|Scientific Name:||Pocillopora elegans|
|Species Authority:||Dana 1846|
|Red List Category & Criteria:||Vulnerable A4ce ver 3.1|
|Assessor/s:||Hoeksema, B., Rogers, A. & Quibilan, M.|
|Reviewer/s:||Livingstone, S., Polidoro, B. & Smith, J. (Global Marine Species Assessment)|
This species is widespread and locally common throughout its range. However, it is particularly susceptible to bleaching, disease, crown-of-thorns starfish predation, and extensive reduction of coral reef habitat due to a combination of threats. Specific population trends are unknown but population reduction can be inferred from declines in habitat quality based on the combined estimates of both destroyed reefs and reefs at the critical stage of degradation within its range (Wilkinson 2004). Its threat susceptibility increases the likelihood of being lost within one generation in the future from reefs at a critical stage. Therefore, the estimated habitat degradation and loss of 35% over three generation lengths (15 years) is the best inference of population reduction and meets the threshold for Vulnerable under Criterion A4ce. It will be important to reassess this species in 10 years time because of predicted threats from climate change and ocean acidification.
|Range Description:||In the Indo-West Pacific, this species is found in the Central Indo-Pacific, the Oceanic West Pacific, the Central Pacific, Solomon Islands (Veron and Turak 2006) and Papua New Guinea (Fenner 2003) and the Eastern Pacific.
In the Eastern Tropical Pacific region, this coral species is present in: Mexico: Baja California Sur, Nayarit, Jalisco, Colima, Guerreo and Oaxaca (Reyes-Bonilla et al. 2005, Pérez Vivar et al. 2006, Glynn and Ault 2000); El Salvador* (Reyes-Bonilla and Barraza 2003); Costa Rica: Murcielago Archipelago, Culebra Bay, Peninsula de Osa, Caño Island, and Cocos Island (Cortés and Guzmán 1998); Panama: throughout the Gulfs of Chiriqui and Panama (Holst and Guzmán 1993, Maté 2003); Colombia: Malpelo Island, Gorgona Island, Ensenada de Utría (Zapata and Vargas-Ángel 2003); Ecuador: Salango Island, Los Frailes, Sucre Island and La Plata Island, and throughout the Galápagos Archipelago (except for Fernandina and the west side of Isabela) (Glynn et al. 2001, Glynn 2003, Hickman et al. 2005).
(*Taxonomic determination requires confirmation (Reyes-Bonilla and Barraza 2003).)
Native:American Samoa (American Samoa); Australia; Colombia; Cook Islands; Costa Rica; Ecuador; El Salvador; Fiji; French Polynesia; Guadeloupe; Honduras; Indonesia; Japan; Kiribati; Malaysia; Mexico; Micronesia, Federated States of ; Nicaragua; Niue; Northern Mariana Islands; Palau; Panama; Papua New Guinea; Philippines; Samoa; Singapore; Solomon Islands; Thailand; Tokelau; Tonga; Tuvalu; United States Minor Outlying Islands; Wallis and Futuna
|FAO Marine Fishing Areas:||
Indian Ocean – eastern; Pacific – eastern central; Pacific – northwest; Pacific – southeast; Pacific – western central
|Range Map:||Click here to open the map viewer and explore range.|
This species is locally common in some regions of the central Indo-Pacific and the far eastern Pacific.
The relative abundance of Pocillopora elegans in the Eastern Tropical Pacific region has been categorized as follows:
Abundant: Mexico, Panama and Colombia (Glynn and Ault 2000).
Common: Gulf of California, Revillagigedo Islands in Mexico, and Ecuador including the Galápagos Archipelago (Glynn and Ault 2000, Glynn 2003).
Uncommon: Malpelo Island, Colombia (Glynn and Ault 2000).
Rare: Costa Rica including Cocos Islands (Glynn and Ault 2000).
Nevertheless, according to Guzmán and Cortes (1989), P. elegans is likely to be abundant in mainland Costa Rica but rare in Cocos Island, and common in Panama (Guzmán et al., in prep.).
Glynn et al. (1988) report high rates of pocilloporid coral mortality across the eastern Pacific following the 1982/83 El Niño, ranging from 51% at Caño Island to 76-85% in Panama and 97-100% in the Galápagos Islands (Glynn et al. 1988). However, following this high coral mortality, numerous pocilloporid recruits have been observed at some sites in Costa Rica, Panama and the Galapagos Islands (Glynn unpublished data in Glynn et al. 1991). According to Guzmán (pers. comm.), Pocillopora elegans populations are probably increasing in Costa Rica and Panama following the severe coral mortality of recent ENSO events (1982-83 and 1997-98) and red tides (Guzmán and Cortes 2001). Population trends of P. elegans in the Galápagos Islands are unknown (Edgar and Chiriboga pers. comm.)
In the Galápagos Islands, pocilloporid communities were well developed off northeastern San Cristobal, Espanola and Floreana Island until the 1980s (Glynn 1994, 2003), but disappeared following the 1982-83 ENSO event, with minimal coral recruitment since (Glynn 2003).
In mainland Ecuador at Sucre Island, Machalilla, the predominant frame-building species before 1983 were Pocillopora elegans and Pocillopora damicornis, with the two species generating a rigid framework that covered over 1 ha of bottom (Glynn 2003). However, by 1991 the reef had declined to dispersed colonies of P. damicornis and P. elegans (Glynn 2003).
There is evidence that overall coral reef habitat has declined, and this is used as a proxy for population decline for this species. This species is particularly susceptible to bleaching, disease, and other threats and therefore population decline is based on both the percentage of destroyed reefs and critical reefs that are likely to be destroyed within 20 years (Wilkinson 2004). We assume that most, if not all, mature individuals will be removed from a destroyed reef and that on average, the number of individuals on reefs are equal across its range and proportional to the percentage destroyed reefs. Reef losses throughout the species' range have been estimated over three generations, two in the past and one projected into the future.
The age of first maturity of most reef building corals is typically three to eight years (Wallace 1999) and therefore we assume that average age of mature individuals is greater than eight years. Furthermore, based on average sizes and growth rates, we assume that average generation length is 10 years, unless otherwise stated. Total longevity is not known, but likely to be more than ten years. Therefore any population decline rates for the Red List assessment are measured over at least 30 years. Follow the link below for further details on population decline and generation length estimates.
|Habitat and Ecology:||
This species occurs in shallow reef environments. Pocillopora elegans occurs in all shallow water habitats on coral reefs and coral communities on rocky substrata (Guzmán et al. pers. comm.), to at least 20 m depth, but is most common between 1-10 m depth (Guzmán and Chiriboga pers. comm.). Maximum size is 25 cm.
In the Eastern Tropical Pacific region, Pocillopora elegans is one of the major reef building species (along with P. damicornis), forming intermeshing compact frameworks that can attain 2-3 m in relief (Guzmán and Cortes 1989, Glynn 2001).
Amongst the reef building corals in the Eastern Tropical Pacific region, pocilloporid species have the fastest growth rates (Guzmán and Cortés 1993). Reported growth rates of Pocillopora elegans vary little between locations in the Eastern Tropical Pacific, from 3.1 cm per year in Galápagos to 3.32 cm per year in Costa Rica (Guzmán and Cortés 1993).
P. elegans is a broadcast spawner (Glynn et al. 1991), with the capacity to function as a simultaneous hermaphrodite (Glynn et al. 1991). According to Glynn et al. (1991), larval settlement in the Galápagos Islands presumably has been the predominant mode of recruitment, and the only observed form of recruitment in areas that experienced high mortality (97-100%) in 1983. Asexual reproduction by fragmentation has been reported as an important mechanism for reef recovery in Panama (Glynn et al. 1991). P. elegans, like other pocilloporid species in the eastern Pacific, has low rates of recruitment (Glynn et al. 1991). Histological evidence indicates that spawning is likely to occur during a few days around the new moon (Glynn et al. 1991). Reproductive activity in the eastern Pacific its related to local thermal regimes, with a generally higher incidence of gravid corals at sites with stable, warm water conditions, or during warming periods in areas that experience significant seasonal variation (Glynn et al. 1991). Glynn et al. (1991) conclude that moderate El Nino warming can stimulate gametogenesis in Galápagos pocilloporid corals.
Pocilloporid corals, presumably including P. elegans, are generally amongst the strongest coral competitors with relatively high rates of calcification (Glynn 2001). However, coral species exhibiting high rates of calcification usually have relatively high mortality rates (Glynn 2000). Pocilloporid corals also usually predominate at shallow depths (1-15 m). Amongst the reef building corals in the Eastern Tropical Pacific region, pocilloporid species have the highest growth rates (Guzmán and Cortes 1993). They are the principal framework builders on Panamanian reefs (Glynn 2002).
Pocillopora species are preyed on by at least nine groups of consumers. These vary in their consumption patterns, but include:
a) Species that bite off colony branch-tips: pufferfishes (Arothron), parrotfishes (Scaridae), filefishes (Monacanthidae) (Glynn 2002).
b) Species that scrape skeletal surface: hermit crabs (Trizopagurus, Aniculus, and Calcinus) (Glynn 2002).
c) Species that remove tissues but leave the skeleton intact: gastropods (Jenneria pustulata and Quoyula sp. (Glynn 2002)), buterflyfishes, angelfishes, damselfish (Stegastes acapulcoensis), and Acanthaster planci (Glynn 2002).
d) Species that abrade tissues and skeleton: Eucidaris galapagensis (Glynn 2001).
Jenneria and Acanthaster can kill whole, relatively large (approx. 30 cm in diameter) colonies of Pocillopora (Glynn 2002). Pocilloporid species can have crab (Trapezia sp.) and alpheid shrimp as mutualistic symbionts that protect the coral from the attack of the crown-of-thorns sea star A. planci (Glynn 2001).
Pocilloporid species as well as other major reef building corals within the Eastern Tropical Pacific region (Porites, Pavona, Gardinoseris) catastrophically declined in the Galápagos Archipelago and Cocos Island after 1983. Recovery observed since that time was in large part nullified by the 1997-98 ENSO event (Glynn 2000). According to Glynn et al.(1988), pocilloporid coral mortality in the eastern Pacific was high, ranging from 51% at Caño Island to 76-85% in Panama and 97-100% in the Galápagos Islands (Glynn et al. 1988).
Glynn (1994) suggests that the sea urchin Eucidaris galapagensis (syn. E. thouarsii) provides important biotic control of pocilloporid reef development. This urchin is the most persistent corallivore in the Galapagos Islands, where it is often observed grazing on pocilloporid corals (Glynn 2001).
Overfishing is probably responsible for some ecological imbalance on coral reefs that could prolong recovery from other disturbances (Glynn 2001). Moreover, Edgar et al.(unpublished manuscript) reported that over-exploitation of sea urchin predators (lobsters and fishes), along with ENSO, has a major effect in the condition and distribution of corals in the Galapagos Islands, by increasing the grazer and bioerosion pressure on corals.
Coral mortality associated with phytoplankton blooms has been reported from Caño Island, Costa Rica, and Uva Island, Panama, in 1985; where mortality of pocilloporid species (especially P. capitata and P. elegans) was in the order of 100% and 13% respectively at 3 m depth (Guzmán et al. 1990).
According to Glynn (2001), pocilloporid coral harvesting is an important threat in the Eastern Tropical Pacific region, specially along the continental coast. This activity has virtually eliminated pocilloporid corals from Acapulco (Mexico), Bahia Culebra (Costa Rica), Taboga Island (Panama), and parts of the coast of Ecuador (Glynn 2001). Nevertheless, this activity is now largely excluded from Costa Rica and Panama (Guzmán pers. comm.).
Bryant et al. (1998), based on four anthropogenic factors (coastal development; overexploitation and destructive fishing practice; inland pollution and erosion, and marine pollution), estimated a high threat to coral reefs along the coasts of Costa Rica, Panama and Colombia. High levels of siltation caused by accelerated coastal erosion have degraded coral reefs in Costa Rica, Colombia and Ecuador (Glynn 2001)
Other threats include: a) predation principally by Acanthaster and Jenneria (Glynn 2002, 1994, 2000), and b) harvesting for the curio trade, an activity that has virtually eliminated pocilloporid corals from Acapulco (Mexico), Bahia Culebra (Costa Rica), Taboga Island (Panama), and parts of the coast of Ecuador (Glynn 2001).
In general, the major threat to corals is global climate change, in particular, temperature extremes leading to bleaching and increased susceptibility to disease, increased severity of ENSO events and storms, and ocean acidification.
Coral disease has emerged as a serious threat to coral reefs worldwide and a major cause of reef deterioration (Weil et al. 2006). The numbers of diseases and coral species affected, as well as the distribution of diseases have all increased dramatically within the last decade (Porter et al. 2001, Green and Bruckner 2000, Sutherland et al. 2004, Weil 2004). Coral disease epizootics have resulted in significant losses of coral cover and were implicated in the dramatic decline of acroporids in the Florida Keys (Aronson and Precht 2001, Porter et al. 2001, Patterson et al. 2002). In the Indo-Pacific, disease is also on the rise with disease outbreaks recently reported from the Great Barrier Reef (Willis et al. 2004), Marshall Islands (Jacobson 2006) and the northwestern Hawaiian Islands (Aeby 2006). Increased coral disease levels on the GBR were correlated with increased ocean temperatures (Willis et al. 2007) supporting the prediction that disease levels will be increasing with higher sea surface temperatures. Escalating anthropogenic stressors combined with the threats associated with global climate change of increases in coral disease, frequency and duration of coral bleaching and ocean acidification place coral reefs in the Indo-Pacific at high risk of collapse.
Localized threats to corals include fisheries, human development (industry, settlement, tourism, and transportation), changes in native species dynamics (competitors, predators, pathogens and parasites), invasive species (competitors, predators, pathogens and parasites), dynamite fishing, chemical fishing, pollution from agriculture and industry, domestic pollution, sedimentation, and human recreation and tourism activities.
The severity of these combined threats to the global population of each individual species is not known.
All corals are listed on CITES Appendix II. Parts of the species’ range fall within Marine Protected Areas.
Recommended measures for conserving this species include research in taxonomy, population, abundance and trends, ecology and habitat status, threats and resilience to threats, restoration action; identification, establishment and management of new protected areas; expansion of protected areas; recovery management; and disease, pathogen and parasite management. Artificial propagation and techniques such as cryo-preservation of gametes may become important for conserving coral biodiversity.
|Citation:||Hoeksema, B., Rogers, A. & Quibilan, M. 2008. Pocillopora elegans. In: IUCN 2013. IUCN Red List of Threatened Species. Version 2013.2. <www.iucnredlist.org>. Downloaded on 16 April 2014.|
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