branching coral rubble compared to relatively unpolluted
sites (Holmes et al., 2000).
Studies in the polluted Hong Kong Harbour showed
that stressed corals were more extensively bored than
healthy ones (Dudgeon and Morton, 1982) and weakened
the skeleton making it more susceptible to storm damage.
In contrast low densities of mollusc borers in healthy corals
may actually strengthen the substrate as Lithophaga lines its
burrows with aragonitic secretion (Barthel, 1982).
Both experimental and field studies show that while
eutrophic sites experience high levels of bioerosion
(Reaka-Kudla et al., 1996; Cortes, 1993; Holmes et al.,
2000), pristine sites may also exhibit elevated rates (Pari
et al., 2002). So high rates per se do not always indicate
poor water quality.
Increasing water temperatures
Sites in the Caribbean with extensive stands of the elkhorn
coral Acropora palmate, which were subjected to massive
mortalities and bleaching during the 1980s, now consist of
broken dead stands covered by the encrusting and excavating sponge Cliona tenuis. As the sponge undermines
the branches of coral, they break off and during storms
they are thrown against new coral hosts and the sponge
is able to colonize a new uninfected coral colony. The
times of initial colonization of the corals by the sponges
was related to the timing of hurricanes in the area. In addition to infecting the dead branches, the sponge was also
undermining encrusting and foliose corals settling on the
dead A. palmata, retarding the recovery of these reefs
(López-Victoria and Zea, 2004).
Extensive mortality of corals from bleaching at Uva
Island, Panama caused by the prolonged 1982–1983 El
Niňo event, led to significant increases in echinoid grazers
and increased rates of internal bioerosion led to significant
loss of reef framework and collapse of reef walls (Eakin,
1992).
Kleeman (2008) working on reefs in the Maldives after
a severe bleaching event found dense concentration of
the bivalve Parapholas quadrizonata with boreholes
reaching 80 mm in length and 25 mm in diameter and he
estimated life spans of 3–8 years probably 10 years. Thus
a continual supply of larvae is being produced and as
many of the reefs have less than 50% live coral cover following the severe bleaching event in 1998, this has
resulted in an accelerated rate of loss of reefal substrate
as suitable dead coral substrate is available for colonization by the bivalves (Kleeman, 2008).
Cumulative impacts
Initially many of these anthropogenic impacts were related
to the location of most coral reefs in developing countries
with increasing urbanization and declining water quality
in part due to lack of sewage treatment works, unregulated
coastal development leading to excessive land run off, loss
of riparian vegetation along the rivers flowing onto the
reefs, overfishing and inappropriate fishing techniques
such as dynamite fishing and collecting of coral for building and inappropriate dredging in lagoons. Typically, this
has led to damaged fringing reefs close to centres of population. These same communities were also exploiting the
reefs as tourist attractions earning valuable revenue with
often the nearby hotel developments also impacting on
these reefs. But during the 1980s increasing records of
widespread bleaching of coral were being recorded, while
some bleached coral colonies recovered many did not.
Satellite imaging of surface water temperatures allowed
the areas where bleaching was likely to occur to be determined and subsequent surveys often supported these
predictions. As well as the duration of elevated water temperatures other factors such as water quality seemed to be
involved in determining the recovery of the reef as well as
location of nearby unaffected reefs. Increased incidences
of bleaching around the world highlighted the impact of
climate change on reefs, and it became evident that reefal
ecosystems are one of the most vulnerable to climate
change. A recent vulnerability assessment of the climate
change on the Great Barrier Reef (Johnson and Marshall,
2007) is a sobering analysis of the various aspects of reefs,
which are changing, not just elevated temperatures,
increased storm intensities, increased run off, increasing
alkalinity, rising sea levels, and changes in oceanography
which will impact on recruitment processes, increased disease and invasive species, for example. Many of these
impacts will result in increased amounts of dead coral substrate and thus higher rates of bioerosion of reef framework (Hutchings et al., 2007; Przeslawski et al., 2008).
Loss of reef framework will impact directly on those
organisms which either feed or live on live coral, change
fish communities (Cinner et al., 2009) which will have
serious economic consequences for many people. Tourism
revenues will decline and low lying areas which are currently protected from storms will loose this protection
and low lying areas will become inundated as sea levels
rise. Many of the coastal settlements lack the financial
resources to protect these areas and we are already seeing
that people are being relocated from such areas.
Conclusions
While over the past 20 years or so, our understanding of
the processes of bioerosion has increased considerably,
many gaps remain, especially as to the mechanisms of boring and the interactions between borers and grazers. Our
knowledge is best for the Caribbean, Great Barrier Reef,
and French Polynesia, with little information available
from the Western Indian Ocean, Red Sea, and SE Asia.
There is an urgent need to continue studies on the impact
which water quality and sediment loads have on rates
and agents of bioerosion and how these will change
with increasing ocean acidification and rises in seawater
temperatures.
As reefs are increasingly being subjected to anthropogenic impacts many of which act synergistically, it is
becoming critical to develop monitoring techniques which
152
BIOEROSION
sites (Holmes et al., 2000).
Studies in the polluted Hong Kong Harbour showed
that stressed corals were more extensively bored than
healthy ones (Dudgeon and Morton, 1982) and weakened
the skeleton making it more susceptible to storm damage.
In contrast low densities of mollusc borers in healthy corals
may actually strengthen the substrate as Lithophaga lines its
burrows with aragonitic secretion (Barthel, 1982).
Both experimental and field studies show that while
eutrophic sites experience high levels of bioerosion
(Reaka-Kudla et al., 1996; Cortes, 1993; Holmes et al.,
2000), pristine sites may also exhibit elevated rates (Pari
et al., 2002). So high rates per se do not always indicate
poor water quality.
Increasing water temperatures
Sites in the Caribbean with extensive stands of the elkhorn
coral Acropora palmate, which were subjected to massive
mortalities and bleaching during the 1980s, now consist of
broken dead stands covered by the encrusting and excavating sponge Cliona tenuis. As the sponge undermines
the branches of coral, they break off and during storms
they are thrown against new coral hosts and the sponge
is able to colonize a new uninfected coral colony. The
times of initial colonization of the corals by the sponges
was related to the timing of hurricanes in the area. In addition to infecting the dead branches, the sponge was also
undermining encrusting and foliose corals settling on the
dead A. palmata, retarding the recovery of these reefs
(López-Victoria and Zea, 2004).
Extensive mortality of corals from bleaching at Uva
Island, Panama caused by the prolonged 1982–1983 El
Niňo event, led to significant increases in echinoid grazers
and increased rates of internal bioerosion led to significant
loss of reef framework and collapse of reef walls (Eakin,
1992).
Kleeman (2008) working on reefs in the Maldives after
a severe bleaching event found dense concentration of
the bivalve Parapholas quadrizonata with boreholes
reaching 80 mm in length and 25 mm in diameter and he
estimated life spans of 3–8 years probably 10 years. Thus
a continual supply of larvae is being produced and as
many of the reefs have less than 50% live coral cover following the severe bleaching event in 1998, this has
resulted in an accelerated rate of loss of reefal substrate
as suitable dead coral substrate is available for colonization by the bivalves (Kleeman, 2008).
Cumulative impacts
Initially many of these anthropogenic impacts were related
to the location of most coral reefs in developing countries
with increasing urbanization and declining water quality
in part due to lack of sewage treatment works, unregulated
coastal development leading to excessive land run off, loss
of riparian vegetation along the rivers flowing onto the
reefs, overfishing and inappropriate fishing techniques
such as dynamite fishing and collecting of coral for building and inappropriate dredging in lagoons. Typically, this
has led to damaged fringing reefs close to centres of population. These same communities were also exploiting the
reefs as tourist attractions earning valuable revenue with
often the nearby hotel developments also impacting on
these reefs. But during the 1980s increasing records of
widespread bleaching of coral were being recorded, while
some bleached coral colonies recovered many did not.
Satellite imaging of surface water temperatures allowed
the areas where bleaching was likely to occur to be determined and subsequent surveys often supported these
predictions. As well as the duration of elevated water temperatures other factors such as water quality seemed to be
involved in determining the recovery of the reef as well as
location of nearby unaffected reefs. Increased incidences
of bleaching around the world highlighted the impact of
climate change on reefs, and it became evident that reefal
ecosystems are one of the most vulnerable to climate
change. A recent vulnerability assessment of the climate
change on the Great Barrier Reef (Johnson and Marshall,
2007) is a sobering analysis of the various aspects of reefs,
which are changing, not just elevated temperatures,
increased storm intensities, increased run off, increasing
alkalinity, rising sea levels, and changes in oceanography
which will impact on recruitment processes, increased disease and invasive species, for example. Many of these
impacts will result in increased amounts of dead coral substrate and thus higher rates of bioerosion of reef framework (Hutchings et al., 2007; Przeslawski et al., 2008).
Loss of reef framework will impact directly on those
organisms which either feed or live on live coral, change
fish communities (Cinner et al., 2009) which will have
serious economic consequences for many people. Tourism
revenues will decline and low lying areas which are currently protected from storms will loose this protection
and low lying areas will become inundated as sea levels
rise. Many of the coastal settlements lack the financial
resources to protect these areas and we are already seeing
that people are being relocated from such areas.
Conclusions
While over the past 20 years or so, our understanding of
the processes of bioerosion has increased considerably,
many gaps remain, especially as to the mechanisms of boring and the interactions between borers and grazers. Our
knowledge is best for the Caribbean, Great Barrier Reef,
and French Polynesia, with little information available
from the Western Indian Ocean, Red Sea, and SE Asia.
There is an urgent need to continue studies on the impact
which water quality and sediment loads have on rates
and agents of bioerosion and how these will change
with increasing ocean acidification and rises in seawater
temperatures.
As reefs are increasingly being subjected to anthropogenic impacts many of which act synergistically, it is
becoming critical to develop monitoring techniques which
152
BIOEROSION