ECOLOGICAL MONITORING OF REEF FISHES, INVERTEBRATES AND MACROALGAE AT LORD HOWE ISLAND MARINE PARK, NEW SOUTH WALES, 2006-2010

SURVEY AND REPORTING: 1,2 1 2 2 GRAHAM EDGAR , JOE VALENTINE , TONI COOPER , RICK STUART-SMITH , 3 3 SALLYANN GUDGE , IAN KERR

1 AQUENAL PTY LTD, 244 SUMMERLEAS ROAD, KINGSTON, TAS 7050 2 INSTITUTE FOR MARINE AND ANTARCTIC STUDIES, UNIVERSITY OF TASMANIA, NUBEENA CRES, TAROONA, TAS 7053 3 LORD HOWE ISLAND MARINE PARK, PO BOX 161, LORD HOWE ISLAND, NSW 2898

AUGUST 2011

A QUE NAL PTY LTD

www.aquenal.com.au

DOCUMENT INFORMATION

TITLE: Ecological Monitoring of reef fishes, invertebrates and macroalgae at Lord Howe Island Marine Park, New South Wales, 2006-2010

PUBLISHER: AQUENAL PTY LTD ABN 86 081 689 910 244 Summerleas Road Kingston Tasmania 7050 Phone +61 (0)3 6229 2334 Fax +61 (0)3 6229 2335 E-mail: [email protected] Website: www.aquenal.com.au

REPORT CITATION: Aquenal (2011) Ecological Monitoring of reef fishes, invertebrates and macroalgae at Lord Howe Island Marine Park, New South Wales, 2006-2010. Report for New South Wales Marine Parks Authority by Aquenal Pty Ltd, Kingston, Tasmania, August 2011

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SUMMARY

Quantitative surveys of reef communities across the Lord Howe Island Marine Park (LHIMP) were undertaken during February/March each year from 2006 to 2010, with the exception of 2007. A total of 49 different sites were investigated overall, including 33 core sites resurveyed most years, and up to 46 sites surveyed in a single year. A total of 285 fishes (250 identified to species level), 31 crustaceans, 51 molluscs and 38 echinoderms were recorded, including eight new fish records and the first sighting of the eastern rock lobster Sagmariasus verreauxi at the island. Threatened species observed during surveys included the green turtle (Chelonia mydas), black cod (Epinephelus daemelii), blotched fantail ray (Taeniura meyeni) and humphead wrasse (Cheilinus undulatus). Fish species protected under New South Wales fisheries regulations sighted comprised the elegant wrasse (Anampses elegans), bluefish (Girella cyanea), and Ballina angelfish (Chaetodontoplus ballinae). No introduced fish, invertebrate or algal species was seen during surveys.

Underwater visual censuses indicated that the composition of the fish community changed through time from 2006 to 2010 in SZs, where fishes are fully protected, relative to HPZs, which remain open to fishing. The density of large (>40 cm length) fish increased by nearly an order of magnitude in SZs while remaining stable in HPZs. Densities of bluefish (Girella cyanea) declined dramatically within HPZs while remaining stable in SZs, a possible consequence of overfishing in HPZs. Densities of spotted sawtail (Prionurus maculatus) unexpectedly increased within SZs while remaining stable in HPZs. No changes in macro-invertebrate communities within SZs were detected relative to changes within HPZs. Such changes are, however, predicted through the longer term as predation by increasingly-high densities of large fishes and lobsters in SZs affects populations of invertebrate prey.

A threefold increase in density of the most abundant urchin species—the destructive-grazing black urchin Centrostephanus rodgersii—was recorded between 2006 and 2010, while densities of all four other common urchin species also rose significantly over this same time period. An increasing population of black urchins with associated expansion of ‗barrens habitat‘, in particular, greatly increases extinction risk for the rich local seaweed flora, which includes 47 species not recorded outside the LHIMP.

A total of 124 categories of sessile taxa were identified from photoquadrats. Prominent species recorded comprised the reef-building corals Acropora palifera (6.3%), Porites heronensis (3.9%), Pocillopora damicornis (1.1%), Acropora solitaryensis (0.8%), Cyphastrea cerialia (0.7%), Stylophora pistillata (0.5%) and Acropora yongei (0.5%), the soft corals Cladiella sp. (2.0%) and Xenia sp. (1.7%), and the algae Asparogopsis taxiformis (4.2%), Dictyota sp. (3.7%), Caulerpa racemosa (1.4%), Dilophus sp. (1.0%), Sarcodia ciliata (0.8%), Codium spongiosum (0.8%), Chlorodesmis major (0.4%), and Lobophora variegata (0.4%). These taxa were not evenly distributed across the LHIMP, but responded to a large degree to variation in wave exposure, and generally fell within three major groupings — (i) a coral grouping, (ii) a macro-algal grouping, and (iii) a sponge, hydroid, crustose coralline algae and hydrocoral grouping. Communities of sessile organisms at the Algal Holes, South East Rock and North Bay were distinctly different from each other and from other community types observed.

Total coral cover was stable between survey years, with 20% mean cover of the seabed between 2006 and 2010. By contrast, mean cover of large foliose macro-algae across all sites declined from 36% to 25% over the same period. This decline was not evenly spread across sites, but a precipitous decline decline occurred between 2006 and 2008 at sites with high recruitment of the urchin Tripneustes gratilla, with later major declines of macro-algal cover between 2008 and 2010 at other sites. The most likely cause of the macro-algae decline at T. gratilla outbreak sites is that this urchin overgrazed the seabed, causing expansion of bare areas of substrata at the expense of the foliose macro-algae. Declines in macro-algae at other sites more likely reflected the continuing population increase of other urchin species, particularly Centrostephanus rodgersii.

An additional apparent threat to LHI biodiversity values is coral bleaching, a threat exacerbated by the transitionary oceanographic position of LHI on the Tasman Front. Most coral colonies within the northern and eastern sectors of the Lagoon were greatly affected by an extreme heating event that extended through the first three months of 2010.

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Negligible bleaching had been recorded in prior marine park surveys, although a widespread coral bleaching event was recorded at LHIMP in summer 1998 (P. Harrison unpubl. data), albeit with limited detectable impact on coral cover at that time (Harrison et al. 2011). Bleaching in 2010 was most severe (>90% of coral) amongst sites investigated at North Bay and Sylphs Hole, with Comets Hole, Horseshoe and Signal Point also badly affected (>50%). These sites were the most sheltered, and also were subjected to temperatures above 26 oC for a period that exceeded two months – an apparent threshold. Bleaching disproportionately affected Pocillopora damicornis, Acropora yongei, Seriatopora hystrix and Porites heronensis compared to faviid corals and Acropora palifera, including when the different coral taxa co-occurred at the same sites.

A full evaluation of bleaching impacts on coral communities through the medium term should be undertaken through follow-up surveys at the core ecological monitoring sites in February 2012. In addition to allowing an assessment of coral mortality, such surveys should include fishes and mobile macro-invertebrates to allow assessment of any persistent ecosystem-level impacts. Collection of coincident density data on fishes, macro- invertebrates, corals and macro-algae during the peak of the 2010 bleaching period provides an unprecedented opportunity to assess such broad-scale impacts.

The LHIMP ecological monitoring program has increasingly relied on skilled volunteers associated with the Reef Life Survey program to achieve monitoring goals. Assistance of volunteer divers trained in underwater visual census techniques should continue to be encouraged as it has allowed a major expansion of surveys in both time and space.

Management recommendations arising from the study are summarised as follows: 1. The magnitude of ongoing ecological change between surveys indicates that the ecosystem may be transitioning, and that frequent ecological monitoring is desirable. Follow-up surveys are particularly needed to quantify persistent impacts of the 2010 bleaching event on sessile invertebrate, macroalgal, mobile invertebrate and fish communities. These should be undertaken as soon as possible, ideally February 2012. 2. The reef monitoring program should be extended through the long term, with surveys undertaken at core monitoring sites on at least a three-yearly basis. 3. In intervening years, data collection from core monitoring sites should be undertaken as resources allow, including through facilitation of mechanisms that allow further assistance from volunteer divers trained in underwater visual census techniques. 4. Additional surveys of impacted and reference sites should be undertaken following exceptional events (e.g. oil spills, extreme bleaching). 5. Factors contributing to the formation of urchin barrens, and impacts of urchin barren formation on inshore flora and fauna, should be assessed. 6. An expanded network of long-term monitoring sites should be identified and developed in the intertidal zone. 7. Stakeholder negotiations should be undertaken with the aim of creating a sanctuary zone that includes the unique Algal Holes community type, as this community likely includes many globally endemic species and may be threatened by increasing grazing of sea urchins. 8. With the exception of suggested modification of boundaries associated with the Algal Holes, the boundaries of sanctuary zones should remain stable through the long term. The South East Rock sanctuary zone has particular importance within the existing sanctuary zone network due to exceptionally high fish biomass. 9. The 2006 baseline evaluation of marine pests should be updated through the establishment of a five-yearly monitoring program. 10. Consultation with the fishing sector should be undertaken with respect to management of declining bluefish numbers observed in HPZs. 11. Research studies on topics relevant to LHIMP management should be actively encouraged.

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CONTENTS

Summary ...... 3

1 Introduction ...... 6

Methods ...... 9 1.1 Fish surveys ...... 9 1.2 Macroinvertebrate and cryptic fish surveys ...... 12 1.3 Sessile biota (macroalgae and sessile invertebrate) surveys ...... 12 1.4 Statistical analyses ...... 12

2 Results ...... 15 2.1 Threatened and protected Species ...... 15 2.2 Fish and macro-invertebrate community structure ...... 15 2.3 Community-level effects of MPA zoning regulations ...... 19 2.4 Species responses to MPA zoning regulations ...... 24 2.5 Coral and macroalgal community structure ...... 27 2.6 Inter-annual trends ...... 31 2.7 Coral Bleaching ...... 33

3 Discussion ...... 37 3.1 Distribution of reef communities across marine park zones ...... 37 3.2 Ecological changes related to marine park zoning ...... 38 3.3 Coral bleaching ...... 39 3.4 Expanding urchin barrens and other potential threats ...... 40 3.5 Ecological monitoring within the LHIMP ...... 41 3.6 Summary of recommendations ...... 43

4 Acknowledgments ...... 44

References ...... 44

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1 INTRODUCTION

Lord Howe Island sits on the western margin of the basaltic Lord Howe Rise, an isolated north-south trending underwater volcanic ridge positioned approximately 600 km east of the northern NSW coast (Woodroffe et al. 2006). The crescent-shaped island rises to 800 m height, and is 11 km long and slightly less than 3 km wide, with a shallow (<4 m depth) coral-fringed lagoon extending for ca. 7 km off the western coast.

Oceanographic conditions on the island are greatly affected by the position of the Tasman Front, a boundary between tropical (Coral Sea) and temperate (Tasman Sea) water masses that seasonally moves north and south of Lord Howe Island. North-south oscillations in this front contribute to seasonally-alternating cooler and warmer waters in the region (Nilsson and Cresswell 1981). Water temperatures vary from ca. 17°C in winter to ca. 25°C in late summer (Hutton 1986), although temperatures up to 28°C have been recorded in the sheltered lagoon (see Allen et al., 1976).

The alternating influences of warm and cool currents have created an oceanic transition zone in the region between temperate and tropical biomes (Kennedy et al. 2002). Waters surrounding Lord Howe Island consequently possess rich and unusual biodiversity, and are globally important in several respects (Environment Australia 2000). This importance was recognized by inclusion of the island on the UNESCO World Heritage List in 1982 (Environment Australia 2000). Values that contributed to the World Heritage listing and are specific to the marine environment include:  The unusual combination of tropical and temperate taxa of marine flora and fauna, including many species at their distributional limits, reflecting the extreme latitude of coral reef ecosystems, which comprise the southernmost true coral reef in the world.  The high diversity of marine benthic algae, fishes and marine invertebrates and associated high levels of endemism.

Lord Howe Island marine faunas and floras are exceptionally diverse for an isolated oceanic island. Of the 433 fish species known to occur in inshore waters, the majority are wide-ranging tropical forms, while ca. 10% are found only at Lord Howe Island, southern Australia and/or New Zealand (Allen et al. 1976). Approximately 4% (15 species) of the shore fishes are endemic to the Lord Howe region (including Norfolk Island) and 32% are restricted to the south-western or southern Pacific Ocean (Allen et al. 1976).

At least 65 species of echinoderms, comprising 70% tropical species, 24% temperate species and 6% endemic species, have been recorded from the island (Pollard and Burchmore 1985). Lord Howe Island is also known to support over 1500 species of molluscs, including several endemic species (Hedley and Hull 1912, Iredale 1940, Iredale and Allan 1940, Allen and Paxton 1974, Parker 1993). A total of 83 species of coral from 33 genera in 11 families have also been recorded (Harriott et al. 1993, 1995). While this number is relatively low compared with tropical reefs, it indicates a high diversity considering the island‘s latitude and isolation from other major coral communities. Lord Howe Island coral communities include populations of tropical species at the southern limits of their distribution, as well as subtropical species that are rare or absent on tropical reefs (Harriott et al. 1995).

Algal assemblages at Lord Howe Island are diverse and abundant compared with tropical reefs, with macroalgal species dominant over corals at some sites (Harriott et al. 1995). More than 305 species of benthic algae are present, comprising approximately 65 green algae, 67 brown algae and 173 red algae, and with 47 (15%) endemic species (Millar and Kraft 1993, 1994a, b). The most common genera are the brown algae Dictyota, Sargassum and Lobophora, and the green algae Caulerpa, Ulva, Codium and Chlorodesmis (Harriott et al. 1995). Lord Howe Island is also important because it sits at the extreme latitudinal limit of many algal species and genera. Amongst the green algae, the world‘s highest latitude populations in the genera Neomeris, Boodlea, Valoniopsis, Ventricaria and Trichosolen are located at Lord Howe Island (Millar and Kraft 1994a).

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The New South Wales Government proclaimed the Lord Howe Island Marine Park (State Waters) (LHIMP) in 1999 to protect marine conservation values within 46,000 ha surrounding Lord Howe Island. An associated multiple-use zoning scheme came into force on 1 December 2004 (Figure 1). Adjoining the NSW marine park immediately offshore, the Australian Government proclaimed the Lord Howe Island Marine Park (Commonwealth Waters) in 2000, thereby providing protection to marine life from ocean long-lining and trawling for an additional 300,000 ha.

The state zoning scheme partitioned coastal waters to allow different human usages while minimising threats to marine conservation values, and included restrictions on fishing. The objectives of the LHIMP zoning scheme are: (a) to conserve marine biological diversity and marine habitats, (b) to maintain ecological processes, and, where consistent with the preceding, (c) to provide for ecologically sustainable use of fish and marine vegetation, and (d) to provide opportunities for public appreciation, understanding and enjoyment (Marine Parks Act 1997). The majority (73%) of the marine park (state waters) consists of large Habitat Protection Zones (HPZs), which allow some forms of fishing, including charter fishing and recreational fishing other than spearfishing. Trawling, long- lining, dredging and most traps and nets are prohibited throughout the marine park. Seven Sanctuary Zones (SZs; 27% of LHIMP state waters) are also included in the LHIMP, where fishing and other extractive activities that harm marine life are prohibited. Three small inshore Special Purpose Zones allow for fish feeding. A prohibition on anchoring in lagoon SZs, and for all vessels greater than 5 m also in lagoon HPZs, offers additional protection for benthic habitat. Restrictions on fishing associated with the LHIMP management plan were gazetted, and thereby came into force, in 2005.

Fishing pressure on waters surrounding Lord Howe Island, both before and after declaration of LHIMP, has been low compared to that experienced in mainland Australian waters, with almost all fish taken by handline/rod and reel. Fishing practices recognized as most environmentally harmful—including gill-netting, fish traps, long–lining dredging and trawling—are prohibited under the zoning plan. While no commercial fishing licences currently exist, nine MPA permitted charter vessels operate and sell their product locally – mainly through the restaurant trade. No commercial export of fish is allowed and all fish caught commercially are consumed on the island.

Prior to declaration of LHIMP, areas most heavily fished included waters off the northern coast out to the Admiralty Islands, and the northern, southern and outer lagoonal regions off the west coast. Three locations (Comets Hole, Neds Beach and Erscotts Hole) have acted as de facto sanctuary zones for over 40 years, following recognition by local residents that these sites possess much greater value for tourism than for their fishery resources (Gary Crombie, pers. comm.).

Through collaborative studies involving staff at the Lord Howe Island Marine Park (LHIMP), the Lord Howe Island Board, the University of Tasmania, Reef Life Survey Foundation and Aquenal Pty Ltd, the NSW Marine Parks Authority has provided the focus for a series of ecological monitoring surveys (Aquenal 2006, 2008, 2010, Edgar et al. 2010, Valentine and Edgar 2010). These surveys are aimed in large part at assessing changes in fish, macro- invertebrate, coral and macro-algal communities that develop through time within marine park zones with differing restrictions.

Changes through time in patterns of marine biodiversity have been assessed during four surveys undertaken in the same season (February/March) in the years 2006, 2008, 2009 and 2010. Between 28 and 46 sites were surveyed each year using underwater visual transects, with sites distributed using a sampling design that allowed assessment of the extent of recovery of populations in SZs relative to HPZs. Quantitative surveys were undertaken for fishes, mobile macro-invertebrates, sessile invertebrates and benthic macroalgae. Survey protocols were based on Reef Life Survey methods (www.reeflifesurvey.com), which are similar to those applied widely in marine park assessments around Australia, thereby allowing continental-scale comparisons using standardised methods through the long term (Edgar et al. 1997, Edgar and Barrett 1999, Edgar et al. 2004a).

The primary objectives of the Lord Howe Island marine ecological monitoring program are to: 1. Establish a quantitative reef biodiversity dataset that can be used as a baseline for ongoing assessment of ecological changes associated with climate change and other human impacts. 7

2. Describe patterns of marine biodiversity in shallow reef habitats around Lord Howe Island. 3. Determine whether any major changes in biotic communities have occurred since the initial baseline survey in 2006, particularly changes in relation to the marine park zones 4. Detect the presence of introduced or invasive species, and describe associated ecological impacts. 5. Detect the presence of threatened and protected species, and monitor population trends for such species.

Although no-take sanctuary zones within the LHIMP have only been in existence for five years, hence the magnitude of any ecological changes was not expected to be great (Edgar et al. 2009), a number of specific predictions associated with MPA declaration were tested. These related to recovery of fished populations within SZs when compared to any background environmental changes evident in HPZs, and secondary flow-on effects of increased fish predator densities on prey populations. These predictions were:

1. The abundance and biomass of higher carnivorous and large-bodied fishes has increased through time in SZs relative to HPZs, 2. The abundance and biomass of exploited species (Coris bulbifrons and Girella cyanea) has increased through time in SZs relative to HPZs, 3. Densities of invertebrate prey, particularly urchins, declined as a consequence of increased predator numbers in SZs relative to HPZs, and 4. Densities of macroalgae and coral increased as a consequence of decreased grazing pressure from urchins in SZs relative to HPZs.

Predictions were tested by assessing the significance of changes between the start (2006) and end (2010) of the monitoring because any effect of protection would be expected to increase through time and be largest in the 2006 versus 2010 comparisons. Data from 2006 were considered to represent near baseline conditions in analyses of MPA effects, given that these initial surveys were conducted one year after gazettal of sanctuary zones and ecological changes associated with protection of fishing in MPAs generally take decades to manifest (Edgar et al. 2009).

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METHODS

Field surveys at Lord Howe Island Marine Park were conducted from mid February to early March in 2006, 2008, 2009 and 2010. Densities of fishes, mobile macro-invertebrates and sessile organisms were separately estimated along 50 m transect lines using Reef Life Survey (www.reeflifesurvey.com) methods. Geographical coordinates of sites (projection WGS84) were recorded using handheld Garmin GPS systems accurate to ~10 m (Table 1). A total of 33 subtidal sites were repeatedly investigated and an additional 16 sites surveyed in only one or two years. (Figure 1, Table 1). Each ‗site‘ included transects undertaken to quantify densities of fishes, mobile macro- invertebrates and sessile biota at two different depths (Table 1). Sites were located either within, or adjacent to, seven sanctuary zone ‗locations‘ (Admiralty Islands, North Bay, Sylphs Hole, Lord Howe Island Lagoon, South Coast, Observatory Rock and South East Rock).

Three environmental variables were quantified at each site – depth, underwater visibility, and wave exposure (Table 1). The depth of each transect was recorded from SCUBA gauges. Underwater visibility was estimated as the maximum distance sighted by divers along transect lines. Wave exposure was estimated at each site using a five point ordinal scale based on the following units: 1: highly sheltered conditions with little wave or wind energy; 2: sheltered conditions in lagoonal and other protected environments with little oceanic swell but wind waves; 3: sheltered coast open to limited swell; 4: coast open to moderate swell; and 5: coast open to full oceanic swell from prevailing east and south swell directions.

1.1 FISH SURVEYS

Subtidal ecological surveys were undertaken along two depth contours at a total of 46 sites distributed around Lord Howe Island and the Admiralty Islands (Table 1; Figure 1), and at an additional 3 sites near Balls Pyramid. The general fish censusing protocol at each site involved a diver layer a 50 m transect line along a depth contour on the reef, then swimming back while recording the number and estimated size-category of all fishes sighted within 2.5 m of either side of the diver. The transect block thus encompassed a total reef area of 50 m x 5 m with the water column above to a height of 5 m. The diver next censused an adjacent replicate block by swimming back parallel to the initial transect at a distance of about 6 m from the initial transect line. This up and back procedure was repeated at a second depth contour, generating duplicate transect block data for each of two depths at each site (1000 m2 total). Data were recorded on waterproof paper. Size-classes of total fish length used in the study were 25, 50, 75, 100, 125, 150, 200, 250, 300, 350, 375, 400, 500, 625, 750, 875 and 1000+ mm. Lengths of fish >1 m length were individually estimated.

Fish abundance counts and size estimates were converted to biomass estimates using length-weight relationships presented for each species (in some cases genus and family) in Fishbase (http://www.fishbase.org/search.php). In cases where length-weight relationships were described in Fishbase in terms of standard length or fork length rather than total length (as recorded by divers), additional equations provided in Fishbase allowed conversion between different length metrics. For improved accuracy in biomass assessments, the bias in divers‘ perception of fish size underwater was additionally corrected using relationships presented in Edgar et al. (2004b). Note that estimates of fish abundance made by divers can be greatly affected by fish behaviour for many species (Edgar et al. 2004b); consequently biomass determinations, like abundance estimates, can reliably be compared only in a relative sense (i.e. for comparisons with data collected using the same methods) rather than providing an accurate absolute estimate of fish biomass for a patch of reef.

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Figure 1. Zoning scheme and locations of biodiversity survey sites named in Table 1. Note that marine park zones and sites in the Balls Pyramid region (sites 31, 32, 43) are not included in this figure. Dashed line indicates Lagoon reef edge.

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Table 1. Geographical coordinates (Datum = WGS84), depths of two transects, estimated wave exposure (Exp), underwater visibility (Vis; mean for all survey years), marine park zone (SZ: sanctuary zone, HPZ: habitat protection zone), and name of associated sanctuary zone location (Admiralty Islands & Neds Beach SZ, North Bay SZ, LHI Lagoon SZ, East Coast & Shelf SZ, Sylphs Hole SZ and Balls Pyramid SZ) for each biodiversity survey site.

MPA MPA l Site # Site Label Latitude Longitude zone ocation Depths Exp Vis 1 North Channel N-Channel 31.52379 159.03913 HPZ North Bay 14, 11 3 12 2 North Bommie N-Bommie 31.52352 159.04141 HPZ North Bay 5, 4 3 20 3 Erscotts Blind Passage Ers-BP-D 31.54974 159.06295 SZ Lagoon 8, 5 2 20 4 Erscotts Blind Passage Ers-BP-S 31.54974 159.06295 SZ Lagoon 4, 2 2 15 5 Comets Hole Comets-N 31.53908 159.06543 SZ Lagoon 4, 2 1 25 6 Comets Hole Comets-S 31.53961 159.06598 SZ Lagoon 4, 2 1 8 7 Erscotts Passage South Ers-P-D 31.55193 159.06731 HPZ Lagoon 8, 6 2 10 8 Erscotts Passage South Ers-P-S 31.55193 159.06731 HPZ Lagoon 5, 2 2 16 9 Ruperts Reef Ruperts 31.49935 159.06494 SZ Admiralty 14, 11 3 35 10 Noddy Island Noddy 31.50197 159.06513 SZ Admiralty 8, 5 3 25 11 Little Slope L-Slope-D 31.58355 159.06596 HPZ East 14, 11 4 15 12 Little Slope L-Slope-S 31.58355 159.06596 HPZ East 8, 5 4 15 13 Little Island Little-I 31.57082 159.06824 HPZ East 9, 8 4 15 14 Algal Hole North Algal-N 31.56235 159.06843 HPZ Lagoon 8, 7 3 15 15 Algal Hole South Algal-S 31.56469 159.07015 HPZ Lagoon 5.5, 4.5 3 16 16 Rabbit Island offshore Rabbit 31.53915 159.05341 SZ Lagoon 14, 10 4 10 17 North Head North-H 31.52289 159.04014 SZ North Bay 7, 4 3 10 18 Keyhole North Keyhole 31.49747 159.06767 HPZ Admiralty 13, 7 4 25 19 Sugarloaf west Sugarloaf-D 31.50414 159.06679 SZ Admiralty 14, 11 3 30 20 Sugarloaf west Sugarloaf-S 31.50414 159.06679 SZ Admiralty 8, 5 4 30 21 Big Slope B-Slope 31.59540 159.07875 SZ East 11.5, 10.5 5 30 22 Georges Bay Georges 31.56557 159.09975 SZ East 6, 9 5 12 23 Boat Harbour Boat-H 31.55782 159.09852 HPZ East 8.5, 7.5 4 15 24 Phillip Rock N Phillip-D 31.51721 159.03430 HPZ Admiralty 20, 15 4 20 25 Phillip Rock S Phillip-S 31.51721 159.03430 HPZ Admiralty 10, 6 4 20 26 Sylphs Hole N Sylphs-N 31.52032 159.05466 SZ Sylphs 2.5, 1.7 1 25 27 Sylphs Hole S Sylphs-S 31.52087 159.05425 SZ Sylphs 2, 1.5 1 6 28 Old Gulch N O-Gulch-N 31.51202 159.04371 HPZ Admiralty 11, 6 3 10 29 Old Gulch S O-Gulch-S 31.51221 159.04382 HPZ Admiralty 9, 5 3 10 30 Malabar Mal 31.51059 159.05560 SZ Admiralty 14, 13 3 20 31 Wheatsheaf Rocks Wheatsheaf 31.75636 159.23627 HPZ Balls Pyramid 18, 12 5 35 32 Observatory Rocks Observatory 31.75067 159.23682 SZ Balls Pyramid 22, 10 5 35 33 Signal Point Signal 31.52736 159.05983 HPZ Sylphs 0.5, 0.4 1 7 34 Neds Beach Neds 31.51340 159.06903 SZ Admiralty 10, 11 4 35 35 Middle Beach Middle 31.52310 159.07723 HPZ Admiralty 11, 11.5 4 35 36 Stephens Hole Stephens_N 31.53225 159.05403 HPZ Lagoon 1.5, 1.8 1 9 37 Malabar 2 Mal2 31.51130 159.05615 SZ Admiralty 10, 12 3 20 38 North Bay North_B 31.52113 159.04688 SZ North Bay 1.3, 1.5 1 20 39 Yellow Rock Slope Yellow_R 31.52794 159.04575 HPZ North Bay 7, 10 4 18 40 Horseshoe Horseshoe 31.54252 159.06194 HPZ Lagoon 2, 3 2 8 41 Stephens Hole NE Stephens 31.5332 159.05212 HPZ Lagoon 2.7, 3 2 16 42 Stephens Hole SE Stephens-SE 31.5332 159.05212 HPZ Lagoon 4, 4.1 2 25

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Table 1 (continued) MPA MPA Site # Site Label Latitude Longitude zone location Depths Exp Vis 43 South East Rock SE-Rock 31.78750 159.28145 SZ Balls Pyramid 12, 14 5 30 44 Old Gulch West O-Gulch-W 31.51272 159.04262 HPZ Admiralty 6, 15 3 25 45 Malabar Deep Mal-D 31.50823 159.05395 SZ Admiralty 18, 21 3 24 46 Mutton Bird Island Mutton-Bird 31.54218 159.10646 HPZ East 8, 8.1 4 22 47 Neds Beach Inshore Neds-In 31.51793 159.06675 SZ Admiralty 2, 3 2 11 48 Malabar West Mal-W 31.51139 159.05416 SZ Admiralty 10, 17 3 14 49 Erscotts Ers-H 31.54666 159.06128 HPZ Lagoon 2, 4 2 15

Fish species were subdivided into four trophic categories – herbivore, planktivore, benthic carnivore and higher carnivore – and six distributional categories using information provided in Fishbase (http://www.fishbase.org/search.php). Benthic carnivores were distinguished from higher carnivores on the basis of whether their diet predominantly consisted of molluscs, amphipods, isopods and polychaetes rather than other fishes, squid and decapods.

1.2 MACROINVERTEBRATE AND CRYPTIC FISH SURVEYS

Large macro-invertebrates (large molluscs, echinoderms and crustaceans) and cryptic fishes (i.e. inconspicuous species closely associated with the seabed that were likely to be overlooked during baseline fish surveys) were censused along the 50 m transect lines set at two depths per site for fish surveys. A diver swam along the deeper side of the transect, counting all macro-invertebrates and cryptic fishes within 1 m of the line, then returned along the other side of the line counting animals within the adjacent 50 m x 1 m block.

1.3 SESSILE BIOTA (MACROALGAE AND SESSILE INVERTEBRATE) SURVEYS

The percent cover of different floral and faunal components on the seabed was quantified along the subtidal transect lines that were also set for fish and invertebrate censuses by divers taking digital photoquadrats. Photo images were taken vertically-downward each 5 m or 2.5 m along each transect line from a height aimed at encompassing a 0.5 m x 0.5 m area. In situations of low visibility, it was necessary to photograph a smaller area to allow sufficient resolution for discrimination of small organisms. The number of images per transect increased from 10 per transect in 2006 to 20 per transect in 2008, 2009 and 2010, other than when photographers had wide-angle digital SLR cameras, in which case the greater resolution allowed a smaller number of images (10). The scale of each quadrat is evident from centimetre markings along the transect line. A full set of photoquadrat images is available through the NSW Marine Parks Authority and University of Tasmania.

The percent cover of different macroalgal, coral, sponge and other attached invertebrate species was quantified digitally in the laboratory. A grid of 50 points was superimposed on the quadrat image, and species present under each grid point identified to the highest taxonomic level possible and counted. Substratum classes (e.g. rock, rubble, sand) were also recorded when no living organism was evident under a particular grid point. Because of time and cost constraints, photoquadrat data for 2009 have not yet been analysed, other than the single site investigated in 2009 and no other (site 39, Yellow Rock Slope).

1.4 STATISTICAL ANALYSES

Fishes and mobile invertebrates

Relationships between sites in density of species were initially analysed using ordination with Principal Coordinates Analysis (PCO) (Anderson 2003, Anderson and Willis 2003), as run by the PRIMER+PERMANOVA program 12

(Anderson et al. 2008), based on fish biomass, fish abundance, and invertebrate abundance. Data were converted to a Euclidean distance matrix relating each pair of sites after log (x+1) transformation of raw data. The mean of records for different depths, transect blocks and years were calculated for each species and site, generating a site x species matrix. Contributions of important species to patterns of site similarity were assessed by correlating densities of species with the first two coordinate axes generated by PCO, and plotting the highest correlations as vector plots.

To test whether overall changes in community structure were affected by zoning status, a multivariate permutational analysis of variance (PERMANOVA) was conducted using the PRIMER+ statistical package (Anderson et al. 2008). This mixed-model analysis included the fixed factors ‗zone‘ (2 levels, SZ and HPZ) and ‗year‘ (2 levels, 2006 and 2010), with the random factor ‗site‘ (33 levels) nested hierarchically below both factors. The critical test was provided by the ‗zone x year‘ interaction, which if significant would indicate changes in community structure in SZs relative to HPZs through time. Residuals were permutated under a reduced model with a Type III (partial) PERMANOVA model (Anderson et al., 2008). Data for the years 2008 and 2009 were disregarded for this analysis because not all of the 33 sites surveyed in 2006 were resurveyed in those years, and because any effect of protection would be expected to increase through time and be largest in the 2006 versus 2010 comparisons.

Constrained ordination was used to visualise multivariate patterns associated with changes between years in SZ and HPZ sites where a significant interaction between year and zone was indicated by PERMANOVA. The method used was Canonical Analysis of Principal components (CAP), a variation of traditional canonical discriminant analysis that maximises separation of pre-defined groups using the most important coordinate axes generated by Principal Coordinate Analysis (Anderson and Willis 2003, Willis and Anderson 2003). CAP analysis was conducted to visualise relationships between sites in community structure in relation to zones and years, and also to identify species most strongly correlated with the resultant canonical correlation axes. Thus, the species that showed disproportionately large changes in densities between years at SZ sites could be identified using this analysis. Analyses were based on the same community similarity matrices as used for multivariate PERMANOVA tests.

The same PERMANOVA design as applied for multivariate tests was also used for univariate metrics associated with specific ecological predictions. Analytical outputs (Sum of squares, mean squares, F-values) calculated through this process were identical to those calculated using mixed-model analysis of variance (ANOVA) other than P- values, which were calculated using permutation procedures rather than with Gaussian distribution assumptions (Anderson et al., 2008). As populations of species subjected to fishing recovered within SZs, we predicted increasing elevated total biomass in SZs relative to HPZs of (i) large bodied fishes and (ii) higher carnivorous fishes. Depending on fishing pressure, less targeted groups of fishes could also potentially increase in SZs: (iii) benthic carnivorous fishes (iv) herbivorous fishes, and (v) planktivorous fishes. We also tested for secondary effects of fishing, which can occur when increased numbers of large predators cause a reduction in their prey numbers in SZs (Shears & Babcock, 2003; Pederson & Johnson, 2006). Such effects potentially include declines in densities in SZs relative to HPZs of the common urchin species.

Analysis of fish survey data focused on inshore species of conservation interest and those most exploited by fishers or captured as bycatch (elegant wrasse Anampses elegans, double header Coris bulbifrons, bluefish Girella cyanea and Galapagos shark Carcharhinus galapagensis). The other major species targeted by fishers on inshore reefs, the silver trevally Pseudocaranx dentex, occurred in large schools that were insufficiently frequent for meaningful statistical analysis. The effect of protection status was also investigated in relation to total large fish (>30 cm in length; >40 cm length) and key trophic groups using analyses that included density, biomass and species richness for fishes. Other species analysed included the abundant herbivore Prionurus maculatus and the abundant benthic carnivore Pseudolabrus luculentus. The planktivore Chromis hypsilepis was also considered, since this was the most abundant species across all surveys. For mobile macroinvertebrates, analyses focused on total abundance and species richness. Separate analyses were also conducted for the dominant sea urchin species encountered during the surveys (Centrostephanus rodgersii, Heliocidaris tuberculata, Tripneustes gratilla, Diadema savignyi and Echinostrephus aciculatus).

13

Macroalgae, corals and other sessile invertebrates

Relationships between sites in percent cover of different sessile flora and fauna were initially analysed using two graphical procedures, non-metric Multi-Dimensional Scaling (MDS) and Principal Coordinates Analysis (PCO) (Anderson 2003, Anderson and Willis 2003). These were run by the PRIMER+PERMANOVA program (Anderson et al. 2008). Both procedures reduce multidimensional patterns to two-dimensions, although using quite different approaches. The MDS plot provides the best depiction in two-dimensional space of patterns of similarity between sites, whereas PCO identifies the most important two axes in terms of total variation explained.

Data for both procedures were converted to a Bray-Curtis distance matrix relating each pair of sites after square root transformation of raw data. The transformation was applied to downweight the relative importance of the dominant species at a site, and thus allow less abundant species to also contribute to patterns depicted as plots. The mean of records for different depths, transect blocks, and years were calculated for each taxon and site, generating a site x taxon matrix. Contributions of important species to patterns of site similarity were assessed by correlating densities of species with the first two coordinate axes generated by PCO, and plotting the highest correlations (R > 0.5) as vector plots.

Influences of the three measured environmental covariates – wave exposure, depth and underwater visibility – on site similarity patterns were assessed using DISTLM, a procedure analogous to stepwise regression analysis but with a permutational rather than parametric basis. DISTLM was run using the PRIMER+PERMANOVA program with the same site x taxon similarity matrix as used for MDS and PCO, with covariates added in order of maximum R2 explained.

To test whether groups of major taxa varied significantly between years, a univariate permutational analysis of variance was conducted using PERMANOVA (Anderson et al. 2008). The initial analysis included the fixed factor ‗year‘ (3 levels, 2006, 2008 and 2010) and random factor ‗site‘ (32 levels, comprising the 32 sites surveyed in all three years). The site matrix was calculated using Euclidean distance after square root transformation of percent cover data, with residuals permutated using the default program options, namely a reduced Type III (partial) model (Anderson et al., 2008).

Additional analyses were performed to investigate the potential impacts of the sea urchin Tripneustes gratilla (hereafter Tripneustes) on different major elements of the sessile biota. While generally rare in 2006 surveys, Tripneustes densities increased dramatically at some survey sites in 2008, most notably amongst the Admiralty Islands (Valentine and Edgar 2010). In other LHIMP locations, increases in Tripneustes density were not apparent. The variable spatial pattern of increase provided a ‗natural experiment‘ to explore potential ecological impacts of Tripneustes on subtidal communities. Because data sets were available for sites where Tripneustes recruitment occurred, and also for comparable sites where no Tripneustes recruitment was evident, potential impacts could be examined using an approach analogous to a ‗before-after control impact‘ (BACI) design.

A restricted number of the total available survey sites were used for the urchin impact analysis, comprising the 12 ‗offshore community‘ sites identified in the 2006 surveys, subdivided into three Tripneustes density groups. Multivariate analyses conducted in 2006 showed that these sites were ecologically comparable and possessed communities distinct from other survey sites. The 12 offshore sites were conveniently grouped into three urchin recruitment categories for analysis (Valentine and Edgar 2010): (1) High recruitment sites, where Tripneustes density increased by an average of 50-250 per transect over the 2006-08 period; (2) Low recruitment sites, where Tripneustes density increased by an average of 10 – 50 per transect over the 2006-08 period; and (3) sites where Tripneustes was functionally absent (< 0.5 per transect in 2008). Similar PERMANOVA procedures were used in this analysis as for the inter-annual assessment, with two fixed factors: Tripneustes density (three levels: high, low and absent) and year.

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2 RESULTS

2.1 THREATENED AND PROTECTED SPECIES

Threatened species sighted comprised green turtles (Chelonia mydas), black cod (Epinephelus daemelii and the blotched fantail ray (Taeniura meyeni) on transects, and a single humphead wrasse (Cheilinus undulatus) observed in 2008 at Observatory Rocks near Balls Pyramid outside survey transects (Joe Valentine, personal observations). The latter represents the first record of that species in the Lord Howe Island region. The humphead wrasse is listed on the IUCN Red List (IUCN 2006) as ‗Endangered‘, and the blotched fantail ray, which was recorded at Comets Hole in 2008 and Old Gulch in 2010, as ‗Vulnerable‘. Galapagos sharks were also occasionally observed, both within and outside transects. According to the IUCN Red List, this species ―is classified globally as Near Threatened (just failing to meet Vulnerable A2acd, and likely to be A3d in the near future)‖, hence is regarded as possessing conservation concern at the global level.

Six green turtles were observed on transects at sites widely dispersed towards the north of the marine park. Single individuals were recorded on transects at North Channel, North Head, Old Gulch, Middle Beach, Mutton Bird Island and Malabar West. The green turtle is listed as ‗Vulnerable‘ under New South Wales and Commonwealth legislation, and ‗Endangered‘ on the IUCN Red List. Hawksbill turtles (Eretmochelys imbricata) also occur in the marine park and are listed nationally as a ‗Vulnerable‘ species, although none were recorded on transects they are commonly seen in North Bay and Sylphs Hole and by glass bottom boat tour operators (S. Gudge pers comm.). Black cod were recorded on transects at North Bommie (two individuals), Algal Holes South, Sugarloaf, North Bay, Sylphs Hole and Erscotts Hole. Black cod have been totally protected in NSW waters since 1983, and are listed as ‗Vulnerable‘ under the NSW Fisheries Management Act and by Pogonoski et al. (2002) in a threat assessment of Australian fishes. None of the threatened species recorded were in sufficient numbers for analysis of population trends.

Fish species protected under New South Wales fisheries regulations include the elegant wrasse (Anampses elegans), bluefish (Girella cyanea), which is protected off the NSW continental coast, and Ballina angelfish (Chaetodontoplus ballinae), a species not recorded on transects but with sightings during dives at Observatory Rocks and South East Rock near Balls Pyramid. Booth‘s pipefish Halicampus boothae was recorded at 4 m depth at Comets Hole (two specimens) and 16 m depth at Rabbit Island in 2009. One other species in the protected syngnathid (pipefish) family was observed during surveys in a cave at 16 m depth at Malabar on 2 February 2009. This animal was tentatively identified as Heraldia nocturna; however, that species has not previously been reported from Lord Howe Island and a specimen or photograph is needed to validate the sight record.

No introduced fish or invertebrate species were sighted during surveys.

2.2 FISH AND MACRO-INVERTEBRATE COMMUNITY STRUCTURE

A total of 285 fish taxa (including 250 identified to species level) were recorded during 5-m wide fish transects, and 91 cryptic fishes (67 identified species), 31 crustaceans (15 identified species), 51 molluscs (33 identified species) and 38 echinoderms (26 identified species) during 1-m wide surveys. Mean densities of the various taxa recorded at each of the 49 sites studied are listed in Appendices 1 and 2.

Principal coordinate analysis (PCO) based on fish biomass revealed three underlying fish community types: coral- associated, Admiralty Island and offshore island (Figure 2). The coral-associated community type showed a positive relationship with PCO axis 1 and negative relationship with PCO axis 2, and was most clearly evident at Comets and Sylphs Holes. This community was typified by a diverse community of coral-reef fishes such as the butterflyfish Chaetodon plebeius and the pufferfish Canthigaster valentini, and also included the regional endemic anemonefish Amphiprion maccullochi and cardinalfish Ostorhinchus norfolcensis. The Admiralty Islands community type was most evident off Malabar and Old Gulch. This showed a strong positive relationship with PCO 15

axis 2 but was not characterised by any single species. The offshore island community type, such as present at sites near Balls Pyramid, was negatively associated with PCO axes 1 and 2. This community included a range of species of large body mass and the damselfish Chromis hypsilepis.

Amongst sites considered to represent the offshore community type, the most anomalous community in terms of fish biomass was present at South East Rock. Total fish biomass at this site (2009 kg/1000 m2) was more than twice the biomass estimated at any other site, and was over an order of magnitude higher than for the mean for all other sites (190 kg/1000 m2). Exceptionally high fish biomass at South East Rock was evident at all trophic levels except benthic carnivore, with large carnivorous fishes (133 kg/1000 m2 c.f. a mean of 8 kg/1000 m2 elsewhere), planktivores (1091 kg/1000 m2 c.f. a mean of 52 kg/1000 m2 elsewhere) and herbivores (776 kg/1000 m2 c.f. a mean of 88 kg/1000 m2 elsewhere). The estimated total biomass of benthic carnivores was lower than the mean of sites elsewhere (19 kg/1000 m2 c.f. 42 kg/1000 m2 elsewhere), perhaps because of secondary trophic effects (i.e. high numbers of predators were depressing prey numbers).

16

G H

Plate 1. New fish records for the Lord Howe Island region observed from 2010 surveys: A. Pleurosicya micheli; B. Koumansetta rainfordi; C. Heniochus chrysostomus; D. Eviota distigma; E. Pervagor janthinostoma; F: Halichoeres hortulanus; G: Coryphopterus inframaculatus (all photos by RLS diver Andrew Green); H: Cyprinocirrhites polyactis (photo: G. Edgar).

17

Figure 2 Results of principal components analysis showing relationships in fish biomass between assemblages observed at different sites. Strong correlations (r > 0.55) between densities of individual fish species and first two principal components are shown as an associated vector plot.

Site associations assessed using fish abundance differed slightly from those identified using fish biomass (Figure 3). The greater emphasis placed on small abundant species in this analysis, rather than large infrequent biomass dominants, produced a stronger separation of coral-rich sites than for the analysis with biomass. The coral- associated community was very strongly associated with PCO axis 1, which in turn showed a very high correlation with abundances of reef fishes such as chaetodontids. The Admiralties and offshore community types identified in the fish biomass analysis tended to overlap in the abundance PCO. These sites had disproportionately high abundances of Chromis hypsilepis, Pseudolabrus luculentus, Chrysiptera notialis and Pseudanthias squamipinnis. The other major community type identified in the PCO was most notable at Algal Holes, and was characterised by Stegastes fasciolatus. This community type was largely defined on the basis of a negative association with PCO axis 2 and consequently was not as well defined as those defined on the basis of the more important PCO axis 1 (34.5% of total variation explained, compared with 9.6% for PCO axis 2).

Figure 3 Results of principal components analysis showing relationships in fish abundance between assemblages observed at different sites. Strong correlations (r > 0.65) between densities of individual fish species and first two principal components are shown as an associated vector plot. Macro-invertebrates were distributed between different sites along a community gradient that ranged from sites located on the outer edge of the Lagoon to sites in the Admiralties region, with Sylphs Hole, Comets Hole, Algal Hole North and North Bay grouping in a separate cluster to the top right of the PCO plot (Figure 4). The latter grouping was primarily caused by the near absence of any mobile invertebrates at these sites rather than the presence

18

of particular species. Outer Lagoon sites located at the bottom right of the PCO plot (i.e. positively associated to PCO axis 1 and negatively associated with PCO axis 2; Figure 4) also lacked a characteristic fauna other than high abundance of the urchin Heliocidaris tuberculata. By contrast, the Admiralties community type included several common species that were rare elsewhere such as the seastar Ophidiaster confertus, the crinoids Comanthus wahlbergi and Cenolia glebosus, and the urchins Diadema savignyi and Tripneustes gratilla.

Figure 4 Results of principal components analysis showing relationships in macro-invertebrate abundance between assemblages observed at different sites. Strong correlations (r > 0.65) between densities of individual invertebrate species and first two principal components are shown as an associated vector plot.

2.3 COMMUNITY-LEVEL EFFECTS OF MPA ZONING REGULATIONS

The composition of species within the reef fish community changed significantly in sanctuary zones (SZs) relative to habitat protection zones (HPZs) between 2006 and 2010. This is indicated by a significant ‗zone x year‘ interaction in the PERMANOVA for both fish density and fish biomass data (Table 2). Significant patterns of variation between years and sites were also evident from the PERMANOVA (Table 2). Consistent zonal changes were highlighted in the CAP analysis, which depicts changes in community structure for SZ and HPZ sites between 2006 and 2010 (Figure 5). The fish community at SZ sites changed greatly between 2006 and 2010, with no biotic overlap of sites between years on plots of CAP axes 1 and 2, compared to a small overlap of sites between years for HPZs (Figure 5a). This figure notably also indicated a wide scatter of sites in SZs in 2010, with a number of outlying sites in the bottom left of the plot. This pattern can be interpreted as a subset of sites showing large changes in fish populations in response to protection from fishing whereas other SZ sites showed little MPA-related change.

The CAP vector plot (Figure 2b) indicates a large group of fish species negatively correlated with CAP axes 1 and 2 in the bottom left of the figure. These species, which included Prionurus maculatus and the parrotfishes Chlorurus sordidus and Scarus psittacus, thus occurred disproportionately abundantly at SZ sites in 2010. However, contrary to predictions associated with recovery of fishes in protected zones, this group did not include any species recognised as a major target species for fishers.

Invertebrate community structure varied according to survey year, but changes were consistent between HPZ and SZ sites, and no effects of zoning status were apparent (Table 2). As was the case for fish assemblages, the effect of ‗site (zone)‘ was highly significant (P<0.001), indicative of significant differences between replicate survey sites.

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Table 2. Results of PERMANOVA examining community effects for fish biomass, fish abundance invertebrate abundance at each of 33 core sites surveyed in 2006 and 2010, with the fixed factors ‘zone’ (2 levels; sanctuary zone, habitat protection zone) and ‘year’ (2 levels; 2006, 2010), and random factor ‘site’ nested below. Degrees of freedom for F-tests are 1/31, 31/31, 1/31 and 1/31 for zone, site (zone), year and ‘zone x year’ interaction, respectively. ***, p<0.001; ** 0.001

Trophic level/taxa Zone Site (Zone) Year zone x year error MS F MS F MS F MS F MS Fish biomass 12.18 1.07 11.36 1.90 *** 18.17 3.05 ** 10.77 1.81 * 5.96 Fish density 68.63 1.23 55.75 2.52 *** 107.13 4.84 *** 35.50 1.60 * 22.14 Invertebrate density 16.61 0.84 19.80 5.09 *** 29.35 7.55 *** 2.03 0.52 3.89

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Figure 5 Results of CAP analysis maximizing differences in fish community structure (based on fish abundance) in relation to zoning status and year (2006 & 2010). (b) Correlations of species abundance with the two CAP axes are also shown for species with r > 0.35.

21

Further investigation of MPA effects on fish assemblages focused on different trophic groups and also the abundance of large fishes in two size classes (>30 cm and >40 cm). For most of the metrics considered, significant effects attributable to MPA zoning status were not evident (Table 3, Figures 6-8). The only metric where a significant ‗zone x year‘ interaction was apparent was abundance of fish > 40 cm, while total abundance of herbivores was at the margins of significance (P=0.055). For both of these metrics, abundance increased in sanctuary zones between 2006 and 2010, while remaining comparable in sanctuary zones over the same period. No significant ‗zone x year‘ interactions were evident for fish trophic groups when applied to metrics based on biomass, however, the probability value associated with herbivore biomass (P=0.09) approached the significance value of P < 0.05.

Table 3 Results of PERMANOVA for different fish metrics using data for fish abundance, fish biomass and invertebrate abundance at each of 33 core sites surveyed in 2006 and 2010, with the fixed factors ‘zone’ (2 levels; sanctuary zone, habitat protection zone) and ‘year’ (2 levels; 2006, 2010), and random factor ‘site’ nested below. Degrees of freedom for F-tests are 1/31, 31/31, 1/31 and 1/31 for zone, site (zone), year and ‘zone x year’ interaction, respectively. ***, p<0.001; ** 0.001

Trophic level/taxa Zone Site (Zone) Year Zone * Year error MS F MS F MS F MS F MS Species richness Benthic carnivores 0.54 7.36 * 0.07 1.45 0.35 6.91 * 0.02 0.38 0.05 Higher carnivores 0.90 9.97 ** 0.09 1.84 * 0.10 1.95 0.02 0.40 0.05 Herbivores 0.05 0.51 0.09 1.50 0.23 3.68 0.15 2.37 0.06 Planktivores 0.22 1.35 0.16 3.96 *** 0.53 13.17 ** 0.00 0.06 0.04 Total 0.49 7.45 * 0.07 1.48 0.46 10.31 ** 0.03 0.58 0.04 Density Fish > 30cm 0.83 0.36 2.30 2.31 * 4.63 4.64 * 3.05 3.05 1.00 Fish > 40cm 1.63 0.60 2.69 2.75 ** 11.71 11.99 ** 4.44 4.54 * 0.98 Benthic carnivores 0.03 0.22 0.13 0.76 1.59 9.59 ** 0.27 1.62 0.17 Higher carnivores 9.38 7.57 *** 1.24 3.46 *** 1.10 3.06 * 0.01 0.03 0.36 Herbivores 3.34 4.07 * 0.82 2.21 * 0.02 0.04 1.48 3.99 0.37 Planktivores 2.99 1.98 1.51 2.45 * 3.49 5.65 * 0.01 0.02 0.62 Total 1.25 2.31 0.54 2.08 * 2.05 7.92 * 0.01 0.04 0.26 Biomass Benthic carnivores 2.02 5.32 * 0.38 0.90 0.00 0.01 0.03 0.07 0.42 Higher carnivores 10.98 5.81 * 1.89 1.74 0.39 0.35 0.90 0.83 1.09 Herbivores 0.83 0.44 1.90 1.96 * 1.99 2.05 3.28 3.38 0.97 Planktivores 7.56 7.18 * 1.05 1.94 * 4.36 8.03 ** 0.07 0.14 0.54 Total 3.64 5.80 * 0.63 1.45 1.57 3.61 0.51 1.18 0.43

22

Habitat Protection Sanctuary

16 Benthic carnivores 1.4 )

) Higher carnivores 2 2 1.2 12 1 0.8 8 0.6 4 0.4 0.2 0 0

4 6 Herbivores Planktivores 3

4

Species richness (/250 m m (/250 (/250 richness richness Species Species 2 2 1

0 0 2006 2008 2009 2010 2006 2008 2009 2010 Figure 6 Species richness of fish species belonging to different trophic levels recorded in 2006, 2008, 2009 and 2010 (+ SE of site means), in the two major management zone types.

Habitat Protection 600 Benthic carnivores Sanctuary 25 Higher carnivores 20 400 15

200 10 5

0 0

) )

2 2 200 Herbivores 1800 Planktivores 160 1200 120

80 600

40

Density (/1000 m m (/1000(/1000 Density Density 0 0 160 Fish (>30 cm) 90 Fish (>40 cm)

120 60 80 30 40

0 0 2006 2008 2009 2010 2006 2008 2009 2010

Figure 7 Density of fish species belonging to different trophic levels and size classes recorded in 2006, 2008, 2009 and 2010 (+ SE of site means), in the two major management zones.

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Habitat Protection Sanctuary 70 Benthic carnivores 40 60 Higher carnivores 50 30

40

) ) 2 2 30 20 20 10 10 0 0

160

Herbivores 100 Planktivores Biomass (kg/1000 m (kg/1000 Biomass Biomass (kg/1000 m (kg/1000 Biomass 120 80 60 80 40 40 20

0 0 2006 2008 2009 2010 2006 2008 2009 2010 Figure 8 Biomass of fish species belonging to different trophic levels and size classes recorded in 2006, 2008, 2009 and 2010 (+ SE of site means), in the two major management zones.

When total biomass of all fishes is considered, the ratio of fish biomass in SZs relative to HPZs doubled between 2009 and 2010, from an approximately equal distribution of biomass (i.e. ratio of 1) to a ratio of 2.4. Further monitoring is needed to identify whether this change represents a major shift in the system, or whether the pattern reflects the fortuitous sightings in 2010 of patchily-distributed schools containing numerous large-bodied fishes on transects. When the SZ:HPZ fish biomass ratio is plotted for each year of survey (Figure 9), recovery of fish biomass in SZs is apparent only between 2009 and 2010. 4 A Port Phillip

3 ) 2010 Abrolhos I SE Tasmania

2 Fly Point Governor Is SZ/HPZ Aldinga Port Davey Maria I Kent Group Rottnest I 2006 2008 1 Lord Howe I Batemans Bay 2009

Biomassratio( Jurien Bay Jervis Bay

2 5 10 20 40 Years Figure 9 Changes between years in the biomass ratio relating mean biomass of all fishes sighted on transects in SZs with mean biomass in HPZs. Data have been plotted on the scatterplot for MPA sites across southern Australia, as presented in Edgar and Stuart-Smith (2009).

2.4 SPECIES RESPONSES TO MPA ZONING REGULATIONS

Significant effects attributable to MPA zoning status were indicated by a zone x year interaction in PERMANOVA tests for two of the fish and invertebrate species investigated—Girella cyanea and Prionurus maculatus (Table 4).

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Abundance patterns for the exploited species Girella cyanea (bluefish) were consistent with MPA effects, with a steady decline in HPZ sites since the 2006 survey, while remaining at similar densities at SZ sites (Figure 10). Inspection of the raw data indicated that the large decline in bluefish in HPZs largely resulted from the near disappearance of the species from sites in the southern Lagoon area, particularly near Erscotts Passage South and Algal Holes, where the species was abundant in 2006.

Abundances of Prionurus maculatus increased consistently at various sanctuary zone sites in 2010, while abundances at HPZ sites did not change greatly between 2006 and 2010 (Figure 10). While this change could be interpreted as an effect due to MPA zoning, when all survey periods are considered it is notable that abundance patterns for this schooling species have fluctuated considerably during past surveys at both HPZ and SZ sites. Sea urchins, the most common invertebrate grazers on Lord Howe Island reefs, showed no effects attributable to MPA zoning status when 2006 and 2010 survey data were considered. A notable pattern was an increase in sea urchin density evident between 2006 and 2010. This pattern was consistent across HPZ and SZ sites and was statistically significant for all five common species. While an increase in sea urchin density was apparent when comparing 2006 and 2010 survey data, considerable fluctuations in sea urchin density occurred for some species when all survey years are considered. Centrostephanus rodgersii and to a lesser extent Heliocidaris tuberculata have, however, steadily increased in density since 2006 across both HPZ and SZ sites (Figure 11). The Tripneustes gratilla outbreak described previously (Valentine and Edgar 2010) showed no evidence of further expansion. While densities observed in 2010 remain markedly higher than 2006 levels, population numbers for this species appear to have declined over the past two years (Figure 11).

Table 4. Results of PERMANOVA for different fish and invertebrate species using data for fish and invertebrate abundance at each of 33 core sites surveyed in 2006 and 2010, with the fixed factors ‘zone’ (2 levels; sanctuary zone, habitat protection zone) and ‘year’ (2 levels; 2006, 2010), and random factor ‘site’ nested below. Degrees of freedom for F-tests are 1/31, 31/31, 1/31 and 1/31 for zone, site (zone), year and ‘zone x year’ interaction, respectively. ***, p<0.001; ** 0.001

Species Zone Site (Zone) Year Zone * Year error MS F MS F MS F MS F MS Fish Density Carcharhinus galapagensis 0.08 1.35 0.06 1.46 0.01 0.20 0.01 0.20 0.04 Chromis hypsilepis 6.44 0.48 13.51 9.61 *** 3.97 2.83 0.12 0.09 1.41 Coris bulbifrons 0.50 0.60 0.84 1.22 0.49 0.71 0.52 0.75 0.69 Girella cyanea 0.14 0.09 1.48 1.75 1.82 2.14 8.44 9.93 ** 0.85 Prionurus maculatus 0.08 0.02 3.50 1.65 7.98 3.75 9.05 4.26 * 2.12 Pseudolabrus luculentus 0.08 0.07 1.15 3.55 *** 0.14 0.43 0.55 1.70 0.32 Anampses elegans 0.011 0.31 1.05 3.11 ** 0.012 1.10 0.001 0.04 0.011

Invertebrate Density Centrostephanus rodgersii 0.46 0.06 7.70 7.96 *** 13.43 13.88 *** 0.72 0.74 0.97 Diadema savignyi 0.43 0.17 2.50 4.14 *** 6.52 10.79 ** 0.91 1.50 0.60 Echinostrephus aciculatus 2.16 0.49 4.40 4.67 *** 9.37 9.93 ** 0.42 0.44 0.94 Tripneustes gratilla 3.30 1.26 2.63 1.13 15.13 6.51 * 2.48 1.07 2.32 Heliocidaris tuberculata 18.28 2.02 9.04 13.30 *** 7.54 11.09 ** 0.02 0.02 0.68

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1.8 Carcharhinus Habitat Protection galapagensis Sanctuary 1.2

0.6

0 1400 1200 Chromis hypsilepis 1000 800 600 400 200 0 8 7 Coris bulbifrons 6 5 4 3 2 1 0 16 14 Girella cyanea 12

10 )

2 8 6 4 2 0 70 Prionurus maculatus 60 50

40 Density (/1000 m Density(/1000 30 20 10 0 140 Pseudolabrus luculentus 120 100 80 60 40 20 0 25 Anampses elegans 20 15 10

5 0 2006 2008 2009 2010

Figure 10 Mean abundances (+ SE) of selected fish species recorded per site in 2006, 2008, 2009 and 2010 in relation to the two major management zones.

26

Habitat Protection 250 Sanctuary Centrostephanus rodgersii 200

150

100

50

0 200 160 Heliocidaris tuberculata 120

80

40

0

) ) 2 2 250 Tripneustes gratilla 200

150

100

50 Density (/200 m Density Density (/200 m Density 0 45 Diadema savignyi

30

15

0

100 Echinostrephus aciculatus 80

60

40

20

0 2006 2008 2009 2010 Figure 11 Mean abundance (+ SE) of commonly observed sea urchin species recorded per site in 2006, 2008, 2009 and 2010 in relation to the two major management zones.

2.5 CORAL AND MACROALGAL COMMUNITY STRUCTURE

A total of 124 categories of sessile taxa (including classes of bare habitat) were identified in photoquadrats. The most prominent species identified comprised the reef-building corals Acropora palifera (6.3%), Porites heronensis (3.9%), Pocillopora damicornis (1.1%), Acropora solitaryensis (0.8%), Cyphastrea cerialia (0.7%), Stylophora pistillata (0.5%) and Acropora yongei (0.5%), the soft corals Cladiella sp. (2.0%) and Xenia sp. (1.7%), and the algae Asparogopsis taxiformis (4.2%), Dictyota sp. (3.7%), Caulerpa racemosa (1.4%), Dilophus sp. (1.0%), Sarcodia ciliata (0.8%), Codium spongiosum (0.8%), Chlorodesmis major (0.4%), and Lobophora variegata (0.4%). Important aggregated categories comprised Rubble (13.2%), Red algal spp. (fine turf) (9.8%), Crustose coralline algae (8.8%), Red algal spp. (foliose) (8.4%), Brown algal spp. (foliose) (5.5%), Sand (2.9%), Faviid coral

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spp. (1.6%), Coral (dead) (1.4%), Rock (bare) (1.3%), Acropora spp. (1.3%), Hydroids spp. (1.2%), Red algal spp. (filamentous) (1.1%), Green algal spp. (1.0%), and Sponge spp. (encrusting) (1.0%).

Non-metric multidimensional scaling (MDS) using percent cover of the various sessile categories indicated a major separation of lagoonal sites in the upper left of Figure 12 from other sites. This division is clearly evident when sites are coded for wave exposure. Sheltered sites (exposure categories 1 and 2) within the Lagoon are separate from wave exposed sites (exposure categories 3, 4 and 5), with no apparent overlap. Outlying sites with sessile biotas quite distinct from other sites are South East Rock (site 43), Algal Holes (sites 14 and 15) and Yellow Rock (site 39).

The different major communities of sessile organisms are not evenly distributed within sanctuary and habitat protection zones (Figure 13). Sites with communities falling towards the top of Figure 13 tend to be HPZs, while SZs tend to encompass the community type at the bottom left of the figure. The community within a HPZ at the Algal Holes (sites 14 and 15), in particular, is different from communities present at all SZ sites studied.

Stress: 0.16 8 25 33 24 14 36 38 1 15 41 39 42 49 4 21 2 13 7 34 37 11 22 17 27 3 16 35 12 46 26 30 28 23 20 31 6 29 10 32 48 19 40 5 9 18 44 47 Exposure 45 1 2 3 4 43 5

Figure 12 Results of MDS showing relationships in percent cover between sessile floral and faunal assemblages observed at different sites. Site codes are as shown in Table 1. Different symbols depict different levels of wave exposure, from extremely sheltered (1) to open to full oceanic swell (5).

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8 25 33 24 14 36 38 1 15 41 439 2 42 49 7 13 2234 21 37 1116 17 3 35 27 30 12 46 28 26 23 20 31 6 29 10 32 40 5 48 19 9 18 44 47 45

zone Habitat protection 43 Sanctuary

Figure 13 Results of MDS showing relationships in percent cover between sessile floral and faunal assemblages observed at different sites. Site codes are as shown in Table 1. Different symbols distinguish sites within sanctuary zones from those in habitat protection zones.

When biotic relationships between sites are plotted using PCO (Figures 14 and 15), similar patterns are evident to the MDS plots, although outlying sites 14, 15, 43 and 38 are not so extreme, and the level of overlap between SZ and HPZ sites is reduced (Figure 15). Three major groupings of taxa are evident in vector plots showing correlations with first two PCO axes. Sites at the lower left of the figure are characterised by high cover of the corals Porites heronensis and Pocillopora damicornis, the alga Padina spp., the seagrass Zostera capricorni and sand; sites at the lower right of the figure are characterised by macro-algae (Codium spongiosum, Sarcodia ciliata, Asparagopsis taxiformis and other foliose reds); sites in the upper right of the figure are characterised by sponge species, crustose coralline algae, hydroids, the hydrocoral Stylaster brunneus and the alga Caulerpa peltata.

The overlay plot of correlations with covariates shows the high level of correlation between visibility, depth and wave exposure; sites with high underwater visibility also tend to be the deepest sites surveyed, and to have greatest wave exposure (Figure 14). Additional overlays of correlations with the first two PCO axes for major aggregated taxonomic groups indicate different dominant faunal components at different groups of sites. Sites in the bottom right of the plot are dominated by foliose algae, while corals dominate the sheltered sites to the left of the plot, and soft corals dominate sites in the upper plot. Sites dominated by foliose algae tend to have very little bare substrata, while soft coral-dominated sites tend to have considerable bare space.

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40 Exposure 1 2 43 3 Bare substratum 4 45 44 5 20 Soft corals 19189 47 10 32 28 20 Vis 48 29 31 Depth 5 23 40 Exposure 6 12 3046 26 0 22 1735 Coral 49 11 37 27 73 16 34 42 4 39 21 33 36 41 24 13 38 2 25 15 8 PCO2 (14.0% PCO2 oftotal variation) 14 -20 1 Algae

-40 -20 0 20 40 PCO1 (26.2% of total variation) Figure 14 Results of principal components analysis showing relationships in percent cover between sessile assemblages observed at different sites. Site codes are as shown in Table 1. Correlations relating habitat and biological variates to first two principal components are shown as vector plots. Different symbols depict different levels of wave exposure, from extremely sheltered (1) to open to full oceanic swell (5).

40 zone Habitat protection 43 Sanctuary 45 44 Stylaster brunneus Sponge spp. (erect) Rubble Crustose coralline algae 20 1918 9 Caulerpa peltata 47 10 32 Sponge spp. (encrusting) 28 Hydroids spp. 48 29 31 5 23 20 40 6 12 3046 26 22 35 0 17 37 49 3 113916 Porites heronensis 27 42 36 7 2124 Pocillopora damicornis 4 34 2 Sand 33 41 13 Zostera capricorni 38 25 15 Padina spp. Codium spongiosum Red algal spp. (foliose) 8 Sarcodia ciliata Asparogopsis taxiformis PCO2 PCO2 (14.0% oftotal variation) 14 1 -20

-40 -20 0 20 40 PCO1 (26.2% of total variation) Figure 15 Results of principal components analysis showing relationships in percent cover between sessile assemblages observed at different sites. Site codes are as shown in Table 1. Correlations relating percent cover of taxa to first two principal components are shown as vector plots. Different symbols distinguish sites within sanctuary zones from those in habitat protection zones.

The importance of wave exposure as an environmental covariate was affirmed by DISTLM, with a highly significant influence of exposure on sessile community structure (SS = 7760, Pseudo-F = 9.52, R2 = 0.168, P < 0.001). Both underwater visibility (SS = 3709, Pseudo-F = 4.12, R2 = 0.081, P < 0.001) and depth (SS = 5045, Pseudo-F = 5.78, R2 = 0.110, P < 0.001) also showed R2 values that were highly significant in individual tests; however, these covariates were highly correlated with wave exposure and did not contribute significantly to observed patterns once the effect of wave exposure was included in models (Table 5).

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Table 5. Sequential PERMANOVA tests assessing influence of three environmental covariates—wave exposure, underwater visibility and depth—on sessile community relationships between 49 sites. Variable R2 res.df SS(trace) Pseudo-F P Exposure 0.168 47 7760 9.52 0.001 +Visibility 0.189 46 974 1.20 0.267 +Depth 0.219 45 1348 1.69 0.059

2.6 INTER-ANNUAL TRENDS

Cover of reef-building and soft corals both remained relatively constant over the four year period of assessment (Figure 16); however, mean cover of macro-algae declined through time while the amount of bare substrata increased in compensation. These trends were confirmed by PERMANOVA, which revealed highly significant (p<0.001) changes between years in cover of macro-algae and bare substrata, but not for the other two major sessile groups (Table 6). 60

50

40

30

20 Cover (%) Cover 10

0 06 08 10 06 08 10 06 08 10 06 08 10 Macro-algae Coral Bare substrata Soft coral Figure 16 Changes between years in mean cover of four major benthic groups at the 32 Lord Howe Island sites that have been surveyed in all three years investigated (2006, 2008 and 2010).

Table 6. Results of univariate PERMANOVA based on year (fixed factor with three levels: 2006, 2008 and 2010) and site (random factor with 32 levels) for four major benthic groups. Source df SS MS Pseudo-F P(perm) Macroalgae Year 2 33.2 16.61 11.06 0.001 Site 31 326.6 10.54 7.01 0.001 Residual 62 93.1 1.50 Coral Year 2 1.8 0.91 1.59 0.214 Site 31 183.3 5.91 10.36 0.001 Residual 62 35.4 0.57 Soft coral Year 2 0.2 0.09 0.16 0.888 Site 31 121.7 3.93 6.94 0.001 Residual 62 35.0 0.57 Bare substratum Year 2 17.5 8.76 10.62 0.001 Site 31 150.0 4.84 5.87 0.001 Residual 62 51.1 0.82

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Analysis of data for the 12 offshore sites surveyed in 2006-10 indicated different patterns of change through time at sites with outbreaks of the sea urchin Tripneustes gratilla compared to low Tripneustes density sites and sites where the urchin was functionally absent (Figure 17). Such changes were confirmed by PERMANOVA, which indicated an interaction between Tripneustes group and year for macro-algae and bare substrata (Table 7). Thus, for these two major sessile categories, cover changed between years differently, depending on Tripneustes density. Macro-algae declined rapidly between 2006 and 2008 at high density sites, but with a delayed decline between 2008 and 2010 at low density sites, and also a delayed but lesser decline at sites with Tripneustes absent (Figure 17). The opposite pattern was evident for the category ‗bare substratum‘.

80 30 Soft coral 70 Macro-algae 25 2006 60 2008 20 50 2010 40 15 30 10 20 5 10 0 0 30 80 Coral 70 Bare substratum 25 60

Cover (%) Cover 20 50 15 40 30 10 20 5 10 0 0 06 08 10 06 08 10 06 08 10 06 08 10 06 08 10 06 08 10 High Low Absent High Low Absent Tripneustes Tripneustes Tripneustes Tripneustes

Figure 17 Changes between years in mean cover of four major benthic groups at four Lord Howe Island sites within each of three categories of Tripneustes outbreak for offshore sites surveyed in all three years investigated (2006, 2008 and 2010).

Table 7. Results of univariate PERMANOVA based on Tripneustes presence (high density, low density, absent) and year as fixed factors for four major benthic groups. Source df SS MS Pseudo-F P(perm) Macroalgae Tripneustes 2 50.2 25.10 16.05 0.001 Year 2 45.0 22.51 14.40 0.001 Tripneustes x year 4 19.4 4.85 3.10 0.026 Residual 27 42.2 1.56 Coral Tripneustes 2 10.3 5.17 2.02 0.175 Year 2 0.1 0.04 0.02 0.988 Tripneustes x year 4 1.1 0.28 0.11 0.978 Residual 27 69.1 2.56 Soft coral Tripneustes 2 25.0 12.48 7.33 0.005 Year 2 0.6 0.31 0.18 0.853 Tripneustes x year 4 3.0 0.76 0.45 0.790

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Residual 27 46.0 1.70 Bare substratum Tripneustes 2 27.2 13.58 12.55 0.001 Year 2 22.0 11.00 10.17 0.002 Tripneustes x year 4 17.0 4.25 3.93 0.011 Residual 27 29.2 1.08

2.7 CORAL BLEACHING

In contrast to 2006 and 2008 when negligible (<1%) bleaching of corals was recorded, corals in the Lord Howe Island Lagoon suffered from a major bleaching event between February and April 2010, commencing about four weeks prior to the start of our surveys. This bleaching coincided with an extended period of anomalously high water temperature, warm air temperatures, light winds and low swell. These factors all contributed to abnormal conditions for marine flora and fauna – especially within the Lagoon where reduced flushing occurred during this period. The bleaching event significantly affected corals within shallow water habitats (see Figure 18). The most affected regions were located in the shallow northern and eastern sectors of the Lagoon, where sites surveyed at Sylphs Hole, North Bay, Comets Hole, Signal Point and Horseshoe Reef possessed 100% bleached corals on many quadrats, and >50% bleaching of corals overall (Table 8; Figure 19). Minor bleaching of corals was also noted at sites along the western and southern margins of the Lagoon, and at Malabar, Boat Harbour, Little Slope and Neds Beach (Table 8).

Spatial patterns of coral bleaching corresponded closely with wave exposure, with all highly sheltered sites surveyed exhibiting >50% bleaching (Figure 20). Only one additional site (Horseshoe) recorded a similarly high level of bleached corals, and that site was also sheltered but with more wave action than the highly sheltered sites.

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Figure 18 Bleached corals at North Bay (upper) and Sylphs Hole (lower) (photos G. Edgar).

Table 8. Extent of coral bleaching at sites surveyed in 2010 where some bleaching was noted. Live and bleached coral relate to the full seabed, while percent bleached describes percentage of live corals that were found to be bleached. Live coral Bleached Percent Site # Site (%) coral (%) bleached 27 Sylphs Hole S 0.8 39.0 98.0 38 North Bay 6.0 59.2 90.8 26 Sylphs Hole N 1.7 11.1 86.5 6 Comets Hole 6.7 29.8 81.6 40 Horseshoe 11.3 40.3 78.1 5 Comets Hole 14.2 24.8 63.6 33 Signal Point 4.9 6.0 55.0 7 Erscotts Passage South 29.7 8.7 22.6 41 Stephens Hole NE 12.3 2.3 15.9 3 Erscotts Blind Passage 17.3 3.1 15.2 12 Little Slope 25.5 4.0 13.4 42 Stephen's Hole SE 27.6 3.4 11.0 8 Erscotts Passage South 8.8 0.6 6.4 11 Little Slope 13.1 0.8 5.8 13 Little Island 9.6 0.6 5.5 30 Malabar 19.9 0.9 4.3 23 Boat Harbour 12.1 0.5 4.0 34 Neds Beach 18.2 0.5 2.7 16 Rabbit Island offshore 24.0 0.6 2.6 1 North Channel 5.4 0.1 2.4 48 Malabar West 23.3 0.4 1.7 10 Noddy Island 23.4 0.2 0.8

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Figure 19 Distribution of live (black) and bleached (white) corals around Lord Howe Island Marine Park.

100

80

60

40

20

leached coral (%) coral leached B 0 1 2 3 4 5

Figure 20 Relationship between extent of bleaching of live coral and wave exposure at sites investigated.

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The extent of bleaching varied greatly between coral taxa, with very high levels observed amongst the highly branched corals (Acropora spp. Seriatopora hystrix and Pocillopora damicornis), and also the massive coral Porites heronensis. Much lower levels of bleaching were recorded for other massive and prostrate corals, including Acropora palifera and faviid coral species. Although corals present in the highly-sheltered sites where bleaching was widespread generally differed in taxonomic composition from those at more exposed sites, consistent differences in the magnitude of bleaching between taxa were evident at even the most highly-bleached sites. Acropora palifera and faviid corals, in particular, appeared to be largely resistant from bleaching at all sites, including some sites with high levels of bleaching amongst other taxa (Table 9).

Table 9. Extent of coral bleaching in 2010 for major coral taxa across all sites investigated, and for the seven highly-bleached sites where >50% bleaching of coral present was recorded. Live and bleached coral relate to the full seabed, while percent bleached describes percentage of live corals that were found to be bleached. Coral taxon All sites Highly-bleached sites Live Bleached Percent Live coral Bleached Percent coral (%) coral (%) bleached (%) coral (%) bleached Acropora palifera 6.84 0.11 1.6 0.89 0.06 6.0 Acropora yongei 0.18 1.02 85.2 1.16 6.72 85.2 Other Acropora spp. 1.35 0.15 9.9 0.73 0.97 57.0 Echinophyllia aspera 0.15 0.01 2.7 0.10 0.03 22.2 Faviid coral spp. 2.19 0.05 2.3 0.42 0 0 Pocillopora damicornis 1.13 0.86 43.4 1.11 5.31 82.7 Porites heronensis 1.59 2.54 61.4 1.38 14.49 91.3 Seriatopora hystrix 0.13 0.16 55.8 0.46 1.00 68.4 Stylophora pistillata 0.36 0.14 28.5 0.69 0.86 55.5 Other coral spp. 1.24 0.04 3.2 0.26 0.20 43.6

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3 DISCUSSION

3.1 DISTRIBUTION OF REEF COMMUNITIES ACROSS MARINE PARK ZONES

Analysis of 2006 data indicated that sites across the LHIMP possessed reef communities that could be subdivided into three major types (Edgar et al. 2008, Edgar et al. 2010): (i) a coral-associated central Lagoon community, (ii) a macroalgal-dominated community localised at the southern end of the Lagoon, and (iii) an offshore community type such as present in the Admiralty Islands. Surveys of additional sites in 2008 (Aquenal 2008b) further indicated that the reef community located in North Bay was distinctively different from communities elsewhere.

The PCO analysis of sessile taxa integrated across three survey periods (Figure 14) revealed the same three major groupings of taxa, and these can be regarded as fundamental inshore community types within the region — a coral grouping, a macro-algal grouping, and a sponge, hydroid, crustose coralline algae and hydrocoral grouping. The separation between the coral and the other two groupings appears to be largely driven by wave exposure or a closely-related environmental covariate, whereas the separation between the sessile invertebrate and macro-algal communities does not correspond with any measured environmental covariate (Figure 14). In addition, anomalous community types were evident in North Bay, Balls Pyramid and the outer Lagoon.

Surveys of fishes and mobile invertebrates indicated communities present in the offshore region of Malabar and the Admiralty Islands were different to those present along the outer Lagoon edge. Moreover, when patterns of fish biomass of different species were examined across the region (Figure 2), fish communities near Balls Pyramid (i.e. Observatory Rocks, Wheatsheaf Rocks, South East Rock) were seen to differ from those present elsewhere.

The MDS plot for sessile biota also indicated that communities at the Algal Holes, South East Rock and North Bay were distinctly different from each other and from other community types. The difference between PCO and MDS plots for sessile biota, in terms of relative positions of plotted outliers, is due to MDS showing outliers at considerable distance from other points if possessing a community type unlike those found elsewhere, whereas PCO typically shows such outliers close to other points because their isolation does not relate to information relevant to other sites but pertains to higher order canonical axes that are not figured.

Overall, six major community types are indicated for the wider region, with these community types most strongly represented at: 1. North coast/Admiralty Islands; 2. Algal Holes; 3. Inner Lagoon holes; 4. Outer Lagoon; 5. North Bay; and 6. Balls Pyramid. Species assemblages observed at other sites largely comprised a mixture of species from these six community types.

Importantly, the two most anomalous sites for sessile organisms, South East Rock and North Bay, are protected within SZs from threats associated with fishing, including habitat distrubance through trawling. Four of the six major community types are represented in both SZs and HPZs; consequently, effects of fishing on these community types can be deduced by contrasting ecological changes at sites subjected to these two differing management strategies. However, the macro-algal community, which is most strongly represented at Algal Holes, appears to lack any SZ sites and be open to fishing across its full extent. In addition, the North Bay site does not appear to have any comparable HPZ site as it is fully encompassed within a SZ.

As described in the 2006 survey report (Aquenal 2008a), the apparent absence of the macroalgal community type from a SZ reduces the conservation value of the LHIMP because it compromises both the comprehensiveness of the zoning scheme and the design of the associated monitoring program. The latter represents a real benefit of the LHIMP – through monitoring different management zones, the effects of fishing on marine ecosystems can be quantified, including any interactions between fishing, climate change and introduced species. Although numerous additional sites have been investigated since the 2006 surveys, no site within an SZ was encountered with the macroalgal community type, and the existence of such a site is considered highly unlikely given its localised position on the island‘s coast.

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A major conservation concern specifically associated with macroalgae is that this community type could be adversely affected by expansion of urchin populations, which form barrens when present in high density. Lord Howe Island is possibly the richest subtropical locality worldwide for benthic macroalgae, with at least 318 species present, of which 47 (15%) are not known from elsewhere (Aquenal 2008a). Expansion of urchin barrens habitat has been strongly associated with fishing at the global level (Tegner and Dayton 1999). As populations of larger fishes and lobsters predators are depleted by overfishing, urchin populations expand, converting macroalgal habitat to barrens habitat. Densities of the urchin Centrostephanus rodgersii, a species that is presently causing rapid expansion of barrens along the Australian east coast (Ling 2008, Ling et al. 2009), have increased threefold over the past four years, while significant increases in all other common urchins at LHIMP are also now evident. Thus, protection from all fishing of at least some sites with the macroalgal community present should be considered a management priority to reduce the risk of urchins causing extinction of endemic macroalgal species. In the context of potential algal loss, it is notable that local extinction of a macroalgal species at Lord Howe Island has already been documented. The large brown algae ‗Neptune‘s necklace‘ (Hormosira banksii) was a dominant organism on intertidal shores of the Lagoon in the early twentieth century, but beds declined rapidly due to unknown causes in the 1930s (Lucas 1935) and the species is now regarded as extinct on the island (Millar and Kraft 1994b).

Recent surveys revealed the Horseshoe HPZ site to contain fish and macro-invertebrate communities that approach those found in SZ sites in Sylphs Hole and Comets Hole. Thus, whereas no reference HPZ site had been identified in 2006 to allow ecological comparisons through time for changes evident at the two Lagoon coral dominated sites, Horseshoe provides a basis for such comparisons in future monitoring.

3.2 ECOLOGICAL CHANGES RELATED TO MARINE PARK ZONING

Ecological changes that followed protection from fishing in LHIMP SZs were evident for fishes, whereas no secondary effects amongst macro-invertebrate grazers could be detected. At the community level, the fish assemblage changed significantly between 2006 and 2010 in SZs relative to HPZs. Significantly increased densities of large fishes developed in SZs, and density patterns of Girella cyanea and Prionurus maculatus also changed between zones. Nevertheless, contrary to predictions associated with MPAs, no significant increase in biomass of higher carnivorous species as a group was observed, nor increased densities of the exploited doubleheader wrasse Coris bulbifrons.

No significant effect of zone was detected in the PERMANOVA analyses of fish biomass and fish density where interactive effects of zone and year were identified (Table 2), indicating that the distribution of species in 2006 at the start of monitoring was generally similar between SZs and HPZs, and statistical outcomes are robust. Similarly, univariate tests for densities of fishes > 40 cm, Girella cyanea and Prionurus maculatus did not appear to be confounded by pre-existing differences between sites studied in different management zones. By contrast, mean densities of higher carnivorous fishes were much higher within SZs than HPZs at the beginning as well as the end of the monitoring period, complicating interpretation of results. The lack of response to protection of higher carnivorous fishes as a group, which was counter to expected MPA effects, may be due to (i) the spatial patchiness of higher carnivorous fishes generating large confidence intervals and low power in tests, (ii) insufficient time for populations that have been depressed by fishing to recover, or, most likely, (iii) a combination of these factors. Studies elsewhere indicate that fish populations may take decades to recover from effects of fishing (Edgar et al. 2009, Babcock et al. 2010).

Alternatively, populations of higher carnivorous fishes may not be increasing within SZs relative to HPZs, perhaps because populations were not substantially depressed by fishing prior to declaration of the MPA, illegal fishing continues to depress population numbers, or because individuals of higher carnivorous species such as sharks tend to range widely and are captured while moving outside protected zones. Support for the first of these explanations is provided by the low levels of fishing pressure at Lord Howe Island compared to mainland Australia and the rest of the world, with most local fishing directed towards pelagic species such as yellowtail kingfish. Three shallow demersal fish species are known to be directly targeted – doubleheader, bluefish and silver trevally, and no changes 38

in doubleheader populations were detected in the LHIMP through time. Regardless, low levels of fishing effort may be all that is needed to greatly diminish populations of species with large body size and low population density. Further monitoring should clarify which is the best supported of the various alternative hypotheses that explain the small apparent recovery of predatory fishes.

Change in densities of large fishes within SZs was strongest for the group >40 cm length, which increased almost an order of magnitude between 2006 and 2010 in SZs while showing little change in HPZs (Figure 7). When individuals in the 30-40 cm range were added to the analysis, the significance of results declined. When fishes of all sizes were analysed then no interactive effect of zone x year was found. Data relating to small species, which tended to be extremely abundant, swamped the signal provided by the few large species in the total fish analysis.

Rather than increasing within SZs, the bluefish (Girella cyanea) population showed a greater than tenfold decline in HPZs while maintaining stable populations within SZs. The dramatic decline in HPZs indicates a need for further investigation of possible fishery management issues, particularly given the high conservational and recreational significance of this species. Consultation with local fishers should indicate whether patterns observed at sites studied reflect patterns of abundance across the species‘ full local range, or whether large bluefish populations remain in HPZs at unstudied sites. A particular concern is that fishing effort displaced from SZ locations following declaration of the LHIMP may have been transferred and concentrated in the few known HPZ locations with high bluefish densities, and levels of fishing effort are now well above limits of sustainability at those locations. If this is the case then recovery of populations in HPZs will require highly targeted management action to create conditions that facilitate population recovery of this iconic species, which, at the global level is largely reliant on the Lord Howe Island population for breeding.

One unexpected signal arising from the monitoring program was a steady and significant increase in densities of the spotted sawtail Prionurus maculatus in SZs, which contrasted with stable densities in HPZs. Such a pattern of change would normally be interpreted as population recovery in protected zones following overfishing; however, this surgeonfish is not known to have been targeted by fishers. Perhaps this change simply represents a chance event associated with the distribution of sites surveyed (i.e. a Type I statistical error); alternatively, it may reflect interactive changes amongst populations of fish species that have followed protection of the exploited predatory fishes (see Kellner et al. 2010). Changes in spotted sawtail densities should be investigated through the future because of the possibility that they are MPA related, reflecting food web or other connections in ecosystems that are currently unrecognised.

With respect to new sites investigated in 2010, South East Rock requires particular comment with respect to SZ effects because of its exceptionally high fish biomass. Without ongoing monitoring, the basis of this high biomass cannot be definitively attributed; however, both the isolated location, with strong currents and high productivity benefitting planktivores and their predators, and protection within a SZ, which has prevented the fish-down of large fish species, have probably contributed. Because of the great abundance of large predators and planktivores at this location, and the presence of species not recorded elsewhere (including the first record for the Lord Howe Island region of the hawkfish Cyprinocirrhites polyactis during the 2010 survey (Plate 1), and the only Ballina angelfish sighted at sites investigated), this location possesses particular importance for conservation within the sanctuary zone system.

3.3 CORAL BLEACHING

The focal point of an anomalously warm oceanographic cell was positioned over Lord Howe Island at the time of our February 2010 surveys, with water temperatures elevated by ca. 2oC, and remaining above 25 °C for three months until early April (coralreefwatch.noaa.gov) (Harrison et al. 2011). This period of heat stress coincided with an extended period of several weeks when winds were light and prevailed from the southeast, with little swell, thereby greatly restricting flushing of the Lagoon. The coincidence of these atmospheric and oceanographic events resulted in extreme bleaching of corals within the Lagoon, particularly the northern region that includes Sylphs Hole 39

and North Bay, where bleaching affected 90% of all corals observed on transects. Major bleaching (>50%) extended to all sites investigated in the central Lagoon area also, and minor bleaching was noted at sites in southern and western Lagoon areas. The very small level of bleaching observed outside this area should be considered negligible as occasional background coral mortality unrelated to heating is to be expected.

Temperature loggers located at key monitoring sites through the AIMS sea temperature recording initiative (data can be accessed through http://data.aims.gov.au/aimsrtds/map.xhtml?parameterType=water+temperature&latitude=- 15.623036831528252&longitude=135.966796875&zoom=4&channels=&checked=&from=01-01-1980) indicate the magnitude of the heating anomaly. At monitoring stations at 2.5 m depth in Sylphs Hole and 3 m depth in North Bay, water temperatures remained above 25oC for the three month Jan to March period, with temperatures of ca. 26.5 oC persisting for most of January and February at North Bay and ca. 26 oC at Sylphs Hole. Although also a hot year, temperatures were a degree lower through the same season in 2011.

An additional characteristic of the temperature signal at the two most heavily-bleached locations is diurnal variation of around 2 oC, considerably higher than at other temperature monitoring sites apart from Comets Hole, where 1.5 oC diurnal variation was evident. Water temperatures in early 2010 at Comets Hole persisted around 26.5 oC, albeit with a two week period in early February when temperatures dropped to 25 oC at night. Water temperature records at other Lord Howe Island sites (Algal Holes North, Wheatsheaf, Malabar, Sugarloaf West) remained just over 25 oC during the peak of the 2010 heating period. Thus, effects of bleaching corresponded in a non-linear way to the temperature signal, with a major bleaching threshold occurring when temperatures exceeded 26 oC for over a month. Sites with major bleaching were also characterised by strong diurnal fluctuations in temperature, a consequence of the extremely sheltered conditions present at these sites (Figure 20).

An additional contributing factor to the extreme bleaching observed at northern and eastern Lagoon sites is that the dominant corals at these locations appear to be disproportionately susceptible to bleaching. If faviid corals and Acropora palifera had predominated at these sites then the level of bleaching would presumably have been considerably lower. The few colonies of these corals present appeared to resist bleaching much better than Pocillopora damicornis, Acropora yongei, Seriatopora hystrix and Porites heronensis.

The major outstanding question associated with the 2010 bleaching event is: ―Have bleached corals and associated organisms subsequently recovered, or has severe coral mortality occurred?‖ To answer this question, at least in part, ongoing monitoring of the impact of bleaching on coral communities is presently underway through a long-term program undertaken by Southern Cross University researchers (see Harriott et al. 1993, Harriott et al. 1995, Harrison et al. 1995, Harrison and Carroll 2002, Harrison et al. 2011). A full evaluation of the impact of bleaching on coral communities nevertheless also requires follow-up surveys at the core ecological monitoring sites. In addition to corals, such surveys should include fishes and mobile macro-invertebrates to allow assessment of any ecosystem-level impacts of bleaching. Co-located data on densities of macro-invertebrates and fishes are available for all of the coral transects. As was the case with the invasion of the urchin Tripneustes gratilla to northern LHIMP sites between 2006 and 2008 (Valentine and Edgar 2010), a before/after/control/impact analytical design should be applied to identify impacts of bleaching by quantifying population changes (including fishes and mobile macro- invertebrates) at heavily bleached sites relative to changes at sites with little bleaching.

3.4 EXPANDING URCHIN BARRENS AND OTHER POTENTIAL THREATS

Regardless of the extreme 2010 bleaching event, coral cover remaintained at a constant level of ~20% of the seabed between 2006 and 2010. By contrast, mean cover of large foliose macro-algae declined from 36% to 25% over the same period. This decline was not evenly spread across sites, but a precipitous decline occurred between 2006 and 2008 at sites with high recruitment of the urchin Tripneustes gratilla, with subsequent major declines between 2008 and 2010 at other sites.

Surveys of mobile macro-invertebrates undertaken at the same sites and time indicated that all five common urchin species at LHIMP increased in population numbers between 2006 and 2010. For the two most abundant Lord Howe 40

Island species additional to T. gratilla, the urchin-barren forming species Centrostephanus rodgersii underwent a threefold increase in densities, while populations of Heliocidaris tuberculata increased twofold. The most parsimonious explanation for the decline of macro-algae at T. gratilla outbreak sites is that this urchin overgrazed the seabed, causing expansion of bare areas of substrata at the expense of foliose macro-algae. Declines in macro- algae at other sites more likely reflected population increases of other urchin species, particularly C. rodgersii. Increasing urchin grazing does not, however, appear to have affected coral populations, at least to 2010.

The increase in urchin densities represents a major threat to endemic macroalgae within the LHIMP. While densities of T. gratilla have declined since the recruitment peak in 2008, individual animals have presumably grown to larger body size while densities are much higher than in 2006. The expansion of population numbers of C. rodgersii represents a more serious apparent threat to conservation values of the LHIMP than T. gratilla because C. rodgersii consumes sessile plants and animals to bedrock, and its increase in population density appears sustained and increasing rather than episodic.

Management of expansion of sea urchin barrens will be difficult at the scale of the LHIMP, other than through control on fishing of urchin predators. In addition to doubleheader, lobster may have also historically contributed to controlling urchin numbers, as at other tropical and temperate sites worldwide (Tegner and Dayton 1999); however, densities are currently so low that the species presumably has little functional role now. Present controls include sanctuary zones where all fishing is prohibited, and fishing regulations that restrict the take of doubleheader—the main predator of urchins—to a maximum of one fish per person/day, and prohibit sale of this species. The effectiveness of indirect control of urchin barrens expansion by fishing prohibitions in SZs should translate to increased urchin densities in HPZ relative to SZ sites. However, to the present, increasing urchin densities in HPZs have been matched by increasing densities in SZs.

Direct control of urchin numbers through diver removal or chemical means appears impractical because of the scale of effort required (diver removal) or collateral environmental impacts (chemical control). Nevertheless, consideration of direct intervention may be warranted at the scale of tens of metres for sites with exceptional conservation or tourism value, such as at locations with threatened endemic algal species or at major dive tourism sites.

Other potential threats identified previously that could potentially affect conservation values of the LHIMP included nutrient enrichment of inshore habitats, leading to eutrophication and consequently to overgrowth and mortality of corals, seagrasses and macroalgae (Aquenal 2008a). No gross changes in epiphyte cover were apparent in the 2010 photoquadrats since 2006. Small patches of cyanobacteria and epiphytic algae were seen overgrowing coral and seagrasses at sheltered intertidal and shallow subtidal sites in the northeaster section of the Lagoon (North Bay, Sylphs Hole and Signal Point) and Neds Beach, but such patches covered a small total area (>2% of total transect area).

No introduced fish or invertebrate species was observed during surveys; hence the LHIMP appears to remain largely quarantined from this threat.

3.5 ECOLOGICAL MONITORING WITHIN THE LHIMP

Quantitative ecological monitoring baselines have now been established for both inshore (Aquenal 2008b, a) and offshore (Neilson et al. 2010) ecosystems of the Lord Howe Island Marine Park. The baseline for the inshore LHIMP ecological monitoring program was established using a team of scientific divers; however, volunteer divers now contribute greatly to the monitoring program. Surveys in 2010 were undertaken by two visiting scientists (Graham Edgar and Toni Cooper), two LHIMP staff (Sallyann Gudge, Ian Kerr), one LHI board person (Chris Haselden), two local dive operators/residents (Brian Busteed, Tasman Douglass), and five mainland divers (Ian Shaw, Andrew Green, Bill Barker, Nick Mooney, Chris Preston) trained in Reef Life Survey (RLS) techniques.

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The contribution of RLS divers has allowed a major expansion of effort. With help from good weather conditions, transect data were obtained from a total of 46 sites in 12 field days in 2010, compared to 38 sites covered in 14 days in 2008 and 33 sites in 11 days in 2006. A total of 29 sites were covered by RLS divers in 9 days in 2009. More than two transects were also contributed by RLS divers at some sites. Thus, only two transect depths were undertaken at sites visited by scientific divers, whereas additional transect depths could generally be covered at sites surveyed by RLS divers because of the larger available diver pool and longer total dive time.

The contribution of RLS divers to the monitoring program thus allowed an expansion of site and depth coverage, and also allowed surveys to be undertaken in 2009 when few financial resources could be directed to the ecological monitoring program. The quality of data collected should not have declined during this expansion of effort with RLS divers, given that no significant difference was found in data collected by RLS volunteers when compared to data obtained by professional scientists (Edgar and Stuart-Smith 2009). The contribution of accommodation by the Lord Howe Island Board at the Island research station, and support contributed by dive operators, provided additional impetus to the monitoring process and also led to an expanded survey effort.

Populations of reef communities within the LHIMP appear to undergo much more extreme interannual fluctuations than observed at other Australian marine parks with long-term ecological monitoring programs (see, for example, Barrett et al. 2002, Edgar et al. 2003, Barrett et al. 2007, Keough et al. 2007, Barrett et al. 2009). Lord Howe Island‘s location on the Tasman Front presumably contributes to this variability, with major oceanographic changes occurring across the region from year to year. In order to track the ecological consequences of this variability, which greatly increases extinction risk for species with highly localised distributions, monitoring should occur at Lord Howe Island on a more regular basis than at continental Australian locations. Ideally, monitoring should occur on an annual basis; however, this would entail such a large commitment of available management resources that such a recommendation is likely impractical.

Consequently, we recommend a monitoring cycle of three years, with extended surveys of all core monitoring sites undertaken and reported on a three-yearly basis, but with data infilling in intervening years through support of RLS divers, provided that infilling surveys can be undertaken at minimal cost. Power analyses could usefully be undertaken to refine the periodicity of the monitoring strategy; however, before these are undertaken key management goals relating to effect sizes and desired power to detect change need to be defined. Such goals will be difficult to quantitatively define given the variety of different elements of marine biodiversity that comprise targets of management planning (including protection of endemic, rare, threatened and protected species, and habitat types). Moreover, power analyses are readily conducted only for relatively simple statistical tests involving a single factor and effect size (see, for example, Keough et al. 2007), rather than tests involving interaction between multiple factors, as applied here.

To the present, the core set of sites has comprised the set of 33 sites investigated as a baseline in the first year of monitoring (labelled 1-33 in Table 1). Continuation of monitoring at this set of sites through the long-term clearly has value because of the longevity of the dataset; nevertheless we recommend some refinement to this set of core sites to remove redundancy in zone evaluation, and to better allow comparisons between patterns developing in SZs relative to HPZs. Ideally, at least two core sites should be located for different community types within each SZ, and a similar number of sites in nearby HPZs also regularly monitored, so that the efficacy of each SZ can be individually assessed. The minimum set of 26 core sites that we suggest is integral to ongoing reef monitoring in the LHIMP is listed in Table 10. Criteria used to select this set were: (i) two sites for each community type in each SZ where present, (ii) two sites for each community type and sanctuary zone in an HPZ adjacent to core SZ sites, and (iii) where redundancy exists, remove sites that are most logistically challenging to access or in close proximity to core sites.

During any monitoring period, this set of core sites should be surveyed as first priority, while recognising that additional sites should also be added as weather and logistic constraints allow. Such additional sites should preferentially include the sites numbered from 1 to 33 that have not been identified here as core sites, and also new

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sites that possess a combination of protection, community type and management zone that has not already been covered by duplicate sites (e.g. HPZ at South East Rock).

Table 10. Sites regarded as highest priority sites for long-term monitoring. Site Name Protection Zone Community Latitude Longitude Depth 18 Keyhole North HPZ Admiralty Admiralty 31.49747 159.06767 13, 7 28 Old Gulch N HPZ Admiralty Admiralty 31.51202 159.04371 11, 6 19 Sugarloaf west SZ Admiralty Admiralty 31.50414 159.06679 14, 11 30 Malabar SZ Admiralty Admiralty 31.51059 159.0556 14, 13 40 Horseshoe HPZ Lagoon Inner Lagoon 31.54252 159.06194 2, 3 5 Comets Hole SZ Lagoon Inner Lagoon 31.53908 159.06543 4, 2 6 Comets Hole SZ Lagoon Inner Lagoon 31.53961 159.06598 4, 2 14 Algal Hole North HPZ Lagoon Macroalgal 31.56235 159.06843 8, 7 15 Algal Hole South HPZ Lagoon Macroalgal 31.56469 159.07015 5.5, 4.5 7 Erscotts Passage South HPZ Lagoon Outer Lagoon 31.55193 159.06731 8, 6 8 Erscotts Passage South HPZ Lagoon Outer Lagoon 31.55193 159.06731 5, 2 3 Erscotts Blind Passage SZ Lagoon Outer Lagoon 31.54974 159.06295 8, 5 16 Rabbit Island offshore SZ Lagoon Outer Lagoon 31.53915 159.05341 14, 10 1 North Channel HPZ North Bay Outer Lagoon 31.52379 159.03913 14, 11 2 North Bommie HPZ North Bay Outer Lagoon 31.52352 159.04141 5, 4 38 North Bay SZ North Bay Outer Lagoon 31.52113 159.04688 1.3, 1.5 31 Wheatsheaf Rocks HPZ Balls Pyramid Balls Pyramid 31.75636 159.23627 18, 12 32 Observatory Rocks SZ Balls Pyramid Balls Pyramid 31.75067 159.23682 22, 10 43 South East Rock SZ Balls Pyramid Balls Pyramid 31.7875 159.28145 12, 14 13 Little Island HPZ East Mixed 31.57082 159.06824 9, 8 23 Boat Harbour HPZ East Mixed 31.55782 159.09852 8.5, 7.5 21 Big Slope SZ East Mixed 31.5954 159.07875 11.5, 10.5 22 Georges Bay SZ East Mixed 31.56557 159.09975 6, 9 33 Signal Point HPZ Sylphs Inner Lagoon 31.52736 159.05983 0.5, 0.4 26 Sylphs Hole N SZ Sylphs Inner Lagoon 31.52032 159.05466 2.5, 1.7 27 Sylphs Hole S SZ Sylphs Inner Lagoon 31.52087 159.05425 2, 1.5

3.6 SUMMARY OF RECOMMENDATIONS

Recommendations described in this report, plus recommendations in previous reports that remain relevant, are summarised as follows: 1. The next survey of monitoring sites within the LHIMP should be undertaken as soon as possible, preferably February 2012, the same month as previous surveys to reduce any seasonal biases in interpretation of data. This survey should be regarded as urgent given that major changes in ecological communities are clearly evident in cover of sessile flora and fauna observed in 2006, 2008 and 2010, indicating that the ecosystem may be transitioning between alternative states, and that a more frequent monitoring program is desirable than necessary for most mainland Australian MPAs. For the sessile biota, major stresses on the system are evident as coral bleaching, decreased densities of macro-algae, and expanding urchin barrens, while other ongoing changes are also occurring amongst the fish and urchin communities. 2. Follow-up surveys are particularly needed to quantify persistent impacts of the 2010 bleaching event on sessile invertebrate, macroalgal, mobile invertebrate and fish communities. 43

3. The reef monitoring program should be extended through the long term, with surveys undertaken at 26 core monitoring sites on at least a three-yearly basis. 4. In intervening years, surveys at core monitoring sites by RLS divers should be facilitated as resources allow, as should data collection from non-core monitoring sites investigated in 2006, and new sites within zones and with community types poorly covered by the set of core sites. 5. Additional surveys of impacted and reference sites should be facilitated following exceptional events (e.g. oil spills, extreme bleaching), and analyses of impacts undertaken using a before/after/control/impact (‗BACI‘) statistical design. 6. An assessment should be undertaken of factors contributing to the formation of urchin barrens, and impacts of urchin barren formation on other floral and faunal species. 7. An expanded network of long-term monitoring sites should be identified and developed in the intertidal zone. 8. A sanctuary zone should be created to include the algal community type near Algal Holes through negotiation with stakeholders. The optimum time for negotiating this management change is during the next LHIMP zoning review, although if major loss of macro-algae at the Algal Holes is detected before this time then negotiation should be brought forward as a matter of urgency. 9. With the exception of suggested modification of boundaries associated with the Algal Holes, the boundaries of sanctuary zones should remain stable through the long term. The South East Rock sanctuary zone has particular importance within the existing sanctuary zone network due to exceptionally high fish biomass and anomalous community structure. 10. The 2006 baseline evaluation of marine pests should be updated through the establishment of a program with comprehensive marine pest monitoring at five-yearly intervals, plus continuation of existing surveys of accessible habitat types (including piers, beach wrack and moorings) in intervening years. 11. Consultation with the fishing sector should be undertaken with respect to management of declining bluefish numbers observed in HPZs. 12. Research studies on topics relevant to LHIMP management should be actively encouraged and supported as financial resources allow. Highly relevant studies include: a. Hydrological studies of subterranean water and nutrient flow into the Lagoon and southwest coastal reefs. b. Effects of nutrient enrichment associated with sewage effluent on coral reefs and seagrass beds. c. Effects of nutrient enrichment on the Algal Holes region. d. Environmental factors contributing to separate development of, and ecosystem differences between, coral and macroalgal community types. e. An assessment of impacts of rodent baits on inshore marine communities.

4 ACKNOWLEDGMENTS

This study would not have been possible without the logistic support provided by the New South Wales Marine Park Authority at Lord Howe Island and the Lord Howe Island Board, who generously provided laboratory and accommodation facilities for divers. The assistance and support of local dive operators (Pro-Dive Lord Howe Island and Howea Divers) who provided valuable local knowledge to assist survey operations is also greatly appreciated, as is the commitment made by Reef Life Survey and local divers (Ian Shaw, Andrew Green, Bill Barker, Nick Mooney, Jemina Stuart-Smith, Chris Preston, Tom Davis, Christo Haselden, Brian Busteed and Tas Douglass) to the project, which included considerable time and travel expense.

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