The key demographic effects that could influence the persistence vs. winnowing of species from the community were:

#### Oceanographic Model of Connectivity

Declining fecundity and increased requirements for connectivity were central issues that could, besides levels of mortality, determine which species would be winnowed from the local fauna. In the oceanographic model, southern PAG showed clear indications of local retention of larvae (**Figure 8**) where between 5 and 36% of all larvae were locally retained in their region of origin. Connectivity between sources was apparent between (i) Dubai, Umm al Quwain and Ras al Khaimah, and (ii) central and eastern Abu Dhabi (**Figure 8**). Connectivity was primarily associated with larvae moving between neighboring subpopulations. Dubai, Umm al Quwain, and Ras al Khaimah sites showed low larval connectivity. From Dubai, 4.6% of larvae moved to Umm al Quwain, 6.61% of larvae released there reached Ras al Khaimah sites. Larval connectivity from west to east was lower. 1.37% of Ras al Khaimah larvae appeared in Umm al Quwain, 1.87% of Umm al Quwain larvae appeared in Dubai. 3.15% of larvae from central Abu Dhabi were apparent in eastern Abu Dhabi, while 2.28% of larvae released in eastern Abu Dhabi moved the opposite way. Western Abu Dhabi seemed isolated from all other sites. There was no larval connection between the adjacent areas of eastern Abu Dhabi and Dubai or Ras Al Khaimah and Hormuz Straits. Most larvae from the Hormuz Straits dispersed into northern Gulf of Oman or Iran.

### DISCUSSION

#### Environmental and Coral Condition

The environmental variable immediately responsible for most coral death in the southeastern PAG is the average August daily maximum temperature, which increased over the past half century. The primary environmental disturbance used in this paper to measure species-specific variance in demographic effects was the first-ever recorded triple bleaching event that occurred in PAG during 2010/11/12 (Riegl and Purkis, 2015). Such bleaching disturbances are predicted to be common from about 2050 onward (Donner et al., 2005; vanHooidonk et al., 2013) and in 2016/17, a double bleaching event has compromised reef health on the hitherto relatively unaffected Great Barrier Reef (Hughes et al., 2017a).

The principle of species winnowing by environmental disturbances is compatible with concepts like "winners and losers" in response to bleaching (Loya et al., 2001; vanWoesik et al., 2011). Concomitant changes in community structure are not merely theoretical expectations but have already been observed in PAG and elsewhere (Kayanne et al., 2002; Purkis and Riegl, 2005; McClanahan et al., 2011; Edmunds et al., 2014; Harii et al., 2014; McClanahan, 2014, among many others) and it has been shown that coral recovery is delayed in a warming ocean (Osborne et al., 2017). Our study demonstrates the primary mechanisms causing delayed recovery, namely reduced population fertility due to tissue death resulting in fewer, and smaller coral colonies (**Figure 6B**). This, in turn, requires more recruitment and more connectivity. The former is difficult to achieve in degraded populations (**Figure 7**), the latter is limited by oceanography (**Figure 8**).

The disappearance of three previously common species, through much of their former distribution in southeastern PAG, supports the results of our study. All were highly susceptible to bleaching (Loya et al., 2001; McClanahan et al., 2011; vanWoesik et al., 2012). Stylophora pistillata was zone-forming in Tarut Bay (Coles et al., 1995) and quite common in Abu Dhabi until 2010 (pers.comm.). It has disappeared from most of Saudi Arabia and Abu Dhabi (Coles, pers. comm.), but is present in cooler waters at Iranian islands near Hormuz Straits (Samimi-Namin pers.obs. in Riegl et al., 2012a). Pocillopora damicornis was zone-forming at Jana, Saudi Arabia (Vogt, 1995) and present throughout Abu Dhabi. No records exist from Saudi Arabia since 1998 or Abu Dhabi since 2010 but it also still occurs in Iran (Samimi-Namin pers.obs. in Riegl et al., 2012a). Montipora spp. (M. aequituberculata and M. circumvallata listed by Vogt (1993); M. aequituberculata, M. informis, and M. spongiosa listed by F. Benzoni in Riegl et al., 2012a) were common around Qatar until the 1980s(Emara et al., 1985) but are no longer present (Sheppard et al., 2010). The most recent record of this genus comprises three colonies from Kuwait (F. Benzoni in Riegl et al., 2012a). Elsewhere it is only recorded from Iran near Hormuz Straits (Samimi-Namin et al., 2009; and Samimi-Namin in Riegl et al., 2012a).

### Matrix-Based Derivation of Coral Population Dynamics

While differential resilience to environmental stresses has been documented, exact demographic mechanisms of persistence or extinction are less clear (but see Edmunds and Elahi, 2007; Edmunds, 2010; Roth et al., 2010; Doropoulos et al., 2015). Dynamics of PAG species, with respect to potential winnowing from the community, fell into two categories. Firstly, those that were immediately affected by environmental disturbance in its first year (2010), and suffered either predominantly wholecolony mortality and some partial mortality (A. downingi) or predominantly partial mortality and only little whole-colony mortality (P. harrisoni). In both species, populations consisted of markedly smaller corals after disturbance and had changed survival probabilities (**Figure 4**). In A. downingi, both average coral size as well as population size shrank, in P. harrisoni population size increased due to abundant colony fission, but average colony size shrank. These corals are referred-to as the "shrinkers." The second group consisted of those species that were not affected in the first, but only the second and/or third disturbance, and while affected by partial mortality, it was less dramatic than in the "shrinkers." This group consisted of D. pallida, P. daedalea, C. microphthalma (**Figure 4**) and is referred to as the "persisters." Interestingly, temperate corals seem to show different reactions to heating. Populations of temperate, shallow water azooxanthellate corals skewed increasingly toward the larger sizes with heating (Goffredo et al., 2008) or showed no effects at all (Caroselli et al., 2012).

Shrinkage by partial mortality is problematic for reef corals, since smaller colonies tend to have a higher mortality rate (Edmunds, 2010). Since fecundity tends to scale with size (Hughes et al., 2000), it necessarily declines in a population of smaller corals. Consequently, significant losses in fecundity were observed (**Figure 6**). However, while sexual fecundity (fertility) was lost, asexual fecundity increased due to abundant fission into surviving tissue remnants. This survival was most marked in the columnar P. harrisoni. Tips of columns tended to survive as isolated ramets, while the areas between columns frequently died (hence the very strong change in the matrix value signifying transition from SC4:SC1, **Figures 4**, **5**). In A. downingi, there was no clear pattern where tissues survived preferentially but regeneration of several ramets, originating from a single parent colony, was frequent. Also the "persisters" formed many new ramets from tissue remnants.

The abundant production of ramets consisting of variablysized tissue fragments was important in the modeling of population dynamics. Due to the difficulty of correctly identifying small recruits, the smallest size class (SC1) integrated young-of-the-year as well as smaller ramets. This had the effect of raising the population's apparent rate of increase (dominant eigenvalue of the transition matrix; Caswell, 2001; **Figure 6A**) after the disturbances. While at first sight counterintuitive, the large number of ramets can indeed increase the numerical frequency of larger colonies as they grow (unless ramets fuse again into bigger colonies), but at the cost of genetic diversity.

Even if population growth rate was boosted temporarily by the ramets, declined fertility of smaller colonies may have the opposite effect in the long term. Smaller colonies tend to be less fertile (sexually), and metapopulations of shrinking corals may find recruit flow among subpopulations increasingly compromised. Primarily asexually reproducing populations would be, at least temporarily, lost to the metapopulation with the consequence of breaking connectivity ratchets and endangering the persistence of downstream populations, if these depended on imported recruitment. The metapopulation model suggested potentially negative effects on A. downingi, a species requiring good connectivity. Shrinking (both numerically and in colony size) populations required help from connected populations and the need for connectivity increased with correlation of disturbance across the metapopulation. Graham and vanWoesik (2013) showed that ramets of some corals (in the family Merulinidae, formerly Faviidae, related to the PAG "persisters"), can remain fertile even if shrunk below puberty-size. If that was also the case in PAG corals, a strong regeneration mechanism would exist, even in the "shrinkers" (especially P. harrisoni) since fertility and connectivity might decline less than in model calculations. Also, vanWoesik et al. (2011) found that small (<5 cm, so SC1) Acropora fared better than many other corals, at least in the short term, and that fragmentation and survival from remnants was a long-term winning strategy. Thus, shrinking may not be all problematic and might include some defense mechanisms. This suggests that hope may exist for the "shrinkers" but most certainly for the "persisters," even if they should experience partial mortality in bad periods.

### Connectivity Considerations

The problems created to PAG corals by declining fecundity would be compounded by a dispersal pattern with strong local retention and connectivity primarily among neighboring sub-populations. Further, if the dominant reproductive mode of the population were to shift, impacts of declining overall fertility could be compounded since brooding corals release fewer larvae into the water column than broadcasting species. In the case of the investigated PAG corals, all were broadcast spawners (Bauman et al., 2011), thus changes in the overall number of larvae available for connectivity here would be a simple function of changes in the fertile coral tissue area. The metapopulation suggests highest connectivity requirement under correlated disturbances, i.e., when neighboring populations are stressed. Tissue mortality in neighboring populations would reduce fertility and thus larvae flow, causing disadvantages especially to populations that are highly self-seeding and the connected downstream populations. The most isolated populations would be winnowed first due to the absence of larval rescue. Highly isolated western Abu Dhabi (**Figure 8**), with higher temperature extremes than eastern Abu Dhabi (Riegl and Purkis, 2012) is home to dense coral growth devoid of Acropora, Monipora, Stylophora, and Pocillopora (Foster et al., 2012; Bauman et al., 2013)—the most heat-sensitive coral genera showing whole-colony mortality and, as demonstrated for A. downingi, a high connectivity requirement. Species winnowing as envisaged in this paper seems to have already occurred in PAG, certainly helped by its apparent reproductive isolation.

The observed and modeled population dynamics of PAG corals suggested a hierarchy of negative demographic effects by environmental stress from the "shrinkers," where A. downingi was worse affected than P. harrisoni, to the "persisters." Within the latter, the hierarchy was less clear, but C. microphthalma appeared as a particularly robust species with the least fertility loss and least changes in demographic parameters. Based on the knowledge of environmental disturbances and demographic responses, a generalized scheme of the winnowing response in PAG coral communities was developed (**Figure 9**). The differentially strong reaction among species to environmental stressors seems to generate disturbance filters related to the thresholds of resistance of individual species. In the discussion regarding changed coral communities in future seas, a better anticipation of which species are likely to get winnowed at specific environmental filters may be a valuable management tool. Many predictions as to impact severity and positioning of potential refuges exist (Riegl and Piller, 2003; vanHooidonk et al., 2013; Cacciapaglia and vanWoesik, 2015) which will be supported and strengthened by an understanding of the demographic realities (such as, shrinking, persisting, and connectivity) that can make or break a refuge.

Since many environmental stressors are ultimately driven by human societal behavior (Hughes et al., 2017b), demonstration of the filters and the reality of species winnowing that cause loss of diversity in reefs can help to predict the evolutionary price tag that will be associated with the absence of effective conservation or climate action or misguided management approaches. PAG

observations and literature (Sheppard, 1988; Sheppard and Sheppard, 1993; Riegl, 1999; Foster et al., 2012; Bauman et al., 2013; Riegl and Purkis, 2015).

has witnessed severe loss of coral reefs (Sheppard et al., 2010; Sale et al., 2011). Only swift and decisive management action utilizing information as presented within this paper can help to stabilize the continuously declining quality of the region's coral reefs—its most biodiverse but most imperiled ecosystem.

#### AUTHOR CONTRIBUTIONS

BR conceptualized the paper and wrote matrix models, GC wrote oceanographic model, SP and AB collected coral data, AB, DF, and SS did meta-analyses. All authors wrote and corrected the manuscript.

### REFERENCES

Caswell, H. (2001). Matrix Population Models. Sunderland: Sinauer, 1–722.

coral range expansion under climate change. Divers. Distrib. 20, 859–871. doi: 10.1111/ddi.12184

of Baiyangdian Lake in northern China. Hydrol. Earth Syst. Sci. 18, 2113–2126. doi: 10.5194/hess-18-2113-2014

**Conflict of Interest Statement:** The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Copyright © 2017 Riegl, Cavalcante, Bauman, Feary, Steiner and Purkis. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) or licensor are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Settlement Patterns of Corals and other Benthos on Reefs with Divergent Environments and Disturbances Histories around the Northeastern Arabian Peninsula

#### Rita Bento1, 2 \*, David A. Feary 3 , Andrew S. Hoey 4 and John A. Burt 5

*1 Abel Salazar Biomedical Sciences Institute (ICBAS), Porto University, Porto, Portugal, 2 Emirates Diving Association, Dubai, UAE, 3 School of Life Sciences, University of Nottingham, Nottingham, United Kingdom, 4 ARC Centre of Excellence for Coral Reef Studies, James Cook University, Townsville, QLD, Australia, 5 Center for Genomics and Systems Biology, New York University Abu Dhabi, Abu Dhabi, UAE*

#### *Edited by:*

*Verena Schoepf, University of Western Australia, Australia*

#### *Reviewed by:*

*Steve L. Coles, Bernice P. Bishop Museum, United States Andrew Halford, Department of Parks and Wildlife, Australia*

> *\*Correspondence: Rita Bento<EMAIL_ADDRESS>
#### *Specialty section:*

*This article was submitted to Coral Reef Research, a section of the journal Frontiers in Marine Science*

*Received: 28 May 2017 Accepted: 05 September 2017 Published: 20 September 2017*

#### *Citation:*

*Bento R, Feary DA, Hoey AS and Burt JA (2017) Settlement Patterns of Corals and other Benthos on Reefs with Divergent Environments and Disturbances Histories around the Northeastern Arabian Peninsula. Front. Mar. Sci. 4:305. doi: 10.3389/fmars.2017.00305* Larval supply is a principal factor determining the establishment, structure, and diversity of sessile benthic assemblages on coral reefs. Benthic reef communities in north-eastern Arabia have been subject to recurrent disturbances in recent years, and subsequent recovery will be, in part, driven by variation in the supply of available colonists. Using settlement tiles deployed seasonally over 1 year at eight sites encompassing three environmentally divergent regions (southern Arabian Gulf, the Musandam Peninsula in the Strait of Hormuz, and the Sea of Oman) we assessed spatial and seasonal variability in settlement of benthic reef organisms. There was strong spatial variation in composition of new colonists among regions, mainly driven by the high abundance of coralline algae in the Arabian Gulf, colonial ascidians on the Musandam Peninsula and barnacles in the Sea of Oman. Seasonal differences in composition of new colonists were less important than regional differences, with seasonal variation in settlement not consistent among regions. The number of corals settling to the tiles was low compared to those reported for other regions, with mean densities ranging from 0 corals m−2 year−1 in the Sea of Oman to 30 (± 0.6 SE) and 38 (± 0.5 SE) in Musandam and the Arabian Gulf, respectively. Peak coral settlement abundance in the Gulf occurred in summer and autumn and in Musandam in spring (averaging 82 and 70 settlers m−2 year−1 , respectively, during the peak settlement season). This work provides the first record of large-scale spatial and seasonal patterns of settlement in north-eastern Arabia and provides valuable information on the supply of settlers available to recolonize heavily disturbed reefs in this region. The extremely low rates of coral settlement suggest that these marginal reefs are likely to be extremely slow to recover from on-going and future disturbances.

Keywords: Arabian Peninsula, benthic communities, coral recruits, settlement tiles, seasonality, sessile invertebrates

## INTRODUCTION

Coral reefs provide a variety of significant socio-economic benefits to coastal communities (Ferrario et al., 2014; Guannel et al., 2016), however, climate change and increasing local pressures are jeopardizing the future of reefs (Ateweberhan et al., 2013; Bruno, 2013; Spalding and Brown, 2015). Over the past several decades, reefs have been increasingly degraded by overfishing, pollution, sedimentation, disease and coral predator outbreaks (Maina et al., 2013; Riegl et al., 2013; Pollock et al., 2014; Wear and Thurber, 2015; Mumby, 2016). These stressors have resulted in an estimated loss of up to 50% of coral cover from many reefs in the past several decades (De'ath et al., 2012; Hughes et al., 2017), with a third of reef-building coral species now considered at risk of extinction (Carpenter et al., 2008).

Disturbances such as bleaching, cyclones and crown of thorns outbreaks that cause large-scale coral mortality are becoming increasingly common (Spalding and Brown, 2015). Whether reefs can recover from these disturbances will depend, in part, on the supply of colonists available to settle to newly opened space that appears on reefs following a disturbance (Gilmour et al., 2013). Both stochastic processes which affect the supply of potential colonists and deterministic processes such as habitat availability, larval preferences, and interactions with existing community members, can influence settlement patterns (Lillis et al., 2016). Post-settlement processes then further shape subsequent community development (Caley et al., 1996). As a result, the trajectory of community recovery following disturbance is highly dynamic, with some studies reporting a relatively rapid return to pre-disturbance assemblage structure (Halford et al., 2004; Adjeroud et al., 2009), while others have observed dramatic and long-term shifts (Roff et al., 2015; Guest et al., 2016).

Historically, studies of the role of supply in recovery dynamics have focused on coral recruitment, and have largely overlooked settlement by other members of the benthic community (Zhang et al., 2014). However, algae, sponges, ascidians, and various other sessile members of the benthos are also abundant and integral members of reef communities that support diverse functional roles (Mallela, 2007; Bell, 2008; Glynn and Enochs, 2011; De Goeij et al., 2013; Enochs and Glynn, 2017), although the role that these organisms play in recovery dynamics is not well understood. Non-coral benthos are typically far more abundant than coral spat in early settlement communities (Dunstan and Johnson, 1998; Díaz-Castañeda and Almeda-Jauregui, 1999; Glassom et al., 2004; Mangubhai et al., 2007; Stubler et al., 2016), and initial colonization patterns by these organisms can strongly influence the trajectory of subsequent community development (Stubler et al., 2016). Non-coral settlement patterns may translate into long-term shifts in reef community structure, as variation in the early recruitment rates of non-coral benthos can considerably influence adult abundance of these organisms (Jackson, 1984; Caley et al., 1996; Cowen and Sponaugle, 2009; Zabin, 2015). Further, many non-coral settlers can alternately inhibit or facilitate subsequent colonization by corals (Dunstan and Johnson, 1998; Mangubhai et al., 2007; Birrell et al., 2008; Arnold et al., 2010; Diaz-Pulido et al., 2010). Thus, developing an understanding of the settlement of non-coral benthos can provide valuable insights into the role that recruitment across the wider benthic community may play in affecting the trajectory of initial community development on reefs following disturbance.

Coral reefs in north-eastern Arabia have been subject to widespread and substantial disturbances in recent years. Recurrent bleaching events and disease outbreaks have heavily affected reefs in the southern Persian/Arabian Gulf (hereafter 'the Gulf') while reefs in the adjacent Sea of Oman have been impacted by a super-cyclone and a large-scale harmful algal bloom (Bauman et al., 2010; Riegl and Purkis, 2015; Burt et al., 2016). All of these disturbances have caused substantial declines in coral cover and shifts in the composition of the wider benthic community (Bento et al., 2016). Recovery of reef communities in subsequent years has been variable, with a return toward pre-disturbance assemblages observed in some locations but not others (Burt et al., 2008, 2011; Bento et al., 2016). Several studies have suggested that this divergence in recovery patterns can largely be attributed to variation in the abundance and composition of juvenile corals that have recruited to these reefs (Burt et al., 2008; Pratchett et al., 2017). However, it is unclear whether these patterns in juvenile corals (up to 5 cm diameter) were primarily shaped by larval supply or postsettlement processes.

The purpose of this study was to investigate spatial and seasonal variability in settlement of benthic organisms on reefs spanning > 750 km of coastline in the north-eastern Arabian Peninsula. Terra-cotta settlement tiles were deployed seasonally over 1 year on reefs in the highly disturbed Gulf and Sea of Oman, and on relatively undisturbed reefs in the Musandam Peninsula. Comparison of seasonal dynamics provides insights into peak settlement periods for sessile benthic invertebrates, while large-scale spatial comparisons provide insights into regional scale disturbance and recovery dynamics.

### MATERIALS AND METHODS

This study was performed across eight sites in three regions spanning >750 km of coastline in the north-eastern Arabian Peninsula: three sites in the southern Arabian Gulf (Saadiyat, Dhabiya, and Ras Ghanada), three sites along the Musandam Peninsula (Al Harf, Falcon Rock, and Coral Garden) and two sites in the Sea of Oman (Dibba and Al Aqa) (**Figure 1**). All sites were fringing reefs at comparable depths (2–8 m) with coral cover varying between ca. 20–60% among reefs. Communities were generally dominated by faviids and poritids in all regions, with other subdominant families differing in relative abundance among sites. A full description of the benthos in each region is provided in Bento et al. (2016).

Settlement of benthic organisms was quantified at all sites using unglazed terracotta settlement tiles (10 × 10 × 1.5 cm) following methods adapted from Mundy (2000). At each site, 30 tiles were attached to the substratum at ca. 5 m depth and spaced 1–3 m apart. Each tile was secured to the reef using a stainless steel stud that was epoxied into the substratum and

passed through a 1 cm hole in the center of each tile. A 2 cm plastic washer was placed over each stud to position the tile 2 cm above the reef substratum. The textured (i.e., corrugated) surface of each tile was always positioned facing the substratum, as previous studies have shown that settlement primarily occurs on the underside of tiles and that textured materials generally have higher settlement than smooth surfaces (Burt et al., 2009). To assess seasonal variation in settlement, tiles were deployed and replaced every 3 months over 1 year. Each deployment represented a specific season: (summer: July–September 2012, autumn: October–December 2012, winter: January–March 2013, and spring: April–June 2013), with all tiles deployed/collected across all sites within ca. 7–10 days at the beginning/end of each season. These four seasons were selected based on the periods of highest (summer) and lowest (winter) sea surface temperatures, as well as the transitional spring and autumn periods that are known to be important discrete spawning and/or settlement seasons for a variety of marine fauna in this region (e.g., Bauman et al., 2011; John, 2012; Howells et al., 2014).

Upon retrieval, the bottom of each tile was photographed with a 10 megapixel Nikon D-80 digital camera fitted with a macro lens. Only the bottom surfaces of tiles were analyzed as virtually all settlement in this region occurs on the bottoms of tiles (Burt et al., 2009; Bauman et al., 2014). Percent cover of various benthos on each tile was calculated by image analyses using the software CPC with Excel extension (Kohler and Gill, 2006), with coverage tabulated from 50 random point intercepts per tile. Substratum and benthic community type was categorized into 15 broad groups and several sub-groups: bare tile, coralline algae, coral, ascidiacea (subgroups: colonial ascidian, solitary ascidian), sponge, cnidaria (subgroups: anemone, hydrozoa, zoanthid), arthropoda (subgroup: barnacle), mollusca (subgroups: bivalve, chiton, gastropoda), annelid (subgroup: polychaeta), bryozoa, other live (subgroups: mobile invertebrate, urchin, other), algae (subgroups: algae, turf algae), sand/silt, non-benthos (subgroups: gap, tape, shadow), and unknown taxa. Due to the small size of coral recruits (age: ≤3 mo) and their relatively limited coverage, density of coral juveniles was tabulated separately. Following photography for community analysis, tiles were cleaned of living tissue in bleach for 24 h, rinsed and air dried before tabulating the number of coral recruits on each tile using an Olympus DP-70 stereo-microscope (40X magnification).

#### Data Analyses

Before statistical analyses, percent cover for the various benthic groups was standardized as a proportion of the total living benthos (i.e., relative abundance), and an arcsine-square root transformation was applied to normalize the data, as recommended by Zar (2010). Coral densities were log(x + 1) transformed. Prior to multivariate analyses, benthic categories occurring in <5% of samples were excluded from analyses to avoid the influence of outliers (McCune et al., 2002).

To explore the overall spatial and seasonal structure of settlement assemblages, and to identify which components of the benthos were driving any settlement differences, multivariate analyses were performed using Primer, v6 (Clarke and Gorley, 2006). Non-metric multi-dimensional scaling (nMDS) analyses based on Bray-Curtis distance matrices were used to illustrate the influence of regions and seasons on benthic community composition, with taxa strongly driving divergence along nMDS axes identified using Pearson rank-correlations (r ≥ ±0.5 for either axis). To ease interpretation, each season was manually shaded in the resulting nMDS and the vector plot overlaid. Spatial and seasonal differences in settlement community structure were tested with a partially-nested permutational multivariate analysis of variance (PERMANOVA) on the main effects of seasons and regions with sites nested within regions. A similarity percentages analysis (SIMPER) was used to assess which benthic components contributed most to the observed variation in settlement community structure (Clarke and Gorley, 2006). Spatial and seasonal differences in settlement community structure were tested with a partially-nested permutational multivariate analysis of variance (PERMANOVA) on the main effects of seasons and regions with sites nested within regions based on our a priori hypotheses. Key benthic groups and sub-groups identified by our multivariate analyses and coral settlement densities were then then tested with univariate PERMANOVAs to identify significant settlement differences.

A multiple regression was also employed to determine whether the density of coral spat was associated with percent cover of any other components of the settlement community (including each live benthos category plus bare tile). Before regression analyses, the normality of the residuals and homogeneity of variances were confirmed by plotting residuals against fitted values and using QQ plots (Zuur et al., 2007).

#### RESULTS

Multivariate ordination of the overall settlement community observed in this study indicated strong differences in community structure between regions, with individual sites clustering as region-specific groups across all seasons with no overlap among regions (**Figure 2**). Vector plots indicated regional differences were primarily driven by variation in the relative abundance of three benthic groups: coralline algae, colonial ascidians, and barnacles (**Figure 2**), although other benthos also made strong contributions (Supplementary Table 1). SIMPER analyses showed that settlement communities in the Gulf were primarily characterized by the presence of coralline algae, polychaetes, and bryozoans, which together contributed >80% to similarity in tile assemblage structure in this region (**Table 1**). In the Musandam region, over half of the similarity in settlement communities was driven by strong contributions from both colonial ascidians and bryozoans, with polychaetes and turf algae making more modest contributions (**Table 1**). Bryozoans were also the most common component of Sea of Oman settlement communities, where they contributed over a third of the similarity in community structure, followed by barnacles and turf algae also playing important roles. The SIMPER analysis showed higher abundances of

coralline algae in the Gulf was the main driver of differences between both Musandam and Sea of Oman communities, where colonial ascidians, bryozoans, and turf algae were more abundant (**Table 1**). Differences between Musandam and the Sea of Oman settlement communities were largely driven by higher cover of colonial ascidians in the Musandam vs. cover of bryozoans, barnacles, polychaetes and turf algae in Sea of Oman sites (**Table 1**).

While assemblage structure primarily grouped regionally, there were also modest seasonal changes in the settlement community within each region (**Figure 2**), although these changes were not consistent within regions as shown by a significant interaction between seasons and regions in PERMANOVA (**Table 2**). In the Gulf the settlement community showed considerable overlap in structure across summer, autumn, and winter, indicating a good degree of similarity across these seasons, with divergence in the remaining season (spring) being primarily due to higher abundances of coralline algae settling into a single site (Ras Ghanada; **Figure 2**) within this season. In the Musandam the settlement community composition was more similar to Gulf communities during the warm summer season but diverged in the cooler autumn and spring seasons, with the greatest divergence from the Gulf occurring in the cold winter season, when all three Musandam sites had highly convergent community structure, with higher settlement of colonial ascidians (**Figure 2**). The settlement community in the Sea of Oman varied spatiotemporally. The community at Dibba was fairly consistent across summer, autumn, and spring, while the community at Al Aqa was comparable to that found within Dibba in spring, but diverged markedly in the summer and autumn (**Figure 2**). In the winter there were dramatic shifts in the settlement community at both of the Sea of Oman sites, with the assemblages at the two sites TABLE 1 | Benthic groups responsible for >90% of within-regions similarities and >70% of among-regions dissimilarities based on SIMPER analysis.

Turf algae 12.3 86.3 Polychaeta 19.2 7.8 12.4 57.6 Barnacle 5.4 91.7 Turf algae 5.6 14.3 11.2 68.8

*Mean relative abundance (as % of live cover) in each region and individual (Cont.) and cumulative (Cum.) contribution to dissimilarity are shown.*

TABLE 2 | Results of the three factor partially nested PERMANOVA analyses assessing differences between regions, seasons and sites (within regions) in benthic community composition.

converging with each other but diverging from the communities observed in all other seasons in this region, due to the settlement of barnacles in both sites. SIMPER analyses (**Table 3**) showed that the major taxa driving seasonal changes largely reflected the same groups that were associated with regional differences (above; **Figure 2**), indicating that fluctuations in the relative abundance of these key regional taxa was the primary driver of seasonal shifts in settlement communities.

SIMPER analyses were further used to examine the main benthos driving seasonal differences within regional settlement communities (**Table 3**). The only benthic group that showed wide seasonal variation in all three regions were the bryozoans; all other benthos that were key drivers of seasonal change were important in just one or two regions. In addition to bryozoans, seasonal differences in the Gulf were mainly attributable to variation in the amount of coralline algae, colonial ascidians, and polychaetes, in the Musandam to colonial ascidians, and in the Sea of Oman to barnacles. These seasonal drivers largely reflected the same benthos associated with regional community differences (above; **Figure 2**).

There was a significant interaction between the main effects of season, region and sites (within region) in structuring settlement of each major benthos (**Figure 3**), indicating that patterns of change were not consistent among the main effects for any of these benthic groups (PERMANOVA Pseudo- F(15, 824): bryozoans = 11.0, polychaetes = 7.2, turf algae = 5.0, coralline algae = 3.9, barnacles = 7.4, colonial ascidian = 2.7; p < 0.001 for each]. Bryozoans showed strong but inconsistent seasonal fluctuations in all regions, with highest cover in the winter in the Gulf and Musandam (**Figures 3A,B**), while densities peaked in the autumn in the Sea of Oman when cover was nearly 50% higher than peak densities in the other regions (**Figure 3C**). While coralline algae cover was negligible in other regions, it

TABLE 3 | Benthic taxa identified as important seasonal drivers of divergence in community structure in SIMPER.

*Only benthic groups strongly contributing to the dissimilarity (*δ >*15%) in each pairwise comparison are shown.*

was the most abundant benthic component on tiles in the Gulf, and it doubled in cover from winter to summer (mean cover: 16.3% ± 2.2 SE to 31.6% ± 2.7 SE respectively; **Figure 3A**). Colonial ascidians were considerably more abundant in the Musandam than other regions, explaining its identification as a major distinguishing taxa for tiles in this region in earlier SIMPER analyses, and cover was generally comparable across seasons for this group (**Figure 3B**). In the Sea of Oman barnacles heavily dominated settlement in winter, when they comprised over half of the benthos at each site and were substantially more widespread than in other seasons; barnacles were uncommon in Musandam and nearly non-existent in the Gulf.

#### Coral Recruits

In total, 216 coral recruits were observed on the 845 tiles deployed across the 8 sites throughout the study. No coral settlers were observed across the entire study in the Gulf of Oman, and there was inconsistent and high variation in coral settlement among sites and seasons within the Gulf and Musandam (**Figure 4**; **Table 4**). In the Gulf, mean settlement was highest in autumn but this was mainly due to a large pulse at Saadiyat reef (90% of spat this season); a second peak in settlement also occurred in summer, with a small number of spat recorded in spring as well. In the Musandam, coral settlement mainly occurred in the spring, although spat were observed in low densities in all seasons in at least one site.

Multiple regression revealed that coral density was unrelated to cover of other benthic components [including bare tile space; R 2 = 0.27; F(10, 96) = 0.736, p = 0.69], reflecting the considerable variability and low incidence of coral settlement observed in this study.

### DISCUSSION

Coral reef communities in north-eastern Arabia have been subject to various large-scale disturbances in recent years. To date, there had been limited knowledge of the spatial and seasonal patterns of settlement of coral and non-coral benthos, limiting our understanding of the important role that supply may play following disturbance. The results of this study show that settlement of corals was extremely low and that non-coral benthic settlement was highly variable between regions and across sites within regions.

The most striking result of this study was the strong regionally structured patterns of non-coral settlement, with regional differences persisting across seasons throughout the study. These large-scale differences in non-coral settlement mirror the divergence of benthic community structure on reefs across these regions (Bauman et al., 2013; Bento et al., 2016), and is likely driven by the highly divergent environmental conditions among these seas, which have also been subject to varying degrees of disturbance.

The southern Gulf represents one of the most extreme coral reef environments on earth, with sea surface temperatures ranging >25◦C annually, daily mean maxima >34◦C for several months during summer, and salinities that consistently exceed 44 PSU (Sheppard et al., 1992; Coles, 2003; Foster et al., 2012). In addition, these reefs have experienced numerous large-scale

bleaching events over the past two decades (Riegl and Purkis, 2015), resulting in widespread shifts in benthic community structure from which there has been only limited recovery (George and John, 2000; Sheppard and Loughland, 2002; Burt et al., 2008, 2011). In the wake of earlier mass bleaching events, coralline algae dramatically increased in abundance on southern Gulf reefs (George and John, 2000, 2002). In the current study, coralline algae dominated the settlement community on southern Gulf reefs (covering 28% of tile surfaces, on average), and it was the primary differentiator in settlement community structure from the other regions (where it covered <3.5% of tile surfaces in both areas). Thus, the presence of coralline algae as a major component of the settlement community in the southern Gulf appears to be a persistent, long-term characteristic of this highly disturbed, extreme environment.

In contrast to the Gulf, the Sea of Oman has environmental conditions that are more benign. Due to its greater depth and exchange with the wider Indian Ocean, SSTs in the Gulf of Oman are less extreme (mean summer maxima <32◦C, range 10◦C annually) and salinity is comparable to oceanic conditions (37 PSU) (Foster et al., 2012; Howells et al., 2014). Its waters are also highly productive compared with the Gulf and Musandam as a result of monsoon-induced upwelling (Sheppard et al., 1992). Although reefs in the Sea of Oman have experienced widespread disturbance in the past decade as a result of cyclone storm damage and a hypoxic event associated with an algal bloom (Bauman et al., 2010; Burt et al., 2016), the frequency of impacts to coral reefs here has not been as severe as in the Gulf and there are indications that recovery of reef communities is underway (Bento et al., 2016; Pratchett et al., 2017). Barnacles were one of the primary drivers of the divergence of Sea of Oman reefs from the other regions, with barnacles here covering nearly a quarter of settlement tiles (22%) on average, compared with low abundance in the Musandam (5% cover) and a near absence in the Gulf (0.2% cover). Barnacles are reported to be among the most abundant members of settlement communities in Oman (Wallström et al., 2011; Dobretsov et al., 2013; Polman et al., 2013; Dobretsov, 2015), likely a reflection of their success as filterfeeders in this high-productivity environment (Sheppard et al., 1992), suggesting that the high barnacle abundance we observed is a result of long-term supply dominance in the Sea of Oman.

The Musandam Peninsula sits at the interface between the Arabian Gulf and the Sea of Oman at the Strait of Hormuz. Environmental conditions in the Musandam are generally comparable to the Sea of Oman, although productivity is lower due to a lack of monsoonal upwelling in this area (Sheppard and Salm, 1988; Sheppard et al., 1992; Reynolds, 1993). In addition to having relatively benign environmental conditions, Musandam reefs have also escaped the various large-scale disturbances that affected reefs in the adjacent seas, and reef communities here have among the highest coral cover and diversity in northeastern Arabia and are considered among the most pristine in the region (Sheppard et al., 2010; Bento et al., 2016; Burt et al., 2016). Colonial ascidians were the primary driver of divergence of Musandam reefs from those in the other regions, with these organisms covering nearly a third of Musandam tiles (30.7% cover), nearly triple the coverage in the other regions

TABLE 4 | Results from PERMANOVA tests comparing mean coral recruits densities among regions, sites and seasons.

*Values in bold are significant (p* < *0.05).*

(8.6–12.5% cover). Ascidians are relatively common members of the benthic community on reefs in the Musandam (R. Bento, unpubl. data), particularly compared to reefs in the southern Gulf or Sea of Oman where they are virtually undetected in benthic surveys (Burt et al., 2011; Grizzle et al., 2016). Given their short pelagic duration and limited larval swimming ability, ascidians generally have a relatively localized dispersal (Shanks et al., 2003; Weersing and Toonen, 2009), suggesting that the high abundance of colonial ascidians observed on Musandam settlement plates likely relates to their high abundance in the wider reef community.

While regional differences primarily structured settlement communities, there were also modest within-region shifts in settlement communities over the course of the year. Much of these within-region shifts were related to fluctuations in the abundance of bryozoans, a dominant member of the community across all regions, as well as to fluctuations in the relative abundance of those same key benthic categories that drove between-region differences. In all three regions the period of peak bryozoan settlement coincided with a reciprocal decline in the abundance of the region-specific key taxa; peak bryozoan abundance in the winter coincided with the lowest annual mean cover of coralline algae in the Gulf and colonial ascidians in the Musandam, while peak bryozoan settlement in the autumn in the Sea of Oman and coincided with a decline of barnacle cover on tiles. In all cases, peak coverage of bryozoans occurred in the season with maximum chlorophyll-a concentrations for every region (Nezlin et al., 2010; Piontkovski et al., 2011; Moradi and Kabiri, 2015), suggesting that while bryozoans settle year round, their success in these particular seasons may be related to enhanced post-settlement growth due to higher planktonic food availability. Together, these observations suggest that while there are distinct regional settlement signatures, seasonality in bryozoan settlement is largely responsible for modulating withinregion temporal dynamics. It should be noted, however, that these results represent seasonal settlement data from a single year. While this is useful information given the lack of data available previously, benthic recruitment is highly variable from year to year and it is unknown whether the results of this study are representative of longer-term patterns. A multi-year year settlement study to assess the role of recruitment in sessile benthic population dynamics is warranted.

### Coral Settlement Patterns

Coral settlement was low across all three regions in this study. Across the full year of study, coral densities were minimal in the Gulf and Musandam (mean: 38 and 30 coral settlers m−2 year−1 , respectively, across the year, and averaging 82 and 70 settlers m−2 year−1 even when considering only the peak settlement season), and corals were entirely absent from settlement tiles in the Sea of Oman throughout the study. The observed densities were substantially lower than has commonly been reported in tropical reef environments (Seychelles: 595 spat m−2 year−1 , Chong-Seng et al., 2014; Indonesia: 286–705 m−2 year−1 , Sawall et al., 2013; Kenya: 101–908 m−2 year−1 , Mangubhai et al., 2007), and less than half of densities reported for comparable high latitude marginal reefs (Eilat: 190 m2 year−1 , Glassom and Chadwick, 2006; Solitary Islands: 132 spat m−2 year−1 , Harriott and Banks, 1995; Taiwan: 111 spat m−2 year−1 , Soong et al., 2003). The observed densities are also low compared with a recent study of coral settlement in Dubai in the southern Gulf where densities of 121 coral settlers m−2 year−1 were reported (Bauman et al., 2014), although those data were mainly collected on breakwaters that have been suggested to entrain eggs, potentially enhancing settlement compared with what would occur on natural reefs (Burt et al., 2009). Overall, the low densities noted in this region are similar to that seen in heavily degraded, highly disturbed reef environments (Singapore: 55 m−2 year−1 , Bauman et al., 2015; Florida: 38 m−2 year−1 , Van Woesik et al., 2014). Together, the low densities observed in the Gulf and Musandam and the complete absence of coral settlement in the Sea of Oman suggest that there is cause for concern for the regeneration of degraded coral reef communities within the Arabian Peninsula.

The low abundance of coral spat observed here is unlikely to be due to interactions with other members of the settlement community. The abundance of coral recruits was unrelated to cover of any other benthos in our analyses, and bare space was relatively common (ranging from 55% in the Gulf to 20% in the Sea of Oman), suggesting ample availability of habitat for coral settlement. Instead, the low coral settlement likely reflects a limited supply of coral larvae. Low coral settlement has been reported earlier in the southern Gulf (Bauman et al., 2014), and is likely the result of depressed fecundity of corals being exposed to recurrent bleaching events and extreme environmental conditions (Riegl and Purkis, 2015; Howells et al., 2016). Although coral cover is high in the Musandam (Bento et al., 2016; Burt et al., 2016), we observed low levels of coral settlement. This suggests that low coral recruitment is characteristic of reefs in northeastern Arabia, and the high coral cover in the Musandam is likely the result of the relatively low levels of disturbance in this area rather than high larval supply. The implication is that the relatively pristine reefs in the Musandam may be highly vulnerable to any future disturbances that may occur, as any recovery would be potentially limited by low recruitment levels. The absence of coral settlement to tiles in the Sea of Oman was surprising. Reproductive studies occurring near these sites have shown that common coral species were spawning during this study in April 2013 (Howells et al., 2014), and fecundity was substantially higher than conspecifics in the southern Gulf (Howells et al., 2016). Additionally, juvenile surveys on these sites in 2012 showed that corals were recruiting to these reefs, although densities were half of that observed in the southern Gulf (Pratchett et al., 2017), suggesting that recruitment is impaired here. Reefs in the Sea of Oman sites have low coral cover (18%) relative to the southern Gulf and the Musandam (56 and 58%, respectively, Bento et al., 2016), mainly as a result of a algal bloom in 2009 when 50–90% of corals were lost from reefs (Bauman et al., 2010; Foster et al., 2011). As a result, while individual corals here are fecund, reef-wide reproductive output for a variety of dominant coral species is low as a consequence of the limited number of fecund adults in the community (Howells et al., 2016), potentially explaining the absence of coral settlers on tiles and the relatively low overall juvenile recruitment rates observed by Pratchett et al. (2017). It should be noted, however, thatrecovery on highly disturbed reefs typically takes 10–15 years (Purkis and Riegl, 2005; Burt et al., 2008; Ateweberhan et al., 2011), even in areas where larval supply remains depressed for up to 6 years (Gilmour et al., 2013). As the current study was performed 8 years after the catastrophic HAB event, this suggests that there is hope that recovery may occur in the near future. It is unclear whether the absence of coral settlement observed here extends to other parts of the Sea of Oman. A recent survey of coral reefs 400 km southwest around Muscat, Oman, suggested that many coral reefs there continue to have high coral cover despite showing indications of localized decline in some areas over the past two decades, with available evidence suggesting that declines were primarily due to a recent cyclone rather than the dramic HAB induced loss that impacted the coral communities on reefs studied here (Coles et al., 2015). An assessment of the reproductive capacity and settlement patterns of corals on reefs elsewhere in the Sea of Oman is highly warranted.

Coral reefs around the world are becoming increasingly degraded as a result of climate change and localized anthropogenic impacts (Hughes et al., 2010; Pandolfi et al., 2011) with widespread shifts toward a dominance of noncoral benthos (Colvard and Edmunds, 2011; Kelmo et al., 2013, 2014). While many studies have focused efforts on understanding the dynamics of coral settlement following disturbance (Glassom et al., 2004; Abelson et al., 2005; Green and Edmunds, 2011; Sawall et al., 2013; Bauman et al., 2014), few studies have explored the role that non-coral benthos may play in affecting early recovery dynamics on reefs (Colvard and Edmunds, 2011; Luter et al., 2016). Non-coral benthos are among the first colonists to settle on substrates opened by disturbance, and typically reach an abundance and coverage that greatly exceeds that of coral spat (Dunstan and Johnson, 1998; Díaz-Castañeda and Almeda-Jauregui, 1999; Glassom et al., 2004; Mangubhai et al., 2007; Stubler et al., 2016). Given that many of these non-coral benthos can alternatively facilitate or inhibit subsequent coral recruitment in the space opened by disturbance, their presence could dramatically impact the trajectory of subsequent recovery on disturbed reefs. Our results showed that the initial settlement community was highly region-specific, with non-coral benthos being the primary members of the settlement community in all areas. In some regions, these initial communities were dominated by taxa known to facilitate settlement of coral larvae (for example, over a quarter of Gulf tile space was covered by crustose coralline algae, a well-known inducer of larval coral settlement (Ritson-Williams et al., 2014; Tebben et al., 2015), suggesting conditions favorable to coral colonization following disturbance. In others, however, early settlement communities were dominated by fauna known to inhibit coral larval settlement (e.g., ascidians covered a third of tiles in the Musandam, and various ascidians are known to produce allelochemicals that can inhibit coral settlement (Chadwick and Morrow, 2011). Our knowledge of the types of benthic organisms and the mechanisms that they use to interact with coral larvae is relatively under-developed (Ritson-Williams et al., 2009), but our results here suggest that such information is essential if we are to understand how the initial settlement community may affect the trajectory of early community development. Initial settlement communities will, of course, be shaped by post-settlement processes that will likely lead to mature assemblages that are markedly different in their structure. But because this first stage of development provides the foundation open which all post-settlement processes can act, it represents a critical bottleneck that has important implications for whether or not a disturbed reef has the capacity to recover.

#### AUTHOR CONTRIBUTIONS

RB: Responsible for data collection, analysis and writing process. DF and AH: Responsible for reviewing the data analysis and writing. JB: Responsible for data analysis and writing.

#### ACKNOWLEDGMENTS

This research was supported by the Ford Environmental Grant, Biosphere Expeditions, New York University Abu Dhabi and the Australian Research Council (DE130100688). We thank

#### REFERENCES

Fujairah Municipality, Dibba Municipality, and Oman Ministry of Environment and Climate Affairs for research permits. We would also like to thank the NYU Abu Dhabi Institute for field support, and particular gratitude are extended to Grace Vaughan and Dain McParland for field support.

#### SUPPLEMENTARY MATERIAL

The Supplementary Material for this article can be found online at: http://journal.frontiersin.org/article/10.3389/fmars. 2017.00305/full#supplementary-material

for coral assemblages across the United Arab Emirates. Mar. Pollut. Bull. 114, 1031–1035. doi: 10.1016/j.marpolbul.2016.11.036

**Conflict of Interest Statement:** The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

The reviewer AH declared a past co-authorship with one of the authors DF to the handling Editor, who ensured that the process met the standards of a fair and objective review.

Copyright © 2017 Bento, Feary, Hoey and Burt. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) or licensor are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Coral Community Structure and Recruitment in Seagrass Meadows

Kathryn E. Lohr 1 \*, David J. Smith2 , David J. Suggett 3 , Matthew R. Nitschke3 , Alex J. Dumbrell 2 , Stephen Woodcock 4 and Emma F. Camp2, 3

1 Program in Fisheries and Aquatic Sciences, School of Forest Resources and Conservation, Institute of Food and Agricultural Sciences, University of Florida, Gainesville, FL, United States, 2 School of Biological Sciences, University of Essex, Essex, United Kingdom, 3 Climate Change Cluster, University of Technology Sydney, Ultimo, NSW, Australia, 4 School of Mathematical and Physical Sciences, University of Technology Sydney, Ultimo, NSW, Australia

Coral communities are increasingly found to populate non-reef habitats prone to high environmental variability. Such sites include seagrass meadows, which are generally not considered optimal habitats for corals as a result of limited suitable substrate for settlement and substantial diel and seasonal fluctuations in physicochemical conditions relative to neighboring reefs. Interest in understanding the ability of corals to persist in non-reef habitats has grown, however little baseline data exists on community structure and recruitment of scleractinian corals in seagrass meadows. To determine how corals populate seagrass meadows, we surveyed the established and recruited coral community over 25 months within seagrass meadows at Little Cayman, Cayman Islands. Simultaneous surveys of established and recruited coral communities at neighboring back-reef sites were conducted for comparison. To fully understand the amount of environmental variability to which corals in each habitat were exposed, we conducted complementary surveys of physicochemical conditions in both seagrass meadows and back-reefs. Despite overall higher variability in physicochemical conditions, particularly pH, compared to the back-reef, 14 coral taxa were capable of inhabiting seagrass meadows, and multiple coral families were also found to recruit to these sites. However, coral cover and species diversity, richness, and evenness were lower at sites within seagrass meadows compared to back-reef sites. Although questions remain regarding the processes governing recruitment, these results provide evidence that seagrass beds can serve as functional habitats for corals despite high levels of environmental variability and suboptimal conditions compared to neighboring reefs.

Keywords: seagrass, coral, recruitment, environmental variability, marginal habitats, benthic ecology, climate change, marine chemistry

### INTRODUCTION

Scleractinian corals occupy a range of habitat types and form the architectural framework within different reef zones (e.g., Goreau, 1959; Logan, 2013). However, corals also colonize non-reef habitats, such as mangrove systems and seagrass beds that can exhibit high levels of variability in abiotic conditions relative to neighboring reefs (e.g., Riegl, 1999; Perry and Larcombe, 2003; Yates et al., 2014; Camp et al., 2016a). Despite the presence of corals in these non-reef habitats, many assessments of coral communities focus only on fore-reef zones (e.g., Mumby et al., 2007; Castillo et al., 2012). Few studies exist to characterize coral populations outside such classical reef systems.

Edited by:

Hajime Kayanne, The University of Tokyo, Japan

#### Reviewed by:

Steeve Comeau, University of Western Australia, Australia Susana Enríquez, Universidad Nacional Autónoma de México, Mexico

> \*Correspondence: Kathryn E. Lohr<EMAIL_ADDRESS>
#### Specialty section:

This article was submitted to Coral Reef Research, a section of the journal Frontiers in Marine Science

Received: 01 May 2017 Accepted: 16 November 2017 Published: 30 November 2017

#### Citation:

Lohr KE, Smith DJ, Suggett DJ, Nitschke MR, Dumbrell AJ, Woodcock S and Camp EF (2017) Coral Community Structure and Recruitment in Seagrass Meadows. Front. Mar. Sci. 4:388. doi: 10.3389/fmars.2017.00388

**128**

To understand the importance of non-reef habitats for corals, it is critical to establish baseline data on the structure of coral assemblages in such habitats. Habitat monitoring and assessment data is widely acknowledged as a critical factor in the success of conservation initiatives and adaptive management (e.g., Jackson et al., 2014). However, prioritizing indicators for monitoring ecosystem health can be challenging (Borja et al., 2014). Live coral cover alone is increasingly acknowledged as an insufficient metric for characterizing the ability of a coral-bearing habitat to persist over time (Hughes et al., 2010). For example, recruitment is recognized as a valuable metric for evaluating coral community function, as it can provide insight into a system's capacity for replenishment, a critical component of resilience (Nyström et al., 2008; Hughes et al., 2010). A holistic examination of a coral community should therefore characterize both established and recruited coral assemblages.

Here we focus on seagrass beds that, to-date, have been understudied with respect to their suitability as a non-reef coral habitat. Typically, seagrass beds have not been considered important coral habitats due to a lack of solid substrate (Jackson, 1986; Manzello et al., 2012). Coral communities inhabiting seagrass meadows have been reported from both the Caribbean (Rützler and Macintyre, 1982; Johnson, 1992) and Indo-Pacific (Higuchi et al., 2014; Camp et al., 2016a). These studies have focused on the ecology of free-living coral species (Johnson, 1992) and physiology of corals (Higuchi et al., 2014; Camp et al., 2016a,b) in such habitats, or have described coral colonies in seagrass meadows as part of larger reef systems (Rützler and Macintyre, 1982). Johnson (1992) investigated coral recruitment within a seagrass habitat, however this study focused on a single coral species rather than the coral community as a whole. Others have briefly noted the presence of corals living among seagrass during unrelated experiments (e.g., Wulff, 2008).

For coral communities in seagrass beds, measurements of physicochemical conditions are important given the characteristic variability in ocean chemistry associated with these sites (Higuchi et al., 2014; Camp et al., 2016a). Seagrass systems routinely experience highly variable environmental conditions relative to coral reefs, such as reduced pH and increased temperature (Semesi et al., 2009; Challener et al., 2016), which could be critical in shaping resident coral communities. Studies of corals living among these high-variability systems can therefore also provide insight into which coral species are able to successfully acclimatize or adapt to survive under such conditions (Barshis, 2015). We are unaware of any study that has comprehensively characterized coral community structure within a seagrass meadow, particularly in the context of adjacent reef communities and the physicochemical variability within the system. A lack of baseline data on the composition and condition of coral communities in seagrass meadows inhibits our ability to understand how these systems may change over time.

We conducted a 25-month study to characterize the coral community at understudied seagrass sites in a location with minimal direct anthropogenic impact (Little Cayman, Cayman Islands). To fully understand how coral communities in seagrass meadows differ from those of local reefs, adjacent back-reefs were also surveyed for comparison. We assessed key attributes to provide a robust characterization of each site, specifically: (i) physicochemical conditions, (ii) benthic structure, and (iii) coral recruitment, including the relationship between the established and recruited coral community. Our study thus provides detailed insight into an important but poorly understood coral habitat, as well as a comprehensive comparison to an adjacent but very different shallow reef habitat. In doing so, we provide novel information on how environmentally-variable seagrass beds can function as a coral habitat.

## METHODS

### Study Site

This study was conducted off the north coast of Little Cayman, Cayman Islands (19.6897◦N, 80.0367◦W), at three shallow (≤ 2 m) sites with adjacent back-reefs and seagrass meadows (**Figure 1**). The first site was located in Grape Tree Bay, (GTB), a lagoon approximately 0.10 km wide and 1.50 km long, situated within the Bloody Bay Marine Park. The two remaining sites were located at either end of the larger Mary's Bay, approximately 0.15 km wide and 4.00 km long. The two sites within Mary's Bay were designated as Mary's Bay West (MBW) and Mary's Bay East (MBE). At each site, three sampling replicates, spaced 50 m apart, were selected within each habitat type [i.e., seaward back-reef (n = 3) and inshore seagrass meadow (n = 3); **Figure 1**]. Across sites, seagrass was comprised of 88.0 ± 1.5% Thalassia testudinum and 12 ± 0.6% Syringodium filiforme.

### Physicochemical Conditions

To characterize the physicochemistry of each site, discrete seawater samples were collected over two 6-week periods, one in the wet season (July-August) and one in the dry season (February–March) of 2012. Within each sampling period, three 24 h sampling sessions were conducted during both neap and spring tidal phases. Samples were collected every 3 h over the 24 h period at each site and replicate (n = 48 total samples per site, per sampling session). In addition, diurnal (1 h before sunrise and 1 h after sunset) open-ocean water samples were collected (n = 6 samples per site) to act as an off-shore reference for the ocean chemistry. Discrete water samples were collected from a depth of ca. 0.5 m following Carbon Dioxide Information Analysis Center (CDIAC) standards. Samples were collected in borosilicate bottles and fixed with HgCl2 [0.05 ml of saturated (aq) solution] (Dickson et al., 2007). Stoppers were inserted to ensure a gas tight seal and samples were stored in the dark until they were returned to the laboratory (within 30 min) for analysis.

An additional seawater sample was collected at each replicate during every sampling period to allow immediate measurements of pH, temperature, oxygen, and salinity to obtain the in situ conditions used for calculation of carbonate parameters. An OrionTM 5 Star meter (Thermo ScientificTM) fitted with a pH/temperature probe and DO probe (Thermo ScientificTM) and a handheld refractometer to measure salinity (ExTech R ) were used throughout. The pH probe was calibrated with TRIS buffer using the potentiometric technique and the total scale (pHT). Water flow was also measured at each replicate following every sample collection using a mechanical flow meter (General

FIGURE 1 | Study sites on Little Cayman Island. Map of Little Cayman Island (A) located within the Caribbean Sea (upper inset), yellow highlighting denotes the location of marine parks. Three locations, Mary's Bay East (MBE, circles) and West (MBW, squares), and Grape Tree Bay (GTB, triangles) were chosen as study sites. Back-reef (light gray areas with dashed lines, lower inset) and seagrass (green areas, lower inset) sites were selected within each study location (n = 3 per site spaced 50 m apart, lower inset). Typical benthic scenes are indicated for back-reef and seagrass sites in (B,C), respectively. The base data for the map (A) were collected from www.openstreetmap.org (© OpenStreetMap contributors) under the Open Database License and customized in Adobe Illustrator (version 16).

Oceanics). All laboratory and field equipment were calibrated as recommended by their instruction manuals and/or CDIAC protocols (Dickson et al., 2007).

Samples were returned to a climate-controlled laboratory where pH was re-measured to check for pH drift using a pH electrode (Thermo ScientificTM) calibrated with TRIS buffer using the potentiometric technique and the total scale (pHT). An open-cell potentiometric titration procedure was used to measure total alkalinity (AT) using the Gran method to determine the second end point of the carbonate system. AT of all samples was determined using a titrator (Metrohm) with an accuracy and precision of ca. ≤ 2 µmol kg−1 as verified with certified reference materials distributed by A. Dickson (Scripps Institution of Oceanography). All carbonate parameters [pCO2, TCO2 and aragonite saturation (arg)] were calculated with CO2SYS from AT and pHT (Riebesell et al., 2000), and in situ temperature, salinity, and sampling depth (m) as a proxy for pressure (Lewis and Wallace, 1998). For CO2SYS the dissociation constants provided by Mehrbach et al. (1973) were used for carbonic acid (as refined by Dickson and Millero, 1987), and those provided by Dickson (1990) were used for boric acid. In addition, HOBO R data loggers (Onset R Computer Corporation) were used to measure light (in lux subsequently converted to PAR, see Long et al., 2012) in the back-reefs and seagrass meadows for a period of 25-months beginning in January 2012. Loggers were screwed onto a PVC mount installed at each site using a galvanized nail and marine epoxy. Loggers were cleaned and data was downloaded monthly in situ using a HOBO R Waterproof Shuttle (Onset R Computer Corporation). Fouling on the loggers was found to alter light quality after the first 10 days following cleaning, therefore only light data collected within these 10 day periods were analyzed (n = 250 days of light data for the full 25-month experiment).

### Benthic Surveys

Benthic surveys were conducted in February 2012 and February 2013. To characterize the benthos of the back-reef and seagrass meadow, continuous-line intercept video transects were conducted using a high-definition video camera with an underwater housing (Canon R ). At each site and habitat type, three 50 m transects were laid parallel to shore. The start and end points of each transect were recorded using a handheld GPS (Garmin R ) and marked in situ using semi-permanent metal pickets. Benthic composition was quantified using categories defined by Atlantic and Gulf Rapid Reef Assessment protocol version 5.4 (Lang et al., 2010), and scleractinian corals were identified to species level.

The size-frequency, density, and health of established scleractinian coral populations were also assessed within each habitat. Established coral populations refer to all coral colonies identified on natural substratum within each habitat (e.g., not recruited on to settlement tiles, see proceeding section). A 20 m2 (5 × 4 m) quadrat was placed at the start point of each transect, and species, size, and condition of each coral (≥1 cm) was recorded. Colonies were measured using calipers or measuring tape. Colony condition characteristics recorded included depigmentation and evidence of new, transitional, or old mortality (as defined by Lang et al., 2010).

#### Recruitment

Within this study, recruited coral populations refer to corals recruited onto settlement arrays. Settlement arrays were deployed in January 2012 to assess scleractinian coral recruitment patterns within each habitat type. Pre-conditioned limestone settlement tiles were used, as they have demonstrated coral recruitment rates comparable to natural substrate (e.g., Salinas de León et al., 2011). Arrays consisted of six unglazed limestone tiles (10 × 10 × 1 cm) with 0.8 cm holes drilled in the center (modified from Mundy, 2000; Salinas de León et al., 2011). Each tile was labeled for identification and attached to a PVC base with a plastic cable tie. Tiles were conditioned for 1 month in their respective target habitats in Grape Tree Bay prior to installation. Arrays were installed using galvanized nails through either end of the PVC base. The PVC base created a 2.5 cm gap between the tiles and the substrate, which has been shown to facilitate coral settlement (Harriott and Fisk, 1987). Three settlement arrays were deployed at every sampling replicate at each of the three study sites (n = 18 tile arrays per site, n = 54 arrays (324 tiles) total). Arrays were spaced a minimum of 5 m apart (Salinas de León et al., 2011).

Tiles two, three, and six (selected at random prior to deployment) were removed and replaced with new, conditioned tiles 1 month following installation and monthly thereafter. Tile removal allowed for the timing and frequency of recruitment to be assessed. Removed tiles were examined microscopically (20x) in the laboratory under blue light (Leica Microsystems). Tiles with coral recruits present were laid in a single layer in a bath treated with 10% aq chlorine solution for 48 h to remove organic matter and examined with the microscope (per Schmidt-Roach et al., 2008). Recruits belonging to the families Acroporidae, Agariciidae, Faviidae (now part of family Mussidae, see Budd et al., 2012), Poritidae, and Siderastreidae were identified using photomicrographs published previously (English et al., 1994; Babcock et al., 2003; de Putron, 2005, 2007). Recruits belonging to other families (for which photomicrographs were unavailable), or those that could not be identified, were recorded as other (per Schmidt-Roach et al., 2008). All tile arrays were removed after 25 months and analyzed following the same protocol as the monthly assessments.

#### Statistical Analysis

Mean and coefficient of variation (CV) of each abiotic factor (i.e. pHT, temperature, AT, arg, salinity, and water velocity) were compared between back-reefs and seagrass meadows and relative to the open-ocean sites using ANOVA and post hoc Tukey tests. Benthic structure of each habitat was compared between years using a Principal Component Analysis (PCA), with unpaired t-tests on the first and second principal components extracted for each site. Data were square-root transformed and clusters (80 and 90% similarity) were generated from a Bray-Curtis similarity matrix of all benthic data. The PCA vectors were generated by Pearson's Correlations exceeding R > 0.6, between plot ordinations and benthic categories. To characterize the established coral populations of each habitat, coral abundance, disease prevalence, species richness, diversity (Shannon-Wiener Index) and evenness (Shannon's Evenness) were calculated. The abundance of established corals per year, for each site, were compared by a 2 × 2 contingency table and a Pearson's Chisquared test. Size-frequency of the established coral communities of each habitat was binned into size categories; larger categories were pooled together to avoid the problems associated with low expected counts (Starnes et al., 2012). The maximum diameter for each colony was used for comparison to ensure standardization among colonies with different growth forms (per Suggett et al., 2012). Fits of exponential decay across the counts in each bin were performed for both 2012 and 2013 to assess the population size-classes (Santangelo et al., 2004). The log count in each bin was plotted against the middle size category, with fits only done on categories until the first zero bin was reached; thus, avoiding issues with non-defined logarithms. Population sizes between years, for each habitat, were also assessed initially by a 2 × 2 × 5 contingency table and then with Pearson's chi-squared test.

To analyze recruitment, all tile data was pooled and considered as a mean per site (see Hurlbert, 1984). The density of recruits was determined per family. For both the recruited coral population (determined from settlement tiles) and established coral populations (determined from the quadrat surveys), the proportional abundance of each species was determined for the back-reef and seagrass meadow habitats. Differences between the established and recruited coral populations were assessed through the z-test. Data was pooled from the two communities to obtain a pooled sample abundance of:

$$p\_{d\overline{f}\overline{f}} = \frac{p\_{\text{rec}}N\_{\text{rec}} + p\_{\text{est}}N\_{\text{est}}}{N\_{\text{rec}} + N\_{\text{est}}}$$

where prec is the sample proportion for the recruited population, Nrec is the sample size for the recruited population, pest is the sample proportion for the established population, and Nest is the sample size for the established population.

With a pooled standard error of:

$$SE\_{diff} = \sqrt{p\_{d\tilde{\mathcal{Y}}} (1 - p\_{d\tilde{\mathcal{Y}}}) \left(\frac{1}{N\_{\text{rec}}} + \frac{1}{N\_{\text{est}}}\right)}$$

Using test statistic z = pest−prec SEdiff , p-values were calculated for each family for the null hypothesis that the recruited and established coral population proportions were equal. At the end of the study, the recruited benthic taxa were determined by the CPCe analysis and compared between habitat types (n = 9 replicates per habitat type) using MANOVA with post-hoc Tukey tests. Statistical analysis was performed using R (R Core Team, 2014) and Graphpad Prism (version 7.01). Data were assessed and found to meet the assumptions of linearity, independence, homogeneity of variance, and normality.

### RESULTS

#### Physicochemical Conditions

Both the back-reef and seagrass meadow sites had significantly greater variability (CV) in abiotic conditions compared to the open-ocean sites (**Tables 1**, **2**). With the exception of salinity, all mean abiotic variables were also different between the seagrass meadows and back-reefs compared to the open-ocean reference (**Tables 1**, **2**). Specifically, mean pHT was elevated at both of the inshore habitats (back-reef = 8.17 ± 0.01 and seagrass meadow = 8.16 ± 0.01) relative to the open-ocean (8.12 ± 0.01),

#### TABLE 1 | Physicochemical data for the seagrass and back-reef habitats on Little Cayman, Cayman Islands, British West Indies.

Data from the open-ocean reference sites is also provided for comparison. The carbonate chemistry data shown is the mean (± SE) and range of n = 3 days per season and lunar cycle (discrete water samples: n = 432 for the back-reef and seagrass and n = 54 for the open-ocean). Light data was obtained from in situ HOBO® data loggers that were set to log at 1 min intervals over the study duration. Data shown is for the first ten days of each month (due to fouling, see Methods, with n = 24-33 days per season and lunar cycle).

irrespective of season or phase of the lunar cycle (**Tables 1**, **2**). The elevation in mean pHT and corresponding depletion in pCO2 was significant enough to elevate mean arg in the seagrass meadow and back-reef habitats relative to the open-ocean sites [F(2, 105) = 166.45, p < 0.001, **Tables 1**, **2**].

The seagrass meadows had greater variability in environmental conditions across the measured parameters than the back-reefs (see **Table 1**), with pHT the most variable condition measured. Specifically, the pHT range in the seagrass meadows over diurnal (0.51 ± 0.07), seasonal (0.62 ± 0.03) and lunar cycle (0.58 ± 0.04) time scales was ca. double that of the back-reefs (diurnal 0.22 ± 0.02, seasonal 0.29 ± 0.04, lunar 0.24 ± 0.05). Both habitats experienced a greater range of pHT conditions during the summer season and on spring tides (**Table 1**), with the seagrass meadows reaching a maximum pHT of 8.28 ± 0.02 around mid-day and a low of 7.80 ± 0.01 at night

(A) Results from analysis of variance (ANOVA) comparing daily mean values of abiotic variables among habitat types. Seasonal and lunar data were grouped by site to generate n = 36 daily means per habitat type. Post-hoc grouping letters denote differences among habitat types identified using Tukey's post-hoc test. (B) Results from ANOVA comparing the coefficient of variation (Cv) for abiotic variables among habitat types. Seasonal and lunar data were grouped by site to generate n = 36 values per habitat type. Post-hoc grouping letters denote differences among habitat types identified using Tukey's post-hoc test.

(**Figure 2**). Oxygen co-varied with pH, with maximum levels experienced in the day (7.11 ± 0.2) and lowest levels occurring at night (4.6 ± 0.1). Mean values of temperature, light, AT, salinity, and water velocity did not differ between the back-reefs and seagrass meadows (**Table 1**).

#### Benthic Structure

In the seagrass meadows, benthic cover was primarily photoautotrophs (seagrass and non-calcifying algae), with the back-reefs predominantly comprised of abiotic-hard and abioticsoft ground (**Figure 3**). Across years there was little variation in the benthic composition of each site (see Supplementary Figure 1). Only the GTB back-reef site was different between 2012 and 2013 [two-tailed t-test, t(4) = 3.79, p < 0.05], resulting from changes in the composition of photoautotrophs and a 1.7 fold increase in scleractinian coral cover in 2013.

The back-reefs had ca. twice the mean coral cover of the seagrass meadows (back-reef: 6.5 ± 1.2%, seagrass: 2.7 ± 0.4%) and coral community composition within each habitat did not differ between years (**Table 3**). Coral species richness, diversity, and evenness were also higher in the back-reefs compared to the seagrass meadows in both years (**Table 3**). Although coral cover was higher in the back-reefs, there was a greater density of corals in the seagrass meadows than in the back-reefs (back-reef 119.5 ± 12.8, seagrass 236 ± 95.6 per 20 m2 ). No dependence between year and habitat type was detected for absolute counts of corals. Coral colonies were relatively small in both habitats, however faster exponential decay rates suggests that seagrass communities were dominated by higher numbers of small coral colonies, while counts were more even across size classes in the backreefs (**Figure 4**). Size profile of colonies did not differ significantly across the two years of study. Coral disease prevalence was low across habitats for both years (<1%). However, the prevalence of old mortality was higher on the back-reefs compared to the seagrass meadows [back-reef: 14 ± 0.07%, seagrass: 2 ± 0.02%; F(2, 36) = 253.46, p < 0.001].

Seagrass coral communities were comprised almost entirely of Poritidae (49.0 ± 0.02%) and Siderastreidae (49.8 ± 0.04%), while these families comprised 65.0 ± 0.03% of total coral cover on the back-reefs (followed by Faviidae which contributed 11.4 ± 0.01%). The dominant coral species in the back-reefs for both years was Porites astreoides (2012: 32 ± 0.7%, 2013: 36 ± 0.4%), while Siderastrea radians was the dominant species in the seagrass meadows both years (2012: 49 ± 0.3%, 2013: 52 ± 0.5%). Supplementary Table 1 provides a full list of coral species observed in each habitat across both years of study.

#### Recruitment

Coral recruitment onto settlement tiles was only observed during the months of August and September (in both 2012 and 2013). More coral spats recruited to the tile arrays in the back-reefs compared to the seagrass meadows [**Figure 5A**; t(17) = 8.84, p < 0.001]. A greater diversity of corals also recruited to the back-reef compared to the seagrass meadows (**Figure 5**). Poritidae and Siderastreidae were the most common families observed on the tiles from both habitat types (**Figure 5**). In general, the proportional density of coral species in the

recruited populations mirrored the proportional density in the established populations (**Figure 5**). Acroporidae and Agariciidae were absent from both the recruited and established population in the seagrass. However, Faviidae was absent from only the established population in the seagrass (**Figure 5C**). In the backreefs, the proportional density of Acroporidae, Agariciidae and "other" recruits were significantly higher compared to that of the established population (p < 0.01). Also, on the back-reef the proportional density of Faviidae recruits was significantly lower compared to that of the established population (p < 0.01). Both the back-reefs and seagrass meadows had more corals that belonged to families other than Acroporidae, Agariciidae, Poritidae, Siderastreidae and Faviidae (and thus fell into the "other" category) in the recruited population compared to the established population (p < 0.001).

Analysis of the composition of all benthic taxa on the experimental tiles following the full 25-month study revealed significant differences between the back-reefs and seagrass meadows [F(4, 13) = 200.73, p < 0.01; Supplementary Figure 2]. Notably, the percent cover of recruited corals was less than 1% in both habitats. Also, cover of calcareous Rhodophyta (i.e., crustose coralline algae) was significantly higher on tiles in the back-reefs compared to the seagrass meadows (post-hoc Tukey test: p < 0.05).

#### DISCUSSION

Seagrasses are recognized as important primary producers (Duarte and Chiscano, 1999) and for their roles in sediment stabilization and carbon sequestration (Duarte et al., 2013), but the role of seagrass meadows as a habitat in which scleractinian corals can recruit, grow, and reproduce has not been well-studied. We observed up to 14 established scleractinian taxa living

FIGURE 3 | Benthic composition of back-reef and seagrass sites. Principal co-ordinate analysis (PCA) plot (A) of the benthic categories pooled across both 2012 and 2013 (using all species and abiotic categories) per sites with 80% (solid) and 90% (hashed) similarities shown. Pearson's correlations that exceed R > 0.6 are represented as vectors in blue. The proportional composition of each benthic category pooled across years is shown in (B). Note: invertebrates other than the Cnidarians are not visible due to their low density.

within seagrass meadows, despite considerable physicochemical variability. Our data also demonstrate that the larvae of multiple coral families are able to recruit into seagrass meadows. These findings are consistent with previous reports of established coral communities among seagrasses (Rützler and Macintyre, 1982; Johnson, 1992; Higuchi et al., 2014).

#### Physicochemical Conditions

Seagrass beds were characterized by higher variability in both temperature and pH compared to the adjacent back-reef. This is

TABLE 3 | Summary of established coral community characteristics in the back-reef and seagrass habitats of Little Cayman, Cayman Islands.

consistent with previous studies of physicochemical conditions at this study site (Barry et al., 2013; Camp et al., 2016a,b). High physicochemical variability has also been reported for shallow seagrass beds globally; a diel temperature range of as much as 6 ◦C has been reported for shallow seagrass beds (Semesi et al., 2009), consistent with the range of up to 6.7◦C recorded in the present study. Similarly, mean diel fluctuations of 0.34 (Saderne et al., 2013) to 0.35 pH units (Challener et al., 2016) have been reported previously from studies of macrophyte beds. One study reported diel fluctuations in excess of 1 pH unit (Semesi et al., 2009). The pH range recorded in Little Cayman seagrass beds (0.51–0.62 pH units) is therefore consistent with previous reports, and is similar to the pH range reported from a study of Grape Tree Bay in 2011 (0.47 pH units, Barry et al., 2013). These substantial fluctuations in pH result from changes in pCO2 driven by seagrass photosynthesis and respiration (Semesi et al., 2009; Saderne et al., 2013; Challener et al., 2016). Covariance of DO with pH is also consistent with previous reports (Barry et al., 2013), and supports the idea that seawater chemistry is altered by seagrasses (Challener et al., 2016). Although physicochemical conditions fluctuated less on the back-reef, variability was still high compared to the open ocean reference site. This is likely due to the effects of the adjacent seagrass meadow; seagrasses have been shown to alter carbonate chemistry of seawater on adjacent reefs (Manzello et al., 2012).

Living under more variable physicochemical conditions can place an enhanced energetic burden on physiological maintenance (Hofmann and Todgham, 2010). Although extreme temperatures are known to be detrimental to coral populations (Eakin et al., 2009), some populations have shown capability for adapting or acclimating to temperature extremes (Craig et al., 2001; Palumbi et al., 2014; Schoepf et al., 2015). Thermal tolerance is also known to vary among coral taxa (Darling et al., 2012), and this variability could therefore explain why some taxa were only observed in the less-variable back-reef in the present study. Similarly, although some sites have been shown to maintain high coral cover and diversity in low-pH conditions (e.g. Golbuu et al., 2016), reduced pH has been shown to limit the abundance (Price et al., 2012) and diversity (Fabricius et al., 2011; Crook et al., 2012) of calcifying organisms at other locations. In the present study, pH at the seagrass sites routinely dropped to values considered unfavorable for most corals (e.g., <7.8, **Figure 2**; see also Camp et al., 2016a), and therefore may have resulted in environmental filtering against some coral species (Sommer et al., 2014). The ability to tolerate reduced pH varies among coral species (McCulloch et al., 2012), and importantly, low pH appears to inhibit settlement of Acropora palmata,

FIGURE 4 | A comparison of coral colony sizes in back-reef and seagrass habitats. Exponential decay model of coral colony size vs. the natural log (Ln) of the frequency of corals within each size class for 2012 (solid lines) and 2013 (hashed lines) in back-reef (circles) and seagrass (squares) habitats. Data are pooled across study sites (n = 3) and 95% confidence bands of each model are shown by areas colored corresponding to symbol colors.

potentially through alteration of larval settlement processes or chemical settlement cues from crustose coralline algae (Albright et al., 2010). Together these factors governing sensitivity provide one possible explanation for the lack of Acroporidae spats either on settlement tiles or in the established coral community at the seagrass sites despite the presence of established and recruited Acropora spp. in the adjacent back-reef. It is however critical to note that the ability of seagrasses to alter seawater carbonate chemistry is dependent on the seagrass species present, water depth, water residence time, and a variety of other local conditions (Unsworth et al., 2012). It is therefore possible that seagrass meadows in other locations could harbor a higher diversity of coral species.

#### Benthic Structure

Previously published data on coral community composition within seagrass meadows is limited, however our observations of Manicina areolata (Johnson, 1992), Porites furcata (Wulff, 2008), and S. radians (Rützler and Macintyre, 1982) populating seagrass meadows are consistent with previous Caribbean studies. In contrast to our observations, Rützler and Macintyre (1982)report occasional Acropora cervicornis colonies within seagrass patches in Belize; however this data was published prior to the widespread

regional decline of Caribbean acroporids (Aronson and Precht, 2001), which could partially explain its absence within seagrass meadows in our study. Low mean coral cover within seagrass meadows in the present study (2.7 ± 0.4%) is also generally consistent with, although somewhat lower compared to, reported coral cover in Indian Ocean (4.20% coral cover) and Pacific Ocean (5.40% coral cover) seagrass meadows (Camp et al., 2016a).

Although size data for corals in seagrass meadows are also rare, previous studies suggest that coral colonies within seagrass in the Caribbean are generally small (Johnson, 1992; Wulff, 2008), consistent with our findings of relatively small colonies in Little Cayman seagrass meadows. Small colony size is characteristic of species that are better able to survive poor conditions (Darling et al., 2012) that may occur frequently in shallow seagrass meadows, including sedimentation, thermal stress, and reduced pH. Small size is also characteristic of the dominant coral species in the seagrass meadow, S. radians (Lirman and Manzello, 2009). Decreased colony size has also been directly associated with low pH conditions (Crook et al., 2012), which regularly occurred in the seagrass meadows studied.

Although the seagrass meadows and back-reefs assessed in this study were in close proximity to one another, species richness and diversity of the established coral community were lower in seagrass meadows. Given that availability of hard substrate is fundamental in structuring benthic communities (Jackson, 1977; Taylor and Wilson, 2003), higher availability of hard substrate in the back-reef is likely a significant factor in the observed higher cover and diversity of corals. Hard substrate is characteristic of reefs, as reef structure is formed by deposition of calcium carbonate skeleton by living corals (Hallock, 2015). In contrast, seagrasses generally require soft sediment, and therefore seagrass meadows typically lack abundant hard substrate (Hemminga and Duarte, 2000). Many of the common coral species in the seagrass meadows studied (i.e., S. radians, Porites divaricata, P. porites) are capable of recruiting to suboptimal habitats (Darling et al., 2012). S. radians in particular has been shown to be highly tolerant of sediment burial (Lirman and Manzello, 2009), which is likely frequent in the soft-sediment seagrass meadow.

#### Recruitment

Recruitment onto experimental tiles during the 25-month study was only observed during the months of August and September, which is consistent with observations of multiple species spawning in the Cayman Islands during these months in 2012 and 2013 (C. McCoy, Cayman Islands Department of Environment, personal communication), and with published reports for spawning times among Caribbean broadcastspawning corals (e.g., Jul-Sept, Szmant, 1986). However, our results contrast with previous studies that have documented yearround planulation in Caribbean brooding species with varying species-specific annual peaks in larval release (Szmant, 1986; Chornesky and Peters, 1987; Soong, 1991; Schlöder and Guzman, 2008). Recently reported coral recruit densities vary among sites within both the Caribbean [e.g., 0.14–1.17 spat m−2 (Edmunds et al., 2014), 30 ± 46 – 236 ± 143 spat m−2 (Humanes and Bastidas, 2015)] and the Indo-Pacific [e.g., 1.42 ± 0.33 – 1.72 ± 0.64 spat m−2 (Salinas de León et al., 2011), 54.74 spat m−2 (Bauman et al., 2015)]. Johnson (1992) reported recruitment ranging from <1 to 5 spat m−2 for M. areolata in a seagrass meadow. Recruit densities in our study (i.e., back-reef 1.2 ± 0.05 spat m−2 , seagrass: 0.5 ± 0.03 spat m−2 ) fall within the range reported from Caribbean sites in the published literature.

Factors governing coral recruitment are complex, and recruitment success can be temporally and spatially variable due to factors such as temporal variability in fecundity and physical disturbances (Fitzhardinge, 1985; Williams et al., 2008). The presence or absence of other species can also influence recruitment. For example, crustose coralline algae and microbial biofilms are known to provide cues that induce coral settlement (Heyward and Negri, 1999; Webster et al., 2004). Though our current data set provides new information regarding coral recruitment into seagrass meadows and adjacent back-reefs, it precludes fully understanding the complex nature of differences in recruitment between habitats.

### Comparison of Established and Recruited Communities

Although the diversity and density of coral recruits differed between the seagrass and back-reef, community composition of recruited corals largely mirrored that of established corals within sites. This contrasts with multiple studies of coral reefs, which have reported variability in the composition of established and recruited coral communities (e.g., Dunstan and Johnson, 1998; Bramanti and Edmunds, 2016). However, our findings are consistent with those of Johnson (1992), who found that recruitment rates for M. areolata in a seagrass bed related to local density of reproductive adult colonies. The established and recruited coral communities in seagrass meadows were comprised almost entirely of brooding species (**Figure 5**), which are believed to have a relatively short dispersal distance as a result of rapid settlement following release of planulae (Szmant, 1986; Harriott, 1992; Johnson, 1992; Carlon and Olson, 1993; Kenkel et al., 2013). Brooded planulae also typically receive zooxanthellae from their parent colonies in Caribbean species (Szmant, 1986; Thornhill et al., 2006), and planulae may therefore have a physiological advantage to settle locally, as symbionts and hosts are both adapted to local abiotic conditions (Kenkel et al., 2013). Short dispersal distance combined with local adaptation may therefore explain why the recruited population was highly similar to the established population in each habitat, and also why brooding Agariciidae were not found to disperse from the back-reef to the seagrass.

Despite overall similarity between the recruited and established community within habitats, differences were apparent for some coral families. These intra-habitat differences may also be explained by differences in reproductive strategy. Broadcast spawning corals are particularly susceptible to "Allee effects" and subsequent sexual recruitment failure, especially when densities of mature colonies are low (Knowlton, 2001). For example, Acroporidae are known to have low sexual reproductive success and recruitment in the Caribbean, and thus rely heavily on asexual fragmentation as a reproductive mechanism (Tunnicliffe, 1981; Richmond and Hunter, 1990; Williams et al., 2008). Reliance on asexual fragmentation could explain why relatively high cover of established Acroporidae was observed in the back-reef despite low levels of recruitment across both years of study. Asexual fragmentation is also used as a reproductive mechanism in some poritids, such as P. furcata (Highsmith, 1982). This could potentially explain the abundance of Poritidae in the established population in both the back-reefs and seagrass meadows, as fragmentation can play a role in establishing local dominance by certain species (Highsmith, 1982). Local abundance of Poritidae achieved through fragmentation could ultimately lead to higher rates of sexual reproduction and subsequent recruitment, which was also observed in both habitats. In contrast, lower relative cover of established Faviidae compared to the recruited population could suggest post-settlement mortality, which was found to be common for Orbicella faveolata (Szmant and Miller, 2006; Miller, 2014). Dispersal of coral larvae from distant reefs could also explain observed differences in community composition between established and recruited corals (Harriott and Fisk, 1988). We currently lack data to determine whether outer reef sites may have served as a larval source for recruits in either habitat, however recent work has demonstrated high levels of connectivity between reefs (Schill et al., 2015); it is therefore conceivable that larval exchange between the outer reef, back-reef, and seagrass meadow is possible.

### CONCLUSIONS AND FUTURE CONSIDERATIONS

Our study demonstrates that multiple coral species are able to live within and recruit to seagrass meadows, suggesting that corals are capable of using such systems as a habitat regardless of naturally high variability in environmental conditions relative to neighboring reefs. Fluctuating physicochemical conditions in seagrass meadows can expose corals to both periods of high temperature and periods of low pH, which could enhance tolerance to thermal and chemical stress, respectively (Camp et al., 2016a). Evidence suggests that transgenerational acclimatization is possible following exposure to extreme temperature and pH conditions (Putnam and Gates, 2015); however, exposure of the corals P. astreoides and A. palmata to environmental variance found within seagrass meadows resulted in no enhanced resistance to subsequent temperature and pH stress (Camp et al., 2016b). Furthermore, some seagrass meadows have been shown to buffer against ocean acidification by elevating mean pH in their vicinity, potentially improving the ability of downstream corals to calcify (Manzello et al., 2012). The potential biogeochemical-services seagrass meadows can provide under climate change remains unresolved. Given that seagrasses will be potentially selected for under future climates (Garrard and Beaumont, 2014), they may continue to provide habitat for some coral species. Clearly large gaps still remain in understanding the nature of coral populations in these habitats and the extent to which they can persist in a changing climate. Further studies are needed across other geographic locations to fully understand the extent to which seagrass beds may serve as coral habitats. Importantly however, the added role of seagrass systems in providing coral habitat emphasizes the critical importance in managing and conserving these sites.

#### AUTHOR CONTRIBUTIONS

All authors have agreed to be listed on and approve this manuscript. EC, DJSu, and DJSm conceived and designed the study; KL and EC collected the data; EC, AD, MN, and SW analyzed the data; MN prepared all figures; KL, EC, and DJSu wrote the manuscript; all authors discussed and critically edited the manuscript.

#### ACKNOWLEDGMENTS

The authors acknowledge C. Manfrino for contributions toward the funding, resources, and context necessary for the completion of this project. We would also like to thank IBEX Earth for their financial support. The authors are grateful for logistical support from the staff of the Little Cayman Research Centre,

#### REFERENCES

local volunteers, as well as technical staff from the University of Essex. Sincere thanks also go to the Cayman Islands Department of Environment for issuing two permits required to conduct this study. The permit for settlement tile installation and monitoring was issued to C. Manfrino and K. Foster, and an additional permit for ongoing monitoring and abiotic sampling was issued to E. Camp. This project was part of the Central Caribbean Marine Institute's Climate Change and Coral Reef Stress Program and represents contribution 2017:04.

#### SUPPLEMENTARY MATERIAL

The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fmars. 2017.00388/full#supplementary-material

between populations of the mustard hill coral (Porites astreoides) from different reef environments. Mol. Ecol. 22, 4335–4348. doi: 10.1111/mec.12391

hermatypic coral Montastraea faveolata and Acropora palmata in the Florida keys, USA," in Proceedings of the 10th International Coral Reef Symposium (Okinawa).

**Conflict of Interest Statement:** The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Copyright © 2017 Lohr, Smith, Suggett, Nitschke, Dumbrell, Woodcock and Camp. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY). The use, distribution or reproduction in other forums is permitted, provided the original author(s) or licensor are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Physiological and Biogeochemical Responses of Super-Corals to Thermal Stress from the Northern Gulf of Aqaba, Red Sea

#### Andréa G. Grottoli 1 \*, Dan Tchernov 2 and Gidon Winters 3

1 School of Earth Sciences, Ohio State University, Columbus, OH, United States, 2 Leon H Charney School of Marine Science, University of Haifa, Haifa, Israel, 3 The Dead-Sea and Arava Science Center (ADSSC), Tamar Regional Council, Neve Zohar, Israel

#### Edited by:

Emma Camp, University of Technology, Sydney, Australia

#### Reviewed by:

Anastazia T. Banaszak, National Autonomous University of Mexico, Mexico Matthew Ryan Nitschke, University of Technology, Sydney, Australia

> \*Correspondence: Andréa G. Grottoli<EMAIL_ADDRESS>
#### Specialty section:

This article was submitted to Coral Reef Research, a section of the journal Frontiers in Marine Science

Received: 26 April 2017 Accepted: 19 June 2017 Published: 04 July 2017

#### Citation:

Grottoli AG, Tchernov D and Winters G (2017) Physiological and Biogeochemical Responses of Super-Corals to Thermal Stress from the Northern Gulf of Aqaba, Red Sea. Front. Mar. Sci. 4:215. doi: 10.3389/fmars.2017.00215 Mass coral bleaching is increasing in frequency and severity, leading to the loss of coral abundance and diversity. However, some corals are less susceptible to bleaching than others and can provide a model for identifying the physiological and biogeochemical traits that underlie coral resilience to thermal stress. Corals from Eilat in the Gulf of Aqaba in the northern Red Sea do not bleach unless seawater temperatures are sustained at +6 ◦C or higher above their average summer maximum. This extreme thermal tolerance qualifies these as super-corals, as most corals bleach when exposed to temperatures that are only +1–2◦C above their thermal maximum. Here, we conducted a controlled bleaching experiment (+6 ◦C) for 37 days (equivalent to 32◦ heating weeks) on three species of corals from Eilat: Stylophora pistillata, Pocillopora damicornis, and Favia favus. To assess the response of the holobiont to thermal stress, the following variables were measured on each coral: endosymbiotic algal cell density, Chlorophyll a, endosymbiotic mitotic cell division, total lipids, protein, carbohydrate, and the stable carbon (δ 13C) and oxygen (δ 18O) isotopic composition of the skeleton and the δ 13C of the animal host tissue and endosymbiotic algae. While all three species appeared visibly bleached, their physiological and biogeochemical responses were species-specific. S. pistillata catabolized lipids but still maintained total energy reserves and biomass. Increases in both skeletal δ 13C and δ 18O indicates that calcification declined in this species. P. damicornis was the least affected by bleaching. It maintained its total energy reserves and biomass, and isotopic evidence suggests that it maintained calcification and was not dependent on heterotrophy for meeting metabolic demand when bleached. Finally, F. favus catabolized protein and carbohydrates, and suffered losses in total energy reserves and biomass. Nevertheless, isotopic evidence suggest that photosynthesis and calcification were maintained, and that this species has a high baseline heterotrophic capacity. Thus, just like their non-super-coral conspecific counterparts, maintaining energy reserves and biomass, and heterotrophic capacity appear to be traits that underlie the thermal tolerance of these super-corals from Eilat. Given the high thermal tolerance of these super-corals, these populations could provide viable seed stock for repopulating coral losses on other reefs.

Keywords: coral bleaching, heterotrophy, energy reserves, isotopes, resilience, physiology, biogeochemistry, Red Sea

#### INTRODUCTION

Increased atmospheric carbon dioxide (CO2) is causing the oceans to warm and to become more acidic. Over the course of this century, tropical seawater temperatures are expected to rise by 1–3◦C coupled with a drop in pH of 0.2–0.4 (IPCC, 2013). At this rate, tropical reefs are expected to experience significant declines in coral abundance, coral diversity, and reef growth during the second half of this century (e.g., Hoegh-Guldberg et al., 2007; Eakin et al., 2009; Veron et al., 2009; Hoegh-Guldberg, 2011). While coral calcification is projected to decline by 5% by 2034, severe bleaching will be annual by 2030 in some regions and globally by 2055 (van Hooidonk et al., 2014). Recent studies have shown that the warming of tropical oceans is a much more immediate threat to the survival of coral reefs than is ocean acidification (e.g., Cantin et al., 2010; Frieler et al., 2012; Chua et al., 2013). Corals can survive slight decreases in calcification, but it is not likely that they can physiologically survive severe bleaching on an annual basis.

Elevated temperatures can cause corals to bleach: a physiological response to sustained elevated temperatures that results in a breakdown of the coral symbiosis, the loss of the coral's algal endosymbiont partners, which results in a white appearance. The consequences of bleaching to the coral are reduced coral growth, disease, reduced fertility, and death (e.g., Brown, 1997; Hoegh-Guldberg, 1999; Omori et al., 1999; Buddemeier et al., 2004; Tchernov et al., 2011). The frequency of bleaching events appears to be increasing (Eakin et al., 2009; Heron et al., 2016) and back-to-back bleaching years have been observed on basin-wide scales (coralreefwatch.noaa.gov). Annual bleaching is predicted to occur by 2030 in some regions (van Hooidonk et al., 2013, 2014), resulting in large coral losses.

Interestingly, there are corals that are less susceptible to and/or recover more quickly from bleaching (e.g., Loya et al., 2001; Grottoli et al., 2006, 2014; Fine et al., 2013). Resilience to bleaching has been associated with high energy reserves, heterotrophic plasticity, and endosymbiont type (Rowan et al., 1997; Baker et al., 2004; Tchernov et al., 2004; Berkelmans and van Oppen, 2006; Grottoli et al., 2006, 2014; Rodrigues and Grottoli, 2007; Anthony et al., 2009; Guest et al., 2012). Corals with elevated levels of energy reserves (i.e., lipids, protein, carbohydrates) either take longer to bleach, bleach less severely, and/or recover more quickly from bleaching than corals with lower energy reserves because they have an energy source to buffer them against losses in photosynthetically derived fixed carbon (e.g., Rodrigues and Grottoli, 2007; Anthony et al., 2009; Grottoli et al., 2014; Schoepf et al., 2015). Corals that increase their intake of heterotrophic carbon (i.e., zooplankton, dissolved and particulate organic carbon) or increase the proportionate contribution of heterotrophic vs. photoautotrophic carbon in their tissues when bleached, are able to partially or fully supplement the deficit in their carbon budgets due to declines in photosynthesis when bleached, and are able to recover more quickly from bleaching (Grottoli et al., 2006, 2014; Rodrigues and Grottoli, 2006; Palardy et al., 2008; Levas et al., 2013, 2016). Corals that are able to shuffle or switch their dominant endosymbiont type for a thermally tolerant one are less sensitive to repeat exposures to thermal stress (e.g., Rowan et al., 1997; Baker et al., 2004; Berkelmans and van Oppen, 2006; Grottoli et al., 2014). In addition, there are regions such as the northern region of the Gulf of Aqaba, where entire communities of corals can already tolerate extreme warming conditions, thought the underlying traits that produce that extreme tolerance is poorly understood. These corals may provide seed stock for restoring dying reefs as well as be a model system for identifying the physiological and biogeochemical traits that underlie coral resilience to thermal stress.

Corals in the northern tip of the Gulf of Aqaba in the Red Sea do not bleach when exposed to temperatures that are ≥6 ◦C above their average summer time maximum temperatures for 3–4 weeks (Fine et al., 2013) coupled with extreme salinities of 41ppt (Paldor and Anati, 1979). This is remarkable resilience to extreme increases in seawater temperature given that most corals bleach when seawater temperatures exceed summertime maximum temperature by as little as 1–2◦C for as little as 2–4 weeks (e.g., Hoegh-Guldberg, 2011). Fine et al. (2013) proposed that this extreme thermal tolerance in these northern Red Sea corals is due to natural selection processes on the larvae transported to the northern section of the Gulf of Aqaba through the very hot straights of Bab el Mandeb in the southern Red Sea. However, the underlying physiological traits responsible for the resilience of these corals is not fully understood. We hypothesized that like their less thermally tolerant counterparts, maintenance of energy reserves and biomass, as well as heterotrophic plasticity, are all traits responsible for extreme temperature tolerance in super-corals of the northern Gulf of Aqaba. To test this hypothesis, bleaching was induced by thermal stress in an experiment conducted on three species of corals (Stylophora pistillata, Pocillopora damicornis, and Favia favus) from the Gulf of Aqaba near Eilat, Israel and the physiological and biogeochemical responses of the host and algal endosymbiont were assessed.

### MATERIALS AND METHODS

Two fragments from five healthy colonies of S. pistillata and P. damicornis (5–7 cm tall), and two fragments for seven healthy colonies of F. favus (6–8 cm in diameter) were collected between 2 m and 4 m depth in the Gulf of Eilat (northern Gulf of Aqaba), Israel (29◦ 30′ 5.55′′ N, 34◦ 55′ 2.49′′ E) in mid-February 2005 and glued to labeled gray PVC tiles using underwater epoxy putty. All three species are common in this region.  P. damicornis is also a branching coral, is a member of a genus that reproduces by both broadcasting and brooding (though spawning has never been documented for this species in the Red Sea), and is dominated by the endosymbiont Symbiodinium clade C (Shlesinger and Loya, 1985; Karako-Lampert et al., 2004; Fine et al., 2013; Sawall et al., 2015; Ziegler et al., 2017). F. favus is a mounding coral, a broadcast spawner, and dominated by the endosymbiont Symbiodinium clade C (Shlesinger and Loya, 1985; Karako-Lampert et al., 2004; Fine et al., 2013). Symbiodinium clade D has never been reported for healthy or bleached corals in this region (Fine et al., 2013).

One fragment from each colony was placed in each of two 95 liter outdoor flow-through seawater tanks at the Interuniversity Institute for Marine Science of Eilat, Israel at the northern tip of the Gulf of Aqaba, Red Sea, and allowed to acclimate to the tanks and recover for 2 weeks. The tanks were shaded with black shading nets to simulate photosynthetically active radiation levels at collection depth, and the seawater was filtered through a small coarse gravel filter (70–200 µm pore size), allowing only small organic particulates and dissolved organics into the tanks. Thus, heterotrophic food sources available to the corals were limited to dissolved and particulate organic matter that came through the tank system. Coral were not supplemented with zooplankton. Flow rates were maintained in each tank at 2 l/min allowing each tank to fully flush every 47 min. On 28 February 2005, the seawater temperature was gradually increased over 3days using three 300W ATMAN aquarium heaters per tank yielding an average daily temperature of 32◦C, while the second tank remained at ambient seawater temperature as controls with an average temperature of 26◦C. The rate of temperature increase in the 32◦C tank was chosen to minimize heat shock and also get the tank up to temperature relatively quickly. Since these corals are very tolerant of elevated temperatures, the rate of increase was not detrimental to these corals. The original temperature logger data has been lost, but the typical standard deviation of average daily temperatures in experimental tanks at the institute are ± 0.5◦C. On 6 April 2005 (37 days later), the fragments were collected, frozen at −80◦C, and shipped frozen to the Grottoli laboratory for analyses. In the lab, a quarter of each fragment was used for the following analyses: (1) chlorophyll a (Chl a) concentration, (2) total lipid, soluble animal protein, and soluble animal carbohydrate concentrations, (3) endosymbiotic algae concentrations and mitotic indexes, and (4) stable isotopic analyses of the host tissue, endosymbiotic algae, and underlying skeleton.

## Algal Endosymbiont Analyses and Coral Energy Reserves

Chl a was extracted from ground samples in 100% acetone and the absorbance measured at 663 and 630 nm wavelengths using a Spectronic 20 Genesis visible spectrophotometer. The Chl a concentration was calculated using equations from Jeffrey and Humphrey (1975) corrected for turbidity with the absorbance value measured at 750 nm. Algal endosymbionts were separated from the animal tissue by centrifugation and three subsamples were counted with a hemocytometer and light microscope. The total number of algal cells and the number of mitotically dividing algal cells were recorded.

Total lipids were extracted in a 2:1 chloroform/methanol solution, washed in 0.88% KCl, and then dried to a constant weight (method modified from Grottoli et al., 2004). Soluble animal protein was extracted using the bicinchoninic acid method (Smith et al., 1985) with bovine serum albumin as a standard (Pierce BCA Protein Assay Kit). Soluble animal carbohydrate was extracted using the phenol-sulfuric acid method (Dubois et al., 1956) with glucose as a standard. All analyses were made from whole ground coral samples, converted into Joules (Gnaiger and Bitterlich, 1984), and standardized to both grams of ash free dry weight (gdw) and to surface area as measured using the aluminum foil technique (Marsh, 1970). However, normalization to ash free dry weight is more robust when comparing species with different tissue depths, morphology, and polyp structure than the same data normalized to surface area (Edmunds and Gates, 2002). Therefore, only results of data standardized to ash free dry weight are reported here (However, note that all statistics were also performed on the area normalized data and yielded almost identical patterns). Total energy reserves were calculated as the sum of total lipids, proteins, and carbohydrates. Total dry biomass was standardized to surface area. These analytical methods have been successfully used in corals before (e.g., Rodrigues and Grottoli, 2007).

#### Stable Isotopic Analyses

Coral tissue was completely removed from the skeleton using a Water-pik (Johannes and Wiebe, 1970). The tissue slurry was separated into animal host and algal endosymbiont fractions by centrifugation using established methods (Rodrigues and Grottoli, 2006). Briefly, coral slurry was homogenized to break open the animal cells, then centrifuged at 4,000 rpm for 5 min to separate the animal host and endosymbiotic algal fraction. The supernatant (i.e., animal host) was decanted and the algal pellet resuspended in deionized water and centrifuged again to remove any remaining animal host particles. The supernatants from both centrifugations were combined. Each organic fraction was isolated onto a pre-burned glass fiber filter under vacuum, and individually combusted in a Costech Elemental Analyzer. The resulting CO2 gas was analyzed for δ 13C [δ 13C = permil deviation of 13C:12C relative to Vienna Peedee Belemnite Limestone standard (VPDB)] with a Thermo Delta Advantage IV via a Thermo ConFlow III open split in the Grottoli lab. At least 10% of all samples were run in duplicate. The standard deviation of repeated measurements of the USGS24 standard for δ 13C was ±0.08 h. The host tissue isotope values were reported as δ 13Ch while the algal endosymbiont isotope data were reported as δ 13Ce. In general, as the incorporation of isotopically depleted zooplankton and other heterotrophic carbon sources (i.e., dissolved and particulate organic carbon) into coral tissues increases, the δ 13Ch decreases (Rodrigues and Grottoli, 2006; Levas et al., 2013; Schoepf et al., 2015). As the rate of photosynthesis increases, isotopic fractionation decreases, and endosymbiotic algae incorporate relatively less 12C than 13C into their tissues, resulting in increased δ 13Ce (Muscatine et al., 1989; Rodrigues and Grottoli, 2006). The difference between the δ 13Ch and δ 13Ce were also calculated to assess the proportionate contribution of heterotrophy and photoautotrophy in the carbon budget (sensu Muscatine et al., 1989). Specifically, decreases in δ 13Ch–δ 13Ce are interpreted to mean that the proportionate contribution of heterotrophic carbon to coral tissues has increased.

The skeleton was prepared for δ 13C (δ 13Cs) and δ 18O (δ 18Os)[δ 18O = permil deviation of 18O:16O relative to VPDB] analysis by shaving the top 100–200 µm of skeletal material off the tip of each coral fragment using a Dremel tool with a diamond-tipped drill bit. Approximately ∼80 µg of skeletal material was acidified with 100% ortho-phosphoric acid in an automated Kiel III carbonate device, and the δ 13C and δ 18O values of the resulting CO2 were measured using a Finnigan MAT 252 mass spectrophotometer at Stanford University. At least 10% of all samples were run in duplicate. The standard deviation of repeated measurements of an internal standard was ±0.03 h for δ 13C and ± 0.05 h for δ 18O. Skeletal values were reported as δ 13Cs and δ 18Os . In general, as the rate of photosynthesis increases or as the incorporation of heterotrophic carbon into coral tissues decreases, the pool of inorganic C available for calcification through animal respiration becomes enriched resulting in increased δ 13Cs (Grottoli and Wellington, 1999; Grottoli, 2002). In addition, variability in δ 18Os results from temperature-induced kinetic fractionation and seawater δ 18O (Epstein et al., 1953). As temperature increases, δ 18Os decreases (Weber and Woodhead, 1972; Kim and O'Neil, 1997). Since seawater δ 18Os is constant in this study because the source seawater was the same in all tanks, it did not influence the δ 18Os values.

#### Statistical Analysis

Univariate two-way analysis of variance (ANOVA) was used to test the effects of species and temperature on Chl a, algal endosymbiont, mitotic index, lipid, protein, carbohydrate, total energy reserves, and tissue biomass concentrations as well as δ 13Cs , δ 18Os , δ 13Ch, δ 13Ce, and δ 13Ch–δ 13Ce. Prior to statistical analysis, all data was tested for normality using a Shapiro-Wilk's test and homogeneity of variance was assessed with plots of expected vs. residual values. Any data failing to meet this assumption was log transformed and then met the assumptions of normal distribution. A posteriori slice tests (i.e., tests of simple effects, Winer, 1971) determined if bleached and control averages significantly differed within species. Bonferroni corrections were not used due to increased likelihood of false negatives (Quinn and Keough, 2002). With only one tank for treatment and one tank for controls, we were unable to assess tank effects independent of treatment effects. However, every aspect of both tanks that were measured were the same (i.e., seawater source, flow, light, position relative to the sun's arc) except for temperature. In addition, one bleached P. damicornis and two control F. favus fragments were lost during shipping or storage, yielding an unbalanced number of samples in the treatment and controls for these two species. Statistical analyses were generated using SAS software, Version 8.02 of the SAS System for Windows. Values of p < 0.05 were considered significant.

To get an overall assessment of the effect of treatment across all variables, non-parametric multi-variate statistics were employed. The non-parametric analyses were conducted on three sub-sets of the data: (1) all of the physiology data, (2) all of the isotopic data, and (3) all of the physiology and isotopic data combined. Since two samples did not have isotopic data, the multi-variate analyses that included the isotopes were comprised of 29 corals whereas the analyses of the physiological variables alone included 31 corals. Euclidean-distance based resemblance matrix was constructed using normalized data of all of the measured variables. Non-parametric multidimensional scaling (NMDS) plots were generated to graphically represent relationships between each coral sample across all treatments and species. Analysis of similarities (ANOSIM) was then used to test for the effect of species and treatment. ANOSIM simultaneously uses all variables measured, and tests whether all data (i.e., all physiological and/or isotopic) from one coral group differs from another coral group. The ANOSIM pairwise test statistic R ranges from 0 (no difference and complete overlap between groups) to 1 (maximum difference and no overlap between groups) and is a strong indicator of separation between groups, especially at low sample sizes such as in this study, as long as the overall ANOSIM test statistic R is significant (Clarke and Gorley, 2006). Multivariate analyses were conducted using the software package Primer V6, and for the ANOSIMs a p < 0.05 was considered significant. Finally, since fragments pairs of each colony were included in both treatments, differences between treatment and control for any variable were due to treatment alone, independent of genotype.

### RESULTS

The temperature exposure of the treatment corals was equivalent to 32◦ heating weeks (DHW) (coralreefwatch.noaa.gov) relative to the controls. All three species in the elevated temperature treatment (i.e., 32◦C) appeared pale or white and as such were bleached, whereas all of the corals in the control treatment (i.e., 26◦C) were brown in color and appeared non-bleached.

#### Endosymbiotic Algae

Average endosymbiotic algal concentrations dropped to 5, 49, and 19% of control concentrations in S. pistillata, P. damicornis, and F. favus, respectively (**Table 1**, **Figure 1A**). At the same time, Chl a concentrations in bleached S. pistillata and P. damicornis decreased to 37 and 33% of control values, respectively (**Table 1**, **Figure 1B**), and mitotic divisions were 26-fold lower in bleached S. pistillata and 9-fold lower in bleached F. favus corals relative to

TABLE 1 | Results of three two-way ANOVAs testing for the effect of species (S), and temperature (T) on average chlorophyll a concentration, endosymbiotic algal density and mitotic index.

Effects of species (S) and temperature (T) were fixed and fully crossed. Endosymbiontic algal concentration and mitotic index values were log transformed to meet assumptions of normal distribution before ANOVA analysis. df, degrees of freedom; SS, type III sum of squares of the main effects. Significant effects are in bold.