SEASONAL AND TEMPORAL CHANGES IN REEF FAUNA AT JULIAN ROCKS 2010 TO 2013

Report to Cape Byron Marine Park, NSW Department of Primary Industries

Brigitte Sommer, Maria Beger and John Pandolfi School of Biological Sciences, The University of Queensland September 2014

Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Table of contents

1 Executive Summary...... 1 2 Introduction ...... 3 2.1 Aims and deliverables...... 5 3 Methods ...... 5 3.1 Field methods ...... 5 3.2 Desk top studies and statistical analyses ...... 6 3.2.1 Spatial and temporal patterns in coral and benthic communities...... 6 3.2.2 Spatial and temporal patterns in fish communities...... 7 4 Results ...... 8 4.1 Benthos...... 8 4.1.1 Patterns in coral and benthic community structure ...... 8 4.1.2 Multivariate community structure ...... 13 4.1.3 Temporal trends in benthic fauna...... 15 4.2 Fish ...... 18 4.2.1 Patterns in fish species composition ...... 18 4.2.2 Patterns in fish abundance...... 19 4.2.3 Patterns in fish biomass ...... 22 4.2.4 Multivariate community structure ...... 24 5 Discussion...... 29 5.1.1 Outlook and implications for management...... 32 6 References...... 33

ii Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

1 Executive Summary

Cape Byron Marine Park is centrally located within the tropical to temperate transition zone in eastern Australia. Julian Rocks, a small offshore island in the Marine Park, provides habitat for a unique combination of tropical, subtropical and temperate taxa, including non-reefal coral assemblages and critically endangered grey nurse sharks. To investigate seasonal and temporal dynamics of rocky reef communities at Julian Rocks, benthic and fish communities at two sites (Nursery, False Trench) were surveyed biannually, in March and August, between 2010 and 2013. Species composition, abundances and fish biomass were recorded along five replicate 25m long belt transects using standard methods for benthos, hard corals, and fishes.

Benthic communities at Julian Rocks were dominated by macroalgae, with turfing algae, species from the family Dictyotaceae, and red algae most abundant. Benthic invertebrates (other than hard corals), including soft corals, sponges, anemones, ascidians, zoanthids, hydrozoans, bryozoans, barnacles and tubeworms were the second most abundant benthic category at all sites and surveys, followed by scleractinian corals. Relative abundance of macroalgae and other invertebrate taxa was variable among sites and survey periods.

Scleractinian corals in the region occur at or close to the southern limits of their distribution, in conditions that are considered marginal for coral reef growth. Hard corals at Julian Rocks do not form accreting coral reefs and comprised a maximum of 6.8% ± 0.83 of benthic cover in the underlying surveys. A total of 19 scleractinian coral species from eleven genera, within seven coral families were recorded in transects at Julian Rocks. Over 40% of coral species were tropical, including Acropora valida, Favites abdita, Stylophora pistillata, Goniopora djiboutiensis, Montrastrea curta, Acantastrea echinata and Turbinaria peltata. Most tropical coral species were rare, with only one or two colonies recorded in transects. Instead, the bulk of coral abundance consisted of cosmopolitan and subtropical coral species, including Goniastrea australensis, Turbinaria frondens, T. mesenterina, Acropora solitaryensis and the subtropical endemic Pocillopora aliciae (formerly recorded as P. damicornis). In terms of coral life-history-strategies, coral assemblages were dominated by ‘stress-tolerant’ species, a life-history strategy favourable in marginal and disturbed environments.

Fish assemblages at Julian Rocks were characterised by strong tropical influence, with 72.9% of species having tropical affiliation. Abundant fish species were mainly small-bodied planktivores and micro-invertebrate predators, which is consistent with trophic patterns at higher latitude NSW locations, and may result from increased productivity from the regional upwelling of nutrient enriched waters. Consistent with latitudinal abundance gradients in herbivorous fishes, herbivores had intermediate abundances at Julian Rocks, and the large schools of roving herbivores typical of tropical coral reefs were absent. In contrast to the dominant influence of small-bodied planktivores on fish abundance patterns at Julian Rocks, large-bodied predator and apex predator species were the main drivers of patterns in fish biomass, which differed among sites. Julian Rocks provides important habitat for several species of sharks, including leopard shark (Stegostoma fasciatum), spotted (Orectolobus maculatus) and ornate (Orectolobus ornatus) wobbegongs, and the critically

1 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 endangered grey nurse shark (Carcharias taurus), for which Julian Rocks provides critical habitat under the NSW Fisheries Management Act 1994.

Surprisingly, no seasonal signal in benthic or fish communities was recorded from the underlying data. This stability in reef communities at Julian Rocks may be linked to topographically induced, current-driven upwelling of colder and nutrient enriched waters throughout the year, as the East Australian Current accelerates over the narrowing continental shelf at Cape Byron. No interannual differences among benthic and fish communities were detected, except for a significant decline in coral cover at the Nursery in August 2012. Overturned and broken coral colonies, as well as large amounts of sand at the Nursery suggest that this decline may have been the result of a storm event preceding the August 2012 surveys. This highlights the vulnerability of fauna at Julian Rocks to local disturbance and highlights the need to minimise existing stressors (e.g. from habitat degradation, fishing), to enable biotic communities to withstand and recover from disturbances.

2 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

2 Introduction

Understanding the processes that govern the distribution and organisation of biological communities is a prerequisite to predicting how they may respond to climate change and for their management and conservation (Guisan & Rahbek 2011). Ecosystems at the transition of biogeographical zones, in particular, are projected to undergo multidimensional changes, both through direct climate change impacts and through a shift in the distribution and abundances of species (Parmesan & Yohe 2003; Poloczanska et al. 2013). Climate change therefore has the potential to fundamentally alter ecosystems in the subtropical-to-temperate transition zone, which harbour a unique overlap of tropical, subtropical and temperate species, many of which occur at the margins of their ranges and environmental tolerances.

Eastern Australia is under the influence of the East Australian Current (EAC) which transports tropical waters to cooler temperature regions (Roughan & Middleton 2004; Hayes et al. 2005; Roughan et al. 2011) and enables coral assemblages in pockets of suitable habitat. The EAC reaches its maximum strength in the subtropical-to- temperate transition zone, where it drives complex oceanographic conditions (Ridgway and Dunn 2003; Hayes et al 2005), as well as latitudinal and cross-shelf patterns in the distribution of benthic and fish assemblages (Harriott et al. 1994; Harriott & Banks 2002; Booth et al. 2007; Dalton & Roff 2013; Sommer et al. 2014). Oceanographic conditions fluctuate seasonally and between years and are strongly influenced by El Niño Southern Oscillation patterns (Zann 2000). The EAC peaks in February and is weakest in winter, when the flow reduces by up to 50% (Hayes et al. 2005). The EAC also impacts the transport of larvae (Booth et al. 2007; Noreen et al. 2009; Noreen et al. 2013) and coastal connectivity, which varies with latitude, in- and offshore position, seasonality, and in relation to climatic oscillations such as El Niño and La Niña (Roughan et al. 2011).

Subtropical environments are cooler, receive less light and are exposed to greater variability and seasonality than tropical environments (Willig et al. 2003), creating marginal environmental conditions for many tropical taxa. While coral reefs tend to occur in warm, clear, shallow, logographic, fully saline and aragonite supersaturated seas (Buddemeier 1997; Kleypas et al. 1999), high-latitude coral assemblages, exist at the margins of their distributions, in close proximity to their environmental limits (Kleypas et al. 1999). The world’s southernmost true coral reefs are located at offshore Lord Howe Island, Australia, at 31°33’S (Veron & Done 1979; Harriott et al. 1995). The southernmost true coral reef along the eastern Australian coast is situated at Flinders Reef (26°59’S, 153°29’E) in southeast Queensland. Nevertheless, diverse and abundant non-reefal coral communities occur in pockets of suitable habitat along the eastern Australian coast and extend to high latitudes, well south of the Great Barrier Reef (GBR). These non-reefal coral communities are distinguished from true coral reefs based on their inability to accrete calcium carbonate reefs (Buddemeier & Smith 1999). Instead they form low relief veneers of living coral growing on non-reefal substrata that follow the existing seafloor morphology. The reasons for the lack of reef accretion, despite the often high species richness and abundance of scleractinian corals, are still not well established (Buddemeier & Smith 1999; Harriott 1999; Kleypas et al. 1999; Harriott & Banks 2002), and it is likely that cumulative stress from a combination of physical and

3 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 biological factors limits coral reef development (Kleypas et al. 1999; Harriott & Banks 2002).

Coral assemblages in the subtropical-to-temperate transition zone are characterised by the biogeographical overlap of tropical species at or near the southern limits of their distribution; subtropical species, which tend to be rare or absent at lower latitudes, including the endemic species Pocillopora aliciae (Schmidt-Roach et al. 2013); and cosmopolitan species, which tend to be common throughout their geographical distribution (Veron 1993, 2000; Sommer et al. 2014). Although coral species richness generally declines with increasing latitude (Veron & Minchin 1992; Veron 1993; Harriott & Banks 2002; Harrison & Booth 2007; Sugihara et al. 2009; Sommer et al. 2014), coral cover of high-latitude reefs is frequently comparable to that of tropical coral reefs (Harriott et al. 1994; Harriott et al. 1995; Sugihara et al. 2009; Thomson & Frisch 2010; Dalton & Roff 2013; Sommer et al. 2014).

High-latitude coastal reefs of eastern Australia are characterised by widely distributed, generalist, stress-tolerant coral species with massive and horizontally spreading morphologies, and by diminishing influence of tropical coral species at higher latitudes and closer to the mainland (Sommer et al. 2014). Patterns in the distribution of species with different characteristics are strongly shaped by local environmental conditions, suggesting that characteristics linked to energy acquisition and physical stability may be particularly important for coral survival in these marginal, high-latitude environments (Sommer et al. 2014). Apart from difference in coral community structure, high-latitude reefs are furthermore distinguished from tropical coral reefs by greater abundance of other taxa including macroalgae (Johannes et al. 1983; Veron & Minchin 1992; Sommer unpublished data) and invertebrates such as soft corals, ascidians, sponges, echinoderms and barnacles (Harriott et al. 1994; Harriott et al. 1995; Harriott et al. 1999; Sommer unpublished data; Schleyer et al. 2008).

Similarly, sub-tropical fish assemblages harbour tropical, subtropical and temperate species and vary in both species richness and composition along latitudinal gradients (Turpie et al. 2000; Edgar et al. 2004; Smith et al. 2008). In Australia for example, the abundance of tropical species declines with increasing latitude, and the relative influence of tropical, subtropical and temperate species depends on latitude, exposure, and local oceanographic conditions (Booth et al. 2007; Figueira & Booth 2010; Malcolm et al. 2010). The proportion of tropical and temperate fish species varies seasonally, with recruitment of tropical species in summer, and subsequent demise of some species in winter (Booth et al. 2007), as winter temperatures affect performance and over-winter survival of tropical fish species (Figueira et al. 2009; Figueira & Booth 2010). Despite strong evidence for the role of the EAC in the transport of tropical fish species at a coastal scale, individual recruitment events are independent of local increases in temperature, suggesting that recruitment of tropical vagrant fish species largely depends on outside sources (Booth et al. 2007).

Compared to tropical ecosystems, the processes that drive community organisation of these transitional ecosystems are still poorly understood. High-latitude reefs experience a greater range and seasonality of environmental conditions, and temperature extremes may limit the establishment of viable populations (Figueira & Booth 2010) and lead to mortality of tropical (e.g. coral bleaching at Lord Howe Island; Harrison et al. 2011) and temperate species (e.g. range contraction of a

4 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 temperate seaweed; Smale & Wernberg 2013) at their southern and northern range limits, respectively. Studying the temporal dynamics of biogeographic transition zones is therefore an important research goal towards creating a better understanding of these highly complex and dynamic ecosystems.

Cape Byron Marine Park is centrally located within the tropical to temperate transition zone in eastern Australia. The region is influenced by the EAC, which frequently drives upwelling of colder and nutrient enriched waters as it accelerates over the narrowing continental shelf at Cape Byron (Oke & Middleton 2000; Roughan & Middleton 2002, 2004; Everett et al. 2014). Julian Rocks, a small island approximately two kilometres offshore of Byron Bay, provides habitat for a unique combination of tropical, subtropical and temperate taxa, including non-reefal coral assemblages and critically endangered grey nurse sharks. Byron Bay is a popular tourist destination, and whilst protected from extractive activities in no-take zonation, the subtropical reefs of Julian Rocks experience very high levels of diver visitation year round. In the mid 1990s, Harriott et al. (1999) conducted baseline surveys of coral and benthic communities at Julian Rocks and Middle Reef. Over two decades later (Hartley & Gorton 2008) investigated disease dynamics at heavily and rarely dived sites at Julian Rocks and established a relationship between diver visitation rates and coral health. During 2009-2010 the Byron Bay Underwater Group conducted timed swim surveys of fish assemblages using the roving diver technique (Hammerton 2010) and found that fish assemblages differed among habitat protection zone and sanctuary zone habitats. Emphasis was placed on fish families targeted by recreational fishers and on vulnerable or threatened species.

2.1 Aims and deliverables The overall goal of this project is to identify temporal ecological dynamics of reef fauna at Julian Rocks, Cape Byron Marine Park, for the period 2010 to 2013. Specifically, this report aims to establish ecological baselines for fish, coral and benthic communities, by undertaking the following tasks: Processing and CPCe analysis of benthic transect photographs for 5 visits to ‘The Nursery’ and 2 visits to the ‘False Trench’ sites at Julian Rocks over 3 years between 2010 and 2012. Analysis of fish communities for 8 visits to ‘The Nursery’ and 2 visits to the ‘False Trench’ sites over 4 years between 2010 and 2013: species richness, abundance, biomass. Analysis of benthic communities over time: hard coral species richness and diversity; abundance of hard coral, soft coral, sponges, other benthic invertebrates, and macroalgae. Production of a preliminary report ‘seasonal and temporal changes in reef fauna at Julian Rocks: 2010-2013’. Production of final report.

3 Methods

3.1 Field methods Benthos and demersal fishes were surveyed along five replicate 25m belt transects at approximately 8m depth at Julian Rocks between 2010 and 2013. The fish

5 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 observer recorded all fishes, their abundance and size classes in a 5m belt while simultaneously unrolling the transect tapes (625m2 per site). The benthic observer undertook photographic surveys along the same transect lines using a Canon Powershot S90 digital camera in an Ikelite housing, equipped with a wide angle conversion lens and two Ikelite strobes. Photographic methods are widely used in coral reef monitoring, as they provide a permanent record and outperform visual methods in terms of accuracy and time-efficiency (Leujak & Ormond 2007). The benthic observer took twenty-five, non-overlapping 1 m2 size photographs along each transect line and used an L-shaped graduated stick (see Plate 1) to maintain consistent distance from the substratum (approximately 80 cm) and to provide a scale in the photographs for taxonomic identification. Photographs were taken parallel to the substratum to avoid image distortion. Additional macro photographs were taken of coral colonies that require examination of corallite details for species identification (e.g. species from the family Acroporidae). Fish surveys were conducted at the Nursery in March and August 2010, 2011 and 2012, and in March 2013; and at the False Trench site in August 2010 and March 2011. Benthic surveys were conducted at the Nursery in March and August 2010 and 2011, and in August 2012; and at the False Trench site in August 2010 and March 2011.

3.2 Desk top studies and statistical analyses 3.2.1 Spatial and temporal patterns in coral and benthic communities Benthic photographs were analysed using the image analysis software Coral Point Count with Excel extensions (CPCe) (Kohler & Gill 2006). Twenty random data points were overlaid onto each photograph and the category underlying each point was identified and recorded in the categories: scleractinian corals, soft corals, sponges, macroalgae, other benthos, dead corals, and other substrata (Plate 1). Scleractinian corals were identified to species level wherever possible, otherwise to genus and growth form (e.g. branching, corymbose, digitate, table and plate Acropora).

To investigate zoogeographical patterns of corals, we assigned each coral species to one of three zoogeographical groups based on the distributional information in taxonomic monographs (Veron 1993; Wallace 1999; Veron 2000): tropical for taxa classified as ‘common or sometimes common in the tropics’, and that ‘occur in the tropics and are generally rare or uncommon’; subtropical for taxa described as ‘rare or uncommon in the tropics and common in high latitudes’; cosmopolitan for taxa described as ‘common in the tropics and especially common in high latitudes’, ‘common throughout their geographical distribution’, and taxa that ‘occur around Australia’ (see also Sommer et al. 2014).

To evaluate life-history strategies of corals recorded at Julian Rocks each coral species was categorised as ‘competitive’, ‘weedy’ or ‘stress-tolerant’ (Darling et al. 2012; Darling et al. 2013). The competitive life-history-strategy describes fast growing, large, branching and plating species that reproduce by broadcast spawning and occur at shallow depth. Weedy species mainly comprise brooding species of smaller colony size. The stress-tolerant strategy includes slow growing, broadcast spawning species of predominantly domed morphologies, large corallites and high fecundity (Darling et al. 2012; Darling et al. 2013).

Coral species richness and abundances of benthic taxa (i.e. mean percentage cover)

6 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 were calculated for each site. Analysis of variance (ANOVA) was conducted to test for differences in coral species richness, coral cover, macroalgae cover, other benthic invertebrate cover and substrata cover among sites, years and seasons. When significant differences were recorded in initial tests, post-hoc TukeyHSD tests were used to further investigate the nature of these differences. Distance based redundancy analysis (dbRDA) was used to test for differences in multivariate coral and benthic community structure among sites, years and seasons. Due to differences in the number of visits to the Nursery (5 visits) and the False Trench (2 visits) sites, univariate and multivariate tests for differences among years and seasons were only conducted for data from the Nursery. Similarity percentage (SIMPER) analysis was used to investigate the taxa that best characterised the nature of the differences. Differences in community structure among sites, survey years and seasons were visualised using non-parametric multidimensional scaling ordinations (nMDS).

Plate 1. Benthic photograph in CPCe with 20 random datapoints.

3.2.2 Spatial and temporal patterns in fish communities Mean species richness, abundance (i.e. fish counts) and biomass (i.e. fish count x size class) were calculated for each site for each visit. Each species was categorised as either tropical, subtropical or temperate and based on its feeding strategy as planktivore, herbivore, detritivore, corallivore, omnivore, predator or apex predator. Analysis of variance (ANOVA) was conducted to test for differences in species richness, abundance and biomass among sites, years and seasons. Distance based redundancy analysis (dbRDA) was used to test for differences in multivariate fish community structure (abundance and biomass) among sites, years and seasons. Due to differences in the number of visits to the Nursery (7 visits) and the False Trench (2 visits) sites, univariate and multivariate tests for differences among years and seasons were only conducted for data from the Nursery. SIMPER analysis was used to investigate the taxa that best characterised the nature of the differences. Differences in community structure among sites, survey years and seasons were visualised using non-parametric multidimensional scaling ordinations (nMDS).

7 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

4 Results

4.1 Benthos 4.1.1 Patterns in coral and benthic community structure

Macroalgae was the most abundant benthic category recorded during all surveys at the Julian Rocks and False Trench sites, and covered between 57.4% ± 2.13 and 72.8% ± 3.01 of the area surveyed, followed by other invertebrate taxa (8% ± 1.17 to 25.5% ± 4.83), and corals (2.9% ± 1.31 to 6.8% ± 0.83) (Figs. 1 to 7). Benthic cover was greatest at the False Trench site, where benthic organisms covered up to 98.6% of the area surveyed (Figs 1 and 8). Benthic organisms covered all available rocky substrate (i.e. no bare rock was recorded), with sediment (i.e. dead shell pieces) and sand the only substrate categories (Fig. 8).

Fig. 1. Mean percent cover of benthic categories at the two study sites (N = Nursery, T = False Trench) recorded in surveys between 2010 and 2012.

Macroalgae communities were dominated by turfing algae (13.5% ± 1.89 to 44.4% ± 6.85), species from the family Dictyotaceae (14.6% ± 1.52 to 33.1% ± 4.8) and red algae (Amphiroa anceps, red algae other, and encrusting coralline algae; 1.9% ± 0.35 to 35.7% ± 1.98). Non-coral benthic invertebrates contributed up to 25.7% ± 2.95 of benthic cover and comprised diverse assemblages of soft corals, sponges, anemones, ascidians, zoanthids, hydrozoans, bryozoans, barnacles, tubeworms, and echinoderms (Fig. 3). The abundance of other invertebrate taxa was higher at the False Trench site (albeit not significant; Fig. 3 and Table 4), except during March 2010 when particularly high abundance of barnacles (18.6% ± 3.88) was recorded at the Nursery.

8 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig. 2. Mean percent cover of macroalgae at the two study sites (N = Nursery, T = False Trench) recorded in benthic surveys between March 2010 and August 2012.

Fig. 3. Mean percent cover of soft corals, sponges and other invertebrates at the two study sites (N = Nursery, T = False Trench).

9 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

A total of 19 scleractinian coral species from 11 genera, within seven coral families were recorded during the surveys at Julian Rocks, ranging between nine species recorded at the False Trench site in March 2011 and 15 species recorded at the Nursery in August 2010 (Figs. 4 and 5). Mean coral species richness at the False Trench site was significantly lower than at the Nursery (Table 4). Mean hard coral cover ranged between 2.9% ± 1.31 at the Nursery in August 2012 and 6.8% ± 0.83 at the Nursery in August 2011 (Fig. 4), and did not differ significantly among sites (Table 4). Dendrophylliidae (genus Turbinaria) and Faviidae were the most abundant families (Fig. 5). Generally, most coral species were rare, with the species Goniastrea australensis, Turbinaria frondens, T. mesenterina, T. peltata, Acanthastrea echinata, Acropora solitaryensis, and the subtropical endemic Pocillopora aliciae the most abundant species recorded in transects.

Eight of the 19 coral species recorded in transects were tropical (42.1%), seven subtropical (36.8%), and four cosmopolitan (21.1%), with cosmopolitan and subtropical species generally most abundant (Fig. 6). In terms of coral life-history- strategies (Darling et al. 2012; Darling et al. 2013), 13 species were ‘stress-tolerant’ (68.4%), four ‘competitive’ (21.1%) and two ‘weedy’ (10.5%). Stress-tolerant species made up the greatest proportion of coral cover at both sites and in all survey periods, and were most abundant at the False Trench site (Fig. 7).

Fig. 4. Mean percent cover of hard corals (columns) and coral species richness (numerals) at the two study sites (N = Nursery, T = False Trench) recorded in benthic surveys between March 2010 and August 2012.

10 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig. 5. Mean percent cover of hard coral families (columns) and coral family richness (numerals) at the two study sites (N = Nursery, T = False Trench) recorded in benthic surveys between March 2010 and August 2012.

Fig. 6. Relative abundance of cosmopolitan, subtropical and tropical corals at the two study sites (N = Nursery, T = False Trench) recorded in benthic surveys between March 2010 and August 2012.

11 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig. 7. Relative abundance of corals with weedy, competitive and stress-tolerant life- history-strategies at the two study sites (N = Nursery, T = False Trench) recorded in benthic surveys between March 2010 and August 2012.

Fig. 8. Mean percent cover of the substrata sediment and sand at the two study sites (N = Nursery, T = False Trench) recorded in benthic surveys between March 2010 and August 2012.

12 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

4.1.2 Multivariate community structure

Coral community structure differed among sites (Table 3; Fig. 9). The results of SIMPER analysis show that differences in the abundances of the species Goniastrea australensis, Turbinaria frondens, T. peltata, Acropora solitaryensis and T. radicalis explained over 60% of the dissimilarity among sites (Table 1). Multivariate coral community structure did not separate clearly among seasons and among years (Fig. 9, Table 3).

Fig. 9. nMDS ordination of coral community structure at the Nursery (triangles) and False Trench (circles) sites. Colour codes: red = winter; grey = summer. Blue circles and text = abbreviated coral species names. See Table 1 for full species names.

Table 1. SIMPER analysis for coral communities showing the amount of dissimilarity in coral community structure among the sites Nursery and False Trench explained by individual coral species. Zoogeography (Veron 1993, 2000) and Life-History- Strategies (Darling et al. 2012, 2013) of species are also shown. Mean cover Cumulative Spp. Life-History- Mean cover Species Zoogeography False contribution to abbrev. Strategy Nursery Trench dissimilarity Goniastrea australensis GSAU cosmopolitan stress-tolerant 1.06 2.35 23.4% Turbinaria frondens TUFR cosmopolitan stress-tolerant 0.54 0.98 37.1% Turbinaria peltata TUPE tropical stress-tolerant 0.56 0.00 47.1% Acropora solitaryensis ASOL subtropical competitive 0.45 0.03 55.0% Turbinaria radicalis TURA subtropical competitive 0.35 0.00 61.4% Turbinaria patula TUPA subtropical stress-tolerant 0.34 0.00 67.2% Acanthastrea echinata ACEC tropical stress-tolerant 0.38 0.55 72.9%

13 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Pocillopora aliciae POCA subtropical weedy 0.46 0.48 77.6% Psammacora superficialis PSSU tropical stress-tolerant 0.13 0.25 82.1% Acanthastrea lordhowensis ACLH subtropical stress-tolerant 0.07 0.28 86.0% Favites abdita FTAB tropical stress-tolerant 0.00 0.15 89.0% Goniopora djiboutiensis GPDJ tropical stress-tolerant 0.16 0.00 91.9% Acropora valida AVAL tropical competitive 0.14 0.03 94.2% Turbinaria mesenterina TUME cosmopolitan competitive 0.60 0.70 96.5% Acanthastrea hillae ACHI subtropical stress-tolerant 0.08 0.00 97.9% Plesiastrea versipora PLES cosmopolitan stress-tolerant 0.04 0.00 98.6% Stylophora pistillata STYL tropical weedy 0.06 0.05 99.2% Acanthastrea bowerbanki ACBO subtropical stress-tolerant 0.02 0.03 99.6% Montastrea curta MACU tropical stress-tolerant 0.02 0.00 100.0%

Benthic community structure (i.e. including all taxa from the benthic categories corals, hard corals, soft corals, sponges, other invertebrate taxa and algae) showed some separation among sites, albeit not significant (Table 4, Fig. 10). The results of SIMPER analysis show that over 80% of dissimilarities among sites were due to differences in the abundance of macroalgae and other benthic invertebrate taxa (Table 2). Hard corals contributed less than five percent to mean dissimilarities among sites (Table 2). Multivariate benthic community structure did not separate clearly among seasons and among years (Fig. 10, Table 3).

Fig. 10. nMDS ordination of benthic community structure at the Nursery (triangles) and False Trench (circles) sites. Colour codes: red = winter; grey = summer. Red cross = corals, blue plus sign = other benthic invertebrate, green star = algae.

14 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Table 2. SIMPER analysis for benthic communities showing the amount of dissimilarity in community structure among the sites Nursery and False Trench explained by individual benthic taxa. Cumulative Benthic Mean cover Mean cover Taxa contribution to category Nursery False Trench dissimilarity Red algae other Algae 9.77 22.65 18.8% Turf algae Algae 26.11 14.00 37.5% Dictyotaceae Algae 22.06 19.53 46.8% Barnacles Invertebrate 6.98 1.15 55.0% Sponges encrusting Invertebrate 0.67 6.40 63.2% Soft corals Invertebrate 1.53 6.93 70.9% Amphiroa anceps Algae 4.24 5.70 75.6% Coralline encrusting Algae 3.34 4.83 78.9% Ascidian colonial Invertebrate 0.14 2.40 82.1% Hydroid Invertebrate 0.92 2.28 84.2% Goniastrea australensis Coral 1.06 2.35 86.1% Caulerpa Algae 0.26 1.10 87.4% Ascidian solitary Invertebrate 0.56 1.43 88.7% Sponge submassive Invertebrate 0.98 1.58 89.7% Turbinaria frondens Coral 0.54 0.98 90.8% Turbinaria peltata Coral 0.56 0.00 91.5% Anemone Invertebrate 0.47 0.00 92.2% Urchin Invertebrate 1.07 0.83 92.8%

Table 3. Distance-based redundancy analysis (dbRDA) testing for differences in multivariate coral and benthic community structure among sites (Nursery, False Trench), survey years at the Nursery (2010, 2011, 2012) and seasons at the Nursery (summer, winter). Significant p values (p <0.05) highlighted in boldface type.

Model Df Variance F-value P

Coral community ~ Site 1 0.22519 3.0347 0.04 Residual 5 0.37102 Coral community Nursery ~ Year 2 0.22071 1.7845 0.07 Residual 2 0.12368 Coral community Nursery ~ Season 1 0.07653 0.8571 0.70 Residual 3 0.26787 Benthic community ~ Site 1 0.13169 2.2852 0.17 Residual 5 0.28815 Benthic community Nursery ~ Year 2 0.19148 2.5228 0.07 Residual 2 0.07890 Benthic community Nursery ~ Season 1 0.045279 0.6116 0.56 Residual 3 0.222104

4.1.3 Temporal trends in benthic fauna

Results of statistical tests of temporal changes in benthic fauna were the same when survey data from both sites were pooled compared to when tests were conducted for surveys from the Nursery only. For brevity, only the results for the Nursery data are presented here. No seasonal signal was detected for any of the benthic parameters tested (coral and benthic community structure, Table 3; coral species richness, coral

15 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 cover, algae cover, other invertebrate cover, substrata cover, Table 4). Mean coral cover differed among survey years (Table 4) with a modest increase between 2010 and 2011 and a subsequent decline between 2011 and 2012 (Table 4, Fig. 11a). No accompanying changes in coral species richness were recorded during these times and abundance of algae and other benthic invertebrates remained relatively stable (Table 4). It is of note, however, that the area covered by soft sediment (particularly sand, Fig.8) increased significantly in 2012 (Table 4, Fig. 11b).

Table 4. Analysis of variance (ANOVA) testing for differences in coral species richness, coral over, macroalgae cover, other invertebrate cover and substrata cover among sites (Nursery, False Trench), survey years at the Nursery (2010, 2011, 2012) and seasons at the Nursery (summer, winter). Where the main effect was significant, results of TukeyHSD post-hoc tests are also shown. Significant p values p <0.05 highlighted in boldface type. Sums of Mean Model Df F-value P squares squares Richness ~ Site 1 19.56 19.56 10.52 0.023 Residuals 5 9.3 1.86 Richness Nursery ~ Year 2 6.3 3.15 2.52 0.284 Residuals 2 2.5 1.25 Richness Nursery ~ Season 2 0.3 0.3 0.106 0.766 Residuals 4 8.5 2.833 Coral cover ~ Site 1 0.257 0.2568 0.128 0.735 Residuals 5 10.012 2.0023 Coral cover Nursery ~ Year 2 8.491 4.246 35.35 0.028 Residuals 2 0.240 0.120 2010-2011 0.209 2010-2012 0.045 2011-2012 0.025 Coral cover Nursery ~ Season 1 1.257 1.257 0.504 0.529 Residuals 3 7.475 2.492 Algae cover ~ Site 1 6.72 6.72 0.183 0.687 Residuals 5 183.52 36.7 Algae cover Nursery ~ Year 2 56.25 28.12 0.472 0.679 Residuals 2 119.07 59.53 Algae cover Nursery ~ Season 1 24.14 24.14 0.479 0.539 Residuals 3 151.18 50.39 Invertebrate cover ~ Site 1 125 125 3.444 0.123 Residuals 5 181.5 36.3 Invertebrate cover Nursery ~ Year 2 76.87 38.44 0.779 0.562 Residuals 2 98.67 49.34 Invertebrate cover Nursery ~ Season 1 108.49 108.49 4.854 0.115 Residuals 3 67.05 22.35 Substrata ~ Site 1 201.9 201.89 3.848 0.107 Residuals 5 262.3 52.47 Substrata cover Nursery ~ Year 2 260.78 130.39 247.2 0.004 Residuals 2 1.05 0.53 2010-2011 0.456 2010-2012 0.004 2011-2012 0.003 Substrata cover Nursery ~ Season 1 41.84 41.84 0.571 0.505 Residuals 3 219.99 73.33

16 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig. 11. Results of TukeyHSD posthoc tests of differences in (a) mean coral cover and (b) mean substrata cover between years. Mean differences in cover and 95% confidence intervals are shown for change in cover between years.

17 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

4.2 Fish

4.2.1 Patterns in fish species composition

A total of 199 fish species were recorded during transect surveys at the Nursery and False Trench sites in the period 2010 to 2013, ranging between 46 species recorded at the False Trench site in winter 2010 and 89 species recorded at the Nursery in summer 2012 (Fig. 12). A total of 145 species (72.9%) were tropical, 37 (18.6%) were subtropical, and 17 species (8.5%) were temperate (Fig. 12). Total species richness was generally higher in summer than in winter, but this trend was not significant (Table 5). Species richness was generally variable, but not significantly different among years and survey sites (Table 5). Almost half (45.7%) of the species recorded in transects were predators, followed by planktivores (21.6%), herbivores (15.1%), omnivores (9.1%), corallivores (4%), apex predadors (2.5%), and detritivores (2%; Fig. 13).

Fig. 12. The total number of fish species (species richness) recorded at the Nursery (N) and False Trench (T) sites over time, broken down into the zoogeographic categories tropical (red), subtropical (green) and temperate (blue).

18 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig. 13. The total number of fish species recorded at the Nursery and False Trench sites during surveys between 2010 and 2013, broken down feeding functional groups.

4.2.2 Patterns in fish abundance

Fish abundance was highly variable and did not differ significantly among sites, survey years or seasons (Table 5, Fig. 14). Abundance also fluctuated for zoogeographic groups of tropical, subtropical and temperate species. Tropical species were most abundant at the Nursery up until summer 2012. In winter 2012 and summer 2013 temperate species were most abundant. At the False Trench site tropical species were most abundant in summer and subtropical species were most abundant in winter, with temperate species the least abundant group during both surveys (Fig. 14).

During all nine surveys, fish abundance was dominated by planktivores, (predominantly damselfishes), followed by predators. The predator group mainly included smaller wrasses that predate on micro- and macro invertebrates, not necessarily fishes. In contrast, fish communities harboured relatively low abundances of herbivores and very low abundances of corallivores, detritivores, omnivores and apex predators (Figs. 15 & 16). The black-tipped bullseye (Pempheris affinis), eastern pomfret (Schuettea scalaripinnis), Tarwhine (Rhabdosargus sarba), the neon damselfish (Pomacentrus coelestis) and the sapphire damselfish (Pomacentrus pavo) were the five most abundant species (mean across all surveys).

19 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig 14. Total fish abundance (and SD) over time for the Nursery and False Trench.

Fig 15. Relative abundance of fish zoogeographic groups of tropical (red), subtropical (blue) and temperate (green) species over time for the Nursery (N) and False Trench (T).

20 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig 16. Abundance of fish feeding functional groups for the Nursery (logarithmic scale).

Fig 17. Abundance of fish feeding functional groups for the False Trench site (logarithmic scale).

21 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

4.2.3 Patterns in fish biomass

While patterns in fish abundance are driven by small-bodied species (mainly planktivores, Figs 16 & 17), patterns in fish biomass are driven by large-bodied and often aggregating predatory species, as well as apex predators (e.g. leopard sharks, nurse sharks, wobbegong sharks) that are only supported in low numbers at each site (Figs. 20 & 21). Total fish biomass was significantly higher at the False Trench site than at the Nursery (Fig. 18, Table 5), which was mainly due to high biomass of Rhabdosargus sarba, the Tarwhine, as well as other predators and sharks. Rhabdosargus sarba, a tropical generalist predator, was responsible for the high proportion of biomass of tropical species at the Nursery (Fig. 19). Biomass did not differ significantly among seasons or years (Table 5).

Fig 18. Total fish biomass (and SD) over time for the Nursery and False Trench sites.

Fig 19. Proportion of fish biomass by zoogeographic groups over time.

22 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig 20. Biomass of fish feeding functional groups over time for the Nursery (logarithmic scale).

Fig 21. Biomass of fish feeding functional groups over time for the False Trench site (logarithmic scale).

23 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Table 5. Analysis of variance (ANOVA) testing for differences in fish species richness, fish abundance, and fish biomass among sites (Nursery, False Trench), survey years at the Nursery (2010, 2011, 2012, 2013) and seasons at the Nursery (summer, winter). Significant p values p <0.05 highlighted in boldface type. Sums of Mean Model Df F value p squares squares Richness ~ Site 1 492.1 492.1 3.364 0.109 Residuals 7 1023.9 146.3 Richness Nursery ~ Year 3 532.4 177.48 1.965 0.297 Residuals 3 271 90.33 Richness Nursery ~ Season 1 278.7 278.7 2.655 0.165 Residuals 5 524.7 105 Abundance ~ Site 1 11136 11136 0.172 0.691 Residuals 7 454292 64899 Abundance Nursery ~ Year 3 360612 120204 4.276 0.132 Residuals 3 84334 28111 Abundance Nursery ~ Season 1 39034 39034 0.481 0.519 Residuals 5 405912 81182 Biomass ~ Site 1 2.04E+12 2.04E+12 46.53 0.000 Residuals 7 3.07E+11 4.39E+10 Biomass Nursery ~ Year 3 7.82E+10 2.61E+10 0.393 0.768 Residuals 3 1.99E+11 6.63E+10 Biomass Nursery ~ Season 1 9.50E+09 9.50E+09 0.178 0.691 Residuals 5 2.68E+11 5.35E+10

4.2.4 Multivariate community structure

4.2.4.1 Fish abundance

In terms of fish abundance, spatial and temporal differences in multivariate community structure were overwhelmingly driven by small-bodied planktivorous species. The nMDS ordination in Fig. 22 and statistical tests (Table 10) illustrate clear separation of community structure for the Nursery and False Trench sites, but no interannual or seasonal separation for the Nursery (Table 10). The results of SIMPER analysis shows that over 50% of the dissimilarity among sites for fish abundance were due to greater abundance of Tarwhine (Rhabdosargus sarba), eastern pomfret (Schuettea scalarpinnis), blackfin dartfish (Ptereleotris evides) and smallscale bullseye (Pempheris compressa) at the False Trench, and greater abundance of black-tipped bullseye at the Nursery (Table 6). While there was no clear seasonal pattern in fish abundances (Fig. 22; Table 10) and a large number of species contributed to 80% of the dissimilarity between summer and winter, Table 7 highlights several tropical species that only occurred in summer (black rabbitfish, false fusilier, blue and gold fusilier, yellowtail sergeant) and several species that only occurred in winter (pigmy sweeper, yellowtail kingfish, silver seabream).

24 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Fig. 22. nMDS ordination of fish abundance community structure at the Nursery (triangles) and False Trench (circles). Colour codes: red = winter; grey = summer. Grey crosses = fish species.

Table 6. SIMPER analysis showing the amount of dissimilarity in fish community structure among the sites Nursery and False Trench that is explained by the abundance of individual fish species. Mean Mean Cumulative Species Common name Food group abundance abundance contribution to Nursery False Trench dissimilarity Rhabdosargus sarba Tarwhine Predator 0.4 160.5 17.7% Pempheris affinis Black-tipped bullseye Planktivore 153.8 0.0 31.9% Schuettea scalaripinnis Eastern pomfret Planktivore 38.1 141.8 45.6% Ptereleotris evides Blackfin dartfish Planktivore 0.0 55.0 52.3% Pempheris compressa Smallscale bullseye Planktivore 0.0 35.0 56.0% Pomacentrus coelestis Neon damselfish Planktivore 33.0 6.9 59.2% Pomacentrus pavo Sapphire damselfish Planktivore 27.1 9.0 61.8% Microcanthus strigatus Stripey Omnivore 3.1 21.5 64.4% Abudefduf whitleyi Whitleys sergeant Planktivore 22.1 4.5 66.9% Lutjanus russellii Moses' snapper Predator 0.0 20.0 69.3% Scorpis lineolata Silver sweep Planktivore 2.9 23.3 71.6% Atypichthys strigatus Australian mado Planktivore 22.0 5.0 73.5% Paracaesio xanthura False fusilier Planktivore 7.3 10.5 75.3% Dascyllus trimaculatus Threespot dascyllus Planktivore 12.3 0.5 76.6% Siganus fuscescens Black rabbitfish Herbivore 9.9 0.5 77.8% Parupeneus signatus Blackspot goatfish Predator 12.4 6.7 78.8% Dascyllus reticulatus Reticulate dascyllus Planktivore 8.8 0.0 79.7% Pseudocaranx dentex White trevally Predator 6.8 5.5 80.5% Parapriacanthus ransonneti Pigmy sweeper Planktivore 6.7 0.0 81.3%

25 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Table 7. SIMPER analysis showing the amount of dissimilarity in fish community structure among summer and winter at the Nursery that is explained by the abundance of individual fish species. Mean Mean Cumulative Food Species Common name abundance abundance contribution to group summer winter dissimilarity Pempheris affinis Black-tipped bullseye Planktivore 167.2 135.9 24.3% Schuettea scalaripinnis Eastern pomfret Planktivore 64.4 3.0 32.6% Abudefduf whitleyi Whitleys sergeant Planktivore 1.6 49.4 39.5% Pomacentrus pavo Sapphire damselfish Planktivore 38.8 11.4 44.9% Atypichthys strigatus Australian mado Planktivore 22.9 20.8 49.2% Siganus fuscescens Black rabbitfish Herbivore 17.3 0.0 52.7% Pomacentrus coelestis Neon damselfish Planktivore 35.2 30.1 56.0% Paracaesio xanthura False fusilier Planktivore 12.8 0.0 58.4% Parapriacanthus ransonneti Pigmy sweeper Planktivore 0.0 15.7 60.7% Parupeneus signatus Blackspot goatfish Predator 17.1 6.1 62.8% Dascyllus reticulatus Reticulate dascyllus Planktivore 7.3 10.8 64.8% Pseudocaranx dentex White trevally Predator 3.8 10.8 66.4% Caesio caerulaurea Blue and gold fisilier Planktivore 9.4 0.0 67.9% Seriola lalandi Yellowtail kingfish Predator 0.0 6.3 69.3% Dascyllus trimaculatus Threespot dascyllus Planktivore 14.6 9.3 70.5% Amphipriona kindynos Barrier reef anemonefish Herbivore 7.3 2.3 71.7% Cirrhilabrus punctatus Dotted wrasse Planktivore 1.5 6.7 72.7% Prionurus microlepidotus Sawtail surgeonfish Omnivore 6.4 5.7 73.7% Abudefduf vaigiensis Indo-Pacific sergeant Planktivore 3.4 2.7 74.6% Pagrus auratus Silver seabream Predator 0.0 4.1 75.4% Chromis nitida Barrier reef chromis Planktivore 2.8 0.7 76.1% Macropharyngodon meleagris Blackspotted wrasse Predator 2.4 3.9 76.8% Thalassoma amblycephalum Bluntheaded wrasse Planktivore 1.7 5.2 77.4% Labroides dimidiatus Bluestreak cleaner wrasse Predator 5.1 2.2 78.0% Chaetodon guentheri Crochet butterflyfish Omnivore 4.3 1.2 78.5% Microcanthus strigatus Stripey Omnivore 3.9 2.1 79.1% Scorpis lineolata Silver sweep Planktivore 3.2 2.5 79.6% Abudefduf notatus Yellowtail sergeant Planktivore 2.3 0.0 80.2%

4.2.4.2 Fish biomass

At the biomass level, spatial and temporal differences in community structure were driven by large bodied predatory and apex predatory species. The nMDS ordination in Fig. 23 illustrates that patterns in fish community structure differed among sites (Table 10) and to a lesser extent also among seasons at the Nursery (p = 0.05, Table 10), with winter surveys at the Nursery intermediate between summer surveys at the Nursery and the two surveys conducted at the False Trench site. The results of SIMPER analysis show that differences in community structure among sites for fish biomass were overwhelmingly due to predatory and apex predatory species (Table 8). For example, the Tarwhine, Rhabdosargus sarba, contributed over two thirds of the differences among the Nursery and the False Trench sites. In addition, ornate wobbegong and leopard sharks, red morwong, Moses’ snapper, eastern pomfret and silver sweep contributed to greater fish biomass at the False Trench site. Grey nurse and spotted wobbegong sharks, as well as eastern shovelnose rays, yellowtail kingfish, pink snapper and brown sweetlips had greater biomass at the Nursery

26 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

(Table 8). Seasonal differences at the Nursery were largely due to grey nurse sharks, ornate wobbegongs, yellow kingfish and pink snapper in winter, and greater biomass of spotted wobbegongs, leopard sharks, eastern shovelnose rays and ribbontail stingrays in summer (Table 9).

Fig. 23. nMDS ordination of fish biomass community structure at the Nursery (triangles) and False Trench (circles) sites. Colour codes: red = winter; grey = summer. Grey crosses = fish species.

Table 8. SIMPER analysis showing the amount of dissimilarity in fish community structure among the sites Nursery and False Trench that is explained by the biomass [in g] of individual fish species. Mean Mean Cumulative Species Common name Food group biomass biomass contribution to Nursery False Trench dissimilarity

Rhabdosargus sarba Tarwhine Predator 5,460 1,080,000 67.4% Orectolobus ornatus Ornate wobbegong Apex 13,900 123,000 74.2% Carcharias taurus Grey nurse shark Apex 85,400 0 78.6% Cheilodactylus fuscus Red morwong Predator 2,640 72,500 83.0% Stegostoma fasciatum Leopard shark Apex 17,500 70,100 87.1% Aptychotrema rostrata Eastern shovelnose ray Predator 62,700 0 90.6% Orectolobus maculatus Spotted wobbegong Apex 47,300 41,500 93.3% Seriola lalandi Yellowtail kingfish Predator 21,800 0 94.7% Lutjanus russellii Moses' snapper Predator 0 19,100 95.9% Schuettea scalaripinnis Eastern pomfret Planktivore 456 7,330 96.3% Pagrus auratus Pink snapper Predator 4,680 990 96.6%

27 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

Plectorhinchus gibbosus Brown sweetlips Predator 4,610 0 96.9% Scorpis lineolata Silver sweep Planktivore 341 3,620 97.2% Prionurusmicro lepidotus Australian sawtail Omnivore 5,110 2,060 97.4%

Table 9. SIMPER analysis showing the amount of dissimilarity in fish community structure among seasons at the Nursery that is explained by the biomass [g] of individual fish species. Mean Mean Cumulative Species Common name Food group biomass biomass contribution to summer winter dissimilarity Carcharias taurus Grey nurse shark Apex 0 199,000 24.0% Orectolobus maculatus Spotted wobbegong Apex 79,300 4,670 39.8% Aptychotrema rostrata Eastern shovelnose ray Predator 110,000 0 54.9% Seriola lalandi Yellowtail kingfish Predator 0 50,900 67.5% Orectolobus ornatus Ornate wobbegong Apex 3,700 27,400 75.4% Stegostoma fasciatum Leopard shark Apex 30,600 0 81.5% Pagrus auratus Pink snapper Predator 0 10,900 84.5% Rhabdosargus sarba Tarwhine Predator 3,900 7,530 86.7% Plectorhinchus gibbosus Brown sweetlips Predator 2,860 6,940 88.6% Taeniura lymma Ribbontrail stingray Predator 4,290 0 89.9% Prionurus microlepidotus Sawtail surgeonfish Omnivor 4,100 6,460 90.9%

Table 10. Distance-based redundancy analysis (dbRDA) testing for differences in multivariate fish community structure for abundance and biomass data among sites (Nursery, False Trench), survey years at the Nursery (2010, 2011, 2012, 2013) and seasons at the Nursery (summer, winter). Significant p values (p <0.05) highlighted in boldface type.

Model Df Variance F P

Abundance data Fish community ~ Site 1 0.63474 2.9765 0.037 Residual 7 1.49274 Fish community Nursery ~ Year 3 0.65981 1.037 0.5 Residual 3 0.63626 Fish community Nursery ~ Season 1 0.19061 0.8621 0.54 Residual 5 1.10546

Biomass data Fish community ~ Site 1 0.82394 2.5409 0.031 Residual 7 2.26989 Fish community Nursery ~ Year 3 1.0299 0.8366 0.8 Residual 3 1.231 Fish community Nursery ~ Season 1 0.58881 1.7607 0.054 Residual 5 1.67212

28 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

5 Discussion

Located in the heart of the tropical-to-temperate transition zone in eastern Australia, the rocky reefs of Julian Rocks boast a unique biogeographical overlap of species, many of which are at or close to the edges of their ranges. Transect surveys at two sites at Julian Rocks between 2010 and 2013 have identified 19 species of scleractinian corals, diverse algal and invertebrate communities, and 199 species of fishes. The make-up of the fauna at Julian Rocks is characteristic for biogeographic transition zones, including a very interesting combination of tropical, subtropical and temperate species. Benthic assemblages at Julian Rocks are dominated by macroalgae, contain a high percent cover of non-coral benthic invertebrates and relatively low abundances of hard coral. High percent cover of macroalgae is consistent with other subtropical rocky reefs in eastern Australia (Sommer unpublished data) and elsewhere (e.g. Hawaii, Vroom & Braun 2010). While high macroalgae cover is generally considered undesirable and an indicator of poor health on tropical coral reefs, subtropical reefs tend to naturally feature high abundance of fleshy macroalgae, interspersed with patches of high coral cover (Vroom & Braun 2010).

Over 40% of the coral species recorded in our surveys at Julian Rocks were tropical, including species such as Acropora valida, Favites abdita, Stylophora pistillata, Goniopora djiboutiensis, Montrastrea curta, Acantastrea echinata and Turbinaria peltata. Most tropical coral species were very rare at these sites, with only one or two colonies recorded during the surveys. The bulk of coral abundance was instead made up by cosmopolitan and subtropical coral species, including Goniastrea australensis, Turbinaria frondens, T. mesenterina, Acropora solitaryensis and the subtropical endemic Pocillopora aliciae (Schmidt-Roach et al. 2013) close to the northern limit of its distribution. These patterns of commonness and rarity are consistent with the distributional patterns of tropical, subtropical and cosmopolitan coral species in this zone of biogeographical overlap, where tropical species, although often the most speciose group, tend to be rare and locally distributed (Sommer et al. 2014). The cosmopolitan species Goniastrea australensis, Turbinaria mesenterina and T. frondens, on the other hand, are among the most widely distributed and abundant species in the region, as well as the subtropical endemic Pocillopora aliciae, which reaches great abundance as far south as Port Stephens (32°42’S) (Sommer et al. 2014).

A recent latitudinal study of coral distributions in the subtropical-to-temperate transition zone between the Sunshine Coast (26°36’S) and Port Stephens (32°42’S) identified the Tweed-Byron region as the region along the latitudinal gradient where the relative abundance of coral zoogeographic groups switches from being dominated by tropical species to the north (relative abundance of 50-90% at outer Moreton Bay and Sunshine Coast locations) to less than 20% relative abundance of tropical coral species to the south (except at North Solitary Island, where the relative abundance of tropical corals exceeded 40%) (Appendix C in Sommer et al. 2014). This dramatic decline in the influence of tropical coral taxa in the Cape Byron region is likely linked to a change in oceanographic conditions stemming from topographically induced, EAC driven upwelling of colder and nutrient enriched waters as the EAC accelerates over the narrowing continental shelf at Cape Byron (Oke & Middleton 2000; Roughan & Middleton 2002; Everett et al. 2014). Upwelling- favourable conditions at Cape Byron prevail 60% of the time (Everett et al. 2014),

29 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 influencing thermal and nutrient environments (Oke & Middleton 2000; Everett et al. 2014) and likely also the fauna in the region.

Coral reefs are usually best developed in warm, clear, shallow seas with ample light, low nutrient concentrations, normal oceanic salinity and supersaturation of calcium carbonate levels (Harrison & Booth 2007). Coral communities at Julian Rocks (and in coastal NSW in general) do not form accreting coral reefs and occur in environmental conditions that are considered marginal for coral survival and reef growth (Guinotte et al. 2003). The strong prevalence of upwelling conditions might thus explain the moderate diversity and abundance of scleractinian corals in the Tweed-Byron region, compared with rocky reefs to the north (e.g. outer Moreton Bay) and south (e.g. Solitary Islands). Moreover, coral species of stress-tolerant life-history-strategy, (Darling et al. 2012; Darling et al. 2013), were most speciose and abundant in this study, which is consistent with marginal environmental conditions favouring stress- tolerant corals over more sensitive competitive species in this biogeographic transition zone (Sommer et al. 2014). The stress-tolerant strategy is deemed favourable in chronically harsh conditions and includes slow-growing, broadcast spawning species of predominantly domed morphologies, large corallites and high fecundity (Darling et al. 2012).

Low abundance of most coral species suggests that corals at Julian Rocks rely on dispersal and replenishment from more distant populations on the Great Barrier Reef (GBR) or from other subtropical locations (Sommer et al. 2014). It is possible that in addition to creating marginal environmental conditions for corals in the region, the strong persistence of upwelling conditions (Everett et al. 2014) may disrupt connectivity with these source populations and thereby limit the diversity and abundance of corals at Julian Rocks. Moreover, the temporal persistence of upwelling in the region might help explain the weak to absent seasonal signal identified for fish and benthic communities, respectively, in the underlying study, as well as the high abundance of macroalgae. Along with nearshore rocky reefs in the Solitary Islands, which are also influenced by the intrusion of cooler waters (Malcolm et al. 2011), Julian Rocks features among the highest relative abundance of macroalgae among subtropical eastern Australian rocky reefs (Sommer unpublished data).

Previous transect surveys at 11 Julian Rocks sites in 1994 (Harriott et al. 1999) recorded pooled mean coral cover of 4.8%, which is similar to coral cover of between 2.9% ± 1.31 and 6.8% ± 0.83 recorded in the underlying study. This suggests that coral abundance has fluctuated around these levels for at least the past two decades, with episodic decline and recovery following disturbance events. These patterns are consistent with temporal stability of benthic assemblages at eastern and western Australian high-latitude reefs (Dalton & Roff 2013; Speed et al. 2013). Toppled boulders and coral colonies, as well as widespread coral mortality (Plate 2) and high abundance of sand (Fig. 8) recorded at the Nursery in August 2012 suggest that a storm event may have caused the significant drop in coral cover to 2.9% ± 1.31 (Figs. 4 and 11; Table 4). In combination with the observed low taxonomic richness and abundance of corals at Julian Rocks, and their likely reliance on propagules from core habitats, this highlights severe vulnerability of coral assemblages to local disturbance, such as storm events and breakage by inexperienced SCUBA divers. The use of different survey methods precludes direct comparison of our results with the findings of Hartley and Gorton (2008). Nevertheless, we encountered evidence of

30 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 coral breakage and mortality during our surveys at the Nursery, which was likely caused by contact with novice divers. Consistent with the findings of Hartley and Gorton (2008) breakage was mostly evident for fragile branching and plating species, such as the endemic Pocillopora aliciae, and tropical Acropora corals. The high visitation rates of often inexperienced SCUBA divers at the sheltered Nursery site may therefore depress coral cover at the Nursery and potentially also at other heavily dived sites.

Plate 2. Storm and sediment damaged Turbinaria colony at the Nursery in August 2012.

Fish assemblages at Julian Rocks were characterised by strong tropical influence, with 72.9% of species having tropical affiliation. This is comparatively higher than in the Solitary Islands Marine Park, where 50% of species were tropical (Malcolm et al. 2010). Surprisingly, no seasonal patterns in fish zoogeography were evident from our surveys, with high relative abundance of tropical species also in winter, and generally strong variability among surveys. Abundant fish species were mainly small-bodied planktivores or predators of mobile invertebrates, which is consistent with trophic patterns at higher latitude NSW locations (Curley et al. 2002; Malcolm et al. 2010) and other high-latitude reefs elsewhere (Ferreira et al. 2004). This may relate to increased productivity from the regional upwelling of nutrient enriched waters, which supports thriving populations of planktivores and predators of invertebrates. Despite the high abundance of macroalgae, herbivores only had intermediate abundance, and the large schools of roving herbivores typical of tropical coral reefs, such as acanthurids and scarids (Choat et al. 2004; Ferreira et al. 2004), were absent at Julian Rocks. This is consistent with latitudinal patterns in herbivorous fishes

31 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013

(Meekan & Choat 1997; Cole 2001; Ferreira et al. 2004; Malcolm et al. 2010). Nevertheless, fleshy macroalgae provide habitat for many macro- and microinvertebrates and abundant macroalgae at Julian Rocks may thus indirectly support high prevalence and biomass of predatory species.

In contrast to the dominant influence of small-bodied planktivores on fish abundance patterns at Julian Rocks, large-bodied predator and apex predator species were the main drivers of patterns in fish biomass. Julian Rocks provides important habitat for several species of sharks, including leopard shark (Stegostoma fasciatum), spotted (Orectolobus maculatus) and ornate (Orectolobus ornatus) wobbegongs, and the critically endangered grey nurse shark (Carcharias taurus), for which Julian Rocks provides critical habitat under the NSW Fisheries Management Act 1994. Due to their large body size and their trophic position, apex predators (i.e. sharks) tend to be less abundant, as any reef can only support limited numbers.

Fish biomass was significantly higher at the False Trench site, which may relate to greater habitat complexity conducive to shark resting habitat at that site. Over two thirds of the differences in community structure between the Nursery and False Trench sites in terms of biomass, was due to greater biomass of Tarwhine, Rhabdosargus sarba, at the False Trench site. Tarwhine belongs to the Sparidae (Bream) family and tends to occur on shallow coastal reefs and in estuaries (Edgar 2012), where it feeds on mobile benthic invertebrates, such as molluscs (Froese & Pauy 2013). Tarwhine is an important commercially and recreationally fished species, whose exploitation status is fully fished (DPI 2014). There was a weak (albeit not significant) seasonal signal of fish biomass patterns at the Nursery, with greater biomass of grey nurse sharks (Carcharias taurus), yellowtail kingfish (Seriola lalandi), ornate wobbegongs (Orectolobus ornatus), and pink snapper (Pagrus auratus) in winter; and spotted wobbegong (O. maculatus), eastern shovelnose ray (Aphychotrema rostrata), leopard shark (Stegostoma fasciatum), and ribbontail stingray (Taeniura lymma) in summer.

5.1.1 Outlook and implications for management

Ecosystems at the transition of biogeographical zones are projected to undergo multidimensional changes, both through direct climate change impacts and through indirect impacts from distributional and abundance shifts of species, as well as associated changes in species interactions (Parmesan & Yohe 2003; Poloczanska et al. 2013; Vergés et al. 2014). Recent observations from Australia highlight that tropicalisation has already commenced, with several tropical Acropora coral species shifting their ranges poleward (Marsh 1993; Baird et al. 2012) and tropical fishes surviving in temperate regions over winter when winter temperatures exceed survival thresholds (Figueira & Booth 2010). While these and other tropical species may thus benefit from warming oceans at their southern range edges, temperate species may experience range contractions at their northern limits of their distribution. This is exemplified by an extreme warming episode in Western Australia in 2011, which caused the range contraction of a habitat-forming, temperate seaweed (Scytothalia doycarpa, Smale & Wernberg 2013).

Potential changes of subtropical ecosystems of eastern Australia are closely linked to future changes of the EAC, which has already been strengthening and penetrating

32 Seasonal and temporal changes in reef fauna at Julian Rocks, 2010 to 2013 further south (Cai et al. 2005; Ridgway 2007; Wu et al. 2012), with a trend for warmer winter sea surface temperatures of 0.7 to 1.5°C for locations between the Solitary Islands Marine Park (30°S) and Merimbula (36°53’S) between 1870 and 2000 (Figueira & Booth 2010). Nevertheless, due to a complex interplay of EAC flow, eddies and upwelling, changes along the eastern Australian coast will likely be heterogeneous. For example, strengthening of the EAC could increase the nearshore incursion of cold and nutrient-rich waters in upwelling areas (Roughan & Middleton 2004; Everett et al. 2012; Everett et al. 2014). As the topographically induced upwelling in the Cape Byron region is linked to variability in the strength of the EAC, it is thus likely that intensified EAC flow will affect upwelling patterns and cause increased surface concentrations of chlorophyll a in the region and possibly greater poleward advection (Everett et al. 2014). While this may constrain tropicalisation in the region relative to localities outside of upwelling areas, it could enhance productivity and biological significance for subtropical and temperate taxa.

Upwelling areas have been proposed as possible refugia from thermal stress in warming oceans (West & Salm 2003), and although protective capacity needs to be assessed for each upwelling area individually in terms of the temporal variability of warming and upwelling (Chollett et al. 2010), the high proportion of upwelling- favourable conditions at Cape Byron – in fact, the highest proportion between 28º and 37.5ºS latitude (see Table 1 in Everett et al. 2014) – may confer particular conservation significance to this region as a thermal refuge for tropical, subtropical and temperate species, as surrounding waters become warmer. Julian Rocks currently provides critical habitat for the critically endangered grey nurse shark (Carcharias taurus, NSW Fisheries Management Act 1994), and important habitat for many commercially and recreationally fished species, including Tarwhine (Rhabdosargus sarba), pink snapper (Pagrus auratus), and yellowtail kingfish (Seriola lalandi), as well as scleractinian corals at or close to their southern range limits. The drop in coral cover in August 2012 highlights the vulnerability of fauna at Julian Rocks to disturbance and underscores the need to minimise existing stressors (e.g. from habitat degradation, fishing) both locally and also of potential source populations, to enable them to withstand and recover from disturbances. Stringent protection of these habitats is therefore the most likely avenue to enhancing their persistence, and their suitability as potential refugia under climate change (Beger et al. 2014).

6 References

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