A Study on the Recovery of Tobago's Coral Reefs Following the 2010 Mass Bleaching Event

A study on the recovery of Tobago's coral reefs following the 2010 mass bleaching event

Salome Buglass

BSc. Geography, University College London, United Kingdom, 2009

A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF

THE REQUIREMENTS FOR THE DEGREE OF

Master of Science

THE FACULTY OF GRADUATE AND POSTDOCTORAL STUDIES

(Geography)

The University of British Columbia

(Vancouver)

December 2014

The rise of ocean temperatures globally has become a grave threat to coral reefs, as it is increasing the severity and frequency of mass coral bleaching events and post-bleaching coral mortality. The continued existence of productive coral reefs will rely on corals’ ability to undergo recovery. In 2010, Tobago’s coral reefs were exposed to severe heat stress leading to mass bleaching of up to 29-60% of colonies at observed sites. This study evaluated the impact of coral bleaching and recovery of coral communities across three major reef systems in Tobago that differ in their exposure to terrestrial runoff. Assessments were done on the 1) density and composition of coral juveniles to characterise the levels of recruitment, 2) sedimentation rates and composition to understand its potential impact on recovery, and 3) species’ size frequency distributions in 2010, 2011 and 2013 to examine temporal changes among population size structure.

In 2013, low juvenile densities were observed (5.41 ± 6.31 m-2) at most reef sites, which were dominated by brooding genera while broadcasting genera were rare. Sediment material, measured in May and June (end of Tobago’s dry season) was mostly terrigenous and deposited at rates below coral stress threshold levels at most sites. Out of 27 species populations assessed between all sites, 4 populations mean colony size had significantly changed by the bleaching event, and only changed 5 populations over the two following years. The few populations that were significantly altered (mainly S. siderea and M. faveolata) after the bleaching saw a rise in small sized colonies, mostl likely as a result of colony fragmentation.

This study highlights that recovery via sexually produced recruits among broadcasting species was limited. While sedimentation rates were low, it is likely they are significantly higher throughout the rainy season, thus a long-term sedimentation study is highly recommended. Most coral populations resisted significant alteration from heat stress in 2010. However, given that future thermal stress is projected to become more intense, this study shows that mass bleaching disturbance could lead to decline coral population’s mean colony size, which could affect coral recovery as smaller colonies are less fecund.

Preface

This thesis is based on field data collected in 2013 in collaboration with the Institute of Marine Affairs (IMA) under the guidance and supervision of Dr. Simon Donner. All data from the field were collected by myself with the assistance of volunteers. Benthic percent cover and coral colony size frequency data for 2010, 2011 and 2012 were previously collected by the IMA as part of their Biodiversity and Ecology Research Programme. Laboratory sample processing was performed partly in the IMA’s Biodiversity and Ecology Laboratory and on UBC Vancouver campus. I undertook all laboratory sample processing, analysis of data, and writing of the thesis manuscript.

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Table of Contents

Abstract ...... ii

Preface ...... iii

Table of Contents ...... iv

List of Tables ...... vi

List of Figures ...... vii

Acknowledgements ...... viii

Chapter 1. Introduction ...... 1

Chapter 2. An assessment of coral juvenile community and sediment deposition among three major reef systems in Tobago ...... 4

2.1 Introduction ...... 4

2.2 Methods ...... 6

2.2.1 Study area ...... 6

2.2.2 Sedimentation assessment ...... 8

2.2.3 Juvenile community survey ...... 10

2.2.4 Statistical analysis ...... 11

2.3 Results ...... 11

2.3.1 Juvenile density and composition ...... 11

2.3.2 Characterization of sedimentation ...... 14

2.4 Discussion ...... 16

Chapter 3. Using coral size distribution to assess the recovery from mass bleaching in the southern Caribbean ...... 21

3.1 Introduction ...... 21

3.2 Methods ...... 23

3.2.1 Benthic cover survey ...... 24

3.2.2 Colony size frequency survey ...... 25

3.2.3 Statistical analysis ...... 25

3.3 Results ...... 26

3.3.1 Changes in percent coral cover ...... 26

3.3.2 Changes in coral population structure and community composition ...... 27

3.4 Discussion ...... 32

Chapter 4. Conclusion ...... 36

Bibliography ...... 39

Appendices ...... 46

List of Tables

Table 1. Information on the six reef sites where juvenile assessment was undertaken and sediment traps were installed ...... 8

Table 2. Sieved size groups and sediment classes according to the Wentworth size class system...... 10

Table 3. Coral juvenile data per site...... 11

Table 4. Number of juvenile taxa found at each site...... 12

Table 5. Mean and standard deviations of sediment measurements ...... 15

Table 6. Mean percent cover of live coral and (±) standard deviation estimated at each site and year ...... 27

Table 7. Total number of coral colony and species recorded per site ...... 27

Table 8. Summary of the colony abundance and size data collected for each dominant species present at each reef sites...... 46

Table 9. Significant comparison of size frequency distributions and colony size between years (2010, 2011 and 2013) determined using Kolmogorov-Smirnov test (KS) and Kruskal-Walis (KW) test respectively...... 47

Table 10. Significant comparison size frequency distributions between reef sites determined using Kolmogorov-Smirnov test (KS) test respectively...... 48

List of Figures

Figure 1. Map of Tobago and location of studied reef systems and site ...... 7

Figure 2 Proportion of counted juveniles that were produced from broadcasting or brooding reproductive strategies ...... 12

Figure 3. Relative abundance of major taxa (genus) groups at each site for (A) juvenile population based from count data and (B) adult population based on percent cover assessed in 2013 (see Chapter 3 for data collection methods) ...... 13

Figure 4. Correspondence analysis (CA) biplot ...... 14

Figure 5. Boxplot of sediment accumulation rates per site ...... 15

Figure 6. Stacked barplots of average percent of sediment (left) composition and (right) particle size distribution from sediment data collected May-June 2013 ...... 15

Figure 7. Map of Tobago and location of studied reef systems and sites ...... 24

Figure 8. Size frequency distributions of coral taxa at sites with significant differences between years, as determined by the Kolmogorov-Smirnov test ...... 29

Figure 9. Boxplot and mean size (white filled dots) per species per site indicating changing trends in colony size between each year...... 30

Figure 10. Non-metric multidimensional (NMDS) scaling using Bray-Curtis dissimilarities plot of the qualitative changes among the coral communities at each of the sites per year (named and colour coded) across Tobago...... 31

vii

Acknowledgements

I am really thankful to the many people who supported and inspired me while undertaking my master’s degree and writing up my thesis. I especially thank my supervisor Dr. Simon Donner for taking me on as his student and for his sound encouragement and guidance throughout the entire process. I am also grateful for the time and support I received my committee member Jennifer Williams. Additionally, I am very thankful to Jahson Alemu I for assisting me in field and providing me with key information about Tobago’s coral reefs.

My gratitude is also extended to my many fellow geography students and friends for their help and support, especially Leonora King, Christopher Quick, Lawrence Bird and David West. Thanks also go to my father David Buglass and Giordano Mitchell for proofreading. Further thanks go to the staff, students and faculty of the Geography Department, particularly for the support by Suzanne Lawrence, Sandy Lapsky and Stefanie Ickert. Particular thanks go to my family and friends who have been so encouraging and supportive throughout my years as a graduate student.

Finally this work would not have been possible without the funding from TerreWEB, BRITE, and from all those individuals who generously donated to the crowdfunding campaign to help finance my field work.

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Chapter 1. Introduction

Caribbean coral reefs represent less than 10% of all the world’s tropical coral reef systems, nonetheless they sustain critical habitats for the region’s marine biodiversity and provide vital goods and services for over 43 million people (Wilkinson & Souter 2008). Many coral reefs in this region, however, are at high risk of being degraded (Burke et al. 2011). Since the 1970s many Caribbean reefs have experienced unprecedented levels of decline in coral cover, from about 50% to 10%, and are instead turning into algal dominated environments (Gardner et al. 2003; Roff et al. 2011). This ecological deterioration has been attributed to the reduction of herbivory, due to overfishing and the regional die-off of grazing urchins in the 1980s, increased terrestrial runoff, marine pollution and disease outbreaks (Hughes 1994; Jackson et al. 2014). Furthermore, increased warming of ocean temperatures, driven by anthropogenic climate change in the last two decades, has increased the frequency of mass coral bleaching events, which are often followed by significant coral mortality. Consequently, these heat stress events have exacerbated the decline of coral communities and pose a grave threat to the already fragile coral reef ecosystems of this region.

Considering that the frequency and intensity of mass coral bleaching events are likely to increase in the near future (Donner et al. 2005; Hoegh-Guldberg et al. 2007), the post-bleaching recovery of schleractinian coral population is critical to the survival of productive coral reefs in the Caribbean. Following post-bleaching mortality, ideally, corals progressively recover to their pre-disturbance state (Gilmour et al. 2013). Alternatively, however, the coral community composition changes, due to differential bleaching impact and reproductive success among coral species (Shenkar et al. 2005; Obura 2005), or reefs become colonized by algae and sponges due to corals inability to undergo recovery (Norström et al. 2009; McClanahan 2000). Post-disturbance recovery relies on coral communities re-growing and colonizing the reef via sexual recruitment (Pearson 1981). In turn this process is determined by the diversity, abundance and size of surviving coral colonies, their ability to grow and reproduce, their ability of larva to settle and survive, as well as the post-settlement survival of recruits (Tamelander 2002; Baker et al. 2008; Crabbe 2009). Additionally, recovery can also be very site specific due to secondary disturbances affecting coral community dynamics, especially in the case of disturbances that undermine coral reproductive processes such as terrestrial runoff and overfishing (Burt et al. 2008). Consequently, the extent and direction of coral recovery is not easily predictable. Thus, to improve our understanding of post-bleaching recovery it is important to assess the bleaching impact on coral assemblages and their ability to sexually

1 reproduce within their given environment (Irizarry-soto & Weil 2009; Birrell et al. 2005; Smith et al. 2005).

In this thesis, I focus on understanding post-bleaching recovery among the most southern Caribbean coral reef systems that fringe the island of Tobago. Tobago’s coral reefs have undergone the same degradation trajectory as the majority of their Caribbean counterparts. Nor were they spared from undergoing mass bleaching during the Caribbean wide ocean thermal stress events in 1998, 2005 and 2010 (M. Eakin et al. 2010). In 2010 Tobago’s reefs were reported to undergo severe bleaching of up to 29-60% of colonies at observed sites (Alemu I & Clement 2014). Though bleaching induced mortality was estimated at only 2-8% of corals (Alemu I & Clement 2014), a disease outbreak recorded following the bleaching likely lead to further mortality (Alemu I 2011). Currently, little is known about the post-bleaching impact or recovery process among Tobago’s coral communities. Most data on the health and disturbance history of Tobago’s coral reefs is related to changes in percent benthic cover, with the exception of one study done on recruitment (on artificial substrate) and growth modelling (Mallela & Crabbe 2009). Furthermore, due to continuous coastal development taking place along the island’s south-western coast, concerns have arisen about the impact of increased terrestrial runoff on some of the island’s coastal ecosystems (Parkinson 2010; Lapointe et al. 2010; Mallela et al. 2010). Undoubtedly, these terrestrial runoff flows may also shape the recovery trajectory of some of the island’s coral reefs.

This thesis explores the impacts on and recovery of coral reef communities across three distinct reef systems adjacent to land characterized by different levels of settlement and infrastructure development. The first objective (Chapter 2) is to quantify the juvenile communities across these different reef systems, as a high abundance of coral juveniles indicates that coral populations are able to reproduce sexually and recruit, and thus able to support recovery processes (Arnold 2011; Huitric & Mcfield 2000; Ritson-williams et al. 2009). Given that high levels of sedimentation impede coral growth, reproduction and recruitment (Fabricius 2005; Miller et al. 2000; Wittenberg & Hunte 1992) a second objective is to investigate the rate and composition of sedimentation across the three distinct reef systems (Chapter 2). The final objective of this study is to assess the impact and recovery of the bleaching event on species’ population demographics by examining the changes in the size and frequency distributions of species’ populations before and after the bleaching event and in the two years following the bleaching.

2 Tobago’s coral reefs represent some of the most understudied reefs among Caribbean (Alemu I & Clement 2014). This study is one of the few in the region that is able to examine temporal changes following a mass-bleaching event. It provides valuable baseline datasets and analysis of juvenile communities, sediment composition and deposition rates, and corals’ population size structure among some of Tobago’s key reef systems. Additionally, this research supports a growing body of research using population size structure to research effects of disturbance on coral population dynamics.

3 Chapter 2. An assessment of coral juvenile community and sediment deposition among three major reef systems in Tobago

2.1 Introduction

Tropical coral reefs are highly productive ecosystems that act as important natural assets, which provide vital goods and services to coastal communities. This is especially true for small island states in the Caribbean, such as Tobago, which rely heavily on healthy coral reefs for protecting coastal property from storms, supporting near-shore fisheries and attracting tourism (Burke et al. 2008). Despite their value, in the last three decades Caribbean reefs have experienced a dramatic decline in coral cover, from approximately 50% to 10%, and consequently have gone from being coral-dominated to algae-dominated reefs (Côté & Darling 2010). This shift has been attributed to historical overfishing leading to the reduction of large sized grazing herbivore fish populations, die-off of grazing Diadema urchins in the early 1980s, increased sedimentation and nutrient enrichment, and the spread of coral disease (Hughes & Connell 1999; Norström et al. 2009; Jackson et al. 2014). In the last two decades, mass bleaching events are often followed by substantial coral mortality, have become an additional threat, further testing the resilience of the remaining Caribbean coral communities (Hoegh- Guldberg 1999; Eakin et al. 2010). Bleaching events are irrefutably linked to ocean warming as a result of global climate change, and there is strong evidence that the frequency and intensity of mass coral bleaching events are likely to increase in the near future (Donner et al. 2005; Hoegh- Guldberg et al. 2007).

Coral communities’ ability to recover following disturbances like bleaching events will determine the long-term survival of these ecosystems. Maintenance of coral populations depends critically on corals' sexual production of larvae, recruitment and capability of recruits to survive and grow into adult colonies (Hughes & Tanner 2000; Arnold 2011). A variety of biotic and abiotic factors across time and space can impact the success of recruitment and post- settlement survival, including the size and health of parent colonies, availability and complexity of substrate, competition, predation and light availability (Sammarco 1985; Tamelander 2002; Babcock & Smith 2000). Coral larvae and recruits thus tend to be very susceptible to mortality (Trapon et al. 2013). In comparison to Indo-Pacific reefs, post-disturbance recovery of Caribbean coral reefs has generally been low (Baker et al. 2008). According to the literature this disparity is driven by high algal cover and land-based marine pollution, making Caribbean reefs

4 hostile environments for recruitment and juvenile survival (Arnold & Steneck 2011; Jackson et al. 2014).

Increased terrestrial runoff along Caribbean coastlines, has been a growing problem in a region where continuous coastal development and agriculture intensification is taking place (Burke et al. 2011; Begin 2012; Hernandez et al. 2009). Regardless of whether sediments settle or remain in suspension, they can affect all growing stages of a coral’s lifecycle. Turbidity decreases light penetration, reducing coral growth (Cortes & Risk 1985; Fabricius et al. 2003), whilst particles deposited on coral colonies weakens their health and in excess can smother corals (Hernandez et al. 2009; Erftemeijer et al. 2012). Corals’ reproductive and recruitment stages are believed to be especially sensitive to sedimentation levels, as reefs with high sedimentation levels have low recruitment and juvenile densities (Fabricius 2005). It is likely that larvae and recruits are very susceptible to being smothered or damaged by sediment, and particles covering hard surfaces inhibit the settling of larvae (Babcock & Smith 2000; Torres & Morelock 2002; Fabricius 2005). Additionally, recent studies indicate that sedimentation suppresses herbivory and promotes the growth of algae, thereby reducing settling space for recruitment (Goatley & Bellwood 2013). Sedimentation-based stress on coral depends on multiple physical factors such as the timing and amount of sediment deposition, reef depth and distance to shore, ocean currents and waves, and sediment grain size and composition (Hernandez et al. 2009; Abdullah et al. 2011; Waheed et al. 1998). The size of sediment particles is a particularly important property as it determines the transport mode and potential impact on coral communities. For instance, corals can remove sand-sized grains with more ease than very fine sediment like silt/clay (Weber et al. 2006). Fine sediment also tends to carry greater concentrations of toxic contaminants that can be lethal to corals (Fabricius et al. 2003; Rogers 1990).

Most of Tobago’s reefs share the same history of degradation as the rest of Caribbean and thus are characterized by low coral cover that ranges from 10-30% (Mallela et al. 2010). Recently there have been rising concerns that increased terrestrial runoff, due to continuous urban developments, may play an important role in shaping the trajectory of some the island’s coral reefs (Parkinson 2010; Lapointe et al. 2010; Mallela et al. 2010). Data about sedimentation on Tobago’s reefs is limited to one study, however , which found sediment rates along 11 separate reef systems to be all below levels that tend to stress corals (Mallela et al. 2010). No further investigation was conducted into the sediments’ particle size distribution or composition, which can provide key information such as the origin of sediments (Begin 2012)

5 Tobago’s reefs were affected by the regional mass bleaching events in 1998, 2005 and most recently in 2010. The heat stress in 2010 lasted almost 5 months, during which 29 to 60% of coral communities among Tobago’s reefs experienced severe bleaching. While bleaching induced mortality was about 2-8% (Alemu I & Clement 2014), the bleaching event was followed by a disease outbreak possibly causing further mortality (Alemu I 2011). Knowledge about coral recruitment processes on Tobago’s coral reefs comes exclusively from studies assessing recruitment rates on artificial substrate, which quantify the early settling stages of coral recruits (Babcock & Smith 2000). These studies have indicated that the number of recruits settling on some of Tobago’s reefs has decreased over last two decades (Mallela & Crabbe 2009). However no research has yet focused on the early life-stages of corals on natural substrata, which can be done by assessing the juvenile populations. As juveniles are 4-7 years old, they serve as a proxy measurement of the integrated outcome of corals sexual reproduction, larval settlement, and recruits post-settlement survivorship within a multiple year time period (Vermeij et al. 2011)

Given that both bleaching and disease have negative impacts on the fecundity of coral colonies (Weil & Vargas 2009) and that recruitment and post settlement survival often can be adversely impacted by sediment deposition, the objectives of this study were to: (1) quantify juvenile densities and taxa composition, and (2) estimate the rate and composition of sedimentation across three distinct reef system adjacent to land with different levels of urbanization. 2.2 Methods

2.2.1 Study area

Tobago is a 300 km2 large hilly island of volcanic origin that is surrounded by fringing shallow reefs. These reefs evolved under the influence of nutrient and sediment rich inputs from the Orinoco and Amazon Rivers, and consequently Tobago’s coral communities have lower species diversity in comparison to other Caribbean reefs (Moses & Swart 2006; Lapointe et al. 2003; Potts et al. 2004). Whilst the majority of Tobago is covered in forest and shrubs lands, the south-western part of the island has undergone significant urbanization and agricultural development. The study was conducted on three major reef systems, chosen for local perceived importance and their potentially different exposure to sedimentation due to coastal land uses. These included Caribbean Sea facing Buccoo and Culloden Reef, and Atlantic Sea facing Speyside Reef (Figure 1).

Figure 1. Map of Tobago and location of studied reef systems and sites

Buccoo Reef is comprised of five large, sloping reef flats covering about 7 km2 and is Tobago’s only official marine protected park (since 1973). Despite being a major economic asset, attracting over 10,000 visitors annually, this park has received little protection or safeguarding against land-based pollution (Lapointe et al. 2010). In the last three decades, Land adjacent to this reef has experienced rapid urbanization and untreated sewage and uncontrolled storm waters drain into Buccoo Bay (Potts et al 2004, Lapointe et al. 2010, Parkinson 2010). The horseshoe shaped reef of Culloden covers ~5.8 ha and is located in a remote bay surrounded mostly by forested hills (Laydoo 1991). The bay is accessible via a dirt road and human activities are limited to occasional recreational divers, artisanal fishing and boat anchorage. Speyside features a large network of fringing reefs along small islands and rocky outcrops on the north- eastern side of the island. Like Culloden, Speyside coastal lands remain relatively undeveloped comprising of a hilly forested landscape apart from Speyside village (a fishing community) and 2 medium-sized hotels.

7 Table 1. Information on the six reef sites where juvenile assessment was undertaken and sediment traps were installed between 8-12m depths

Reef system/site Coordinates %Coral cover / Distance to shore and potential sources of Dominant coral taxa sediment (Parkinson 2010)

Buccoo Reef Outer Buccoo (OB) 11°11.371’ N 20.24±5.83 ~1.5km from mainland. Two streams and 60°49.412’ W M.faveolata, S. siderea, urban wastewaters discharge in Buccoo bay C.Natans, Agaricia spp. (Parkinson 2010). Western Buccoo (WB) 11°11.043’ N 13.64±8.38 60°50.782’ W M.faveolata, D.strigosa, C.natans, S. siderea Culloden Reef Culloden East (CE) 11°14.833’ N 12.19±5.23 ~280m from mainland. One stream terminates 60°45.086’ W M.faveolata, in this bay. Dirt road access to the bay. M.cavernosa, A. palmate, D. strigosa Culloden West (CW) 11°14.982’ N 16.02±7.57 60°44.904’ W M. faveolata, D. strigosa, C.natans, M. cavernosa Speyside Reef Black Jack Hole (BJH) 11°18.072’ N 13.01±7.12 ~1km from mainland. Paved road all along the 60°31.223’ W M.faveolata, S.siderea, coast line. Doctors River and several small P.astreoides, M.alcicornis streams drain into Speyside bay. Patches forest have been Angel Reef (AR) 11°17.688’ N 46.42±14.67 increasing cleared for growing crops 60°30.001’ W M.mirabilis, M. faveolata, M.cavernosa, C.natans

Two study sites were established at each reef system (Figure 1, Table 1). At the six reef sites juvenile population and sedimentation assessments were carried out from May to June in 2013 using SCUBA. All data were collected at the six sites between the depths of 8-12m during the months of May and June 2013, the end of the dry season in Tobago. 2.2.2 Sedimentation assessment

Sediment traps were deployed for roughly one month, following published methods (English et al., 1977, Hill and Wilkinson 2004, and Storlazzi et al 2009). Each sediment trap consisted of three cylindrical 5-cm-wide and 20-cm-long PVC pipes with wire mesh at the top to allow for sediment deposition but deter large marine organisms. Zip ties were used to fix three pipes onto a 1 m tall metal rod to form a trap set. At each site three sediment trap sets were hammered into non-living reef substrate, leaving the pipe traps about 0.75 m above the substrate. Traps were spaced ~ 30 m from each other, maintaining adequate coverage of the reef site.

8 From the 14th to 26th of May 2013, three trap sets (9 tubes) were set up at each site – a total of 54 individual tubes. After 30-37 days (20-27th of June 2013) the sediment traps were recovered by capping each pipe underwater before bringing them up to the surface. A total of 45 tubes were recovered. Only 6 and 5 tubes were recovered safely from Culloden East and West, respectfully, as the rest were either dislodged or disappeared. After collection the content from each tube washed out with distilled water and filtered through a funnel and filter paper. Sediment samples were dried in an oven at 60°C for three hours, allowed to cool and were sealed in zip-lock bags for transport to Trinidad. At the laboratory the sediments samples were rinsed out twice with distilled water to remove salts and oven-dried at 105°C overnight. The dry sediment was transferred from filter paper into a petri dish with a fine brush and kept in desiccator overnight. The total weight of the sediment collected from each pipe trap was determined using an analytical balance. Sedimentation rates (mg cm-2d-1) were determined by dividing the dry weight (in mg) by the area of the sediment trap aperture width (in cm2) over the duration (in days) the sediment trap was installed at each site (Abdullah et al. 2011).

Sediment composition was analyzed employing loss on ignition (LOI) methods to determine what fraction of the sediment was composed of organic and carbonate matter, leaving the remaining non-carbonate material as the terrigenous fraction of the sample. It is important to note that this method is predominantly used by paleolimnologists on lake core samples and is known to provide rough estimates of sediment composition (Santisteban et al. 2004). As many individual pipe traps contained <2 grams of sediment, samples from each trap set were pooled to form a composite sample. About 3 grams of each composite sediment sample were placed in a pre-weighed ceramic crucible. To determine dry sediment weight all samples were dried in an oven at 105°C for three hours (to remove moisture) and cooled in a desiccator before weight was recorded. Afterwards organic matter was combusted from the samples by placing them in a furnace at 550°C for six hours, cooling in a desiccator and then weighed (Brooks et al. 2007). This procedure was followed by ashing samples at 925°C to determine carbonate content (Luczak and Kupka 1997). The percent of organic and carbonate content was calculated using the following formula (Heiri et al. 2001):

Organic content (%) = ((DW105 – DW550)/DW105)×100 �� Carbonate content (%) = ((DW550– DW950)/DW105) × ! ×100 ��! Where DW105 is the dry weight of the sample before combustion, DW550 is the dry weight after combustion at 550°C, DW950 is the dry weight after combustion at 925°C and CO3/CO2 is the ratio between the molecular weights of CO3 (60u) and CO2 (44u) being 1.36 (as cited in

9 Veres 2002). A total of 16 composite sub-samples were ashed. Particle size analysis was conducted using the wet sieving method (Syvitski 2007), using ~0.5 grams from each composite sediment sample, which were separated into five fractions (Table 2). Before sieving, samples were dried for 3 hours to remove all moisture and weighed.

Table 2. Sieved size groups and sediment classes according to the Wentworth size class system.

Grain size groups sieved Sediment type >500 μm coarse sand 500-250μm medium sand 250-125μm fine sand 125-63μm very fine sand <63μm silt-clay

Dry samples were then emptied into a glass beaker filled with distilled water and left for an hour for the sediment particles to disaggregate. The content was then poured through the stacked sieve set, in the order of mesh sizes 500, 250, 125 and 63 microns. Care was taken to ensure that no particles remained in the beaker by washing with a fine tip squeeze bottle. Each sieve was carefully separated, placed in a ceramic beaker to retain the washings, and set to dry in an oven for 3 hours at 100°C until all water was evaporated. Once dry, the sediment particles were moved from the sieves into the ceramic dish using a fine brush to ensure every particle was retained. The weight of each sediment sample fraction was determined using a high precision analytical balance. 2.2.3 Juvenile community survey

Juvenile coral colonies, hereafter referred to as juveniles, were enumerated and identified by carefully scrutinizing the reef benthos in sixty randomly-placed 0.25m² quadrats per site (Carpenter & Edmunds 2006). Quadrats were dropped from a height of 3m above the substrate while ensuring that no quadrats overlapped, as per McClanahan (2000). A pilot survey of this method was undertaken in advance to determine sample sizes by plotting the running mean against the number of quadrats surveyed (Edmunds et al. 1998). Juveniles were defined as any colony of large-sized coral taxa visible to the naked eye with a maximum diameter of 5 cm (e.g., Montastraea spp., Diploria spp., and Siderastrea spp., etc.) or of 2 cm in diameter for small sized coral taxa (Porites spp., Favia fragum, and Agaricia spp., etc.). This distinction was made as small-sized corals tend to be sexually mature adults once larger than 2 cm (Chiappone & Sullivan 1996; Miller et al. 2000; Irizarry-soto & Weil 2009). Small living coral fragments that are not a product of sexual reproduction were omitted from the count (Trapon et al. 2013). Juveniles were identified to the species level when possible and to the genus when they were

10 taxonomically difficult to distinguish. Care was taken to inspect cryptic habitats such as crevasses and underneath shelves. If more than three quarters of the benthic cover within the quadrats was covered by non-settling substrate, e.g. sand or living coral, the quadrat was moved to the side or the aforementioned process was repeated (Edmunds et al. 1998). 2.2.4 Statistical analysis

All summary statistics and analyses were done in R version 2.15.1. All juvenile and sediment data were tested for normality using the Shapiro-Wilk test and homogeneity of variance using graphical methods. Sedimentation rate data were found to be normally distributed, although juvenile data did not follow a normal destruction. Differences in sedimentation rates per site were tested employing a one-way ANOVA, followed by a Tukey HSD post-hoc test to detect pairwise differences. Differences in juvenile densities per site were tested using the non- parametric methods. A Kruskal-Wallis test was employed followed by Dunn’s post-hoc pairwise test (using packages multcomp 1.2-17 and coin 1.0-23) (Miller et al. 2000). Correspondence analysis (CA) was performed on eight genus groups found at each site, after removing all rare taxa, to explore the distribution of species composition across all sites (Irizarry-soto & Weil 2009) using ordination analysis tools in the Vegan Package in R software. 2.3 Results

2.3.1 Juvenile density and composition

A total of 428 of juveniles between the sizes of ≥0.5 to ≤5 cm were counted over a total area of 90 m2 (15 m2 at each site). Overall, mean juvenile density was 5.4± 6.3 m-2 and ranged from 0 to 32 m-2. The juvenile distribution was patchy with 20-49% of the 0.25 m2 quadrats featuring no juvenile colonies across the assessed sites. The Buccoo Reef sites had the lowest abundance of juveniles, number of taxa and diversity (Table 3).

Table 3. Coral juvenile data per site (15m2 surveyed area).

Site No. of juveniles Density(±SD)m-2 No. of taxa Shannon H' Outer Buccoo 59 3.9± 4.8 5 1.8 Western Buccoo 50 3.3± 4.2 6 3.7 Culloden East 64 4.3± 4.2 7 4.2 Culloden West 99 6.6± 7.7 8 5.6 Black Jack Hole 69 4.6± 5.9 8 5.4 Angel Reef 146 9.7± 7.9 10 4.0

11 Whilst Angel Reef had the highest abundance and richness in taxa, Culloden West and Black Jack Hole had the most even diversity according to the Shannon Diversity Index. Mean juvenile abundances were similar between most sites except at Angel Reef, where abundances were significantly higher when compared to all other sites (Dunn’s test, p < 0.05) except at West Culloden.

Table 4. Number of juvenile taxa found at each site.

Taxa/site OB WB CE CW BJH AR Agaricia sp. 51 24 34 26 1 85

Colpophyllia sp. - 2 - - - - 160 Reproducon mode Diploria sp. - 7 8 16 15 1 140 Broadcast

Eusmilia fastigiata - 1 - - - - 120 Brooder Favia fragum - - 6 1 8 3 100 Madracis decatis - - 4 3 4 5 80 Madracis mirabillis - - - - - 27 60 Montastraea cavernosa - - 1 5 2 - 40 Montastraea sp. 2 - - - 5 5 Total juveniles counted 20 Mycetophyllia sp. 1 - - - - - 0 Porites astreoides 1 - - 3 27 8 OB WB CE CW BJH AR Porites porites - - - - - 1 Scolymia sp. - 2 9 19 - 2 Figure 2 Proportion of counted juveniles that Siderastrea sp. 4 14 2 26 7 9 were produced from broadcasting or brooding Total 59 50 64 99 69 146 reproductive strategies

The majority of the juvenile colonies belonged to brooding taxa (72.9%), such as from the genera Agaricia, Porites, Madracis, Scolymia, and Favia. Broadcasting juvenile taxa represented the minority (27.1 %) such as Siderastrea, Diploria, Montastrea and Colpophyllia. Brooding taxa dominated at Outer Buccoo, Culloden East and Angel Reef while West Culloden, Western Buccoo and Black Jack Hole sites had similar proportions of brooders and broadcasters, due to moderate abundances of Diploria and Siderastrea juveniles (Table 4, Figure 2). The predominant species were Agaricia spp. (45.4%) followed by Siderastrea spp. (12.7%), Diploria spp. (9.7%), and Porites spp. (8%). The remaining species altogether only represented 24.2% of all juveniles. Agaricia spp. was most predominant at most sites with the exception of Black Jack Hole, where Porites spp. (mainly P. astreoides) dominated.

12 A 100

%Juvenile taxa 20

0 OB WB CE CW BJH AR

The assessed juvenile taxa composition differed greatly from the adult coral community composition (Figure 3). Dominant adult coral species, like Montastrea spp., were unrepresented in the juvenile sample counted. On the other hand adult coral from the Agaricia genus represent less than 5% of adult coral cover at each site, despite being the most dominant juvenile group. The small-sized brooding Scolymia spp., had moderate abundances of juveniles, mainly at Culloden sites. The abundance of Siderastrea and Diploria juveniles were the only genera to somewhat reflect the adult percent cover.

Correspondence analysis shows the relationship between the abundance of juvenile taxa (8 genus groups) across the six sampled sites (Figure 4). The ordination plot show close correspondence between Outer Buccoo and Angel Reef due to their similar high abundance of Agaricia. West Buccoo and Culloden differ from the aforementioned sites due to their higher abundance of Scolymia, Siderastrea and Diploria. The far right position of Black Jack Hole shows this site’s taxa composition differs from all other sites because of high Porites abundance and a lack of Agaricia juveniles.

Figure 4. Correspondence analysis (CA) biplot showing the ordination of in situ juvenile genera in 2013 along the first and second axis, which explains 51% and 31% of the variance, respectively. (the group “other” compromises of Favia fragum, Eusmilia fastigiata, Colpophyllia natans, and Mycetophyllia ssp.)

2.3.2 Characterization of sedimentation

The average sediment accumulation rate was 5.6 ± 4.2 mg cm-2 d-1 (Table 5). Sediment rates were not significantly different at all sites, with the exception of Culloden West (Tukey HSD, p>0.05). Accumulation rates at Culloden West were four times higher than at all other sites (15.1 ± 2.9 mg cm-2 d-1). While sedimentation rates were higher in one trap-set (17.6 ±0.9 mg cm-2 d-1) than the other (12.5 ±0.4 mg cm-2 d-1), the overall accumulation rates in all tube traps were much higher than those rates measured in the other sites. Whilst at Buccoo sediment rates did not vary greatly among trap-sets, the opposite was true at Culloden and Speyside (Fig.5).

The composition analysis (Table 5, Figure 6) indicated that terrigenous material dominated at all sites and represented the most dominant fraction of the collected material, ranging between 51.6 ±0.2-72.98 ±0.3 % being highest at Culloden Reef which were the sites nearest to mainland. The total percentage of carbonate materials ranged from 22.03 ±0.7 - 40.1 ±0.2 %, and were highest at Buccoo sites. Organic matter represented <8% of the sediment composition of the samples at all the sites; this small percentage could be made up of the thin layers of turf that had grown inside the pipe trap. Apart from the greater proportion of terrigenous material being slightly higher at the Culloden sites, overall the composition of sediment collected did not vary greatly at each site.

14 Table 5. Mean and standard deviations of sediment measurements

Outer Buccoo Western Buccoo Culloden East Culloden West Black Jack Hole Angel Reef No. of traps recovered 9 9 5 6 8 8 Mean weight(g) 2.5 ± 0.73 3.2 ± 0.5 3.1 ± 1.0 9.2 ± 1.8 3.62 ± 2.2 2.7 ± 1.3 Rate (mg cm-2 d-1) 3.4 ± 0.9 4.3 ± 0.7 4.8 ±1.6 15.1 ±2.9 4.9 ±3.1 3.6 ±1.8 Sediment composition (%) Terrigenous 56.5 ± 0.2 51.6 ± 0.3 70.98 ± 0.6 72.98 ± 0.3 65.32 ± 2.4 58.76 ± 4.6

CaCo3 35.6 ± 0.5 40.5 ± 0.2 23.49 ± 0.3 22.03 ± 0.7 26.96 ± 3.9 35.90 ± 5.7 Organic 7.9 ± 0.2 7.9 ± 0.2 5.53 ± 0.1 4.99 ± 0.4 7.72 ± 1.6 5.34 ± 1.1 Grain size distribution (%) Course sand >500 µm 1.4 ± 0.1 5.5 ± 1.7 3.3 ± 1.5 3.5 ± 3.1 15.9 ± 5.9 8.8 ± 1.8 Medium sand 250-500µm 3.1 ± 1.3 10.5 ± 1.6 7.6 ± 1.1 4.1 ± 0.0 20.2 ± 2.2 21.7 ± 4.1 Fine sand 125-250µm 22.6 ± 4.5 18.2 ± 1.6 26.0 ± 1.0 29.8 ± 1.3 35.0 ± 3.8 42.9 ± 11.2 Very fine sand 63-125µm 21.8 ± 8.9 29.0 ± 0.7 44.4 ± 6.8 45.5 ± 2.4 20.5 ± 5.2 15.3 ± 5.3 Silt/clay <63 µm 51.0 ± 10.7 36.7 ± 4.3 18.8 ± 3.3 17.1 ± 0.7 8.4 ± 4.9 11.3 ± 7.6

1 - d

2 - Sediment rate mg cm

Figure 5. Boxplot of sediment accumulation rates per site

% Distribuon of parcle sizes % Sediment composion 0 20 40 60 80 100 0 20 40 60 80 100

OB OB

WB WB

CE CE

CW CW

BJH BJH

AR AR

coarse sand mediam sand ﬁne sand very ﬁne sand silt/clay Terrigenous Calcareous Organic Figure 6. Stacked barplots of average percent of sediment (left) composition and (right) particle size distribution from sediment data collected May-June 2013

15 The sediment grain size distributions differed between the three reef systems. Culloden and Buccoo Reef, both on the western side of Tobago facing the Caribbean Sea, had a greater proportion of very fine sand (63-125-µm) and silt/clay (<63μm). At Buccoo Reef sites the proportion of silt/clay particles was the highest and represented the largest grain size fraction (36-50%). At the Culloden sites the dominant grain size fraction was very fine sand (44-45%). The Speyside sites, on the north-eastern side of the island, were dominated by fine sand sediment, 250-500µm (35-44%). Additionally, these sites had the largest proportion of coarser materials, >125μm, and had almost twice the amount of course and medium sand (>500- 250µm) in comparison with the other four sites. Although offshore from Tobago, Black Jack Hole and Angel Reef are 50m or less from the shore of small islands, so there is a greater potential for coarse and medium sand particles reaching the fore reef zones with enough current and wave energy. 2.4 Discussion

While healthy coral reefs typically have high numbers of juvenile coral colonies, at degraded reefs their numbers tend to be limited (Jackson et al. 2014). From reviewing the literature across the Caribbean, Ruiz-Zarate and Gonzales (2004) found that juvenile densities ranged from 13–274 m-2 in what used to be healthier reefs up before the die-off of the grazing urchin Diadema antillarum in the 1980s. Most studies since then have reported much lower densities that range between 0.8-12 juveniles m-2. This decline has been attributed to the decline in parental stock, decrease in herbivory, increase in algal growth and terrestrial runoff (Jackson et al. 2014). Our study found juvenile density across Tobago’s major reef system to be low, averaging around 5.71±2.39 m-2, which was similar to the densities found in Florida (Moulding 2005; Miller et al. 2000; Chiappone & Sullivan 1996), Bermuda (Smith 1997), Cayman Islands (Manfrino et al. 2013) and Belize (Irizarry-soto & Weil 2009).

Densities did not vary among the different reef sites assessed with the exception of Angel reef, where juvenile abundances were significantly higher. This difference may be explained by Angel Reef’s higher level of coral cover, which, based on 2013 assessments (see Table 1 in Chapter 2), represents about 46% of the reef benthos, whereas at the other sites it was ≤ 21%. Additionally, algal coverage was <14% of Angel Reef’s benthos as opposed to 28-51% at the other reef sites. This supports the notion that higher coral cover and lower algal cover is a more hospitable environment for larvae and juveniles to develop (Jackson et al. 2014).

16 Based on a growth rate of 0.37 to 0.73 mm/month (Birkeland 1977) most juveniles recorded in this study were about 5 to 10 years old. Accordingly, the assessed juvenile communities were likely impacted by the 2005 bleaching event, which was followed by a coral disease outbreak (Bouchon et al. 2008). Whilst coral juveniles tend not to suffer greatly from bleaching induced mortality (Mumby 1999; Shenkar et al. 2005), the bleaching of adult colonies however is known to reduce coral’s reproductive output in the years following the event (Ritson-williams et al. 2009; Ward et al. 2000a; Mallela & Crabbe 2009). Broadcasting taxa in the Caribbean are especially affected, as most bleaching events occur during their yearly spawning period between August and October (Szmant & Gassman 1990). Thus it is likely that the juvenile densities reported in this study were even lower than usual, due to the impact of the 2005 bleaching event. This notion is supported by studies measuring recruitment rates on tiles at Buccoo Reef. In the early 1990s, mean recruitment rates were 188 m-2year-1 (Laydoo 1993), which by 2007 was reduced to 103 m-2year-1 (Mallela & Crabbe 2009), indicating the decline in reproductive output over the last two decades and following the 2005 bleaching event. Consequently, it is likely that the 2010 bleaching event also has had a negative impact on the reproductive output of corals, leading to a continuously low juvenile community, which in turn has the potential to slow down post-disturbance recovery.

The juveniles assessed comprised primarily of brooding genera, particularly from the genus Agaricia, which generally contribute relatively little to overall reef-building processes (Hughes & Tanner 2000). While juveniles form the largest reef-building taxa on Tobago were very rare, e.g. Montastraea and Colpophyllia, a substantial number of Siderastrea and Diploria juveniles were found at some sites. Other recent studies have also found moderate juvenile abundances of these two taxa (Miller et al. 2000; Moulding 2005; Vermeij et al. 2011). Additionally in-situ post- settlement studies have found Siderastrea and Diploria recruits to have high survival rates in their post-settlement stage (Irizarry-soto & Weil 2009). The community composition did not vary greatly across sites, except at Black Jack Hole, where brooding genus Porites was most dominant while Agaricia juveniles were scarce. This disparity can be explained by the fact that the Agaricia adult population was also very sparse at this site. As in most Caribbean reefs the relative abundance of most juveniles did not match with those of the adult coral community. Pioneer genera like Agaricia and Porites were over-represented while large-sized genera like Montastraea and Colpophyllia were underrepresented. Exceptions were the relative abundances for juveniles of the reef-building the genera Siderastrea and Diploria, which did match the adult community at some sites.

17 The juvenile community composition found in this study is typical among Caribbean reefs (Chiappone & Sullivan 1996; Edmunds et al. 1998; Irizarry-soto & Weil 2009). Unlike Pacific coral reefs, where the most common corals also have the most abundant juveniles (Miller et al. 2000) juvenile communities in this region tend to be a direct function of the taxa’s different life- history characteristics and reproductive strategies (Bak & Engel 1979; Miller et al. 2000). Weedy small-size brooding genera such as Agaricia and Porites tend to dominate in terms of juvenile abundance as they are very fecund and have high recruitment but also have high mortality rates. Conversely, broadcasting corals, such as most of Tobago’s key framework building taxa, have low recruitment rates but are more resistant to disturbances and thus live longer and grow larger (Wittenberg & Hunte 1992; Hughes & Jackson 1985). The key reproductive difference between these two groups is that brooding taxa reproduce on a monthly basis by undergoing self-fertilisation and releasing well-developed larvae that are able to settle quickly. On the other hand, broadcasting corals only spawn once a year by releasing gametes that need to undergo external fertilisation to form larvae, which then take over a week to settle. This reproduction strategy typically produces fewer larvae and tends to be less successful at settling. However, once settled their recruits and juveniles are more resistant to disturbances than recruits from brooding taxa (Szmant 1986; Ritson-williams et al. 2009).

Even though low recruitment of massive species is not unusual among Caribbean juvenile communities, juveniles of the Montastraea genus were exceptionally rare considering it has highest parental stock -- this genus accounts for >60% of the coral cover of most of Tobago’s reefs (Alemu I & Clement 2014). The implication of this genus’s low ability to recover from disturbances via sexual reproduction and recruitment is that it could lead to a continuous decline of coral reefs, as it is possibly Tobago’s most important reef framework-building taxa. While the low number or lack of sexually produced recruits by this genus is well documented across Caribbean reefs (Hughes & Tanner 2000; Vermeij et al. 2011; Irizarry-soto & Weil 2009), the exact reasons remains unknown. Studies on some Montastrea species’ fecundity indicate that the problem does not appear to lie in the reproductive output of its colonies but the processes following spawning, i.e. fertilization, larval settlement and/or post recruitment survival (Szmant 1988). Ritson-Williams et al. (2009) suggest that since the Pleistocene it’s likely that many Caribbean framework building species have evolved with low sexual recruitment levels, as they instead rely on the production of recruits from fragmentation or budding. However, this asexual reproductive strategy no longer appears viable among Caribbean reefs where coral cover has dramatically declined.

18 While sedimentation is known for impeding reproduction and recruitment processes, our study revealed that sedimentation rates, at less than <5 mg cm-2 d-1, were low on most Tobago’s reef sites. Though composition analysis indicates that most of the sediment was of terrigenous (land) origin, the daily rates were still below Rogers’ (1990) threshold rate of 10 mg cm-2 d-1, which suggests that Tobago’s coral communities are probably not greatly impacted by sedimentation during the last months of the dry season. Whereas at Culloden West sedimentation rates were exceptionally high in comparison with the other sites (>15 mg cm-2 day-1), juvenile abundance was not significantly different at this site. Additionally, the high rates reported at site were likely so high due to high current driven turbulence as only one trap set was recovered as the other two were uprooted an.

We observed that sediment grain size analysis at Buccoo Reef sites, which had the lowest juvenile density and diversity, had the highest proportions of silt/clay (at least 20% higher than at the other sites). This is most likely a result of these sites being adjacent to urbanized land and exposure to point and non-point source pollution, especially considering that silt/clay, due to its lightness, is the land-based sediment material with the highest potential to drift all the way from shores to reefs. Though both coarse and fine-grained sediments have the potential to harm corals, studies indicate that fine sediment tends to be more harmful. Fine sediments particularly attenuates light in the water column, as they can remain suspended longer in the water column, reducing photosynthesis (Abdullah et al. 2011). Whereas corals tend to remove coarser grained sediment more readily than fine particles, fine sediment is easily re-suspended, which means the same grains can impact the coral communities more than once (Hernandez et al. 2009; Weber et al. 2006). Additionally, clay and silt-sized sediment is more likely to carry organic chemical contaminants and toxins, which in turn are understood to increase coral mortality (Fabricius et al. 2003).

Even though we found that sedimentation deposition was low, it is likely that deposition is much greater during the rainy season, which tends to occur from May until November. Considering that most broadcasting taxa spawn within this time frame it is therefore crucial that assessments be made of sediment deposition throughout the rainy season. Field and experimental studies have found that sedimentation, and the associated toxins, are known to prevent fertilization, increases the mortality of newly settled recruits and that particles covering hard surfaces inhibit the ability of larvae to settle (Babcock & Smith 2000; Torres & Morelock 2002; Fabricius 2005).

19 The disparities of sedimentation rates that we found within the same reef system, reef sites and even trap-sets indicate that sediment deposition can differ greatly among Tobago’s reefs over a narrow spatial scale (Fig.1). This may be due to the prevalence of stronger currents and wave energy at these reefs (Laydoo 1991). However, it is possible the small-scale variation in measured rates is an artefact of the sampling method. We recommend that further sedimentation rate data be collected to produce a long-term sediment accumulation profile that covers both the rainy and the dry seasons (Hernandez et al. 2009). Additionally, this study highlights the value of analyzing grain size distribution and composition of sediment samples, as this allows researchers to distinguish the origin and size of sediment particles.

In conclusion, the low densities of sexually produced juveniles, especially of large broadcasting species, indicate that post-disturbance recovery among coral communities will probably be at best slow if not very limited. Nonetheless, we found evidence that Siderastrea and Diploria species were still successfully recruiting and surviving. It is likely that mass bleaching events are reducing both parental stock (coral cover) and coral’s reproductive output, leading to lower juvenile densities. It is crucial that other disturbances that further impair recruitment process at a local scale, such as terrestrial run off and overfishing, are well understood and managed. Even though our study indicates that sedimentation rates during the dry season appear to be minimal, it is essential that further data be collected during the rainy season.

20 Chapter 3. Using coral size distribution to assess the recovery from mass bleaching in the southern Caribbean

3.1 Introduction

Caribbean coral reefs are among the most heavily impacted marine ecosystems on the planet (Bellwood et al. 2004; Edmunds & Elahi 2007; Alvarez-Filip et al. 2011). Over 70% of coral cover is estimated to have been lost in the last the decades and become replaced by macroalgae and turf dominated environments (Gardner et al. 2003). Such ecological phase-shifts have been attributed to historical reduction of large sized grazing herbivore fish populations due to overfishing, the die-off of grazing Diadema urchins in the early 1980s, water pollution, and coral disease outbreaks (Mallela et al. 2010; Weil 2001; Norström et al. 2009; Hughes 1994). In addition, over the last few decades climate change driven mass bleaching events have increased in frequency and intensity exacerbating the decline of coral communities (C. M. Eakin et al. 2010; Donner et al. 2007).

The existence of productive and attractive reefs in the Caribbean is contingent on the growth of hard corals. Therefore, there is a strong need to understand how mass bleaching events are impacting coral reefs and their ability to recover back into their pre-disturbance state. Studying the population size structure of corals can provide valuable demographic information about coral communities and population dynamics (Smith et al. 2005; Crabbe 2009; McClanahan et al. 2008). Coral life-history processes are strongly related to colony size, thus data on taxa’s colonies size frequency distribution can reflect corals responses to environmental stress (Vermeij & Bak 2000). Though labour and time intensive, collecting coral size frequency data can provide insight into past and future patterns of growth and mortality (Bak & Meesters 1998; Meesters et al. 2001; McClanahan et al. 2008).

Scleractinian coral populations tend to be positively skewed as they are comprised mainly of small to medium sized colonies with relatively fewer large colonies (Babcock 1991; Meesters et al. 2001; McClanahan et al. 2008; Adjeroud et al. 2007). The typical population size structure, however, can vary among different taxa due to their different life history traits such as fecundity, growth rates and susceptibility to morality (Meesters et al. 2001). Additionally, intraspecific population size structure can vary as a result of different environmental conditions as well as disturbance histories (De Lins Barros & De Oliveira Pires 2006; Adjeroud et al. 2007).

21 Research on Caribbean reefs have found that coral population structure tends to become negatively skewed in reefs with poor of water quality, as large colonies will be more likely to survive harsh reef conditions than smaller coral and conditions for recruitment become impaired (Bak & Meesters 1999; Bak & Meesters 1998).

Coral bleaching, a paling caused by the loss of the symbiotic micro-algae (Symbiodinium) that reside in coral tissue, leaves corals in energy deficit and consequently impairs growth and reproduction, making them vulnerable to disease and mortality (Ward et al. 2000b). Loss of live coral cover due to complete or partial colony mortality following mass coral bleaching can have a significant impact on the population size structure of different coral species (Shenkar et al. 2005). As coral’s fecundity is determined by the number of sexually mature and healthy polyps on a colony, bleaching that leads to the decline of the size and number of colonies can impact a population’s reproductive output and thus may alter the dominance and persistence of certain coral taxa (Hughes 1984). The few studies which have assessed the effect of mass bleaching on coral size distribution have found that many taxa’s mean colony size declines, due to the loss of large sized corals to partial or complete mortality resulting in an increase in small sized colonies (McClanahan et al. 2008; Crabbe 2009).

The purpose of this study was to examine changes in corals population size structure to assess the impact of climate change and recovery among coral communities of the southern Caribbean reefs of Tobago. The corals among these reefs were studied over a period during which they endured a mass bleaching event in 2010. Tobago’s fringing reefs have experienced the same degradation patterns as the rest of the region’s reefs due to the continuous overfishing, terrestrial runoff, wastewater discharge and poor coastal natural resource management (Mallela et al. 2010). Thus these reefs are also characterized by low coral cover dominated by the more persistent massive and encrusting types of coral species. Though corals were known to have bleached in 1998, 2005 was the first well recorded bleaching event in Tobago followed by another in 2010. A rapid bleaching response assessment was done throughout the bleaching period in 2010 recorded bleaching severity as high as 29-69% across Tobago’s reefs coral communities and about 2-8 % coral mortality (Alemu I & Clement 2014). Recordings of disease infected corals (Alemu I 2011) additionally indicated that coral mortality probably further increased after the bleaching event, as was the case in 2005 (Harding et al. 2008). Little is known, however, about the impact of bleaching events on the population dynamics of Tobago’s coral communities. Most knowledge about the health and disturbance history of Tobago’s coral reefs is based on changes in percent benthic cover. Only one study has

22 explored the impact of bleaching events on Tobago’s coral population dynamics, based on recruitment and growth modeling (Mallela & Crabbe 2009).

Here, we quantified the population size structure of Tobago’s dominant coral species before the bleaching episode in September, 2010, immediately after the bleaching ended in March, 2011 to assess impact of the bleaching-induced mortality, and in May, 2013 to assess change in the three years since the event. Furthermore, comparisons were made between reef systems located adjacent to urban coastal land vs. rural forested land to assess how coral population structure differs generally and in the face of a bleaching event. Benthic percent cover data was also re-surveyed in order to compare if any recovery has taken place. The results of this study provide insight into the ability of different dominant Caribbean coral species to persist in the face of warming ocean temperatures.

3.2 Methods

Tobago is the smaller sister island of the nation Trinidad and Tobago, part of the Lesser Antilles island arc, located close to the South American mainland. Due to Tobago’s proximity the Orinoco and Amazon River deltas its fringing coral reefs systems evolved under the influence of nutrient and sediment rich flushes. Consequently, Tobago’s coral communities are characterised for having lower species diversity in comparison to other Caribbean reefs (Moses & Swart 2006; Potts et al. 2004). Tobago is a hilly 300 km2 island of volcanic origin covered mostly by forest and shrubs lands, though the south-western part of the island has undergone significant urbanisation and agricultural development.

The study was conducted on three major reef systems: Buccoo Reef, Culloden Reef and Speyside (Figure 1). Buccoo and Culloden Reef are both along the south-western coast of the island, facing the Caribbean Sea. While Culloden Reef’s bay borders forested hills with minor developments of gravel roads, coastal lands adjacent to Buccoo Reef have become heavily urbanized and wastewaters become directly discharged into Buccoo’s bay. Speyside features a large network of fringing reefs along small islands and rocky outcrops on the north-eastern side of the island (Laydoo 1991). These reefs are highly valued by the recreational diving community. Speyside coastal lands remain relatively undeveloped comprising of a hilly forested landscape apart from Speyside village (a fishing community) and two medium-sized hotels.

Figure 7. Map of Tobago and location of studied reef systems and sites

Two study sites were established at each reef system (Figure 7). Outer Buccoo and Western Buccoo represent reef systems exposed to land pollution from run off and sewage waters. Culloden East and Culloden West act like controls to the Buccoo sites, as they experience similar marine conditions but less land-based pollution. Black Jack Hole and Angel Reef at Speyside represent reefs exposed to the Atlantic Ocean that are also less affected by marine pollution. Surveys on corals colony size and the reefs benthic cover were collected at each reef site from May to June in 2013 using SCUBA. The data are compared to previous surveys carried out since 2010 at the same sites, except Angel Reef, gathered by the Trinidad and Tobago Institute of Marine Affairs (IMA) as part of their annual monitoring programme. 3.2.1 Benthic cover survey

Benthic percent cover was estimated following the photo-quadrat method as described by Hill and Wilkinson (2004). Ten randomly-placed 10 m transects were carried out at each site between the depth of 8 and 12 m. Along each transect 1m2 non-overlapping photo-quadrats were taken. For each photoquadrat benthic cover was identified under 60 random points in each image using Coral Point Count with the Excel–extensions (Kohler & Gill 2006), following the protocol developed by the IMA (Alemu I & Clement 2012). Surveys from previous years, carried out by the IMA, employed similar methods, however they used permanent transects

24 which they revisited in each survey. Additionally, only five transects were completed in their 2010 survey. 3.2.2 Colony size frequency survey

Along each 10m transect the length of all adult coral colonies ≥5 cm were recorded, that lay within 50 cm on each side of the transect tape, following Done et al. (2010). All measured colonies were identified to species levels expect for the genus Agaricia due to identification difficulties. Colonies with >50% of their living tissue within the belt transect area were included. Colonies divided by partial mortality into separate patches of living tissue that were >3 cm apart from each other were considered separate colony entities and were measured individually (McClanahan et al. 2008; Adjeroud et al. 2007; Done et al. 2010). To identify changes in the coral community and population structure since the 2010 bleaching event we used similar datasets collected by local Institute of Marine Affairs (IMA) in September 2010 (before bleaching induced coral mortality) and in March 2011 (after bleaching induced coral mortality). The previous surveys collected length and width of each living colony along four 10 m by 2 m belt transects. The mean of the width and length measurements was employed in this analysis. As surveys in 2010 and 2011 covered a total of 80 m2, eight transects were randomly subsampled from the 10 transects assessed in 2013.

3.2.3 Statistical analysis

All data analysis was carried out in in R version 2.15.1. Colony size data were used to calculate mean size, standard deviation, standard error, median, skewness and skewness standard error for each dominant taxon per site and year. Skewness values greater than two times the skewness standard error were considered to be significantly skewed from normal (McClanahan et al. 2008). For each taxon percent cover estimates were calculated per site following the index developed by Done et al (2010) (which was based on Marsh et al 1984):

� = 100×(��!/4)/[(� + �)�], where C is the percentage cover, N is the number of colonies, D is the mean dimension of the N colonies, W is the width of the belt transect in centimeters (100 cm) and L is the total length of the belt transects. This equation makes the assumption that all colonies are circular in shape with a diameter equal to the mean lateral dimension recorded and that the width is the width of the actual belt plus the mean diameter of all colonies to account for colonies extending beyond the 100 cm belt width. Though this index tends to over-estimate true percent of coral

25 cover, the index provides us with comparable values to assess the changes occurring in the total live coral cover per species and sites.

To test whether each species’ mean size and size frequency distribution differed between the years and between the sites, each distribution was first tested for normality using Shapiro-Wilk test and for homogeneity of variance by Levene’s test. The majority of distributions, before and after log-transformation, were not normal and did not exhibit homoscedasticity. Therefore, non-parametric Kruskal–Wallis tests were employed to test for the significant differences in size across each years and sites (McClanahan et al. 2008), followed by post-hoc pairwise Mann- Whitney U-tests comparisons, and Kolmogorov-Smirnov tests to test for differences in size frequency distribution (Adjeroud et al. 2007). To avoid Type I errors across multiple comparison tests, critical values for all tests were adjusted using the Bonferroni correction, resulting in an α-level of 0.0167.

The impact of the bleaching event on the coral community of each site over the three years was visualised with a non-metric multidimensional scaling (NMDS) based on Bray-Curtis dissimilarities of each species abundances per site, using vegan package version 2.0-10 (Borcard et al. 2011). Changes in mean percent live coral cover per site (based on photoquadrat benthic assessment data using number of transects as sampling units) were calculated using a one-way ANOVA followed by a Tukey’s honest significant difference test, after testing assumptions of homogeneity of variance and normality were met using graphical methods. Angel Reef was excluded from all historical analysis, as data was not available.

3.3 Results

3.3.1 Changes in percent coral cover

The annual estimated percent live coral cover, based on the photo-quadrat assessments, before the bleaching event in 2010 across the five sampled reefs ranged from 15.6 ±10.7 to 28.3 ±11.0 (Table 1). After the 2010 bleaching event, mean coral cover declined by ≥35% at all sites except Blackjack Hole, but this decline was only statistically significant at the Culloden sites (Tukey HSD, <0.05). Coral cover appears to have undergone little recovery by between 2011 and 2013; no significant change in coral cover was determined throughout this recovery period. Although there is no historical data for Angel Reef, percent live coral cover at this site in 2013 was three times higher than at all other sites, including Black Jack Hole, which is less than 1km away.

26 Table 6. Mean percent cover of live coral and (±) standard deviation estimated at each site and year. Values with the same letter subscript indicated significant pairwise comparison (p<0.05)

Year (# of transects) Site 2010(5) 2011(10) 2012(10) 2013(10) Outer Buccoo 28.3 ±11.0 18.1 ±12.0 19.8 ±9.7 20.2 ±5.9 Western Buccoo 20.2 ±10.6 12.5 ±9.2 13.3 ±8.7 13.6 ±8.4 Culloden East 24.3 ±10.8abc 14.3 ±5.7a 13.9 ±7.3b 12.2 ±5.23c Culloden West 27.6 ±5.8abc 12.5 ±5.5a 10.9 ±4.5b 16.0 ±7.6c Black Jack Hole 15.6 ±10.7 17.2 ±5.8 13.0 ±7.1

Angel Reef 46.4 ±14.6

3.3.2 Changes in coral population structure and community composition

Overall 3671 scleractinian coral colonies were measured among which 27 species. Total colony abundance at each reef site after the bleaching event only changed by 10% or less at the Buccoo sites and Culloden West, but increased by 23% at Black Jack Hole and decreased by 26% at Culloden East (Table 2). However from 2011 to 2013, the number of colonies increased markedly at most sites, except Western Buccoo. Changes in species richness showed no consistent pattern of change across sites.

Table 7. Total number of coral colony and species (in parenthesis) recorded per site

2010 2011 % change between Sites 2013 (Pre-bleaching) (Post-bleaching) 2010-11 2011-13 Outer Buccoo 192 (16) 209 (12) 381 (14) 9% 82% Western Buccoo 264 (11) 248 (14) 230 (15) -6% -7% Culloden East 186 (16) 138 (14) 246 (15) -26% 78% Culloden West 99 (10) 109 (13) 328 (16) 10% 201% Black Jack Hole 200 (10) 245 (14) 292 (11) 23% 19% Angel Reef 304 (14)

The most dominant coral species in 2010, based on abundance among the five surveyed sites were: Montastraea faveolata, Diploria strigosa, Siderastrea siderea, Agaricia spp., Porites astreoides, Colpophyllia natas, Montastraea cavernosa, and Diploria labyrinthiformis. M. faveolata was by far the most dominant species at all sites in terms of percent cover and in many cases also dominated in colony abundance. Most of these abundant species were massive and encrusting coral types, and belong to the Faviidae family with the exception of S. siderea, P. astreoides and Agaricia spp.. Species with large populations at all sites (≥12 colonies) included M. faveolata, S. siderea and D. Strigosa (except at Black Jack Hole). M. cavernosa was common only at Culloden sites and Black Jack Hole, while C. natas and D. labyrinthiformis were

27 only abundant enough at the Buccoo sites. Weedy species like Agaricia spp. frequented at all sites except Black Jack Hole, where P. astreoides was more common despite being sparsely present at the other sites (see Appendix Table A1 for summary statistics of each dominant species per site).

The changes among each dominant species over time varied greatly per site in terms of abundance and percent cover (determined from size frequency data). We generally found that after the bleaching event, the majority of coral populations at each site saw decline in total percent cover as determined from the size frequency data. While colony abundance also declined among most population, many remained relatively unchanged or increased in abundance, even among some populations that saw a decline in percent cover. For example, the abundance of colonies for M. faveolata at Black Jack Hole and C. natans at Outer Buccoo increased after the bleaching event but their percent cover almost halved. However, most Agaricia spp. and P. astreoides populations did not decline in abundance nor percent cover.

Before the bleaching event, the coral populations at each site were mostly dominated by small to medium sized colonies, as the skewness of most size distributions was significantly positive or slightly positive (Appendix Table A1). The few normally skewed distributions compromised of the populations of P. astreoides, C. natas, D. labyrinthiformis, and S. siderea at Culloden sites and Western Buccoo. After the bleaching event, we found that skewness became slightly less positive or tended towards normality among some species (Appendix Table A1). However some other coral populations, like that of M. faveolata at most sites, became more positively skewed after the bleaching.

At each site we found that the size frequency distributions for the majority of species did not significantly change (KS-test, P<0.016) between 2010 and 2011 (Appendix Table A2). Significant change was found among three species: Agaricia spp. at Outer Buccoo, S. siderea at Western Buccoo and M. faveolata at Black Jack Hole. These populations showed a decline in the large size classes and an increase in the smaller size classes (Figure 8). Mean colony size (diameter) decreased among some species and sites by 2011 (Figure 9), especially among S. siderea, Agaricia spp., P. astreoides and C. natans. However among most species mean colony size did not significantly differ (Mann-Whitney test, P<0.016) after the bleaching event. The only colony mean sizes that did significantly declined were for Agaricia spp. at Outer Buccoo, S. siderea at Western Buccoo and M. faveolata at Outer Buccoo and Black Jack Hole (Figure 9).

Figure 8. Size frequency distributions of coral taxa at sites with significant differences between years, as determined by the Kolmogorov-Smirnov test (*)

*2010-13 *2010-11 *2010-11 *2010-11

2010-13 2011-13

2010-13

*2010-11 *2011-13 *2010-13 *2011-10 *2010-13

2010-13 2010-13 2011-13

*2010-13

2011-13

Figure 9. Boxplot and mean size (white filled dots) per species per site indicating changing trends in colony size between each year. Letter in each plot indicate if there was a significant difference (Mann-Whitney U test, P< 0.016) between 2010-2011 (A), 2010-2013 (B), 2011-2013 (C).

Between 2011 and 2013 we observed that colony abundance among most coral populations had increased, though quite a few remained unchanged. While, percent cover did increase among some populations, in many cases it remained unchanged and in some it declined. We noted that abundance of P. astreoides and Agaricia spp. more than doubled by 2013 at most sites and their percent cover also increased or stayed the same. On the other hand colony

30 abundance and cover among C. natas, D. labyrinthiformis and S. siderea populations remained relatively stable or declined. We also noted Black Jack Hole most of its dominant species, M. faveolata, M cavernosa and S. siderea, which lost ≥45% of its percent cover by 2011, showed no recovery by 2013 and M. cavernosa had declined further.

By 2013 most species population distributions became more positively skewed than in 2011 and 2010, indicating a greater proportion of smaller sized colonies with some exceptions (Appendix Table A1). Additionally, a visible rise of smaller size classes was noted among the size frequency distribution of many species (Figure 8). However, we found that for the majority of species no significant difference was detected among the size frequency distribution and mean colony size between 2011 and 2013. The only significant changes were found among the populations of M. faveolata, S. siderea and P. astreoides at Black Jack Hole, as well as for Agaricia spp. at Culloden West and D. strigosa at Culloden East (Appendix, Table A2). All of these populations by 2013 showed an increase in the smaller size classes large size classes and some a decline in the larger size classes (Figure 8). More populations experienced significant changes in their size frequency distributions and mean colony size from before bleaching in 2010 to 2013, than from after bleaching in 2011 to 2013 (Appendix, Table A2).

Figure 10. Non-metric multidimensional scaling using Bray-Curtis dissimilarities plot of the qualitative changes among the coral communities at each of the sites per year (named and colour coded) across Tobago. NMDS stress = 0.133.

31 In terms the community species composition at each site, we found the species that were dominant in 2010 continued to be dominant by 2013. The nMDS analysis, based on colony abundances of all species present per site, illustrates (Figure 4) how the community composition among most sites remained relatively similar across the three assessed years, especially among the Culloden Sites. The greatest community composition change from 2010 to 2013 occurred at Black Jack Hole possibly due to an increase in P. astreoides and decline of M. cavernosa. The community composition shift at the Buccoo sites by 2013 also stand out, which was most likely due to an increase in abundance of Agaricia spp. Intraspecific comparisons of populations between the different reefs sites also showed that species size frequency distributions did not significantly differ between most sites. The few significant differences were mostly found in 2013 (Appendix Table 3). 3.4 Discussion

Even though severe mass bleaching events are often followed by extensive loss of coral mortality (Hughes et al. 2003; Baker et al. 2008), the overall impact of this bleaching event appears to have been low for most of the assessed coral populations. This may be related to the fact that coral cover at these reef systems was already low, having been affected by a historical loss of herbivores, water pollution, and more recently from the 2005 bleaching event and subsequent disease outbreaks. Thus it is likely that by the time of the 2010 event coral assemblages were already narrowed down to more resistant species, like the massive species of Montastraea, Diploria, Colpophyllia and Siderastrea that have thick tissues and large inter- corallite spacing (Baker et al. 2008; Ritson-williams et al. 2009).

We found that following the heat stress in 2010 coral populations showed signs of having experienced some bleaching-induced mortality. Many species experienced a decline in colony abundance; percent cover and mean colony size by 2011, symptomatic of corals having suffered complete mortality and/or partial mortality. Nonetheless, almost all populations’ size frequency distributions and colony mean size did not significantly differ between 2010 and 2011, indicating that the bleaching event did not have a major impact on the population size structure of most species.. The composition of the dominant coral community also remained largely unchanged following the bleaching event. Additionally, mean percent cover, based on photoquadrat transects, only declined significantly at Culloden sites, whilst no significant changed was detected at the other three sites. However, the lack of statistical significance in the decline at the other sites may be an artefact of the sampling method; given that coral cover

32 is low across Tobago and that several species are rare at some sites, the individual transects may be too short to be representative of coral cover or population size structure.

By 2013, most species population size structure did not significantly differ, suggesting that most populations remained unchanged throughout the two years following the bleaching event. We did observe that in many populations, though colony abundance increased by 2013, this rise was not always accompanied by an increase in percent cover, indicating that perhaps coral colonies may have been affected by further fragmentation in the years following the bleaching. Additionally, the skewness of most populations in 2013 became more positive and there was a noticeable increase in the smallest size class (5-15cm) among many species. The rise of smaller size classes could suggest that there has been a input of juvenile colonies that have grown large enough (<4cm) in the past two years becoming part of adult demographic, which would be indicative of the first steps of post-bleaching recovery taking place. However, findings from Chapter 2 indicate that recruits/juveniles densities were very low throughout Tobago and mainly compromised of brooding genera Agaricia or Porites, which means this was unlikely the case for most other species. It is possible that the rise in small sized colonies between 2011 and 2013 among the massive species may have instead resulted from colonies undergoing fragmentation due to partial mortality driven by other disturbances, such as predation, sedimentation or disease. Disease could be the most likely explanation, given that the IMA recorded a rise in disease infected corals after the bleaching event, and after the 2011 survey used in this study (Alemu I 2011). Many studies have found that disease outbreaks tend to be the primary cause of post-bleaching coral mortality, whilst few corals actually die of bleaching itself (Miller et al. 2006; Wilkinson & Souter 2008; Brandt & McManus 2009)

Though the response to the 2010 bleaching event varied greatly among and within species and sites, some noticeable patterns of change were detected across all sites. S. siderea, one of Tobago’s main reef-building species, was most impacted by the 2010 bleaching event. The decline of the mean colony size continued after the bleaching event, likely due to the post- bleaching fragmentation discussed above. Other Caribbean studies have reported S. siderea to be among the species most susceptible to bleaching, bleaching induced mortality (van Hooidonk et al. 2012; Oxenford et al. 2007) and becoming infected by diseases that typically occur following bleaching (Gochfeld et al. 2006).

Significant population changes were also noted among the weedy species Agaricia spp. and P. astreoides (abundant only at Black Jack Hole). While their populations remained relatively unchanged following the bleaching event (2011), by 2013 abundance of both corals more than

33 doubled in size and increased in percent cover with a conspicuous rise in the smaller size classes. This rise in smaller size colonies most likely came from an input of juveniles. These brooding species tend be have high recruitment and post-settlement survival rates as their recruits initially grow quickly in comparison to other taxa (Arnold & Steneck 2011). This is supported by our finding in Chapter 2, as the genera Agaricia and Porites compromised of between 30-80% of all juveniles at the sites at which they were present. Thus, these results indicate that both Agaricia spp. and P. astreoides appear to favor post-disturbance reef conditions highlighting their niche as opportunist and weedy species (Hughes & Jackson 1985).

Our results also revealed that population size structures between the assessed reef sites did not significantly differ, despite Buccoo reefs being adjacent to urbanized land and its sites were likely to be more environmentally stressed due to poor water quality (Lapointe et al. 2010). Species’ population size structure at all sites tended to be positively skewed as the majority were dominated by small to medium sized colonies throughout the three years. The only site that set itself apart was Black Jack Hole, which out of all sites most of its dominant coral species’ population size structures became altered following the bleaching event. The populations of M. faveolata, M. cavernosa and S. siderea, saw a decline in mean colony size and percent cover after the bleaching by 2011 and in the two years following. This indicated that the coral populations likely suffered from bleaching-induced partial and complete mortality, and the surviving individuals may have succumbed after the bleaching event. Correspondingly, Alemu I and Clement (2014), who carried out the bleaching assessment at the time of the 2010 heat stress, found that reefs in Speyside had the highest bleaching and mortality response among the three reef systems studied (same ones as in this study).

Taken together, the results indicate that acute disturbances, such as bleaching events, can lead to the decline of the mean colony size of species’ populations due to the dominance of smaller sized colonies as a result of fragmentation. Similar findings were established among Kenya reefs based on the impact of a bleaching event in 2005 using the similar definitions of an independent coral colony (McClanahan et al. 2008). Coral colonies that have experienced partial mortality need to prioritize their energy sources to recover from lesions and often postpone reproduction (Hughes et al. 2003). Therefore, considering that smaller colonies produce fewer gametes and are also more susceptible to other disturbances like disease or sedimentation (Graham & van Woesik 2013), a reduction of mean colony size can lead to a reduction in fecundity and overall resilience to disturbance. This will slow down post- disturbance recovery as it declines the chances newly dead corals becoming rapidly recolonized

34 by coral recruits and instead enabling the invasive establishment algae and sponges (Connell 1997). Thus, overall acute disturbances that drive fragmentation such as bleaching, pose a large threat Tobago’s fragile communities.

In conclusion, from assessing the changes in population size structure among Tobago’s most dominant coral species, we found that bleaching events caused by heat stress, such as in 2010, act as chronic disturbances as they lead to the slow shrinkage of Tobago’s coral communities. We found many populations of Tobago’s coral communities resisted become heavily impacted by the heat stress. Nevertheless, we also found evidence of some population’s mean colony size decreasing, including most S. siderea populations and those of Black Jack Hole’s dominant species. We also found that the majority of population that experienced post-bleaching decline instead of recovering, the mean colony size continued to the declined in the two years following the bleaching event. Considering that heat stress events are predicted to be more frequent and intense in the next 20 to 30 years (Donner et al. 2007; Hoegh-Guldberg et al. 2007), it is likely that the population size structure of Tobago’s coral community will become dominated by smaller sized colonies leading to the shrinkage of populations’ mean colony size and overall size. As this will impact the reproductive output of coral communities, species that are naturally more fecund such as brooding species Agaricia spp. and P. astreoides are likely going to become more dominant.

35 Chapter 4. Conclusion

In this study I assessed the recovery of coral communities of three representative reef systems around the southern Caribbean island of Tobago following the 2010 mass bleaching. To understand which coral taxa are producing sexual recruits that can grow successfully surviving on among Tobago’s reefs in 2013 I assessed the abundance and composition of the juvenile community at each site (Chapter 2). As coral growth and regeneration is affected by the impact of sedimentation, deposition rates and the composition of sediment deposited at each reef site were also evaluated (Chapter 2). Finally, to assess the impact and recovery of the bleaching event, I examined the changes in the size population structure of Tobago’s dominant coral taxa using data collected before the bleaching episode in 2010, immediately after bleaching ended in 2011, and two and a half years after the bleaching event in 2013 (Chapter 3).

The results suggest very slow recovery of coral cover through substrate re-colonization via the sexual production of recruits since the bleaching event. Similarly to other Caribbean reefs, we found low juvenile densities across most of the Tobago’s reefs. The juvenile coral community was dominated by weedy brooding taxa, while broadcasting corals represented the minority of juveniles found. Juveniles of the key framework building coral taxa in Tobago, Diploria and Siderastrea, were however present, but rare. The disparity between brooding vs. broadcasting juvenile abundances is consequence of the different life-histories among these two types of corals. Nevertheless, the overall low abundance of juveniles is a sign that the vitality of coral populations across Tobago’s reefs and/or their environmental conditions have become so altered that they are no longer suitable for recruitment processes.

The sedimentation assessment found that between May and June of 2013, sedimentation deposition rates were below levels that tend to subject coral communities to stress at most sites. The exception was the Culloden West site where rates were three times higher than at the other sites; these higher sedimentation rates were likely caused by strong currents occurring at that site. A larger fraction of the sediment deposited at all reefs sites was likely coming from land, as most of it was composed of terrigenous material. Consequently, it is likely that sedimentation rates to be significantly higher during the rainy season, as terrestrial leaching and runoff from land into coastal zones tends to be much higher. Thus, it is likely that corals communities experience seasonal sedimentation stress, especially those near urbanized landscapes. Though I found sedimentation rates did not differ between reefs sites near urban or rural land, there was more silt/clay sized material at rural sites. Additionally the highest

36 juvenile density and diversity of juvenile corals were only at reefs adjacent to undeveloped land, whilst sites at Buccoo reef had the lowest density. I highly recommend that a long-term sediment accumulation assessment be done among Tobago’s reefs that cover both the rainy and the dry seasons. Especially considering that most broadcasting taxa spawn during the rainy season as released gametes, larvae and recruits are particularly susceptible to becoming damaged by sedimentation (Ritson-williams et al. 2009).

From examining changes in the population size structure before and after the bleaching event we found no significant change in the population structure of the majority of dominant coral species after undergoing severe heat stress in the late summer of 2010. However, those populations that did significantly change, apart from experiencing declines in their total coral cover, had reduced mean colony size due to complete or partial mortality shrinking or fragmenting larger sized colonies into smaller ones. For example, this was the case for S. siderea, which was found to be among the most sensitive taxa to bleaching mortality across Tobago’s reefs, as well as for key reef building taxa (M. faveolata, M. cavernosa and S. siderea) at Black Jack Hole. This indicates that in the event of bleaching being more severe in the future, it may lead to coral populations’ mean colony size declining due to extensive fragmentation and complete mortality among primary reef building coral species.

In the two years following bleaching event at most sites no change was observed among the population size structure for the majority of coral species. This is not surprising, given slow coral growth rates; changes should again be examined over a longer time period, such as 5 to 8 years. Nonetheless, we did find that many taxa by 2013 had experience an increase in the abundance of small sized colonies, without a major change in coral cover, and leading to a significant decline in colony size among a few taxa. This increase in smaller sized colonies could have been caused by colonies experiencing post-bleaching fragmentation due to other secondary disturbance such as disease. A major implication of coral population becoming more positively skewed (i.e. dominated by smaller sized corals) is a decreases the fecundity of coral populations. This, in turn, can affect the long-term recovery of the reefs as it reduces the larvae and recruitment output slowing down coral populations overall regeneration. Among the only taxa that may have been starting to recover were Agaricia spp. and P. astreoides, which had high number of juveniles at the reef sites where they were present.

Many Caribbean reefs’ coral communities over the last decades have become narrowed down to the most resistant species. Though these communities can avoid major mortality following mass bleaching, they still become weakened and can experience enough individual full and

37 partial mortality to significantly change the population size structure. Thus, mass bleaching events greatly threaten the future persistence and health of Caribbean coral reefs like Tobago. This study indicates that across Tobago’s different reef sites, the bleaching disturbance can lead to a dominance of smaller size coral colonies, which could negatively affect the reproductive output. This could further decrease these coral communities already low ability to develop coral juveniles. Given the evidence that recruitment processes are already stunted among Tobago’s coral communities, as has been observed for other Caribbean reefs, it is essential to improve the current health of coral communities to increase the chances of successful recruitment taking place. Thus, it is paramount that local coastal ecosystem conservation and management efforts strive towards increasing herbivory and reducing sedimentation and nutrient enrichment.

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45 Appendices

Table 8. Summary of the colony abundance and size data collected for each dominant species present at each reef sites, which includes: number of colonies (n), percent cover (%), mean size (mean), standard deviation (sd), and skewness (sk). If skewness was significantly positive skewness values are underlined.

M. faveolata S. siderea

site year n %cover mean sd sk site year n %cover size sd sk

AR 2013 83 4.8 27.3 22.7 0.7 AR 2013 29 0.7 16.8 13.2 1.5

BJH 2010 51 6.6 43.4 43.6 1.6 BJH 2010 70 4.4 28.6 23 1.8

BJH 2011 98 4 22.5 23.3 3.2 BJH 2011 52 2.7 25.8 21.3 1.9

BJH 2013 99 3.4 20.4 26.1 4.8 BJH 2013 82 2.2 17.9 16.2 2.9

CE 2010 29 2.8 36.7 42.6 2.4 CE 2010 22 1.3 27.8 18.9 1.1

CE 2011 13 0.6 24.4 30.8 2.5 CE 2011 12 0.2 12 6.1 0.4

CE 2013 52 4.2 33.2 50.5 5.5 CE 2013 18 0.6 20.5 13.5 0.5

CW 2010 24 1.3 26.9 25.5 2.4 CW 2010 11 0.4 21.7 24.7 2.5

CW 2011 6 0.6 38.6 17.3 1.5 CW 2011 14 0.9 28.2 17.9 1.1

CW 2013 48 9.5 56.2 65.6 3.3 CW 2013 16 0.8 24.9 13 0.3

OB 2010 46 4.4 36.7 34.1 2.6 OB 2010 35 3.5 37.4 27 2.1

OB 2011 102 7.9 32.3 46.8 4.2 OB 2011 21 1.3 28.2 25.3 2.6

OB 2013 121 12.1 37.4 41.1 2.5 OB 2013 49 2.5 25.4 20.1 1.6

WB 2010 105 10.5 37.4 28.8 1.8 WB 2010 40 4.4 39.5 25.2 1.1

WB 2011 59 4.2 30.7 24.8 2.1 WB 2011 36 2 27.1 21 1.2

WB 2013 59 10.7 53.1 56.7 2 WB 2013 23 0.6 17.9 12.6 1.5

C. natans D. labyrinthiformis

site year n %cover mean sd sk site year n %cover mean sd sk

OB 2010 14 3.1 60.5 49.3 1.5 OB 2010 21 1.2 26.6 20.5 1.8

OB 2011 17 1.8 38.6 26.3 1.1 OB 2011 9 0.6 30.7 16.2 0.4

OB 2013 22 2.5 40.1 41.6 3.5 OB 2013 6 0.3 26.7 15.5 0.7

WB 2010 34 7.5 59.8 61.3 1.7 WB 2010 13 0.7 25.3 20.7 2.4

WB 2011 16 2.3 46.5 43.5 0.9 WB 2011 16 1 29.1 17.6 0.6

WB 2013 22 3.6 50.1 57.4 1.6 WB 2013 13 0.4 18.4 7.2 1.7

Agaricia spp. D. strigosa

site year n %cover mean sd sk site year n %cover mean sd sk

AR 2013 67 0.5 9.5 6.4 2.2 BJH 2013 14 0.1 10.3 15.9 3.6

BJH 2010 17 0.1 6.5 2.5 1.4 CE 2010 47 1.6 20.2 12.4 2.6

BJH 2013 1 0 4 NA NA CE 2011 29 0.6 16.1 6 0.8

CE 2010 23 0.1 7.7 4.4 2.4 CE 2013 60 1.2 15.1 22.5 6.9

CE 2011 19 0.1 7.9 4.1 1.4 CW 2010 17 0.2 11.2 3.8 0.8

46 CE 2013 31 0.2 7.7 4.1 3.1 CW 2011 33 0.7 15.2 8.8 0.6

CW 2010 8 0.1 9.4 4.5 0.8 CW 2013 121 3.4 18.3 14 3

CW 2011 16 0.1 9.1 2.6 0.1 OB 2010 26 0.8 18.9 14.6 3.2

CW 2013 52 0.2 7.1 4.7 3.8 OB 2011 12 0.5 22 16.5 1.4

OB 2010 26 0.3 11 9.5 3.2 OB 2013 17 0.7 23.1 18.7 2.2

OB 2011 28 0.1 6.6 3.3 2 WB 2010 28 0.6 16.1 6.4 0.6

OB 2013 122 0.6 7.6 6.2 3.7 WB 2011 47 0.8 13.8 8.8 1.2

WB 2010 22 0.2 10.7 6.6 0.9 WB 2013 39 1 17.2 10.9 1.4

WB 2011 43 0.2 7.3 3 1.9 M. cavernosa

WB 2013 50 0.2 6.8 2.8 1.3 site year n %cover mean sd sk

P. astreoides BJH 2010 25 1.2 25 21.3 2.4

sites year n %cover mean sd sk BJH 2011 13 0.4 19.3 18.2 1.9

AR 2013 32 0.6 14.2 8.7 2 BJH 2013 4 0.1 15.8 5.9 -0.7

BJH 2010 24 0.3 12.3 4.1 0.5 CE 2010 28 1.2 23.6 16.3 1.3

BJH 2011 33 0.4 12 5.4 0.4 CE 2011 32 1.2 21.8 11.8 1.1

BJH 2013 65 0.5 9.1 3.6 0.8 CE 2013 31 1.7 26.7 18.8 1.3

CE 2010 10 0.2 16.2 3.4 -2.2 CW 2010 27 1.3 24.4 21.2 2

M. mirabilis CW 2011 22 0.8 20.7 12.1 1.2

sites year n %cover mean sd sk CW 2013 23 1.7 31 45.2 3.3

AR 2013 46 18.9 88.9 41.5 2.7

Species Site year KS P-value KW P-value

Agaricia spp. Outer Buccoo 2010-2011 0.00 0.00

M. faveolata Outer Buccoo 2010-2011 NA 0.01

M. faveolata Black Jack Hole 2010-2011 0.01 0.00

S. siderea Western Buccoo 2010-2011 NA 0.00

S. siderea Culloden East 2010-2011 0.01 0.00

Agaricia spp. Culloden West 2011-2013 0.00 NA

D. strigosa Culloden East 2011-2013 0.01 0.00

M. faveolata Black Jack Hole 2011-2013 0.01 0.04

P. astreoides Black Jack Hole 2011-2013 0.00 0.01

S. siderea Black Jack Hole 2011-2013 0.00 0.00

47 Agaricia spp. Outer Buccoo 2010-2013 0.00 0.00

D. strigosa Culloden East 2010-2013 0.00 0.00

D. strigosa Culloden West 2010-2013 0.01 NA

M. faveolata Culloden West 2010-2013 0.00 0.00

M. faveolata Black Jack Hole 2010-2013 0.00 0.00

P. astreoides Black Jack Hole 2010-2013 0.01 0.00

S. siderea Black Jack Hole 2010-2013 0.00 0.00

S. siderea Outer Buccoo 2010-2013 0.01 0.01

S. siderea Western Buccoo 2010-2013 0.00 0.00

Table 10. Significant comparison size frequency distributions between reef sites determined using Kolmogorov- Smirnov test (KS) test respectively. Abbreviated site codes: OB is Outer Buccoo, WB is Western Buccoo, CE is Culloden East, CW is Culloden West, and BJH is Black Jack

Species Year KS P-value >0.0167

D. strigosa 2010 CE-CW M. faveolata 2011 OB-BJH M. faveolata 2011 WB-BJH S. siderea 2011 CE-CW S. siderea 2011 WB-BJH D. strigosa 2013 CE-OB M. faveolata 2013 WB-BJH M. faveolata 2013 WB-OB M. faveolata 2013 CW-EC