CALIFORNIA STATE UNIVERSITY, NORTHRIDGE
ANTHROPOGENIC STRESSORS AND THE IMPACTS ON SCLERACTINIAN AND
NON-SCLERACTINIAN TAXA
A thesis submitted in partial fulfillment of the requirements
for the degree of Master of Science in Biology
By
Ashley Elizabeth Potter
December 2018
This thesis of Ashley E. Potter is approved by:
Robert C. Carpenter, PhD. Date
Hollie Putnam, PhD. Date
Peter J. Edmunds, PhD., Chair Date
California State University, Northridge
ii ACKNOWLEDGEMENTS
First and foremost, I would like to thank my super awesome advisor, Dr. Pete Edmunds who has always encouraged me to a better scientist and a better person, through his own actions. Pete’s mentorship has given me the tools I need to prepare for challenges and opportunities. He has also invested a tremendous amount of time and energy in order to allow me to have opportunities that further my career as a scientist. He is remarkably dedicated to all his students and their futures and even though he is a rock star, he always makes time for us. I feel very proud to have done a master’s with Pete and will use the skills I have learned for the rest of my life. Thank you so much Pete.
I would like to thank my committee members, Dr. Robert Carpenter and Dr. Hollie Putnam putting up with me and providing me with their time and counsel. Their guidance was invaluable to my thesis work at CSUN and aided in making me a better scientist. Thank you to Hollie who helped support me as an undergraduate and really influenced my decision to join Pete’s lab.
I am so eternally grateful to have such amazing, inspiring fellow graduate students and faculty in the Biology department at CSUN who went out of their way to help me with my thesis work. First, I am so grateful to ALL the members of the polyp lab (past and present) that I interacted with during my time at CSUN. Specifically, I would like thank Steve Doo for being my zen rock out in the field, Jesse Bergman who I just find really cool, and Hannah Nelson for going out of her way to help me with statistics. I also want to acknowledge my new lab mates (Megan Williams and Jack Girard) who have been a huge support system while writing my thesis. Furthermore, I am thankful to Nyssa Silbiger, Casey terHorst, and Georgios Tsounis for working with me on my statistics and bearing with all of my questions. All of you have been tremendously supportive and encouraging and I am so lucky to have been chosen for this lab.
Finally, I would like to thank my family for their never-ending love and support through my masters. My sister (Shelby) whom I look up to and who inspires me to live my best life. My mom (Cookie) who always believed in me. My dad (Ray) who kept me grounded with my head on straight. My brother (Matt) who was always my super support system. And a very special and warm thank you to my partner in crime, Paco, who has endless understanding, support, and patience for me and my scientific pursuits.
This work was made possible by financial support from the National Science Foundation to the Mo’orea Coral Reef, Long-Term Ecological Research site (OCE 16-37396). The work was also partially funded by the CSUN Research and Graduate Studies Office, and CSUN Associated Student.
iii TABLE OF CONTENTS
Signature Page……………………………………………………………………….ii
Acknowledgements…………………………………………………………….…....iii
Abstract……………………………………………………………………………....v
Chapter 1: General Introduction…………………………………………………1
Chapter 2: Effects of natural disturbances non-scleractinian benthic invertebrates Introduction……………………………………………….….13 Methods…………………………………………………...….17 Results………………………………………………………..20 Discussion……………………………………………………22 Figures/Tables ………………………………………………..28
Chapter 3: Characterizing the Metabolic Gradient of Two Reef Corals: Understanding More than Just Highs and Lows Introduction………………………………………………… 35 Methods……………………………………………...……...38 Results ………………………………………………………47 Discussion…………………………………………………..49 Figures/Tables………………………………………………59
Chapter 4: Concluding Remarks………………………………………………..68
Literature Cited……………………………………………………….……………..72
iv ABSTRACT
Anthropogenic stressors and the impacts on scleractinian and
non-scleractinian taxa
By
Ashley E. Potter
Master of Science in Biology
Coral reefs are one of the most diverse ecosystems in the world and this diversity is largely due to the abundance and range of invertebrate taxa that live within the reef.
Stony corals and sponges are two benthic taxa that provide numerous benefits to coral reef communities that can be affected by anthropogenic disturbances. This thesis is comprised of two studies that address the ecological or physiological implications of anthropogenic disturbances on shallow coral reef communities. The first study took an ecological approach by addressing two morphologically distinct sponge species (Ircinia campana and Niphates digitalis) and their associated fauna, in St. John, US Virgin
Islands, following two category five hurricanes. N. digitalis was 52% more abundant
(individual m-2) on the reefs of St John then I. campana. Hurricanes effects did not significantly decrease density of either sponge, but they did decrease sponge size for both species. N. digitalis housed more invertebrate
fauna (Ophiothrix suensoni, and Pelia mutica) per unit length, then I. campana but this epifauna-host relationship did not appear to be obligate and changed over time.
The second study took a physiological approach by measuring, for the first time, the metabolic scope of two corals, Pocillopora verrucosa and Acropora pulchra, that are important components of the resilient reefs of Mo’orea, French Polynesia. Using a
v metabolic uncoupler to elicit the maximum metabolic rate (MMR), and starvation to elicit the basal metabolic rate (BMR), I described the gradient of metabolism and placed the metabolic rates associated with polyp expansion and digestion along this gradient. The metabolic scope ([MMR-BMR]/BMR) was ~136% for P. verrucosa and ~251% for A. pulchra, with metabolism increasing with expansion ~48-94%, and ~18-43% with digestion. I also explored size dependent metabolism and assessed if scaling differs based on organismic activity. The metabolic scaling exponents differed significantly among several treatments (b = 0.2 to 1) and differed from the exponents (b = 0.67 to 1) proposed by the metabolic level boundary (MLB) hypothesis. Variation in the scaling exponent was potentially caused by behavioral differences among the polyps. Together the results from Mo’orea, French Polynesia, and St. John US Virgin Islands, demonstrate the importance of studies that address the extent to which scleractinian and non-scleractinian taxa are being impacted by natural disturbances. They also highlight the need for ecological and physiological studies on coral reefs to determine how these stressors effect coral reef community structure and function.
vi CHAPTER 1
General Introduction
Anthropogenic Stressors on Coral Reefs
Coral reefs are one of the most diverse ecosystems in the world, and are often compared to their terrestrial equivalent, the tropical rain forest (Connell 1997). This high diversity is due largely to the abundance and types of invertebrates that live within coral reefs (Glynn and Enochs 2011; Romero et al. 2015). Scleractinians, or hard corals, are critical invertebrates on coral reefs because they are essential ecosystem engineers (Jones et al. 1994). This means that they provide resources to other species, such as complex three-dimensional habitats that make up the reef structure (Jones et al. 1994). Their framework provides coastal protection, tourism, fisheries, raw materials for biomedical research, habitat, feeding, spawning, and nurseries (Moberg and Folke 1999). Because of their ecological importance, scleractinians have garnered much research attention over the last few decades. Coral reef ecosystems are constantly being challenged by a suite of natural and anthropogenic stressors that include, but are not limited to, ocean acidification (Hoegh-Guldberg et al. 2007), eutrophication (Fabricius 2005), disease
(Birkeland 1997), rising sea surface temperature (Hoegh-Guldberg et al. 2007), and hurricanes (Harmelin-Vivien 1994), all of which can restructure coral reef communities.
As some of these disturbances continue to increase in intensity due to climate change
(Solomon et al. 2007; Knutson et al. 2010), it is critical to understand to what extent corals are being impacted by these events, both ecologically and physiologically, and to determine how this effects coral reef community structure and function.
1 Coral Reef Communities in Transition
The increase in intensity of stressors in coastal tropical seas may cause major or minor shifts in coral reef community structure depending on the severity, duration, and return time of the perturbations (Woodley et al. 1981; Gardner et al. 2005). In recent decades, there have been ecologically significant losses in coral cover on the Great
Barrier reef (De’ath et al 2012; Hughes et al. 2018), Florida Keys (Porter et al. 2001;), and large portions of the Caribbean (Gardner et al. 2003; Jackson et al. 2014). These effects range from 20 to > 90% decline in coral cover (Porter et al. 2001; Gardner et al.
2003; De’ath et al 2012; Edmunds 2013; Jackson et al. 2014; Hughes et al. 2018).
Major drivers of the decline in coral cover throughout the Caribbean can be attributed to local disturbances including the regional-scale mortality of Diadema antillarum in 1983 (Levitan 1988; Edmunds and Carpenter 2001), tropical storms
(Woodley et al. 1981; Rogers et al. 1991), and disease (Aronson and Precht 2001; Miller et al. 2009). Depending on the nature of the shifts in community structure that have occurred on Caribbean reefs, they can be termed “phase shifts” or “alternative stable states” (Dudgeon et al. 2010). A phase shift occurs when there is a change in the community structure in response to a persistent change in the environment, whereas an alternative stable state occurs when there is more than one stable state occurring in the same area under the same conditions at different times (Dudgeon et al 2010). Current examples of changing communities support the phase shift hypothesis on coral reefs, based on fossil records (Lighty 1980) and data that suggests shifts back to coral dominated communities can occur (Fong et al. 2006; Dudgeon et al. 2010). With
2 increasing number and intensity of natural disturbances affecting coral reefs, there is greater evidence suggesting that shifts in community structure from coral dominated to algal (Hughes 1994 but see Bruno et al. 2009), soft coral (Stobart et al. 2005), or even sponge dominated systems (Ward-Paige et al. 2005; Bell et al. 2013), might be the future of Caribbean coral reefs.
After Hurricane Hugo hit the US Virgin Islands on September 17th and 18th 1989, both
Rogers et al. (1991) and Edmunds and Witman (1991) reported seeing fragmented and overturned coral colonies, sedimentation smothering live corals, and poor water quality due to high turbidity from rainfall (Edmunds and Witman 1991; Rogers et al. 1991). On the south coast of St. John, Hurricane Hugo caused mean percent live coral cover at Yawzi Point (a long-term study site that monitored live coral cover and density and data at 10-12 m depth had been collected prior to Hurricane Hugo (Rogers et al. 1991) to decline from 19.8% (June 1989) to 12.3% (November 1989), with no indication of recovery a year later (11.7%, November 1990) (Rogers et al. 1991). Macroalgae subsequently invaded and colonized the open substrata, and with a decline in abundance of herbivorous fishes (Beets and Rogers 2002), the macroalgae went largely uncontrolled and quickly proliferated over the bare substratum (Rogers and Miller 2006). In another example of disturbance effecting a Caribbean reef, Aronson et al. (2002) discovered that after the 1998 bleaching event on Channel Cay reef in Belize, hard coral cover decreased from ~45% to 5%, but sponge cover increased from 15% to 43% from 1998 to 2001.
Other taxa, such as macroalgae and CTB (CCA, turf, bare rock), declined in cover from
1998 to 2001 (Aronson et al. 2002). Together, these results suggest that benthic taxa
3 other than coral may have the capacity to dominate coral reef systems, especially after high intensity disturbances like large scale bleaching events and tropical storms
(Norström et al 2009).
Anthropogenic stressors affect invertebrate taxa associated with coral reefs, including scleractinian corals, bivalves, and sponges (Harvell et al. 2002; Webster 2007), and with increasing intensity of tropical storms (Solomon et al. 2007), the risk of physical disturbances on benthic invertebrate taxa is expected to increase. Severe storms can influence the community structure of benthic invertebrates through increased wave action, coastal runoff, sedimentation, and turbidity (Rogers 1990). Turbidity due to increased suspended organic particles may contribute positively (Anthony and Fabricius
2000) to some benthic taxa but negatively to others (Elfwing et al. 2003). Edmunds and
Gray (2014) reported seeing increased sedimentation on the reefs of St. John due to greater rainfall associated with four storms that affected St. John in 2010. They linked the disturbances to a decline in coral recruitment and an increase in densities of suspension- feeding polychaetes (Edmunds and Gray 2014). Conversely, Elfwing et al. (2003) saw lower growth rates in giant clams in the Philippines that were exposed to enhanced seawater turbidity attributed to natural and anthropogenic causes. Members of the invertebrate suspension feeding guild on shallow coral reefs (e.g. polychaetes, ascidians, bryozoans, crustaceans, sponges, bivalves,) have developed specialized mechanisms for intaking phytoplankton (Riisgård and Larsen 2010). Increases in sedimentation may allow for an increase in population of suspension feeders if the quality of the sediment particles are nutritionally beneficial.
4
Caribbean sponges play important functional roles on coral reefs. They alter reef complexity by binding coral rubble together (Wulff 1984) and degrading parts of the reef through bioerosion (Wulff 1984). They have the capacity to filter water (Reiswig 1971;
McMurray et al. 2014) and recycle nutrients that can be used by other trophic levels
(Diaz and Rützler 2001; Rix et al. 2015; McMurray et al. 2017). They can also provide places for settlement and habitat structure for other benthic invertebrates (Diaz and
Rützler 2001; Bell 2008). These functional roles are of particular importance after environmental perturbations like hurricanes. Yet, despite their importance to coral reef ecosystems, the impact of anthropogenic stressors on sponges is still under-represented in coral reef research. This deficiency could be due to difficulties in identifying sponge species without using microscopic analyses of spicules, reliable field guides, and/or lack of genetic tools (Diaz and Rützler 2001).
Natural disturbances, like hurricanes, affect sponges differently than scleractinian corals (Rogers et al. 1991; Wulff 1995). In the Caribbean, many sponges are able to recruit and grow quickly allowing them to recolonize benthic surfaces faster than scleractinians following natural disturbances (Wulff 1994), in some cases allowing for higher densities of sponges after the disturbance versus before. Wulff (1994) addressed the abundance of erect sponges on the shallow reefs (1-9 m depth) around the San Blas
Islands after Hurricane Joan struck in October 1988. After surveying for density of sponges, she found a greater density of juvenile sponges after the hurricane then before
(Wulff 1994). This may indicate that while hurricanes and other natural disturbance can
5 dislodge and fragment sponges with relative ease (Woodley et al. 1981; Wulff 1994), their ability to repair (Woodley et al. 1981), recruit, and colonize quickly allows them to be early settlers after major perturbations (Wulff 1994).
Due to their abundance and morphological complexity on many coral reefs, sponges can provide unique habitat structure for a wide diversity of taxa that live on and within the spongin matrix (Wulff 2006; Bell 2008). These taxa are diverse and range from microscopic prokaryotes to macroscopic invertebrates that fill functional roles extending from symbiosis, to mutualism, to commensalism (Bell 2008). Sponges may fulfill more than one functional role in interaction with their associated taxa, which makes it difficult to determine the ecological importance of such relationships (Bell
2008). For example, the sponge brittle star Ophiothrix lineata relies heavily on the tube sponge, Callyspongia vaginalis, as habitat structure on Caribbean reefs. In the Florida
Keys, results showed that 99% of the O. lineata were living within individual C. vaginalis, indicating that O. lineata is a species-specific obligate (Henkel and Pawlik
2005). Another study done by Hendler (1984) on the Belize Barrier Reef revealed that O. lineata and C. vaginalis live together in a cleaning symbiosis, because the gut contents of
O. lineata shows that it consumes particles that adhere to the sponge surface, thereby cleaning the inhalant surfaces of the sponge (Hendler 1984). In Chapter 2, I describe: (1) the susceptibility of sponges differing in size and morphology, to the effects of the hurricane, and (2) the effect of morphology and hurricane effects on the presence of sponge associated epifauna.
6 Global Climate Change
In recent decades, the effects of global climate change (GCC) have become more evident to scientists in both terrestrial and aquatic ecosystems (Walther et al. 2002;
Hughes et al. 2017; Nolan et al. 2018). Global warming, caused by greenhouse gases, causes increasing sea surface temperatures, and the proximal cause of this warming, high concentrations of atmospheric CO2, also leads to ocean acidification (OA) (Zeebe 2012).
Both rising temperature and OA can have negative impacts on coral reefs through coral bleaching and reduced net calcification (Kroeker et al. 2010; Lesser 2011). Warming sea surface temperatures may have profound implications for coral that already live close to the upper end of their thermal tolerance limits (Jokiel and Coles 1990). Mass bleaching events have occurred on reefs throughout the world, notably along the Great Barrier Reef
(Marshall and Baird 2000) and throughout the Caribbean (Goreau et al. 2001), and many have been linked to increased sea surface temperatures (Brown 1997, Lesser 1997, Gates et al. 1992, Hoegh-Guldberg 1999, Rowan 2004). Working synergistically with increasing water temperatures is OA, which is the result of atmospheric CO2 equilibrating with seawater to depress seawater pH and reduce net calcification of biogenic calcifiers
(Doney et al. 2009).
Physiological Responses of Corals to GCC
Global sea surface temperatures have been on the rise for decades and are projected to rise another 5ºC by the year 2100, assuming a higher emissions scenario
(IPCC, 2018). Bleaching is one of the main stress responses exhibited by corals, and large-scale episodes are caused by increasing sea surface temperatures that result in
7 expulsion (or release) of the symbiotic Symbiodinium that live within the host tissue of most tropical corals (Rowan 2004). Even though signs of bleaching have been found in many corals, not all stony corals bleach in response to thermal stress (Bruno et al. 2007;
Hoegh-Guldberg et al. 2007). When there are changes to the environment involving temperature, the effects can modify physiological processes in a wide variety of deterministic ways as described by the Arrhenius function and thermal response curves
(Prosser, 1991). Typically, reaction rates involved in calcification, photosynthesis, respiration, and metabolism are all affected by temperature and, therefore, are affected by global climate change (Jokiel & Guinther 1978; Warner et al. 1996; Reynaud et al. 2003;
De’ath et al. 2009; Edmunds et al. 2011).
Metabolic Plasticity
Metabolic plasticity is the adjustment of an organism’s metabolism (i.e. aerobic respiration) in response to changes in the chemical or physical environment of the organism (Sokolova and Portner, 2002). The range over which metabolism can vary differs inter- and intra-specifically causing, the power output (joules/s) of the organism to vary. Power output is critical for the cellular machinery of scleractinian corals, thus variation in power output is important to evaluate as it allows for the interpretation of biological implications of the metabolic rate. Depending on the life style (i.e., active or sedentary) of the organism, maximal metabolic rates can represent a 2- to 100-fold increase in basal metabolic rates (Willmer et al. 2000). The maximum metabolic rate
(MMR) is reached when the organism has reached its greatest capacity to supply the energy (Joules) required for mechanical and chemical work, whereas basal metabolic rate
8 (BMR) is reached when the organism has decreased its mechanical and chemical work to the minimum metabolic level required for life, with no spontaneous activity (Willmer et al. 2000). The difference between the maximum oxygen consumption (MMR) of an organism at the highest activity level and the minimum oxygen consumption at the lowest activity level (BMR) defines the metabolic scope (Bishop 1999). Metabolic scope indicates the capacity to increase ATP production through aerobic respiration to meet the greatest possible demand for mechanical and chemical work. Anaerobic respiration is an additional process that supplies energy for work, but generally it is assumed to be trivial in terms of the amount of energy it provides (Willmer et al. 2005) relative to the daily energy demand.
Although metabolic scope is defined commonly for vertebrates (Bishop 1999;
Clark et al. 2013) including human athletes (Åstrand and Rodahl 1986), where values of
12-24-fold are common (Åstrand and Rodahl 1986; Bishop 1999; Clark et al. 2013), it rarely has been defined in invertebrates, and never for scleractinian corals. This is unfortunate because understanding the metabolic scope of scleractinian corals can allow us to assess the current metabolic rate of corals on the reef, which can reveal how they are functioning in their current environments. After defining the metabolic scope, a conceptual model of energy output as a function of time (joules/s) can be used to determine the duration that a coral can persist under metabolic rates based on the size of the energy reserves and the capacity to replace them (Peterson et al 2010).
9 MMR and BMR reflect the end points on a gradient of metabolic activity, along which aerobic metabolism can reflect differential energy demands, assuming anaerobic respiration is trivial in terms of supplying ATP to meet the energetic costs of mechanical and chemical work. For scleractinian corals, routine activities (i.e. digestion, polyp expansions/retraction) reflect different processes that alter metabolism. Digestion incurs energetic costs associated with the breakdown and re-assembly of metabolites (i.e. protein), that collectively is termed Specific Dynamic Action, SDA (Jobling 1983). This set of processes (i.e., SDA) ultimately can lead to an increase in metabolic rate that can range from 20% in humans to ≥ 600% in snakes (Secor 2009). Polyp expansion is a prerequisite for particulate feeding in many coral (Lewis and Price 1975). It is also an activity that alters the diffusive boundary layers (DBL) and shortens the diffusion of pathways through tissue, increasing the surface area over which diffusion of solutes from seawater can occur (Patterson 1992; Shashar et al. 1993). Conversely, polyp retraction decreases the surface area over which diffusive exchange with seawater can occur, potentially causing aerobic metabolism to decline under conditions of mass transfer limitation for O2 (Patterson 1992).
After characterizing a gradient under which both routine (i.e., polyp expansion/retraction) and extreme (MMR/BMR) metabolic rates can be measured and compared, an energy budget equation can be used. A simple energy budget can describe the energy used for metabolism (R = C - (F + U + P)). Where R is respiratory loss, C is consumption, F is faecal loss, U is excretion, and P is production or synthesis. Using the energy budget equation can give insight into how much the coral is consuming (i.e.,
10 heterotrophic and autotrophic resources) versus how much output (i.e., energy in Joules) is being released and used by other trophic levels (Crossland et al. 1991) and these how inputs and outputs change based on the metabolic activity. Chapter 3 describes a series of experiments that address the metabolic plasticity of two reef building corals, Pocillopora verrucosa and Acropora pulchra, that were conducted to better to understand variation in metabolism of scleractinian corals. These experiments were conducted to quantify the variation in metabolic rates established by a range of energy demands, which together, defines the metabolic scope.
Anthropogenic Factors and Present-Day Oceans
The objectives of this study were to broadly examine the effects of anthropogenic stressors on tropical coral reefs. This goal was achieved through a program of investigation focused firstly, on the response of coral reef communities to disturbances
(i.e. analyses of sponges in St. John), and secondly, on the metabolic capacity of individual corals in these systems to respond to energetic demands created by physiological stress (i.e. through an analysis of MMR and BMR in Mo’orea). First, I investigated the changes in non-scleractinian invertebrate communities after a major perturbation impacted the reefs of St. John, US Virgin Islands, by examining: (1) the susceptibility of sponges differing in size or volume and morphology, to effects of the hurricane disturbances, (2) the effects of morphology and hurricane disturbances on the presence of sponge-associated epifauna, and (3) the trend of change in population density of sponge communities following natural disturbance and periods of heavy rainfall periods. Second, I investigated the metabolic plasticity of two reef building corals,
11 Pocillopora verrucosa and Acropora pulchra, in Mo’orea, French Polynesia, by: (1) characterizing a metabolic gradient using metabolic rates established by differing energy demands, and (2) using differences in nubbin size to explore size-dependent metabolism and assess if scaling differs based on organismic activity. Together, these studies highlight the importance of research on the ecology and physiology of coral reef communities and to what extent these communities are impacted by anthropogenic stressors.
12 CHAPTER 2
Effects of natural disturbances on non-scleractinian benthic invertebrates
Introduction
Severe tropical storms represent major disturbances shaping the structure and function of multiple ecosystems (Gardner et al. 2005; Fabricius et al. 2008). Their major destructive effects arise from the consequences of high wind speeds, torrential rain, mudslides, large waves, and storm surges (Knutson et al. 2010). Variation in the intensity and return time of these disturbance effects has long been thought to play fundamental roles in maintaining tropical diversity (Connell 1978). Therefore, the strong likelihood that storms will increase in severity under climate change scenarios is cause for concern
(Emmanuel 1987; Cheal et al. 2017). Arguably, these concerns are most acute in the marine environment where coral reefs are uniquely subject to damage by major storms
(Fabricius et al. 2008).
Major tropical storms (i.e., cyclones, typhoons, and hurricanes) can cause a wide variety of damage on coral reefs based on storm intensity, reef exposure, and proximity to the eye of the storm (Woodley et al. 1981). In the benthic realm, the extent of the damage is strongly dependent on the community structure prior to the storm (Zhang et al. 2014), which depends, in part, on the time since the last major event (Woodley et al. 1981).
Most damage results from the physical effects of large waves, leading to extensive coral colony breakage (Woodley et al. 1981), dislodgement of corals from the substratum
(gen), sandblasting of their tissue (Woodley et al. 1981), and burial by rearrangement of
13 benthic sediment (Rogers et al. 1991). Additional mortality can arise from the consequences of heavy rainfall, leading to low salinity events (Haapkylä et al. 2011; Bahr et al 2015), acute terrigenous sedimentation (Rogers 1990), and delayed onset of coral diseases (Knowlton et al. 1981; Haapkylä et al. 2011). Together, the damage to reef communities can range from minimal (Rogers et al. 1984) to severe (Glynn et al. 1964), which can cause rates of recovery of the impacted reefs to vary dramatically (Stoddard
1974: Hughes et al. 2018).
Although there is a long history of documenting the effects of storms on stony corals (Rogers et al. 1991; Hughes 1994; Edmunds 2013), many of the other organisms composing reef communities remain understudied. This is unfortunate because benthic taxa other than corals can have ecological significance to reef communities, including sea urchins, macroalgae, and sponges (Levitan 1988; Edmunds and Carpenter 2001; Wulff
2001). Therefore, the likelihood that storms severely impact these organisms has important consequences. In the Caribbean, sponges play particularly well-developed functional roles on coral reefs, extending from filtration and particle entrapment
(McMurray et al. 2014 and references therein), to binding mobile substrata together
(Wulff 1984), recycling organic carbon (Rix et al. 2015), primary production (Southwell et al. 2008), and habitat provisioning (Bell 2008). Presumably, these functions are interrupted when major storms dislodge and fragment sponges (Wulff 1994), although following the damage, the role of sponges in binding substrata together can be particularly important (Wulff and Buss 1979; Wulff 1984).
14 Storms have the capacity to hinder sponges in their role as habitat for other taxa.
This could have particularly important consequences, because of the wide diversity of taxa living on and within sponges (Rützler 1976; Villamizar and Laughlin 1991; Ribeiro et al. 2003; Bell 2008). The taxa that associate with sponges can range from microscopic prokaryotes (Taylor et al. 2007) to macroscopic invertebrates (e.g., crustaceans, ophiuroids, and cnidarians), which interact with their host sponges through mutualistic, commensal, or parasitic relationships (Wulff 2006; Bell 2008). A leading determinant of the fauna inhabiting sponges, is sponge size and morphology, with vase and goblet sponges inhabited by taxa differing from those occurring in more globose species
(Koukouras et al. 1992; Henkel and Pawlik 2014; but see Ribeiro et al. 2013). As the impact of storms on sponges is dependent on the size and aspect ratio of the sponges affected, it is likely that storms will have predictable effects on sponges. Additionally, storm impacts can affect sponge-associated fauna, based on the susceptibility of individual sponges to drag and dislodgement (Wulff 1994).
St. John, in the US Virgin Islands, provides an interesting location in which to explore the effects of storms on the sponge communities of coral reefs, because the ecological history of these communities is well known (Randall 1961; Rogers et al. 1991;
Edmunds 2002), including the effects of the last severe hurricane in 1989 (Edmunds and
Witman 1991; Rogers et al. 1991). Hurricane Hugo (Category 4) passed directly over St.
Croix, US Virgin Islands, causing damage to hard coral and soft-bottom sponge and gorgonian communities (Rogers et al. 1991; Hubbard et al. 1991). Due to the intensity of
Hurricane Hugo, some reefs surrounding the nearby islands, St. John and St. Thomas,
15 were also severaly impacted. Colvard and Edmunds (2010) reported a 64% increase in sponge abundance on the shallow reefs of St. John, US Virgin Islands from 1992 (3 years after Hugo) to 2004 where coral cover remained ~ 5%. Nearly three decades later in
2017, Hurricanes Irma (Category 5) and Maria (Category 5) hit the USVI within 10 days causing destruction of reefs in St. John (Edmunds 2019).
The present study focuses on the effects of the recent storms on two common species of vase sponges on shallow Caribbean coral reefs, Ircinia campana (0.92 sponges m2) and Niphates digitalis (1.72 sponges m2) (Edmunds unpublished). Ircinia campana was chosen based on its flexible skeleton made of fine spongin fibers and spicules that are resistant to tearing, breaking, or other types of damage caused by a natural disturbance (Wulff 1994). This flexible skeleton also helps prevent I. campana from being toppled, compared to other species (Bergquist and Bedford 1978; Wulff 1994). The resistance of the skeletal structure allows the morphology (i.e., wide-open cup with a small base) to persist in shallow water (Wulff 1994), moreover, this morphology provides microhabitat structure for other invertebrates that can live within or upon the spongin matrix (Wendt et al. 1985; Prentiss and Harris 2011). Additionally, I. campana is defended chemically by secondary metabolites which protect it from predation (Walters and Pawlik 2005). Niphates digitalis was chosen because it exhibits a shorter tube morphology compared to I. campana, with a rough exterior and an opening encircled by spines that are connected by a stiff membrane (Humann et al. 2002). Niphates sp. also supports epifauna, that can include mainly brittle stars (i.e. Ophiothrix suensoni and
Ophiothrix lineata), and sponge zoanthids (Parazoanthus parasiticus) (Wulff 2006;
16 Henkel and Pawlik 2011; Evans 2012), however, N. digitalis is not chemically defended from predation (Walters and Pawlik 2005). Additionally, both sponge species were chosen based on their high abundance on Caribbean coral reefs (Humann et al. 2002;
Wulff 2006; Villamizar et al. 2014).
Documenting and evaluating spatiotemporal changes in sponge density and community composition following natural disturbances is critical in understanding how these events change benthic coral reef communities. Using surveys completed both before (August 2017), two months after (November 2017), and 11 months after (August
2018) two major storms, three questions were addressed: (1) did susceptibility to the storms differ between sponge taxa due to storm effects on individuals differing in size.
Due to differing aspect ratios, size was measured as linear for N. digitalis and volumetric for I. campana, (2) did susceptibility to the storms differ between sponge taxa due to storm effects on individuals differing in morphology, and (3) did morphology have an effect on the presence of epifaunal invertebrates living on the sponges, was the presence of invertebrates altered by storm effects.
Materials and Methods
Overview
The reefs of St. John have been studied since the 1950’s (Randall 1961) and monitored since 1987 (Rogers and Miller 2006; Edmund 2013). Since then, researchers have gathered time-series data, which allows them to detect changes to reef communities after natural disturbances over longer periods of time (Edmunds and Witman 1991;
17 Rogers et al. 1991). Surveys of two shallow Caribbean vase sponges, Ircinia campana and Niphates digitalis, were completed in August 2017 (pre-hurricanes), November 2017
(2 months after the hurricanes), and August 2018 (11 months after the hurricanes) around
Great Lameshur Bay in St. John, US Virgin Islands, which is within the Virgin Islands
National Park (VINP). These two species were chosen based on their ease of identification, abundance on Caribbean reefs including those around St. John (Edmunds unpub.), differences in the composition and organization of their skeletal structures
(Wulff 1994; Humann et al. 2002), and their capacity to house invertebrate epifauna
(Henkel and Pawlik 2011; Wulff 2006; Villamizar et al. 2013). Six random sites were initially surveyed on shallow fringing reefs (5-9 m depth) of Great Lameshur Bay in
August 2017, stretching from White Point to East Cabritte. Six additional sites were selected at random for sampling in November 2017 and August 2018 to compare with time series data that dates back to 1992 (Fig. 1).
Sponge Abundance
To quantify sponge density and their response to hurricane disturbance, a band transect (10 x 5 m) was used at each site (n = 4 per site) at a depth of 5-9 m, with bands deployed parallel to the shore. Sponges were counted at each site and averaged per transect, which served as a statistical replicate. Sites were sampled before (August 2017),
2 months after (November 2017), and 11 months after (August 2018) Hurricanes Irma and Maria. Six sites were sampled in August 2017, with four additional sites added in
November 2017 and August 2018 (Fig. 1).
18
Size
The volume of Ircinina campana and size of Niphates digitalis were estimated to assess size distributions in St. John, and detect differences in size after the hurricanes.
Sponge volume (cm3) for I. campana was estimated as a hollow cone by calculating the
2 2 difference between two cones ((π x r1 x (h1/3)) - (π x r2 x (h2/3))), where h1 is the height of the entire sponge, r1 is the outer radius of the entire sponge, h2 is the height of the inner cone, and r2 is the inner radius of the bowl. N. digitalis has a narrower tube morphology with infoldings and invaginations that do not conform to a cone or cylinder like I. campana. Size of N. digitalis was calculated as the average of its height and width of opening (cm).
Invertebrate Assemblage
Sponges have the capacity to provide critical habitat for a wide diversity of taxa that may live on or within the sponge tissue (Wulff 2006; Bell 2008). Abundance and morphology of sponge individuals influences the type and number of taxa that are found in association with their biomass (Koukouras et al. 1992). Since this research was conducted in the Virgin Islands National Park (VINP), it was not possible to manipulate sponges in the field or collect and observe them in a laboratory, so invertebrates living inside the matrix of the sponges were not accounted for in this study. The presence of invertebrate epifauna based on sponge morphology was examined by observing, identifying (i.e. to genus), and recording the invertebrate species living on the surface inside the osculum of the sponge. Paguristes erythrops, Calcinus tibicen, Ophiothrix
19 suensoni, and Pelia mutica were found in the N. digitalis and Paguristes erythrops,
Calcinus tibicen, and Ophiothrix suensoni were found in I. campana. When comparing the abundance of invertebrate epifauna between N. digitalis and I. campana, the density of invertebrates was standardized to the length of each sponge (cm).
Statistical Analyses
To examine differences in the density and size of I. campana and N. digitalis following a major natural disturbance, a univariate permutational multivariate analysis of variance (PERMANOVA) was used. A univariate PERMANOVA was used to test for differences in mean density of invertebrates (per cm of sponge) due to sponge species
(i.e. I. campana and N. digitalis), invertebrate epifauna (C. tibiens, O. suensoni, P. mutica), and time.
The data analyzed with a univariate PERMANOVA (tested in a permutational framework yielding a Pseudo-F and P-perm) (Anderson 2001), were log-transformed or square root transformed to ensure that the assumptions were met. All analyses were conducted using the open source software R ver. 3.5.1 (R Core Team 2018 [Vegan,
Permute, Lattice]),
Results
A PERMANOVA showed that mean density of sponges (individuals m-2) differed significantly between sponge species (Pseudo-F1,110 = 26.1, P = 0.001), but not between times (Pseudo-F2,110 = 0.973, P = 0.34), or the interaction term between species and time
20 (Pseudo-F2,110 = 1.02, P = 0.36) (Fig. 2). Before Hurricanes Irma and Maria (August
2017), the mean density of N. digitalis was 79% higher than I. campana on the shallow reefs of St. John.
A PERMANOVA indicated that the mean volume (cm3) of Ircinia campana
(Pseudo-F2,285= 13.4, P = 0.001) (Fig. 3), and mean size (height, cm) of N. digitalis
(Pseudo-F2,529 = 74.9, P = 0.001) (Fig. 4) were significantly affected by time. A post-hoc
Tukey HSD test (α = 0.05), revealed that mean size of I. campana declined significantly between August 2017 (267 cm3) and August 2017 (121 cm3) and the mean size of N. digitalis declined significantly between August 2017 (6.97 cm) and November 2017
(5.69 cm).
Of the sponges sampled in August 2017, 51% (n = 189 sponges) of Niphates digitalis and 22% (n = 165 sponges) of Ircinia campana contained epifaunal invertebrates
(Paguristes erythrops, Calcinus tibicen, Ophiothrix suensoni, Pelia mutica), but these percentages declined to 24% (n = 381 sponges) and 8% (n = 123 sponges) in August
2018, respectively. Of the sponges that contained invertebrates, 39% of N. digitalis and
22% of I. campana contained more than one invertebrate organism in August 2017. A
PERMANOVA indicated that mean number of invertebrates per cm of sponge
(standardized to height) was significantly affected by sponge species (Pseudo-F1,2399 =
32.9, P = 0.001), and type of invertebrate species (Pseudo-F2,2399 = 14.2, P = 0.001), but was not affected by time (Pseudo-F2,2399 = 0.092, P = 0.75). There was also a significant interaction between sponge species and invertebrate species on mean density of
21 invertebrates per cm of sponge (Pseudo-F2,2399 = 5.6, P = 0.02), but the three-way interaction between sponge species, time, and type of invertebrate was not significant
(Pseudo-F1,2399 = 0.26, P = 0.64) (Figure 5). A Post-hoc Tukey HSD test (α = 0.05) indicated that there were significantly more Ophiothrix suensoni (P < 0.05) and Pelia mutica (P = 0.044) per cm of sponge in N. digitalis then I. campana, but the abundance of Calcinus tibiens and Paguristes erythrops (P = 0.98) (cm-1 of sponge) were similar between sponge species. Overall, N. digitalis had a significantly greater abundance (94% higher) of invertebrate fauna (cm-1) then I. campana.
Discussion
Natural disturbances can play an important role in maintaining species diversity among many ecosystems (Connell 1978) and the likelihood that natural disturbances, like hurricanes, will increase in severity over time (Emmanuel 1987) may cause shifts in community structure and composition (Dudgeon et al. 1999). Storms effects on scleractinian corals have been documented widely in the literature (Rogers 1991;
Edmunds and Witman 1991; Edmunds 2013) but such effects on other benthic invertebrate taxa remain understudied. Caribbean sponges are one of the most abundant taxa on coral reefs (Dubinsky and Stambler 2011) and provide numerous benefits to coral reef communities (Bell 2008; Wulff 2006), but they are susceptible to dislodgement and fragmentation by major storms (Woodley et al. 1981; Wulff 1994). Despite susceptibility to physical forces, many sponge species have the capacity to recruit and colonize benthic surfaces quickly following natural disturbances (Wulff 1994). Major storms also cause increased sedimentation due to rain induced terrestrial runoff (Rogers 1990, Rogers et al.
22 1991; Fabricius 2005; Edmunds and Gray 2014). This increase in turbidity can be detrimental to corals and other invertebrate fauna (Rogers et al. 1991; Edmunds and
Witman 1991), but can be beneficial to members of the suspension feeding guild, like sponges (Fabricius 2005; Edmunds and Gray 2014).
The present study was motivated by the morphological differences of two
Caribbean vase sponge, their capacity to house invertebrate epifauna, and the impacts that two category five hurricanes had on these sponge species in St. John, US Virgin Islands.
This study tested the hypotheses that: (1) susceptibility to the storms differed between sponge taxa over time due to storm effects on individuals differing in size, (2) did susceptibility to the storms differ between sponge taxa due to storm effects on individuals differing in morphology and (3) did morphology have an effect on the presence of epifaunal invertebrates living on the sponges, was the presence of invertebrates altered by storm effects. I tested the hypotheses by first surveying several sites in the Virgin Island
National Park (VINP) for the density, size and invertebrate assemblages within two
Caribbean vase sponges Ircinia campana and Niphates digitalis.
In the case of the first hypothesis, overall sponge density varied between species with, N. digitalis being more abundant than I. campana, among the sites sampled in St.
John. However, losses from these sponge populations due to Hurricanes Irma and Maria were not significant, meaning that the hurricane disturbance did not significantly affect the density of either sponge. This could be due to the resiliency of the siliceous spicules of N. digitalis or fine spongin fibers of I. campana (Humann et al. 2002; Wulff 1994). I.
23 campana has an extensible and flexible skeleton that may be beneficial for preventing toppling from wave forces and resistant to tearing or ripping (Berquist 1978; Wulff
1994), however once toppled and removed from the substratum, observed I. campana appeared to be unable to reattach to the substratum, which may lead to mortality. After
Hurricane Allen impacted Jamaica in 1989, Woodley et al (1981) reported seeing fracturing and tearing of many different sponges in Discovery Bay (15 m depth), but sponges with a tough texture, like Ircinia sp., were structurally unaffected. Conversely,
N. digitalis is shorter and has more rigid skeletal tissue. It is positioned lower on the substratum, typically on the vertical surface of large boulders, which may offer protection from wave forces. N. digitalis is able to regenerate some tissue, within a few days, after effects of predation (Pawlik et al. 2008) and necrosis (Wulff 2013).
To gain further insight into the likely cause of no difference in density
Additionally, in a survey quantifying washed up sponges on the shore of Great Lameshur
Bay November 2017 (two months after Hurricanes Irma and Maria), 1.96 pieces of sponge m-2 were found washed up on the beach. A subsample of these sponges revealed that ~90% of the pieces were rope sponges in the genus Aplysina. A repeat of this survey in August 2018 showed an increase to 14.23 pieces of sponge m-2. Out of the sponges surveyed in August 2018, ~56% were rope sponges, ~12% were vase sponges, and 32% were unidentifiable fragments. This pattern coincides with literature that suggests that rope sponges are more likely to be affected by storm effects that cause drag and dislodgement (Woodley et al. 1981) and sponges with higher spongin content are less likely to wash up on the beach (Wulff 1994; Ávila et al. 2011) The susceptibility of rope
24 sponges to waves may help to explain why the densities of I. campana and N. digitalis, with a vase/tube morphology, were not significantly affected by the recent hurricanes in
St. John.
Volume and Size
Size can determine the effects of storm on sponges due to the roles of drag and dislodgement on some sponge species (Wulff 1994). The present results show that the mean individual volume (cm3) of I. campana, decreased from August 2017 (267 cm3) to
November 2017 (120.8 cm3), indicating that I. campana with a larger volume preferentially were removed by the storms, most likely because they were exposed to enhanced drag. Mean size (cm) of N. digitalis decreased from August 2017 (7 cm) to
November 2017 (5.7 cm) to August 2018 (4.6 cm), indicating that larger individuals of this species were also impacted by storm effects. However, there was an increase in abundance of smaller sponges indicating that N. digitalis was able to recruit and colonize quickly by settling or reattaching to recently cleared benthic space (Wulff
1995). Additionally, N. digitalis typically release brooded larvae in July and August, during the daytime between 7-11am (Lindquist et al. 1997). Because of the photonegative behavior of N. digitalis larvae, they tend to settle in low light microhabitats (Lindquist et al. 1997). With the increase in suspended sedimentation and reduced underwater light intensity that occurred in the weeks following Hurricanes Irma and Maria (Edmunds et al. in prep), N. digitalis larvae may have been able to settle in light exposed shallow areas that were previously inhospitable to larvae. This could explain why there was an increase in density of smaller N. digitalis after a heavy rainfall
25 period (September 1 – November 30, 2017) that caused increased suspended particulate
(Edmunds et al. in prep) with potentially high nutrient content.
Invertebrate Assemblage
As a dominant component of the benthos, sponges have the capacity to provide habitat structure to many different taxa (Pawlik 1983; Duffy 1992; Koukouras et al. 1992;
Henkel and Pawlik 2014). Overall, N. digitalis contained more Ophiothrix suensoni
(sponge brittle star) and Pelia mutica (cryptic treardrop crab) per cm of sponge
(standardized to height) then I. campana (Fig. 5). Research suggests that N. digitalis is a commonly used as a microhabitat structure by the sponge dwelling brittle star, Ophiothrix sp. (Henkel and Pawlik 2005). Results showed that 99% of the O. lineata were living within individual C. vaginalis, indicating that O. lineata is a species-specific obligate
(Henkel and Pawlik 2005).
To date there have been no studies examining the relationship between Pelia mutica and select sponge species. Pelia mutica is a type of decorator crab decorates its carapace with easily torn materials taken from sponge, algae, bryozoans, and ascidians
(Wicksten 1980). Decorating the carapace with sponges is of particular value because sponges produce secondary metabolites that are unpalatable to many fish species potentially creating a defense mechanism against predation (Wicksten 1980; Wadell and
Pawlik 2000). Sponge-decorated carapaces are also useful as camouflage when living within or around the host sponge (Wicksten 1980). Calcinus tibicen and Paguristes erythrops (hermit crabs) were similar in abundance between I. campana and N. digitalis
26 and before and after Hurricanes Irma and Maria. This is not surprising considering that hermit crabs, such as Calcinus tibicen, often form numerous associations with other taxa ranging from incidental, to facultative, to obligate (Ross and Sutton 1961; Williams and
Dermont 2004 [and references therein]; Brown and Edmunds 2013). These associations indicate that hermit crabs, like Calcinus tibicen, may move from one habitat to another quite frequently looking for food, shelter or other resources to exploit.
Documenting and evaluating spatiotemporal changes in sponge density and community composition following natural disturbances is critical in understanding how these events shape benthic coral reef communities. Sponges are critical components of a reef that have numerous functional roles (i.e. habitat structure, reef complexity, nutrient recycling). These results suggest that some sponge species may be more resilient to natural disturbances than other taxa, including scleractinian corals. The results also suggest that certain sponges may be more morphologically able to provide habitat structure then other sponge species. Additionally, sponge communities may benefit from storm effects supplied by natural disturbances (i.e. increased sedimentation), whereas other reef building communities (i.e. coral) may suffer. Overall, sponge communities provide numerous benefits that are crucial for the structure and function of coral reef communities. With the increase in natural disturbance intensity, sponges may be becoming more abundant on impacted reefs and thus more research regarding the implications of sponge dominated reefs is needed.
27 Figures
Figure 1. Map of sponge sampling locations along the south shore of St. John, US Virgin Islands. (Blue and Yellow dots = sampled August 2017, November 2017, and August 2018; Black dots = sampled November 2017 and August 2018)
28
0.3
y
t i
s 0.2
n Time e
17−Aug
D
e 17−Nov
g 18−Aug
n 0.1
o
p S
0.0 IC ND Species
Figure 2. Bar graph showing mean (±S.E., [n = 4, except for Nov 17 [n = 2]) sponge density (individual m-2) of Niphates digitalis (ND) and Ircinia campana (IC), two months before (Aug-17), two months after (Nov-17), and eleven months after (Aug-18) Hurricanes Irma and Maria.
29
a 5 b b
) 4 3
m Time c
( 3 17−Aug
e 17−Nov m 18−Aug
u 2
l o V 1
0
17−Aug 17−Nov 18−Aug Time Figure 3. Mean (±S.E., [Aug 17: n = 165, Nov 17: n = 49, Aug 18: n = 71]) volume (cm3) of I. campana at sheltered and exposed sites two months before (Aug-17), two months after (Nov-17) and eleven months after (Aug-18) Hurricanes Irma and Maria hit St. John.
30 2.0 a
b
1.5 b
) Time m
c 17−Aug ( 1.0
17−Nov
e z
i 18−Aug S 0.5
0.0 17−Aug 17−Nov 18−Aug Time Figure 4. Mean (±S.E., [Aug 17: n = 190; Nov 17: n = 116; Aug 18: n = 224]) size (cm) of N. digitalis two months before (Aug-17), two months after (Nov-17), and eleven months after (Aug-18), Hurricanes Irma and Maria.
31
0.08 m
c 0.06
r
e
p
s
e t
a 0.04 August 2017
r b
e November 2017
t
r e
v 0.02 August 2018
n I
0 C. tibiens O. suensoni P. mutica Invertebrate Species
0.08 m
c 0.06
r
e
p
s
e August 2017 t
a 0.04
r November 2017
b
e t
r August 2018 e
v 0.02
n I
0 C. tibiens O. suensoni P. mutica Invertebrate Species
Figure 5. Mean (±S.E., [Aug 17: n = 189 (ND), n = 165 (IC); Aug 18; n = 381 (ND), n = 123 (IC)]) number of invertebrates (cm-1) inside individual A) Niphates digitalis and B) Ircinia campana, two months before (Aug 2017), two month after (Nov 2017), and eleven months after (Aug 2018) Hurricanes Irma and Maria.
32
a b c )
d e
Photo 1. Niphates digitalis with associated invertebrate epifauna (a and b) Pelia mutica and Calcinus tibicen, (c) Ophiothrix suensoni, and (d) tagged N. digitalis with its (e) associated invertebrate assemblage, Pelia mutica and Ophiothrix suensoni.
33 b) c)
a)
Photo 2. a) Ircinia campana with its associated epifauna, b,c) Calcinus tibicen.
34 CHAPTER 3
Characterizing the Metabolic Gradient of Two Reef Corals: Understanding More than Just Highs and Lows
Introduction
Metabolism or the “fire of life” can be defined as the summation of anaerobic and
aerobic respiration that generates ATP for chemical and mechanical work and is involved
in many important biological processes (Kleiber 1961; Brown et al. 2004; Glazier 2010).
It can reflect differential flux of respiratory substrates through glycolysis and aerobic
respiration, and these fluxes drive differential use and exchange of transformed energy
within the organism and the environment (Brown et al. 2004). Scleractinians can
demonstrate a wide range of metabolic rates depending on metabolic activity, allowing
metabolism to be plastic (Fry 1947). This means that metabolism can vary in response to
changes in the chemical or physical environment (Sokolova and Portner, 2002). Scientists
have studied metabolic plasticity in scleractinians for decades, focusing on changes in
metabolism due to light intensity (Porter et al. 1984), temperature (Coles and Jokiel 1977;
Jokiel and Cole 1990; Haryanti and Hidaka 2015), morphology (Bruno and Edmunds
1998; Hoogenboom et al. 2008), and flow rate (Bruno and Edmunds 1998; Patterson and
et al. 1991), but less attention has been paid to magnitude of the range of metabolic rates
corals can exhibit or the ecological and physiological fitness consequences of the
metabolic range.
Organismic metabolic activity can vary inter- and intra-specifically causing, the
power output (Joules/s = Watts) of the organism to vary. Power output is critical for the
35 cellular machinery of scleractinians corals, and the variation in which is important to measure experimentally as it allows for the interpretation of the biological implications of metabolic rate. The maximum metabolic rate (MMR) is reached when the organism has reached its greatest capacity to supply the energy (Joules) required for mechanical and chemical work, whereas basal metabolic rate (BMR) is reached when the organism has decreased its mechanical and chemical work to the minimum metabolic level required for life, with no spontaneous activity (Willmer et al. 2000; Savage et al. 2004). Depending on the life style (i.e. active or sedentary) of the organism, maximal metabolic rates can represent a 2- to 100-fold increase from basal metabolic rates (Willmer et al. 2000).
The difference between BMR and MMR defines the metabolic scope, which is the capacity to increase the flux of metabolites through aerobic respiration to meet the greatest possible demand for mechanical and chemical work (Willmer et al. 2000).
Metabolic scope is an important parameter for studying organismic success because it demonstrates the capacity of an organism to increase maximum aerobic state when transitioning from resting to active (Savage et al. 2004). Depending on the lifestyle of the organism, the metabolic scope can vary dramatically. For example, the metabolic scope for active organisms like fish, birds, reptiles and mammals can vary from 8-20, whereas sessile organisms like barnacles and periwinkles it, is only ~ 2-4 (Åstrand and
Rodahl 1986; Bundle et al., 1999; Willmer et al. 2005).
Because metabolism is a fundamental process of life, understanding the relationship between whole-organism metabolic rate and body mass (i.e. metabolic scaling) is also
36 key in understanding how activity affects metabolic rate (Brown et al. 2004). The relationship between metabolic rate and body mass can be expressed as a power function
R = aMb, where R is the metabolic rate (aerobic respiration), a is a constant, M is the body mass, and b is the scaling exponent. This power function quantifies metabolic scaling (i.e., how biological processes covary with size) through a continuously distributed, and rigorously defined dependent variables (i.e., b).
Understanding how corals respond physiologically to anthropogenic stressors such as temperature and ocean acidification is critical in projecting long-term survival of corals and the reefs they build. An important part of this effort is obtaining ecologically relevant measurements of metabolism that accurately reflect the in situ daily energy requirements of corals. In the present study, for the first time, the metabolic scope of two scleractinian corals was measured, with these analyses focusing on Pocillopora verrucosa and
Acropora pulchra that are important components of the coral reef community in
Mo’orea, French Polynesia. Using a metabolic uncoupler (i.e., 2,4-dinitrophenol) to elicit the maximum metabolic rate (MMR), and starvation to elicit a basal metabolic rate
(BMR), I describe the gradient of metabolism from lowest to highest, and place the metabolic rates associated with polyp expansion and digestion along this gradient. The goals of this study were to: (1) characterize a metabolic gradient by determining if metabolic rates differ based on metabolic activities that require differential energy demands, and (2) use differences in coral size to explore size-dependent (i.e., biomass) metabolism and assess if scaling differs based on organismic activity.
37 Materials and Methods
Fragments of Pocillopora verrucosa and Acropora pulchra were used in this experiment. Both species are ecologically important and common shallow water reef building corals in the Indo-Pacific (Done et al. 1991). Colonies of P. verrucosa were identified based on the morphological traits identified in Veron (2000). P. verrucosa colonies (120-170 mm diameter) were collected from 1-2 m depth in the back-reef on the
North shore of Mo’orea, French Polynesia. Coral were collected on two occasions, one from 23–27 January 2018 and one on 6-10 April 2018 (17º28’31.94”S, 149º48’59.40”W), with corals from the two periods used in the same experiment to measure aerobic respiration. Colonies were selected haphazardly, ~3 m apart to reduce chance that the sampled colonies were clone-mates arising from fragmentation. Fragments of Acropora pulchra were collected at 5 m, from the common garden (17º29’01.80”S,
149º49’59.05”W) on 18 April 2018 and were harvested from colonies that had been on the common garden for 8 months. These colonies had been collected from several fringing reefs (1-2 m) on the north shore. Colonies were selected from several genotypes that were present on the common garden. Coral fragments were removed from parent colonies with a hammer and chisel (P. verrucosa) or pliers (A. pulchra) and transported to the UC Berkeley Richard B. Gump South Pacific Research Station.
To test for size-dependent differences in metabolic rate, coral fragments in two size classes (1-2.5 and 3-4.5 cm in length) were prepared using a band saw (Gryphon
Diamond Band Saw, Model C-40). This preparation resulted in damage to the tissue around the skeletal base only, and the skeleton was affixed to a base. The fragments were
38 fixed in a vertical orientation to a plastic base using Coral Glue (ECOTECH Elements), thereby creating nubbins. After the glue set, the nubbins were placed in a 150 L seawater tank at 29.16 ± 0.03 °C (± S.E. n = 300), which was similar to the temperature conditions found in the back reef 28.93 ± 0.01 °C (± S.E. n = 1875) (MCR LTER). Temperature was measured daily with a digital thermometer (Fisher Scientific, 15-077-8). The nubbins were left in the acclimation tank to recover from fragmentation and transportation for 5-7 days. The tanks were illuminated by SOL White LED, 6000 K, Aqua-Illumination lights, that were operated at a target photon flux density (PFD) of ~700 μmol photons m-2 s-1.
Light was measured daily in the tank, and mean PFD over the course of the experiment was 717 ± 11 μmol photons m-2 s-1 (± SE n = 35, measured with a Li-Cor 4-π quantum sensor LI-193). Tanks were illuminated on a 12:12 h light:dark photoperiod to mimic the light regime in the back reef of Mo’orea in February to April. Seawater was supplied to the tanks at 12 L h-l, and was pumped from ~12 m depth in Cook’s Bay and filtered through a ~500 μm sand filter. A nubbin from each size class was randomly placed in one of four 150 L tanks that either had light (~700 μmol photons m-2 s-1) (used to establish a maximum metabolic treatment) or dark conditions used to establish a basal metabolic treatment). The seawater in each tank was chilled, heated, and circulated independently with a water pump that pumped water through a chiller (Aqua Logic’s Multi-Temp and
Titan Series Chiller). A total of 5 nubbins from each size class (1-2.5 and 3-4.5 cm in length) and colony were randomly placed in one of four tanks, and the position of the nubbins was shifted daily to decrease tank effects due to light and flow.
39 Corals were exposed to 12 h of darkness before dark respiration was measured, with this duration selected to prevent photosynthetic production of oxygen by the Symbiodinium in the coral tissue (Patterson et al. 1991). Respiration was measured as oxygen uptake with a
Ruthenium based optrode (FOXY-R, 1.58mm diameter, Ocean Optics, Dunedin, FL,
USA) connected to a spectrophotometer (USB200, Ocean optics) combined with a light source (USB-LS-450, ocean optics) and connected to a computer with the manufacturer’s software (OOISensor). The probe was calibrated using a zero solution (~ 2 mL of 0.01M sodium tetraborate with a few crystals of sodium sulphite). The probe was rinsed in DI water and placed in the 100% solution (water-saturated air) at the treatment temperature.
Respiration measurements were taken in a confined respirometer consisting of a circular acrylic chamber that was 7.8 cm diameter and had a volume of 270 mL. Water displacement of corals within the chamber was recorded and accounted for in calculations of respiration rate. The chamber was surrounded by a water jacket that was used to keep the seawater at 29.2 °C (same temperature as the treatment tanks), and a stir bar in the base of the chamber was used to provide water motion around the enclosed coral. Water within in the respirometer’s water jacket was circulated using a Fisher Scientific Isotherm
Refrigerated recirculating bath to keep it at a constant temperature. The stir bar was operated at ~15 rpm to create a water flow of ~25 cm/s to create ecologically relevant flow at 2 m depth in the back reef. Flow speed in the chamber was measured by photographing hydrated brine shrimp eggs in the chamber (Sebens and Johnson, 1991).
The duration of incubation in the chamber varied from 5-45 min, or until the oxygen saturation declined to 85%. Respiration was not recorded below this saturation to avoid a confounding effect on respiration rate (Edmunds and Davies 1986). O2 saturations were
40 -1 -1 recorded as % O2 time and converted to concentration (μmol O2 mL ) using O2 solubility (Garcia and Gordon, 1992). The rate of oxygen consumption was calculated from a linear regression model of O2 concentration over time, and rates were corrected using a blank respiration chamber filled with filtered seawater only.
Measuring Maximum Metabolic Rate
One common means to stimulate maximum metabolic rate is to use a metabolic uncoupler that separates the electron chain of the TCA cycle from proton gradients, thus causing aerobic metabolism to operate at a maximal rate without producing ATP
(Willmer et al. 2009). A variety of chemicals have been used for this purpose (Chalker and Taylor 1975; Blackstone 2003), with one common reagent being 2,4-Dinitrophenol
(2,4-DNP) (Immers and Runnstrom 1960; Rognstad and Katz 1986). For example, when administered at 5 x 10-5 M to sea urchin eggs, respiration increased by 170% (Immers and
Runnstrom, 1960). Metabolic uncouplers have to be administered with caution, however, for the concentration applied can cause confounding effects (i.e., increase or decrease in aerobic respiration), and the carrier (i.e., ethanol) also can have effects that are in addition to the uncoupler solute (2,4-DNP). For these reasons, applications of uncouplers to invertebrates need a series of preliminary experiments to establish optimal concentrations and to test for confounding effects of the carrier solvent.
To elicit a maximum metabolic rate in P. verrucosa and A. pulchra, 2,4-
Dinitrophenol was applied. A 2,4-Dinitrophenol titration initially was performed to determine the concentration that yielded the maximum metabolic response for both
41 species. Coral nubbins were dosed with four different 2,4-DNP concentrations (Immers and Runnstrom 1960) and incubated in a 270-mL chamber for ~15 min before dark respiration was measured. The concentration that yielded the highest aerobic respiration was 1x10-4 M, and respiration subsequently decreased at 2.5x10-4 M, indicating a threshold response. Ethanol (500 μl of 70%) was used as a carrier to dissolve 2,4-
Dinitrophenol in filtered seawater (seawater alone was not sufficient to completely dissolve the 2,4-DNP). The effect of 500 μl of 70% ethanol on aerobic respiration was compared to a control treatment (no ethanol) and indicated no significant effect (P > 0.05, n = 4).
Measuring Basal Metabolic Rate
To stimulate a basal metabolic response, one common method applied to symbiotic corals is to prevent photosynthetic production of algal symbionts by keeping corals in darkness for a period of time specific to that species (Jacobson et al. 2016).
Many hard corals rely on their algal symbionts (Symbiodinium) to provide food resources
(Falkowski et al. 1984), and through the photosynthetic capabilities of the Symbiodinium, corals in shallow water can receive the majority (60-90%) of their daily needs for carbon resources (Muscatine et al. 1981; Falkowski et al. 1984). Fluctuations in light intensity varies photosynthetic rate, which alters the supply of carbon to the coral host and causes the host to rely on other methods (heterotrophy) to supply carbon for respiration
(Falkowski et al. 1984).
To measure basal metabolic rate in P. verrucosa and A. pulchra, nubbins were starved of autotrophic resources by keeping them in darkness within tanks (150 L) fully
42 screened from ambient light with black plastic sheets. Seawater was pumped from ~12 m depth in Cook’s Bay and filtered through a ~500 μm sand filter. All other conditions (i.e. temperature and flow) remaining the same as stated above. To determine the appropriate time in darkness to elicit a basal metabolic response, nubbins were placed in darkened tanks for 24, 48, 72, or 96 h in April, and aerobic respiration was recorded at each time as described above. Nubbins were fixed in 10% formaldehyde in seawater for 24-48hr then decalcified in 5% HCl for 48hrs (Davies 1980). The tissue layer was washed with fresh water and transferred to a 10 mL tube where it was homogenized with an ultrasonic dismembrator (Fisher 12-338-550) that was fitted with a 3.2-mm diameter probe. Volume
(ml) of the tissue homogenate was recorded to calculate total Symbiodinium in the tissue following symbiont cell counts. Six replicate counts of the Symbiodinium in the slurry were completed using a haemocytometer. The time in darkness that yielded the lowest minimum metabolic rate was 72-96 h. Therefore, the nubbins were incubated in darkness for 96 h before a basal metabolic rate was determined by the method described above.
Measuring Routine Metabolic Rate
MMR and BMR are the extremes of metabolic scope, however there are other activities that alter metabolic rate through differential energy demands, and these can be useful to consider when characterizing a metabolic gradient. For example, polyp expansion is a routine activity necessary to capture prey, expand the surface area for photon capture by symbionts, and irrigate the gastrovascular cavity, and this activity increases the surface area of the tissue over which resources can transfer to and from the environment (Shashar et al. 1993). Expansion alleviates mass transfer limitations of the
43 flux of important metabolites (e.g., O2), thus aerobic respiration can increase (Patterson
1992). Conversely, polyp contraction limits the tissue surface area over which resources can be transferred, thus enhancing mass transfer limitation and causing aerobic respiration to be depressed.
To compare the aerobic respiration of coral with polyps expanded versus polyps retracted, aerobic respiration was measured twice for each nubbin. The first time, nubbins of P. verrucosa and A. pulchra were placed in the respiration chamber and allowed to incubate for ~30-60 minutes until full polyp expansion was observed. Full polyp expansion was observed under a dissecting scope and consisted of seeing polyps elevated above the orals disk forming a “cone-like mouth” (Lewis and Price 1975). Respiration rate of corals with expanded polyps was measured as described above. The nubbins were placed back into the same treatment tanks for 10-15 days to re-acclimate to the treatment tank conditions. After the 10-15 days, the same nubbins were placed back in the respiration chamber for the second time and allowed to incubate for ~30 minutes. The surface area of the nubbins was gently disturbed with a pipet tip to promote polyp retraction. Once ~80% of the polyps were retracted, aerobic respiration was measured again, and this rate was used to characterize routine activities of P. verrucosa and A. pulchra. Nubbins were continuously observed, during dark respiration, to ensure that polyps remained retracted.
Digestion is another metabolic activity that can cause fluctuations in aerobic respiration through protein synthesis, degradation, and turnover (Jobling 1983; Secor
44 2009). To measure aerobic respiration during digestion, nubbins of P. verrucosa and A. pulchra were allowed to feed on live Artemia sp. (brine shrimp), and oxygen uptake was recorded ~1-2 hours after the corals had captured brine shrimp. To promote feeding, the nubbins were incubated in a 240 mL chamber containing 24-48hr-old brine shrimp (~ 20 ml-1) for 1-1.5 h. After 1 h, nubbins were examined under a dissecting microscope for feeding behavior. The assumption of feeding behavior was observed as polyps extended, tentacles on the polyps were open, and slight mucus production from the polyps (Lewis and Price 1975). Nubbins were removed from the chamber containing brine shrimp and placed in a respiration chamber with fresh seawater. At this stage, corals were assumed to have fed upon Artemia shrimp such that a portion (i.e. ~50%) of the polyps were actively engaged in digestion. Based on visual observation, ~40-50% of polyps were engaged in feeding behavior within 90 minutes of exposure to brine shrimps. Nubbins were rinsed and placed in fresh seawater and incubated in the respiration chamber for ~30 minutes, under the same conditions mentioned above, and dark respiration measured.
Quantifying Biomass
Aerobic respiration was normalized to animal biomass to facilitate comparison of metabolic rate among corals and treatments, and also to allow comparison of empirical respiration with values published in the literature in which biomass is a frequent means of normalization of physiological rates in reef corals. Biomass was quantified by fixing P. verrucosa and A. pulchra nubbins in 10% formaldehyde in seawater for 24-48hr then decalcifying in 5% HCl for 48hrs (Davies 1980). The tissue was then dried at 60 ºC (to a constant weight) for 24-48hrs and weighed (g). Biomass was also used to explore the
45 relationship between body mass (i.e. biomass) and metabolic rate, and to evaluate how this relationship changes with metabolic activity.
Statistical Analysis
For physical parameters of the experiment, temperature and light were compared between tanks using an analysis of variance (ANOVA). To test the effect of different 2,4- dinitriphenol concentrations on aerobic respiration of P. verrucosa and A. pulchra, an
ANOVA was used. To examine the effect of the ethanol carrier on aerobic respiration, a t-test was used. To test for differences in aerobic respiration and Symbiodinium density as a function of duration in darkness, an ANOVA was used. Differences in aerobic respiration (standardized to surface area) as a function of metabolic treatment was analyzed using ANOVA.
Analysis of covariance (ANCOVA) was used to test for the effects of biomass and metabolic treatment on scleractinian aerobic respiration. Separate ANCOVAs were conducted for each species (P. verrucosa and A. pulchra) with metabolic rate as the dependent variable, and biomass was the covariate, and treatment (categorical) as the independent variables.
All data analyzed with ANOVA and ANCOVA. Log transformations were used, if necessary, to meet assumptions of normality and homogeneity of variances, that were tested through the graphical analysis of residuals. All analyses were conducted using the open source software R package ggplot ver. 3.5.1. (R Core Team 2018).
46
Results
Efficacy of Treatments
Mean temperature of the all four tanks were 29.16 ± 0.03 °C (n = 300).
Temperature was significantly different among tanks (F3,296 = 4.4, P = 0.004) (Table 1).
A Post hoc Tukey’s HSD test (α = 0.05) revealed that tank 3 (29.04 ± 0.03 °C) was lower in temperature than the other tanks. Out of the four tanks used, two tanks were illuminated to ~717 ± 11 μmol photons m-2 s-1 (n = 35) and two were dark with no illumination and a black tarp over the top to prevent light intrusion. Light intensity between the tanks illuminated with light were not significantly different (F1,52= 2.1, P =
0.15) (Table 1).
2,4-Dinitrophenol Titrations
Aerobic respiration of the corals was affected by 2,4-dinitrophenol, and it differed significantly among 2,4-DNP concentrations in Pocillopora verrucosa (F3,12 = 6.024, P =
0.009) (Fig. 1a) and Acropora pulchra (F3,12 = 17.72, P < 0.001) (Fig. 1b). The effect was strongly concentration dependent, with no effects at 0 M (i.e. control [seawater]) and the strongest effects at 1 x 10-4 M, with a reduced stimulative effect at 2.5 x 10-4 M.
Overall, 2,4-DNP at a final concentration of 1 x 10-4 M increased the mass-specific aerobic respiration of P. verrucosa by ~80%, and for A. pulchra the same concentration increased mass-specific aerobic respiration by ~800%. Aerobic respiration was not significantly affected by the 70% ethanol that was used as a carrier to dissolve the 2,4-
DNP (t = -1.7, df = 3, P = 0.2) (Fig. 4).
47 Symbiodinium density
Symbiodinium densities of freshly collected corals were 0.43 ± 0.05 x 106 cells cm-2 in P. verrucosa, and 1.04 ± 0.03 x 106 cells cm-2 in A. pulchra, and in both cases, densities quickly declined when corals were subject to darkness. Symbiodinium density was significantly affected by duration of exposure to darkness in P. verrucosa (F3,10 =
10.5, P = 0.002) and A. pulchra (F2,9 = 12.05, P = 0.003) (Fig. 4), and in both species they declined after 24 - 96 hours in darkness. After 96 h in darkness, mean Symbiodinium declined by 67% in P. verrucosa and 33% in A. pulchra.
Effects of activity on aerobic respiration
The aerobic respiration of the corals strongly differed among levels of activity of the coral holobiont, with lowest values at ~ 96 h in darkness for P. verrucosa (F4,13 =
5.76, P = 0.007) (Fig. 3a) and A. pulchra (F2,9 = 9.5, P = 0.006) (Fig. 3b). There were significant differences in aerobic respiration, standardized to surface area (cm2), when compared among all treatment for P. verrucosa (F4,60 = 9.3, P < 0.05) and A. pulchra
(F4,33 = 12.3, P < 0.05) (Fig. 6).
Metabolic Scaling
There was a strong dependence of aerobic respiration on the biomass of the coral, with double logarithmic plots of metabolic rate against biomass revealing scaling that sometimes was statistically indistinguishable from 1 (i.e., the relationships were isometric), and sometimes was distinct from 1 (i.e., the relationships were allometric)
(Fig. 5).
48
An ANCOVA indicated that the metabolism for P. verrucosa under the 2,4-DNP treatment was significantly higher than the metabolism of the corals in the dark treatment
(F1,25 = 57.2, P < 0.05) but metabolism in both treatments scaled the same way with respect to biomass (F1,24 = 0.7, P = 0.4) (Fig. 5a). The metabolism of fed corals was significantly higher compared to starved corals (F1,23 = 7.7, P = 0.01), but metabolism in both treatments scaled in the same way with respect to mass (F1,22 = 1.9, P = 0.2) (Fig.
5b). Metabolism increased more rapidly with biomass in corals with polyps expanded compared to corals with polyps contracted (F1,20 = 8.2, P = 0.009) (Fig. 5c).
For A pulchra, metabolism increased more rapidly with biomass in corals under the 2,4-DNP treatment compared to corals in the dark treatment (F1,19 = 21.4, P < 0.05)
(Fig. 5d). Additionally, metabolism increased more rapidly with biomass in corals that were fed compared to corals that were starved (F1,15 = 14.5, P = 0.002) (Fig. 5e). The metabolism of A. pulchra with polyps expanded was significantly higher than the metabolism of A. pulchra with polyps retracted (F1,17 = 52.3, P < 0.05), but respiration in both treatments scaled in the same way (i.e., with similar scaling exponents, b) with respect to mass (F1,16 = 0.005, P = 0.9) (Fig. 5f).
Discussion
While the effects of light, temperature, and morphology on metabolism have been studied extensively in scleractinians through measurements of aerobic respiration (Porter et al. 1984; Coles and Jokiel 1977; Jokiel and Cole 1990; Haryanti and Hidaka 2015;
49 Bruno and Edmunds 1998; Hoogenboom et al. 2008), comparing these measurements within coral species is difficult due to differing metabolic states (i.e. degree of activity).
The present study examines the metabolic plasticity of Pocillopora verrucosa and
Acropora pulchra, by: (1) characterizing the metabolic gradient using metabolic rates established by differing energy demands, and (2) using differences in coral size to explore size dependent metabolism under varying metabolic activities. In this study, I demonstrate that some scleractianian corals are capable of metabolic plasticity and that the metabolic scope differs between species.
Metabolic Gradient
I characterized a metabolic gradient by first defining the metabolic scope [MMR-
BMR] for P. verrucosa and A. pulchra. I then compared the MMR and BMR to metabolic rates simulated through the routine activities of polyp expansion, polyp contraction, and digestion. Overall, I found that metabolic rate (aerobic respiration) can vary inter- and intra-specifically for scleractinian corals wherein both A. pulchra and P. verrucosa have the capacity to exhibit a range of metabolic rates when exposed to different metabolic treatments (Table 2). The maximum metabolic rate (MMR), elicited from a metabolic uncoupler, revealed the greatest capacity for mechanical and chemical
-2 -1 -2 - work in both A. pulchra (2.42 μmol O2 cm hr ) and P. verrucosa (1.04 μmol O2 cm hr
1). The metabolic uncoupler, 2,4-dinitrophenol, has been used in several experiments to examine maximum power production of invertebrates, which leads to increases in aerobic respiration (Immers and Runnstrom 1960). Understanding the maximum metabolic capabilities of scleractinians, provides a foundation for projecting maximum metabolic
50 power output (Watts) under different regimes of anthropogenic stressors (i.e. thermal stress).
Metabolic rates for routine activities were simulated by the routine activities of polyp expansion, digestion, and polyp retraction. The metabolic rates during these activities did not differ from each other among replicate nubbins, in terms of work (J/hr).
The metabolic rate upon exposure to 2,4-DNP was significantly higher than the metabolic rate of the feeding treatment for both species. This was interesting considering that one study related the maximum ingestion of food to elicit a maximum metabolic rate in the stony corals, Porites porites, Manicina areolata, and Montestrea cavernosa due to their capacity to ingest several times as many calories as they lose through respiration (Coles
1969). Additionally, because polyp expansion can alleviate mass transfer limitation in corals (Patterson 1992), and polyp retraction typically increases mass transfer limitation
(Patterson 1992), I expected the capacity for work to be significantly different between polyp expansion and polyp retraction. In contrast to this expectation, the metabolic rates of expansion and retraction were not statistically different from each other in either species.
The minimum capacity for work was simulated in the dark treatment, which
-2 elicited the lowest (i.e., basal metabolic rate (BMR)) for A. pulchra (0.69 μmol O2 cm
-1 -2 -1 hr ) and P. verrucosa (0.44 μmol O2 cm hr ). Simulating the minimum capacity for work is beneficial for studying coral physiology under metabolic depression (MD)
(Guppy and Withers 1999; Jacobson et al. 2016). When resources (i.e. light or
51 heterotrophic resources) are limited, some scleractinians have the capacity to depress their area-normalized respiration rates (Jacobson et al. 2016). Starvation, caused by resource limitation, may shift energy allocation into primarily maintenance (instead of structural or reproductive), which can have implications for the coral success if duration of starvation increases (Jacobson et al. 2016).
Metabolic Scope
Metabolic scope can be an indicator of organismic success when measuring metabolic rate because it can be linked to whole-animal performance and fitness by examining an individual’s ability to perform aerobic activities (Pörtner and Knust 2007;
Donelson et al. 2012). However, the extent to which metabolic scope matters is context specific and can strongly dependent on environmental conditions the organism experiences. The present results indicate that the metabolic scope for A. pulchra is 1.3
[251% increase from BMR to MMR] and 0.6 for P. verrucosa [136% increase from BMR to MMR]. For corals who live under varying environmental conditions, a lower metabolic scope may indicate limitations on an individual’s ability to mitigate potential anthropogenic stressors (Pörtner and Knust 2007). Understanding how corals mitigate the effects of stressors (i.e. high temperature, salinity, pH) through metabolism may be critical in projecting future performance and fitness of coral and coral reefs (Pörtner and
Farrell 2008).
Since the metabolic scope has never been quantified in scleractinian corals, the metabolic scope (MMR/BMR) of A. pulchra and P. verrucosa was compared to other
52 taxa in order to determine if the ability to upregulate metabolism to meet increased energy demands was the similar across taxa. The metabolic scope of P. verrucosa and A. pulchra (2-3) reveals a comparable ability to upregulate metabolism when compared to other marine invertebrates (i.e. barnacles and periwinkles) (2-4), but this ability is modest when compared to terrestrial invertebrates (i.e. fruit fly and butterfly) (13-170), and non- invertebrate taxa (i.e. mammals, birds, reptiles, and fish) (8-20) (Table 3). The difference in metabolic scope between these taxa are consistent with lifestyle (i.e. activity level), considering that marine invertebrates, like barnacles and periwinkles, tend to be sessile or slow moving, live in cooler aquatic habitats then their terrestrial counterparts within in the same geographic area, and typically exhibit lower internal temperatures during activities than terrestrial invertebrates or vertebrates in the same geographic location
(Willmer et al. 2000). Active invertebrates (i.e. insects and flatworms) and vertebrates
(i.e. mammals and reptiles) have the ability to increase their metabolic rate through movement (i.e. running or flying), which can widen the range of metabolism associated with transitioning from resting to maximum activity (Willmer et al. 2000).
After defining the metabolic scope, energy output as a function of time (joules/s) can be calculated to determine the duration that a coral can persist (i.e. lifespan) under metabolic rates based on the size of the energy reserves and the capacity to replace them
(Peterson et al 2010). Since scleractinians have the capacity to exhibit metabolic plasticity, and their respiration rate is varying continuously based on the metabolic activity (i.e. oxygen uptake and energy released), it is critical to also calculate the duration a coral can sustain metabolic rates. Duration of sustainable metabolic rate is key
53 information for defining the energy production capacity associated with aerobic respiration and how sustainable metabolic rate can fluctuates over the course of a year, under environmental conditions. Knowing the capacity of aerobic respiration to support the energy demands of varying levels of muscular and chemical work can indicate how parameter values within the energy budget or flux of energy can be altered for that particular metabolic activity. The energy budget can describe the energy used for metabolism (R = C - (F + U + P)). Where R is respiratory loss, C is consumption, F is faecal loss, U is excretion loss, and P is production or synthesis. Using the energy budget equation can give insight into how much energy the coral consumes (i.e. heterotrophic and autotrophic resources) versus how much energy the coral outputs (i.e. Joules), which can be utilized to scale coral community metabolism (Crossland et al. 1991).
Additionally, there is still a need to examine how this changes over the course of a year, with changing coral biomass, symbiont density, or type of symbiont.
Metabolic Scaling
Metabolic scaling has been studied extensively in unitary organisms (Muñoz and
Cacino 1989; Reich et al 2006; Glazier 2010), but less attention has been paid to scaling in marine organisms with a colonial modular (CM) design (Hughes and Hughes 1986;
Edmunds and Burgess 2016; Burgess et al., 2017;). This lack of attention could indicate the widely accepted idea that all organisms scale to the power of ¾, or Kleiber’s Law, which states that as the log metabolic rate increases by 4-fold, log metabolic rate increases by 3-fold (Kleiber 1932). However, there have been deviations from the ¾ scaling exponent for colonial modular organisms (Burgess et al. 2017). Colonial modular
54 organisms, like corals, were once considered to be the exception to the ¾ power law due to the addition of identical modules that make up a colony. Individual modules are subject to allometric constraints (i.e. b = 0.75) due to surface area and volume constraints, however, when identical modules are added to the colony, biomass will increase without increasing the mass-specific metabolic rate of the colony causing the whole colony to scale isometrically (i.e. b = 1) (Hughes 2005; Burgess et al. 2017). This does not consider the developmental and functional difference among modules. A recent review of metabolic scaling in CM designs found that the slope of the scaling relationship
(b) of metabolism on body mass is more variable in modular organisms (0.66 – 0.93) then unitary organisms (0.62 – 0.69) (Burgess et al. 2017).
Variation in the scaling exponent (b) for aerobic respiration was much greater than expected based on Kleiber’s Law (¾) and ranged from 0.29 to 1.07. Additionally, there were significant differences in the metabolic scaling exponent when comparing treatments (i.e. 2,4-DNP and starvation; polyp expansion and retraction; digestion and starvation) and this differed among species. For A. pulchra, b was higher in the 2,4-DNP treatment (b = 1.07) then in the starvation treatment (b = 0.29), and b was also greater in the feeding treatment (b = 0.82) than in the starved treatment (b = 0.29). However, b did not differ between the polyp expansion treatment (b = 0.48) and polyp retraction treatments (b = 0.47), but metabolism was greater in corals with expanded polyps. For P. verrucosa, the significant differences in b were opposite from A. pulchra. Metabolism was greater for the 2,4-DNP treatment compared to the starvation treatment and greater for the polyp expansion treatment than for the polyp retraction treatments, with no
55 difference in b. However, b was higher in the polyp expansion treatment (b = 0.93) then the polyp retraction treatments (b = 0.5).
Detection of high variation in the scaling exponent (b) in the present study in part underscores the challenges of measuring metabolic rate in colonial modular organisms.
This is because while the scaling exponent of metabolism (b) classically should address the size-dependence of basal metabolic rate (BMR) (Glazier 2010), it is rare for the role of activity in determining metabolic rates to be considered in corals, or other taxa. This variation in the scaling exponents has bearing on one theory involving metabolic scaling: the metabolic level boundary (MLB) hypothesis (Glazier 2005). The MLB hypothesis has been posited to account for the variation in b (i.e. 0.67-1) based on organismic activity.
The organismic activity (L) determines what is most influential on the scaling exponent
(i.e. volume or surface area) (Glazier 2010). Surface-area diffusion of metabolites and waste products are more dependent on metabolic scaling when the organism is at a high metabolic activity (Glazier 2010). Volume-related demands associated with the organism
(i.e. maintenance) are more influential when the organism is at low metabolic activity
(Glazier 2010). In unitary organisms, it is common for the scaling exponent (b) to be < 1
(i.e., ~ 0.75) when resting metabolic rates (RMR) are measured as a function of biomass, whereas b is more likely to equal one when maximum metabolic rates (i.e., MMR) are measured as a function of biomass (Glazier 2010). According to the MLB hypothesis, the scaling exponent (b) is biologically linked to the metabolic level (i.e. overall activity [L]) of the scaling relationship (Glazier 2010).
56 Other than the role of overall metabolic activity (L) driving departures of b for metabolism from expected values of 0.67-1.00 (Glazier 2010), there are several additional features that could drive further departures. For example, behavior of the coral polyps could drive variation in b. Colonial modular organisms (i.e., polyps in corals) produce colonies consisting of iterated modules that can function independently but are typically connected physically and physiologically (Burgess et al. 2017). If the polyps that make up a colony are functionally similar and comparably sized, the whole colony metabolic rate is expected to scale isometrically (b =1) (Hughes and Hughes 1986;
Burgess et al. 2017). However, coral polyps that make up a colony are not all functionally, physiologically, or geometrically equal. Some polyps may vary in nutritional state, reproductive state, or age which may cause differential resource demands (Harvell 1994). For example, Dudgeon et al. (1999) found that not non-feeding polyps on the outer edge of the colony of the colonial hydrozoan, Podocoryne carnea, were transferred food from feeding polyps through a connected network within the colony. Because these non-feeding polyps are not using muscle-driven flow to capture and transport food resources and are provided with food, their energy expenditure is less than their feeding counterparts (Dudgeon et al. 1999). Behavioral differences among polyps in a single colony may cause the whole colony metabolic rate to scale allometrically (b ≠ 1). These differences (i.e. functional, physiological, geometric) make it difficult to assess metabolic scaling in colonial modular organisms because resources acquisition and use may not be equal in all polyps (Dudgeon et al. 1999; Burgess et al.
2017).
57 Understanding how corals respond physiologically to anthropogenic stressors, such as temperature and ocean acidification, aid in the capacity to make projects of long- term survival of corals and the reefs they build. A critical part of this effort is obtaining ecologically relevant measurements of metabolism that accurately reflect the in situ daily energy requirement of corals. Metabolism or the “fire of life” is the driving force of chemical reactions in the body (Kleiber 1961). Metabolic plasticity is a critical part in an organism’s ability to acclimate to changes in physical or chemical environments.
Understanding the metabolic scope is also key in assessing organismic performance and success under future anthropogenic stressors. A. pulchra and P. verrucosa produce a range of metabolic rates that reflect differential energy demands.
Additionally, measuring the metabolic scope of scleractinian corals can help determine the duration that coral can persist under certain metabolic rates based on the size of the energy reserves and the ability to replace them (Peterson et al 2010). This can be assessed using an energy budget that uses daily inputs and outputs and how differing metabolic rates can affect those inputs and output (i.e. Joules). The metabolic scaling exponent (b) of A. pulchra and P. verrucosa varied between different treatments in the present study, and also differed from the values proposed, under BMR and MMR, in the
MLB hypothesis. While I cannot rule out other explanations for variation in the metabolic scaling exponent, behavior-mediated differences among the polyps provides a compelling basis under which we can further explore metabolic scaling in coral. Furthermore, a larger range of sizes will be critical to explore the scaling exponents reported in the present study and more research is necessary to explore the MLB hypothesis (sensu
Glazier 2010) in colonial modular organisms.
58 Figures
Trial 1 (Jan 2018) Trial 2 (Apr-May 2018)
Temp erature 29.14 ± 0.02 29.13 ± 0.02
Photon Flux Desnity 717 ± 11 730 ± 4 Table 1. Conditions recorded three times a day for 28 days in temperature (oC) -2 -1 and photon flux density (μmol photons m s ). Values shown are mean ± SE.
59
b c A) B) 1.2
4
e
t
e
t a
a 3
1.0 R
R
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c
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M M a a 1 0.6 0 5x10−5 1x10−4 2.5x10−4 0 5x10−5 1x10−4 2.5x10−4 Concentration (M) Concentration (M) -2 -1 Figure 1. Mean metabolic rate (μmol O2 cm hr ) of A) Pocillopora verrucosa and B) Acropora pulchra nubbins as a function of 2,4-Dinitrophenol concentration. Mean ± SE shown (n = 4 for all treatments).
60
Pocillopora verrucosa A) B) Acropora pulchra a a 1.1 ab
0.6 abc 1.0
e
e
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a 0.9 ab
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0.5 0hr 24hr 48hr 72hr 96hr 0hr 24hr 96hr Time Time a a D)
5e+05 C)
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24hr 48hr 72hr 96hr 0hr 24hr 96hr Time Hours -2 -1 Figure 2. A-B) Mean metabolic rate (μmol O2 cm hr ) and C-D) Symbiodinium density (cm-2) of P. verrucosa and A. pulchra as a function of duration in darkness (0, 24, 48, 72, 96 h). Mean ± SE shown (n = 4 for all treatments except for 96 h [n = 2 for P. verrucosa]).
61 ab A) a ab
1.00 ab e
t ab
a
R
c
i ab ab
l 0.75
o
b
a t
e b M 0.50
8AM 11AM 2PM 5PM 8PM 11PM 2AM 5AM Time
2.5 B)
e 2.0
t
a
R
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i l
o 1.5
b
a
t
e M 1.0
0.5 8AM 5PM Time -2 -1 Figure 3. Mean metabolic rate (μmol O2 cm hr ) of A) P. verrucosa every 3 hours over 24 hours and B) A. pulchra at 8am and 5pm. Mean ± SE shown (n = 4 for all treatments).
62
6
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. 0 Ethanol No Ethanol
Treatment
-2 -1 Figure 4. Mean metabolic rate (μmol O2 cm hr ) of P. verrucosa dosed with 500 uL of 70% ethanol and no ethanol (n = 4 for all treatments).
63
Pocillopora verrucosa Acropora pulchra 2,4-DNP: y = 1.0457x + 1.5768; R² = 0.78 Starved: y = 0.8492x + 1.2186; R² = 0.70
a) b)
Fed: y = 0.5403x + 1.2007; R² = 0.69 Fed: y = 0.8228x + 1.5025; R² = 0.85 Starved: y = 0.8491x + 1.2186; R² = 0.70 Starved: y = 0.2978x + 0.5702; R² = 0.75
c) d)
Expanded: y = 0.9339x + 1.5123; R² = 0.84 Expanded: y = 0.4859x + 1.4504; R² = 0.85 Contracted: y = 0.4144x + 1.1162; R² = 0.5 Contracted: y = 0.4773x + 1.1108; R² = 0.73 e)
f)
-1 Figure 5. Log metabolic rate (μmol O2 h ) as a function of log biomass (g) of P. verrucosa and A. pulchra under metabolic treatments: (a-b) 2,4-DNP and starved [Dark]; (e-f) polyps contracted and polyps expanded; (c-d) fed and starved.
64
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. 0 2,4−DNP Expansion SDA Retraction Dark
Treatments
-2 -1 Figure 6. Mean metabolic rate (μmol O2 cm hr ) of A) P. verrucosa and B) A. pulchra under different metabolic treatments (metabolic uncouper [2,4-DNP], polyp expansion, specific dynamic action [SDA], polyp retraction, and darkness) (n = 14 [P. verrucosa] and n = 8 [A. pulchra]).
65
-2 -1 Table 2. Mean metabolic rate (μmol O2 cm hr ) of two species of coral, Acropora pulchra and Pocillopora verrucosa, under differing metabolic treatments.
66
Factorial Species Aerobic Scope Percent Increase
Marine Invertebrates Barnacle 4 314 Periwinkle 2 100 Scleractinian Coral A. pulchra 3 251 P. damicornis 2 136
Fish, Reptiles Salmon 8 650 Iguana 17 1,680 Turtle 20 2,033
Birds, Mammals Hummingbird 15 1,400 Dog 12 1,112 Human 14 1,291
Table 3. Factorial aerobic scope (MMR/BMR) and percent increase (([MMR- BMR]/BMR)*100) (Aerobic scope values taken from Willmer et al. 2005).
67 CHAPTER 4
Concluding Remarks
Coral reefs are one of the most diverse ecosystems in the world due to the abundance and range of diverse fauna that make up a reef (Connell 1997; Glynn and
Enochs 2011). Corals are critical ecosystem engineers that provide the reef with resources, such as complex three-dimensional habitat structures (Jones et al. 1994). Many coral reefs are being impacted by anthropogenic stressors, from rising sea-surface temperatures (Hoegh-Guldberg et al. 2007) to eutrophication (Fabricius 2005), to hurricanes (Harmelin-Vivien 1994). In areas like the Great Barrier Reef, Florida Keys, and the Caribbean, coral cover has declined significantly over the last few decades, with some areas declining up to >90% (Jackson et al. 2014; Hughes et al. 2018). Drivers of coral cover decline in the Caribbean include disturbances such as the Diadema antillarum die-off (Levitan 1988; Edmunds and Carpenter 2001), severe tropical storms (Woodley et al. 1981; Rogers et al. 1991), and disease outbreaks (Aronson and Precht 2001; Miller et al. 2009). These disturbances can cause shifts in the community composition of a reef
(i.e. phase shifts), and with increasing severity and duration of disturbance events, communities may shift from coral dominated to communities dominated by other taxa
(Hughes 1994; Stobart et al. 2005; Ward-Paige et al. 2005; Bell et al. 2013). As natural disturbances continue to increase in intensity, due to climate change, it is important to know how coral reef communities are being impacted both ecologically and physiologically.
68 Much like corals, sponges play important functional roles on coral reefs extending from binding substratum together, filtering water, and providing critical habitat structure
(Wulff 1984; McMurray et al. 2014; Bell 2008). These roles can be particularly important after natural disturbances events, such as hurricanes. Hurricanes have been known to fragment, overturn, or smother coral colonies with sediment, causing high mortality
(Edmunds and Witman 1991; Rogers et al. 1991) with little chance of recovery.
However, sponges have the capacity to rapidly repair damaged tissue, recruit and colonize benthic surfaces quickly, and potentially use suspended sedimentation as a food source (Reiswig 1971; Wulff 1994; Duckworth et al. 2006). However, some sponge species are still at risk of mortality through storm effects based on their size and morphology (i.e. wide and narrow) (Woodley et al 1989; Wulff 1994).
In chapter 2, I assess temporally-mediated change in density of two Caribbean vase sponges in St. John, US Virgin Islands, likely driven by hurricane induced physical drivers, to determine if size and morphology mitigate storm effects. I also assessed how differences in morphology alter the invertebrate associations between the sponges and determine how this was changed by hurricane effects. My study demonstrates that sponge density did not change over time, even following major disturbance events (i.e.
Hurricanes Irma and Maria), but sponge size did differ. The present study also demonstrated that N. digitalis contained more invertebrate epifauna (per cm) then I. campana,and the abundance of this invertebrate epifauna did not change over time.
69 In chapter 3, I examined the metabolic plasticity of two scleractinian corals by assessing the difference in metabolic rate under differing metabolic treatments elicited differential energy demands. I then used nubbins differing in size to explore size- dependent metabolism and how it varies based on metabolic activity. Exposing A. pulchra and P. verrucosa to a suite of metabolic treatments caused significant differences in metabolic rates and power output. Metabolic scope can be an indicator of metabolic fitness when measuring metabolic rate, because it can be linked to whole-animal performance and fitness by examining an individual’s ability to perform aerobic activities
(i.e. buildup and break down of proteins) (Pörtner and Knust 2007; Donelson et al. 2012).
Metabolic scaling differed between several treatments for A. pulchra (2,4-DNP versus dark and digestion versus starvation) and P. verrucosa (expanded versus contracted), and several of the metabolic scaling exponent (b) deviated from the exponents predicted by the MLB hypothesis (Glazier 2010). A possible reason for deviations in the scaling exponent (b), from the exponents proposed by MLB hypothesis (b = 0.67 - 1) could be due to differences (i.e. physiological, developmental, geometric) among the polyps.
Theoretically, if the polyps are physically and physiologically similar, whole colony metabolic rate is expected to scale isometrically (b =1) (Hughes and Hughes 1986;
Burgess et al. 2017). However, coral polyps that make up a colony are not all functionally or physiologically equal. Some polyps may vary in their nutritional state, reproductive state, or age, which may cause differential resource demands (Harvell 1994; Dudgeon et al. 1999; Burgess et al. 2017), so whole colony metabolic rate is likely to scale allometically (b ≠ 1).
70 As anthropogenic stressors continue to increase in intensity and duration, coral reefs may continue to change. Scleractinian corals have been declining in abundance on many coral reefs over the last few decades (Gardner et al 2003; De’ath et al. 2012;
McClenachan et al. 2017) and are projected to experience continuing declines as anthropogenic stressors increase. Other non-scleractinian invertebrate fauna may benefit from decreases in overall coral cover, but can also be negatively impacted by stressors.
Together, these studies highlight the importance of ecological and physiological studies on scleractinian and non-scleractinian invertebrates present on coral reefs and the impacts of anthropogenic stressors on these invertebrates. The increase in sponge density post hurricane, in St. John US Virgin Islands, shows resistance of certain sponge populations to natural disturbances (i.e. hurricanes), which may cause shifts in coral reef community structure and composition as coral cover continues to decline. Additionally, determining the range of metabolic rates and metabolic scope of scleractinian corals in Mo’orea,
French Polynesia can give insight into organismic fitness of individual corals on the reef.
Finally, the findings from both studies highlight that multiple organisms within a coral reef community should be studied in order to understand the broader impacts of climate change on coral reefs.
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