major currents (Ferraroli et al. 2004, Polovina et al. 2004). It must be supposed that alteration of major current systems may challenge the navigational abilities of marine turtles, deflect turtle movements and change availability of planktonic prey that concentrate at oceanographic features so resulting in range shifts and changes in abundance (Luschi et al. 2003, Robinson et al. 2005). Mixed Layer Depth

Little is known about the pelagic stage of turtles life history but this may last for up to 15 years, if not more (Carr 1987). At this time the turtles are foraging in surface and subsurface waters around oceanic fronts for mainly gelatinous plankton (Carr 1987, Polovina et al. 2000, Witherington 2002). Adult leatherbacks and loggerheads also concentrate at and move along oceanic fronts where their main prey, gelatinous plankton, are concentrated (Ferraroli et al. 2004, Polovina et al. 2004). Alterations of primary productivity driven by changes in mixed layer depth will impact on food availability for turtles in the open oceans.


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Sea Level

Sea level rise will be a major threat for turtle breeding beaches, particularly on beaches where coastal development acts as a barrier constraining landward movement of beaches or hindering natural accretion of beach material and evolution of beach morphology (Fish et al. 2005). Losses of nesting beaches may be substantial for a small rise in sea level (Fish et al. 2005). Green turtles are herbivores feeding in inshore shallow waters on seagrass beds and algae and these areas may also be at high risk from sea level rise (Duarte 2002, also see Seagrass chapter). Cyclones and Storms

Turtles tend to nest just above the high water mark so storm surges, exceptionally high tides and heavy rainfall can inundate nests. This can result in drowning of embryos within eggs or erode sand dunes exposing the egg clutches to the elements and to predators (Milton et al. 1994). High numbers of stranded dying and dead turtles have been recorded after cyclones and large storm surges which can cause damage, stress, starvation and death (Limpus & Reed 1985, Carr 1987). Storms and cyclones will also erode or damage nesting beaches as well as coastal feeding grounds such as seagrass meadows (see Seagrass chapter). UV radiation

Although there is presently no evidence this will impact on turtles directly, it will affect prey items that turtles feed on, such as seagrasses and algae (Dawson & Dennison 1996, El-Sayed et al. 1996, Häder et al. 1998). Increased UV radiation will also impact on plankton communities, altering community diversity and composition (El-Sayed et al. 1996), these effects will be transmitted through food webs to marine turtles. Being air breathers, turtles spend considerable periods of time at the sea surface so it is expected they may incur UV radiation-induced damage.

12.4 Observed impacts

There is evidence of forcing by climate on turtle reproduction (discussed below), and it has been suggested that an increase in green turtle populations in the southern Great Barrier Reef is due to an increase in the frequency of El Niño-Southern Oscillation (ENSO) anomalies resulting in environmental conditions favourable for breeding success (Chaloupka & Limpus 2001). Reports from the northern hemisphere indicate marine turtle populations may already be responding to increasing temperatures. From records over the past century, most sightings of marine turtles in UK waters have been in the past 40 years and sightings are increasing suggesting a shift or expansion in distributions (Robinson et al. 2005). There is compelling evidence of climate change effects on turtle phenology with loggerhead turtles in Florida, USA, nesting progressively earlier concurrent with an increase in sea surface temperatures (Weishampel et al. 2004). The incidence of disease in marine turtles has drastically increased over the last few decades which may be indirectly related to warming temperatures although pollution is likely to be a major factor (see section 4b below, Robinson et al. 2005).


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Distribution/Abundance/Phenology

Synchronicity observed in green turtle reproduction at widely seperated rookeries suggests large-scale environmental forcing (Limpus & Nicholls 1988, Chaloupka 2001). Interannual fluctuations in numbers of green turtles nesting in rookeries within the Great Barrier Reef were positively correlated with the Southern Oscillation, which will influences temperature and rainfall in the central and eastern Pacific, with a lag of two years (Limpus & Nicholls 1988). The lag may represent the time required to acquire fat reserves for reproduction and migration (Limpus & Nicholls 1988). Variation in winter sea surface temperature anomalies partly explains inter-nesting intervals of a Costa Rican population of green turtles with large 2-year re-migration probabilities related to warm winter sea surface temperatures (Solow et al. 2002). Modelling studies suggest breeding intervals (time between nesting years) are determined by resource provisioning on adult feeding grounds and potential reproductive success (Hays 2000, Rivalan et al. 2005). The frequency of strong ENSO anomalies is predicted to increase under climate change scenarios, which may favour green turtle breeding. Synchronicity with ENSO has not been observed for other species of marine turtles but it has been suggested that as herbivores, coupling with the environment is tighter for green turtles (Broderick et al. 2001). Increases in air temperatures expected to occur over Australia in the next hundred years could drastically affect the sex ratio of turtle hatchlings. No long-term trends have been detected where records of sex ratios are available so far (Chaloupka & Limpus 2001). Annual monitoring of the sex-ratio of hatchlings from a population of freshwater painted turtles (Chrysemys picta) in the USA revealed sex ratio was highly correlated with July air temperature (Janzen 1994). An increase of 4°C, which is predicted to be highly likely over the next 100 years, would eliminate production of males in this population. What is the future for marine turtle populations? Reconstruction of incubation temperatures over past 150 years at a green turtle rookery on Ascension Island (South Atlantic) revealed a warming of nests, further warming may potentially jeopardise male production at cooler nest sites (Hays et al. 2003). One possible strategy to counteract rising temperatures would be earlier nesting so temperatures during the critical sex determining period would still produce both sexes (Shine 1999). Median nesting date was found significantly correlated with May sea surface temperatures in a population of loggerheads nesting in Florida, USA (Weishampel et al. 2004). Over a 15 year period, median nesting date became earlier by 10 days so sea turtle phenology is already shifting with climate.

Consequences for Biodiversity

Marine turtles are important for maintaining biodiversity in seagrass and coral reef ecosystems. Regular cropping of seagrasses by green turtles increases productivity and improves nutritional attributes of seagrasses (Aragones 2000, Aragones & Marsh 2000). This is beneficial not only for the turtles but also for other herbivores within the seagrass system and is particularly important where there are no other large herbivores or periodic natural disturbances such as cyclones (Aragones 2000). For example, recent die-offs of seagrass beds in Florida waters has been attributed to the ecological extinction of green turtles (Jackson et al. 2001). The large-scale removal of green turtles by historical overfishing led to increased deposition of plant material and


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subsequently increased hypoxia and an increase in seagrass diseases. Given the high levels of anthropogenic-induced threats to marine ecosystems, further reductions in turtle populations should be avoided.

12.5 Threats other than climate change

The major threats to marine turtles breeding in Australia, other than climate change, have been identified as bycatch in fisheries; unsustainable levels of harvest by neighbouring countries, predation of eggs by feral animals, coastal development and habitat loss, marine debris and pollution (Environment Australia 1998). Fishing

Commercial fishing operations, such as shrimp trawling and long-lining, kill many thousands of marine turtles every year (Robins 1995, Poiner et al. 1996, Hays et al. 2003b, Kaplan 2005). As air-breathers turtles are drowned in nets, they may also be captured alive and killed by crew, be landed and sold, be injured or killed on release from nets, be injured or killed by associated gear and ships propellers (Robins 1995, Cheng & Chen 1997, Chaulopka et al. 2004b). Strikes by both commercial and recreational vessels kill large numbers of turtles in coastal waters (Hazel & Gyuris 2006). The worldwide decline in leatherback turtles is largely attributed to high mortality rates caused by interactions with fisheries (Ferraroli et al. 2004). In some parts of the world turtles, mainly green turtles and olive ridley turtles, are themselves the focus of fisheries for their meat, and eggs are also commercially harvested from nests (Poiner et al. 1996, Kaplan 2005). Hawksbills are valued for their mottled carapaces which are marketed as ‘tortoiseshell’ (Limpus & Miller 1990) but international trade in tortoiseshell is now being halted. Fishing mortality may be high in areas of the Indo-Pacific, for example in Indonesia it is estimated more than 70,000 turtles are taken annually (Poiner et al. 1996). Large mortality rates generated as bycatch in commercial fisheries targeting other species (Robins 1995, Poiner et al. 1996). As turtles nesting on Australian beaches turtles make long distance migrations throughout the Indo-Pacific (Limpus et al. 1992) fishing-induced mortality potentially threatens Australian populations. There is some dispute as to whether green turtle populations are endangered, many local populations are under threat but globally this species is increasing (Broderick et al. 2006). Green turtle populations around the world are showing strong recovery following cessation or reduction of harvesting (Bjorndal et al. 2000a, Chaloupka & Limpus 2001, Balazs & Chaloupka 2004, Broderick et al. 2006).

Other Anthropogenic factors

Marine turtles have complex life histories involving long migrations so face a variety of threats at different stages of their life (Bowen et al. 2005). Coastal development is a major threat to breeding sea turtles with the disappearance of nesting beaches, disturbance by humans and artificial lighting all reducing reproductive success (Lohmann & Lohmann 1996, Tuxbury & Salmon 2005, Ozdilek et al. 2006). Shading of beaches by artificial structures during nesting or changes in beach sediments will alter sex ratio of hatchlings (Mrosovsky et al. 1995, Hays et al. 2001).


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Pollution is a major threat for all turtle species. All species are prone to eating plastics and other anthropogenic debris, particularly plastic bags and plastic ropes that are mistaken for jellyfish (Carr 1987, Witherington 2002). Even a small amount of ingested debris can kill turtles obstructing the digestive gut (Bugoni et al. 2001). Sub-lethal effects, such as dietary dilution when non-nutritive debris replaces nutritious food in the gut, affects growth and reproduction and may lead to long term impacts on marine turtle populations (Carr 1987, McCauley & Bjorndal 1999). Fibropapillomatosis is a disease caused by a herpesvirus that causes tumours to grow both internally and externally and may eventually lead to death (Jones 2004). This disease was first documented in 1930s and was rare until early 1980s but now is at epidemic proportions in many turtle populations (Jones 2004). It is common in green turtles, olive ridley turtles and loggerheads but has yet to be recorded in hawksbills (Adnyana et al. 1997, Jones 2004). Young mature turtles are most frequently affected and affected turtles tend to have multiple tumours (Adnyana et al. 1997, Foley et al. 2005). The prevalence of the tumours in young turtles suggests the turtles are infected when they recruit to nearshore waters, the disease is common in regions densely populated by humans suggesting prolonged exposure to anthropogenic pollutants may be responsible (Jones 2004, Ene et al. 2005, Foley et al. 2005). Further, phylogeny of the viruses indicates that outbreaks are unlikely to be caused by virulent mutations of the virus but are more likely to be the result of environmental or ecological factors (Herbst et al. 2004). The increase of this disease since the 1980s coincides with rising temperatures recorded over the last few decades so may also be indirectly related to climate change (Robinson et al. 2005).

Main Findings: Turtles • Expected changes

o Warming will expand or shift distributions southwards o Nesting and feeding habitats will be lost or altered o Nesting may occur earlier o Sex ratios of turtle hatchlings will become skewed towards females o Alteration of currents will impact on distributions, migration and foraging o Reproductive success will increase/decrease

• Observed changes within the Australian region o Evidence of large-scale climate forcing on reproductive success on turtles o Strong female bias in hatchling sex ratio associated with high SSTs

recorded at some turtle rookeries o Breeding success related to SST in certain turtle populations


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13. IMPACTS OF CLIMATE CHANGE ON SEABIRDS

Anthony Richardson, Elvira S Poloczanska & David Milton, CSIRO Marine and Atmospheric Research [email protected]

13.1 Overview

Seabirds are birds adapted to life in the marine environment. Most species nest in colonies varying in size from a few dozen to many millions of individuals. They occur around Australia, and feed both at the ocean's surface and below it. Seabirds can be highly pelagic, coastal, or spend a part of the year away from the sea entirely. Many also undertake long annual migrations. A diverse seabird fauna breeds on mainland and island coastlines around Australia. The world’s largest colony of crested terns (Sterna bergii) occurs in the Gulf of Carpentaria in Australia’s north (Walker 1992), and almost a quarter of all albatross species have nesting sites on islands around Tasmania and on Macquarie Island. Planktivorous seabirds nesting in Australia are found only in the southern waters but can occur in high numbers. For example, an estimated 23 million short-tailed shearwaters (Puffinus tenuirostris) nest in south-east Australia every year (Ross et al. 2001). Some species of seabirds are highly valued by coastal indigenous communities for their cultural and spiritual significance and may also be hunted for food. They also benefit local economies through ecotourism; colonies of little penguins on Phillip Island near Melbourne bring in millions of dollars annually for local communities. Seabirds are efficient integrators of ecosystem health, as many feed on small pelagic fish and zooplankton and thus are sensitive to changes at lower trophic levels. They are frequently used as indicators of the state of the marine environment.

13.2 Introduction

Definition and distribution Seabirds vary greatly in their lifestyle, behaviour and physiology, but often have strikingly similar adaptations to cope with the same environmental problems and feeding niches. Seabirds live longer, breed later and have fewer young than their terrestrial counterparts, but they invest a great deal of time in the young they do have. Most species nest in colonies varying in size from a few dozen to many millions of individuals. They feed both at the ocean's surface and below it, and some even feed on each other. Seabirds can be highly pelagic, coastal, or in some cases spend a part of the year away from the sea entirely. Many also undertake long annual migrations, crossing the equator or circumnavigating the Earth.

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A diverse seabird fauna breeds on mainland and island coastlines around Australia. For example, the Houtman Abrolhos Islands on the west coast are an important nesting area for Australian seabirds in terms of biomass and species diversity (Ross et al. 2001). The world’s largest colony of crested terns (Sterna bergii) occurs in the Gulf of Carpentaria in Australia’s north (Walker 1992). Almost a quarter of albatross species have nesting sites on islands around Tasmania and on Macquarie Island. Planktivorous seabirds nesting in Australia are found only in the southern waters but can occur in high numbers. For example, an estimated 23 million short-tailed shearwaters (Puffinus tenuirostris) birds nest in south-east Australia every year (Ross et al. 2001). Roles and ecosystem services (ecological, social, economic)

Some species of seabirds are highly valued by coastal indigenous communities for their cultural and spiritual significance and may also be hunted for food. They also benefit local economies through ecotourism. For example, the colonies of little penguins on Phillip Island near Melbourne are a popular tourist attraction bringing in millions of dollars annually.

Seabirds are efficient integrators of ecosystem health, as many feed on small pelagic fish and zooplankton and thus are sensitive to changes at lower trophic levels. Seabirds are considered useful indicators of the state of fish stocks (Cairns 1987, 1992).

General impacts of climate change on seabirds

Climate change is also likely to influence phenology and distributions of all species; globally there is evidence that birds are nesting earlier in response to warming and ranges are extending or shifting poleward. Increases in sea level will lead to substantial loss of nesting sites and inshore foraging habitats for some seabirds, particularly where coastal development constrains evolution of shorelines. The greatest threat to seabirds is likely to come from climate induced changes of food resources in their critical habitats. Australasian gannet populations in Australia are increasing, which has been attributed partly to the recent high frequencies of strong ENSO anomalies favouring food availability. For many other species, such as the wedge-tailed shearwater in the Great Barrier Reef, reproductive success will decline as warming reduces prey availability. Alteration of prey availability will ultimately affect distributions, abundance, migration patterns and community structure of higher trophic levels.


Sea surface temperature Although larger marine species, such as seabirds and marine mammals that can range over many thousands of kilometres, may be able to rapidly shift their distributions with climate change, many are restricted by habitat requirements at particular life stages such as availability of nursery areas, feeding grounds or breeding grounds. The relationship between temperature and fish production and distribution is apparent over the decadal time scale where oceanographic (temperature and productivity) regime shifts regulate zooplankton biomass, and fisheries catches and seabird abundances


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(Beamish et al. 1997, Mantua et al. 1997, McGowan et al. 1998, Beamish et al. 1999, Koslow et al. 2002). Other climate stressors There is a variety of other climate stressors that regulate bird population dynamics. Rising sea-level is a threat to bird species that nest on low-lying coastal areas. Australian examples are the Little Kingfisher Alcedo pusilla pusilla and Spangled Drongo Dicrurus bracteatus carbonarius nesting on low lying islands in the Torres Strait, and the Lesser Noddy Anous tenuirostris melanops that nests in mangroves (Garnett & Crowley 2000). Cyclones can cause catastrophic destruction of breeding colonies and death of individuals for northern birds such as the Lesser Noddy Anous tenuirostris melanops (Garnett & Crowley 2000). In tropical Australia, the highly cyclical rainfall, with practically all rain falling during the wet season, influences the timing of breeding of birds (Garnett & Crowley 2000, Whitehead & Saalfeld 2000).

13.4 Observed impacts of climate change (or relationships)

Many seabirds are restricted to coastal and shelf waters during the breeding season when adults forage in waters adjacent to nesting sites. Changes in ocean temperatures and climatic conditions around Western Australia are responsible for observed range expansions of the Red-tailed Tropicbird Phaethon rubricauda, Bridled Tern Sterna anaethus and the Roseate Tern Sterna dougallii (Dunlop and Wooller 1986). These range expansions may compromise the habitats of the southerly cool-water species (Dunlop and Wooller 1990). The establishment of nesting colonies are constrained by factors such as temperature that will determine prey availability and the distributions of breeding adults foraging in coastal waters (Jaquemet et al. 2005, Weimerskirch et al. 2005). Wedge-tailed shearwaters Puffinus pacificus are only found over waters with surface temperatures exceeding 20°C minimum (Surman & Wooller 2000). The recent growth of nesting colonies of wedge-tailed shearwaters in south-western Australia may be due to a southerly movement from more northerly colonies as temperatures rise (Bancroft et al. 2004). The population of Australasian gannets Morus serrator that breed in south-east Australia has increased by approximately 6% per year since 1980 with new breeding sites being established as nesting space becomes limited (Bunce et al. 2002). This increase appears to be associated with a long-term warming trend and a concurrent increase in the abundance of small pelagic prey fish, principally pilchards Sardinops sagax. There are also likely to be marked changes in the phenology or timing of lifecycle events of seabirds. Mean egg laying dates of many bird species around the world have advanced considerably in response to increasing temperatures (Archaux 2003, Both et al. 2004, Both et al. 2005, Moller et al. 2006). Migratory species are arriving earlier and leaving later (Mason 1995, Crick et al. 1997, Lehikoinen et al. 2004, Marra et al. 2005, Jonzén et al. 2006). Most evidence is from the Northern Hemisphere but a similar pattern has recently been found in Australian migratory wetland birds such as the curlew sandpiper Calidris ferruginea and the double-banded plover Charadrius bicinctus (Chambers 2005, Beaumont et al. 2005). It is assumed that similar changes are occurring in Australian seabirds. Protracted breeding seasons and southward range expansions observed in some of seabird species in Western Australia are likely to be a response to changing climate (Dunlop & Wooller 1986, also see Chambers et al. 2005).

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Breeding success of Little Penguins Eudyptula minor in the Bass Strait, south-eastern Australia are correlated with sea temperatures and mean laying dates are earlier in warmer years (Chambers 2004). Warmer waters may also be causing Short-tailed Shearwaters Puffinus tenuirostris to fly greater distances for food for chick provisioning (Olsen et al. 2003). Abnormally warm SSTs on the Great Barrier Reef during the summer of 2002 led to reduced provisioning for chicks and reproductive failure in the Wedge-tailed Shearwater Puffinus pacificus (Smithers et al. 2003).


There are several other stressors that act synergistically with the direct effect of climate changes on birds. Overfishing small pelagic fish species can lead to a reduction in seabird numbers or reduced breeding success (Crawford 1991, Furness & Tasker 2000, Furness 2002). Longlining can also kill large numbers of seabirds such as albatrosses attracted to the bait, although modern longlining techniques are minimising these losses. Changes is discarding by fisheries can impact seabird populations, discards from fisheries are a key component of the diet in many seabird species globally (Votier et al. 2004). The indirect effect of climate change on the productivity of lower trophic levels is one of the most important factors controlling seabird populations. Finally, loss or changes of intertidal fauna and flora will also have repercussions for terrestrial species such as wading shorebirds for which the intertidal zone provides valuable forage (Bradley & Bradley 1993, Kendall et al. 2004).

Main Findings: Seabirds • Expected changes

o Warming will expand or shift distributions southwards o Nesting and feeding habitats will be lost or altered o Nesting may occur earlier o Alteration of currents will impact on distributions, migration and foraging o Reproductive success will increase/decrease

• Observed changes within the Australian region o Breeding success related to SST in certain seabird populations o Earlier laying dates recorded in some seabird populations o Evidence of distributional changes in tropical seabird populations


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14. REFERENCES

References are arranged by review section. There are a total of 13 review sections.

14.1 Phytoplankton

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Antia, A.N., Koeve, W., Fischer, G. Blanz, T., Schul-Bull, D., Scholten, J., Neuer, S., Kremling, K., Kuss, J., Peinert, R., Hebbeln, D. Bathmann, U., Conte, M., Fehner, U. 2001. Basin-wide particulate carbon flux in the Atlantic Ocean: Regional export patterns and potential for atmospheric CO2 sequestration. Global Biogeochemical Cycles 15 (4), 845-862.

Barbieri, E.S., Villafane, V.E. & Helbling, W. 2002. Experimental assessment of UV effects on temperate marine phytoplankton when exposed to variable radiation

regimes. Limnology and Oceanography 47, 1648-1655. Beaugrand, G. 2004. The North Sea regime shift: evidence, causes, mechanisms and

consequences. Progress in Oceanography 60, 245-262. Behrenfeld, M., Hardy J., Gucinski H., Hannemann, A., Lee, H. II, & Wones A. 1993.

Effects of Ultraviolet-B radiation on primary production in the South Pacific Ocean. Marine Environmental Research 35, 349-363.

Blesinger, S., Wilhelm, C., & Kesselmeier, J. 1994. Exchange of reduced sulfur compounds between marine phytoplankton and the ocean. In The Proceedings of EUROTRAC Symposium ’94, P.M. Borrel (ed.), The Hague, 559-562.

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Fernández-Torquemada, Y. & Sánchez-Lizaso, J.L 2005. Effects of salinity on leaf growth and survival of the Mediterranean seagrass Posidonia oceanica L.: Delile. Journal of Experimental Marine Biology and Ecology 320, 57-63.

Fonseca, M.S. & Bell, S.S. 1998. Influence of physical setting on seagrass landscapes near Beaufort, North Carolina, USA. Marine Ecology Progress Series 171, 109-121.

Fonseca, M.S. & Cahalan, J. A. 1992 A preliminary evaluation of wave attenuation by four species of seagrass. Estuarine, Coastal and Shelf Science 35, 565-576.

Gacia, E. & Duarte, C.M. 2001. Sediment retention by a mediterranean Posidonia oceanica meadow, the balance between deposition and resuspension. Estuarine Coastal and Shelf Science 52, 505-514.

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Gattuso, J.P., Frankignoulle, M. & Wollast, R. 1998. Carbon and carbonate metabolism in coastal aquatic ecosystems. Annual Review of Ecology and Systematics 29, 405-434.

Gillanders, B.M. & Kingsford, M.J. 2002. Impact of changes in flow of freshwater on estuarine and open coastal habitats and the associated organisms. Oceanography and Marine Biology: an Annual Review 40, 233-309.

Gray, C.A., Chick, R.C. & McElligott, D.J. 1998. Diel changes in assemblages of fishes associated with shallow seagrass and bare sand. Estuarine Coastal and Shelf Science 46, 849-859.

Häder, D.P., Kumar, H.D., Smith, R.C. & Worrest, R.C. 1998. Effects on aquatic ecosystems. Journal of Photochemistry and Photobiology B 46, 53-68.

Haywood, M.D.E. Vance, D.J. & Loneragan, N.R. 1995. Seagrass and Algal Beds as Nursery Habitats for Tiger Prawns Penaeus semisulcatus and P. esculentus in a Tropical Australian Estuary. Marine Biology 122, 213-223.

Heck, K.L., Hays, G. & Orth, R.J. 2003. Critical evaluation of the nursery role hypothesis for seagrass meadows. Marine Ecology Progress Series 253, 123-136.

Hemminga, M.A., Marba, N. & Stapel, J. 1999. Leaf nutrient resorption, leaf lifespan and the retention of nutrients in seagrass systems. Aquatic Botany 65, 141-158

Hindell, J.S. 2006. Assessing the trophic link between seagrass habitats and piscivorous fishes. Marine and Freshwater Research 57, 121-131.

Inglis, G.J. & Smith, M.P.L. 1998. Synchronous flowering of estuarine seagrass meadows. Aquatic Botany 60, 37-48.

Invers, O. Romero, J. & Pérez, M. 1997. Effects of pH on seagrass photosynthesis, a laboratory and field assessment. Aquatic Botany 59, 185-194.

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Jernakoff, P., Brearley, A. & Nielsen, J. 1996. Factors affecting grazer-epiphyte interactions in temperate seagrass meadows. Oceanography and Marine Biology: an Annual Review 34, 109-162.

Jernakoff, P. & Nielsen, J. 1997. Plant-animal associations in two species of seagrasses in Western Australia. Aquatic Botany 60, 359-376.

King, R.J., Adam, P. & Kuo, J. (1990) Seagrasses, Mangroves and Saltmarsh Plants. In Biology of Marine Plants (Eds. M.N. Clayton & R.J. King) Longman Cheshire Pty Ltd, Melbourne, Australia, 213-239.

Kirkman, H. 1997. Seagrasses of Australia, Australia: State of the Environment Technical Paper Series (Estuaries and the Sea), Department of the Environment, Canberra, Australia. 36pp.

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Lee Long, W.J.L., Mellors, J.E. & Coles, R.G. 1993. Seagrasses between Cape York and Hervey Bay, Queensland, Australia. Australian Journal of Marine and Freshwater Research 44, 19-31.

Lemmons, J.W.T.J., Clapin, G., Lavery, P. & Cary, J. 1996. Filtering capacity of seagrass meadows and other habitats of Cockburn Sound, Western Australia. Marine Ecology Progress Series 143, 187-200.

Loneragan, N.R., Kenyon, R.A., Staples, D.J., Poiner, I.R. & Conacher, C.A. 1998. The influence of seagrass type on the distribution and abundance of postlarval and juvenile tiger prawns Penaeus esculentus and P. semisulcatus in the western Gulf of Carpentaria, Australia. Journal of Experimental Marine Biology and Ecology 228, 175-195.

Loneragan, N.R., Kenyon, R.A., Haywood, M.D.E. & Staples, D.J. 1994 Population-Dynamics of Juvenile Tiger Prawns (Penaeus esculentus and P. semisulcatus) in Seagrass Habitats of the Western Gulf of Carpentaria, Australia. Marine Biology 119, 133-143.

Longstaff, B.J. & Dennison, W.C. 1999. Seagrass survival during pulsed turbidity events, the effects of light deprivation on the seagrasses Halodule pinifolia and Halophila ovalis. Aquatic Botany 65, 105-121.

Longstaff, B.J., Loneragan, N.R., O'Donohue, M.J. & Dennison, W.C. 1999. Effects of light deprivation on the survival and recovery of the seagrass Halophila ovalis (RBr). Hook. Journal of Experimental Marine Biology and Ecology 234, 1-27.

Lourey, M.J., Dunn, J.R. & Waring, J. 2006. A mixed-layer nutrient climatology of Leeuwin Current and western Australian shelf waters: Seasonal nutrient dynamics and biomass. Journal of Marine Systems 59, 25-51.

Marbà, N. & Walker, D.I. 1999. Growth, flowering, and population dynamics of temperate Western Australian seagrasses. Marine Ecology Progress Series 184, 105-118.

Marsh, J.A., Dennison, W.C. & Alberte, R.S. 1986. Effects of temperature on photosynthesis and respiration in eelgrass Zostera marina (L.) Journal of Experimental Marine Biology and Ecology 101, 257-267.

Masini, R.J. & Manning, C.R. 1997. The photosynthetic responses to irradiance and temperature of four meadow-forming seagrasses. Aquatic Botany 58, 21-36.

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Pergent, G., Rico-Raimondino, V. & Pergent-Martini, C. 1997. Fate of primary production in Posidonia oceanica meadows of the Mediterranean. Aquatic Botany 59, 307-321.

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Short, F.T. & Wyllie-Echeverria, S. 1996. Natural and human-induced disturbance of seagrasses. Environmental Conservation 23, 17-27.

Skilleter, G.A., Cameron, B., Zharikov, Y., Boland, D. & McPhee, D.P. 2006. Effects of physical disturbance on infaunal and epifaunal assemblages in subtropical, intertidal seagrass beds. Marine Ecology Progress Series 308, 61-78.

Smith, K.A. & Sinerchia, M. 2004. Timing of recruitment events, residence periods and post-settlement growth of juvenile fish in a seagrass nursery area, south-eastern Australia. Environmental Biology of Fishes 71, 73-84.

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Soluna, A., Sekiguchi, Y. & Nakata, K. 2004. Modeling and evaluating the ecosystem of sea-grass beds, shallow waters without sea-grass, and an oxygen-depleted offshore area. Journal of Marine Systems 45, 105-142.

Udy, J.W. & Dennison, W.C. 1997. Growth and physiological responses of three seagrass species to elevated sediment nutrients in Moreton Bay, Australia. Journal of Experimental Marine Biology and Ecology 217, 253-277.

Udy, J.W., Dennison, W.C., Lee Long, W.J., & McKenzie, L.J. 1999. Responses of seagrass to nutrients in the Great Barrier Reef, Australia. Marine Ecology Progress Series 185, 257-271.

Uhrin, A.V. & Holmquist, J.G. 2003. Effects of propeller scarring on macrofaunal use of the seagrass Thalassia testudinum. Marine Ecology Progress Series 250, 61-70.

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Walker, D.I & Prince R.I.T. 1987. Distribution and biogeography of seagrass species on the northwest coast of Australia. Aquatic Botany 29, 19-32.

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Watson, R.A., Coles, R.G. & Lee Long, W.J. 1993. Simulation estimates of annual yield and landed value for commercial penaeid prawns from a tropical seagrass habitat, northern Queensland, Australia. Australian Journal of Marine and Freshwater Research 44, 211-219.

Williams, S.L. 1988. Disturbance and recovery of a deep-water Caribbean seagrass bed. Marine Ecology Progress Series 42, 63-71.

Zieman, J.C. 1976. The ecological effect of physical damage from motor boats on turtle grass beds in southern Florida. Aquatic Botany 2, 127-139.

Zann, L. 2000. The Eastern Australian region, a dynamic tropical/temperate biotone. Marine Pollution Bulletin 41, 188-203.

14.4 Mangroves

Alongi, D.M. 2002. Present state and future of the world’s mangrove forests. Environmental Conservation 29, 331-349.

Badola, R. & Hussian, S.A. 2005. Valuing ecosystem functions: an empirical study on the storm protection function of Bhitarkanika mangrove ecosystem, India. Environmental Conservation 32, 85-92.

Blasco, F., Saenger, P. & Janodet, E. 1996. Mangroves as indicators of coastal change. Catena 27, 167-178.

Bloomfield, A.L. & Gillanders, B.M. 2005. Fish and invertebrate assemblages in seagrass, mangrove, saltmarsh, and nonvegetated habitats. Estuaries 28, 63-77.

Costanza, R., d’Arge, R., de Groot, R., Farber, S., Grasso, M., Hannon, B., Limburg, K., Naeem, S., O’Neill, R.V., Paruelo, J., Raskin, R.G., Sutton, P. & van den Belt, M. 1997. The value of the world’s ecosystem services and natural capital. Nature 387, 253-260.


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Crowley, G.M. 1996. Late quaternary mangrove distribution in northern Australia. Australian Systematic Botany 9, 219-225.

Davis, R.A. 1995. Geologic impact of Hurricane Andrew on everglades coast of southwest Florida. Environmental Geology 25, 143-148.

de Lange, W.P. & de Lange, P.J. 1994. An appraisal of factors controlling the latitudinal distribution of mangrove (Avicennia marina var. resinifera) in New Zealand. Journal of Coastal Research 10, 538-548.

Detres, Y., Armstrong, R.A. & Connelly, X.M. 2001. Ultraviolet-induced responses in two species of climax tropical marine macrophytes. Journal of Photochemistry and Photobiology B 62, 55-66.

Doyle, T.W., Smith, T.J. & Robblee, M.B. 1995. Wind damage effects of Hurricane Andrew on mangrove communities along the southwest coast of Florida. Journal of Coastal Research 21, 159-168.

Diop, E.S., Soumare, A., Diallo, N. & Guisse, A. 1997. Recent changes of the mangroves of the Saloum River estuary, Senegal. Mangroves and Salt Marshes 1, 163-172.

Duarte, C.M., Middelburg, J.J. & Caraco, N. 2005. Major role of marine vegetation on the oceanic carbon cycle. Biogeoscience 1, 173-180.

Duke, N.C. 1992. Mangrove floristics and biogeography. In Tropical Mangrove Ecosystems (eds A.I. Robertson & D.M. Alongi), Coastal and Estuarine Studies 41, 63-100.

Duke, N.C., Bell, A.M., Pederson, D.K., Roelfsema, C.M. & Nash, S.B. 2005. Herbicides implicated as the cause of severe mangrove dieback in the Mackay region, NE Australia: consequences for marine plant habitats of the GBR World Heritage Area. Marine Pollution Bulletin 51, 308-324.

Duke, N.C., Ball, M.C. & Ellison, J.C. 1998. Factors influencing biodiversity and distributional gradients in mangroves. Global Ecology and Biogeography Letters 7, 27-47.

Ellis, J., Nicholls, P., Craggs, R., Hofstra, D. & Hewitt, J. 2004. Effects of terrigenous sedimentation on mangrove physiology and associated macrobenthic communities. Marine Ecology Progress Series 270, 71-82.

Ellison, J.C. 1998. Impacts of sediment burial on mangroves. Marine Pollution Bulletin 37, 420-426.

Ellison, J.C. 1993. Mangrove retreat with rising sea-level, Bermuda. Estuarine Coastal and Shelf Science 37, 75-87.

Ellison, J.C. & Stoddart, D.R. 1991. Mangrove ecosystem collapse during predicted sealevel rise – Holocene analogs and implications. Journal of Coastal Research 7, 151-165.

Ellison, A.M. & Farnsworth, E.J. 1993. Seedling survivorship, growth and response to disturbance in Belizean mangal. American Journal of Botany 80, 1137-1145.

Ellison, A.M. & Farnsworth, E.J. 1996. Spatial and temporal variability in growth of Rhizophora mangle saplings on coral cays: links with variation in insolation, herbivory and local sedimentation rate. Journal of Ecology 84, 717-731.

Ellison, A.M. & Farnsworth, E.J. 1997. Simulated sea level change alters anatomy, physiology, growth and reproduction of red mangrove (Rhizophora mangle L.). Oecologia 112, 435-446.

Evans, M.J. & Williams, R.J. 2001. Historical distribution of estuarine wetlands at Kurnell Peninsula, Botany Bay. Wetlands (Australia) 19, 61-71.

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Farnsworth, E.J., Ellison, A.M. & Gong, W.K. 1996. Elevated CO2 alters anatomy, physiology, growth and reproduction of red mangrove (Rhizophora mangle L). Oecologia 108, 599-609.

Field, C.D. 1995. Impact of expected climate change on mangroves. Hydrobiologia 285, 75-81.

Furukawa, K., Wolanski, E. & Mueller, H. 1997. Currents and sediment transport in mangrove forests. Estuarine Coastal and Shelf Science 44, 301-310.

Harty, C. 2004. Planning strategies for mangrove and saltmarsh changes in southeast Australia. Coastal Management 32, 405-415.

Harty, C. & Cheng, D. 2003. Ecological assessment and strategies for the management of mangroves in Brisbane Water- Gosford, New South Wales, Australia. Landscape and Urban Planning 62, 219-240.

Imbert, D. & Menard, S. 1997. Vegetation structure and primary production in mangroves of Baie de Fort-de-France, Martinique. Biotropica 29, 413-426.

Kathirisan, K. & Bingham, B.L. 2001. Biology of mangroves and mangrove ecosystems. Advances in Marine Biology 40, 81-251.

Kathirisan, K. & Rajendran, N. 2005. Coastal mangrove forests mitigated tsunami. Estuarine, Coastal and Shelf Science 65, 601-606.

Kathirisan, K., Rajendran, N. & Thangaduri, G. 1996. Growth of mangrove seedlings in the intertidal area of Vellar estuary, southeast coast of India. Indian Journal of Marine Sciences 25, 240-243.

Loneragan, N.R., Adnan, N.A., Connolly, R.M. & Manson, F.J. 2005. Prawn landings and their relationship with the extent of mangroves and shallow waters in western peninsular Malaysia. Estuarine Coastal and Shelf Science 63, 187-200.

Lugo, A.E., Brown, S. & Brinson, M.M. 1998. Forested wetlands in freshwater and salt-water environments. Limnology and Oceanography 33, 894-909.

Manson, F.J., Loneragan, N.R., Skilleter, G.A. & Phinn, S.R. 2005. An evaluation of the evidence for linkages between mangroves and fisheries: a synthesis of the literature and identification of research directions. Oceanography and Marine Biology: an Annual Review 43, 483-513.

Moorthy, P. & Kathiresan, K. 1997. Influence of ultraviolet-B radiation on photosynthetic and biochemical characteristics of a mangrove Rhizophora apiculate. Photosynthetica 34, 465-471.

Parkinson, R.W., Delaune, R.D. & White, J.R. 1994. Holocene sea-level rise and the fate of mangrove forests within the wider Caribbean region. Journal of Coastal Research 10, 1077-1086.

Rogers, K., Wilton, K. M. & Saintilan, N. 2006. Vegetation change and surface elevation dynamics in estuarine wetlands of southeast Australia. Estuarine Coastal and Shelf Science 66, 559-569.

Saintilan, N. & Williams, R. J. 1999. Mangrove transgression into saltmarsh environments in south-east Australia. Global Ecology and Biogeography 8, 117-124.

Thom, B.G. 1982. Mangrove ecology – a geomorphological perspective. In Mangrove Ecosystems of Australia (ed B.F. Clough), Australian National University Press, Canberra.

Twilley, R.R. & Chen, R. 1998. A water budget and hydrology model of a basin mangrove forest in Rookery Bay, Florida. Marine and Freshwater Research 49, 309-323.


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Woodroffe, C. 1992. Mangrove sediments and geomorphology. In Tropical Mangrove Ecosystems (eds A.I. Robertson & D.M. Alongi), Coastal and Estuarine Studies 41, 7-42.

Woodroffe, C.D. & Grime, D. 1999. Storm impact and evolution of a mangrove-fringed chenier plain, Shoal Bay, Darwin, Australia. Marine Geology 159, 303-321.

14.5 Kelps

Abbott, I.A. & Hollenberg, G.J. 1976. Marine Algae of California, Stanford, California, USA: Stanford University Press.

Adey, W.H. & Steneck, R.S. 2001. Thermogeography oxer time creates biogeographic regions: A temperature/space/time-integrated model and an abundance-weighted test for benthic marine algae. Journal of Phycology 37, 677-698.

Andrew, N.L. 1993. Spatial heterogeneity, sea-urchin grazing, and habitat structure on reefs in temperate Australia. Ecology 74, 292-302.

Barry, J.P., Baxter, C.H., Sagarin, R.D. & Gilman, S.E. 1995. Climate-related, long-term faunal changes in a California rocky intertidal community. Science 267, 672-675.

Bernstein, B.B. & Jung, N. 1980. Selective pressures and coevolution in a kelp canopy community in Southern California. Ecological Monographs 49, 335-355.

Bischof, K., Hanelt, D. & Wiencke, C. 1998. UV-radiation can affect depth-zonation of Antarctic macroalgae. Marine Biology 131, 597-605.

Bodkin, J.L. 1988. Effects of kelp forest removal on associated fish assemblages in central California. Journal of Experimental Marine Biology and Ecology 117, 227-238.

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14.6 Rocky shores

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Valentine, J.P. & Johnson C.R. 2003. Establishment of the introduced kelp Undaria pinnatifida in Tasmania depends on disturbance to native algal assemblages. Journal of Experimental Marine Biology and Ecology 295, 63-90.

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14.7 Corals

Baker, A.C. 2003. Flexibility and specificity in coral-algal symbiosis: Diversity, ecology, and biogeography of Symbiodinium. Annual Review of Ecology Evolution and Systematics 34, 661-689.

Bellwood, D.R., Hughes, T.P., Folke, C. & Nystrom, N. 2004. Confronting the coral reef crisis. Nature 429, 827-833.

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Thresher, R.E., Rintoul, S.R., Koslow, J.A., Weidman, C., Adkins, J. and Proctor, C. 2004. Oceanic evidence of climate change in southern Australia over the last three centuries. Geophys. Res. Lett. 31, L07212.

Wantièz L, HarmelinVivien M, Kulbicki M 1996. Spatial and temporal variation in a soft-bottom fish assemblage in St Vincent Bay, New Caledonia. Marine Biology 125, 801-812.

Ware, D.M., 1995. A century and a half of change in the climate of the North East Pacific. Fisheries Oceanography, 4, 267-277.

14.11 Pelagic fishes

ABARE (2005). Australian Fisheries Statistics 2004. Canberra. Arnold, G. and H. Dewar (2001). Electronic tags in marine fisheries research: a 30-year

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Bakun, A. and S. J. Weeks (2004). Greenhouse gas buildup, sardines, submarine eruptions and the possibility of abrupt degradation of intense marine upwelling ecosystems. Ecology Letters 7: 1015-1023.

Block, B. A., D. P. Costa, G. W. Boehlert and R. E. Kochevar (2003). Revealing pelagic habitat use: the tagging of Pacific pelagics program. Oceanologica Acta 5(5): 255-266.

Brown, A. (2005). Age and growth of jack mackerel off the east coast of Tasmania. Honours Thesis, School of Zoology, University of Tasmania.

Bunce, A. (2004). Do dietary changes of Australasian gannets (Morus serrator) reflect variability in pelagic fish stocks? Wildlife Research 31(4):383-387.

Campbell, R. A. and A. Hobday (2003). Swordfish - Environment - Seamount - Fishery Interactions off eastern Australia. Report to the Australian Fisheries Management Authority, Canberra, Australia.

Campbell, R.A. (1999). Long term trends in yellowfin tuna abundance in the south-west Pacific: with an emphasis on the eastern Australian Fishing Zone. Final report to the Australian Fisheries Management Authority, Canberra, 105pp.

Caton, A. and K. McLoughlin, Eds. (2004). Fishery Status Reports 2004: Status of Fish Stocks Managed by the Australian Government., Bureau of Rural Sciences, Canberra.

Chavez, F. P., J. Ryan, S. E. Lluch-Cota and M. Niquen (2003). From anchovies to sardines and back: multidecadal change in the Pacific Ocean. Science 299: 217-221.

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Cury, P. and C. Roy (1989). Optimal environmental window and pelagic fish recruitment success in upwelling areas. Canadian Journal of Fisheries and Aquatic Sciences 46(4): 670-680.

Cury, P., A. Bakun, R. J. M. Crawford, A. Jarre, R. A. Quiñones, L. J. Shannon and H. M. Verheye (2000). Small pelagics in upwelling systems: patterns of interaction and structural changes in "wasp-waist" ecosystems. ICES Journal of Marine Science 57(3): 603-618.

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Gunn, J. and B. Block (2001). Advances in acoustic, archival, and satellite tagging of tunas. Tuna, Academic Press. 19: 167-224.

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Jackson, G. D., B. McGrath Steer, S. Wotherspoon and A. J. Hobday (2003). Variation in age, growth and maturity in the Australian arrow squid Nototodarus gouldi over time and space—what is the pattern? Marine Ecology Progress Series 264:57-71.

Jacobson, L. D., J. A. A. De Oliveira, M. Barange, M. A. Cisneros-Mata, R. Félix-Uraga, J. R. Hunter, J. Y. Kim, Y. Matsuura, M. Ñiquen, C. Porteiro, B. Rothschild, R. P. Sanchez, R. Serra, A. Uriarte and T. Wada (2001). Surplus production, variability, and climate change in the great sardine and anchovy fisheries. Canadian Journal of Fisheries and Aquatic Sciences 58: 1891-1903.

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Polacheck, T. (2006). Tuna longline catch rates in the Indian Ocean: Did industrial fishing result in a 90% rapid decline in the abundance of large predatory species? Marine Policy 30: 470-482.

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14.12 Marine turtles

Adnyana, W., Ladds, P.W. & Blair, D. 1997. Observations of fibropapillomatosis in green turtles (Chelonia mydas) in Indonesia. Australian Veterinary Journal 75, 737-742.

André, J. Gyuris, E. & Lawler, I.R 2005. Comparison of the diets of sympatric dugongs and green turtles on the Orman Reefs, Torres Strait, Australia. Wildlife Research 32, 53-62.

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Bels, V.L., Davenport, J. & Renous, S. 1998. Food ingestion in the estuarine turtle Malaclemys terrapin: comparison with the marine leatherback Dermochelys coriacea. Journal of the Marine Biological Association of the UK 78, 953-972.

Binckley, C.A., Spotila, J.R., Wilson, K.S. & Paladino, F.V. 1998. Sex determination and sex ratios of Pacific leatherback turtles, Dermochelys coriacea. Copeia 2, 291-300.

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Bjorndal, K.A., Bolten, A.B. & Martins, H.R. 2000b. Somatic growth model of juvenile loggerhead sea turtles Caretta caretta: duration of pelagic stage. Marine Ecology Progress Series 202, 265-272.

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Bowen, B.W., Bass, A.L., Soares, L. & Toonen, R.J. 2005. Conservation implications of complex population structure: lessons from the loggerhead turtle (Caretta caretta). Molecular Ecology 14, 2389-2402.

Brand-Gardner, S.J., Lanyon, J.M. & Limpus, C.J. 1999. Diet selection by immature green turtles, Chelonia mydas, in subtropical Moreton Bay, south-east Queensland. Australian Journal of Zoology 47, 181-191.

Broderick, A.C., Godley, B.J. & Hays, G.C. 2001. Trophic status drives interannual variability in nesting numbers of marine turtles. Proceedings of the Royal Society of London Series B 268: 1481-1487.

Broderick, A.C., Frauenstein, R., Glen, F., Hays, G.C., Jackson, A.L., Pelembe, T., Ruxton, G.D. & Godley, B.J. 2006. Are green turtles globally endangered? Global Ecology and Biogeography 15, 21-26.

Bugoni, L., Krause, L. & Petry, M.V. 2001. Marine debris and human impacts on sea turtles in southern Brazil. Marine Pollution Bulletin 42, 1330-1334.

Carr, A. 1987. New perspectives on the pelagic stage of sea turtle development. Conservation Biology 1, 103-121.

Carr, A., Carr, M.H. & Meylan, A.B. 1978. The Ecology and Migrations of Sea Turtles, 7: The west Caribbean green turtle. Bulletin of the American Museum of Natural History 162, 1-46.

Chaloupka, M. 2001. Historical trends, seasonality and spatial synchrony in green sea turtle egg production. Biological Conservation 101, 263-279.

Chaloupka, M. 2002. Stochastic simulation modelling of southern Great Barrier Reef green turtle population dynamics. Ecological Modelling 148, 79-109.

Chaloupka, M. & Limpus, C. 2001. Trends in the abundance of sea turtles resident in southern Great Barrier Reef waters. Biological Conservation 102, 235-249.

Cheng, I.J. & Chen, T.H. 1997. The incidental capture of five species of sea turtles by coastal setnet fisheries in the Eastern waters of Taiwan. Biological Conservation 82, 235-239.

Collins, J., Klomp, N. & Birckhead, J. 1996. Aboriginal use of wildlife: past, present and future. In Sustainable use of wildlife by aboriginal peoples and Torres Strait islanders (Eds. Bomford, M. and Caughley, J.) Australian Government Publishing Service, Canberra p14-36.

Davenport, J. 1997. Temperature and the life-history strategies of sea turtles. Journal of Thermal Biology 22, 479-488

Dawson, S.P. & Dennison, W.C. 1996. Effects of ultraviolet and photosynthetically active radiation on five seagrass species. Marine Biology 125, 629-638.

Duarte, C.M. 2002. The future of seagrass meadows. Environmental Conservation 29, 192-206.

Dutton, P.H., Bowen, B.W., Owens, D.W., Barragan, A. & Davis, S.K. 1999. Global phylogeography of the leatherback turtle (Dermochelys coriacea). Journal of Zoology 248, 397-409.

Eckert, S.A. 2002. Distribution of juvenile leatherback sea turtle Dermochelys coriacea sightings. Marine Ecology Progress Series 230, 289-293.

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Hewavisenthi, S. & Parmenter, C. J. 2001. Influence of incubation environment on the development of the flatback turtle (Natator depressus). Copeia 2001: 668-682.

Hewavisenthi, S. & Parmenter, C. J. 2002. Thermosensitive period for sexual differentiation of the gonads of the flatback turtle (Natator depressus Garman). Australian Journal of Zoology 50, 521-527.

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Holland D.L., Davenport, J. & East, J. 1990. The fatty acid composition of the leatherback turtle Dermochelys coriacea and its jellyfish prey. Journal of the Marine Biological Association of the UK 70, 761-770.

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Kwan, D. 1994. Fat reserves and reproduction in the green turtle, Chelonia mydas. Wildlife Research 21, 257-266.

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Lewison R.L., Freeman S.A. & Crowder L.B. 2004. Quantifying the effects of fishing on threatened species: the impacts of pelagic longlines on loggerheads and leatherback sea turtles. Ecology Letters 7, 221-231.

Limpus, C.J. 1992. The Hawksbill Turtle, Eretmochelys imbricata, in Queensland: Population Structure within a Southern Great Barrier Reef Feeding Ground. Wildlife Research 19, 489-506.

Limpus, C.J. 1993. The Green Turtle, Chelonia mydas, in Queensland - Breeding Males in the Southern Great Barrier Reef. Wildlife Research 20, 513-523.

Limpus, C.J. & Miller, J.D. 1990. The use of measured scutes of hawksbill turtles, Eretmochelys imbricata, in the management of the tortoiseshell (Bekko) trade. Australian Wildlife Research 17, 633-639.

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Impacts of climate change on Australian marine life-Part A: Executive Summary

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