Word - 8934 KB - Department of the Environment

Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species

DECEMBER 2011

PRODUCED BY Morgan Pratchett, James Cook University

FOR the Department of Sustainability, Environment, Water, Population and Communities

ON BEHALF OF the State of the Environment 2011 Committee

Citation

Pratchett M. Vulnerability and status of marine fishes for the Australian State of the

Environment Report 2011—tropical species. Report prepared for the Australian Government

Department of Sustainability, Environment, Water, Population and Communities on behalf of the State of the Environment 2011 Committee. Canberra: DSEWPaC, 2011.

© Commonwealth of Australia 2011.

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. Requests and enquiries concerning reproduction and rights should be addressed to Department of Sustainability, Environment, Water, Populations and

Communities, Public Affairs, GPO Box 787 Canberra ACT 2601 or email public.affairs@environment.gov.au

Disclaimer

The views and opinions expressed in this publication are those of the author and do not necessarily reflect those of the Australian Government or the Minister for Sustainability,

Environment, Water, Population and Communities.

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.

Cover image

Sea anemone and Clownfish, Great Barrier Reef, QLD

Photo by GBRMPA

Australia ■ State of the Environment 2011 Supplementary information i

Preface

This report was commissioned by the Department of Sustainability, Environment, Water,

Population and Communities to help inform the Australia State of the Environment (SoE)

2011 report. As part of ensuring its scientific credibility, this report has been independently peer reviewed.

The Minister for Environment is required, under the Environment Protection and Biodiversity

Conservation Act 1999 , to table a report in Parliament every five years on the State of the

Environment.

The Australia State of the Environment (SoE) 2011 report is a substantive, hardcopy report compiled by an independent committee appointed by the Minister for Environment. The report is an assessment of the current condition of the Australian environment, the pressures on it and the drivers of those pressures. It details management initiatives in place to address environmental concerns and the effectiveness of those initiatives.

The main purpose of SoE 2011 is to provide relevant and useful information on environmental issues to the public and decision-makers, in order to raise awareness and support more informed environmental management decisions that lead to more sustainable use and effective conservation of environmental assets.

The 2011 SoE report, commissioned technical reports and other supplementary products are available online at www.environment.gov.au/soe

Australia ■ State of the Environment 2011 Supplementary information ii

Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species

Executive Summary

This report presents a new framework for assessing the vulnerability and population status of Australian marine fishes based on i) inherent vulnerability to extinction, ii) current population status, and iii) population resilience. This assessment considered 3-4 distinct components for each indicator, which were assessed using a four-point scale (e.g., very good, good, poor and very poor).

Assessments of vulnerability and population status were completed for two species of coral reef fishes, i) the redfin butterflyfish ( Chaetodon lunulatus Quoy and

Gaimard 1824) and ii) the leopard coral trout ( Plectropomus leopardus Lacèpde

1802). The inherent vulnerability to extinction for both species was considered Low , given their reasonably large geographic ranges and ability to utilise a wide range of different reef habitats. The current population status for both C. lunulatus and P. leopardus was considered Good , with no evidence of long-term, reef-wide declines in abundance.

The two species considered in this report ( C. lunulatus and P. leopardus ) are each facing distinct threats, due to habitat degradation (especially coral loss) and direct fisheries exploitation, respectively. Current fisheries for Plectropomus spp. on the Great Barrier Reef do appear to be sustainable, and populations exhibit considerable resilience. With the recent expansion of no-take marine reserves populations on reefs closed to fishing have recovered very quickly. In contrast, C. lunulatus appears to have Poor population resilience, with no recovery observed >5 years after severe coral bleaching in the central Great Barrier Reef.

As habitat perturbations become more frequent and more severe, population resilience will have increasing importance in the persistence of populations and species. If the time taken for complete recovery of local populations exceeds the time between successive disturbances, then populations will inevitably decline. In the case of C. lunulatus , population resilience appears to be constrained by i) strong reliance on live corals (for food and settlement), such that recovery rates are limited by recovery in the local abundance of suitable coral species, and ii) seemingly low levels of connectivity and larval exchange among disparate populations (on widely separated reefs) and certainly among populations in different geographic regions

(e.g., the Great Barrier Reef versus Papua New Guinea). In this case, the large

Australia ■ State of the Environment 2011 Supplementary information 1

Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species geographic range of C. lunulatus is unlikely to offer any insurance against localised extirpations, thereby highlighting the need to consider population resilience when assessing vulnerability and population status of marine fishes.

There are numerous existing frameworks used to assess the vulnerability of species, based mostly on differential susceptibility to specific disturbances (e.g., fishing) and/ or inherent vulnerability to extinction. Population resilience (especially the capacity for populations to recover in the aftermath of disturbances) has not been explicitly considered in previous assessments, but is fundamental in assessing the fate of populations and species. Comprehensive assessments of inherent vulnerability, population status, and population resilience need to be completed

Australian marine fishes to identify species that are most at risk from increasing degradation and loss of natural habitats, direct exploitation, and climate change.



1. Background

1.1 Current threats to marine environments

The degradation and loss of natural habitats, due to the ever-expanding ecological footprint of humans, is causing population declines, range contractions, and extinction of species in both terrestrial (e.g., Dirzo and Raven 2003) and aquatic environments (e.g., Roberts and Hawkins 1999). During this “Anthropocene Era”, the extinction of species in different parts of the world has coincided very closely with the timing of first human contact (Dirzo and Raven 2003). On Pacific Islands, for example, 10% of birds (~1,000 species) were eliminated within several decades of human colonization (Pimm et al. 1994). Since 1500 A.D. there have been a total of

811 documented species extinctions (Hilton-Taylor 2000), though it is widely recognized that extinction rates have been grossly underestimated in taxa such as insects and plants. Moreover, population extinctions (or localised extirpation of species) are occurring at 10-100 times the rate of species extinctions, and are no less important for local biodiversity and ecosystem function (Ehrlich and Daily 1993).

Reducing and reversing local and global rates of biodiversity loss represents a significant, but critical environmental challenge. Importantly, the loss of species may have far-reaching ramifications for ecosystem processes (Thomas et al. 2004,

Worm et al. 2006), potentially causing the loss of key functions, and reducing both productivity and resilience. The major cause of documented species extinctions and extirpations, is habitat loss and fragmentation, which affects >80% of threatened species (Hilton-Taylor 2000). Other major causes of documented and predicted species extinctions (in order of importance) are i) direct exploitation, and ii) introduction of alien and invasive species (Dirzo and Raven 2003). Anthropogenic climate change is also increasingly being implicated in the extirpation, if not extinction, of species, and is expected to emerge as a major cause of species losses given sustained and ongoing changes in environmental conditions (Thomas et al.

2004).

Until recently, marine species were considered much more resilient to global extinctions compared to terrestrial and freshwater counterparts (Roberts and

Hawkins 1999), largely due to their large geographic ranges (Gaston 1994). It was


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species generally assumed that marine species had unlimited access to vast ocean areas, and could therefore persist in refuge populations that were isolated from major environmental perturbations or anthropogenic disturbances. It was also assumed that marine environments would be much more ‘buffered’ against environmental change (Carlton 1993), and that high diversity of marine species would confer a high degree of functional redundancy, whereby some species are expendable (Steele

1991). There have however, been a significant number of extinctions and near extinctions reported among marine species (Roberts and Hawkins 1999), mostly due to overexploitation and/ or habitat alteration and destruction. For the most part, it is marine species with restricted distributions that are threatened with extinction (e.g.,

Roberts and Hawkins 1999, Munday 2004). However, large-scale environmental perturbations (e.g., due to climate change) and widespread destruction of specific habitat types are increasingly threatening global populations of even wide-ranging marine species (e.g. Carpenter et al. 2008).

Coastal marine habitats are particularly exposed to global climate change, as well as more direct anthropogenic disturbances, which have caused rapid and accelerating loss of critical habitat types. The global extent of seagrass beds is declining at 1.5% (or 110 km 2 ) per year (Waycott et al. 2009). For coral reefs,

Wilkinson (2004) estimated that 20% of the world’s coral reefs have already been destroyed, whereby coral cover has declined by >90% and there is limited prospect of recovery. Coral cover has declined by 20-

90% on a further 50% of the world’s coral reefs, and these reefs may be destroyed by 2050 (Wilkinson 2004). In addition, anthropogenic disturbances are contributing degradation and increased fragmentation of coastal habitats. Strong directional changes in the structure of benthic communities have been reported for seagrass beds (Duarte 2002), kelp forests (Steneck et al. 2002), mangroves (Alongi 2002) and coral reefs (Hughes et al. 2003), where structurally complex, yet seemingly fragile, habitat-forming species are increasingly replaced by more ephemeral species that contribute little to overall habitat structure (Wilson et al. 2006). Declines in the quantity and quality of habitat within these ecosystems have significant subsequent impacts on the biodiversity and abundance of motile organisms, such as fishes and motile invertebrates (Wilson et al. 2006, Pratchett et al. 2008a, 2009a). Increasing fragmentation of habitat patches


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species may further limit the potential for remnant populations to repopulate locations where species are extirpated (Hughes et al. 2003, 2005).

Historically, the greatest threat to marine fishes was from direct exploitation

(e.g., Cheung et al. 2005, Newton et al. 2007). Coastal fisheries throughout the world are generally regarded as unsustainable, if not already grossly overexploited (Pauly et al. 2002, Worm et al. 2009). Inshore fisheries and specifically, coral reef fisheries have already collapsed in 18% of tropical island countries, and are fully exploited or overexploited in a further 17% of countries (Newton et al. 2007). Moreover, destructive fishing practices and direct habitat alteration, combined with extrinsic contributors to habitat degradation (pollution, sedimentation, eutrophication, and climate change), have greatly increased the impact of human populations on marine fishes. Importantly, habitat degradation affects a much greater range of different fishes compared to fisheries extraction, as changes to habitat structure not only reduce availability of resources, but may influence the outcomes of key biological interactions, such as competition and predation (e.g., Coker et al. 2009). Fisheries exploitation typically targets larger individuals and species at higher trophic levels

(Pauly et al. 1998), whilst habitat alterations and destruction may affect an altogether different component of fish assemblages (Wilson et al. 2008, Graham et al. 2011a).

However, fishing and habitat-degradation tend to co-occur, which is likely to lead to comprehensive declines in the abundance of fishes, especially in heavily populated regions of the world (Wilson et al. 2008, Graham et al. 2011a).

Marine fishes are important, not only in sustaining global fisheries (Worm et al. 2009), but also in maintaining ecosystem function. On coral reefs, Bellwood et al

(2004) showed that certain fishes fulfil ecological functions that are fundamental in maintaining ecosystem state. Declines in the abundance of fishes that fulfil these functions often results in a phase shift to a less desirable ecosystem states (Pandolfi et al. 2003). For example, depletion of herbivorous fishes may cause a transition from coral to macroalgal dominance. The loss of an entire functional group, especially those comprising multiple species, may appear unlikely except during extreme (severe or prolonged) disturbances (e.g., Hughes 1994). However, functional redundancy among coral reef fishes may have been significantly overstated (e.g., Bellwood et al. 2006); even in high diversity systems, some


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species functions may be performed by just one species (Bellwood et al. 2006). It is also important to separate functional redundancy from response diversity (Elmqvist et al.

2003). If entire groups of fishes all respond to a disturbance in the same way (e.g., if all species are extirpated following climate-induced coral bleaching), then ecological functions will stop irrespective of how many species fulfil that role (Graham et al.

2011a). Thus, functional redundancy in the absence of response diversity will give a false sense of security (Bellwood et al. 2004). Even for groups with high functional redundancy and response diversity, Naeem et al. (1994) suggested that all species are important and individually contribute to increased efficiency in biogeochemical and trophic functions (see also Tilman and Downing 1994).

The purpose of this report is to consider a modified framework for assessing the vulnerability and population status of marine fishes in Australia, based on i) inherent vulnerability to extinction, ii) current population status, and iii) population resilience. Population resilience has never before been explicitly considered in species and population assessments, but encapsulates several key elements that determine vulnerability to disturbances and the long-term persistence of populations and species. This assessment will consider several different components of population resilience, as well as inherent vulnerability and current population status, which will be assigned one of four different grades (e.g., very good, good, poor and very poor), following the Great Barrier Reef Outlook Report (Great Barrier Reef

Marine Park Authority 2009). Assessment grades are intended to reflect the current status of Australian populations of marine fishes, as well as their likelihood of becoming threatened within a timeframe relevant to the biology of the species

(e.g.,10-30 years), following Hilton-Taylor (2000).

1.2 National State of the Environment Reporting – Marine Ecosystems

The Australia State of the Environment Report (e.g., Beeton et al. 2006) is intended as a “comprehensive [national] assessment of the current condition of

Australia’s environment, which also identifies key pressures on the environment”.

Since 1996, there have been three Australian State of the Environment Reports, and each have included a specific chapter on marine and/ or coastal environments: i)

Chapter 8

– Estuaries and the Sea

in State of Environment Advisory Council (1996), ii) Thematic Findings – Coasts and Oceans in Australia State of the Environment



Committee (2001), and iii) Chapter 6 - Coasts and Oceans in Beeton et al (2006).

Each of these reports is based on the “Pressure – State (or Condition) – Response” reporting model, identifying key pressures imposed by anthropogenic activities (e.g., agriculture and fisheries) on key aspects of the environment (e.g., biodiversity), and the institutional and individual responses (e.g., legislation) required to reverse and/ or minimize sustained and ongoing pressures.

The major pressures identified in each of the three previous reports (Table 1) relate to i) coastal development, ii) contaminants and pollution (or declining water quality), iii) fishing and aquaculture, iv) shipping, and v) introduced pests (mostly caused by shipping activity). The Australia State of the Environment Committee

(2001) explicitly recognised the role of existing management and governance (or legislative) structures as a key limitation, as well as key contributor, to effective marine management. By 2006, there was a strong emphasis on the threat posed by global climate change (specifically, increasing atmospheric concentrations of carbon dioxide) to specific ecosystems and species, but Beeton et al. (2006) did not directly assess effects of climate change on marine ecosystems or species (Table I). Recent changes in the state of governance and management structures, as well key climatic variables (mainly, sea surface temperatures) will be fundamental to assessing the state of marine ecosystems and species in 2011.



Table I.

Summary of major pressures (and state variables) identified in each of the three previous Australia State of the Environment Reports

State of Environment Report 1996

Pressure

Coastal development

Contaminants

Fishing

Mining

Transport and shipping

State

Human populations in non-metropolitan coastal zone grew 95%

(2 million people) between 1971 and 1991

450km 2 -(0.9%) of seagrass beds have been destroyed by habitat alteration and dredging in the last decade

Up to 17% of mangrove forests have been removed during land clearing and reclamation

River inflow and waste water discharges as one of the most significant pressures on Australia’s coastal ecosystems, causing for example, algal blooms

80% of contaminants originate from land

Each year Australia’s sewerage systems discharge around 10

000 tonnes of phosphorus and 100 000 tonnes of nitrogen

Long-term degradation of coral reef environments (including, extensive loss of corals) is apparent at some inshore locations, probably due to sedimentation

National fisheries production increased from 73 000 tonnes in

1964 –65 to 195 000 tonnes in 1994–95

15 (out of 17) commercial fisheries species are fully or overexploited

Virtually all estuarine, near-shore and off-shore areas support fishing.

Incidental catches (e.g., turtles) and habitat modification are having significant effects on non-target species

Nesting populations of loggerhead turtles have declined >90% from 1977 to 1191 at major breeding sites

Commercial prawn trawling in the Gulf of Capenteria in 1991 is estimated to have killed up to 67,000 sea snakes

576 dugongs were killed in shark nets from 1964 to 1988.

Sand mining and dredging causes localised habitat degradation at both mining and dumping sites

86% of oil and 75% of natural gas comes from offshore wells.

1,100 offshore wells have been drilled and extracted 2,8000 million barrels of oil

800 barrels of oil released (spilt) into environment

~12,000 ships arrive in Australian ports (total of 68 ports) from overseas each year

Ports recognised as the most disturbed marine environments in

Australia

At least 55 introduced species of fishes and invertebrates, and a number of toxic algae, have been introduced in Australia via ballast water discharge

In 1991, ~155 million tonnes of ballast water (78% of from offshore ports) was discharged in Australian ports




Pressure State

Habitat degradation

Most habitats poorly studied and may be highly vulnerable

~50% of estuaries are degraded, owing to land-use practices or human settlement

Australia’s coral reefs degraded mostly by sediment and nutrient runoff

Human population growth greatest on Australia’s coastal strip

Increasing incidence and extent of acid (sulphate) soils

Coastal settlement and development

Declining water quality

Fisheries and aquaculture

Nutrient and sediment loads increasing

Many landowners unaware (or apathetic) of downstream consequences of poor land use

Many fisheries fully or overexploited

Aquaculture production increasing 14% per year

Introduced pests ~200 pest species have been introduced, mostly through ballast water

Pest species contribute greatly to deterioration of marine habitats and native species

Significant threat of further introductions

Marine industry development

Shipping and port infrastructure require extensive and ongoing dredging

Four [oil spill] incidents in six years

Marine resource management

~80 international agreements relate to use of oceans

The High Court establishes that the Native Title Act 1993 recognises native title rights to the sea

Australia’s ocean policy (including support for integrated oceans management) released in 1998

194 marine protected areas, covering 60 million ha, representing an increase of 17.6 million since 1996


Pressure

Urban development

Agriculture

Fishing

State

6.0% of Australia’s coastline has been developed, which is an increase of 30% (from 4.6%) since 1980.

Approximately 19,000 tonnes of phophorus and 141,000 tonnes of nitrogen are discharged to rivers and oceans each year

17 (out of 74) Commonwealth-managed fisheries species are overexploited



Aquaculture Australia contributes 0.1% of global aquaculture production, but the industry (and international exports) is growing rapidly

Aquaculture accounts for 1/6 of the volume and 1/3 of the vaule of total seafood production

Coastal and marine pollution

Introduced marine species

Climate and carbon dioxide*

11 serious oil spills have occurred since 1970, resulting in

>21,094 tonnes of oil to be released (spilt) in coastal environments

At least 250 (and as many as 500) non-native species have been introduced to Australia via ballast water discharge

Net emissions are estimated to have increased by 2.3 per cent to a total of 564.7 million tonnes (Mt) CO2-e from 1990 to 2004

The energy sector accounts for 68.6% of Australia's net emissions

Emissions for 2020 are projected to reach 122% of the 1990 level

*Not included within pressures that are specifically relevant to coasts and oceans, but recognises the threat posed from sustained and ongoing increases in CO

2

emissions.

Each of the former State of the Environment Reports (e.g., Beeton et al.

2006; Table II) has utilised a unique combination of indicators to assess the population status of marine fishes (and other organisms). These indicators vary greatly in their application and specificity for assessing populations status, ranging from summary statistics derived independent assessments focussed in the global vulnerability of marine species (CO-02 Number of marine species that are endangered or threatened and changes in population/ distribution of selected threatened species; Beeton et al. 2006), to explicit trends in abundance for select groups of fishes (Table II). There are also a number of indicators relevant to the specific threats (e.g., fisheries and climate change) identified for certain species.

Until now however, there has not been a comprehensive framework developed for assessing the vulnerability and status of individual populations and species that is directly relevant to the Australian State of the Environment Report. This is especially true for species that occur in Australia and also have more widespread geographic ranges (cf. endemic species), and species not currently considered to be threatened.

Table II.

Indicators used in the Australia State of Environment Report 2006 (Beeton et al.

2006) that are relevant to assessing the population status of marine fishes.



Indicators

Condition of the ocean and coastal waters - Condition of species, habitats and ecosystems

CO-01 Trends in selected groups of coastal and marine species and habitats

Justification - In the absence of any way of assessing the overall condition of species, habitats and ecosystems across all Australian coastal land and coastal and marine waters, considering the condition of a selection of key species, groups of species, habitats and ecosystems is probably the closest we can come to an indicator.

Corals, mangroves, seagrass and kelp forests have been selected as significant indicators for biodiversity generally because they are groups of species that also represent habitat for a wide range of other species.

Fish and bird species have selected because they are visible, well up in the food chain (and therefore vulnerable to pressures operating deeper in the food chain) and some measurement of populations of fish and bird species is undertaken.

Conclusions - We can not, at this stage, even in the rare cases where we know changes are happening, be sure whether changes in either extent of the selected habitats, or in populations of particular species, are indicative of healthy or unhealthy changes for their supporting and supported ecosystems. A precautionary approach would suggest that if, on balance, in the context of a range of anthropogenic pressures, more species and habitats seem to be declining than expanding, it probably does not bode well for the condition of ecosystems more broadly.

CO-02 Number of marine species that are endangered or threatened and changes in population/ distribution of selected threatened species

Justification - Trends in population and distribution of selected threatened species is limited as an indicator because positive changes in populations of such species may reflect only efforts to save that particular species, rather than improvements in the condition of marine species, habitats and ecosystems more generally.

However, the declining range in which particular species are found can be read as broadly indicative of a decline in marine ecosystems.

Conclusions - The total number of species that are considered to be threatened has steadily increased since 1993: the listings of these species do not disaggregate whether they are terrestrial and marine. Also, increases in numbers of listed species may reflect an increase in scientific knowledge and/or levels of public concern, rather than real changes in the level of threat.

Much of the data on Trends in population and distribution of selected threatened species are limited to the Great Barrier Reef. However, these also show that several threatened marine species such as dugongs and several species of marine turtle appear to be declining, as are some other species of marine mammal, such as the Indo-Pacific humpback dolphin.

BD-02 Conservation status of nationally significant species and ecological communities, compared with previous years

Justification - The number of species that are threatened or endangered is a surrogate indicator for decline in species richness more generally

Conclusions - There has been a slight increase in the number of species listed under Australian government legislation as either vulnerable, endangered or critically endangered for all groups.



The indicator reports only species that have been listed as threatened and reflects the state of knowledge rather than the state of species.

Pressures on marine biodiversity: pressures of fishing

CO-16 Status of Australian Fisheries

Justification - Status of commercial fisheries does not give a comprehensive picture of the broader condition of marine species and ecosystems, but commercial fish species are important components of marine ecosystems and serious declines are likely to be indicative of broader ecological change. Status of commercial fisheries is assessable on the basis of catch changes, whereas the condition of non-commercial fish populations is much more difficult of assess.

Conclusions - There has been a sustained increase in the number of fisheries deemed to be overfished (from 5 in 1992 to 17 in 2004). There is also inadequate information to assess the status of many (40) different fisheries stocks in 2004.

Pressures on biodiversity - Climate variability

CO-76 Examples of the impact of climate variability on selected coastal and marine species, habitats or ecosystems

Justification - Some species and habitat types may be particularly susceptible to changes of climate, and may provide sentinel indicators for more widespread or less readily detectable changes. Coral bleaching is a particularly visible impact that can be caused by thermal change. Mangrove incursion into terrestrial rainforest or saltmarsh in tropical areas may also be indicative of the pressure of changing climate on ecosystems.

Conclusions – Coral bleaching, mangrove incursion into rain forest and salt marsh, and a range of other changes in the population, distribution and condition of selected species, groups of species and habitats may be indicative of the pressure of climate change and climate variability on ecosystems.

Response of biota – Fish (Freshwater)

IW-31 Fish - Abundance and distribution

Justification - The health of populations and communities of native fish species is critical to the overall health of the aquatic system. Fish occupy trophic levels from top predators down to herbivores and detritivores and play a vital role in maintaining the ecological integrity of river ecosystems. Fish species may be affected by a range of pressures including significant changes to water flow, damage to riparian zones, removal of in-stream habitats, sedimentation, lowered water quality, thermal pollution and barriers to fish passage.

Conclusions – The only data available are from NSW, where condition of fish communities continues to decline, with reduced abundance or distribution.

Pressures include river flow alteration, habitat loss, water quality, introduced species and stocking and translocation.



1.2 Marine fishes in Australia

There are currently 4,482 described species of Australian fishes (Hoese et al. 2006), but this number poor reflects the total biodiversity of Australian marine and freshwater fishes. There are, for example, at least 300 species still awaiting identification or description, and many more species are likely to be revealed following extensive searches in deep water habitats and isolated locations. Most

(68%) of the described fishes (approximately 3,029 species) are found in tropical marine environments, and mostly on coral reefs. Of the remaining species, 166 species (7%) are restricted to freshwater, and 1,222 (27%) are found exclusively in sub-tropical or temperate habitats (Hoese et al. 2006). The diversity of fishes in tropical environments is more than twice that of temperate environments, though

Australia is a global hotspot for biodiversity of temperate fishes. The temperate fish fauna also comprise very high levels of endemism. Overall, 24% of the Australian fish fauna. The lowest level of endemism (8%) occurs on coral reefs, whereas 46-

47% of fishes in subtropical and temperate locations are endemic, increasing to 55% among those fishes found on temperate rocky reefs (Hoese et al. 2006).

There are approximately 1,610 species of fishes found within the vicinity of

Australian coral reefs (Hoese et al. 2006), including those species that live on or over coral reef habitats, as well as those species that associate with adjacent habitats, such as inter-reefal sand flats or seagrass beds. Choat and Bellwood (1991) define coral reef fishes as those species that are intimately associated and strongly dependent upon the unique biogenic features of coral reefs. More specifically, coral reef fishes are expected to settle directly in reef habitats or recruit to coral reefs prior to maturation, and rely on reef structures for food or shelter (Choat and Bellwood

1991). Fishes that rely on coral reef habitat food include both those species that directly feed on or from the reef benthos (including, corals, algae or associated micro-invertebrates), as well as those that benefit from the prey associated with reefgenerate habitats (including planktivorous fishes that exploit hydrological fronts and benthic feeding fishes that exploit sediment fields adjacent to carbonate structure).

The major families of coral reef fishes (Acanthuridae, Chaetodontidae, Scaridae/

Labridae, Siganidae, Pomacentridae and Pomacanthidae) are thus, highly


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species dependent coral reef ecosystems, and will be adversely affected by major declines in the quantity or quality of coral reef habitats.

Coral reef fishes did not necessarily evolve within coral reef ecosystems.

Even among those fishes that are intimately associated with coral reef habitats, many important families (e.g., Scaridae, Acanthuridae, and even Chaetodontidae) appear to have their origins in non-reef habitats, such as algal reefs or deep-water rocky habitats. Many of these fishes colonised coral reefs only relatively recently, as recently as five million years ago (Bellwood and Wainwright 2002). However, once they had colonised reef habitats, many of these fishes diversified and adapted to their new habitat, and are now intimately dependent on coral reefs. Among those fishes with the greatest reliance on coral reefs, are the butterflyfishes (family

Chaetodontidae), many of which have evolved to feed on scleractinian corals

(Bellwood et al. 2010). This review will focus on individual species from two different families (Chaetodontidae and Serranidae), which were chosen to represent extremes among coral reef fishes. Butterflyfishes (family Chaetodontidae) are relatively small, site attached fishes, typical of many smaller coral reef fishes that have a very strong reliance on coral reef habitats. Butterflyfishes are mostly benthic feeders and collectively consume a wide range of benthic and sessile prey, including corals, algae, and small motile crustaceans (Pratchett 2005). Groupers (family Serranidae) are typically large, predatory fishes, which exploit a range of different coastal environments. Groupers are often among the most important species (by value and volume) in coral reef fisheries (Morris et al. 2000). Importantly, there has been significant research interest in both these families of fishes (and therefore, considerable data on key biological and ecological variables), but for very different reasons.

1.2.1. Chaetodon butterflyfishes (family Chaetodontidae)

The family Chaetodontidae are a relatively diverse family of percoid fishes comprising 122 extant species. Butterflyfishes are characterised by deep compressed bodies, small protractile mouths and bristle-like teeth (Allen et al. 1998).

The family is dominated by fishes of the genus Chaetodon , which are among the most conspicuous inhabitants of coral reef environments (Figure 1). The majority of butterflyfishes are pair forming, presumably for the purpose of monogamous mating



(e.g., Fricke 1986, Pratchett et al. 2006a). These fishes are particularly well known and have been extensively studied because of their tendency to feed on scleractinian corals (e.g., Reese 1977, Harmelin-Vivien and Bouchon-Navaro 1983,

Pratchett 2005).

Coral feeding represents a relatively unique and potentially important trophic link between scleractinian corals and higher consumers (Cole et al. 2008).

Scleractinian corals are responsible for producing a significant component of the exogenous carbon on coral reefs (Reaka-Kudla 1997), which may ultimately enter trophic webs through incidental ingestion of coral mucous that settles on reef substrates (Wild et al. 2004). However, corallivorous fishes make this carbon immediately available to higher consumers. While many early researchers suggested that there were very few fishes are capable of feeding on corals (e.g., Randall 1974), coral-feeding has recently been reported for 128 species of coral reef fishes from 11 families (Cole et al. 2008). Butterflyfishes (family Chaetodontidae) account for 52%

(64/ 123 species) of fishes that to feed on corals (Cole et al. 2008). At least 50% of the butterflfyshes that inhabit coral reefs feed on hard (Scleractinian) or soft

(alcyonarian) corals. The remaining species are mostly benthic -feeding carnivores, which feed on small discrete invertebrates (e.g., polychaetes), though some species

(11 species) are planktivorous (Sano 1989).

Coral feeding infers a very high dependence upon corals, which has lead to suggestions that changes in the abundance (or behaviour) of butterflyfishes may provide an effective indicator for declines in coral cover (e.g., Samways 2005, Khalaf and Crosby 2005), or declining condition of entire coral reef ecosystems (Crosby and

Reese 1996, 2005). Butterflyfishes, especially highly specialised coral-feeding species, are certainly sensitive to changes in the availability of scleractinian corals

(e.g., Crosby and Reese 2005, Pratchett et al. 2006b, Graham 2007). However, responses of butterflyfishes to coral loss are very complex, highly species-specific, and often delayed, which greatly reduces their effectiveness as indicators of coral health Even among obligate hard-coral feeders, the extent to which coral cover

(versus recruitment rates or availability of other resources) limits population size may vary in time and space (Cox 1994).



Figure 1. The vagabond butterflyfish , Chaetodon vagabundus , exhibiting strong pair bonding that is typical of these fishes. Photograph by S.

Walker.

1.2.2. Plectropomus groupers (family Serranidae)

The family Serranidae is one of the largest and most diverse families of perciform fishes, comprising nearly 500 species (Randall 2005). The family comprises very large carnivorous species (e.g., Epinephephelus lanceolatus ) as well as several smaller planktivorous species (e.g., Pseudoanthias spp.). However, the best known species, and most conspicuous components of reef fish assemblages, are the larger Epinephalus and Plectoropomus species (subfamily Epinephelinae) that are among the largest coral reef fishes. As dominant predators, the epinepheline serranids (groupers) also play a major ecological role in structuring reef fish assemblages (e.g., Almany 2003).



Figure 2. The barcheek coral trout, Plectropomus maculatus , which is the most common coral trout in nearshore habitats. Photograph by R. Evans.

Epinepheline serranids are among the most important and valuable tropical fisheries species, and have been heavily exploited throughout the world (Morris et al.

2000, Pogonoski et al. 2002). In particular, groupers are the most intensively exploited group for the live fish trade (Morris et al. 2000), and the high prices paid for individual fishes (up to US$10,256) places considerable pressure on target species

(Sadovy and Vincent 2002). Since the late 196 0’s the demand for groupers (and other high-value species, such as Chelinus undulates ) has resulted in sequential over-exploitation of reefs and locations at increasing distances from the central market (in Hong Kong), including the Phillipines, Indonesia, Solomon Islands, and

Fiji. Many ephiepheline serranids also have biological characteristics that make them particularly vulnerable to fishing. In particular, many of the most heavily exploited species, tend to form very large spawning aggregations that can be readily targeted by local fishers (Morris et al. 2002). The Nassau grouper ( Epinephelus striatus ), for example, may form aggregations of tens of thousands of individuals, and individuals migrate up to 240 km to form aggregations. Heavy fishing of spawning aggregations is the major reason for the catastrophic decline in populations of E. striatus , which is now regarded as Endangered throughout the West Indies (Beets and Hixon 1994).



Reproductive biology of fishes in the family Serranidae is diverse and complex (Pears 2005). Most serranids are protogynous hermaphrodites, changing sex from female to male, as has been shown for two species of coral trout

Plectropomus laevis and P. leopardus (Adams 2002). There are however, a number of exceptions and variations to the generalised reproductive pattern. For example,

Sadovy and Colin (1995) deduced that the E. striatus is gonochoristic, whereby juvenile fishes become males or females upon maturity, and there is no evidence of sex change in the field. As a consequence, it is difficult to generalise about the reproductive biology for the large number of species (75%) for which nothing is currently known about reproductive biology. Also, most of the biological research on serranids has been conducted in the Caribbean and for species that do not occur in

Australia (e.g., Sadovy and Colin 1995), even though the highest diversity of serranids is found within the Indo-Pacific.



2. Assessing vulnerability and status of marine fishes

Previous methods used to establish priorities for species conservation have focused mostly on the inherent biological characteristics of species that result in and an increased risk of overall (species-level) extinction (e.g., Hilton-Taylor 2000:

Appendix I). These factors (e.g., the size of the global population and geographic extent) are intended to provide an indication of the relative risk of global extinction, without any explicit consideration of proximal causes for declines in their abundance, or extrinsic threats to populations and species. It is expected, for example, that species with smaller geographic ranges, and global populations comprising fewer reproductively mature individuals, would be much more vulnerable (and more likely to go extinct) given any natural or anthropogenic, acute or chronic, direct or indirect disturbance (Gaston 1994). The inherent vulnerability imposed by small population size, restricted geographic ranges (or extent) and area of occupancy, were formalized by IUCN (Appendix I), and are now central to assessing the threatened status of extant species for prioritising conservation.

There are a great many biological traits (in addition to geographic range, area of occupancy and population size) that may influence susceptibility to environmental disturbances and vulnerability of extinction (Roberts and Hawkins

1999, Graham et al. 2011a). These characteristics can be broadly divided into those characteristics (e.g., population size, gender ratio, as well as size and age structure) that affect the viability and persistence of populations and determine the risk of localized extirpation, versus those characteristics (range size, occupancy, and connectivity) that affect persistence versus extinction for the entire species (Graham et al. 2011a). It is also important to recognise that vulnerability of populations and species may vary through time due to either natural fluctuations in population dynamics, or chronic pressures that may cause gradual changes in underlying population dynamics, and therefore vulnerability. Assessing temporal trends in abundance and population dynamics is thus critical to provide an accurate and up-todate account of a species’ vulnerability.



2.1 Inherent vulnerability and extinction risk

In terrestrial environments, it is species with small populations, restricted geographic ranges, and limited ecological versatility, that are most at risk of extinction from large-scale environmental perturbations and increasing habitat degradation (Owens and Bennett 2000, Julliard et al. 2003, Williams et al. 2006).

Similarly, in marine systems it is rare, endemic and highly specialised fishes that have recently disappeared or are committed to extinction (Hawkins et al. 2000,

Munday 2004). Importantly, species with multiple traits that predispose them to extinction (e.g., restricted geographic ranges and small population size) face a disproportionate risk of extinction (e.g., Williams et al. 2006). Previous SoE reports have relied on formal threatened species assessment to provide indicators of inherent vulnerability (e.g., BD-02 Conservation status of nationally significant species and ecological communities, compared with previous years ; Table 2) but independent assessments of inherent vulnerability are likely provide a more up to date and regionally focussed (i.e. Australian based) indicators of population status

(Gärdenfors 2001).

2.1.1 Geographic range

Geographic ranges of marine fishes are mostly very large, but vary greatly

(Hughes et al. 2002, Bellwood et al. 2005). Forcipiger flavissimus , for example, is the most widespread butterflyfish species (geographic range of 1.06

× 10 8 km 2 ) with a circum-tropical distribution and very large latitudinal range (from 34 o N to 32 o S). At the other extreme, there are some fishes with extremely small geographic ranges

(<1 km 2 ), and these species may be much more susceptible to extinction due to an increased probability that any given disturbance may make the species unviable

(Gaston 1994). Several recently discovered coral-dwelling gobies ( Gobiodon spp.) are known from only one site in Papua New Guinea (Munday 2004). The abundance of one of these species ( Gobiodon sp A) declined precipitously following extensive mortality of its preferred host coral on nearshore reefs, and it may be threatened with extinction due to ongoing habitat degradation throughout its known range (Munday

2004).



For coral reef fishes, Hawkins et al. (2000) defined restricted range species as those with ranges <80,000 km 2 and estimated that 24% of coral-reef fishes (n =

1,677 species) fit within this category. There are however, striking taxonomic differences in patterns of range size (Hughes et al. 2003, Figure 7). Pratchett et al.

(2008) examined inter-specific variation in range size for four major families of coralreef fishes (Acanthuridae, Chaetodontidae, Pomacanthidae, and Pomacentridae) and 25.6% of species (158/ 612 species) had geographic ranges <80,000 km 2 , but the proportion of restricted range species ranged from <7% for the Acanthuridae,

21% for Chaetodontidae and Pomacanthidae and up to 34% for Pomacentridae. The mean size of fishes in the family Pomacentridae is also much smaller fishes than for

Acanthuridae, Chaetodonitidae and Pomacanthidae, and there are several published studies showing that large-bodied species tend to have larger geographic ranges

(e.g., for fishes, Pyron 2001). It is to be expected that other families dominated by small-bodied fishes (e.g., Gobiidae and Blennidae) will also have high proportion of species with relatively small geographic ranges.

Small range size per se does not increase a species susceptibility to acute disturbances (e.g., climate-induced coral bleaching) but means that geographically restricted disturbances can have potentially dire consequences for the species rather than just local populations. It is clear however, that the incidence of large-scale (near global) disturbances is increasing, largely due to the increasing importance of climatic disturbances. However, these disturbances vary greatly in their impacts among regions. Also, susceptibility of fishes to extinction is probably more dependent upon their geographical location rather than range size (Hawkins et al.

2000, Roberts et al. 2002). Species whose restricted geographic ranges are centered within areas subject to disproportionate effects of climate change and/or direct anthropogenic disturbances are at much greater risk than similarly restrictedrange species located in relatively unaffected locations and/ or areas devoid of additional anthropogenic stresses (Downing et al. 2005, Graham et al. 2007). The severe 1982 El Niño is blamed for the extinction of

Azurina eupalama , a plankton feeding damselfish that was endemic to the Galapagos Islands (Jennings et al.

1994). Meanwhile, A. hirundo is an ecologically equivalent species that persists in the Guadalupe and Revillagigedo islands (Allen 1991).



2.1.2 Population size

Risk of extinction is intuitively much greater for populations comprising few individuals compared with larger populations. Most importantly, slight fluctuations in the abundance of already small populations may mean that populations are no longer viable (Gaston 1994). There are many examples of ’rare’ coral-reef fishes that are rare not only because they are geographically restricted, but also because they have consistently low abundance throughout their geographic ranges (Jones et al.

2002). Chaetodon bennetti , for example, is relatively widespread but never common

(Allen et al. 1998). Average densities of C. bennetti across three geographically separated locations in the southern Pacific (<1 individual ha-1) are an order of magnitude lower than those of the next rarest species (M. Pratchett and M.

Berumen, unpublished data). Some of these species are naturally rare, and are likely to have persisted at low population size for considerable time. However, increasing habitat degradation and direct anthropogenic pressures may fundamentally alter the long-term viability of species with very small paopulations. We are also most concerned about species that were formerly common, but are now rare (Table 3).

For terrestrial species there is a consistent and often striking positive relationship between geographic ranges and population abundance (Gaston 1998,

Lawton 1999). For coral-reef fishes, however, previous studies have failed to detect any relationship between geographic range size and local abundance (Jones et al.

2002). For both Chaetodontidae and Acanthuridae, which are the only families of fishes for which there are good data on local abundance of multiple species across geographically widespread locations, there is no apparent relationship between geographic ranges of occurrence and mean abundance (Pratchett et al. 2008). It is possible that restricted range species are much more common than expected due to historical effects of extinction filtering (Williams et al. 2006), whereby extant species with restricted ranges have only persisted by virtue of their high abundance.

However, despite a generally poor relationship between range size and abundance among reef fishes (Jones et al. 2002), there are coral-reef fish species that are geographically restricted and locally rare. Perhaps these species are destined for extinction.



Extreme rarity can predispose a species to extinction, but common species are also likely to disappear if they are highly susceptible to particular disturbances.

Recurrent disturbances, predicted to result from sustained and ongoing climate change, are likely to have successive and cumulative effects on highly susceptible species (Smith and Buddemeier 1992), irrespective of their geographic range or global population size. For example, species of fishes (and corals) that experienced significant declines in abundance during the 1998 global mass bleaching may become extinct if locally severe and geographically extensive mass bleaching reoccurs within the time required for populations to recover. Acropora hyacinthus , for example, is currently listed as “near threatened” (Aeby et al. 2008) despite being one of the most widespread of all coral species. Coral-reef fishes that are at greatest risk from climate change are those that are directly dependent on scleractinian corals, either for food, habitat or recruitment (Pratchett et al. 2008). Given sustained and ongoing climate change, it is quite possible that these fishes may be threatened with extinction.

2.1.3 Ecological versatility

Ecological specialisation increases extinction risk because reliance on a limited of resources increases vulnerability to resource depletion. It is possible that some specialist species may shift their resource use if preferred resources decline in abundance (e.g., Pratchett et al. 2004), but it is clear that at least some fishes are strict specialists that use only a narrow range of resources (e.g., Munday 2004) and loss of these habitats will have serious implications for their persistence (Munday

2002). Pratchett et al. (2006b) showed that among coral-feeding butterflyfishes, highly specialised coral-dependent species became locally extinct following coral loss caused by mass bleaching. In contrast, more generalist (ie. ecologically versatile) species were relatively unaffected, presumably because they were able to exploit alternate prey resources as preferred corals were depleted (Pratchett et al.

2004, Pratchett et al. 2008; Figure 3). Specialisation is increasingly recognised as a key factor affecting relative risk of extinction among species within phylogenetic or ecological groupings (e.g., Musick 1999), but quantifying habitat or dietary specialisation does require considerable research effort, and patterns of resource use, let alone their degree of specialisation, are only known for a very restricted set


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species of marine fishes (e.g., butterflyfishes - Pratchett 2007, cardinalfishes - Gardiner and

Jones 2005, anemonefishes

– Elliot et al 1999, gobies – Munday et al. 1997).

-10

-20

-30

-40

-50

-60

-70

-80

-90

-100

0 20 40

No. of coral species consumed

60

Figure 3 . Susceptibility to disturbances versus ecological specialisation among coral-feeding butterflyfishes. Modified from Pratchett et al (2008a).

Ecological theory predicts that specialists should have smaller ranges, compared to generalist species, as ranges will be more constrained by the distribution of a few key resources (Brown 1984). If this relationship holds, highly specialised species may face a ‘double jeopardy’ of extinction or even a ‘triple jeopardy’ where specialist species are also rare (Hawkins et al. 2000, Munday 2004).

All the evidence to date suggests that coral-dependent fishes do not, on average, have smaller geographic ranges than other coral-reef fishes, possibly because the coral species used have widespread distributions themselves, or because the fishes change their patterns of resource use in accordance with the relative abundance of different coral species in different geographic regions. For example, Chaetodon trifascialis is the most specialised of coral-feeding fishes, feeding almost exclusively


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species on Acropora hyacinthus (Pratchett 2005), and yet it is also the most widely distributed of butterflyfish species (Allen et al. 1998). The widespread distribution of its preferred prey ( A. hyacinthus ) probably contributes to the relatively large geographic range of C. trifascialis . However, Jones et al. (2002) specifically tested whether specialist species have more restricted geographic ranges and found no clear relationship between geographic range size and specialisation for either butterfly fishes or anemone fishes.

Specialist species are also expected to be less common compared with generalist counterparts because their populations are more likely to be constrained by the abundance of specific resources (Brown 1984). There is some evidence among coral-dependent fishes that specialists have smaller populations than generalists but the pattern is far from consistent. Butterflyfishes that are specialist coral feeders are often locally abundant and can have higher densities than many generalist species (Jones et al. 2002). For example, C. trifascialis is not only among the most widely distributed of butterflyfish species (Allen et al. 1998), it also frequently ranks in the top three most abundant butterflyfishes throughout its geographic range (Jones et al. 2002).

2.1.4 Resource availability

Understanding species-specific responses to environmental perturbations and resource depletion requires extensive knowledge of patterns of resource use

(e.g., species level of specialization; Munday 2004), but it is also important to consider specific affects of disturbance on critical resources (McKinney 1997, Feary et al. 2007). Specialist species are expected to be much more vulnerable to habitatdegradation compared to species with generalised habitat requirements because their abundance tends to be more limited by habitat-availability and they have limited capacity to use alternate habitats following habitat loss (Munday 2004). However, it is also possible that specialist species may escape any effects from major disturbances if they utilize habitats or resources that are generally not susceptible to disturbance.

For highly specialised species with known patterns of resource use (e.g.,

Pratchett 2005), changes in the availability of essential resources may provide one of


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species the most important indicators for assessing the vulnerability, especially in the absence of rigorous or comprehensive data on population trends. For example, C. trifascialis was recently listed on the IUCN Red List as Near Threatened (Carpenter and Pratchett 2009) not due to documented declines in it s’ own abundance, but based on documented declines in the abundance of its’ preferred prey. The justification stated that, Chaetodon trifascialis has a strong dependency on a species of coral ( Acropora hyacinthus ) that is listed as Near Threatened (and just outside criteria for Vulnerable ) and although it has been seen to feed on at least 14 other coral species, all of these corals have shown substantial population declines throughout the Indo-Pacific. We infer that population declines of C. trifascialis are similar to those of A. hyacinthus (and other species it feeds on) and therefore list this species as Near Threatened (and close to Vulnerable ) (Carpenter and Pratchett

2009).



Table III . Criteria used to assess distinct components for the population and species-level indicator of inherent vulnerability to extinction for Australian marine fishes.

Indicator Very low Low High Very high

Geographic range

Population size

Ecological versatility

Resource vulnerability

Occurs in multiple ocean basins

Always abundant, and >75% of natural abundance

Above average for the family

Generally common, and

>50% natural abundance

Below average for family

Generally uncommon, or

>50% natural abundance

Moderately specialised

Extremely generalist

Primary resources are widely distributed, common and not threatened

Exhibits versatility in patterns of resource use

Primary resources are widely distributed, common, but are under some threat

Primary resources are not widely distributed or common, and are under some threat

<80,000km 2

Always extremely rare, or >25% natural abundance

Highly specialised

Primary resources are not widely distributed or common, and are under extreme threat

2.2 Current population status

2.2.1 Population trends

Trends (persistent changes in a given direction) in the abundance of a population or species provide one of the most direct measures with which to assess their long-term fate. In particular, sustained and ongoing declines in the size of the global population (especially, where the cause(s) of declines are not known) are a strong indicator that the species is facing extinction unless effective and immediate conservation action can reverse declines. The risk of extinction is further exacerbated by the fact that as populations decline, their geographical range also tends to decline, as individuals persist in only the most favourable habitats (Swain and Wade 1993). Documented declines in the abundance of species (where data exists) are one of the major criteria used to categorise the vulnerability status of threatened species (Appendix I). However, rates of change in population size must be considered relative to the generation times of specific species (or other intrinsic measures of population dynamics).



The IUCN assesses changes in the abundance of species over ten years or three generations, allowing for very protracted changes in abundance for extremely long-lived species. This is very relevant in the context of fishes, because the longevity (and generation times) of individual species vary by several orders of magnitude among sympatric species, ranging from 59 days for the coral reef blenny,

Eviota sigillata (Depczynski and Bellwood 2005), up to >40 years for several large species of Epinephelus (e.g., Pears et al. 2006). In the case of Epinephelus fuscoguttatusi, individuals contribute very little to reproductive output until they are

>9 years of age (Pears et al. 2006). Assessing population trends for these species should therefore, be conducted over 27 years (whereby the minimum generation time is assumed to be 9 years), though it is unlikely that data will be available to readily assess changes in the abundance of individual species over these timeframes.

For Australian marine fishes, one of the best resources for assessing largescale and long-term changes in abundance of coral reef fishes is the Long Term

Monitoring Program (LTMP) undertaken by the Australian Institute of Marine Science

(AIMS). A total of 214 species from 10 families (including Chaetodontidae and

Serranidae) have been surveyed annually from 1993 to 2005 and then every second year until 2011 (18 years). The LTMP surveys 47 reefs, in six latitudinal sectors of the GBR (Cooktown-Lizard Island, Cairns, Townsville, Whitsunday, Swains and

Capricorn-Bunker) and three cross-shelf positions (innershelf, mid-shelf and outershelf) (Emslie et al. 2011). Coral reef fishes are surveyed along permanent 50 m transects running parallel to the reef crest, which are established at replicate sites on the reef slope (6-9 metres deep) on the north-east flank of each reef. Five replicate transects are surveyed at each of three sites on each reef. Larger fishes

(Acanthuridae, Chaetodontidae, Labridae, Lethrinidae, Lutjanidae, Scaridae,

Serranidae, Siganidae, Zanclidae) are surveyed in a 5 m wide belt along each transect, whilst smaller fishes (Pomacentridae) are counted in a 1 m wide belt

(Sweatman et al. 2008).

2.2.2 Known threats

Knowledge of key threats to populations and species, or the environments in which they live, greatly informs assessments of vulnerability and population status.



More importantly, this also allows for projected declines in the abundance of population or species, based on current or potential threats. Of the 35 species of epinepheline groupers considered as Vulnerable by Morris et al. (2000), 16 were listed under criteria A2d, which assumes that there will be future population declines due to over-exploitation. Even where there is no evidence of over-exploitation, or current levels of exploitation do not pose a significant threat to local populations, some species are considered Vulnerable because they are extremely vulnerable to even very low levels of fishing pressure (e.g., Epinephelus tukula ) or it is expected that fishing pressure will increase markedly in the future (Morris et al. 2000).

Of the 133 population or species extinctions documented among marine fishes, exploitation is the principle contributor identified for 55% of cases, and a further 37% of species were lost due to habitat loss (Dulvy et al. 2003). Moreover,

Myers and Worm (2005) have shown that the global abundance of large predatory fishes has declined by at least 90% over the last 50 – 100 years, largely due to fishing. In Australia, fisheries exploitation is well below the level of many neighbouring countries, which is largely due to relatively the low population density relative and jurisdiction over very large tracts of marine habitat. However, Australian marine fisheries target a very restricted set of species, which has caused overexploitation and marked population crashes (e.g., Pogonoski et al. 2002). There are

12 species of marine fishes (and 33 freshwater species) considered to be

Endangered , Vulnerable or Conservation Dependant under the Australian

Environment Protection and Biodiversity Conservation (EPBC) Act 1999. Of these, 6 species (50%) are threatened mainly due to overexploitation or fisheries related mortality, including the orange roughy ( Hoplostethus atlanticus ), the grey nurse shark ( Carcharias taurus ) and great white shark ( Carcharodon carcharias ).

The other major threat to marine fishes (especially estuarine or nearshore species) is habitat degradation caused by the climate change and more direct anthropogenic disturbances, which compound upon natural disturbances (e.g.,

Hawkins et al. 2000, Pogonoski et al. 2002, Dulvy et al. 2003). Coral reef ecosystems, in particular, are subject to frequent and often catastrophic disturbances caused by a variety of different agents including severe tropical storms (cyclones), freshwater plumes, temperature extremes, and infestations of the coral eating crown-


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species of-thorns sea star, Acanthaster planci . These acute, but increasingly frequent, disturbances often cause marked reductions in the abundance of reef-building corals

(e.g., Hughes et al., 2003; Hoegh-Guldberg et al., 2007), which are the major architects and significant contributors to endogenous carbon production on coral reefs. Changes in the physical and biological structure of benthic reef habitats also have further, often detrimental, effects on other reef associated fishes (e.g., Wilson et al. 2006, Pratchett et al. 2008, Coker et al. 2009).

Climate-induced coral bleaching represents one of the most significant and increasingly prevalent disturbances to coral reef ecosystems, which not only causes extensive coral mortality, but also reduces the abundance of many other coral reef organisms that are reliant on corals for food, shelter or recruitment (e.g., Sano et al.

1989, Wilson et al. 2006. Pratchett et al. 2008). Climate-induced coral bleaching kills corals, but leaves the underlying skeleton completely intact (Hoegh-Guldberg, 1999).

Exposed coral skeletons are then subject to a whole suite of bio-eroding organisms that undermine the structural integrity of these carbonate structures (Hutchings,

1986). Reef habitats with reduced topographic complexity typically support lower fish abundance, fewer species, and increased evenness (Gratwicke and Speight 2005,

Graham et al. 2006). Global climate change is being caused by anthropogenic forcing of the climate system (Houghton et al. 2001), and not only are atmospheric concentrations of greenhouse gases rising, but the rate is accelerating (e.g. Canadell et al. 2007). Increases in atmospheric temperatures are expected to continue throughout next century, and are expected to accelerate over the next two decades due (Houghton et al. 2001). As a consequence, even if climatic impacts are not yet apparent (or have had minor influence compared to other more direct anthropogenic disturbances) the effects of global climate change on ecosystems, communities and species will become increasingly important in the coming decades. Australia’s Great

Barrier Reef and other important reef ecosystems (Ningaloo and many offshore coral reef systems) have been largely spared from devastating effects of climate change that have already been witnessed on reefs in the Indian Ocean and Caribbean (e.g.,

Graham et al. 2008). However, sustained and ongoing climate change will increasingly threaten marine fishes on Australian coral reefs over coming decades.



2.2.3 Population structure

Aside from immediate and detectable declines in the abundance of species, disturbances (fisheries exploitation and/ or habitat degradation) may significantly alter population structure. For example, populations subject to strong fishing pressure are expected to be comprised of fewer and smaller individuals, have reduced survivorship, and also, contribute very little to reproductive output and populations replenishment (Begg et al. 2005, Evans et al. 2008). For protogynous hermaphrodites, such as coral trout (Adams 2003), fishing will disproportionately affect the abundance of larger, older males (Vincent and Sadovy 1998). This may cause females to change sex earlier, thereby reducing the average age and size of females, which will greatly reduce reproductive output (Adams 2002). Evans et al

(2008) showed in Lutjanus carponatatus that larger females make a disproportionate contribution to reproductive output, whereby a 50% in total length may translate to a

100-fold increase in egg production. Even in cases where the size or age at which females change sex in inflexible, reproductive output of fished populations may be increasingly limited by the lack of males (Vincent and Sadovy 1998, Adams 2002).

Most studies ascertain effects of disturbances on coral reef fishes by quantifying changes in their distribution and abundance (e.g., Harmelin-Vivien and

Laboute 1986; Williams 1986; Kokita and Nakazono 2001). This implicitly assumes that such disturbances will lead to widespread mortality and/or migration. It is very likely however, that major changes in the distribution and abundance of fishes will be preceded by more subtle changes in population structure or individual fitness (e.g.,

Pratchett et al. 2004, 2006b). Pratchett et al. (2004) showed that there was no shortterm decline in the abundance of an obligate coral-feeding butterflyfish ( Chaetodon lunulatus ) despite a 55% decline in coral cover caused by mass bleaching. However,

C. lunulatus did exhibit significant declines in physiological condition (Pratchett et al.

2004), which contributed to reduced survival and eventual population declines

(Pratchett et al. 2006b). Similarly, reductions in live coral may limit settlement and recruitment for fishes that are otherwise unaffected by coral depletion (Booth and

Beretta 2002, Jones et al. 2004, Feary et al. 2007). Declines in the abundance of recruits and juvenile fishes will inevitably lead to population declines, but these effects may not be immediately apparent (Graham et al. 2007, Feary et al. 2007).



Populations with highly skewed size or age structure in either direction (ie. a lack of juveniles or limited large reproductive individuals) are likely to be subject severe population pressures or en route to extirpation. Several such anomalies in population structure may be used as an early warning of the poor status or compromised health of current populations (Table 4).



Table IV.

Criteria used to assess distinct components for current population status of

Australian marine fishes. Where there are detectable declines in the abundance of fishes, the rates used ( ≥30% and ≥50%) correspond with IUCN criteria for vulnerable and endangered species, respectively (Appendix I).

Indicator

Population trends

Extent of known threats

Population structure

Very good

Population very stable or increasing through time

No apparent threats to

Australian populations

Known threats have limited or very localised effects on populations

Population structure is normal and healthy throughout

Good

Populations fluctuating, but no apparent trend

Anomalies in population structure are rare or very localised

Poor

≥30% decline over the last 10 years or three generations

Multiple different or significant threats affect >30% of populations or individuals

Very poor

≥50% decline over the last 10 years or three generations

Multiple different or significant threats affect virtually all populations or individuals

Populations commonly have i) highly biased sex ratio, ii) strongly skewed size or age structure, or one or few cohorts dominate, or iii) compromised health and low local reproductive output

Virtually all populations have i) highly biased sex ratio, ii) strongly skewed size or age structure, or one or few cohorts dominate, or iii) compromised health and low local reproductive output

2.3 Population resilience

Increases in the frequency and severity of environmental perturbations are inevitable, due to climate change (Hoegh-Guldberg et al. 1999, Sheppard 2003,

Donner et al. 2005), and increasing direct anthropogenic pressure on marine species and ecosystems (Newton et al. 2007, Bell et al. 2009). As such, the persistence of fish populations and species will depend upon their resilience, and especially their capacity for recovery following recurrent disturbances. Holling (1973) defined ecological resilience as the amount of disturbance that a population, community or ecosystem can withstand without changing self-organizing processes and structures.

He also noted that the well-being of the world (and specially engendered animals) is not adequately described by concentrating on the current status of species, rather


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species we must look more deeply at the fundamental population dynamics (e.g., fecundity and reproductive output) in order assess the ultimate fate of populations and species

(Holling 1973).

2.2.1 Rates of recovery

Resilience to disturbances may be assessed based on i) resistance, which is the ability of populations and species to withstand major disturbances, and ii) recovery, which is the rate at which disturbed populations or species reassemble and retain their former structure and abundance. Most studies exploring effects of acute disturbances (e.g., severe tropical storms) on marine fishes focus almost exclusively on differential susceptibility (ie. resistance) among populations and species (e.g.,

Williams 1986, Cheal et al. 2002, Jones et al. 2004, Graham et al. 2006), yet established life-history theory on r-selected versus K-selected species (MacArthur and Wilson 1967) shows that persistence in highly disturbed environments is enhanced by rapid recovery in the aftermath of disturbances, not resistance. For fisheries species, there has been considerable research into recovery potential of population following fisheries closures (e.g., Hutchings and Reynolds 2004), and there are an increasing number of studies exploring recovery trajectories for fishes following localised environmental perturbations (e.g., Halford et al. 2004). For the most part, recovery is expressed as the time (in years) for the abundance of fishes to return to pre-disturbance levels (e.g., Halford et al. 2004).

Few fish species recover rapidly in the aftermath of heavy fisheries exploitation. Hutchings and Reynolds (2004) showed that most (85%) of fish populations exhibit limited or incomplete even after 15 years of fisheries closures following stock collapses. The potential recovery for fish populations is clearly related to the level of exploitation, and especially the extent to which reductions in stock size limit subsequent population growth due to the Allee effect (Hutchings and Reynolds

2004). Similarly, recovery of fishes affected by environmental perturbation depends on the severity and extent of habitat loss. In general, fishes appear to be fairly resilient to small-scale or moderate disturbances, readily returning to their predisturbance abundance and structure (Connell and Sousa 1983). Such disturbances generate significant spatial and temporal heterogeneity in the dynamics of populations and communties (Pickett and White 1986) but rarely have long-lasting


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species effects. Large-scale or severe disturbances, or the combined effects of multiple disturbances, may however, lead to long-term or permanent changes in the abundance of species (Berumen and Pratchett 2006). Permanent changes in the abundance of species (often referred to as “phase shifts”) tend to be associated with chronic environmental degradation.

Aside from the level of population depression, the recovery rates of individual species are also influenced by their life history, the degree habitat alteration and availability of essential resources, as well as changes to species assemblages and inter-specific competition. On coral reefs, disruption of fish communities (e.g., due to bleaching or outbreaks of A. planci ) may lead to sustained differences in species composition, even if benthic communities fully recover

(Berumen and Pratchett 2006). Pre-emption of space by new species might retard the recolonisation of some previous residents, although competitive hierarchies can eventually re-instate species as they gradually recruit back to the reef (Munday

2001). The recovery of specialist species will also be critically dependant on the recovery rates of their preferred corals (Halford et al. 2004). In the absence of any further disturbances, it may take as little as 5 years for coral cover to return to predisturbance levels (e.g. Halford et al., 2004), though this will depend on the severity and spatial extent of coral loss, which dictates the ability of surviving corals to recover, reproduce and reseed affected areas (Hughes and Connell 1999). Recovery will be much faster if at least some coral colonies survive the bleaching (e.g., Baird and Marshall 2002), because growth of surviving corals leads to more rapid increases in coral cover compared with settlement and subsequent growth of new individuals (Connell et al. 1997).

2.2.2 Reproduction and recruitment

In fisheries management, the concept of resilience generally refers to the capacity for populations to replenish themselves at normal versus depleted stock sizes (Musick 1999). This clearly relates to the capacity for recovery in the aftermath of disturbances (discussed above) and is influenced by species-specific reproduction and recruitment dynamics. Of greatest concern for the management of endangered species is the strength and nature of the relationship between stock size (local abundance of reproductive adults) and recruitment (Musick 1999). Most fishes tend


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species to have very high reproductive output, but exhibit considerable interannual variation in recruitment (e.g., Doherty 1991), which is often attributed to stochastic environmental factors that affect dispersal and survival of offspring. Stockrecruitment relationships may also be decoupled due to high levels of larval dispersal and strong inter-connection among disparate populations (e.g., Hughes et al. 2000).

If larvae are widely dispersed, the local production of offspring by sessile or sedentary adults will not be correlated with local recruitment; locally derived larvae go elsewhere, and recruits come from afar. Nonetheless, at larger spatial scales there may be some detectable relationship between the size of the spawning stock

(i.e., the amount of larval production) and the amount of recruitment. It is expected therefore, that moderate declines in stock size will not adversely affect recruitment rates (Musick 1999), such that these fish populations would be relatively resilient to moderate levels of fishing and/ or environmental perturbations. Major exceptions include fishes (e.g., sharks) with low fecundity and limited dispersal. The grey nurse shark ( C. taurus ), for example, gives birth to just two pups per litter and only reproduces once every two years (Gilmore et al. 1983). As a result, annual rates of population increase are very slow, greatly reducing its ability to sustain fishing pressure, and this species is now listed as Vulnerable (Pollard and Smith 2005).

The supply of larvae is fundamental in sustaining adult populations marine fishes, but the importance of larval supply in determining patterns of abundance depends on the extent to which population size is limited by larval supply relative to events occurring during or after settlement (Doherty 1998). For populations where enhanced recruitment would lead to greater adult abundance (“recruitment limited”:

Doherty 1998), changes to reproductive output and recruitment success will directly affect population size, and even species persistence. Aside from population connectivity (discussed below) there are three elements of reproduction and recruitment that will influence the vulnerability of species to disturbance (including direct exploitation and environmental perturbations): i) per capita reproductive output

(including individual fecundity and any intrinsic constraints on the occurrence of reproduction or fertilisation success), ii) average and inter-annual variation in rates recruitment, and iii) strength and scale of stock-recruitment relationships.

Importantly, these variables are not static, and may themselves change in response to changes in population status. In group-spawning fishes, reproductive strategies


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species and per capita reproductive success are strongly influenced by group size (e.g.,

Warner 1982), but even for monogamous species reproductive success may be enhanced by the density of con-specifics (Thresher 1985), due to reduced predation rates. In general, there is limited knowledge of the extent to which marine fish populations are limited by recruitment, let alone local reproductive output (Caley et al. 1996). However, there is significant knowledge on the reproductive behaviour and/ or recruitment dynamics of many different fishes, which could be used to provide a qualitative assessment of populations resilience and vulnerability (Table 5).

2.2.3 Population connectivity

Coastal habitats (e.g., seagrass beds, coral reefs, and rocky reefs) almost invariably occur as a series of discrete habitat patches separated at varying distances by alternative habitat types (e.g., sandy plains) that are generally uninhabitable for the majority of species (Kritzer and Sale 2004). The persistence of populations and species is therefore, dependent upon the number and size of different habitat patches, the abundance of fishes within each habitat, and critically, the level of connectivity among habitat patches. At the level of individual habitat patches (e.g., individual reefs), connectivity relates to the proportion of new recruits that originate from external patches, compared to the proportion of recruits that originate from the natal reef. Populations that suffer severe declines in abundance will be slow to recover if recruitment is predominantly from local sources (i.e. high levels of self-recruitment). If however, significant levels of recruitment comes from outside sources then recovery may occur via the long distance dispersal of recruits from unaffected populations (Hughes et al. 2005, Jones et al. 2009). Recent studies on population connectivity among coral reef fishes provide strong evidence for ecologically significant levels of self-recruitment (Swearer et al 2002, Jones et al.

2009). On the other hand, some recruits must travel vast differences to maintain the genetic integrity of widespread species (e.g., Palumbi 2003). It is currently unknown, however, what is the mean dispersal distance for recruits originating from external sources, or more importantly, what are the relative rates of dispersal over varying distances (Kritzer and Sale 2006).

Connectivity between disparate populations directly influences species persistence and population stability in two ways; i) high levels of population


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species connectivity may help rescue declining populations and enable recolonisation of habitat patches following localised extinction (Hanski 1999, Pannell 2003, Jones et al. 2009), and ii) high levels of connectivity also contribute to increased genetic diversity within sub-populations, which can have independent effects on the resilience of populations (Frankham 2005; Frankham 2010). Low gene flow can lead to genetic drift and loss of genetic diversity (Keller and Waller 2002), reducing adaptive capacity and increasing the vulnerability of local populations to disturbance.

Loss of genetic diversity (due to limited genetic exchange with disparate populations and/ or population bottlenecks that result when relatively few individuals contribute to the genetic diversity of recovering populations) can increase extinction risk as it reduces the ability of populations to evolve and cope with environmental change

(Spielman et al. 2004, Frankham 2010).



Table V.

Criteria used to assess distinct components for current population status of

Australian marine fishes. Where there are detectable declines in the abundance of fishes, the rates used ( ≥30% and ≥50%) correspond with IUCN criteria for vulnerable and endangered species, respectively (Appendix I).

Indicator

Observed recovery

Reproductive mode and recruitment

Population connectivity

Very good Good Poor Very poor

Full recovery typically takes <5 years

Full recovery typically takes

<10 years

Full recovery takes >10 years, or is constrained by extrinsic factors

Recovery is very slow (>15 years) or never complete

Very high levels of reproduction and recruitment

(no evidence of recruitment- limitation)

Reproductive output and recruitment are rarely limiting

Low reproductive output and/ or recruitment generally limit population recovery

Very low reproductive output and/ or recruitment already constrain population size

High connectivity and weak genetic structure among regions

High connectivity and weak genetic structure within regions

Limited evidence of connectivity (or strong genetic structure) within regions

Limited evidence of connectivity (or strong genetic structure) at the level of reefs



3. Specific species assessments

3.1 Redfin butterflyfish ( Chaetodon lunulatus )

3.1.1 Inherent vulnerability

The redfin butterflyfish, Chaetodon lunulatus Quoy and Gaimard 1824 is distributed throughout the Indo-Pacific Archipelago and eastern Pacific Ocean (Allen et al. 1998). Its’ geographic range is estimated to be 56.6 million km 2 from ranges estimates by Jones et al. (2002) based on projection of distribution maps from Allen et al. (1998). This is more than twice the average range size for butterflyfishes (21.2 million km 2 ) and well outside the range (0.8 million km 2 ) considered to be indicative of restricted range species (Hawkins et al. 2000). The geographic range of this species is estimated to encompass nearly 200,000km 2 of coral reef habitat, including the entire area of Australia’s Great Barrier Reef. C. lunulatus is also among the most abundant butterflyfish species on reefs throughout its’ geographic range (e.g.,

Anderson et al. 1981; Bouchan-Navaro et al. 1985). Although rarely dominant (in terms of abundance), C. lunulatus is common across a wide range of different habitats (Pratchett and Berumen 2008), which contributes greatly to overall population size. In the absence of major disturbances, densities of C. lunulatus are typically 25-40 fish per 1000m 2 (Pratchett et al. 2004, Emslie et al. 2011).

Chaetodon lunulatus is one of 25 species of obligate coral feeding butterflyfishes (Cole et al. 2008), and feeds exclusively on living tissues from scleractinian corals (Harmelin-Vivien and Bouchan-Navaro 1983, Pratchett 2005).

Even though C. lunulatus feeds mostly on Acropora spp. (Figure 4), it is among the most generalist of obligate coral-feeding butterflyfishes (Pratchett et al. 2004,

Pratchett 2005). At Lizard Island, C. lunulatus was seen to consume at least 52 different species of scleractinian coral and in relatively equal proportions, whereas the more specialised species fed predominantly on just 1-2 different coral species

(Pratchett 2005). It is also one of the only species of coral-feeding butterflyfishes that will readily consume massive Porites . This flexibility in the dietary composition enables C. lunulatus to withstand short term or moderate declines in the availability of coral prey (Pratchett et al. 2004).



Figure 4. The redfin butterflyfish, Chaetodon lunulatus , taking bites from a colony of

Acropora . Photograph by M. Pratchett.

In the most recent Red List Assessment (Rocha et al. 2009), Chaetodon lunulatus was listed as “Least Concern”. The justification for this assessment stated that “while there have been declines documented in some areas (e.g., Pratchett et al. 2006b), these are not believed to have substantially affected the global population. In addition, it has a wide distribution, large population and no apparent major threats other than coral loss”.



Table VI . Indicators of inherent vulnerability to extinction for Australian populations of the redfin butterflyfish, Chaetodon lunulatus . The level of confidence in assigning assessment grades (high, medium, low) depends largely on the availability of quantitative data.

Inherent vulnerability (species) Assessment Grade

Indicator Summary

Very low

Low High Very high

Geographic range

Population size


The overall geographic range is 56.6 million km 2 , extending throughout the

Pacific

Normal densities of C. lunulatus on the

Great Barrier Reef are 25-40 fish per

1,000m 2 , and very consistent across a range of different habitats

C. lunulatus consumes virtually all available corals and is thus, resilient to all but the most severe disturbances


Marked changes in coral composition are very likely, but extensive loss of all coral species is unlikely at least in the short to medium term

Overall Assessment

GOOD – Chaetodon lunulatus has a low risk of extinction due to its’ large geographic range, but its’ strong reliance on corals and potential for extensive coral loss put it at slightly higher risk.

High confidence – The quality and extent (spatial, temporal and/ or taxonomic) of data enables firm predictions about population status

Medium confidence – There are appropriate data with which to complete assessments, but the generality of these findings is unclear

Low confidence – No credible data sources are available, but anecdotal evidence or experience gained from related species suggests there is or is not cause for concern

3.1.2 Current population status

Chaetodon lunulatus , like all obligate coral-feeding butterflyfishes, is adversely affected by severe disturbances (e.g., coral bleaching) that cause extensive coral loss (e.g., Pratchett et al. 2006b). However, given their generalised patterns of prey use, only the most severe disturbances (e.g., outbreaks of A. planci or extensive coral bleaching) have any discernible effect on the abundances of C.


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species lunulatus (Emslie et al. 2011), and even when they are effected, declines in abundance of C. lunulatus often lag behind that of other more specialised coralfeeding butterflyfishes (Pratchett et al. 2006b). At Trunk Reef, in the central GBR, coral cover declined >90% between 2000 and 2005 (Pratchett et al. 2006b) due to severe coral bleaching in 2001-02 (Berkelmans et al. 2004) followed by outbreaks of

A. planci in 2003-2004 (Sweatman et al.

2005). Corresponding with these extensive declines in live coral cover, there was initially no change in abundance from 2000-

2002, but mean densities of C. lunulatus declined by 97% from 2002-2005 (Figure 5;

Pratchett et al. 2006b). Such declines in the abundance of C. lunulatus are however, fairly localised, and despite sustained and ongoing declines in total coral cover on the Great Barrier Reef (Bellwood et al. 2004, Sweatman et al. 2011) densities of C. lunulatus have not changed from 1994 to 2009 (Figure 5). If anything the long-term trend in the abundance of C. lunulatus shows a moderate increase in mean densities.

20

15

10

5

0

1990 1995 2000 2005 2010

Figure 5 . Temporal trends in the abundance of Chaetodon lunulatus , based on surveys conducted in 6-9 m on the north-east flank at 47 reefs across the length and breadth of the

Great Barrier Reef. Data was provided by M. Emslie from the AIMS-LTMP.

Sustained declines in the live coral cover on the Great Barrier Reef, caused by increasing cumulative impacts from natural disturbances, climate change and


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species direct anthropogenic disturbances (Bellwood et al. 2004, Sweatman et al. 2011) pose a significant threat to populations of C. lunulatus , as well as many other coraldependent fishes. To date, the major causes of coral depletion on the Great Barrier

Reef have been i) cyclones and ii) outbreaks of A. planci (Osborne et al. 2011).

However, it is likely that climate change will become increasingly important over the next two decades. Based projected increases in sea surface temperatures, massbleaching episodes are expected to occur at least once every two years by 2030

(Donner et al. 2005). By 2050, most coral reefs are expected to be subject to annual thermal anomalies equivalent to those that cause global devastation in 1998 (Hoegh-

Guldberg 1999). The resulting extent of coral depletion may however, be partially reduced by acclimation and adaptation among coral populations and communities

(Hughes et al. 2003). Moreover, susceptibility to bleaching, and particularly proportional mortality due to bleaching, varies greatly within and among coral genera

(e.g., Marshall and Baird 2000, Loya et al. 2001). Therefore, increased incidence of climate-induced coral bleaching is more likely to cause marked changes in the structure of coral assemblages, rather than killing all corals over similar time frames

(Hughes et al. 2003). While the relative abundance of different corals is almost certain to change, it is likely that scleractinian corals will remain a dominant component of reef benthos (at least on mid-shelf and offshore reefs) for the foreseeable future. However, it is unclear to what extent the availability of corals limits current populations, let alone future populations of C. lunulatus .

Despite their abundance and prominence in reef fish assemblages, the demography and population structure of Chaetodon butterflyfishes have been very poorly studied. Chaetodon lunulatus is one of the only butterflyfishes for which there have been any life-history studies (e.g., Berumen 2005, Pratchett et al. 2006a), but the spatial and temporal extent of sampling is very limited. Berumen (2005) constructed size-at age curves for 41 fishes collected from Lizard Island, northern

Great Barrier Reef. These data showed that C. lunulatus live in excess of 12 years, and have strongly asymptotic (acanthuroid) growth. From the samples collected at

Lizard Island, the maximum size was ~104mm total length and individuals tend to reach 92% of their maximum size within the first 2 years (Berumen 2005). Most fishes reach sexual maturity at 80 –84 mm SL (Pratchett et al. 2006a), corresponding with an age of 1-2 years. Great Barrier Reef populations of C. lunulatus appear to


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species have good representation of both young and old individuals (Berumen 2005), and approximately equal abundance of both sexes (Pratchett et al. 2006a).

Table VII . Indicators of current population status for Australian populations of the redfin butterflyfish, Chaetodon lunulatus .

Population status

Indicator Summary

Very good

Assessment Grade

Good Poor Very poor

Population trends


Despite sustained losses of live coral on the Great Barrier Reef, densities of C. lunulatus have not changed from 1993 to 2009.

The biggest threat to C. lunulatus is acute disturbances (e.g., coral bleaching) that cause coral loss. These disturbances are expected to increase in severity and spatial extent throughout coming decades


Information on demographics and population structure are extremely limited, but there is one study showing good representation of old (up to 13 years) and young cohorts

Overall Assessment

GOOD – Australian populations of Chaetodon lunulatus appear very healthy, though there is a considerable threat posed by climate change, which is expected to cause frequent episodes of extensive coral loss.




3.2.3 Population resilience

Ongoing monitoring of C. lunulatus at Trunk Reef shows no recovery in the abundance of C. lunulatus >5 years after the stock collapsed (Figure 6). Since 2005, coral cover has increased three-fold, but there has been corresponding recovery in the abundance of C. lunulatus (Figure 6). These findings contradict previous studies


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species that showed rapid recovery among butterflyfish populations, which closely matched proportional increases in the abundance of live coral cover (e.g., Halford et al.,

2004). Similarly, Emslie et al. (2011) found high levels of resilience among coral feeding butterflyfishes, whereby disturbances had only moderate and short-lived effects on the abundance of C. lunulatus.

Recovery of butterflyfish populations at

Trunk Reef appears to be limited by a lack of new recruits. Recruitment rates by butterflyfishes are generally low (Pratchett et al. 2008b) and may have been further constrained by limited coral cover at potential settlement sites, as well as widespread depression of breeding populations.

40

35

30

Coral cover

Butterflyfish

10

8

25

20

15

6

4

10

5

2

0 0

2000 2002 2005 2008 2009 2010

Figure 6 . Temporal variation in the abundance of Chaetodon lunulatus , at Trunk Reef in the central Great Barrier Reef. Significant declines in the abundance of corals, and therefore, C. lunulatus were caused by bleaching (2001-02) and subsequent outbreaks of A. planci (2003-

2005) (Pratchett et al. 2009b)

No attempt has been made to directly quantify the proportion of recruits from local versus external sources for C. lunulatus , though Almany et al (2008) revealed high rates (63%) of self-recruitment in C. vagabundus at Kimbe Bay, northern Papua

New Guinea. Using allele frequencies of 12 microsatellite loci and >350 bases of mitochondrial control region sequence, Lawton et al. (in review) explored genetic structure of C. lunulatus from five locations across the Pacific Ocean (Lizard Island,

North Great Barrier Reef; Heron Island, South Great Barrier Reef; Kimbe Bay, Papua



New Guinea; Noumea, New Caledonia; and Moorea, French Polynesia). Significant genetic differentiation was detected among locations in both the microsatellite and mtDNA data, suggesting that there is limited genetic exchange at this large scale.

Together these data suggest that there may be low levels of connectivity among disparate populations (on widely separated reefs) and certainly among populations in different geographic regions (e.g., the Great Barrier Reef versus Papua New

Guinea). If so, the large geographic range of C. lunulatus is unlikely to offer any insurance against localised extirpations.



Table VIII . Indicators of population resilience for Australian populations of the redfin butterflyfish, Chaetodon lunulatus .

Population resilience Assessment Grade

Indicator

Observed recovery

Summary

Recovery is constrained by recovery rates for prey corals, which varies with the severity and spatial extent of disturbances.

Very good

Good Poor Very poor


C. lunulatus forms monogamous breeding pairs and spawns monthly throughout much of the year.

Recruitment rates are relatively low

(<5% of adult population), but local populations appear to be maintained by relatively low recruitment.


Recent genetic studies (microsattelites) reveal high levels of genetic structure among populations in the south Pacific

Overall Assessment

POOR – Australian populations of Chaetodon lunulatus exhibit strong recovery in the aftermath of major disturbances.



Low confidence – No credible data sources are available, but anecdotal evidence or experience gained from related species suggests there is or isn’t cause for concern



3.2 Leopard coral grouper ( Plectropomus leopardus )

3.2.1 Inherent vulnerability

The leopard coral grouper (also known as common coral trout),

Plectropomus leopardus Lacèpde 1802 is found mainly in the western Pacific from southern Japan to Australia, and eastward to the Caroline Islands and Fiji (Cornish and Kiwi 2004).

Its’ geographic range is estimated to be 22.7 million km 2 (Jones et al. 2002). This is greater than the average geographic range for epinepheline groupers (19.1 million km 2 ; Jones et al. 2002) and well outside the range (0.8 million km 2 ) considered to be indicative of restricted range species (Hawkins et al. 2000).

Within Australia, P. leopardus is common along the length of the Great Barrier Reef

(mostly on mid- and offshore reefs), and also occurs throughout northern Australia and Western Australian. Plectropomus leopardus is not common, but densities on the Great Barrier Reef (approximately 3.5 per 1,000 m²) are much higher than elsewhere throughout it’s geographic range (Ayling et al . 2000).

Figure 7. The leopard coral grouper (also known as common coral trout), Plectropomus leopardus . Photograph by A. Frisch.



Plectropomus leopardus is strongly associated with coral reef habitats (e.g.,

Kingsford 2009)

, but its’ specific reliance on habitat and prey resources requires further investigation. Comprehensive dietary analyses based on stomach contents of adult individuals revealed that P. leopardus feed mostly on fishes, but also some crustaceans and other invertebrates (St John et al. 2001). Pomacentridae and

Clupeidae dominated the fish component of the diet, suggesting that P. leopardus will opportunistically exploit both reef-based and pelagic food sources (St John et al.

2001), such that these fishes might expected to be insensitive to changes in the reef habitats and the abundance of reef-associated prey. It appears however, that that coral trout (particularly, P. maculatus ) settle in close proximity with live branching corals and rely heavily on reef-based sources of prey, at least as juveniles (C. Wen and M. Pratchett, unpublished data). There is also evidence that expected increases in the abundance of Plectropomus sp. following fisheries closures at the Keppel

Islands (inshore Great Barrier Reef) were not realised, due to degradation of coral reef habitats that coincided with the implementation of fisheries closures (Russ et al.

2008; Figure 7). At most locations (where there was little or no loss of coral habitat),

Russ et al (2008) documented a 53-68% increase in the abundance of coral trout within 1.5-2 years after fisheries closures, whereas at the Keppel Islands the abundance of coral trout declined 19% despite fisheries closures.

Plectropomus leopardus is listed as Near Threatened in the IUCN red list

(Cornish and Kiwi 2004). This conclusion was largely based on purported declines in the abundance of P. leopardus on the Great Barrier Reef, Australia. However, the reference cited, Ayling et al. (2000) actually shows that declines in the abundance of

P. leopardus on the Great Barrier Reef are fairly localised (see section 3.2.2). Given its’ large size, importance in coral reef fisheries, and tendency to form large spawning aggregations, Morris et al. (2000) suggested that P. leopardus should listed as Vulnerable under criteria A2d: projected future population decline due to actual or potential levels of exploitation.



Table IX . Indicators of inherent vulnerability to extinction for Australian populations of the leopard grourper, Plectropomus leopardus .

Inherent vulnerability (species)

Indicator

Geographic range

Population size

Summary

The overall geographic range is 22.7 million km 2

Average densities of P. leopardus on the

GBR are generally low, though it found in a wide range of habitats

Very low

Assessment Grade

Low High Very high


Very versatile in terms of diet, but may be reliant on corals at settlement


Marked changes in coral composition are very likely, but extensive loss of all coral species is unlikely at least in the short to medium term

Overall Assessment

GOOD – Plectropomus leopardus has a reasonably large geographic range, but its’ low abundance and potential reliance on corals makes increases the risk of local extinction.




3.2.2 Current population status

Plectropomus leopardus is the major target species of the reef line fishery in

Australia, and accounts for nearly 50% of the total commercial catch (e.g., Mapstone et al. 1996). Commercial and recreational line fishing can have a marked effect on coral trout stocks (Ayling and Ayling 1998) whereby fishers removed almost 80% of adult stock in 12 months (estimated at over 12,000 fish) on Bramble Reef when it was to fishing in 1995. However, commercial fisheries for P . leopoardus on the Great

Barrier Reef are generally considered to be sustainable (Mapstone et al 1996). Most epinepheline fishes are long lived and slow growing (Morris et al. 2000), but coral


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species trout grow rapidly and are relatively short lived; Growth is very rapid over the first three years of life and the mean total length of 3+ year old fish is around 40 cm total length (TL) (Ferreira and Russ 1994).

The size limit for P. leopardus in the Great

Barrier Reef (GBR) line fishery is 38 cm TL, meaning that fishes caught may be as little at 2-3 years of age. Fisheries independent measures of population status (e.g., underwater visual census) show that densities of P. leopardus are certainly more common on reefs closed to fishing, compared to fished reefs (e.g., Williamson et al.

2004). In areas with highest fishing pressure (in reefs off Townsiville) there is also evidence of a recent decline in the abundance of P. leopardus (Ayling et al. 2000).

However, reef-wide surveys undertaken by the AIMS-LTMP, do not reveal any consistent long-term decline from 1993 to 2011 (Figure 6).

1,5

1

0,5

0

1991 1996 2001 2006 2011

Figure 8. Temporal trends in the abundance of Plectropomus leopardus , based on surveys conducted in 6-9 m on the north-east flank at 47 reefs across the length and breadth of the

Great Barrier Reef. Data was provided by M. Emslie from the AIMS-LTMP.

Aside from directly removing individuals and reducing population size, fishing is considered to have potentially detrimental consequences on the reproductive capacity of protogynous populations by causing sperm limitation via the disproportionate removal of large males. Adams et al. (2000) tested this by comparing population structure of P. leopardus on 4 reefs that had been closed to


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species fishing for a minimum of 8 years, and four fishes open to fishing. As expected,

Adams et al. (2000) found a much a greater proportion of males above the minimum size and age of harvest on reefs closed to fishing. However, differences between open and closed reefs were not consistently expressed as a reduction in the average male size/age or a highly female biased sex ratio, as might have been expected if fishing removes a disproportionate number of large males. In all, population structure of P. leopardus was more variable between different reefs and locations, rather than between fished versus unfished reefs within each region (Adams et al. 2000). One obvious difference was that females were proportionately more abundant and significantly larger and older on closed reefs (Adams et al. 2000). This is may be expected to affect reproduction and local recruitment (discussed below), as large females make a disproportionate contribution to reproductive output of most marine fishes (e.g., Evans et al. 2008).



Table X . Indicators of current population status for Australian populations of the leopard coral grouper, Plectropomus lepoardus

Population status Assessment Grade

Indicator Summary

Very good

Good Poor Very poor

Population trends


Despite sustained fisheries, reef-wide populations on the GBR show no evidence of decline. There are no data on population trends for trout populations in other parts of Australia, but they are certainly less abundant in

Western Australia

The biggest threat to P. leopardus is fishing (commercial and recreational), which affects >30% of populations.

Improved data is needed to assess the impact of recreational fisheries, but current fishing effort appears sustainable.


Fished populations have far fewer large individuals, but there were no consistent differences in population structure

Overall Assessment

GOOD – The abundance of Plectropomus leopardus fluctuates through time, but there is no evidence of a systematic decline in abundance. The main threat to P. leopardus is widespread fishing, but this has not adversely affected population dynamics.




3.2.3 Population resilience

The 2004 rezoning of the Great Barrier Reef, whereby many reefs previously open to fishing were placed within no-take marine reserves, provided a unique opportunity to test the population resilience of coral trout (Russ et al. 2008). At most of these reefs (with the exception of Keppel Islands, discussed above), the abundance and biomass of Plectropomus sp. have recovered very rapidly (Russ et


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species al. 2008). Within 1.5-2 years, densities of coral increased 57-75% in seven of eight regions (Figure 9). The proportional increases in localised densities of coral trout on reefs that were recently closed to fishing were also very consistent among reefs separated by up to 1,000 km, whereas there was little or no change in densities on reefs that remained open to fishing (Russ et al. 2008).

Figure 9 . Changes in the densities of Plectropomus leopardas (number per 1000m 2 ) on reefs that were recently closed to fishing, following rezoning of the Great Barrier Reef in

2004. Data from Russ et al. (2008).

Fishes of the family Serranidae are considered highly vulnerable to recruitment overfishing, owing to their tendency to form spawning aggregations

(Morris et al. 2000). Spawning activity by P. leopardus is concentrated, but no limited to, a small number of spawning sites at each reef (Samoilys 1997). Fishes aggregate around spawning sites for an average of 5 days, corresponding with each new moon during the spawning period from September to November. Fishers tend to be well aware of the position of primary spawning sites, and the timing of spawning aggregations. Targeted fishing is believed to have contributed to the loss of at least one major spawning aggregation of P. leopardus on the Great Barrier Reef (Cornish and Kiwi 2004).

A major objective of spatial management (e.g., implementation of no-take fisheries reserves) is to protect a portion of critical spawning stocks to ensure supply


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species of larvae to fished areas via larval dispersal. For P. leopardus , Adams et al. (2000) showed that females are significantly larger and older on reefs that were closed to fishing, such that reproductive output of populations protected from fishing would be expected to be considerable higher compared to fished populations. There is however, very little empirical data to test whether increased reproductive output of populations in no-take marine reserves contributes to larval supply in nearby fished locations. Quantifying recruitment rates of P. leopardus is particularly difficult due to the apparent scarcity and crytptic nature of recruits. At reefs closed to fishing, populations appear to be “sustained by a trickle of recruits” (Kingsford 2009), and

Willamson et al (2004) found no difference in the number of small coral trout (<10 cm

TL) on reefs that were closed versus open to fishing.

Genetic studies suggest that there is strong genetic structure, and limited larval exchange, at the scale of geographical regions (e.g., between populations on the Great Barrier Reef versus New Caledonia), but there is limited genetic structure, and high levels of larval exchange within regions. van Herwerden et al (2009) sampled P. leopardus from four regions (six locations) and partial mtDNA D-loop sequences revealed six strong differentiated populations. There was however, limited genetic structure between the northern and southern Great Barrier Reef, and evidence of larval transport from New Caledonia to the Great Barrier Reef (van

Herwerden 2009). Simialrly, phylogenetic and population genetic analyses, using mitochondrial DNA (mt DNA) from the control region, revealed limited genetic structure within and between distinct reefs for the inshore coral trout, P. maculatus

(Evans et al. 2010). These results suggest that there is a high level of connectivity among reefs on the Great Barrier Reef, which would enable replenishment of fished populations by larval supply from populations within no-take marine reserves.



Table XI . Indicators of population resilience for Australian populations of the leopard coral grouper, Plectropomus lepoardus .

Population resilience Assessment Grade

Indicator

Observed recovery

Summary

String recovery has been observed following the implementation of recent fisheries reserves, though recovery can be constrained by the habitat quality

Very good


Spawning activity is concentrated, but no limited to, a small number of spawning sites per reef.

Recruitment rates appear to be naturally low, but there is no evidence of recruitment-limitation at fished reefs

Good Poor Very poor


Recent genetic studies reveal high levels of structure among, but not within, geographical regions

Overall Assessment

GOOD – Populations of Plectropomus leopardus appear to be very resilient to relaxation of fishing presusure. There is considerable potential for fishing to reduce local reproductive output, but recruitment is maintained by high connectivyt among reefs



Low confidence – No credible data sources are available, but anecdotal evidence or experience gained from related species suggests there is or isn’t cause for concern



4. Conclusions

Over-exploitation and habitat degradation are the two primary drivers of population declines and species extinctions across a wide range of habitats (Ludwig et al. 1993, Brooks et al. 2002, Dulvy et al. 2003). Both these disturbances may have devastating effects on the abundance of populations or species, but also have far reaching effects on community structure and ecosystem processes (Wilson et al.

2008). Over-exploitation has caused extirpation or extinction of many target species, with secondary effects on community composition via the release of meso-predators

(e.g., Crooks and Soule 1999) and prey (e.g., Frank et al. 2005). This may have cascading effects on lower trophic levels, affecting diversity and primary productivity of the system (Crooks and Soule 1999, Schmitz et al. 2000). Habitat alteration and degradation (e.g. through direct and indirect anthropogenic disturbances) also have obvious impacts on community structure and dynamics (Walther et al. 2002).

Clearing of Australian temperate woodlands, for example, has reduced habitat heterogeneity and structural complexity resulting in species losses among reptiles

(Brown 2001) and birds (Jansen and Robertson 2001). The relative importance of over-exploitation versus habitat degradation varies among populations and species

(Wilson et al. 2008, Graham et al. 2011a), but when operating in concert, overexploitation (top-down effects) and habitat degradation (bottom-up) can have strongly interacting and potentially dire consequences for biodiversity and ecosystem function (Pace et al. 1999, Travis 2003).

Coastal fisheries tends to target larger individuals and species at higher trophic levels (Pauly et al. 1998), whilst habitat alteration and environmental perturbations tend to have the greatest affect on highly specialised and small-bodied benthic feeders (Pratchett et al. 2008a). Among coral reef fishes, Graham et al.

(2011a) found a strongly convex relationship among the species affected by fisheries exploitation versus climate-induced habitat degradation, whereby species are vulnerable to one or other threat, but rarely both. This finding reduces the possibilities of strong synergistic effects of fishing and climate change on individual species. However, among those fishes that are vulnerable to fishing are key functional groups (e.g., roving herbivores) that provide resilience against sustained effects of climate change (e..g, Hughes et al. 2007, 2010). Graham et al. (2011a)


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species suggested therefore, that effective fisheries management will buffer against the negative effects of climate-induced habitat degradation, and buy much needed time to address the more complex global problem of reducing atmospheric CO

2

levels.

Herbivorous fishes, such as parrot fishes (e.g., Bellwood et al. 2003) and rabbit fishes (Pratchett et al. 2011a) are heavily exploited throughout the Indo-Pacific, but herbivorous fishes are rarely targeted in Australia and comprise a very minor component (<0.5%) of coral reef fisheries. Rather, fisheries exploitation is limited to large piscivorous fishes, such as coral trout ( Plectropomus spp.) and other

Serranidae, as well as Lethrinidae and Lutjanidae (e.g., Williamson et al. 2004). As such, fisheries management may offer very little in indirect benefits for benthic habitats and coastal ecosystems (Graham et al. 2011b). Conversely, management actions that reduce and reverse widespread habitat degradation will contribute greatly to sustained ecological and economic benefits from coastal ecosystems

(Pratchett et al. 2011b).

Australia’s aquatic ecosystems are unique, supporting a high diversity of species and high levels of endemism (Unmack 2001, Poloczanska et al. 2007), but severe and wide spread habitat degradation poses a considerable threat to

Australia’s fish diversity, and this is expected to worsen with ongoing climate change

(Pratchett et al. 2011b). For the most part, effects of climate change compound upon habitat degradation and losses that have occurred as a result of more localised anthropogenic disturbances, such as exploitation, pollution, and habitat modification

(e.g., Hughes et al. 2003). The cumulative effects of exploitation, pollution, and habitat modification on individual species may also make them much more vulnerable to climate change (e.g. Wooldridge and Done 2009). Immediate reductions in global greenhouse gas emissions are critical to reduce long-term and potentially catastrophic impacts arising from severe climate change (Veron et al.

2009). However, there are also considerable opportunities to reduce and reverse anthropogenic degradation of coastal habitats at local and regional scales, which may have the added benefit of increasing resilience to future effects of sustained and ongoing climate change (Hughes et al. 2003).

As increasing anthropogenic pressures and ongoing climate change pose even greater threats to natural ecosystems, effective management and conservation


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species of marine species requires a rigorous framework for predicting which species are most vulnerable and at greatest risk of extirpation and extinction. Aside from increased research to understand the factors that determine inter-specific differences in the vulnerability of species (e.g., Graham et al. 2011a), this requires an explicit understanding of contemporary factors that threaten marine ecosystems and species, as well as ongoing studies to assess changes in the population status of individual species. The two species considered in this report ( C. lunulatus and P. leopardus ) are facing significant threats, due to habitat degradation (especially coral loss) and direct fisheries exploitation, respectively. However, the inherent vulnerability to extinction for both these species is considered Low , which is mostly reflective of their reasonably large geographic range and ability to utilise a wide range of different reef habitats. The current population status of both species is also considered Good , with no evidence of long-term and large-scale declines in abundance despite the aforementioned threats to local populations. However, C. lunulatus and P. leopardus appear to differ in their population resilience; C. lunulatus appears to have Poor resilience, based on limited recovery in the aftermath of severe coral bleaching in the central Great Barrier Reef. In contrast, populations of

Plectropomus spp. have recovered very quickly following fisheries closures on the

Great Barrier Reef, and if current population growth is sustained (> 30% per year,

Russ et al. 2008) they are expected to attain densities seen within established notake marine reserves (e.g., Williamson et al. 2004) within 5-10 years.

As habitat perturbations become more frequent and more severe, population resilience will have an increasing importance in species persistence versus extinction. Two critical components of resilience are; i) resistance, which is the ability of populations or species to survive acute environmental perturbations, and ii) recovery, which is the rate at which organisms recruit and grow to replenish depleted populations, which together determine the long-term persistence of populations and species (Elmqvist et al. 2003). A primary determinant of resistance to sustained and ongoing habitat degradation (and associated resource depletion) will be the ecological versatility of fishes and their ability to use alternative resources if the abundance of preferred prey or habitat declines (Pratchett et al. 2004). Over longer time periods, the recovery of depleted populations will depend on the persistence of viable brood stock (either locally or at other well connected locations) and sustained


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species recruitment (Pannell 2003). In the case of C. lunulatus , population resilience appears to be constrained by i) strong reliance on live corals (for food and settlement), such that recovery rates are limited by recovery in the local abundance of suitable coral species, and ii) seemingly low levels of connectivity and larval exchange among disparate populations (on widely separated reefs) and certainly among populations in different geographic regions (e.g., the Great Barrier Reef versus Papua New

Guinea). As such, Australian populations of C. lunulatus will be extremely vulnerable to increases in the diversity, frequency and severity of disturbances that are expected to occur as a consequence of climate change (e.g., Donner et al. 2005).

Moreover, the large geographic range of C. lunulatus is unlikely to offer any insurance against localised extirpations, thereby highlighting the need to consider population resilience when assessing vulnerability and population status of marine fishes.



References

Adams S (2002) The reproductive biology of three species of Plectropomus (Serranidae) and responses to fishing . PhD thesis, James Cook University, Townsville.

Admas S, Russ GR, Mapstone BD, Davies CR (2000) Geographic variation in the sex ratio, sex specific size, and age structure of Plectropomus leopardus (Serranidae) between reefs open and closed to fishing on the Great Barrier Reef. Canadian Journal of Fish and Aquatic Sciences 57: 1448-1458.

Aeby G, Lovell E, Richards Z, Delbeek JC, Reboton C, et al. (2008) Acropora hyacinthus . In:

IUCN 2010. IUCN Red List of Threatened Species. Version 2010.4. < www.iucnredlist.org/apps/redlist/details/133479/0>

Allen GR, Steene R, Allen M (1998) A Guide to Angelfishes and Butterflyfishes.

Odyssey

Publishing, Perth, Australia.

Almany GR (2003) Priority effects in coral reef fish communities. Ecology 84: 1920-1935.

Alongi DM (2002) Present state and future of the world’s mangrove forests. Biological

Conservation 29: 331 –349.

Anderson GRV, Ehrlich AH, Ehrlich PR, Roughgarden JD, Russell BC, Talbot FH (1981)

The community structure of coral reef fishes. American Naturalist 117: 476-495.

Australia State of the Environment Committee (2001) Australia state of the environment

2001 - Independent Report to the Commonwealth Minister for the Environment and

Heritage . CSIRO Publishing on behalf of the Department of the Environment and

Heritage, Canberra.

Ayling AM, Ayling AL (1998) Bramble Reef replenishment area: third post-opening survey .

Report to the Great Barrier Reef Marine Park Authority, Townville.

Ayling T, Samoilys M, Ryan D (2000) Trends in common coral trout populations on the Great

Barrier Reef . Department of Primary Industries, Queensland, Brisbane.

Baird AH, Marshall PA (2002) Mortality, growth and reproduction in scleractinian corals following bleaching on the Great Barrier Reef. Marine Ecology Progress Series 237:

133-141

Beeton RJS, Buckley KI, Jones GJ, Morgan D, Reichelt RE, Trewin D (2006) Australia State of the Environment 2006 . Independent report to the Australian Government Minister for the Environment and Heritage, Department of the Environment and Heritage,

Canberra.



Beets J and Hixon MA (1994) Distribution, persistence, and growth of groupers (Pisces:

Serranidae) on artificial and natural patch reefs in the Virgin Islands. Bulletin of Marine

Science 55: 470-483.

Begg GA, Mapstone BD, Williams AJ, Adams S, Davies CR, Lou DC (2005) Multivariate lifehistory indices of exploited coral reef fish populations used to measure the performance of no-take zones in a marine protected area. Canadian Journal of

Fisheries and Aquatic Sciences 62: 679-692.

Bell JD, Kronen M, Vunisea A, Nash WJ, Keeble G, Demmke A, et al. (2009) Planning the use of fish for food security in the Pacific. Marine Policy 33: 64-76.

Bellwood DR, Wainwright PC (2002) The history and biogeography of fishes on coral reefs.

Pages 5-32 In: Sale PF (ed) Coral Reef Fishes. Dynamics and Diversity in a Complex

Ecosystem . Academic Press, San Diego, USA.

Bellwood DR, Hoey AS, Choat JH (2003) Limited functional redundancy in high diversity systems: resilience and ecosystem function on coral reefs. Ecology Letters 6: 281-285.

Bellwood DR, Hughes TP, Folke C, Nystrom N (2004) Confronting the Coral Reef Crisis.

Nature 429: 827-833.

Bellwood DR, Hughes TP, Connolly SR, Tanner J (2005) Environmental and geometric constraints on Indo-Pacific coral reef biodiversity. Ecology Letters 8: 643-651.

Bellwood DR, Hughes TP, Hoey AS (2006) Sleeping functional group drives coral reef recovery. Current Biology 16: 2434-2439.

Bellwood DR, Klanten S, Cowman PF, Pratchett MS, Konow N, et al. (2010) Evolutionary history of the butterflyfishes (f: Chaetodontidae) and the rise of coral feeding fishes.

Journal of Evolutionary Biology 23: 335-349.

Berkelmans R, De’ath G, Kininmonth S, Skirving WJ (2004) A comparison of the 1998 and

2002 coral bleaching events on the Great Barrier Reef: spatial correlation, patterns, and predictions. Coral Reefs 23: 74 –83.

Berumen ML (2005) The importance of juveniles in modeling growth: butterflyfish at Lizard

Island. Environmental Biology of Fishes 72: 409-413.

Berumen ML, Pratchett MS (2006) Recovery without resilience: persistent disturbance and long-term shift in the structure of fish and coral communities at Tiahura Reef, Moorea.

Coral Reefs 25: 647-653.

Booth DJ, Beretta GA (2002) Changes in a fish assemblage after a coral bleaching event.

Marine Ecology Progress Series 245: 205 –212.



Bouchon-Navaro Y, Bouchon C, Harmelin-Vivien ML (1985) Impact of coral degradation on a Chaetodontid fish assemblage (Moorea, French Polynesia). Proceedings of the 5th

International Coral Reef Symposium 5: 427-432.

Bozec YM, Doledec S, Kulbicki M (2005) An analysis of fish-habitat associations on disturbed coral reefs: chaetodontid fishes in New Caledonia. Journal of Fish Biology

66: 966-982

Brooks TM, Mittermeier RA, Mittermeier CG, Da Fonseca GAB, Rylands AB, et al. (2002)

Habitat loss and extinction in the hotspots of biodiversity. Conservation Biology 16:

909 –923.

Brown GW (2001) The influence of habitat disturbance on reptiles in a Box –Ironbark eucalypt forest of south-eastern Australia. Biodiversity and Conservation 10: 161-76.

Brown JH (1984) On the relationship between abundance and distributions. American

Naturalist 124: 255-279.

Cadoret L, Aderjoud M, Tsuchiya M (1999) Spatial distribution of chaetodontid fish in coral reefs of the Ryukyu Islands, southern Japan. Journal of the Marine Biological

Association of the United Kingdom 79: 725-735.

Caley MJ, Carr MH, Hixon MA, Hughes TP, Jones GP, et al. (1996) Recruitment and the local dynamics of open marine populations. Annual Review of Ecology and

Systematics 27: 477-500.

Carlton JT (1993) Neoextinctions in marine invertebrates. American Zoologist 33: 499-507.

Carpenter KE, Pratchett M (2009) Chaetodon trifascialis . In: IUCN 2010. IUCN Red List of

Threatened Species. Version 2010.4. www.iucnredlist.org/apps/redlist/details/165712/0

Carpenter K, Abrar A, Aeby G, Aronson RB, Banks S, et al. (2008) One-third of reef-building corals face elevated extinction risk from climate change and local impacts Science

321: 560-563.

Cheal AJ, Coleman G, Delean S, Miller I, Osborne K, Sweatman H (2002) Responses of coral and fish assemblages to a severe but short-lived tropical cyclone on the Great

Barrier Reef, Australia. Coral Reefs 21: 131 –142.

Cheung WWL, Pitcher TJ, Pauly D (2005) A fuzzy logic expert system to estimate intrinsic extinction vulnerabilities of marine fishes to fishing. Biological Conservation 124: 97-

111.



Choat HJ, Bellwood DR (1991) Reef fishes: their history and evolution. Pages 39-66 In: Sale

PF (ed) The Ecology of Fishes on Coral Reefs . Academic Press, San Diego, USA.

Cornish A, Kiwi LK (2004). Plectropomus leopardus . In: IUCN 2010. IUCN Red List of

Threatened Species . Version 2010.4. www.iucnredlist.org/apps/redlist/details/44684/0

Coker DJ, Pratchett MS, Munday PL (2009) Coral bleaching and habitat degradation increase susceptibility to predation for coral-dwelling fishes. Behavioural Ecology 20:

1204-1210

Cole AJ, Pratchett MS, Jones GP (2008) Diversity and functional importance of coral-feeding fishes on tropical coral reefs. Fish and Fisheries 9: 286-307.

Connell JH, Sousa WP (1983) On the evidence needed to judge ecological stability or persistence. American Naturalist 121: 789-824.

Connell JH, Hughes TP, Wallace CC (1997) A 30-year study of coral abundance, recruitment, and disturbance at several scales in space and time. Ecological

Monographs 67: 461-488.

Cox E (1994) Resource use by corallivorous butterflyfishes (Family Chaetodontidae) in

Hawaii. Bulletin of Marine Science 54: 535-545.

Crosby MP, Reese ES (1996) A Manual for Monitoring Coral Reefs with Indicator Species:

Butterflyfishes as Indicators of Change on Indo Pacific Reefs . Office of Ocean and

Coastal Resource Management, National Oceanic and Atmospheric Administration,

Silver Spring, MD, USA.

Crooks KR, Soule ME (1999) Mesopredator release and avifaunal extinctions in a fragmented system. Nature 400: 563 –566.

Crosby MP, Reese ES (2005) Relationship of habitat stability and intra-specific population dynamics of an obligate corallivore butterflyfish. Aquatic Conservation: Marine and

Freshwater Research 15: S13-S25.

Depczynski M, Bellwood DR (2005) Shortest recorded vertebrate lifespan found in a coral reef fish. Current Biology 15: R288-R289.

Dirzo R, Raven PH (2003) Global state of biodiversity and loss. Annual Review of

Environment and Resources 28: 137-167.

Doherty PJ (1991). Spatial and temporal patterns in the recruitment of a coral reef flsh.

Pages 216-293 In: Sale PF (ed) The Ecology of Fishes on Coral Reefs . Academic

Press, San Diego, USA.



Doherty PJ (1998) Recruitment-limitation: definitions, predictions and tests. Pages 129-131

In: Jones GP, Doherty PJ, Mapstone BD, Howlett L (eds) ReeFish ’95: Recruitment and Population Dynamics of Coral Reef Fishes . CRC Reef Research Centre,

Townsville, Australia.

Donner SD, Skirving WJ, Little CM, Oppenheimer M, Hoegh-Guldberg O (2005) Global assessment of coral bleaching requires rates of adaptation under climate change.

Global Change Biology 11: 2251-2265.

Downing N, Buckley R, Stobart B, Éclair L, Teleki K (2005) Reef fish diversity at Aldabara

Atoll, Seychelles, during the 5 years following the 1998 coral bleaching event.

Philosophical Transactions of the Royal Society of London A 363: 257-261.

Duarte CM (2002) The future of seagrass meadows. Environmental Conservation 29: 192 –

206.

Dulvy NK, Sadovy Y, Reynolds JD (2003) Extinction vulnerability in marine populations. Fish and Fisheries 4: 25-64.

Ehrlich PR, Daily CG (1993) Population extinction and saving biodiversity. Ambio 22: 64-68.

Elliot JK, Loughheed SC, Bateman B, McPheee LK, Boag PT (1999) Molecular phylogenetic evidence for the evolution of specialization in anemonefishes. Proceeding of the Royal

Society B 266: 677-685.

Elmqvist T, Folke C, Nyström M, Peterson G, Bengtssson J, et al. (2003) Response diversity, ecosystem change, and resilience.

Frontiers in Ecology and the Environment

1: 488-494.

Emslie MJ, Pratchett MS, Cheal AJ (2011) Effects of different disturbance types on butterflyfish communities of Australia’s Great Barrier Reef. Coral Reefs 30: 461-471.

Evans RD, Russ GR, Kritzer JP (2008) Batch fecundity of Lutjanus carponotatus

(Lutjanidae) and implications of no-take marine reserves on the Great Barrier Reef,

Australia. Coral Reefs 27: 179-189.

Evans Rd, van Herwerden L, Russ GR, Frisch AJ (2010) Strong genetic but not spatial subdivision of two reef fish species targeted by fishers on the Great Barrier Reef.

Fisheries Research 102: 16-25.

Feary DA, Almany GR, McCormick MI, Jones GP (2007) Habitat choice, recruitment and the response of coral reef fishes to coral degradation. Oecologia 153: 727 –737.



Ferreira BP, Russ GR (1994) Age validation and estimation of growth rate of the coral trout

Plectropomus leopardus (Lacepede 1802) from Lizard Island, Northern Great Barrier

Reef. Fishery Bulletin 92: 46-57.

Frank KT, Petrie B, Choi JS, Leggett WC (2005) Trophic cascades in a formerly coddominated ecosystem. Science 308: 1621-1623.

Frankham R (2005) Genetics and extinction. Biological Conservation 126: 131-140.

Frankham R (2010) Challenges and opportunities of genetic approaches to biological conservation. Biological Conservation 143: 1919-1927.

Fricke HW (1986) Pair swimming and mutual partner guarding in monogamous butterflyfish

(Pisces, Chaetodontidae): a joint advertisement for territory. Ethology 73:307 –333

Gärdenfors U (2001) Classifying threatened species at national versus global levels. Trends in Ecology and Evolution 16: 511-516.

Gardiner NM, Jones GP (2005) Habitat specialisation and overlap in a guild of coral reef cardinal fishes (Apogonidae). Marine Ecology Progress Series 305: 163-175.

Gaston KJ (1994) Rarity . Chapman and Hall, London.

Gaston KJ (1998) Species-range size distributions: products of speciation, extinction and transformation. Philosophical Transactions of the Royal Society B 353: 219-230.

Gilmore RG, Dodrill JW, Linley PA (1983) Reproduction and embryonic development of the sand tiger shark, Odontaspis taurus (Rafinesque). Fishery Bulletin 81: 201-225.

Golbuu Y, Victor S, Penland L, Idip D, Emaurois C, et al. (2007) Palau’s coral reefs show differential habitat recovery following the 1998-bleaching event. Coral Reefs 26: 319-

332.

Graham NAJ, Wilson SK, Jennings S, Polunin NVC, Bijoux JP et al. (2006) Dynamic fragility of oceanic coral reef ecosystems. Proceedings of the National Academy of Sciences,

USA 103: 8425 –8429.

Graham NAJ, Wilson SK, Jennings S, Polunin NVC, Robinson J et al. (2007). Lag effects in the impacts of mass coral bleaching on coral reef fish, fisheries, and ecosystems.

Conservation Biology 21: 1291 –1300.

Graham NAJ, McClanahan TR, MacNeil MA, Wilson SK, Polunin NVC et al. (2008) Climate warming, marine protected areas and the ocean-scale integrity of coral reef ecosystems. PLoS ONE 3(8): e3039.



Graham NAJ, Chabanet P, Evans RD, Jennings S, Letourneur Y, et al. (2011a) Extinction vulnerability of coral reef fishes. Ecology Letters 14: 341-348.

Graham NAJ, Ainsworth TD, Baird AH, Ban NC, Bay LK, et al. (2011b) From microbes to people: Tractable benefits of no-take area for coral reefs. Oceanography and Marine

Biology: An Annual Review 49: 177-148.

Great Barrier Reef Marine Park Authority (2009) Great Barrier Reef Outlook Report 2009 .

Great Barrier Reef Marine Park Authority, Townsville, Australia.

Halford A, Cheal AJ, Ryan D, Williams DM (2004) Resilience to large-scale disturbance in coral and fish assemblages on the Great Barrier Reef. Ecology 85: 1892 –1905.

Hanski IA (1999) Metapopulation ecology . Oxford University Press, Oxford, United Kingdom.

Harmelin-Vivien ML, Bouchon-Navaro Y (1983) Feeding diets and significance of coral feeding among Chaetodontid fishes in Moorea (French Polynesia). Coral Reefs 2: 119-

127

Harmelin-Vivien ML, Laboute P (1986) Catastrophic impact of hurricanes on atoll outer reef slopes in the Tuamotu (French Polynesia). Coral Reefs 5: 55-62.

Hawkins JP, Roberts CM, Clark V (2000). The threatened status of restricted-range coral reef fish species. Animal Conservation 3: 81 –88.

Hilton-Taylor C (2000) IUCN Red List of Threatened Species . World Conservation Union,

Gland Switzerland.

Hoegh-Guldberg O (1999) Climate change, coral bleaching and the future of the world’s coral reefs. Marine and Freshwater Research 50: 839−866.

Hoese DF, Bray DJ, Paxton JR, Allen GR (2006) Fishes. Pages 1-2178 In: Beesley PL ,

Wells A (eds) Zoological Catalogue of Australia. Volume 35 . ABRS and CSIRO

Publishing, Australia.

Holling CS (1973) Resilience and stability of ecological systems. Annual Review of Ecology and Systematics. 4: 1-23.

Houghton JT, Ding Y, Griggs DJ, Noguer M, van der Linden PJ, et al. (2001) IPCC Third

Assessment Report: Climate Change 2001.

Cambridge University Press, Cambridge,

UK.

Hourigan TF, Tricas TC, Reese ES (1988) Coral reef fishes as indicators of environmental stress in coral reefs. Pages 107-135 In: Soule DF, Kleppel GS (eds) Marine organisms as indicators . Springer-Verlag, New York.



Hughes TP (1994) Catastrophes, phase shifts, and large-scale degradation of a Caribbean coral reef. Science 265: 1547-1551.

Hughes TP, Connell JH (1999) Multiple stressors on coral reefs: a long-term perspective.

Limnology and Oceanography 44: 932-940.

Hughes TP, Baird AH, Dinsdale EA, Moltschaniwskyj NA, Pratchett MS, et al. (2000) Supplyside ecology works both ways: the link between benthic adults and larval recruits.

Ecology 81: 2241-2249.

Hughes TP, Bellwood DR, Connolly SR (2002) Biodiversity hotspots, centres of endemicity, and the conservation of coral reefs. Ecology Letters 5: 775-784.

Hughes TP, Baird AH, Bellwood DR, Card M, Connolly SR, et al. (2003) Climate change, human impacts and the resilience of coral reefs. Science 301: 929-933.

Hughes TP, Bellwood DR, Folke C, Steneck RS, Wilson J (2005) New paradigms for supporting the resilience of marine ecosystems. Trends in Ecology and Evolution 20:

380-386.

Hughes TP, Rodrigues MJ, Bellwood DR, Ceccarelli D, Hoegh-Guldberg O, et al. (2007)

Phase shifts, herbivory, and the resilience of coral reefs to climate change. Current

Biology 17: 360 –365.

Hughes TP, Graham NAJ, Jackson JBC, Mumby PJ, Steneck RS (2010) Rising to the challenge of sustaining coral reef resilience. Trends in Ecology and Evolution 25: 633-

642.

Hutchings JA, Reynolds JD (2004) Marine fish population collapse: consequences for recovery and extinction risk. Bioscience 54: 297-309.

Jansen A, Robertson AI (2001) Riparian bird communities in relation to land management practices in floodplain woodlands of south-eastern Australia. Biological Conservation

100: 173-185.

Jones GP, Munday PL, Caley JM (2002) Rarity in coral reef fish communities. Pages 81-101

In: Sale PF (ed) Coral Reef Fishes: Dynamics and Diversity in a Complex Ecosystem .

Academic Press, Sydney, Australia.

Jones GP, McCormick MI, Srinivasan M, Eagle JV (2004) Coral decline threatens fish biodiversity in marine reserves. Proceedings of the National Academy of Sciences,

USA 101: 8251 –8253.

Jones GP, Srinivasan M, Almany GR (2007) Population connectivity and conservation of marine biodiversity. Oceanography 20: 100 –111.



Jones GP, Almany GR, Russ GR, Sale PF, Steneck RS, et al. (2009) Larval retention and connectivity among populations of corals and reef fishes: history, advances and challenges. Coral Reefs 28:307-325.

Julliard R, Jiguet F, Couvet D (2003) Common birds facing global changes: what makes a species at risk? Global Change Biology 10: 148 –154.

Keller LF, Waller DM (2002) Inbreeding effects in wild populations. Trends in Ecology and

Evolution 17: 230-241.

Khalaf MA, Crosby MP (2005) Assemblage structure of butterflyfishes and their use as indicators of Gulf of Aqaba benthic habitat in Jordan. Aquatic Conservation: Marine and Freshwater Ecosystem s 15: S27 –S43.

Kingsford MJ (2009) Contrasting patterns of reef utilization and recruitment of coral trout

( Plectropomus leopardus ) and snapper ( Lutjanus carponotatus ) at One Tree Island, southern Great Barrier Reef. Coral Reefs 28: 251-264.

Kokita T, Nakazono A (2001) Rapid response of an obligately corallivorous filefish

Oxymonacanthus longirostris (Monacanthidae) to a mass coral bleaching event. Coral

Reefs 20: 155-158.

Kritzer JP, Sale PF (2004) Metap opulation ecology in the sea: from Levins’ model to marine ecology and fisheries science. Fish and Fisheries 5: 131-140.

Kritzer JP, Sale PF (2006) The metapopulation ecology of coral reef fishes. Chapter 2 In:

Kritzer JP, Sale PF (eds) Marine Metapopulations . Elsevier Academic Preee

Publications, USA.

Lawton JH (1999) Are there general laws in ecology? Oikos 84: 177-192.

Lewis AR (1997) Recruitment and post-recruitment immigration affect the local populations size of coral reef fishes. Coral Reefs 16: 139-149.

Loya Y, Sakai K, Yamazota K, Nakano Y, Sambali H, et al. (2001) Coral bleaching: the winners and the losers. Ecology Letters 4: 122-131.

Ludwig D, Hilborn R, Walters C (1993) Uncertainty, resource exploitation, and conservation: lessons from history. Science 260: 17-36.

MacArthur RH, Wilson EO (1967) The Theory of Island Biogeography . Princeton University

Press, Princeton, NJ, USA.

Mapstone BD, McKinlay JP, Davies CR (1996) A description of the commercial reef line fishery logbook data held by the Queensland Fisheries Management Authority.

Report


Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species to the QFMA from the CRC Reef Research Centre and the Department of Tropical

Environment Studies and Geography, James Cook University, Townsville, Australia.

Marshall PA, Baird AH (2000) Bleaching of corals on the Great Barrier Reef: differential susceptibilities among taxa. Coral Reefs 19:155-163.

McKinney ML (1997) Extinction vulnerability and selectivity: combining ecological and paleontological views. Annual Review of Ecology and Systematics 28: 495-516.

Morris AV, Roberts CM, Hawkins JP (2000) The threatened status of groupers

(Epinephelinae). Biodiversity and Conservation 9: 919-942.

Munday PL (2002) Does habitat availability determine the geographical-scale abundance of coral-dwelling fishes? Coral Reefs 21: 105-116.

Munday PL (2004) Habitat loss, resource specialisation, and extinction on coral reefs. Global

Change Biology 10:1642-1647.

Munday PL, Jones GP, Caley MJ (1997) Habitat specialisation and the distribution and abundance of coral-dwelling gobies. Marine Ecology Progress Series 152: 227 –239.

Munday PL (2001) Fitness consequences of habitat selection and competition among coraldwelling fish. Oecologia 128: 585 –593.

Musick JA (1999) Criteria to define extinction risk in marine fishes. Fisheries 24: 6-14.

Myers RA, Worm B (2005) Extinction, survival or recovery of large predatory fishes.

Philosophical Transactions of the Royal Society B 360: 13-20.

Naeem S, Thompson LJ, Lawler SP, Lawton JH, Woodfin RM (1994) Declining biodiversity can alter the performance of ecosystems. Nature 368: 734-737.

Newton K, Cote IM, Pilling GM, Jennings S, Dulvy NK (2007) Current and future sustainability of island coral reef fisheries. Current Biology, 17: 655-658.

Öhman MC, Munday PL, Jones GP, Caley MJ (1998) Settlement strategies and distribution patterns of coral-reef fishes. Journal of Experimental Marine Biology and Ecology 225:

219-238.

Osborne K, Dolman AM, Burgess SC, Johns KA (2011) Disturbance and the dynamics of coral cover on the Great Barrier Reef (1995-2009) PloS ONE 6: e17516.

Owens IPF, Bennett PM (2000) Ecological basis of extinction risk in birds: Habitat loss versus human persecution and introduced predators. Proceedings of the National

Academy of Sciences, USA 97: 12144-12148.



Pace ML, Cole JJ, Carpenter SR, Kitchell JF (1999) Trophic cascades revealed in diverse ecosystems. Trends in Ecology and Evolution 14: 483-488.

Palumbi SR (2003) Population genetics, demographic connectivity, and the design of marine reserves. Ecological Applications 13: S146-S158.

Pandolfi JM, Bradbury RH, Sala E, Hughes TP, Bjorndal KA, et al. (2003) Global trajectories of the long-term decline of coral reef ecosystems. Science 301: 955-958.

Pannell JR (2003) Coalescence in a metapopulation with recurrent local extinction and recolonization. Evolution 57: 949-961.

Pauly D, Christensen V, Dalsgaard J, Froese R, Torres F (1998) Fishing down marine food webs. Science 279: 860-863.

Pauly D, Christensen V, Guenette S, Pitcher TJ, Sumaila R, et al. (2002) Towards sustainability in world fisheries. Nature 418: 689 –695.

Pears RJ (2005) Comparative demography and assemblage structure of serranid fishes: implications for conservation and fisheries management. PhD thesis, James Cook

University, Townsville

Pears RJ, Choat JH, Mapstone BD, Begg GA (2006) Demography of a large grouper,

Epinephelus fuscoguttatus , from Australia’s Great Barrier Reef: implications for fishery management. Marine Ecology Progress Series 307: 259-272.

Pickett STA, White PS (1986) The Ecology of Natural Disturbance and Patch Dynamics .

Academic Press, Ontario.

Pimm SL, Moulton MP, Justice LJ (1994) Bird extinctions in the Central Pacific.

Philosophical Transactions of the Royal Society, London, Series B 344: 27 –33.

Pogonoski JJ, Pollard DA, Paxton JR (2002) Conservation Overview and Action Plan for

Australian Threatened and Potentially Threatened Marine and Estuarine Fishes .

Environment Australia, Canberra ACT, Australia.

Pollard D, Smith A (2005) Carcharias taurus . In: IUCN 2010. IUCN Red List of Threatened

Species. Version 2010.4. www.iucnredlist.org/apps/redlist/details/3854/0

Poloczanska ES, Babcock RC, Butler A, Hobday AJ, Hoegh-Guldberg O, et al. (2007)

Climate change and Australian marine life. Oceanography and Marine Biology: An

Annual Review 45: 409-480.

Pratchett MS (2005) Dietary overlap among coral-feeding butterflyfishes (Chaetodontidae) at

Lizard Island, northern Great Barrier Reef. Marine Biology 148, 373-382.



Pratchett MS (2007) Dietary selection by coral-feeding butterflyfishes (Chaetodontidae) on the Great Barrier Reef, Australia.

Raffles Bulletin of Zoology 14: S161-S166.

Pratchett MS, Berumen ML (2008) Inter-specific variation in distributions and diets of coral reef butterflyfishes (Teleostei: Chaetodontidae). Journal of Fish Biology 73: 1730-

1747.

Pratchett MS, Wilson SK, Berumen ML, and McCormick MI (2004) Sub-lethal effects of coral bleaching on an obligate coral feeding butterflyfish. Coral Reefs 23: 352-356.

Pratchett MS, Pradjakusuma OA, Jones GP (2006a) Is there a reproductive basis to solitary versus pair-formation in coral reef fishes? Coral Reefs 25: 85-92.

Pratchett MS, Wilson SK, Baird AH (2006b) Declines in the abundance of Chaetodon butterflyfishes (Chaetodontidae) following extensive coral depletion. Journal of Fish

Biology 69: 1269-1280.

Pratchett MS, Munday PL, Wilson SK, Graham NAJ, Cinner JE, et al. (2008a) Effects of climate-induced coral bleaching on coral-reef fishes: ecological and economic consequences. Oceanography and Marine Biology: An Annual Review 46: 251-296.

Pratchett MS, Marnane MJ, Berumen ML, Eagle JE, Pratchett DJ (2008b) Habitat associations of juvenile versus adult butterflyfishes. Coral Reefs 27: 541-551.

Pratchett MS, Wilson SK, Graham NAJ, Munday PL., Jones GP, et al. (2009a) Multi-scale temporal effects of climate-induced coral bleaching on motile reef organisms. Pages

139-158 In: van Oppen M, Loug Jh (ed). Coral Bleaching: Patterns and Processes,

Causes and Consequences . Springer-Verlag, Heidelberg.

Pratchett MS, Baird AH, McCowan DM, Coker DJ, Cole AJ, et al. (2009b) Protracted declines in coral cover and fish abundance following climate-induced coral bleaching on the Great Barrier Reef. Proceedings of the 11th International Coral Reef

Symposium 1: 1042-1046

Pratchett MS, Munday PL, Graham NAJ, Kronen M, Pinica S, et al. (2011a) Coastal

Fisheries. Chapter 9. In: Bell J, Johnson J, Hobday AH (eds) Climate change and the vulnerability of Pacific fisheries . Secretariat for the Pacific Community, Noumea.

Pratchett MS, Bay LK, Gehrke PC, Koehn J, Osborne K, et al. (2011b) Contribution of climate change to degradation and loss of critical fish habitats in Australian aquatic environments. Marine and Freshwater Research In press



Pyron M (2001) Relationships between geographical range size, body size, local abundance, and habitat breadth in North American suckers and sunfishes. Journal of

Biogeography 26: 549-558.

Randall JE (1974) The effect of fishes on coral reefs. Proceedings of the 2nd International

Coral Reef Symposium 1: 159-166.

Randall JE (2005) Reef and Shore Fishes of the South Pacific: New Caledonia to Tahiti and the Pitcairn Islands .

University of Hawai’i Press, Honolulu, USA.

Reaka-Kudla ML (1997) The global biodiversity of coral reefs: a comparison with rain forests. Pages 83-108 In: Reaka-Kudla ML, Wilson DE, Wilson EO (eds) Biodiversity

II: Understanding and Protecting our Natural Resource. Henry/National Academy

Press, Washington DC, USA.

Reese ES (1977) Coevolution of corals and coral feeding fishes of the family

Chaetodontidae. Proceedings of the 3rd International Coral Reef Symposium 1: 267-

274.

Riegl B, Piller WE (2003) Possible refugia for reefs in times of environmental stress.

International Journal of Earth Science 92: 520-531.

Roberts CM, Hawkins JP (1999) Extinction risk in the sea. Trends in Ecology and Evolution

14: 241-246

Roberts CM, McClean CJ, Veron JEN, Hawkins JP, Allen GR, et al. (2002) Marine biodiversity hotspots and conservation priorities for tropical reefs. Science 295: 1280-

1284.

Rocha LA, Pyle R, Craig MT, Pratchett M (2009) Chaetodon lunulatus . In: IUCN 2010. IUCN

Red List of Threatened Species. Version 2010.4. www.iucnredlist.org/apps/redlist/details/165704/0/classify

Russ GR, Cheal AJ, Dolman AM, Emslie MJ, Evans RD, et al. (2008) Rapid increase in fish numbers follows creation of world's largest marine reserve network. Current Biology

18: R514-R515.

Sadovy Y, Cheung WL (2003). Near extinction of a highly fecund fish: the one that nearly got away. Fish and Fisheries 4: 86-99.

Sadovy Y, Colin PL (1995) Sexual development and sexuality in the Nassau groper. Journal of Fish Biology 46: 961-976.



Sadovy YJ, Vincent ACJ (2002) Ecological issues and the trades in live reef fishes. Pages

391-420. In: Sale PF (ed) Coral Reef Fishes. Dynamics and Diversity in a Complex

Ecosystem . Academic Press, San Diego, USA.

Samways MJ (2005) Breakdown of butterflyfish (chaetodontidae) territories associated with the onset of a mass coral bleaching event. Aquatic Conservation: Marine and

Freshwater Ecosystems 15: S101 –S107.

Samoilys MA (1997) Periodicity of spawning aggregations of coral trout, Plectropomus leopardus (Pisces: Serranidae) on the northern Great Barrier Reef. Marine Ecology

Progress Series 160: 149-159.

Sano M (1989) Feeding habits of Japanese butterflyfishes (Chaetodontidae). Environmental

Biology of Fishes 25: 195 –203.

Schmitz OJ, Hamback PA, Beckerman AP (2000) Trophic cascades in terrestrial systems: a review of the effects of carnivore removals on plants. The American Naturalist 155:

141-153.

Sheppard CRC (2003) Predicted recurrences of mass coral mortality in the Indian Ocean.

Nature 425: 294-297.

Smith SV, Buddemeier RW (1992) Global change and coral reef ecosystems. Annual

Reviews of Ecology and Systematics 23: 89-118.

Spielman D, Brook BW, Frankham R (2004) Most species are not driven to extinction before genetic factors impact them. Proceedings of the National Academy of Sciences, USA

101: 15261-15264.

State of the Environment Advisory Council (1996) Australia State of the Environment 1996 .

An independent report presented to the Commonwealth Minister for the Environment.

CSIRO, Collingwood, Australia.

St John J, Russ GR, Brown IW, Squire LC (2001) The diet of the large coral reef serranid

Plectropomus leopardus in two fishing zones on the Great Barrier Reef, Australia.

FisherY Bulletin 99: 180-192.

Steele JH (1991) Marine functional diversity. Bioscience 41: 470-474.

Steneck RS, Graham MH, Bourque BJ, Corbett D, Erlandson JM, et al. (2002) Kelp forest ecosystems: biodiversity, stability, resilience and future. Environmental Conservation

29: 436-459.



Swain DP, Wade EJ (1993) Density-dependent geographic distribution of Atlantic cod

( Gadus morhua ) in the southern Gulf of St. Lawrence. Canadian Journal of Fisheries and Aquatic Sciences 50: 725-733.

Sweatman H, Burgess S, Cheal A, Coleman G, Delean S, et al. (2005). Long-term monitoring of the Great Barrier Reef. Status Report Number 7, 2005 . Australian

Institute of Marine Science, Townsville, Australia.

Sweatman H, Cheal A, Coleman G, Emslie M, John K et al. (2008) Long-term monitoring of the Great Barrier Reef: Status Report Number 8, 200 8. Australian Institute of Marine

Science, Townsville, Australia.

Sweatman HAP, Delean S, Syms C (2011) Assessing loss of coral cover on Australia’s

Great Barrier Reef over two decades, with implications for longer-term trends. Coral

Reefs 30: 521-531.

Thomas CD, Cameron A, Green RE, Bakkenes M, Beaumont LJ, et al. (2004) Extinction risk from climate change. Nature 427: 145-147.

Thresher RE (1985) Distribution, abundance, and reproductive success in the coral reef fish

Acanthochromis polyacanthus . Ecology 66: 1139-1150.

Tilman D, Downing JA (1994) Biodiversity and stability in grasslands. Nature 367: 363-365.

Travis JM (2003) Climate change and habitat destruction: a deadly anthropogenic cocktail.

Proceedings of the Royal Society B: Biological Sciences 270: 467-473.

Unmack PJ (2001) Biogeography of Australian freshwater fishes. Journal of Biogeography

28: 1053-1089. van Herwerden L, Choat JH, Newman SJ, Leray M, Hollersoy G (2009) Complex patterns of population structure and recruitment of Plectropomus leopardus (Pisces:

Epinephelidae) in the Indo-West PaciWc: implications for Wsheries management.

Marine Biology 156: 1595-1607.

Veron JEN, Hoegh-Guldberg O, Lenton TM, Lough JM et al. (2009) The coral reef crisis:

The critical importance of <350 ppm CO

2

. Marine Pollution Bulletin 58: 1428-1436.

Walther G-R, Post E, Convey P, Menzel A, Parmesan C, et al. (2002) Ecological responses to recent climate change. Nature 416: 389-395.

Warner RR (1982) Mating Systems, sex Change and sexual Demography in the rainbow wrasse, Thalassoma lucasanum . Copeia 3: 653-661.



Waycott M, Duarte CM, Carruthers TJB, Orth RJ, Dennison WC, et al (2009) Accelerating loss of seagrasses across the globe threatens coastal ecosystems. Proceedings of the

National Academy of Science, USA 106: 12377-12381.

Wilkinson CR (2004) Status of Coral Reefs of the World: 2004 Australian Institute of Marine

Science, Townsville, Australia.

Williams DM (1986) Temporal variation in the structure of reef slope fish communities

(central Great Barrier Reef): short term effects of Acanthaster planci infestation.

Marine Ecology Progress Series 28:157-164.

Williams YM, Williams SE, Alford RA, Waycott M, Johnson CN (2006) Niche breadth and geographical range: ecological compensation for geographical rarity in rainforest frogs.

Biology Letters 2: 532-535.

Williamson DH, Russ GR, Ayling AM (2004) No-take marine reserves increase abundance and biomass of reef fish on inshore fringing reefs of the Great Barrier Reef.

Environmental Conservation 3: 149-159.

Wilson SK, Graham NAJ, Pratchett MS, Jones GP, Polunin NVC (2006) Multiple disturbances and the global degradation of coral reefs: are reef fishes at risk or resilient? Global Change Biology 12: 2220-2234.

Wilson SK, Fisher R, Pratchett MS, Graham NAJ, Dulvy NK, et al. (2008) Exploitation and habitat degradation as agents of change within coral reef fish communities. Global

Change Biology 14: 2796-2809.

Wild C, Huettel M, Klueter A, Kremb SG, Rasheed MYM, et al. (2004) Coral mucus functions as an energy carrier and particle trap in the reef ecosystem. Nature 428: 66-70.

Wooldridge SA, Done TJ (2009) Improved water quality can ameliorate effects of climate change on corals. Ecological Applications 19: 1492-1499.

Worm B, Barbie EB, Beaumont N, Duffy JE, Folke C, et al. (2006) Impacts of biodiversity loss on ocean ecosystem services. Science 314: 787-790.

Worm B, Hilborne R, Baum JK, Branch TA, Collie JS, et al. (2009) Rebuilding coastal fisheries. Science 325: 578-585.



Appendix I. IUCN Red List categories and criteria, following IUCN (2000).

Critically Endangered

(CR)

Endangered (EN)

Vulnerable (VU)

Population Decline Population size Geographic extent Area of occupancy Other

≥80% population decline over the last

10 years or three generations, or >90% population decline if cause of decline are known and manageable





Population size <50 mature individuals , or <250 individuals if expected to decline

>25% within 3 years, or highly vulnerable population structure

Population size <250 mature individuals , or <2500 individuals if expected to decline


Population size

<1,000 mature individuals , or

<10,000 individuals if expected to decline


Extent of occurrence estimated to be

≤100km 2 , and highly fragmented, declining, or fluctuating.


≤5,000km 2 , and highly fragmented, declining, or fluctuating.



Area of occupancy estimated to be






Probability of extinction is ≥50% within 10 years or three generations.



Australia ■ State of the Environment 2011

Supplementary information 78

Word - 8934 KB - Department of the Environment

Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species

Executive Summary

1. Background

2. Assessing vulnerability and status of marine fishes

3. Specific species assessments

4. Conclusions

References

Related documents

Products

Support

Word - 8934 KB - Department of the Environment

Vulnerability and status of marine fishes for the Australian State of the Environment report 2011 – tropical species

Executive Summary

1. Background

2. Assessing vulnerability and status of marine fishes

3. Specific species assessments

4. Conclusions

References

Related documents

Add this document to collection(s)

Add this document to saved

Suggest us how to improve StudyLib