|Scientific Name:||Pocillopora damicornis|
|Species Authority:||(Linnaeus, 1758)|
|Red List Category & Criteria:||Least Concern ver 3.1|
|Assessor(s):||Hoeksema, B.W., Rogers, A. & Quibilan, M.C.|
|Reviewer(s):||Livingstone, S., Polidoro, B. & Smith, J.|
The most important known threat for this species is extensive reduction of coral reef habitat due to a combination of threats. Specific population trends are unknown but population reduction can be inferred from estimated habitat loss (Wilkinson 2004). It is widespread and common throughout its range and is very resistant to bleaching. Therefore, the estimated habitat loss of 20% from reefs already destroyed within its range is the best inference of population reduction since it may survive in coral reefs already at the critical stage of degradation (Wilkinson 2004). This inference of population reduction over three generation lengths (15 years) does not meet the threshold of a threat category and this species is Least Concern. However, because of predicted threats from climate change and ocean acidification it will be important to reassess this species in 10 years or sooner, particularly if the species is also observed to disappear from reefs currently at the critical stage of reef degradation.
|Range Description:||In the Indo-West Pacific, this species occurs in the Red Sea and the Gulf of Aden, the Southwest and Northwest Indian Ocean and the Arabian/Iranian Gulf, the Central Indian Ocean, the Central Indo-Pacific, Tropical and Sub-tropical Australia, Southern Japan and the South China Sea, the Oceanic west Pacific, the Central Pacific, the Hawaiian Islands and Johnston Atoll, the Far Eastern Pacific, and Easter Island (Glynn 2003, Glynn et al. 2007).
In the Eastern Tropical Pacific region, the species has been reported from: Mexico: Baja California Sur, Nayarit, Jalisco, Colima, Michoacán, Guerrero and Oaxaca (Reyes-Bonilla 1998, Reyes-Bonilla and Lopez-Perez 1998, Reyes-Bonilla 2003, Calderon-Aguilar 2005, Reyes-Bonilla et al. 2005, Glynn et al. 2007); El Salvador: Del Amor beach, Los Cóbanos (Reyes-Bonilla and Barraza 2003); Costa Rica: Islas Murcielago Archipelago, Bahia Culebra, Bahia Brasilito, Samara, Cabo Blanco, Punta Leona, Herradura, Manuel Antonio, Punta Uvita, Peninsula de Osa, Golfo Dulce, Caño Island, and Cocos Island (Cortés and Guzmán 1998, Glynn et al. 2007, Guzmán and Cortés 2007); Panama: throughout the Gulfs of Chiriqui and Panama (Maté 2003, Glynn 1997, Glynn et al. 2007); Colombia: Gorgona Island, Ensenada de Utría and Tebada (Zapata and Vargas-Ángel 2003, Glynn et al. 2007); Ecuador: Salango Island, Los Frailes, Sucre Island and La Plata Island, and throughout the Galápagos Archipelago (Glynn et al. 2001, Glynn 2003, Hickman et al. 2005, Glynn et al. 2007).
Native:American Samoa (American Samoa); Australia; Bahrain; British Indian Ocean Territory; Cambodia; Chile; Christmas Island; Cocos (Keeling) Islands; Colombia; Comoros; Cook Islands; Costa Rica; Djibouti; Ecuador; Egypt; El Salvador; Eritrea; Fiji; French Polynesia; Guadeloupe; Guam; Honduras; India; Indonesia; Iran, Islamic Republic of; Iraq; Israel; Japan; Jordan; Kenya; Kiribati; Kuwait; Madagascar; Malaysia; Maldives; Marshall Islands; Mauritius; Mayotte; Mexico; Micronesia, Federated States of ; Mozambique; Myanmar; Nauru; New Caledonia; Nicaragua; Niue; Norfolk Island; Northern Mariana Islands; Oman; Palau; Panama; Papua New Guinea; Philippines; Pitcairn; Qatar; Réunion; Samoa; Saudi Arabia; Seychelles; Singapore; Solomon Islands; Somalia; Sri Lanka; Sudan; Taiwan, Province of China; Tanzania, United Republic of; Thailand; Tokelau; Tonga; Tuvalu; United Arab Emirates; United States Minor Outlying Islands; Vanuatu; Viet Nam; Wallis and Futuna; Yemen
|FAO Marine Fishing Areas:||
Indian Ocean – eastern; Indian Ocean – western; Pacific – eastern central; Pacific – northwest; Pacific – southeast; Pacific – southwest; Pacific – western central
|Range Map:||Click here to open the map viewer and explore range.|
This is a common species.
Specific information for the Indo-West Pacific populations may be available since this is a widely studied species.
The relative abundance of Pocillopora damicornis in the Eastern Tropical Pacific region has been categorized as follows:
Abundant: from Nayarit to Oaxaca, Mexico (Reyes-Bonilla 2003); Panama, Costa Rica and Colombia (Cortes and Guzmán, pers. comm.; Glynn and Ault 2000). Highly recovering in Caño Island, Costa Rica (Guzmán and Cortes 2001) and in Panama (Guzmán et al. in prep.).
Common: Gulf of California (Reyes-Bonilla 2003, Glynn and Ault 2000).
Uncommon: Revillagigedo Islands, Mexico (Reyes-Bonilla 2003), and Ecuador including the Galápagos Archipelago (Glynn and Ault 2000, Glynn 2003).
Rare: Del Amor beach, El Salvador (Reyes-Bonilla and Barraza 2003).
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 Galápagos Islands (Glynn unpublished data in Glynn et al. 1991). According to Cortés and Guzmán (pers. comm.), Pocillopora damicornis seems to be very abundant and recovering in Panama, Costa Rica and Colombia, despite severe coral mortality after the ENSO events (1982-83 and 1997-98).
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). No live Pocillopora damicornis has been seen in the once coral-filled lava rock pools at Punta Espinosa, Fernandina, Galápagos Islands since 1983 (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 more resilient to some of the threats faced by corals and therefore population decline is estimated using the percentage of destroyed reefs only (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 of 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 all shallow water habitats from exposed reef fronts to mangrove swamps and wharf piles. This species is found in mono-specific stands or multi-species reefs throughout its range from near the surface to a maximum depth of 20 m. It is commonly found from 1-15 m, rarely 18-20 m, in the South China Sea and Gulf of Siam. (Titlyanov and Titlyanova 2002) This species is considered to be a main reef-framework builder and is found from 0.5-6 m of Panama (Sheppard 1982). This species is relatively tolerant of sedimentation and low salinity as long as there is adequate water motion. Colonies reproduce by fragmentation and by sexual reproduction (broadcast spawning) (Hodgson 1998).
In the Eastern Tropical Pacific region the species has not been reported from mangrove environments (Cortés and Guzmán pers. comm.) but it is one of the major reef building species, forming intermeshing compact frameworks that can attain 2-3 m in relief (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 damicornis vary substantially between locations in the Eastern Tropical Pacific, from 1.27 cm per year in Colombia to 3.96 cm per year in Panama (Guzmán and Cortes 1993).
Pocilloporid corals, presumably including P. damicornis, 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 (1-15m). 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).
P. damicornis 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. damicornis, 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.
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).
This is the most commonly harvested species in the area. Like other coral species it is collected to be used directly as construction material, to make lime, which is used to make concrete and to sell as curios. Live corals are collected for sale to the aquarium trade (Hodgson 1998).
This species is targeted for the aquarium trade. Indonesia is the largest exporter with an annual quota of 6,000 live pieces in 2005. The total number of corals (live and raw) exported for this species in 2005 was 15,321.
This species exhibited variable bleaching (0-50%) and low mortality in the 1998 bleaching event in Palau (Brunno et al. 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 3m 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.
Having timely access to national-level trade data for CITES analysis reports would be valuable for monitoring trends this species. The species is targeted by collectors for the aquarium trade and fisheries management is required for the species, e.g., MPAs, quotas, size limits, etc. Consideration of the suitability of species for aquaria should also be included as part of fisheries management, and population surveys should be carried out to monitor the effects of harvesting. Recommended conservation measures include population surveys to monitor the effects of collecting for the aquarium trade, especially in Indonesia.
|Citation:||Hoeksema, B.W., Rogers, A. & Quibilan, M.C. 2014. Pocillopora damicornis. The IUCN Red List of Threatened Species. Version 2015.2. <www.iucnredlist.org>. Downloaded on 29 July 2015.|
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