2014 MARINE ECOLOGICAL SURVEY OF ELIZABETH AND MIDDLETON REEFS, LORD HOWE COMMONWEALTH

Andrew S. Hoey*, Morgan S. Pratchett, Jacob Johansen, and Jessica Hoey

FINAL REPORT – June 1st, 2014

Produced for The Department of the Environment

*Corresponding author – Dr Andrew Hoey, ARC Centre of Excellence for Reef Studies, James Cook University, Townsville QLD 4811. E-mail: [email protected], Telephone/ Fax: (07) 47815979 / 47816722

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In responding to a tender from the Department of Environment, a team of researchers representing the ARC Centre of Excellence for Coral Reef Studies at James Cook University (JCU) completed surveys of the coral reef fauna at Elizabeth and Middleton Reef. The field team comprised Professor Morgan Pratchett, Dr Andrew Hoey, Dr Jacob Johansen and Mrs Jessica Hoey. All the abovementioned researchers contributed to the preparation of this report.

© Commonwealth of 2014 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 inquiries concerning reproduction and rights should be addressed to the Commonwealth Copyright Administration, Attorney General’s Department, Robert Garran Offices, National Circuit, Barton ACT 2600 or posted at http://www.ag.gov.au/cca

The views and opinions expressed in this publication are those of the authors and do not necessarily reflect those of the Australian Government.

This report has been produced for the sole use of the party who requested it. The application or use of this report and of any data or information (including results of experiments, conclusions, and recommendations) contained within it shall be at the sole risk and responsibility of that party. JCU does not provide any warranty or assurance as to the accuracy or suitability of the whole or any part of the report, for any particular purpose or application.

Address all correspondence regarding this report to Dr Andrew Hoey. E-mail: [email protected]

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1 Executive Summary

Elizabeth and Middleton Reefs are of considerable ecological importance. They are the southernmost open ocean platform reefs in the world, and support a unique mix of tropical and subtropical , including several endemic species. Due to their isolation they also provide a refuge for the ‘vulnerable’ black cod. The purpose of this study was to undertake rigorous surveys of shallow coral reef habitats at Elizabeth and Middleton Reefs to assess the current condition of these habitats and detect any changes in the period since the last survey, in March 2011.

Field-based surveys of habitat structure (percentage cover of hard and soft , sponges and algae, topographic complexity), coral replenishment (juvenile corals), coral health (the incidence of coral disease and coral predation), functionally important reef fishes (e.g., surgeonfishes, parrotfishes, drummers), endemic fishes, and invertebrates (holothurians, urchins and the corallivorous crown-of-thorns starfish) were undertaken at Elizabeth and Middleton Reefs from 4th to 9th March, 2014. Two depth zones (shallow reef crest and deep reef slope) were sampled at each of 16 sites, encompassing exposed reef fronts, back reefs and lagoonal habitats. Apex predators, including Galapagos sharks and Black cod, were also sampled at each site, using replicate 250 - 500m transects.

Main findings include:

• Mean cover of hard (scleractinian) corals (Elizabeth Reef: 29.2%; Middleton Reef: 19.3%) was slightly lower than 2011 levels, notably within lagoonal habitats.

• Coral cover at both Elizabeth and Middleton Reefs is lower than has been recorded on the southern Great Barrier Reef and , suggesting that coral assemblages are recovering from previous disturbances. The protracted recovery appears to be related to low rates of coral replenishment and coral growth.

• The health of existing corals was generally high, with a low incidence of and coral disease, and very low densities of coral predators, such as the crown-of-thorns starfish. Consequently these were not likely to be having a

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major influence on coral assemblages, and seemed unlikely to be the cause of observed reductions in coral cover.

• Cover of macroalgae (conspicuous seaweeds >5 mm in height) was generally low (ca. 10%) and displayed limited change to 2011 levels. The exception to this was the back reef on Middleton Reef where macroalgae doubled from 2011 and currently covers over 2/3 of the shallow habitats at these sites.

• The density and biomass of functionally important herbivorous fishes, namely parrotfishes, surgeonfishes and drummers, was generally stable between 2011 and 2014. The only exception was a significant reduction in the abundance of juvenile parrotfish within the lagoons, and appears to be related to a reduction of branching corals in these habitats.

• Densities of endemic fishes ( mccullochi, tricinctus, and Coris bulbifrons) varied greatly among study sites, but were generally greater than or equal to densities recorded in 2011. The densities of C. tricinctus and C. bulbifrons increased on both reefs from 2011 to 2014, while densities of A. mccullochi increased on Middleton and remained stable on Elizabeth Reef.

• Densities of the Galapagos reef shark were extremely high in 2014, and were approximately double those recorded in 2011 and up to 36 times higher than recorded in 2006. All observed were small immature individuals suggesting that the shallow water environments of Elizabeth and Middleton Reefs are an important nursery habitat for this species.

• Densities of black cod, listed as vulnerable in NSW waters and under the Environment Protection and Biodiversity Conservation Act 1999, remained relatively unchanged between 2011 and 2014.

• Elizabeth and Middleton Reefs appear to support healthy fish assemblages and coral assemblages in a state of recovery. The key threats to these reefs are likely to be i) increasing seawater temperatures, ii) physical damage from severe storms, and iii) outbreaks of crown-of-thorns starfish. The impact of these threats is likely to be magnified, relative to low latitude reefs, by the limited capacity for recovery (slow growth and limited replenishment of corals) following such disturbances.

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2 Recommendations

1. Given the isolation and apparent constraints on population and community resilience of reef corals and associated organisms at Elizabeth and Middleton Reefs, anthropogenic disturbances should be minimized through the maintenance of current protection and regulation.

2. Ongoing and regular (at least every 2-3 years) monitoring is critical to assess effects of historical (outbreaks of Acanthaster planci) and emerging threats (e.g., climate change) to corals, and to monitor population fluctuation in key fish species, such as the Galapagos shark and Black cod. This monitoring is a key requirement for effective long-term management of these unique assemblages.

3. In addition to monitoring the population status of key benthic and fish taxa (i.e., corals, macroalgae, herbivorous fish, apex predators), process-orientated monitoring of growth, mortality, and recruitment of these groups would greatly improve our understanding of the dynamics and resilience of these unique coral reef assemblages. This would however, require increased frequency and duration of survey trips. In particular, there is a clear need to understand the determinants of macroalgal abundance and biomass, and the rates of recruitment, growth and mortality of coral assemblages.

4. The current survey highlighted the benefits of matching previous sampling methods and sites, facilitating rigorous temporal comparisons of fish and benthic communities for the first time. Future sampling should continue to use sampling methods that are compatible (if not identical) with the past two surveys (2011 and 2014) and, as a minimum, continue to survey the same sites to maximise detection of long-term trends. Implementation of comparable sampling methods across all Commonwealth Marine Reserves would also enable meaningful spatial and temporal comparisons.

5. A lack of access to data from previous surveys limited our capacity to evaluate long-term trends at these reefs. To facilitate spatial and temporal comparisons, all data arising from these surveys should be made publicly available in a centralised and accessible database, or within an appropriate existing

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database. At the very minimum, all reports relevant to each of the offshore Commonwealth reserves should be digitised and made publically available by the Department of the Environment.

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Table of Contents

1. Executive Summary 3 2. Recommendations 5 3. Introduction 12 3.1 Elizabeth and Middleton Reefs 14 (Lord Howe Commonwealth Marine Reserve) 3.2 Objectives and scope 17 4. Methods 18 4.1 Habitat structure 18 4.2 Reef fishes 22 4.3 Unitary invertebrates (e.g., sea cucumbers) 23 4.4 Species richness 23 5. Findings 24 5.1 Habitat structure 24 5.1.1 Coral cover 24 5.1.2 Macroalgal cover 26 5.1.3 Temporal changes in coral and macroalgae 26 5.1.4 Topographic complexity 31 5.2 Replenishment and resilience of coral populations 32 5.2.1 Juvenile corals 32 5.1.2 Coral predators 39 5.1.3 Coral condition 40 5.3 Coral reef fishes 43 5.3.1 Roving herbivorous fishes 43 5.3.2 Endemic fishes 53 5.3.3 Apex predators 59 5.3.4 New species records 62 5.4 Unitary invertebrates 63 5.4.1 Sea cucumbers (Holothurians) 63 5.4.2 Sea urchins (Echinoids) 65 6. Conclusions 69

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6.1 Habitat structure 69 6.2 Reef fishes 73 6.3 Unitary invertebrates 77 7. Acknowledgements 80 8. References 81

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List of Figures

Figure 1: Map showing the location of the study sites. (a) Geographic location of 19 Elizabeth and Middleton reefs off Australia’s east coast. (b) Locations of the nine study sites on Middleton Reef. (c) Location of the seven study sites on Elizabeth Reef. Note that site 6 on Elizabeth Reef was separated into shallow (6s) and deep (6d) due to the lack of shallow and deep habitat in close proximity. (d) Schematic diagram of reef cross-section showing the location of the three habitats and two depths sampled. Figure 2. Spatial variation in the cover of (a) scleractinian (hard) corals and (b) 25 macroalgae among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Figure 3. Temporal changes (2011 versus 2014) in average cover (±SE) of (a) 27 scleractinian (hard) corals, and (b) macroalgae across three different habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Figure 4. Temporal changes in mean coral cover (±SE) at selected sites from 30 Middleton Reef. Data for 1994 are from Choat et al. (unpublished data), and 2006 from Choat et al. (2006) Figure 5. Temporal changes (2011 versus 2014) in the topographic complexity 31 (±SE) of reef substratum across three different habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Topographic complexity was assessed visually using a 6-point scale (following Wilson et al. 2007). 0 = flat homogenous habitat, 5 = exceptionally complex Figure 6. Composition of juvenile coral assemblages surveyed on Elizabeth and 33 Middleton Reefs in March 2014. Data are total number of juvenile corals recorded in each across 124 transects. Figure 7. Spatial variation in the density of juvenile scleractinian (hard) corals 34 among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE. Figure 8. Temporal changes (2011 versus 2014) in the densities of juvenile 36 scleractinian corals (< 50mm diameter) across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE. Figure 9. Relationships between the density of juvenile corals and the cover of 38 (a) live scleractinian corals, (b) macroalgae, and (c) bare space (a proxy for the availability of potential settlement sites for coral larvae). Each point represents the mean of 4 transects at each depth at each of 16 sites. triangles: Elizabeth Reef, circles: Middleton Reef. Colours represent the three habitats shown in Figure 1. Yellow: exposed reef front; Blue: back reef; Green: lagoon. Figure 10. Spatial variation in the density of corals displaying physical damage 42 consistent with fish feeding (i.e., cropped) among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE. Figure 11. Spatial variation in (a) the abundance, and (b) the biomass of roving 45 herbivorous fishes among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE. The dashed lines represent overall reef means.

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Figure 12. Temporal variation (2011 versus 2014) in the mean abundance of 48 roving herbivorous fishes across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. (a) Total herbivorous fish biomass, (b) Grazing fish biomass, (c) Browsing fish biomass. Error bars represent 1 SE. Figure 13. Temporal variation (2011 versus 2014) in the mean biomass of roving 49 herbivorous fishes across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. (a) Total herbivorous fish biomass, (b) Grazing fish biomass, (c) Browsing fish biomass. Error bars represent 1 SE. Figure 14. Principal component analysis (PCA) showing variation in the 51 composition of roving herbivore assemblages between years, habitats and depths. Symbols indicate sampling year; triangles - 2011, circles - 2014. Colours indicate habitat; dark green – lagoon slope, light green – lagoon crest, dark blue – back reef slope, light blue – back reef crest, bright yellow – exposed reef crest, dull yellow – exposed reef slope. Analysis was based on the covariance matrix of the mean biomass (log-transformed) of each species at each site. Figure 15. Relationships between the herbivore biomass and the cover of (a) 52 macroalgae, (b) live coral, and (c) bare space (the sum of algal turfs and CCA). Each point represents the mean of 4 transects at each depth at each of 16 sites. triangles: Elizabeth Reef, circles: Middleton Reef. Colours represent the three habitats shown in Figure 1. Yellow: exposed reef front; Blue: back reef; Green: lagoon. Figure 16. Relative abundances of members of the (a) 54 (damselfishes), and (b) Chaetodontidae () recorded across all transects on Elizabeth and Middleton Reefs in March 2014. Arrows indicate the relative abundances of the endemic species Amphiprion mccullochi and Chaetodon tricinctus. Data are means ± SE. Figure 17. Spatial variation in the abundance of (a) Amphiprion mccullochi, and 55 (b) Chaetodon tricinctus among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE. Figure 18. Temporal variation (2011 versus 2014) in the mean abundance of (a) 56 Amphiprion mccullochi, and (b) Chaetodon tricinctus across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE. Figure 19. Spatial variation in the abundance of (a) juvenile, and (b) adult Coris 57 bulbifrons among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE. Figure 20. Temporal variation (2011 versus 2014) in the mean abundance of (a) 58 juvenile, and (b) adult Coris bulbifrons across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE. Figure 21. Temporal variation (2011 versus 2014) in the mean abundance of (a) 60 the Galapagos Reef Shark and (b) the Black Cod across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE. Figure 22. Comparison of the size structure of Galapagos Shark (Carcharhinus 61 galapagensis) populations between Elizabeth and Middleton Reefs. Size at birth is 60-80cm, size at maturity is 210-230cm (males) and 250cm (females).

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Figure 23. Comparison of the species composition of sea cucumbers populations 64 between 2011 and 2014 on Elizabeth and Middleton Reefs. The three most abundant species accounted for approximately 85% of all individuals in both sampling years. Figure 24. Temporal variation (2011 versus 2014) in the mean abundance of sea 64 cucumbers (all species combined) across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE. Figure 25. Comparison of the species composition of sea urchin populations 66 between 2011 and 2014 on Elizabeth and Middleton Reefs. Figure 26. Temporal variation (2011 versus 2014) in the mean abundance of the 67 three most abundant sea urchins (a) Echinometra mathaei (b) , and (c) Echinostrephus sp across three habitat types at Elizabeth and Middleton Reefs. Error bars represent 1 SE. Figure 27. Geographic variation in mean (± SE) coral cover between Lord Howe 71 Island, Elizabeth Reef, Middleton Reef, and the southern GBR. Data for Lord Howe Island is for 2010 (Hoey et al. 2011), and data for the southern GBR is for 2005 (Emslie et al. 2008). Figure 28. Geographic variation in mean (± SE) abundance of juvenile corals 73 between three high latitude reefs; Lord Howe Island, Elizabeth Reef, Middleton Reef, and two low latitude reefs; the southern GBR and Moorea, French Polynesia. Data sources are Lord Howe Island (Hoey et al. 2011), southern GBR (Trapon et al. 2013a), and Moorea (Pratchett et al 2011a). Figure 29. Geographic variation in mean (± SE) biomass of roving herbivorous 76 fishes between Lord Howe Island, Elizabeth Reef, Middleton Reef, and the central and northern GBR. The relative contribution of grazing and browsing fishes is shown. Data sources are Lord Howe Island (Hoey et al. 2011), central and northern GBR (Wismer et al. 2009).

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3 Introduction

Corals reefs are critically important, but highly vulnerable ecosystems. Not only do coral reefs support very high biodiversity (e.g., Connell 1978), but they also yield significant goods and services to humankind, estimated to worth in excess of US$30 billion annually (Cesar et al. 2003). Globally, it is estimated that >1 billion people are directly reliant on coral reef fisheries for their food and/ or livelihood. Reef-related fisheries constitute approximately 9–12% of the world’s total fisheries (Smith 1978) and are critically important in sustaining recreational fisheries in many countries. The recreational value of reefs, including income from tourism is enormous (Moberg and Folke 1999). In Australia, the value-added economic contribution of the Great Barrier Reef World Heritage Area to the Australian economy in 2011-12 (mostly through Tourism) was $5.68 billion and it generated almost 69,000 full-time equivalent jobs (Deloitte Access Economics 2013).

Despite their ecological, social and economic importance, coral reefs are being rapidly degraded throughout the world (e.g., Wilkinson 2008). Wilkinson (2008) estimated that 19% of coral reefs have been essentially lost (whereby coral cover has declined by >90% and there is limited prospect of recovery), and unless there are drastic changes in the effectiveness of coral reef management, 60% of reefs are expected to be lost by 2050. The cause(s) of coral reef degradation vary greatly among geographic locations (e.g., Pandolfi et al. 2003). In east Africa, south-east Asia, and the central and southern Caribbean, a disproportionate number of coral reefs have been lost due high levels of sustained human activities, including chronic pollution, eutrophication, sedimentation, overfishing and/or destructive fishing practices (Pandolfi et al. 2003; Wilkinson 2008). Even in the absence of long-term effects of anthropogenic disturbances (e.g., at isolated reefs) reefs are being increasingly affected by the impacts of global climate change (Hughes et al. 2003; Hoegh-Guldberg et al. 2007). It is possible that isolated reefs are even more susceptible to climatic disturbances compared to coastal reef environments because there is limited opportunity for larval replenishment from nearby reefs (Graham et al. 2006; Smith et al. 2008, Gilmour et al. 2013). Many isolated reefs are highly reliant on self-recruitment (e.g., Underwood et al. 2009, Gilmour et al. 2013). The question is, to what extent does the protection of these

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reefs from anthropogenic disturbances offset limitations imposed by isolation (Hughes et al. 2003; Hughes et al. 2010, Gilmour et al. 2013).

The foremost strategy used to manage anthropogenic impacts on coral reef ecosystems is marine protected areas (Graham et al. 2011, Edgar et al. 2014), which limit or prohibit all extractive activities. There is good evidence that the implementation of no-take marine protected areas have significant benefits for abundance and diversity of targeted organisms, especially in large, long- established and well-managed (well enforced) marine parks (e.g., Russ and Alcala 2010, Edgar et al. 2014). However, given the widespread and accelerating degradation of coral reef ecosystems, marine protected areas are increasingly expected to provide additional benefits beyond protection of major targeted species (Graham et al. 2011), specifically related to preserving key functions and maximising ecosystem resilience. For example, marine protected areas are expected to confer some benefit for biological and physical habitat structure, but evidence of higher coral cover, or indeed more rapid recovery among coral assemblages inside marine protected areas is equivocal (Graham et al. 2011). To be really effective in ecosystem-level conservation, Graham et al. (2011) suggest that marine protected areas must be used in conjunction with a range of other management tools, which need to be applied according to specific local environmental and social contexts.

The removal of key functional groups of fishes through fishing activities has underpinned the degradation of coral reef systems worldwide (e.g., Hughes 1994; Rasher et al. 2013). In particular, regional reductions in herbivorous fishes have underpinned to shifts from coral- to macroalgal-dominated reefs in many regions (Hughes 1994; Rasher et al. 2013). Herbivory is widely acknowledged as a key process in maintaining a healthy balance between corals and algae, with large mobile, or roving, herbivorous fishes (namely surgeonfishes, parrotfishes, rabbitfishes and drummers) playing a predominant role (Bellwood et al. 2004, Hughes et al. 2007). On reefs with relatively intact fish assemblages, herbivores rapidly graze the reef substratum. In doing so, these fishes inhibit the proliferation of algal biomass, minimise coral-algal competition (Hughes et al. 2007; Rasher and Hay 2010) and facilitate coral settlement through the provision of suitable

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settlement sites (Hughes et al. 2007; Hoey et al. 2011), thereby promoting the recovery of coral assemblages following disturbance (Mumby and Steneck 2008). The importance of herbivores has often led to the assumption they are a uniform group, however recent studies have highlighted important differences in the ecosystem role of particular species. Herbivorous fishes may be broadly classified into two functional groups based on the algal material they target and their roles in ecosystem processes: grazers and browsers (Hoey and Bellwood 2009; Cheal et al. 2013). Grazing fishes (including scraping and excavating parrotfishes) generally feed on algal turfs, and in doing so limit the expansion and proliferation of macroalgae by removing macroalgal propagules contained within the turfs. The removal of larger, fleshy macroalgae is a separate but critical process on coral reefs and appears to be restricted to a limited suite species, the macroalgal browsers (Bellwood et a. 2006; Hoey and Bellwood 2009).

3.1 Elizabeth and Middleton Reefs (Lord Howe Commonwealth Marine Reserve)

The Lord Howe Commonwealth Marine Reserve covers an area of more than 110,000 km2 and encompasses the former Lord Howe Island Marine Park (Commonwealth Waters) and the former Elizabeth and Middleton Reefs Marine National Nature Reserve. The former Reserve is divided into two zones; the northernmost zones encompassing Middleton Reef and extending southward to latitude 29o53” is a no take zone, which allows “passive use by the public”, but no fishing. The southern “habitat protection zone” surrounding Elizabeth Reef allows limited recreational fishing (through a permit system), but is largely protected by its isolation; Elizabeth Reef is most accessible from Lord Howe Island, which is located 261 km further south. These high latitude reef systems lie close to the boundary between the Coral Sea and the Tasman Sea and thus provide habitats for both tropical and subtropical/warm temperate species. Middleton Reef lies 600 km east of the Australian coast at 29o56'S 159o03'E in the northern Tasman Sea. Both reefs rise independently from deep water (>2000m) and are thus separated by waters of oceanic depth. The nearest major coral formations of the southern Great Barrier Reef are 900 km to the north.

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Elizabeth and Middleton Reefs are of considerable ecological importance. Most notably, these reefs represent the southernmost open ocean platform reefs in the world, and support a unique mix of tropical and subtropical species, which are often at their southernmost and northernmost limits, respectively (Choat et al. 2006). Elizabeth and Middleton Reefs also support a number endemic species or species that are generally rare over most of their range. Due to their relative isolation, Elizabeth and Middleton Reef support populations of species that are otherwise fished throughout much of their range. In particular, black cod (Epinephelus daemelii) occur in relatively high abundances at both Elizabeth and Middleton (Choat et al. 2006), but has been overfished elsewhere in the southwestern Pacific. It is listed as vulnerable in New South Wales waters, and is also listed as ‘vulnerable’ species under the Environment Protection and Biodiversity Conservation Act 1999. Elizabeth and Middleton Reefs are one of the last remaining strongholds for this species (Choat et al. 2006, Hobbs and Feary 2007). For these reasons the maintenance of the Elizabeth and Middleton Reefs Marine National Nature Reserve is critical for the conservation of Australia’s marine biodiversity. Ongoing monitoring of marine resources to document temporal changes in community structure and trigger appropriate management action is crucial to the effective management of the Reserve.

A total of nine scientific surveys have been conducted at Elizabeth and/ or Middleton Reefs since 1979 (Table 1).

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Table 1. Marine surveys undertaken at Elizabeth and/ or Middleton Reef since 1979. For a comprehensive overview of surveys conducted prior to 1979 see Australian Museum (1992).

No. Year Locations Purpose Reference 1 1979 Elizabeth & Middleton Coral diversity Australian Museum 1992 2 1984 Elizabeth & Middleton Coral diversity Australian Museum 1992 3 1981 Middleton Assess impacts of Harriot 1998 COTS 4 1987 Elizabeth & Middleton Biodiversity Australian Museum 1992 assessment 5 1994 Elizabeth & Middleton Marine ecological Choat et al. unpublished survey data 6 2003 Elizabeth Marine ecological Oxley et al. 2004 survey 7 2006 Elizabeth & Middleton Marine ecological Choat et al. 2006 survey 8 2007 Elizabeth & Middleton Rapid assessment of Hobbs and Feary 2007 reef health 9 2011 Elizabeth & Middleton Marine ecological Pratchett et al. 2011b survey

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3.2 Objectives and scope

The purpose of this study was to undertake rigorous surveys of shallow coral reef habitats at Elizabeth and Middleton Reefs, within the Lord Howe Commonwealth Marine Reserve, to assess the current condition of these habitats and explicitly test for any changes in the period since the last survey, in March 2011. Replicate area- based surveys (transect-based) surveys for fishes, corals, benthic invertebrates and habitat structure, were undertaken at the same sites as the 2011 survey. Some of these sites were established in 2006, 2007 and expanded to the current 16 sites in 2011. Sampling locations and methods were matched as closely as possible to those used in 2011, enabling direct comparisons of biological assemblages and habitat structure between these two latest surveys.

The specific requirements of the present survey were: i) To examine benthic cover and composition, including estimates of percentage cover for hard (scleractinian) and soft (alcyonarian) corals, macroalgae, and other sessile organisms ii) To quantify structural complexity of reef habitats iii) To assess coral condition, by quantifying injuries caused by coral disease, corallivorous invertebrates (e.g., Acanthaster plancii and Drupella spp), or reef fishes, iv) To quantify the abundance of small/ juvenile corals (<5cm diameter), as a proxy of coral recruitment and population replenishment, v) To estimate size and abundance of apex predators, including black cod (Epinephelus daemelii) and sharks, vi) To estimate abundances of other reef fishes including the endemic doubleheader wrasse (Coris bulbifrons), rabbitfishes, drummers, damselfishes and butterflyfishes, and vii) To record the abundance of holothurians, urchins and other motile invertebrates.

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4 Methods

This study was conducted at Elizabeth and Middleton Reefs from March 4th to 9th, 2014. Sampling was undertaken at a total of 16 sites (Table 2; Figure 1). At all but one site (Site 3 at Middleton Reef), surveys were conducted within each of two different depth zones, i) the reef crest at 2-4m and ii) the reef slope at 6-10m. In each depth zone at each site, four replicate 50-m transects were laid end-to-end and parallel to the depth contour, with a minimum of 10 metres between adjacent transects (n = 124 transects in total across both reefs). Along each transect (a) benthic cover and composition, (b) reef topographic complexity, (c) juvenile coral densities (as a proxy for coral replenishment), (d) coral condition/health, (e) herbivorous fish assemblages, (f) endemic and site-attached fish densities, (g) apex predator densities, and (h) unitary invertebrates were quantified using the same methods as those employed in the 2011 surveys of the same sites.

4.1 Habitat structure

Benthic cover and composition - To quantify percentage cover of corals, marcoalgae and other sessile organisms we used the point-intercept method, recording the specific organisms or habitat types underlying each of 100 uniformly spaced points (50cm apart) along each transect. All hard (scleractinian) and soft (alcyonarian) corals were identified to genus. For survey points that did not intersect coral, the underlying habitat was be categorised as macroalgae (> 5mm in height; identified to genera), turf algae (< 5mm in height), crustose coralline algae (CCA), rubble or sand. In doing so, these surveys provide data on cover of major habitat forming and other sessile taxa, as a proportion of available substrata.

Topographic complexity - Topographic complexity was estimated visually at the start of each transect, using the six-point scale formalised by Wilson et al. (2007), where 0 = no vertical relief (essentially flat homogenous habitat), 1 = low and sparse relief, 2 = low but widespread relief, 3 = moderately complex, 4 = very complex with numerous fissures and caves, 5 = exceptionally complex with numerous caves and overhangs.

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Figure 1: Map showing the location of the study sites. (a) Geographic location of Elizabeth and Middleton reefs off Australia’s east coast. (b) Locations of the nine study sites on Middleton Reef. (c) Location of the seven study sites on Elizabeth Reef. Note that site 6 on Elizabeth Reef was separated into shallow (6s) and deep (6d) due to the lack of shallow and deep habitat in close proximity. (d) Schematic diagram of reef cross-section showing the location of the three habitats and two depths sampled. Yellow circles = reef front locations, blue squares = back reef locations, and green stars = lagoon locations. The 16 sites are identical to those used in the 2011 surveys (Pratchett et al. 2011b). GPS co- ordinates are provided for each survey site in Tables 2.

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Table 2. Sites surveyed at Elizabeth and Middleton Reefs during the 2014 marine survey (4-9 March, 2014). GPS co-ordinates are given for each site. The sites are the same as those surveyed in 2011

Reef Habitat Site GPS

Elizabeth Exposed Reef Front 1 29º 57.161’S; 159º 01.295’E

2 29º 56.600’S; 159º 01.100’E

3 29º 55.970’S; 159º 01.390’E

Back Reef 4 29º 55.616’S; 159º 02.406’E

5 29º 55.113’S; 159º 03.635’E

Lagoon 6s 29º 56.148’S; 159º 05.346’E

6d 29º 56.079’S; 159º 05.596’E

7 29º 56.170’S; 159º 03.097’E

Middleton Reef Front 5 29º 26.880’S; 159º 03.310’E

6 29º 28.637’S; 159º 07.296’E

7 29º 25.803’S; 159º 07.988’E

8 29º 29.086’S; 159º 04.597’E

9 29º 27.262’S; 159º 08.532’E

Back reef 1 29º 26.610’S; 159º 05.860’E

2 29º 27.177’S; 159º 05.061’E

Lagoon 3 29º 27.223’S; 159º 03.940’E

4 29º2 6.574’S; 159º 06.901’E

Coral replenishment - In addition to adult cover (quantified as % cover of different taxa) we quantified the density (number per 10m2) of juvenile corals at each site, as means for assessing rates of replenishment and population turnover among coral populations. The density of juvenile coral colonies (<5 cm maximum diameter, following Rylaarsdam 1983) were recorded within a 10 x 1 m belt at the start of each transect. The density of juvenile corals does not necessarily provide a good proxy for spatial variation in settlement rates or larval supply (Penin et al. 2010), but is ecologically relevant for assessing population replenishment because

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it encapsulates variation in both settlement rates of coral larvae and early post- settlement mortality, to effectively measure the number of new coral colonies likely to become established in each habitat and location. Spatial variation in densities of juvenile corals is particularly important in assessing the recovery potential of coral assemblages following acute episodic disturbances, such as coral bleaching (Hoey et al. 2011).

Spatial variation and temporal variation in coral cover and other major benthic categories (e.g., macroalgae) were analysed using univariate parametric analyses (ANOVA) following appropriate (square root – arcsine) transformation of proportional cover. Analyses tested for differences between years (2011 and 2014), reefs (Elizabeth and Middleton reefs), reef zones (reef front, back reef, and lagoon), and depths (reef crest and reef slope). To assess temporal trends over longer time scales, the data from the two most recent surveys (2011 and 2014) was compared to data collected in 2006 and 2007. However, differences in specific sites that were surveyed, and changes in the specific survey methodology precluded any statistical comparisons to data collected in 2006 and 2007. Consequently, we provide a graphical presentation of changes in mean coral cover from a limited number of sites from Middleton Reef (sites 3, 4, 5, 6, and 7) that were surveyed in 2006.

Coral health - To assess local impacts of the corallivorous crown-of-thorns starfish (Acanthaster planci), pin-cushion starfish (Culcita novaeguineae) and the corallivorous gastropod Drupella spp., as well as other large and potentially destructive coral predators such as excavating parrotfishes (Choat et al. 2006, Hoey and Bellwood 2008), any evidence of feeding scars were recorded 1-metre either side of the 50-m transect path (i.e., 50 x 2m). Large parrotfish, pufferfish, and triggerfish are all known to feed on live corals (Hoey and Bellwood 2008; Bonaldo et al. 2014), but unlike other such as the butterflyfishes that ‘pick’ at individual polyps when feeding, these larger fishes remove both the coral tissue and the underlying coral skeleton when feeding. When in sufficiently high densities these fishes can have a large impact on coral populations (Hoey and Bellwood 2008; Bonaldo and Bellwood 2011). In addition to signs of corallivory,

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any evidence of adverse coral health, such as coral bleaching, and coral disease were also recorded within the 50 x 2 m belt transects.

4.2 Coral reef fishes

Herbivorous reef fishes – Functionally important (mostly, roving herbivorous) coral reef fishes were surveyed using underwater visual census along 5m wide paths along each 50-m transects at every study site. Each transect consisted of a diver swimming along the depth contour and recording all fishes from the (surgeonfishes), Labridae (parrotfishes), Kyphosidae (drummers), and Siganidae (rabbitfishes) within a 5m belt while simultaneously deploying the transect tape. This procedure minimised disturbance prior to censusing and allowed a specified area to be surveyed. Individual fishes were identified to species and placed into 5cm size categories. Care was taken not to census fish that left and subsequently re-entered the transect area. Fish densities were converted to biomass using published length-weight relationships for each species, following Hoey and Bellwood (2009).

Endemic and site-attached reef fishes – Smaller more site-attached species from the families Chaetodontidae (butterflyfishes) and Pomacentridae (damselfishes), and juvenile double-header wrasse (Coris bulbifrons) were counted along a 2m wide path along each transect (i.e., 50 x 2m) at every study site. Adult double-header wrasse (C. bulbifrons) were surveyed in a 5m wide belt along each transect (i.e., 50 x 5m) at each study site. Mostly, we were interested in the local abundance of endemic species (Table 3).

Table 3: Endemic fishes that were surveyed at Elizabeth and Middleton Reef, following Hobbs and Feary (2007). Species Common name Distribution Coris bulbifrons Doubleheader Lord Howe, Norfolk, Middleton, wrasse Elizabeth, NSW Chaetodon tricinctus Three striped Lord Howe, Norfolk, Middleton, Elizabeth Amphiprion mccullochi McCulloch’s Lord Howe, Middleton, Elizabeth, NSW, anemonefish Norfolk Epinephelus daemelii Black cod, Lord Howe, Middleton, Elizabeth, SE saddletail grouper Australia, Kermadec, Northern New Zealand

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Apex predators - To effectively survey larger and highly mobile reef-associated fishes (namely black cod and sharks) 250 - 500m long underwater visual transects were conducted along the reef at each study site, following Robbins et al (2006). Transects were run approximately parallel with the reef crest and covered a depth range of approximately 4 – 20m. All black cod (E. daemelii) and sharks (primarily the Galapagos shark, Carcharinus galapagensis) were recorded within 10-m of the transect path, giving a total sample area of 10,000m2 (1 hectare) for a 500m transect, and 5,000m2 (0.5 hectare) for a 250m transect. Two transects were conducted at each site, one 500m in length and the other 250m in length.

4.3 Unitary invertebrates

Unitary invertebrates - Counts of unitary invertebrates (mainly holuthurians, urchins, Acanthaster planci and Drupella spp.) were conducted within a 4-m wide belt along each transect. Densities of all functionally important motile invertebrates were recorded 2-m either side of the 50-m transect line (i.e. 50 x 4m), giving a sample area of 200m2 following Pratchett et al. (2011b).

4.4 Species richness

Species richness - Comprehensive lists of all fish and vertebrate species encountered during sampling were compiled, as per Pratchett et al. (2011b). To distinguish between those species that are possible new vagrants versus rare species that have simply gone undetected until now, we assigned all species to one of 4 different abundance categories; Abundant (A) – >500 individuals recorded; Common (C) - >100 individuals recorded; Occasional (O) - 5-100 individuals recorded; Rare (R) –<5 individuals recorded.

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5 Findings

5.1 Habitat structure

5.1.1 Coral cover

Mean cover of hard (scleractinian) corals recorded at Elizabeth reef was 29.18% (±1.70 SE), ranging from 15.31 (±2.04 SE) at lagoon habitats, up to 36.83% (±2.22 SE) at reef fronts (Figure 2a). Coral cover at Middleton reef was generally lower than Elizabeth, averaging 19.03% (±1.32 SE) and much more variable between habitat types, ranging from 8.31% (±1.38 SE) in the back reef, up to 25.50% (±1.32 SE) at reef front habitats (Figure 2a). Total coral cover varied significantly between reefs and habitat zones, but was also influenced by the interaction of reef, habitat and depth (Table 4). At Elizabeth Reef, coral cover on the exposed reef front was greater on the reef slope compared to the crest, while on the back reef and lagoon coral cover tended to be greater on the crest, or broadly comparable among depths (Figure 2a). At Middleton Reef, coral cover on the reef crest was greater than the slope in the lagoon, but lower than the crest on the back reef (Figure 2a).

Table 4: ANOVA comparing hard coral cover between reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). Proportional cover of corals was arcsin-squareroot transformed. Source SS df MS F p Reef 0.64 1 0.64 54.85 0.000 Habitat 0.94 2 0.47 40.14 0.000 Depth 0.00 1 0.00 0.05 0.832 Reef x Habitat 0.32 2 0.16 13.79 0.000 Reef x Depth 0.00 1 0.00 0.00 0.973 Habitat x Depth 0.07 2 0.03 2.79 0.066 Reef x Habitat x Depth 0.13 2 0.07 5.63 0.005 Error 1.31 112 0.01

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Figure 2. Spatial variation in the cover of (a) scleractinian (hard) corals and (b) macroalgae among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE.

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5.1.2 Macroalgal cover

Mean cover of macroalgae at Elizabeth and Middleton reefs was 11.43% (±1.56 SE) and generally similar (<10% cover) across most habitats. There was however very high cover of macroalgae (mostly Codium spp) recorded on the reef crest within back reef (64.25 ± 3.22%) and lagoon (19.38 ±7.60%) at Middleton Reef (Figure 2b). Overall, macroalgal cover varied significantly between reefs, habitats and depths (Table 5), with mean cover of macroalgae being over 2-fold greater on Middleton (15.16 ± 2.61%) than on Elizabeth Reef (6.89 ± 1.17%). These observed differences were directly attributable to the elevated macroalgal cover in the deeper leeward (back reef and lagoon) sites at Middleton (Figure 2b).

Table 5: ANOVA comparing macroalgal cover between reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). Proportional cover of macroalgae was arcsin-squareroot transformed. Source SS df MS F p Reef 0.37 1 0.37 12.42 0.001 Habitat 0.28 2 0.14 4.75 0.010 Depth 1.92 1 1.92 64.51 0.000 Reef x Habitat 0.50 2 0.25 8.34 0.000 Reef x Depth 0.62 1 0.62 20.99 0.000 Habitat x Depth 1.06 2 0.53 17.77 0.000 Reef x Habitat x Depth 0.97 2 0.48 16.30 0.000 Error 3.33 112 0.03

5.1.3 Temporal changes in coral and macroalgae

Coral cover recorded in 2014 at Elizabeth and Middleton Reefs was generally very similar to that recorded in 2011 (Figure 3a), though there have been slight declines at both Elizabeth Reef (2011: 33.3%, 2014: 29.2%) and Middleton Reef (2011: 20.3%, 2014: 19.0%). These estimates equate to an annual geometric rate of decline of 4.3% and 2.1% for Elizabeth and Middleton, respectively. Coral cover displayed limited change on the exposed reef front on both reefs, and the back reef on Elizabeth Reef, however there were significant declines in coral cover in the lagoon at both reefs, and the back reef on Middleton Reef (Table 5, Figure 3a). The greatest decline was recorded in the lagoon at Elizabeth Reef with almost half the coral cover lost over 3 years (2011: 30.44%, 2014: 15.31%). Significant declines in coral cover were also recorded within the back reef (48% decline, 2011: 15.69%, 2014: 8.31%) and lagoon (35% decline, 2011: 23.25%, 2014: 15.08%) at Middleton Reef (Figure 3a).

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Figure 3. Temporal changes (2011 versus 2014) in average cover (±SE) of (a) scleractinian (hard) corals, and (b) macroalgae across three different habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs.

The observed decline in coral cover within the lagoon at both reefs is largely attributable to a reduction in the cover of arborescent within those habitats. In 2011 the lagoon sites at both Elizabeth and Middleton Reefs were dominated by extensive thickets of arborescent Acropora (Pratchett et al. 2011b),

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however some of these thickets had been reduced to rubble by 2014. The causes of such declines are unknown but could be related to a high temperature event or large storm as these corals are highly susceptible to both coral bleaching and storm damage (Marshall and Baird 2000, Madin and Connolly 2006). Alternatively, the reduction in coral cover could be a result of increased macroalgal cover. Many macroalgal species produce allelopathic chemicals that damage corals (Rasher and Hay 2010) and excessive growth and coverage of macroalgae has been shown to reduce the growth, survivorship and fecundity of established coral colonies (e.g., Hughes et al. 2007). However, the reduction in hard coral cover was only accompanied by an increase in macroalgal cover at one of the three locations, the back reef at Middleton Reef (Figure 3b), while macroalgal cover either remained stable (Middleton lagoon) or decreased (Elizabeth lagoon) at the other two locations. It therefore seems likely (although not certain) that the increased macroalgal cover is a consequence rather than a cause of coral mortality. Algal taxa are typically opportunistic, rapidly colonising coral skeletons after the coral tissue dies (Diaz-Pulido and McCook 2002).

Table 6: ANOVA comparing hard coral cover between years (2011 vs 2014), reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). Proportional cover of corals was arcsin-squareroot transformed.

Source SS df MS F p Year 0.16 1 0.16 8.77 0.003 Reef 1.20 1 1.20 65.12 0.000 Habitat 0.88 2 0.44 23.86 0.000 Depth 0.00 1 0.00 0.14 0.705 Year x Reef 0.00 1 0.00 0.08 0.773 Year x Habitat 0.26 2 0.13 7.21 0.001 Reef x Habitat 0.33 2 0.17 9.07 0.000 Year x Depth 0.01 1 0.01 0.38 0.536 Reef x Depth 0.04 1 0.04 2.16 0.143 Habitat x Depth 0.14 2 0.07 3.75 0.025 Year x Reef x Habitat 0.09 2 0.05 2.51 0.083 Year x Reef x Depth 0.04 1 0.04 2.05 0.154 Year x Habitat x Depth 0.30 2 0.15 8.22 0.000 Reef x Habitat x Depth 0.43 2 0.21 11.59 0.000 Y x R x H x D 0.39 2 0.20 10.65 0.000 Error 4.12 224 0.02

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Overall, the cover of macroalgae recorded at Elizabeth and Middleton Reefs in 2014 (11.4 ± 1.6%), was broadly comparable to that recorded in 2011 (10.2 ± 0.9%). There were, however, slight and opposite changes in the cover of macroalgae at each reef, with cover decreasing by 4.4% at Elizabeth Reef (2011: 11.3%, 2014: 6.9%), but increasing by 6.0% at Middleton Reef (2011: 9.2%, 2014: 15.2%) (Table 7, Figure 3b). At Elizabeth Reef, macroalgal cover declined from 2011 to 2014 in the back reef and lagoon, and remained relatively stable on the exposed reef front. At Middleton Reef, macroalgal cover remained relatively stable in the lagoon and exposed reef front habitats, but increased markedly from 15.4 % (± 3.4 SE) in 2011 to 32.7% (± 8.3 SE) in 2014. This increase was especially pronounced in the shallow reef crest environment where macroalgae occupied approximately 2/3 of the total reef substrata (see Section 4.1.2 above).

Table 7: ANOVA comparing macroalgal cover between years (2011 vs 2014), reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). Proportional cover of macroalgae was arcsin-squareroot transformed. Source SS df MS F p Year 0.03 1 0.03 1.04 0.308 Reef 0.10 1 0.10 4.04 0.046 Habitat 0.44 2 0.22 8.59 0.000 Depth 1.78 1 1.78 70.16 0.000 Year x Reef 0.29 1 0.29 11.44 0.001 Year x Habitat 0.08 2 0.04 1.56 0.213 Reef x Habitat 0.53 2 0.26 10.41 0.000 Year x Depth 0.39 1 0.39 15.33 0.000 Reef x Depth 0.79 1 0.79 31.01 0.000 Habitat x Depth 1.87 2 0.93 36.82 0.000 Year x Reef x Habitat 0.10 2 0.05 2.00 0.138 Year x Reef x Depth 0.05 1 0.05 2.08 0.151 Year x Habitat x Depth 0.41 2 0.20 7.99 0.000 Reef x Habitat x Depth 0.91 2 0.45 17.89 0.000 Y x R x H x D 0.19 2 0.10 3.84 0.023 Error 5.68 224 0.03

Examination of trends in coral cover over longer timeframes (1994-2014) at selected Middleton Reef sites revealed that current levels of coral cover are generally higher than those recorded from 1994-2006 (ca. 10%; Figure 4). At two of the exposed reef front sites (site 5 and 7) coral cover has increased slowly and consistently from 1994 to 2014. In contrast, another exposed reef front site (site 6) and the lagoon sites (site 3 and 4) showed greater increases from 2006 to 2011, but have declined again in 2014 (Figure 4).

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Figure 4. Temporal changes in mean coral cover (±SE) at selected sites from Middleton Reef. Data for 1994 are from Oxley et al. (1994), and 2006 from Choat et al. (2006)

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5.1.4 Topographic complexity

Topographic complexity was generally very low across all sites surveyed at Elizabeth and Middleton Reefs, with most sites assigned values of <3 (out of 5). On Elizabeth Reef there was very little change in estimates of topographic complexity among the three habitats, with the reef-wide estimate being 1.82 (±0.01 SE) in 2014 compared to 1.79 (±0.10 SE) in 2011 (Figure 5). On Middleton Reef there were slight increases in estimated complexity within the reef front and back reef habitats, resulting in an increase in the reef average from 1.32 (±0.07 SE) in 2011 to 2.04 (±0.08 SE) in 2014. Topographic complexity is not expected to vary between surveys given limited changes in live coral cover. The reported differences may have resulted from slight differences in the starting position of transects within each habitat, or differences in the methods used. In 2014, visual assessments were made in situ by a diver (always MSP), whereas in 2011 assessments were made from photographs taken at the start of each transect, albeit by the same observer (MSP). Some structural details such as gutters may have been obscured in photographs and led to an underestimate in topographic complexity in 2011.

Figure 5. Temporal changes (2011 versus 2014) in the topographic complexity (±SE) of reef substratum across three different habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Topographic complexity was assessed visually using a 6-point scale (following Wilson et al. 2007). 0 = flat homogenous habitat, 5 = exceptionally complex.

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5.2 Replenishment and resilience of coral populations

5.2.1 Juvenile corals

Recovery rates of coral assemblages are critically dependent upon replenishment (the rate at which new coral colonies recruit, survive and grow to colonise available space), and the regrowth of remnant coral tissues following declines in population size (Bellwood et al. 2004, Hoey et al. 2011). Rates of replenishment depend upon the supply of coral larvae, their successful settlement to the reef, as well as their early post-settlement growth and survivorship (Penin et al. 2010; Trapon et al. 2013a,b). To assess current levels of population replenishment for hard coral communities at Elizabeth and Middleton, we quantified densities of juvenile corals (individuals <50mm maximum diameter) on natural substrata, following Pratchett et al. (2011a). Spatial variation in the densities of juvenile corals is critical for assessing the recovery potential of coral assemblages following acute episodic disturbances, such as coral bleaching (Hoey et al. 2011).

A total of 1,396 juvenile corals from 25 genera were recorded across 124 transects on Elizabeth and Middleton Reefs in March 2014, equating to a mean density of 11.26 juvenile corals per 10m2 (± 0.71 SE). This density is markedly lower than (approximately 1/3) those recorded from tropical reefs using identical methods (ca. 39.3 ind/10m2; Pratchett et al. 2011a), but broadly comparable to densities recorded from Lord Howe Island (7.7 ind/10m2; Hoey et al. 2011). The juvenile coral assemblage was dominated by five coral genera; (19.1%), Porites (15.6%), Montastrea (14.6%), Acropora (14.3%), and Isopora (14.4%). Collectively these genera accounted for almost 80% of all juvenile corals recorded (Figure 6). The composition of juvenile corals differs considerably to that of reefs in the southern, central and northern Great Barrier Reef where Acropora typically accounts for 40-60% of juvenile coral assemblage (Trapon et al. 2013a). The relative low densities of juvenile Acropora and the predominance of brooding coral genera, namely as Pocillopora, Isopora and Porites, suggests that the coral assemblages on Elizabeth and Middleton Reefs may be largely dependent on self- recruitment, as opposed regular larval supply from reefs in the southern GBR.

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Figure 6. Composition of juvenile coral assemblages surveyed on Elizabeth and Middleton Reefs in March 2014. Data are total number of juvenile corals recorded in each genus across 124 transects.

Densities of juvenile corals did not differ between Elizabeth (11.14 colonies /10m2 ± 1.10 SE) and Middleton (11.35 colonies/10m2 ± 0.92 SE), but varied between depths and habitats (Table 8, Figure 7). On both reefs juvenile corals were twice as abundant on the exposed reef front (Elizabeth: 16.63 ±1.75 colonies /10m2; Middleton: 14.40 ±1.17 colonies /10m2) than the back reef (Elizabeth: 7.38 ±0.59 colonies/10m2; Middleton: 7.69 ±1.11 colonies/10m2) or the lagoon (Elizabeth: 6.69 ±1.72 colonies/10m2; Middleton: 6.08 ±1.52 colonies/10m2) (Figure 7). At Elizabeth Reef, juvenile corals were generally more abundant on the reef crest compared to the reef slope (Figure 7), while at Middleton Reef, there was no consistent pattern in densities of juvenile corals between depths. Juvenile coral densities were greater on the reef slope than the reef crest on the back reef, but displayed no differences between depths on the exposed reef front or in the lagoon at Middleton Reef (Figure 7).

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Table 8: ANOVA comparing the density of juvenile corals between reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). The density of juvenile corals was log(x+1) transformed.

Source SS df MS F p Reef 0.01 1 0.01 0.14 0.711 Habitat 4.01 2 2.00 31.89 0.000 Depth 0.23 1 0.23 3.67 0.058 Reef x Habitat 0.04 2 0.02 0.33 0.723 Reef x Depth 0.69 1 0.69 10.91 0.001 Habitat x Depth 0.64 2 0.32 5.08 0.008 Reef x Habitat x Depth 0.05 2 0.02 0.37 0.692 Error 7.04 112 0.06

Figure 7. Spatial variation in the density of juvenile scleractinian (hard) corals among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE.

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There was a small decline in the overall densities of juvenile corals between surveys, from 13.02 ± 0.75 colonies/10m2 in 2011 to 11.25 ± 0.71 colonies/10m2. There was, however, no consistency in the rate of change between reefs or habitats within each reef (Table 9; Figure 8). The density of juvenile corals declined markedly on Elizabeth Reef, from 16.85 to 11.14 colonies/10m2, yet remained largely unchanged on Middleton Reef (10.30 – 11.35 colonies/10m2). The decline on Elizabeth Reef was largely attributable to a 61% reduction in juvenile coral densities on the back reef, from 18.79 colonies/10m2 in 2011 to 7.38 colonies/10m2 in 2014, and to a lesser extent a 22% decline on the exposed reef front (Figure 8). At Middleton Reef a 47% decrease in juvenile coral densities in the lagoon was offset by 39% increase on the exposed reef front (Figure 8).

Table 9: ANOVA comparing the density of juvenile scleractinian corals between years (2011 vs 2014), reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). The density of juvenile corals was log (x + 1) transformed.

Source SS df MS F p Year 0.77 1 0.77 13.92 0.000 Reef 0.36 1 0.36 6.41 0.012 Habitat 4.60 2 2.30 41.33 0.000 Depth 0.01 1 0.01 0.22 0.641 Year x Reef 0.22 1 0.22 3.97 0.048 Year x Habitat 0.56 2 0.28 5.04 0.007 Reef x Habitat 1.02 2 0.51 9.14 0.000 Year x Depth 0.59 1 0.59 10.63 0.001 Reef x Depth 0.08 1 0.08 1.37 0.244 Habitat x Depth 0.71 2 0.35 6.36 0.002 Year x Reef x Habitat 0.55 2 0.27 4.93 0.008 Year x Reef x Depth 0.74 1 0.74 13.28 0.000 Year x Habitat x Depth 0.07 2 0.04 0.63 0.533 Reef x Habitat x Depth 0.06 2 0.03 0.53 0.588 Y x R x H x D 0.24 2 0.12 2.11 0.123 Error 12.02 216 0.06

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Figure 8. Temporal changes (2011 versus 2014) in the densities of juvenile scleractinian corals (< 50mm diameter) across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE.

Spatial variation in the abundance of juvenile corals in 2014 was significantly and positively related to both the cover of live coral and bare space (i.e. the sum of turf algae and CCA) (Figure 9). The positive relationship between live coral cover and the density of juvenile corals suggests that coral cover at individual sites is limited to some extent by the recent replenishment of hard corals. This is particularly evident on the exposed reef fronts of both Elizabeth and Middleton reefs, where generally both coral cover and the density of juvenile corals were greater than the lagoon or back reef habitats (Figure 9a). In addition to the relationship with live coral cover, a positive relationship was evident between the abundance of juvenile corals and the cover of bare space (Figure 9c). While the cover of bare space was typically high across all habitats (>40% cover) the positive relationship with juvenile corals suggests that the availability of suitable settlement sites for coral larvae may be limiting at some sites. This was particularly evident at some of the lagoon and back reef sites on Middleton Reef where low cover of bare space coincided with extremely low densities of juvenile corals (Figure 9c).

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Resilience, and especially recovery, of tropical coral populations is strongly influenced by interactions between corals and macroalgae. Excessive growth and coverage of macroalgae may reduce the growth, survivorship, and fecundity of established coral colonies (Hughes et al. 2007; Rasher and Hay 2010). Perhaps more importantly, macroalgae may limit coral recruitment and the recovery potential of reefs by inhibiting settlement, or smothering new recruits (Hughes et al. 2007). At Elizabeth and Middleton Reef, a negative, but non-significant relationship was found between the cover of macroalgae and the abundance of juvenile corals (Figure 9b). Although non-significant, macroalgae appears to be having some influence on the replenishment of coral populations at Elizabeth and Middleton Reefs. Sites with high macroalgal cover (i.e. Middleton back reef) tended to have low densities of juvenile corals, and conversely sites with high densities of juvenile corals had low cover of macroalgae, in particular the exposed reef front sites (Figure 9b).

In addition, to the relationship with juvenile coral abundances there was a significant negative correlation between the cover of scleractinian coral and the cover of macroalgae (Pearsons correlation, r = -0.517, n = 32, p = 0.002). Interestingly both of these relationships are stronger than those reported in 2011 across identical sites on Elizabeth and Middleton Reefs (juvenile corals: r = -0.14, n = 32, p = 0.46; coral cover: r = -0.27, n = 32, p = 0.14) (Pratchett et al. 2011b). While the causes of the increased cover of macroalgae are unknown, it appears to be having a negative impact on the health of these reefs. While as yet this is not a cause for concern, there is a clear need to investigate the potential causes of the elevated macroalgal cover, in particular at the lagoon and back reef sites on Middleton reef.

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Figure 9. Relationships between the density of juvenile corals and the cover of (a) live scleractinian corals, (b) macroalgae, and (c) bare space (a proxy for the availability of potential settlement sites for coral larvae). Each point represents the mean of 4 transects at each depth at each of 16 sites. triangles: Elizabeth Reef, circles: Middleton Reef. Colours represent the three habitats shown in Figure 1. Yellow: exposed reef front; Blue: back reef; Green: lagoon.

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5.2.2 Coral predators

Corallivorous starfish, such as the crown-of-thorns starfish, Acanthaster planci (L.) and the pincushion starfsh, Culcita novaeguineae Muller & Troschel, represent significant biological threat to Indo-Pacific coral reefs (Pratchett et al. 2011a, 2014; Baird et al. 2013). The crown-of-thorns starfish, in particular, is renowned for its capacity to rapidly increase in densities (i.e., outbreaks) and cause large-scale devastation on tropical coral reefs. During outbreaks, very high densities of A. planci can kill up to 80% of corals across large reef areas (reviewed in Pratchett et al. 2014). Indeed, outbreaks of crown-of-thorns starfish have been identified as one of the major causes of coral decline on the Great Barrier Reef over the past three decades (Osborne et al. 2011, De’Ath et al. 2012). Even at moderate densities, A. planci has the potential to greatly modify coral community structure by selectively feeding on certain corals (Pratchett 2010). Similarly, C. novaeguineae exhibits strong preference for certain corals (Glynn and Krupp 1986) and could potentially contribute to directional shifts in species composition of coral assemblages, however it is not known to occur at largely elevated, or outbreak, densities. Normal densities of C. novaeguineae are typically higher than for non-outbreak populations of A. planci, such that C. novaeguineae may represent a persistent force in structuring coral communities (e.g., Quinn and Kojis 2003). However, few studies have ever considered the ecological impacts of C. novaeguineae (but see Pratchett et al. 2011a).

Densities of invertebrate corallivores (i.e. Acanthaster planci, Culcita novaeguineae, and Drupella spp.) have increased between 2011 and 2014, but remain very low at Elizabeth and Middleton Reefs. Both A. planci and C. novaeguineae were present at Elizabeth and Middleton Reefs in 2014, but as in 2011 they were only recorded in very low numbers. A total of four A. planci were recorded across all 124 transects in 2014 giving an average density of 1.6 starfish per hectare, compared to a single individual (0.4 starfish/ha) in 2011. Three A. planci (25-35 cm diameter) were recorded on the reef slope at site 4 (back reef) on Elizabeth Reef, and a single individual (15 cm diameter) was recorded on the reef slope at site 9 (exposed reef front) on Middleton Reef. A total of six C. novaeguineae were recorded in 2014 (compared to three C. novaeguineae in

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2011), with three individuals recorded at each reef. On both Elizabeth and Middleton Reefs two C. novaeguineae were recorded in the lagoon and one individual on the back reef.

Densities of A. planci recorded during the last two surveys (2014 and 2011) are amongst the lowest ever recorded on Elizabeth and Middleton Reefs. Early qualitative estimates of A. planci suggest there has been a history of outbreak densities on Elizabeth and Middleton, with hundreds of starfish being recorded at various sites in 1981 (Veron and Done unpublished in The Australian Museum 1992), and 1987 (The Australian Museum 1992), with the last known outbreak of A. planci being recorded at Middleton Reef in 1994 (Harriot 1998). There is little doubt that these outbreak densities caused significant coral loss along the exposed reef front and are likely to be a significant factor in the low coral cover recorded on Middleton Reef in from 1994 to 2006 (see Figure 4 above). Further, significant densities of A. planci (114 starfish/ha) were recorded in the vicinity of Site 6 (Exposed reef front) at Middleton Reef and moderate densities (<6 starfish/ha) were also recorded at other exposed locations at both Elizabeth and Middleton Reefsin 2006 (Choat et al. 2006).

In accordance with the low densities of A. planci and C. novaeguineae, the incidence of starfish feeding scars was very low. In total 17 coral colonies were observed to have evidence of starfish feeding scars, with these being recorded along 5% of transects (i.e., 6 of the 124 transects), compared to 16% in 2011. Similarly, evidence of feeding by the corallivorous gastropod Drupella spp. was low, with only 13 coral colonies displaying damage from Drupella spp. These low incidence of feeding scars are likely to reflect the normal ongoing effects of low density populations of different corallivores.

5.2.3 Coral condition

The incidence of coral disease was extremely low on both Elizabeth and Middleton Reefs in 2014, with only 5 infected corals being observed across the 124 transects. Two corals infected with black band disease were recorded with the lagoon (site 7) at Elizabeth Reef, and single coral colonies infected with white syndrome were recorded at the exposed reef front on Elizabeth (site 2) and Middleton Reefs (sites

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8 and 9). Black band disease is produced by a bacterial consortium that forms a conspicuous black band and actively destroys coral tissue as it progresses across a colony, leaving behind the bare coral skeleton (Richardson 2004). White syndrome is a generic term used to describe disease in corals displaying acute tissue loss exposing the underlying white skeleton in the absence of other disease signs. The causative agent for White Syndrome are unresolved but have been related to high prevalence of proteobacteria and/or ciliates (Sussman et a. 2008, Sweet and Bythell 2012).

The most prevalent source of coral damage on Elizabeth and Middleton Reefs was physical damage to the branches, presumably caused by feeding activities of fishes (i.e., ‘cropping’). Across all transects 99 coral colonies were observed to exhibit physical damage consistent with the cropping activities of fishes, equating to a density of <1 damaged colony/100m2. Of these the majority of damaged corals were Acropora (67%), Pocillopora (18%), Isopora (13%), and Stylophora (2%). There was, however considerable spatial variation in the distribution of damaged, or ‘cropped’, corals across both reefs (Table 10, Figure 10). The highest densities of cropped corals were all recorded within the shallow reef crest on the exposed reef fronts of both Elizabeth (1.3 ± 0.5 colonies/100m2) and Middleton (2.7 ± 0.5 colonies/100m2), and the back reef at Elizabeth Reef (1.0 ± 0.6 colonies/100m2). These high incidences of cropped corals corresponded with a high density of excavating parrotfish, namely Chlorurus microrhinos and C. frontalis, within these habitats (see Section 4.3.1 below). The observed physical damage to corals is not a cause for concern and is likely to reflect the normal ongoing effects of natural densities of these fishes.

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Table 10: ANOVA comparing the density of damaged or cropped corals between reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope).

Source SS df MS F p Reef 1.62 1 1.62 0.97 0.327 Habitat 33.03 2 16.51 9.92 0.000 Depth 9.14 1 9.14 5.49 0.021 Reef x Habitat 11.27 2 5.64 3.38 0.037 Reef x Depth 0.42 1 0.42 0.25 0.615 Habitat x Depth 15.25 2 7.63 4.58 0.012 Reef x Habitat x Depth 4.77 2 2.39 1.43 0.243 Error 186.48 112 1.67

Figure 10. Spatial variation in the density of corals displaying physical damage consistent with fish feeding (i.e., cropped) among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE.

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5.3 Coral reef fishes

5.3.1 Roving herbivorous fishes

The overall abundance of herbivorous fishes was very similar between reefs, averaging 54.8 (± 5.8 SE) fishes/250m2 on Elizabeth and 59.6 (± 5.8 SE) fishes/250m2 on Middleton Reef (Table 10, Figure 11). There were, however, significant differences in the abundance of roving herbivores between depths and habitats (Table 10a). The abundance of roving herbivorous fishes was consistently higher on the reef crest than the reef slope at exposed reef front sites on both Elizabeth and Middleton Reefs, but displayed limited difference among depths at either the lagoon or back reef (Figure 11a). This finding is consistent with numerous studies on tropical reefs that have demonstrated the abundance and/or impact of herbivorous fishes is greatest on exposed reef crests and decreases both across the reef and down the reef slope (Fox and Bellwood 2007; Hoey and Bellwood 2008; Hoey et al. 2013).

In contrast to abundance, the biomass of herbivorous fishes differed between reefs, averaging 30.7 (± 4.6 SE) kg/250m2 on Elizabeth and 25.2 (± 3.8 SE) kg/250m2 on Middleton Reef (Table 10b, Figure 11b). In addition to these reef scale difference, there were also significant differences in herbivore biomass among habitats, with the lowest biomass recorded in the lagoon at both reefs (Table 11, Figure 11b). The distinction between abundance and biomass reflects differences in the body size at each location, and is important as it numerous studies have shown that herbivore biomass is a better predictor of ecological impact than abundances (e.g., Bonaldo and Bellwood 2008; Lokrantz et al. 2008; Hoey and Bellwood 2009). The low biomass, but comparable abundances, of herbivorous fishes within the lagoons of Elizabeth and Middleton reefs is reflective of the generally smaller bodied fishes (in particular juvenile parrotfishes) within this habitat, compared to the exposed reef fronts and back reef habitats.

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Table 11: ANOVA comparing (a) the abundance, and (b) the biomass of roving herbivorous fishes between reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). Both abundance and biomass data were log(x +1) transformed.

Source SS df MS F p (a) Abundance Reef 0.02 1 0.02 0.10 0.756 Habitat 0.72 2 0.36 2.25 0.111 Depth 0.28 1 0.28 1.76 0.188 Reef x Habitat 0.28 2 0.14 0.87 0.422 Reef x Depth 0.04 1 0.04 0.24 0.622 Habitat x Depth 1.12 2 0.56 3.46 0.035 Reef x Habitat x Depth 0.17 2 0.08 0.52 0.599 Error 18.08 112 0.16

(b) Biomass Reef 1.76 1 1.76 5.76 0.018 Habitat 10.38 2 5.19 17.03 0.000 Depth 0.80 1 0.80 2.62 0.108 Reef x Habitat 0.98 2 0.49 1.61 0.204 Reef x Depth 1.42 1 1.42 4.66 0.033 Habitat x Depth 0.74 2 0.37 1.21 0.303 Reef x Habitat x Depth 0.39 2 0.19 0.64 0.532 Error 34.14 112 0.31

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Figure 11. Spatial variation in (a) the abundance, and (b) the biomass of roving herbivorous fishes among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE. The dashed lines represent overall reef means.

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Overall, the abundance of herbivorous fishes recorded at Elizabeth and Middleton Reefs in 2014 was broadly similar to that recorded in 2011 (Table 12, Figure 12a), though there have been some declines, albeit non-significant, at both Elizabeth (2011: 67.3 fish/250m2, 2014: 54.8 fish/250m2) and Middleton (2011: 72.8 fish/250m2, 2014: 59.6 fish/250m2). There was, however, a significant decline in the abundance of herbivores within the lagoons at both reefs, while abundances were relatively stable at back reef and exposed reef front sites (Figure 12a). The observed declines in the lagoons was attributable to a reduction small grazing fishes, primarily juvenile parrotfishes, in these habitats (Figure 12b), and is likely to be related to the reduction in coral cover at these sites (see Section 4.1.3 above). Interestingly, there has been a decline in the abundance of grazing fishes at both Elizabeth and Middleton from 2011 to 2014, and a concomitant increase in the abundance of browsing fishes (Figure 12b, c).

Although limited temporal changes were evident in the abundance of herbivorous fishes from 2011 to 2014, there was a 2.3-fold increase in herbivore biomass on Elizabeth Reef, from 13.2 (±1.9) kg/250m2 in 2011 to 30.7 (±4.7) kg/250m2 in 2014 (Table 12, Figure 13). This increase in total herbivore biomass was a result of 6- to 7-fold increases in the biomass of browsing fishes (primarily the Pacific Chub Kyphosus pacificus and the Spotted Sawtail Prionurus maculatus) on the exposed reef front and back reef (Figure 13c). In contrast, there was limited variation in the biomass of herbivorous fishes between years on Middleton (Figure 13).

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Table 12: ANOVA comparing (a) the abundance, and (b) the biomass of herbivorous fishes between years (2011 vs 2014), reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope). Both abundance and biomass data were log(x +1) transformed.

Source SS df MS F p (a) Abundance Year 0.35 1 0.35 2.29 0.131 Reef 0.39 1 0.39 2.59 0.109 Habitat 2.22 2 1.11 7.30 0.001 Depth 1.43 1 1.43 9.43 0.002 Year x Reef 0.20 1 0.20 1.34 0.249 Year x Habitat 1.25 2 0.62 4.10 0.018 Reef x Habitat 0.15 2 0.07 0.49 0.616 Year x Depth 0.20 1 0.20 1.30 0.255 Reef x Depth 0.17 1 0.17 1.11 0.294 Habitat x Depth 1.06 2 0.53 3.47 0.033 Year x Reef x Habitat 0.51 2 0.26 1.68 0.188 Year x Reef x Depth 0.02 1 0.02 0.11 0.739 Year x Habitat x Depth 0.36 2 0.18 1.17 0.313 Reef x Habitat x Depth 0.22 2 0.11 0.73 0.486 Y x R x H x D 0.79 2 0.39 2.59 0.078 Error 34.05 224 0.15

(b) Biomass Year 0.34 1 0.34 1.38 0.241 Reef 0.49 1 0.49 1.96 0.163 Habitat 12.27 2 6.14 24.69 0.000 Depth 3.75 1 3.75 15.07 0.000 Year x Reef 1.38 1 1.38 5.57 0.019 Year x Habitat 2.11 2 1.05 4.24 0.016 Reef x Habitat 5.69 2 2.85 11.46 0.000 Year x Depth 0.45 1 0.45 1.81 0.180 Reef x Depth 1.11 1 1.11 4.45 0.036 Habitat x Depth 0.75 2 0.37 1.50 0.224 Year x Reef x Habitat 1.38 2 0.69 2.78 0.064 Year x Reef x Depth 0.40 1 0.40 1.61 0.205 Year x Habitat x Depth 0.95 2 0.47 1.90 0.152 Reef x Habitat x Depth 0.01 2 0.01 0.02 0.978 Y x R x H x D 0.71 2 0.36 1.43 0.241 Error 55.65 224 0.25

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Figure 12. Temporal variation (2011 versus 2014) in the mean abundance of roving herbivorous fishes across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. (a) Total herbivorous fish biomass, (b) Grazing fish biomass, (c) Browsing fish biomass. Error bars represent 1 SE.

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Figure 13. Temporal variation (2011 versus 2014) in the mean biomass of roving herbivorous fishes across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. (a) Total herbivorous fish biomass, (b) Grazing fish biomass, (c) Browsing fish biomass. Error bars represent 1 SE.

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A principal component analysis (PCA) revealed strong spatial structuring, but limited temporal change, in the taxonomic composition of the herbivorous fish assemblages on Elizabeth and Middleton (Figure 14). The first two axes of the PCA explained 41 % of the total variation in composition of herbivore assemblages, with the reef front crest sites clearly separated from all reef slope and lagoonal sites along the first principal component. The reef front crest sites were characterised by a high biomass of macroalgal browsing fishes (Kyphosus spp and Prionurus maculatus), and large excavating (Chlorurus frontalis, C. microrhinos) and scraping parrotfishes (Scarus altipinnis). In contrast, the lagoonal sites were characterised by a diversity of smaller scraping parrotfishes (Scarus chameleon, S. globiceps, S. psitticus; Figure 14).

Numerous studies have reported negative correlations between the cover of macroalgae and the biomass of herbivorous fishes across a range of scales (e.g., Williams et al. 2001; Wismer et al. 2009). Such relationships are supported by studies that have shown the exclusion of herbivores leads to a proliferation of algal biomass (Hughes et al. 2007). Contrary to expectations there was no relationship between the biomass of roving herbivorous fishes and either the cover of macroalgae or the cover of live coral (Figure 15). There was, however, a significant positive correlation between herbivore biomass and the cover of bare space (i.e., the sum of algal turfs and crustose coralline algae) across Elizabeth and Middleton reefs (Figure 15c). The lack of relationship with either macroalgae or live coral is unexpected, but does not mean that herbivorous fishes are unimportant in the ecology and resilience of coral reefs at Elizabeth and Middleton Reefs. The positive relationship between bare space and herbivore biomass indicates that these fishes are playing some role in shaping benthic communities, but further work is clearly required to understand the ecological functioning of these high latitude reefs, especially what are the determinants of macroalgal abundance and biomass (Choat et al 2006, Hoey et al. 2011; Pratchett et al. 2011b).

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Figure 14. Principal component analysis (PCA) showing variation in the composition of roving herbivore assemblages between years, habitats and depths. Symbols indicate sampling year; triangles - 2011, circles - 2014. Colours indicate habitat; dark green – lagoon slope, light green – lagoon crest, dark blue – back reef slope, light blue – back reef crest, bright yellow – exposed reef crest, dull yellow – exposed reef slope. Analysis was based on the covariance matrix of the mean biomass (log-transformed) of each species at each site.

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Figure 15. Relationships between the herbivore biomass and the cover of (a) macroalgae, (b) live coral, and (c) bare space (the sum of algal turfs and CCA). Each point represents the mean of 4 transects at each depth at each of 16 sites. triangles: Elizabeth Reef, circles: Middleton Reef. Colours represent the three habitats shown in Figure 1. Yellow: exposed reef front; Blue: back reef; Green: lagoon.

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5.3.2 Endemic and site-associated fishes

The endemic anemone fish Amphiprion mccullochi is a relatively minor component of the total pomacentrid community on Elizabeth and Middleton Reefs, with an average density of 0.25 fish/100m2 (Figure 16). In contrast, the endemic butterflyfish Chaetodon tricinctus was the most abundant of 19 butterflyfish species recorded on these reefs with a mean density of 1.47 fish/100m2 (Figure 16).

A. mccullochii is a habitat specialist, being only found in close association with the anemone Entacmea quadricolor. Consequently the distribution of the A. mccullochi is restricted by the availability of its host anemone. At Elizabeth and Middleton Reefs this anemone is restricted to lagoonal areas and the back reef slope, and is reflected in the distribution on A. mccullochi (Figure 17a). The low and variable abundance of A. mccullochi precluded any meaningful analyses. In contrast, the three-striped butterflyfish C. tricinctus was more widely distributed among habitats, being record across all habitats on both reefs except the lagoon reef slope on Middleton reef (Figure 17b). Although there were no clear among-habitat patterns in the distribution of C. tricinctus, the abundance of this species was positively related to live coral cover (Pearsons correlation r = 0.477, p = 0.007) across these reefs. It is likely, therefore, that populations of this species will be susceptible to changes in live coral cover.

The overall abundances of both A. mccullochi and C. tricinctus tended to increase from 2011 to 2014 (Figure 18). The abundance of A. mccullochi was relatively stable on Elizabeth Reef, but showed a 10-fold increase on Middleton Reef over the past 3-years. This increase could possibly be related to an increase in the abundance of the host anemone or greater densities of A. mccullochi with each anemone, however, the increase should be taken in context as it only related to an additional 15 individuals being recorded in 2014 compared to 2011. Chaetodon tricinctus displayed relatively consistent (2.7- to 3.1-fold) increases in abundance across both reefs (ANOVA Year: F1,224 = 8.87, p = 0.003). The largest increases in abundance were recorded on the exposed reef fronts, while there was no significant change within the lagoons (Figure 18b). Interestingly, the changes in abundance of C. tricinctus did not track the changes in live coral cover over the same period (see Section 4.1.3 above).

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Figure 16. Relative abundances of members of the (a) Pomacentridae (damselfishes), and (b) Chaetodontidae (butterflyfishes) recorded across all transects on Elizabeth and Middleton Reefs in March 2014. Arrows indicate the relative abundances of the endemic species Amphiprion mccullochi and Chaetodon tricinctus. Data are means ± SE.

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Figure 17. Spatial variation in the abundance of (a) Amphiprion mccullochi, and (b) Chaetodon tricinctus among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE.

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Figure 18. Temporal variation (2011 versus 2014) in the mean abundance of (a) Amphiprion mccullochi, and (b) Chaetodon tricinctus across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE.

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The endemic double-header wrasse Coris bulbifrons was relatively abundant across both Elizabeth and Middleton Reefs in 2014, with 232 adults and 14 juveniles (<10cm Total Length) being recorded across all transects. Juvenile C. bulbifrons were largely restricted to the back reef and lagoon habitats, with only a single individual being recorded on the exposed reef front (Figure 19a). Adult C. bulbifrons were recorded across all habitats and depths at both Elizabeth and Middleton Reefs, reaching their highest densities within the lagoon at both reefs (Figure 19b).

Figure 19. Spatial variation in the abundance of (a) juvenile, and (b) adult Coris bulbifrons among reefs (Elizabeth and Middleton), habitat zones (Back reef, lagoon, reef front), and depths (crest, slope) in March 2014. Data are means ± SE.

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There were marked increases in the mean densities of both juvenile and adult C. bulbifrons over the past 3-years. The overall densities of juvenile C. bulbifons increased 3.2-fold, with the greatest increase being recorded on Elizabeth Reef (Figure 20a). Similarly, the densities of adult C. bulbifons increased 2.5-fold, with a 3.8-fold increase on Elizabeth compared to a 1.9-fold increase on Middleton Reef (Figure 20b). The increases in adult C. bulbifrons were largely attributed to marked increases within the lagoon at both reefs.

Figure 20. Temporal variation (2011 versus 2014) in the mean abundance of (a) juvenile, and (b) adult Coris bulbifrons across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE.

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5.3.3 Apex predators

Very high densities of Galapagos reef sharks (Carcharhinus galapagensis) were recorded at Elizabeth and Middleton Reefs in 2014. The average abundance of C. galapagensis across all habitats was 12.75 individuals per hectare (± 2.08 SE), compared to 6.81 individuals per hectare (± 1.68 SE) in 2011 (Table 13, Figure 21a). Overall densities of Galapagos reef sharks were significantly greater at Middleton (17.11 individuals/ha ± 2.99 SE) than at Elizabeth (6.58 individuals/ha ± 1.69 SE), though there has been a substantial increase in densities of C. galapagensis at Elizabeth Reef between 2011 and 2014 (Figure 21a). The most pronounced increases were recorded on the back reef, while moderate increases were recorded in the lagoon, and limited change on the exposed reef front on both reefs (Figure 21a). The majority of C. galapagensis recorded were relatively small (< 120cm; Figure 22) with the largest individual recorded being 180cm. There was very little difference in the average size of C. galapagenis on Elizabeth (94.8cm) and Middleton Reefs (97.1cm). Given that C. galapagensis reaches up to 300 cm in length, does not mature until 210-250 cm, and is born at 60-80 cm (Last and Stevens 1994) indicates that the large populations at Elizabeth and Middleton consist of immature individuals, and that these reefs are likely to represent an important nursery area for this species.

Table 13 ANOVA comparing the abundance of the Galapagos reef shark Carcharhinus galapagensis) between years (2011 vs 2014), reefs (Elizabeth and Middleton), and habitats (reef front, lagoon, back reef). Abundances were log(x +1) transformed.

Source SS df MS F p Year 2.45 1 2.45 23.36 0.000 Reef 2.40 1 2.40 22.93 0.000 Habitat 0.63 2 0.31 3.00 0.059 Year x Reef 0.06 1 0.06 0.60 0.442 Year x Habitat 0.91 2 0.46 4.35 0.018 Reef x Habitat 0.40 2 0.20 1.91 0.159 Year x Reef x Habitat 0.02 2 0.01 0.09 0.916 Error 5.14 49 0.10

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Figure 21. Temporal variation (2011 versus 2014) in the mean abundance of (a) the Galapagos Reef Shark and (b) the Black Cod across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE.

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Figure 22. Comparison of the size structure of Galapagos Shark (Carcharhinus galapagensis) populations between Elizabeth and Middleton Reefs. Size at birth is 60- 80cm, size at maturity is 210-230cm (males) and 250cm (females).

Black cod (Epinephelus daemelii) were recorded across all sites at Elizabeth and Middleton Reefs. The average abundance of E. daemelii across all habitats in March 2014 was 2.34 individuals per hectare (± 0.35 SE), which represented a small, but non-significant, increase from the densities recorded in March 2011 (1.88 individuals per hectare ± 0.25 SE; Figure 21b). As shown in 2011, densities of E. daemelii were generally higher at Elizabeth Reef than at Middleton Reef, though these differences were not significant (Table 14).

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Table 14: ANOVA comparing the abundance of the black cod (Epinephelus daemelii) between years (2011 vs 2014), reefs (Elizabeth and Middleton), and habitats (reef front, lagoon, back reef).

Source SS df MS F p Year 2.97 1 2.97 1.01 0.320 Reef 1.33 1 1.33 0.45 0.505 Habitat 0.68 2 0.34 0.12 0.890 Year x Reef 0.70 1 0.70 0.24 0.628 Year x Habitat 6.12 2 3.06 1.04 0.361 Reef x Habitat 2.91 2 1.45 0.49 0.613 Year x Reef x Habitat 6.57 2 3.28 1.12 0.336 Error 144.22 49 2.94

5.3.4 New species records

A total of 270 vertebrate species were recorded at Elizabeth and Middleton Reefs in 2014, including 10 species not previously recorded. These new records increase the total number of reef fishes recorded at Elizabeth and Middleton Reefs to 356 species (Appendix 1). The recent occurrence of tropical reef fishes at these high latitude reefs may reflect the poleward expansion of species ranges due to changing environmental conditions associated with climate change (Feary et al. 2013), but it is also possible that these species are so rare that they have not been encountered during previous surveys. Ongoing monitoring will be required to assess whether these species are increasing in abundance and establishing viable populations at Elizabeth and Middleton (through increased overwinter survival of tropical fishes transported to sub-tropical latitudes by ocean currents; Figueira et al. 2009), or whether they simply reflect stochastic transport and recruitment of individuals from tropical reefs.

The number of fish species recorded and the number of new species is lower than those recorded in 2006 and 2011, however, it is important to recognise that compilation of this species list was given low priority compared to rigorous area- based surveys of specific study species in 2014. Consequently, there was less opportunity for the detection of rare species given that there was no one dedicated to populating this list in 2014. In 2011, by comparison, Dr T. Ayling was specifically tasked with surveying across multiple habitat types to compile a comprehensive list

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of resident fishes, while also conducting timed swims to survey apex predators. The 2014 survey was also shorter and involved fewer divers than the 2011 survey.

5.4 Unitary invertebrates

5.4.1 Sea cucumbers (Holothurians)

A total of 12 species of holothurians have been recorded at Elizabeth and Middleton Reefs. A total of 628 sea cucumbers were recorded across all transects in March 2014. The most common species of sea cucumber recorded in 2014 was Holothuria atra, which accounted for 38.2% (240/628) of individuals (Figure 23). Other common species were Holothuria edulis (25.5%) and Holothuria (Microthele) whitmaei (22.3%). Collectively these three species accounted for >85% of all individuals surveyed. The composition was broadly similar to those recorded in 2011; H. atra (37.3%), H. edulis (32.8%), and H. whitmaei (14.2%).

Overall densities of sea cucumbers (for all species combined) were higher in 2014 (5.07 sea cucumbers per 200m2) than were recorded in 2011 (3.29 sea cucumbers per 200m2), with a marked increase being recorded at Elizabeth Reef (Figure 24). Similar densities were recorded on Elizabeth and Middleton Reefs in 2014 (4.6 – 5.4 sea cucumbers per 200m2), which is in contrast to 2011 when densities on Middleton Reef were approximately double that of Elizabeth Reef. The highest densities were again recorded in the lagoon (Elizabeth: 14.88 individuals/200m2; Middleton: 19.67 individuals/200m2) and to a lesser extent the back reef (Elizabeth: 1.00 individuals/200m2; Middleton: 7.94 individuals/200m2), coinciding with the availability of sand, as opposed to consolidated reef matrix. It should be noted that these surveys were designed to assess consolidated coral reef habitats, as opposed to soft-sediment habitats. As such are likely to have underestimated the reef-wide densities of organisms, such as sea cucumbers, that typically associated with soft sediments.

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Figure 23. Comparison of the species composition of sea cucumbers populations between 2011 and 2014 on Elizabeth and Middleton Reefs. The three most abundant species accounted for approximately 85% of all individuals in both sampling years.

Figure 24. Temporal variation (2011 versus 2014) in the mean abundance of sea cucumbers (all species combined) across three habitat types (see Figure 1 and Table 1 for details) at Elizabeth and Middleton Reefs. Error bars represent 1 SE.

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5.4.2 Sea urchins (echinoids)

Sea urchins (echinoids) can have a major influence on the habitat structure of coral reef environments (e.g., McClanahan and Shafir 1990; Eakin 1996). Like herbivorous fishes, larger species such as Diadema spp. may be important in removing algae that would otherwise inhibit coral growth and/or settlement (Edmunds and Carpenter 2001). Indeed, phase shifts from coral- to macroalgal- dominance in many Caribbean locations were attributed in part to a population crash of sea urchins (Hughes 1994). At high densities, however, intensive grazing by sea urchins may have negative effects on habitat structure, causing significant loss of corals, reducing topographic complexity of reef habitats (McClanahan and Shafir 1990), can lead to a net erosion of the reef carbonates (Glynn et al. 1979; Eakin 1996). Unlike the Caribbean, and some parts of the Indian Ocean, sea urchins are relatively uncommon and have limited influence on coral reef ecology in the Pacific, especially in areas with limited fishing (Bellwood et al. 2003, 2012). Sea urchins are, however, typically more abundant on high latitude reefs, and outbreaks of sea urchins have been recorded in this region. For example, populations of the sea urchin Tripneustes gratilla underwent an outbreak at Lord Howe Island in 2007-08, causing localized declines in the abundance of many red and brown algae (Valentine and Edgar 2010).

A total of 8,858 sea urchins (mostly Echinometra mathaei, Diadema setosum, and Echinostrephus sp.) were counted across 124 transects at Elizabeth and Middleton Reefs in March 2014. The single most abundant species, accounting for 53.4% of all urchins counted (4,735/8,858) was Echinostrephus sp., which are small burrowing urchins with long thin spines (Figure 25). The impact of Echinostrephus on benthic reef assemblages is limited as they rarely, if ever, move outside their burrows to feed. The next most abundant species were Diadema setosum (38.7%) and Echinometra mathaei (7.4%), which are larger, mobile and nocturnally active species, and when in high densities can have a large influence on algal assemblages, benthic composition, and carbonate budgets.

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Overall, the densities of sea urchins was greater on Middleton (104.3 urchins/200m2) than on Elizabeth (31.5 urchins/200m2), and within each reefs were greatest on the exposed reef fronts and back reef habitats, and lowest within the lagoons (Figure 26). The densities of both E. mathaei and Echinostrephus sp decreased significantly from 2011 to 2014 (Figure 26a, c, Table 14). In marked contrast, the abundance of D. setosum increased significantly across both reefs, especially on Middleton Reef where densities on the back reef increased from an already high 78.8 urchins/200m2 in 2011 to 171.0 urchins/200m2 in 2014 (Figure 26b, Table 15).

Figure 25. Comparison of the species composition of sea urchin populations between 2011 and 2014 on Elizabeth and Middleton Reefs.

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Figure 26. Temporal variation (2011 versus 2014) in the mean abundance of the three most abundant sea urchins (a) Echinometra mathaei (b) Diadema setosum, and (c) Echinostrephus sp across three habitat types at Elizabeth and Middleton Reefs. Error bars represent 1 SE.

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Table 15: ANOVA comparing the abundance of the sea urchins Echinometra mathaei, Diadema setosum, and Echinostrephus sp. between years (2011 vs 2014), reefs (Elizabeth and Middleton), habitats (reef front, lagoon, back reef) and depths (crest, slope).

E. mathaei D. setosum Echinostephus sp Source df F p F p F p Year (Y) 1 5.18 0.024 9.76 0.002 1.59 0.209 Reef (R) 1 10.79 0.001 47.83 0.000 0.28 0.599 Habitat (H) 2 10.42 0.000 47.38 0.000 6.88 0.001 Depth (D) 1 0.58 0.447 35.35 0.000 6.51 0.011 Year x Reef 1 1.81 0.180 6.49 0.012 0.00 0.982 Year x Habitat 2 1.53 0.218 6.75 0.001 0.43 0.653 Reef x Habitat 2 4.17 0.017 42.38 0.000 7.00 0.001 Year x Depth 1 1.29 0.258 2.66 0.104 0.16 0.686 Reef x Depth 1 1.27 0.261 30.06 0.000 0.57 0.452 Habitat x Depth 2 5.65 0.004 31.29 0.000 1.78 0.171 Y x R x H 2 1.36 0.259 4.92 0.008 0.00 0.998 Y x R x D 1 2.40 0.123 2.59 0.109 4.05 0.045 Y x H x D 2 1.71 0.183 1.34 0.263 8.22 0.000 R x H x D 2 1.53 0.219 33.67 0.000 0.45 0.635 Y x R x H x D 2 0.37 0.691 1.93 0.147 2.54 0.081 Error 224

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6 Conclusions

Coral reefs are one of the world’s most diverse and productive ecosystems, yet they are also one of the most threatened. The effects of overfishing, pollution, sedimentation, eutrophication, habitat modification, and diseases are being greatly compounded by climate change, and are having a marked effect on the structure and function of coral reef ecosystems globally (Jackson et al. 2001, Pandolfi et al. 2003, Bellwood et al. 2004). Subtropical or high latitude reefs, such as Elizabeth and Middleton Reefs, lie on the latitudinal limit for coral reef growth (Kleypas et al. 1999), and support unique assemblages of both tropical and temperate flora and fauna (Hobbs et al. 2008, Hoey et al. 2011). While not exposed to the frequency or intensity of events affecting tropical reefs, subtropical reefs are nonetheless subject to a range of disturbances, including coral bleaching and disease, storms, and crown-of-thorns starfish outbreaks (e.g., Harriott 1995; Riegl 2003; Harrison et al. 2011). The susceptibility of subtropical reefs is likely to depend on their regenerative capacity following these relatively infrequent disturbances, rather than their ability to resist multiple extreme events. Quantifying spatial and temporal patterns of fish and benthic communities on these high latitude reefs is key to understanding the processes that structure these unique assemblages. Matching sampling locations and methods to those of the previous survey (i.e., 2011) at Elizabeth and Middleton reefs for the first time facilitated rigorous comparisons of temporal change in fish and benthic communities.

6.1 Habitat structure

Elizabeth and Middleton Reefs, although located at the latitudinal limits of coral growth, exhibit many characteristics that are representative of low latitude, tropical reef systems. Hard (scleractinian) corals are key habitat forming organisms on Elizabeth and Middleton Reefs and any changes to the abundance and composition of coral communities are likely to have flow-on effects that threaten the structure and functioning of the entire shallow water ecosystems. Average coral cover on Elizabeth and Middleton is currently low, compared to the southern Great

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Barrier Reef to the north, and Lord Howe Island to the south (Figure 27). Further, coral cover has declined, albeit slightly, since the previous survey in 2011. While the current levels of coral cover (Elizabeth: 29.8%; Middleton 19%) are still above those recorded at a limited number of sites on Middleton Reef in 1994 and 2006 (ca. 10%; Figure 4), the declines are a potential cause of concern and in contrast with the annual increases in coral cover of 6.7% recorded at a limited number of Middleton Reef sites from 2006 to 2011 (Pratchett et al. 2011b). Changes in coral cover were spatially variable with declines being pronounced within the lagoon at both reefs and the back reef at Middleton Reef, while coral cover remained stable or increased slightly at the remaining other locations (Figure 3). Within the lagoon sites large areas of arborescent Acropora thickets present in 2011 (Pratchett et al. 2011b) had been largely reduced to rubble by 2014. The causes of these declines are unclear but the consistency among reefs is indicative of a large-scale disturbance, such as a storm or elevated sea surface temperature.

Perhaps the sites of greatest concern are those on the back reef of Middleton Reef. These sites experienced a 50% decline in coral cover (15.7 to 8.3%), coupled with a 2-fold increase in macroalgal (primarily Codium) cover (15.4 to 32.7%) and a 2-fold increase in the densities of the sea urchin Diadema setosum (78.8 to 171 urchins / 200m2). These are characteristic signs of a stressed reef system (Hughes 1994; Bellwood et al. 2004). The high cover of macroalgae is likely to further reduce the capacity for these sites to recover through reduced settlement and survival of coral larvae (Hughes et al. 2007; Birrell et al. 2008; Mumby and Steneck 2008), and reduced growth and survival of existing corals (Hughes et al 2007; Rasher et al. 2011). Moreover, the high densities of Diadema and the low cover of live coral suggest that these sites may have a negative carbon budget whereby erosion of carbonates by Diadema exceeds the production of carbonates by live corals and other calcifying organisms, as has occurred on several Eastern Pacific reefs (Glynn et al. 1979, Eakin 1996). Again, the causes of the decrease in coral cover and increase in macroalgal cover is unclear. There are a wide range of physical and biological mechanisms that may influence local abundance and species composition of corals and macroalgae in coral reef environments, including herbivory, eutrophication, hydrodynamics and sedimentation (e.g., Payri 1984; Smith et al. 2001; Rasher et al. 2013). The back

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reef sites on Middleton are located either side of the main channel to the lagoon. It is possible that these sites are receiving high sediment and/or nutrient loads during the tidally driven draining of lagoonal waters. A better understanding of the processes that underlie the local distribution and abundance of macroalgae (and therefore, hard corals) on Eliazabeth and Middleton Reefs is necessary for sustainable management of these reefs. This will however, require intensive research and careful experimentation to complement regular monitoring at intervals of 2-3 years.

Figure 27. Geographic variation in mean (± SE) coral cover between Lord Howe Island, Elizabeth Reef, Middleton Reef, and the southern GBR. Data for Lord Howe Island is for 2010 (Hoey et al. 2011), and data for the southern GBR is for 2005 (Emslie et al. 2008).

The low coral cover on Elizabeth and Middleton Reefs compared to those to the north and south (Figure 27) suggest that these reefs have not recovered from previous disturbance/s. Estimates of baseline coral cover are, however, critical in assessing the current state of these reefs. Despite ten surveys to Elizabeth and Middleton Reefs since 1981 the available data on baseline coral cover is limited. Early qualitative estimates at several sites around Elizabeth and Middleton suggested coral cover was ‘high’ in 1981 (Veron and Done unpublished data in Australian Museum 1992). While it is difficult to resolve exactly what coral cover

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was during these early surveys, photographs taken in 1987 (Australian Museum 1992) suggest that it was considerably higher than recorded in recent surveys (2006-2014). The declines in coral cover since 1981 corresponded with high densities of Crown-of-Thorns Starfish in the 1980’s and early 1990’s with >400 individuals being recorded at some sites (Australian Museum 1992; Harriot 1998). While only 4 Crown-of-Thorns Starfish were recorded in the current survey, the threat posed by future outbreaks should not be discounted.

The ability to manage of outbreaks of A. planci will largely depend upon identification of proximal causes of outbreaks at each specific location. On the Great Barrier Reef there have been suggestions that the outbreaks are linked to anthropogenic activities, notably the input of land-based nutrients (Fabricius et al. 2010), and the reduction of predatory fishes through fishing (Sweatman 2008; McCook et al. 2010), however neither of these mechanisms appear particularly relevant to Elizabeth and Middleton Reefs. Previous outbreaks have occurred despite relatively intact assemblages of large predatory fishes (see discussion below), and no input of land-based nutrients.

Although the current coral communities appeared healthy (i.e., low incidence of bleaching, disease, and predation) the threats to coral assemblages at Elizabeth and Middleton reefs (coral bleaching, disease, A. planci outbreaks, storms) are magnified by the limited capacity for recovery in the aftermath of such disturbances. Densities of juvenile corals at Elizabeth and Middleton Reefs (mean = 11.26 juveniles per 10m2) were markedly lower than estimates from low latitude reefs of the southern GBR and Moorea, French Polynesia, but similar to those from Lord Howe Island (Figure 28). These low rates of replenishment appear to be characteristic of high latitude reefs (Harriott 1992; Hoey et al. 2011; Bauman et al. 2014), and coupled with the lower coral growth rates on subtropical reefs (Harriott 1999; Anderson et al. 2012) and isolation (Gilmour et al. 2013) are likely to limit the recovery potential of these reefs following disturbance. Collectively these factors are likely to have contributed to the protracted recovery of coral communities following the crown-of-thorns starfish outbreak in the late 1980’s and early 1990’s.

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Figure 28. Geographic variation in mean (± SE) abundance of juvenile corals between three high latitude reefs; Lord Howe Island, Elizabeth Reef, Middleton Reef, and two low latitude reefs; the southern GBR and Moorea, French Polynesia. Data sources are Lord Howe Island (Hoey et al. 2011), southern GBR (Trapon et al. 2013), and Moorea (Pratchett et al 2011).

6.2 Reef fishes

The removal of key functional groups of fishes through fishing activities has underpinned the degradation of coral reef systems worldwide (e.g., Hughes 1994; Rasher et al. 2013). Reductions in predatory fishes have been related to increased incidence of A. planci outbreaks in the tropical Pacific (Sweatman 2008; Dulvy et al. 2004), and reductions in herbivorous fishes have underpinned to shifts from coral- to macroalgal-dominated reefs in many regions (Hughes 1994; Rasher et al. 2013). For the most part, coral reef fish communities appeared to have changed little over the past 3 years with densities being general similar in 2014 compared to those recorded across the same sites in 2011 (Pratchett et al. 2011b). The major exceptions to this were the Galapagos shark Carcharhinus galapagensis, which increased from 2011 to 2014, and small (mainly juvenile) parrotfish that declined in abundance during this period.

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Densities of C. galapagensis increased 2-fold from 6.8 individuals per hectare in 2011 to 12.8 individuals per hectare in 2014, with in excess of 50 individuals being observed on a single dive. This was on top of the 16-fold increase reported from 2006 to 2011 (Pratchett et al. 2011b). These apparent increases in the densities of C. galapagensis are remarkable, and indicative of the low level of fishing due the geographic isolation and protection status of these reefs. Apex predators are highly susceptible to fishing, and usually the first to be removed from impacted reef systems (Friedlander and DeMartini 2002, Sandin et al. 2008, Ceccarelli et al. 2013). The densities of sharks recorded during the present survey are greater than those recorded for other relatively intact and isolated systems, such as the Chagos Archipelago (Graham and McClanahan 2013) and the Line Islands (Sandin et al. 2008). It is worth noting, however, that all C. galapagensis recorded at Elizabeth and Middleton Reefs were small (<180cm) immature individuals, with several the size of newly born (60-80cm, Last and Stevens 1992) individuals. The high densities and small size of C. galapagensis suggest that Elizabeth and Middleton reefs are an important nursery habitat for this species. If sustained, however, such increases in abundance may impact lower trophic levels through both direct (i.e., predation) and indirect (i.e., behavioural) effects. Importantly, several recent studies have shown that increased densities of apex predators influencing foraging decisions by herbivorous reef fish, altering the intensity and distribution of their feeding impact (Madin et al. 2010; Rizzari et al. 2014). Understanding the direct and indirect effects of the increased density of C. galapagensis will require both dedicated research and continued monitoring at regular intervals.

The black cod (E. daemelii) has been extensively fished and experienced population declines throughout much of its geographic range. The overall abundance of black cod (across all sites) averaged 2.3 individuals per hectare, representing a small, but non-significant, increase from the densities recorded at the same sites in 2011 (Pratchett et al. 2011b) and is within the range of densities recorded by Oxley et al. (2004) and Choat et al (2006). The absence of a decline is a positive but not a cause for complacency. The smallest individual observed during the present surveys was 50cm in length, suggesting that recruitment to the population may be limited. Any future declines, no matter how slight, could threaten the sustainability of this population. Extensive long-term monitoring is critical to

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assess the fate of this species, but more intensive research is also required to assess the natural dynamics of local populations (e.g., recruitment rates, recruitment/nursery habitats, growth and longevity), which will be critical for assessing their overall vulnerability.

The structure and resilience of coral reef ecosystems are fundamentally dependent on the actions of herbivores, with herbivorous fishes playing a predominant role. On reefs with relatively intact fish assemblages, herbivores rapidly graze the reef substratum removing up to 90 percent of the net algal production (Polunin and Klumpp 1992; Ferreira et al. 1998). Such intense herbivory inhibits the proliferation of algal biomass and minimizes algal-coral competition (Hughes et al. 2007; Rasher and Hay 2010) and facilitates coral settlement through the provision of suitable settlement substrata (Hughes et al. 2007; Hoey et al. 2011), thereby promoting the recovery of coral assemblages following disturbance (Bellwood et al. 2004; Mumby and Steneck 2008).

The herbivorous fish assemblages of Elizabeth and Middleton Reefs represent a unique mix of both tropical and temperate species, and appear to have changed little over the past 3 years (Pratchett et al. 2011b). Collectively, the densities of all herbivorous fishes on Elizabeth and Middleton Reefs were broadly comparable to lower latitude reefs of the central and northern Great Barrier Reef and substantially greater than those of Lord Howe Island, 260 km to the south (Figure 28). There were, however, marked differences in the functional composition of herbivorous fish assemblages between reefs. Elizabeth and Middleton Reefs were dominated by macroalgal browsing species, in particular Kyphosus pacificus macroalgal browsing species typically account for < 10% of the herbivore community (Hoey and Bellwood 2010a, b). Studies have shown that these fishes play a critical role in removing macroalgal biomass and potentially reversing phase-shifts on low latitude reefs (Hoey and Bellwood 2011; Rasher et al. 2013), however their role limiting the cover and biomass of macroalgae on higher latitude reefs. For example, the two browsing species Kyphosus pacificus and Prionurus maculatus accounted for >70 % of the total herbivorous fish biomass on Elizabeth and Middleton reefs, but we currently know very little of their biology and ecology. The role these fishes play in limiting the abundance and biomass of macroalgae on high latitude reefs is

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currently unknown and is an area that requires specific targeted research if we are to understand the process that structure these unique reef systems.

Figure 29. Geographic variation in mean (± SE) biomass of roving herbivorous fishes between Lord Howe Island, Elizabeth Reef, Middleton Reef, and the central and northern GBR. The relative contribution of grazing and browsing fishes is shown. Data sources are Lord Howe Island (Hoey et al. 2011), central and northern GBR (Wismer et al. 2009).

Together with the high biomass of browsing fishes, the biomass of grazing fishes on Elizabeth and Middleton Reefs was approximately 1/4 of that on low latitude reefs (Figure 28). Grazing fishes, and in particular scraping and excavating parrotfishes, perform a key role in preventing the establishment and proliferation of macroalgae on tropical reefs (Bellwood et al. 2003; Mumby 2006; Hoey and Bellwood 2008). When present in sufficient densities their intensive feeding not only maintain algal communities in a cropped state, but also provide clean areas of substrata for colonisation by benthic organisms, including corals. While the cover of bare space (primarily turf algae) was negatively related to total herbivore biomass on Elizabeth and Middleton reefs, there was no relationship between the

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density of juvenile corals and the biomass grazing fishes, or all herbivorous fishes collectively. Focused and intensive research will be required to disentangle the complexities of the relationships between herbivorous fishes, macroalgae and the replenishment of corals on these reefs.

Although there were limited changes in biomass and composition of herbivorous fish assemblages over the past 3 years, the reductions in the density of juvenile parrotfish within the lagoons at both reefs is a potential concern. Many parrotfish species, and reef fishes in general, recruit to complex, sheltered habitats as juveniles (e.g., McCormick and Makey 1997; Hoey and Bellwood 2008) and undergo ontogenetic migrations to adult habitats as they grow (e.g., Olds et al. 2014). While the impacts of these reductions are not readily apparent in terms of fish biomass, and thereby ecosystem function, they are likely to have considerable impact on structure and function of future populations of grazing fishes on Elizabeth and Middleton Reefs.

The reduction in juvenile parrotfish coincided with the reduction in arborescent corals at these sites, and is likely a response to the loss of physical structure these provide. Coral loss and associated changes in biological and physical structure of coral reef habitats have an important influence on the abundance and diversity of coral reef fishes (Pratchett et al. 2008). While the loss of live coral cover generally leads to declines in the abundance of fishes that rely on those corals for food and/or recruitment (Wilson et al. 2006, Coker et al. 2012), declines in the structural complexity generally effects a wider range of species and functional groups through reductions in the availability of spatial refuges from predation (Pratchett et al. 2011a). Extensive long-term monitoring is critical to assess the impacts of the loss of juvenile habitat on the dynamics of parrotfish populations on these reefs.

6.3 Unitary invertebrates

Unitary, or mobile, invertebrates contribute significantly to the biodiversity of coral reefs (Stella et al. 2011). Despite the diversity of unitary invertebrates, most research has focused on a limited suite of species, namely echinoderms of commercial value (i.e., sea cucumbers), or ecological importance (i.e., urchins and

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crown-of-thorns starfish). Sea cucumbers (Holuthurians) are exploited throughout their range, primarily as a luxury seafood for Asian markets (Purcell et al 2014). The high market value of these species has led their overexploitation throughout much of the Pacific (Purcell 2014, Purcell et al. 2014). The isolation and protection status of Elizabeth and Middleton Reefs appear to be having positive effects on sea cucumber populations with the overall abundance of sea cucumbers increasing 50% since 2011 (Pratchett et al. 2011b). This included the high-valued and sought after H. whitmaei (listed as endangered on IUCN redlist). The increase in abundance of these species is encouraging, however it is important to note that the recent surveys (2006-2014) have been specifically designed to assess fish and benthic communities within consolidated reef, or coral-dominated, habitats. Due to time constraints there has been minimal focus on areas of unconsolidated sediments (i.e., sand) and as such these surveys are likely to have grossly underestimated the reef-wide abundances of sea cucumbers and other organisms that associate with sand. An accurate assessment of sea cucumber populations on these reefs would require dedicated sampling of shallow reef flat areas and sand- dominated habitats.

The other major group of invertebrates that have received considerable attention are herbivorous sea urchins (e.g., McClanahan and Shafir 1990). Large herbivorous urchins (e.g., Diadema) often increase in densities following the removal of their fish predators, and have been suggested to supplement or replace the ecological function of herbivorous fishes (e.g., Edmunds and Carpenter 2001). However, urchins feed in a fundamentally different way to that of reef fishes. Urchins not only remove algal material from the reef surface when feeding but also burrow into and erode the reef matrix (e.g., Eakin 1996). At high densities they can destabilize the integrity of the reef framework and have the capacity to destroy reefs (Glynn et al. 1979, Eakin 1996). Although often viewed as a healthy state for Caribbean reefs, high densities of urchins on Indo-Pacific reefs is often viewed as a sign of degradation (Bellwood et al. 2004). Sea urchin populations on Elizabeth and Middleton reefs were relatively stable from 2011 to 2014, with one notable exception, a 2-fold increase in the density of Diadema setosum within the back reef at Middleton. Densities of D. setosum were already elevated at this location in 2011, and are currently ca. 20 times greater than other locations across both reefs.

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Echinoids, and echinoderms in generally, are renowned for exhibiting large variations in population densities (Uthicke et al. 2009). The cause of the current high densities of D. setosum at Middleton reef are unknown but will require continued monitoring and focused research to determine the cause and consequences of these urchins on reef structure and function.

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7 Acknowledgements

This research was made possible through the excellent logistical support and assistance provided by the crew of the M.V. Capricorn Star, in particular Scott and Soozi Wilson. Sally-Ann Gudge and Tas Douglass of the Lord Howe Island Marine Park Authority also provided significant assistance and support. We also gratefully acknowledge the guidance and contributions of Kate Dunkerley, which greatly improved the application of this research to the ongoing management of Elizabeth and Middleton Reefs.

This research was commissioned by the Department of the Environment, but the Australian Centre of Excellence for Coral Reef Studies also provided significant financial support.

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Appendix 1. Check list of fishes and other vertebrates recorded at Elizabeth and Middleton Reefs, including Australian Museum List (AM) and results from surveys conducted since 1994. For surveys conducted in 2006, 2011 and 2014 a rough estimate of abundance is given using four categories: Abundant (A) – >500 individuals recorded; Common (C) - >100 individuals recorded; Occasional (O) - 5- 100 individuals recorded; Rare (R) –<5 individuals recorded.

Family Species AM 1994 2003 2006 2011 2014 TOTAL FISH SPECIES: 301 237 179 322 303 270

FISHES:

Triakidae Galeorhinus galeus X

Carcharhinus Carcharhinidae X X X C C galapagensis C C. amblyrhynchos X X

Galeocerdo cuvier X R

Sphyrnidae Sphyrna lewini R

Dasyatididae Dasyatis thetidis X X R R D. brevicauda R

Taeniura mayeni X O O O

Muraenidae Echidna nebulosa R R

Enchelycore ramosa X X X R R

Gymnothorax prasinus X X O

G. eurostus X X O O O G. meleagris X O O G. chilospilus X

G. annasona X

G. porphyreus X

Siderea thrysoidea X O R O Anarchias sp. X

Ophicthyidae Myrichthys colubrinus R

M. maculosus X R R

Muraenichthys X laticaudata Pseudomyrophis sp. X

Plotosidae Plotosus lineatus X O O O

Gobiesocidae Creocele cardinalis X

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Lepadichthys frenatus X

Antennariidae Antennarius coccineus X

Moridae Lotella phycis X

Ophidiidae Brotula multibarbata X

Dinematichthys spp. X

Synodontidae Synodus variegatus X X X C C C S. dermatogenys X X O O O S. cf. binotatus X

Euleptorhamphus Hemirhampidae X X X C O viridis O

Exocoetidae Cheilopogon furcatus X X X C C C

Belonidae Platybelone argala X R O O

Atherinidae Hypoatherina tropicalis X C O O

Holocentridae Myripristes murdjan X X O O O M. kuntee X X

Neoniphon sammara R R Sargocentron diadema. X O O O Plectrypops lima X

Aulostomidae Aulostomus chinensis X X X C C O

Fistulariidae Fistularia commersonii X X X C C C

Syngnathidae Halicampus sp.? R

Cosmocampus X howensis Halicampus brocki X

Caracanthidae Caracanthus unipinna O O O C. maculatus X O

Scorpaenidae Pterois volitans X X X O O O P. antennata X R

Dendrochirus zebra X X R R R Scorpaenopsis diabolus X R R R Scorpaenopsis spp. X

Scorpaena cookii X X R

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Scorpaena sp. X R

Sebastapistes sp. O

Ablabys taeianotus X

Scorpaenodes spp. X

Synanceia verrucosa X R

Serranidae Pseudanthias pictilis X X O O O P. squamipinnis X X X O O O P. tuka R

Cephalopolis argus X X X C C O C. cyanostigma X

C. miniata X X X O O O C. urodeta R

Hypoplectrodes sp. X

Epinephelus fasciatus X X X C O R E. areolatus X

E. merra X X X O O R E. daemelii X X X C C C E. cyanopodus X X O O O E. fuscoguttatus R R R E. maculatus X R R R E. rivulatus X X R

E. tauvina X R

E. polyphekadion X X R O R E. lanceolatus X

E. quoyanus X

Variola louti X X X O C C Plectropomus laevis X R R R P. leopardus X R R R Trachypoma X R R macracanthus R Acanthistius cinctus X X X R R

A. ocellatus X

Grammistes O R sexlineatus R Liopropoma susumi X

Pseudochromid Pseudochromis fuscus O O ae O P. novaehllandiae X X O

Pseudoplesiops X howensis Cypho purpurascens X

Plesiopidae Plesiops sp. X

Belonepterygion Acanthoclinidae X fasciolatum

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Cheilodactylidae Cheilodactylus francisi X X X C C O C. vestitus O O C. ephippium X X X O O O

Cirrhitidae Paracirrhites arcatus X X X O O O P. forsteri X X X O O O Cirrhitus splendens X X X O C C C. pinnulatus R

Cirrhitichthys falco X X X A C C

Apogonidae Apogon aureus O O O A. cf.aureus O

Apogon sp. Large X O barred O A. coccineus X

A. compressus X

A. doederleini X X X R O A. cyanosoma X

A. nigrofasciatus X

A. norfolcensis X X R

A. capricornis O Cheilodipterus X X X R quinquelineata O C. macrodon O O R Fowleria aurita X

Rhabdamia sp. X

Vincentia chrysusus X

Priacanthidae Priacanthus cruentatus X R R R P. macracanthus X

Malacanthus Malacanthidae X O O brevirostris O

Echeneididae Echeneis naucrates X X X O R R

Carangidae Caranx melampygus. X X R

C. lugubris X X X O O R C. ignobilis X

Carangoides X R orthogrammus C. fulvoguttatus R

C. ferdau R R Pseudocaranx dentex X X X C C C Elagatis bipinnulata X X X R O R Seriola lalandi X X X O O C

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S. rivoliana X X X O R O

Arripidae Arripis trutta X

Lutjanidae Lutjanus bohar X X X C C C L. kasmira X O O O Aphareus furca X O R

Aprion virescens X X X R O O Macolor niger R

Paracaesio xanthura X X X A A A

Haemulidae Plectorhynchus picus X X X C C C

Nemipteridae Scolopsis bilineatus X X R O R

Lethrinidae Lethrinus nebulosus X R

Gnathodentex O O aurolineatus O Monotaxis grandoculis X O O R Gymnocranius euanus X X X O O O Gymnocranius sp. X X O O

Mulloidichthys Mullidae X O O vanicolensis O M. flavolineatus O O O Parupeneus X X O O cyclostomus R P. multifaciatus X X X C C O P. pleurostigma X X X C O O P. signatus X X X A A C

Pempheridae Pempheris sp. X O O O

Kyphosidae Kyphosis pacificus X X X A A A K. cinerascens O O K. sydneyanus X O O O K. vaigiensis R

Girellidae Girella cyanea X X X O O O

Scorpididae Scorpis violaceus X X O R Bathystethus cultratus O R Labracoglossa nitida A

Microcanthidae Atypichthys latus X O O Microcanthus strigatus X

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Chaetodontidae Chaetodon auriga X X X C C O Chaetodon baronessa R

Chaetodon bennetti X R

Chaetodon citrinellus X X X O O O Chaetodon ephippium X X X O O R Chaetodon flavirostris X X X C C C Chaetodon guentheri X X X R R O Chaetodon kleinii X X X O O R Chaetodon lineolatus X X X O O O Chaetodon lunula X R R R Chaetodon lunulatus X X X O O O Chaetodon melannotus X X X O O O Chaetodon mertensii X X X C O O Chaetodon X R R ornatissimus Chaetodon pelewensis X X X O O O Chaetodon plebeius X X X C O O Chaetodon R punctatofasciatus R Chaetodon speculum X X R O O Chaetodon tricinctus X X X A A C Chaetodon trifascialis X X C O C Chaetodon ulietensis X X O O R Chaetodon X X X O O unimaculatus R Chaetodon X X X O O vagabundus O C. auriga x C. lunula? R

C. flavirostris x C. R lunula Heniochus X R O chrysostomus R Heniochus singularius R

Heniochus varius R

Forcipiger flavissimus X X X C O O

Chaetodontoplus Pomacanthidae X X X O O conspicillatus O C. meredithi X

Pomacanthus X R R semicirculatus R Centropyge heraldi R R

C. bispinosus X X O O O C. tibicen X X X C C O C. flavissima R

C. bicolor X R

C. vrolikii X X R O R C. flavissima x C. R vrolikii

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Genicanthus X X X O O semicinctus O G. watanabei R R

Pentacerotidae Evistius acutirostris X R

Pentaceros X decacanthus

Pomacentridae Abudefduf vaigiensis X X O O R A. sexfasciatus X R

A. whitleyi X

A. sordidus O

A. vaigiensis x A. R bengalensis Amblyglyphidodon R curacao Amphiprion mccullochi X X X O O O A. latezonatus R R Plectroglyphidodon X R O imparipennis R P. dickii X X X O O O P. johnstonianus X X X O C C P. lacrymatus R

P. leucozonus X O O

Chromis atripectoralis X X X O O O C. viridis X X O O O C. vanderbilti X X X C C O C. agilis X X O R

C. flavomaculata X X X C C O C. iomelas X O O R C. margaritifer X X X O O R C. xanthura R

C. chrysura X O R R C. hypsilepis X X X A A A C. ternatensis X

Chrysiptera notialis X X X A A C C. leucopoma O

C. glauca O R Dascyllus aruanus X X X C C O D. reticulatus X X X O O R D. trimaculatus X X X O O R Neoglyphidodon X X X C C polyacanthus O Parma polylepis X X C C O Pomacentrus parvo R R R P. australis X

P. vaiuli R

P. cf. reidii X

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P. coelestis X X X O O O Stegastes fasciolatus X X X A A A S. gascoynei X X X A A A Stegastes apicalis O Chrysiptera taupon R

Sphyraenidae Sphyraena barracuda R R R

Labridae Bodianus diana R R R B. axillaris X X X O O R B. mesothorax X O R

B. loxozonus X R

B. perdito X X X O O O Choerodon graphicus X

Cheilinus fasciatus R

C. chlorurus X X X O O R C. trilobatus X X O O R Cheilinus undulatus X X R Oxycheilinus X X O O diagramma O O. bimaculatus X X R O O O. unifasciatus X X X O R R Wetmorella R nigropinnata Novaculichthys X X O O taeniourus O Xyrichthys pavo X R R

X. celebicus ? R

Cirrhilabrus punctatus X X C C C C. laboutei X X X O O O Pseudocheilinus X X O O hexataenia O P. evanidus X R

Anampses X X X C C neoguinaicus C A. geographicus X X O O O A. caeruleopunctatus X O O O A. femininus X X X C C C A. elegans X X X C C C A. twistii X

Coris aygula X X X O O O C. bulbifrons X X X C C C C. batuensis X R R

C. dorsomacula X O O O C. gaimard X X O O R C. picta X X X C O O C. pictoides X R

C. sandageri X

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Epibulus insidiator X R R Halichoeres hortulanus X X R R R H. trimaculatus X X C C O H. margaritaceus X X O O O H. nebulosus X O O O H. biocellatus X O O O Hemigymnus X X X O O melapturus O H. fasciatus X X O O R Hologymnosus X X O R annulatus R H. doliatus X R R R H. longipes R R

Suezichthys arquatus X O R R Cheilio inermis X X X R R R Gomphosus varius X X X O O O Macropharyngodon R choati M. meleagris X X X O O O M. negrosensis X X X R R

M. kuiteri X O R Stethojulis bandanensis X X X C C C S. interrupta R R

S. strigiventer X X X R R

Pseudolabrus X X X A A luculentus A Thalassoma lunare X X X C C C T. amblycephalum X X X A A A T. lutescens X X X A A C T. jansenii X X X O C C T. hardwicke X X X O O O T. purpureum X X X C C C T. trilobatum X X C C C T. quinquevittatum X X C C C Labrichthys unilineatus X X O R O Labroides bicolor X X O O O L. dimidiatus X X X A C C L. pectoralis X

Labropsis australis X O R

Pseudodax X X R moluccanus Pseudocoris yamashiroi X X O O Pterogogus cryptus X X R R

Scaridae Cetoscarus bicolor X X R R

Chlorurus microrhinos X X X C C C C. sordidus X X X C C C C. frontalis X X C C C

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Scarus rubroviolaceus X X R R O S. ghobban X X X C C C S. frenatus X X X C C C S. niger X X X O O O S. forsteni X R R S. rivulatus X X R O R S. psittacus X X X A A A S. globiceps X X X C C C S. oviceps X X O O C S. dimidiatus X X R R R S. spinus R R S. longipinnis X X X O O

S. altipinnis X X C C C S. schlegeli X X X A C C S. flavipectoralis R O O S. chameleon X X X C C C Calotomus carolinus R R Hipposcarus longiceps X O O O

Parapercis Pinguipedidae X X X O O hexophthalma O P. cylindrica X O R R P. millepunctata R R

Opistognathidae Opistognathus sp. X R

Limnichthys cf. Creediidae X donaldsoni

Blenniidae Aspidontus taeniatus X R R

Plagiotremus X X X C O tapeinosoma O P. rhinorhinchus R R R Meiacanthus phaeus X X C

Cirripectes alboapicalis X X X O

C. castaneus X X O

C. chelomatus R R Cirripectes sp. X

Ecsenius fourmanoiri X X X A C C Exalias brevis O O O Crossosalarias R macrospilus Istiblennius sp. X O

I. periophthalma X

Salarias fasciatus X O O O Stanulus talboti X X O

Enchelyurus ater X

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Ecsenius axelrodi R Blennidae sp R

Tripterygiidae Norfolkia squamiceps X

Vanclusella sp. X

V. rufopilea X

Nemateleotris Microdesmidae X X O R magnifica R Ptereleotris evides X X X O O O P. zebra X X O O O P. monoptera X X O

Gobiidae Valenciennea strigata R

Amblygobius phalaena R

Gobiodon unicolor O

G. brochus O

G. rivulatus O

G. erythrospilos O

G. quinquestrigatus O

Paragobiodon O echinocephalus Various goby species 23

Zanclidae Zanclus cornutus X X X C C C

Acanthuridae scopas X X X O O O Z. veliferum X X X C O O Acanthurus triostegus X X O O O A. olivaceus X X C O R A. dussumieri X X X C C C A. blochii X X X O O R A. nigroris X O O O A. albipectoralis X X C C C A. mata R R A. nigrofuscus X X X A O C A. pyroferus X O

A. xanthopterus X R R Paracanthurus hepatus X X O R R Ctenochaetus striatus X X C O O C. binotatus X R O R Prionurus maculatus X X X A A A Naso lituratus X O O O N. vlamingi X X R R R N. unicornis X X X C C C N. brevirostris X X X O R R N. tonganus X X O O O

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N. maculatus X R O O N. annulatus R

Siganidae Siganus argenteus X X X O O R S. punctatus R

S. woodlandi R

Scombridae Euthunnus affinis O O O Acanthocybium solandri X

Thunnus albacares X

Coryphaenidae Coryphaena hippurus X R

Pleuronectidae Samariscus triocellatus X

Bothidae Bothus mancus R R

Balistidae Pseudobalistes fuscus X O O O Sufflamen fraenatus X X X C C C S. chrysopterus X X X C C C Melichthys niger X

Balistoides conspicillum R

Balistapus undulatus R

Rhinecanthus X O O rectangulus R R. aculeatus X X R

R. lunula X O O R

Monocanthidae Aleuterus monoceros X

A. scriptus R

Cantherhines pardalis X X C C C C. dumerilii X X O O O C. fronticinctus X R

Pervagor alternans X X R R R P. melanocephalus X X R

P. janthinosoma X R R

Oxymonacanthus X R R longirostris Paraluteres prionurus R R R

Ostraciidae Ostracion cubicus X X X C C C O. meleagris R

Tetradontidae Canthigaster coronata X X O O O C. valentini X X X C C C C. janthinoptera X X R R

C. callisterna X X O R

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Lagocephalus scleratus X

Diodontidae Diodon hystrix X X X C C C D. holocanthus X

Diodon sp R

REPTILES:

Green turtle Chelonia mydas O O Hawksbill turtle Eretmochelys imbricata O

Loggerhead Caretta caretta turtle R

DOLPHINS

Bottlenose Turciops truncatus O dolphin O

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