Conservation planning in a changing world
201
One of the best - known empirical examples of relaxation on an ecological timescale is the loss of bird
species from Barro Colorado Island in Panama. The
island was formerly a hilltop in an area of continuous
neotropical lowland rain forest, but abruptly became
a 15.7 km
2 island when the central section of the
Panama Canal Zone was fl ooded to make Lake Gatun
in 1914.
Of about 208 bird species estimated to have been
breeding on Barro Colorado island immediately following isolation in the 1920s and 1930s, 45 were no
longer present by 1970 (Wilson & Willis, 1975 ).
However, only a minority of these losses were directly
attributable to stochastic processes of relaxation. The
others could be attributed to ecological changes such
as forest regeneration following abandonment of
farming activity, which reduced the availability of open
habitats, or predation by terrestrial mammals (see
review in Whittaker & Fern á ndez - Palacios, 2007 ).
For example, many of the birds lost were typical of
second growth or forest edge, suggesting that the
regeneration of the forest following abandonment of
farming activity must have reduced the availability of
these more open habitats. Additionally, some ground -
nesting species were probably eliminated by their
terrestrial mammalian predators, which became abundant after the disappearance of top carnivores with
large area requirements. This effect, of increasing
numbers of smaller omnivores and predators due to
the absence of large ones, has been termed mesopredator release (Soul é et al. , 1988 ) and has been documented to occur in several other similar contexts (e.g.
Laurance, 2002 ).
A later avifaunal survey of Barro Colorado Island
reported sightings of 218 species from the island or the
waters immediately around it between 1994 and
1996, including fi ve new records, none of which were
thought to be of breeding species (Robinson, 1999 ). As
anticipated from the island theory (Figure 8.1 ), the
rate of species loss appears to have declined over time,
especially for forest - interior birds. However, overall,
species extinctions do appear to have continued to
exceed colonizations.
So, in summary, the isolation of the hilltops to form
this lake - bound island has been followed by around a
century in which the process of relaxation has been
the dominant trend. Future changes in avifaunal
species richness and composition on the island are
likely to be dependent on the extent to which the
It has been argued that it may take several generations for the processes causing relaxation to play
out following habitat destruction and fragmentation,
meaning that there is a substantial lag time between
the initial stimulus and the end of the process of species
losses (Tilman et al. , 1994 ; Ewers & Didham, 2006 ;
Vellend et al. , 2006 ). This creates an ‘ extinction debt ’
– a future ecological cost of habitat destruction that
may not be initially apparent in studies made shortly
after habitat fragmentation has occurred. Indeed,
Brown ’ s (1971) mammal assemblages were hypothesized to still be in the process of relaxation from their
relatively large mountain top habitat islands thousands of years after isolation (Figure 8.5 ; and see
further discussion in Lomolino et al. , 2006 ).
Whether such protracted response times are typical
is unknown, but it does seem highly likely that the true
ecological costs of the historically recent spate of
anthropogenic habitat disturbance, destruction and
fragmentation across the globe are yet to be realized
(see, for example, Figure 8.6 ). It is also noteworthy
that, although the majority of recorded species extinctions since AD 1600 have occurred on oceanic islands,
predictions of increasing numbers of future extinctions suggest a signifi cant shift to continental areas
(Millennium Ecosystem Assessment, 2005 ).
Developing methods to quantify the magnitude and
taxonomic distribution of the extinction debt is clearly
vitally important for effective conservation planning
and prioritization. However, this objective is by no
means simple to attain. Accurate assessment of extinction rates and their extrapolation into the future
requires good quality long - term data on species occurrences – data which are generally lacking, especially
for less conspicuous and/or numerically much more
species rich taxa.
This lack of appropriate knowledge (Chapter 4 ) has
led to an inevitable reliance on indirect measures
and theoretical projections of extinction debt. These
include: species – area models; rates at which well -
known species are shifting to increasingly more threatened categories of conservation concern; extinction
probabilities associated with the IUCN categories of
threat; impacts of projected habitat loss on species currently threatened with habitat loss; and the extrapolation of correlations of species loss with climate change
(e.g. McDonald & Brown, 1992 ; Mace & Kunin, 1994 ;
Pimm & Askins, 1995 ; Thomas et al. , 2004 – for
further discussion see Ladle, 2009 ).
201
One of the best - known empirical examples of relaxation on an ecological timescale is the loss of bird
species from Barro Colorado Island in Panama. The
island was formerly a hilltop in an area of continuous
neotropical lowland rain forest, but abruptly became
a 15.7 km
2 island when the central section of the
Panama Canal Zone was fl ooded to make Lake Gatun
in 1914.
Of about 208 bird species estimated to have been
breeding on Barro Colorado island immediately following isolation in the 1920s and 1930s, 45 were no
longer present by 1970 (Wilson & Willis, 1975 ).
However, only a minority of these losses were directly
attributable to stochastic processes of relaxation. The
others could be attributed to ecological changes such
as forest regeneration following abandonment of
farming activity, which reduced the availability of open
habitats, or predation by terrestrial mammals (see
review in Whittaker & Fern á ndez - Palacios, 2007 ).
For example, many of the birds lost were typical of
second growth or forest edge, suggesting that the
regeneration of the forest following abandonment of
farming activity must have reduced the availability of
these more open habitats. Additionally, some ground -
nesting species were probably eliminated by their
terrestrial mammalian predators, which became abundant after the disappearance of top carnivores with
large area requirements. This effect, of increasing
numbers of smaller omnivores and predators due to
the absence of large ones, has been termed mesopredator release (Soul é et al. , 1988 ) and has been documented to occur in several other similar contexts (e.g.
Laurance, 2002 ).
A later avifaunal survey of Barro Colorado Island
reported sightings of 218 species from the island or the
waters immediately around it between 1994 and
1996, including fi ve new records, none of which were
thought to be of breeding species (Robinson, 1999 ). As
anticipated from the island theory (Figure 8.1 ), the
rate of species loss appears to have declined over time,
especially for forest - interior birds. However, overall,
species extinctions do appear to have continued to
exceed colonizations.
So, in summary, the isolation of the hilltops to form
this lake - bound island has been followed by around a
century in which the process of relaxation has been
the dominant trend. Future changes in avifaunal
species richness and composition on the island are
likely to be dependent on the extent to which the
It has been argued that it may take several generations for the processes causing relaxation to play
out following habitat destruction and fragmentation,
meaning that there is a substantial lag time between
the initial stimulus and the end of the process of species
losses (Tilman et al. , 1994 ; Ewers & Didham, 2006 ;
Vellend et al. , 2006 ). This creates an ‘ extinction debt ’
– a future ecological cost of habitat destruction that
may not be initially apparent in studies made shortly
after habitat fragmentation has occurred. Indeed,
Brown ’ s (1971) mammal assemblages were hypothesized to still be in the process of relaxation from their
relatively large mountain top habitat islands thousands of years after isolation (Figure 8.5 ; and see
further discussion in Lomolino et al. , 2006 ).
Whether such protracted response times are typical
is unknown, but it does seem highly likely that the true
ecological costs of the historically recent spate of
anthropogenic habitat disturbance, destruction and
fragmentation across the globe are yet to be realized
(see, for example, Figure 8.6 ). It is also noteworthy
that, although the majority of recorded species extinctions since AD 1600 have occurred on oceanic islands,
predictions of increasing numbers of future extinctions suggest a signifi cant shift to continental areas
(Millennium Ecosystem Assessment, 2005 ).
Developing methods to quantify the magnitude and
taxonomic distribution of the extinction debt is clearly
vitally important for effective conservation planning
and prioritization. However, this objective is by no
means simple to attain. Accurate assessment of extinction rates and their extrapolation into the future
requires good quality long - term data on species occurrences – data which are generally lacking, especially
for less conspicuous and/or numerically much more
species rich taxa.
This lack of appropriate knowledge (Chapter 4 ) has
led to an inevitable reliance on indirect measures
and theoretical projections of extinction debt. These
include: species – area models; rates at which well -
known species are shifting to increasingly more threatened categories of conservation concern; extinction
probabilities associated with the IUCN categories of
threat; impacts of projected habitat loss on species currently threatened with habitat loss; and the extrapolation of correlations of species loss with climate change
(e.g. McDonald & Brown, 1992 ; Mace & Kunin, 1994 ;
Pimm & Askins, 1995 ; Thomas et al. , 2004 – for
further discussion see Ladle, 2009 ).