Acanthaster planci - General Impacts
Compiled by IUCN SSC Invasive Species Specialist Group (ISSG)
Since the 1960’s outbreaks of the coral-feeding starfish Acanthaster planci have been
recorded throughout the Indo-Pacific region . From the time these outbreak were first
recorded it has been recognised that they pose a threat to the viability of coral reef
habitats and the creatures that depend on them. The impact of outbreaks of the coralfeeding starfish on natural coral assemblages can be severe and long-lasting. On some
reefs up to 90% of live coral cover has been lost, as was the case in areas of Saipan
(Tsuda et al. 1970), the Marshall Islands (Pinca et al. Undated) and Guam (Chesher
1969). The impact of outbreaks can be profound. For example the branching corals of
Iriomote Island (Ryukyu Islands, Japan) were completely decimated by A. planci and
replaced by flat plains of rubble, significantly lower in fish diversity (Sano 2000).
Similarly coral gardens at Tanguisson Reef (Guam) were devastated following the
1960s outbreak of the coral-feeding starfish and the composition of coral communities
shifted from preferred prey species such as Montipora and Acropora to non-preferred
prey species such as Porites, Millepora and Leptastrea (Colgan 1987). A study by
Cameron, Endean and De Vantier (1991) found that coral structure was significantly
different between A. planci affected and non-affected sites. Members of the Poritidae
family were predominant on outbreak-affected reefs, while members of the Faviidae
family were predominant on non-outbreak reefs. In addition few large (old) colonies
occurred on the outbreak reefs, whereas such large corals were common on unaffected
reefs. On the Great Barrier Reef and elsewhere significantly higher abundances of turf
algae are found to occur on reefs affected by starfish predation in comparison with
live coral (Hart and Klumpp 1996). Fortunately, secondary invasion by competitively
strong groups of macro-benthos, such as soft corals or macro-algae, appears not to be
a limiting factor when it comes to coral recovery (Fabricius 1996).
The changes in coral composition mentioned in the last paragraph may be longlasting, as is the case in the western Pacific Islands of Rota, Saipan and Tinian where
non-preferred coral prey such as Poritidae dominate and preferred prey Acroporidae
and Pocilloporidae are kept low by A. planci (Quinn and Kojis 2003). Alternatively,
such changes may be temporary. Data taken from reefs in both Guam and Japan
suggests that coral reefs may recover (in terms of species richness, density and fish
assemblages) from starfish damage in as little as 10 to 20 years (Colgan 1987; Sano
2000). Colgan (1987) commented that the rapid recovery of a coral community from
natural disturbance by A. planci demonstrates that some reef ecosystems have a
greater resilience than once estimated. On the other hand, evidence to the contrary
exists which is both alarming and disheartening. A study by Seymour and Bradbury
(1999) showed that the average reef recovery time on the Great Barrier Reef
(Australia) is lengthening over time, and it is harder for reefs to recover from recent
outbreaks of A. planci than it has been in the past. The authors believe this indicates
that key features of reef community structure have been damaged over time. Lourey
Ryan and Miller (2000) found that for coral cover in areas of the Great Barrier Reef
damaged by A. planci to increase by 30%, it would take an estimated 5 years to 1000+
years. This highlights the variability of rate recovery times between reefs and raises
the possibility that not all reefs will recover from sustained outbreaks of A. planci
(Lourey, Ryan and Miller 2000).
The possible implications of ongoing outbreaks of A. planci are alarming. However
what makes the scenario even more alarming is that, throughout its range, coral reefs
are coming under increasing pressure from human impacts. A recent report
(Wilkinson 2004) predicts that 24% of the world’s reefs are under imminent risk of
collapse through human pressures; and a further 26% are under longer term threat of
collapse. Human impacts exacerbate the effects of natural disturbance such as A.
planci outbreaks by contributing to coral mortality and reducing a reefs ability to
recover. For example, large-scale silt depositions contribute to reef degradation in the
Indian Ocean, where disease, predation and stress are listed as the key factors in
causing coral mortality (Ravindran, Raghukumar and Raghukumar 1999). Similarly,
the degradation of coral reefs in Oman was found to be due to both natural and human
causes, with damage from fishery activities being recorded as the most common
human impact (Al-Jufaili 2005). Other human impacts to coral ecosystems in the area
include coastal construction, recreational activities, oil pollution and eutrophication
(causing nuisance algal growth) (Al-Jufaili 2005). Predation of corals by A. planci,
storm damage, coral diseases and temperature-related stresses were the most
commonly recorded natural impacts to coral reefs in the country (Al-Jufaili 2005). In
the Cocos Island (off the coast of Costa Rica) coral reefs, already significantly
damaged by the 1982-83 El Nino event, became further degraded by high densities of
the corallivores A. planci and Arothron meleagris (the Guinea fowl pufferfish) and the
bio-eroding Diadema mexicanum (the Mexican Pacific sea urchin) (Guzman and
Cortes 1992). Surveys from 1987 found that in some areas of the reef live coral cover
was as low as 3% and scientists believe the coral reefs of the Cocos Islands will need
a time-frame in the order of centuries to recover their original reef-framework and
thickness (Guzman and Cortes 1992). While human impacts on coral reefs can be
mitigated relatively easily by reducing human impacts on the reef ecosystem, it is
difficult to predict outbreaks of A. planci starfish and harder still to manage their
populations once they have reached significant proportions.
Destruction of living coral reef ecosystems is a potential economic disaster for small
isles and atolls of Oceania. Most inhabitants of the region derive almost all their
protein from marine resources. The destruction of reefs would result in the destruction
of fisheries, as well as increasing land erosion along the coast (Chesher 1969). Coralfeeding starfish outbreaks have hindered traditional fishing in Samoa (Birkeland and
Lucus 1990) and elsewhere and dying coral reefs have put livelihoods in jeopardy.
Tour boat operations, diving expeditions, eco-tourism and other tourist attractions
based on reef environments are all at risk of economic loss due damage caused by the
coral-feeding starfish. Popular tourist destinations in the Great Barrier Reef, have
been significantly degraded by A. planci. On Grub Reef, for example, live coral
remaining after an outbreak was so poor that tour boat operators had to cease tours to
the site altogether (Birkeland and Lucus 1990). Such potential outcomes have
prompted control teams in Palau to engage other stakeholders in control efforts,
approaching dive shops and tour operators to "adopt a reef" (Quarterly Report 2002),
a control-management strategy which should be encouraged on a wider scale.
Many factors impact on the size and extent of A. planci outbreaks. The exact cause of
outbreaks have not been determined with any certainty and may in fact vary
depending on local and regional factors. Currently there are considered to be two
main hypotheses for the anthropogenic causes of outbreaks. The first of these is is the
predator removal hypothesis. According to this view decreased post settlement
mortality by the removal of predators such as fish and the shellfish (i.e. the giant
triton) allows adult starfish to persist and build up in numbers on a reef. For instance
some studies have shown that predation is an important determinant of survival rates
of juvenile starfish (Keesing et al. 1996). In Mauritius high A. planci numbers have
been linked to low numbers of its main predator, the mollusc Charonia tritonis
(Triton’s trumpet). In Egypt removal of fish in the families Lethrinidae, Balistidae and
Tetraododontidae have been linked to outbreaks of the coral-feeding starfish(Ormond
et al. 1990, in PERSGA/GEF 2003). The second anthropogenic theory is the nutrient
enrichment hypothesis. This revolves around increase pre-settlement survivorship of
larval COTS onto a reef. In this scenario terrestrial runoff due to extreme rainfall
events or eutrophication causes nutrient enrichment of coastal waters. The increase in
nutrients results in an increase in phytoplankton, upon which the starfish larvae feed,
thereby increases their survival in the water column. Because starfish produce such a
vast quantity of eggs even a small increase in survivorship leads to larger settlement
of larvae onto a reef, which in turn leads to an outbreak. This mechanism has been
implicated in outbreaks in Micronesia and Polynesia (Birkeland 1982). Similarly,
frequent A. planci outbreaks on the Great Barrier Reef have been linked to increased
nutrient delivery from the land (Brodie et al. 2005). River-promoted eutrophication
(algal growth) is a significant factor in the demise of fringing reefs in the inner Great
Barrier Reef lagoon and recorded levels of nano plankton growth in some regions are
sufficient to promote the survival of A. planci larvae and may be implicated to starfish
outbreaks (Bell 1992).
Finally it is worth noting that A.planci is not only a physical danger to corals. The
crown-of-thorns starfish (as its name may suggest) has an array of penetrating spines
which can produce a painful wound, as well as redness, swelling, vomiting, numbness
and paralysis in humans. In at least one case A. planci triggered a very nasty
inflammatory response in a patient, which resulted in local swelling of the hand and
fingers. Even following effective drug treatment the movement of the fingers was still
limited six months later (Adler Kaul and Jawad 2002).