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Marine Pollution Bulletin xxx (2007) xxx–xxx
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Review
Global change and marine communities: Alien species
and climate change
Anna Occhipinti-Ambrogi
*
DET - Dip. di Ecologia del Territorio, Sezione di Ecologia, Università degli Studi di Pavia,Via S. Epifanio 14, I-27100 Pavia, Italy
Abstract
Anthropogenic influences on the biosphere since the advent of the industrial age are increasingly causing global changes. Climatic
change and the rising concentration of greenhouse gases in the atmosphere are ranking high in scientific and public agendas, and other
components of global change are also frequently addressed, among which are the introductions of non indigenous species (NIS) in biogeographic regions well separated from the donor region, often followed by spectacular invasions. In the marine environment, both climatic change and spread of alien species have been studied extensively; this review is aimed at examining the main responses of
ecosystems to climatic change, taking into account the increasing importance of biological invasions.
Some general principles on NIS introductions in the marine environment are recalled, such as the importance of propagule pressure
and of development stages during the time course of an invasion. Climatic change is known to affect many ecological properties; it interacts also with NIS in many possible ways. Direct (proximate) effects on individuals and populations of altered physical–chemical conditions are distinguished from indirect effects on emergent properties (species distribution, diversity, and production). Climatically driven
changes may affect both local dispersal mechanisms, due to the alteration of current patterns, and competitive interactions between NIS
and native species, due to the onset of new thermal optima and/or different carbonate chemistry.
As well as latitudinal range expansions of species correlated with changing temperature conditions, and effects on species richness and
the correlated extinction of native species, some invasions may provoke multiple effects which involve overall ecosystem functioning
(material flow between trophic groups, primary production, relative extent of organic material decomposition, extent of benthic-pelagic
coupling). Some examples are given, including a special mention of the situation of the Mediterranean Sea, where so many species have
been introduced recently, and where some have spread in very large quantities.
An increasing effort by marine scientists is required, not only to monitor the state of the environment, but also to help predicting
future changes and finding ways to mitigate or manage them.
Ó 2006 Elsevier Ltd. All rights reserved.
1. Introduction
Biological invasions are being recognized as an important
element of global change, following the observation of
increasingly spectacular developments of alien species in
various regions of the world. Climate change and other
components of global change such as increasing deposition
of nitrogen and pollutants, and habitat disturbance by
human activities, can affect species distribution and resource
dynamics in both terrestrial and aquatic ecosystems and
*
Tel.: +39 382 984876; fax: +39 382 304610.
E-mail address: [email protected]
consequently can interact with biological invasions (Dukes
and Mooney, 1999; Vitousek et al., 1997).
The effect of climate change and invasive species have
been implicated in the decline and even collapse of several
marine ecosystems (Harris and Tyrrell, 2001; Stachowicz
et al., 2002; Frank et al., 2005) and are known to affect
the presence of pathogens (see Drake et al., this issue).
When many stressors act in synergism they may eventually
have unexpected and irreversible consequences for the
native communities and also may impact economically
valuable human activities such as fisheries in a particular
region (Occhipinti-Ambrogi and Savini, 2003; Whitfield
et al., 2007). This scenario implies a direct economic cost
0025-326X/$ - see front matter Ó 2006 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2006.11.014
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A. Occhipinti-Ambrogi / Marine Pollution Bulletin xxx (2007) xxx–xxx
within marine environments that is related to introduced
species, a cost that is just beginning to be recognized
(Pimentel et al., 2000).
Impact of invasive species has been defined by Parker
et al. (1999), who also discussed, from empirical examples
and theoretical reasoning, a variety of measures of impact.
In the marine environment, Ruiz et al. (1999) describe
impact by alien species and interactions with other stress
factors. Stressing the importance of this man-induced disturbance, Elliot (2003) observes that there are many aspects
in which introduced marine organisms can be regarded as
being no different from chemical pollutants and encourages
the use of the term biological pollution.
On the other hand, beneficial aspects of introductions
are claimed, since intentionally introduced species have significantly contributed to aquaculture production (FAO
DIAS, 1998), as well as fisheries and angling (Minchin
and Rosenthal, 2002). Unintentionally introduced species,
such as the Erythrean species which entered the Mediterranean through the Suez Canal, have become locally of commercial importance (Golani and Ben Tuvia, 1995); the
Mediterranean mussel Mytilus galloprovincialis, accidentally introduced to the West coast of South Africa in the
mid 1970s, was deliberately introduced to the South coast
for mariculture purposes, despite the fact that it had
become invasive, outcompeting local mussels (Branch and
Steffani, 2004).
Ecological principles for predicting the success of invasions in new environments have always been highly
debated, yet have profound implications for management
and control of human mediated introductions. At the same
time, predicting the effects of climate change on invasions is
more difficult, because ecologists must face a complicated
pattern of perturbations, independently affecting species
and ecosystems.
Dukes and Mooney (1999) stated that the emergent pattern of predictions on the effect of global change on invasions leads us to expect that invasive alien species could
be favoured, so that the impacts of invasions on ecosystems
would be exacerbated. In this review some general principles of invasions in the marine environment and some general principles of various elements of global climate change
will be recalled, in order to help the interpretation of phenomena observed in different seas. Special reference is
made to the Mediterranean Sea, which is one of the most
affected seas of the world in this respect, both in terms of
length of time that invasives have been present and in numbers of alien species detected (see Galil, this issue).
2. Biological invasions in the sea
The establishment of NIS in new ecosystems has been
the subject of the so-called ‘‘invasion ecology’’ that has frequently concentrated on the identification of ‘‘invasiveness’’ characteristics of species that proliferate in novel
habitats (Gray, 1986; Lodge, 1993; Williamson, 1996;
Kolar and Lodge, 2001) and/or ‘‘invasibility’’ characteris-
tics of particular habitats that are susceptible to the proliferation of NIS (Wolff, 1999; Shea and Chesson, 2002;
Nehring, 2006).
The role of propagule pressure (i.e. the number of individuals introduced and the number of introduction
attempts) is obviously very important in determining the
success of NIS establishment, although it is not always
taken into account in studies on biological invasions (Williamson, 1996; Ruiz et al., 2000; Ruesink, 2005; Colautti
et al., 2006). In the marine environment, propagule pressure has increased steadily from NIS transport via ships
(Carlton, 1985; Hewitt, 2002; Coutts et al., 2003) and via
aquaculture (Casal, 2006), along with a range of other
activities (Minchin, this issue) such as the opening of navigable canals (Por, 1978; Galil, 2000) and the trade of species for aquaria (Semmens et al., 2004; Calado and
Chapman, 2006; Bolton and Graham, 2006).
Propagule pressure is not easily measurable directly,
except with intentional species introductions (e.g. transplanting of molluscs or release of commercially important
fishes and crustaceans), but can be related with some measure to the intensity of unintentional introductions, such as
the estimated quantities of discharged ballast waters, or
number of boats and ships that might carry fouling on
hulls, or the presence of oyster and mussel cultures that
can favour hitch-hiking species epibionts, or parasites of
imported molluscs.
Given that NIS are introduced non-randomly through
these vectors, failing to consider propagule pressure causes
a bias in the interpretation of observed results. Both for
invasiveness and invasibility, the characteristics of good
invaders and susceptible habitats can be confounded by
propagule bias (Colautti et al., 2006). Although the lack
of knowledge about dose-response relationships between
propagule supply and invasion success represents a critical
gap for invasion science, management actions to reduce
transfers must proceed concurrently with scientific efforts
to evaluate dose-response and efficacy (Ruiz and Carlton,
2003; Lockwood et al., 2005).
So, when we consider the impact of global change, we
first of all take into account the magnitude of release of
NIS propagules, and secondly the effects of climatic
change. Wonham and Carlton (2005) have discussed the literature data of the Northeast Pacific Coast (Northern California to British Columbia) giving indications of the port
traffic and of oyster culture that were in accord with the
observed spatial and temporal pattern of NIS introduction.
In European waters some figures have been published by
Streftaris et al. (2005) and Gollasch (2006). The latter
author estimates that in European coastal waters 39% of
the observed NIS were transported by ships (22% in ballast
waters and 17% as hull fouling), 16% by aquaculture, 9%
were intentionally released, and 24.5% entered through
the Suez canal.
Development of NIS populations is a dynamic process
involving different stages (Colautti and MacIsaac, 2004)
(Fig. 1). Potential invaders begin their life history as prop-
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Population in donor
region
DETERMINANTS
PP+
PP: propagule pressure
uptake
PCF: physical-chemical factors
CI: community interactions
propagules in a
vector
FILTERS
+
PP
±
PCF
vector survival and
release
individuals released
and settled in a new
environment (NE)
+
PP
±
PCF
±
CI
environment
survival and
reproduction
new population in
the NE
+
PP
PCF
±
CI
±
local dispersal
dispersal
environment and
community
suitability
dominant
population in
the NE
in several new
environments
dominant population
PCF±
CI±
in several new
+
PP
environments
environment and
community
suitability
local dispersal
Fig. 1. Stages in the development of an invasive non indigenous species. Transition between stages involves the passing of propagules through filters. The
success in reaching a new stage is determined, besides propagule pressure, by two broad categories of factors : physical–chemical matching and biotic
resistance or facilitation. The determinants may act in a positive (+) or negative ( ) way. Modified from Colautti et al. (2006).
agules residing in a donor region (stage 0), like the port of
call of a ship pumping ballast water or the site of production of bivalve seed shipped for aquaculture. They subsequently pass through a series of filters that may preclude
transition to subsequent stages. Stage 1 corresponds to
the uptake of propagules by the transport vector. After
passing through a survival and release filter, stage 2 is
reached, corresponding to some individuals settled in the
new environment (e.g. metamorphosed larvae in the port
where the ship has deballasted or individuals that have
escaped to the wild from aquaculture facilities).
The passage through the environment survival and
reproduction filter leads to the establishment of a NIS
population (stage 3). The success of the NIS after establishment is governed by two different filters: a ‘‘local
dispersal filter’’, mainly connected to propagule pressure,
determines which stage 3 species reaches stage 4a
(widespread) or which stage 4b species (dominant) can
reach stage 5 (widespread and dominant). The environmental and community suitability filter acts regulating the
passage from stage 3 to stage 4b and from stage 4a and
stage 5.
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As an example, most of the more notorious invaders
(e.g. the Manila clam, Ruditapes philippinarum) form stage
5 populations because they have been introduced repeatedly to ideal habitats, offering niche opportunities. Such
stage 5 species have therefore passed through both the
‘‘local dispersal’’ and ‘‘environmental survival and reproduction’’ filters. On the contrary, other species became
widespread (stage 4a), yet they rarely achieved high densities along the coasts of the Mediterranean, like the algae
Asparagopsis armata and Lophocladia lallemandi (Ribera
Siguan, 2002), which have been repeatedly introduced.
These two examples illustrate the importance of the propagule pressure determinant, compared with determinants
related to biotic and physical–chemical conditions.
Verlaque and Boudouresque (2004) did not find any
connection between the numerous (1 0 5) species of introduced algae and water warming in the Mediterranean,
and most such introductions are better explained by the
presence of intense aquaculture in the hotspots of alien
species.
Finally the introduction of the decapod crustacean
Dyspanopeus sayi in the Northern Adriatic may be an
example of stage 4b population. After their introduction
in the Lagoon of Venice in 1992 (Froglia and Speranza,
1993), it reached high abundances locally (Mizzan, 1995;
Mistri, 2004), then declined and is now rather rare (Mistri
and Mizzan, personal communications).
Climatic change is known to affect many ecological
properties (Walther et al., 2002); it interacts also with
NIS (whose propagule pressure is increasing steadily),
directly by altering physical–chemical conditions (primarily
temperature but also related oceanographic characteris-
tics), and indirectly contributing to the change in the new
communities patterns.
3. Impact of climatic change on marine organisms
(proximate effects)
Anthropogenically induced global climate change has
profound implications for marine ecosystems, well beyond
the direct influence of temperature on marine organisms.
Both abiotic changes and biological responses in the marine environment are complex and deserve a great deal of
research in order to interpret actual patterns and to predict
future changes (Harley et al., 2006).
Following the scheme by Harley et al. (2006) in Fig. 2,
changes in the life cycle of a generic marine species affected
by abiotic environmental changes will be considered first.
These include effects on dispersal and recruitment and on
individual performance.
Climate varies naturally across a range of temporal
scales including seasonal cycles, inter-annual patterns such
as the El Niño Southern Oscillation (ENSO), inter-decadal
cycles such as the North Atlantic Decadal oscillation, and
multimillenial-scale changes such as glacial to inter-glacial
transitions. This natural variability is reflected in evolutionary adaptations of species and biogeography. In the
Mediterranean Sea, for instance, a surface temperature
increase of about 1 °C in the period 1974–2004 has been
reported by Salat and Pascual (2002). In deep waters, a
0.12 °C increase is estimated for the last 40 years (Béthoux
et al., 1990; Béthoux and Gentili, 1996; Goffart et al.,
2002).
CLIMATIC
CHANGE
PROXIMATE ECOLOGICAL RESPONSES
(physiological – morphological - behavioural)
performance of
individuals
larvae/propagules
transport
processes
adults
population
dynamics
EMERGENT ECOLOGICAL RESPONSES
Changes in:
- species distribution
- diversity
- productivity
- microevolution
change in population size
and structure of
community
Fig. 2. Main response pathways of ecological systems to global climatic change. Modified from Harley et al. (2006).
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5
small recruitment
native species 1
medium recruitment
native species 2
no. of recruits
large recruitment
native species 3
native species 4
introduced species
reproductive season
time range
Fig. 3. Simplified patterns of growth of some native and introduced colonial species along the reproductive season. Modified from Stachowicz and Byrnes
(2006).
The effects of warming climate are primarily a cause for
physiological stress, which acts more strongly on species
already near their tolerance limit (Laubier, 2001). Anomalous temperature stress can cause mass mortalities in benthic organisms (Cerrano et al., 2000; Pérez et al., 2000;
Garrabou et al., 2001) that result in empty niches for
new colonisers. However, the responses of the organisms
are often very subtle and cannot be assessed simply by
exposing different life stages to a range of temperatures.
In particular, in seasonal environments where conditions
change predictably, each species may have a temporal or
seasonal niche, caused for instance by differences in growth
rate, patterns of mortality, timing, duration and magnitude
of reproductive output.
Stachowicz et al. (2002) demonstrated temporal complementarity in recruitment for dominant invertebrate species
on hard bottoms, noting that the recruitment pattern of the
NIS Botrylloides violaceus coincided with a period of low
recruitment of other native species of ascidians (Fig. 3).
The competition for open space on the substrate is heavily
influenced by the timing of recruitment, and this in turn is
highly dependent on temperature. Changing seasonal patterns of temperature may favour the settlement of invasive
species in a particular time of the year, and with long lasting consequences in preventing the recruitment of native
species later. The consequences of temperature change also
include vertical stability of the water column and upwelling. Changes in atmospheric circulation might also change
storm frequency and precipitation patterns and alter circulation (Astraldi et al., 1995), and hence the dispersion
routes of introduced species.
Together, increases in global mean temperatures and
elevated CO2 will result in a cascade of physical and chemical changes in marine systems. While the concentration of
CO2 has reached 378 ppm (and other greenhouse gases
have also increased dramatically), global air and sea surface temperatures have risen in the past century by 0.4–
0.8 °C (IPCC, 2001). Continued uptake of atmospheric
CO2 is expected to substantially decrease oceanic pH,
through an increase of surface-water dissolved inorganic
carbon (DIC) and a decrease of carbonate ion concentra-
tion. The dissolution of marine biogenic carbonates neutralizes anthropogenic CO2 and adds to total alkalinity.
The extent of dissolution increases as a function of the
decrease of calcium carbonate saturation state and recent
analyses have shown that much of the CaCO3 exported
out of the surface ocean dissolves in much higher horizons
than previously thought. Also, the saturation horizons of
aragonite, calcite and other minerals essential to calcifying
organisms are changing in several areas of the world
oceans. It has been shown that the calcification rate of
almost all organisms decreases in response to a decreased
CaCO3 saturation even when the saturation level is more
than one. In particular, aragonite producers (including reef
building corals) and magnesian calcite producers (like coralline algae) are expected to be more sensitive to changes in
saturation (Kleypas et al., 1999; Riebesell et al., 2000; Feely
et al., 2004).
In conclusion, climatic driven changes may affect both
local dispersal mechanisms, due to the alteration of current
patterns, and competitive interactions between NIS and
native species, due to the onset of new thermal optima
and/or different carbonate chemistry. The magnitude and
variety of climatically forced changes in the physical environment will provoke responses in the biosphere thus altering the balance of native species vs. NIS, via changes in
population size and effect of interacting species.
The effects of climatic change described above (‘‘proximate effects’’), thus lead to ‘‘emergent’’ patterns such as
changes in species distributions, biodiversity, productivity
and microevolutionary processes (Harley et al., 2006), that
are concomitant with the effects of the introduction of NIS,
especially if they reach stage 4b or 5 (dominant).
4. Impact of global change on marine ecosystems (emergent
properties)
The range expansion of populations of NIS following
climatic changes in their new environment has been proposed as an explanation of the increasing rate of successful
invasions. As said before, only in a few cases has the corresponding change in propagule pressure been taken into
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account, but evidences of changes in the geographic distribution of NIS are increasingly accompanied by observations of warming in particular areas of the sea.
As far as the Mediterranean Sea is concerned, climate
change has been suggested as being particularly important
in the establishment of alien species of plants (Gritti et al.,
2006) and the development of microalgae e.g. species of
Asterodinium (a Dinophyceae genus typical of warm
waters) (Gomez and Claustre, 2003). Recently (summer
2005 and 2006) another introduced Dinophycean, Ostreopsis armata, bloomed repeatedly in the Ligurian sea causing
respiratory diseases to tourists in several occasions. Climate change has also been suggested for the expansion of
the biogeographical range of benthic and nektobenthic
marine species, as recorded in the last decade in the western
Mediterranean (Francour et al., 1994; Vacchi et al., 1999;
Bianchi and Morri, 2000; Laubier et al., 2004) as well as
in other areas (Bianchi, 1997).
Erithrean species (of warm water affinity), established in
different phases after the opening of the Suez canal, have
caused changes in the Levantine part of the Mediterranean
far beyond recorded impacts in other marine ecosystems.
Nearly half of the fish catches along the Israeli coast consist
of Erythrean species (Goren and Galil, 2005). By the 1950s
the sudden increase in the populations of Saurida undosquamis, U. moluccensis was attributed to a rise of 1.0–1.5 °C
in sea temperature during the winter months of 1955 (see
Galil, this issue). The process has accelerated in recent
years, with increasing records of newly discovered Erythrean species and expansion towards other areas of the Eastern (Galil and Zenetos, 2002) and Western Mediterranean
(Harmelin-Vivien et al., 2005; CIESM, 2005).
While the range expansion of species correlated with
changing temperature conditions has been observed in
many cases, this has been rarely subjected to testing
hypotheses explaining this pattern. For instance, Bachelet
et al. (2004) have studied the population of the gastropod
Cyclope neritea, known from the Mediterranean and southern Atlantic coasts of the Iberian peninsula, and which
recently became established in the French coast of the
Bay of Biscayne. They considered the origin of the population, using molecular genetic analyses, and the authors
concluded that multiple introductions had occurred (probably with shipment of oysters from Portugal and the Mediterranean). Also, they studied the competition for food
resources of C. neritea with the overlapping native gastropod Nassarius reticulatus by laboratory experiments, demonstrating a superior ability of the introduced species to
search for and consume carrion. Further Nassarius reticulatus was more subject to digenean trematode parasites.
One of the most debated questions in invasion ecology
has been the relative importance of biotic processes (e.g.
competition and facilitation) of the native–invader relationship and of the abundance of resources in the invaded
habitat. This debate has often postulated that a high resident diversity implied a reduced success in the establishment of NIS (Kennedy et al., 2002).
Stachowicz and Byrnes (2006) provide an illuminating
hypothesis on the interplay of native diversity and resource
availability (free space available for colonisation on the
substratum), analysing the outcome of introduced species
establishment on hard bottoms using a wide array of experimental and observational studies. As described before, the
success of a particular invader depends not only on its
propagule pressure but also on the seasonal timing of the
recruitment: at high levels of open space there was a strong
positive relationship between native and invader richness.
The reduced invader diversity where there was plenty of
open space was likely due to non-selective disturbance
agents; these affect native and invader diversity in a similar
negative manner, resulting in native and invader richness
covarying positively as a reflection of extrinsic factors
(for instance climate).
When resources were limiting (low space availability)
experimental results showed a negative slope in the relationship between native and invader diversity, but in field
observations, the data instead showed a flat relationship.
As diversity increased, temporal complementarity among
species may have increased the consistency of space occupation and decreased the likelihood of invader establishment. The complementarity could have simply balanced
out the effects of extrinsic factors, however, the authors
observed that effects of biotic resistance were offset by the
presence of species (facilitators) that provided secondary
substrate on their shells, thus increasing the surface area
available for colonisation. In summary, an explanation of
the effects of native richness of species on invader richness,
is proposed (Table 1 in Stachowicz and Byrnes, 2006), that
takes into account availability of resources, biotic interactions and extrinsic factors, among which many could be
explained by climatic change. It would be interesting to test
the hypothesis on different assemblages, where the limiting
factors could be different, and especially where climatic
conditions are severe for at least some stages of the organism life cycle.
The most commonly predicted effect of global ocean
warming is a poleward shift in the distributional boundaries of species with an associated replacement of cold
water species by warm water species. However, Schiel
et al. (2005), on the basis of the study of communities colonizing a rocky coastal habitat influenced by the outfall of
a power generating station, rejected this prediction, observing that the responses of the communities to ocean warming were mostly unpredictable and strongly coupled to
direct effects of temperature on key taxa and to indirect
effects operating through ecological interactions. Such
direct BACI (Before After Control Impact) studies are
infrequent and probably difficult to repeat, especially in
areas much larger than the section of coast affected by local
temperature changes, so classic records of shifting biogeographic borders are still useful to assess changes that might
be related to temperature changes. For instance it has been
suggested that thermal pollution from a power plant in
Vladivostok (Russia) can render the outfall basin an inter-
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mediate step for the introduction and acclimation of tropical species, such as Balanus amphitrite and Molgula manhattensis, brought by long distance vessels (Zvyagintsev
et al., 2004).
While marine biological invasions are known to alter
nearshore benthic and pelagic communities and are displacing native species (Carlton, 1989; Carlton and Geller,
1993; Cohen and Carlton, 1998), there is no evidence yet
that aliens have caused large-scale extinctions in recipient
coastal biota (see Carlton, 1993; Leppäkoski et al., 2002;
Wolff, 2000), and more research is needed to clarify this
crucial relation (Gurevitch and Padilla, 2004; Clavero
and Garcı`a-Berthou, 2005). It must be underlined that
the decline in existing populations of native species can
be overlooked and go unnoticed in some cases, due to lack
of previous data or insufficient taxonomic difficulties, as
was suggested by Geller (1999) for the invasion of Mytilus
galloprovincialis in Southern California where it replaced
the native M. trossulus.
Besides the effects on species richness and the correlated
extinction of native species, some invasions may provoke
multiple effects which affect the overall ecosystem functioning, including factors such as material flow between trophic
groups, primary production, relative extent of organic
material decomposition, and extent of benthic-pelagic coupling. Perhaps the most conspicuous example of severe
shifts in ecosystem functioning is the invasion of the American carnivorous comb jelly Mnemiopsis leidyi in the Black
and Caspian Seas, which resulted in serious declines in zooplankton and anchovy populations, causing ecosystemwide cascading effects (Ivanov et al., 2000; Shiganova
et al., 2001). For a discussion of this case see Olenin
et al. (this issue).
Invasions by parasites caused long lasting or even irreversible consequences (Harvell et al., 2002), as in case of
Bonamia ostreae, a disease of the European native oyster
Ostrea edulis, which caused severe decline in their populations and, as a consequence, destruction of native oyster
bed ecosystems (Wolff and Reise, 2002). Other important
changes in biological communities resulting from biological invasions have occurred in estuaries, coastal lagoons
or in waters of variable salinity (Cohen and Carlton,
1998; Olenin and Leppäkoski, 1999; Nehring, 2006). Ruesink et al. (2006) calculated the increase of total primary
production in the estuarine site of Willapa Bay (Washington, USA) caused by the introduction of rooted plants
(>50%) as well as of total secondary production (250%)
due to the introduction of bivalves; they found that environments with empty niches have been colonised by species
that play novel roles in the ecosystem, with further implications for habitat, detritus and filtration.
The case of the Lagoon of Venice has been described by
Occhipinti-Ambrogi (2000). Recently Pranovi et al. (2006)
have assessed the changes in the benthic ecosystem of the
lagoon induced by introduction of Ruditapes philippinarum
and subsequent clam exploitation. Bartoli et al. (2001)
described biogeochemical cycles altered by the farming of
7
R. philippinarum in another lagoon of the Po delta
complex.
Most studies assume that invaders affect negatively
native biota, while a few others contend that aliens in
coastal waters seem to play a beneficial role in ecosystem
functioning. At least in the short term, some invasions have
resulted in a net gain in the number of species present at the
local/regional level (Sax and Gaines, 2003). Some authors
observe that NIS may often be complementary to natives
in their patterns of resource use (Olenin and Leppäkoski,
1999; Reise et al., 2006). Examples of increase of benthic
biomass production and improved filtration activity with
beneficial effects include Dreissena polymorpha in the Baltic
sea (Daunys et al., 2006) and Ensis americanus along coasts
of the North Sea (Armonies and Reise, 1998). A more
problematic view of the invasion of bivalves in an ecosystem lacking dominant filter-feeder organisms is expressed
by Alpine and Cloern (1992).
Also ‘ecosystem engineers’ (Crooks, 2002; Cuddington
and Hastings, 2004) contribute to important changes in
invaded ecosystems: large scale changes in the physical
structure of key habitats in a water body include spawning
grounds, underwater sea grass meadows, and biogenic reefs
(see Wallentinus and Nyberg, this issue). Among ecosystem
engineers, the invasion of the green algae Caulerpa in the
Mediterranean has profound effects both on the habitat
and on ecosystem functioning (see Galil, this issue and Olenin et al., this issue). Caulerpa taxifolia originated from
aquarium release on the coast of France (Jousson et al.,
1998) and has expanded to very large areas of the Ligurian
and Tyrrhenian Seas (Meinesz et al., 2001). It has been
described as ‘‘killer alga’’ (Meinesz et al., 2002) for its
alleged property to cover extensively the substrata where
it had been introduced without leaving any other previously established vegetation, and profoundly affecting all
the biota. It has been regarded as the winner in the competition with the autochtonous Posidonia oceanica that forms
the characteristic Posidonia meadows of the Mediterranean, which is now undergoing a rapid decline. Experimental data and observations in different parts of the
Mediterranean have shown that, even if direct competition
actually exists, Posidonia meadows are overgrown and
replaced by C. taxifolia when they are sparse and in a
regression stage, whereas well preserved and dense populations show a fair resistance to the Caulerpa expansion.
While a possible climate mediated influence has not been
investigated in detail (but see Peirano et al., 2005), it is
likely that the diffusion of C. taxifolia has been favoured
by some weakening of Posidonia due to pre-existing factors. Another factor that might favour C. taxifolia is its
ability to change demographic and life-history traits, as
suggested by Wright (2005) who studied populations in
the 0 stage (source populations) from the subtropical Moreton Bay (Australia) and stage 4b (locally dominant) invasive populations in temperate Southeastern Australia.
The other introduced species is Caulerpa racemosa var.
cylindracea (Verlaque et al., 2003), a south-western Austra-
Please cite this article in press as: Occhipinti-Ambrogi, A., Global change and marine communities: Alien species ..., Mar. Pollut. Bull.
(2007), doi:10.1016/j.marpolbul.2006.11.014
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A. Occhipinti-Ambrogi / Marine Pollution Bulletin xxx (2007) xxx–xxx
lian variety, which is currently achieving dramatic and continuous expansion throughout most of the Mediterranean
Sea and the Atlantic (Piazzi et al., 2001; Verlaque et al.,
2000, 2004). The growth rate of Caulerpa spp. is correlated
with temperature (Komatsu et al., 1997) and this is probably one of the reasons for the success of this species in the
Mediterranean (Raniello et al., 2004; Ruitton et al., 2005).
5. Conclusions
Marine alien species are a component of global change
in all marine coastal ecosystems. The impact of these species on native communities has been evaluated in many
localities all over the world leading to the concept of biotic
pollution. This is especially evident in the Mediterranean
Sea. The global climatic change affecting the earth’s atmospheric and oceanic system interacts in many ways with
global biogeographic changes arising from marine species
translocations. Human activities causing emission of greenhouse gases and transport of species across the oceans are
increasing at a steady pace. Moreover, the consequences on
climate and ecosystems will last for very long time, even if
human activities slow down in the near future. The interactions between the two components of global change are
potentially very important and have been studied in detail
only recently. A common framework for the study of consequences of climatic changes and CO2 increased concentration in the marine environment, and the study of the
dynamics of transport, settlement and invasion of alien
species, is necessary to understand the long term consequences for marine ecosystems, their goods and services.
Recent developments of an observational and experimental nature have highlighted the importance of considering propagule pressure as a prerequisite in all studies of
biological invasions and have indicated a way to disentangle biotic resistance by resident species and/or facilitation
by recent settlers from the effects of abiotic conditions that
can match the ecological niche of the introduced species.
Among these conditions, rapidly evolving ones are the temperature, as well as changing pH and total alkalinity. In
this changing marine landscape, deeper insight is needed
not only to monitor the state of the environment but also
to predict future changes and to mitigate and manage
them.
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Current Biology 17, 360–365, February 20, 2007 ª2007 Elsevier Ltd All rights reserved
DOI 10.1016/j.cub.2006.12.049
Report
Phase Shifts, Herbivory,
and the Resilience of Coral Reefs
to Climate Change
Terence P. Hughes,1,* Maria J. Rodrigues,1,2
David R. Bellwood,1,2 Daniela Ceccarelli,1,2
Ove Hoegh-Guldberg,3 Laurence McCook,1,4
Natalie Moltschaniwskyj,5 Morgan S. Pratchett,1
Robert S. Steneck,1,6 and Bette Willis1,2
1
Australian Research Council Centre of Excellence
for Coral Reef Studies
James Cook University
Townsville QLD 4811
Australia
2
School of Marine and Tropical Biology
James Cook University
Townsville QLD 4811
Australia
3
Australian Research Council Centre of Excellence
for Coral Reef Studies and Centre for Marine Science
University of Queensland
St Lucia
Brisbane QLD 4072
Australia
4
Great Barrier Reef Marine Park Authority
PO Box 1379
Townsville QLD 4810
Australia
5
School of Aquaculture
University of Tasmania, Locked Bag 1-370
Launceston, Tasmania 7250
Australia
6
School of Marine Sciences
University of Maine
Darling Marine Center
193 Clarks Cove Road
Walpole, Maine 04573
Summary
Many coral reefs worldwide have undergone phase
shifts to alternate, degraded assemblages because of
the combined effects of overfishing, declining water
quality, and the direct and indirect impacts of climate
change [1–9]. Here, we experimentally manipulated
the density of large herbivorous fishes to test their
influence on the resilience of coral assemblages in
the aftermath of regional-scale bleaching in 1998, the
largest coral mortality event recorded to date. The experiment was undertaken on the Great Barrier Reef,
within a no-fishing reserve where coral abundances
and diversity had been sharply reduced by bleaching
[10]. In control areas, where fishes were abundant,
algal abundance remained low, whereas coral cover
almost doubled (to 20%) over a 3 year period, primarily
because of recruitment of species that had been locally
extirpated by bleaching. In contrast, exclusion of large
herbivorous fishes caused a dramatic explosion of
*Correspondence: [email protected]
macroalgae, which suppressed the fecundity, recruitment, and survival of corals. Consequently, management of fish stocks is a key component in preventing
phase shifts and managing reef resilience. Importantly,
local stewardship of fishing effort is a tractable goal for
conservation of reefs, and this local action can also
provide some insurance against larger-scale disturbances such as mass bleaching, which are impractical
to manage directly.
Results and Discussion
The ecosystem goods and services provided by healthy
coral reefs are a key component in the economic, social,
and cultural fabric of many tropical maritime countries
[1, 9]. Until recently, land-based pollution and overfishing were considered to be the major threats to coral
reefs. Today, reefs face additional pressure from thermal stress and emergent diseases that are closely linked
to global warming [1–8]. In the most damaging case to
date, 16% of the world’s reefs were impacted in 1997–
1998 by a regional-scale bleaching event that affected
the Great Barrier Reef, vast tracts of the western Pacific,
the Indo-Australian Archipelago, and the Indian Ocean
[1, 10–11]. Climate-change projections indicate that
similar events will reoccur with increased frequency in
the coming decades [2, 12], highlighting the urgency of
developing improved tools for managing reefs in the
face of escalating threats [4–5, 13–15].
Here, we experimentally examine the resilience of
coral-dominated assemblages on the Great Barrier
Reef and the processes underlying a phase shift to
macroalgal dominance (Figure 1A). We define resilience
as the ability of reefs to absorb recurrent disturbances
(e.g., from cyclones, outbreaks of predators, or coral
bleaching events) and rebuild coral-dominated systems.
Loss of resilience can lead to a phase or regime shift to an
alternate assemblage that is typically characterized by
hyperabundances of fleshy seaweeds or other opportunistic species. The experiment was designed to simulate
the depletion of large predatory and herbivorous fishes
caused by chronic overfishing [16–21] and to investigate
their role in the regeneration of reefs after recent mass
bleaching and the mortality of corals (see Experimental
Procedures). The scale and timing of the experiment
allowed us to measure the postbleaching dynamics of
a rich coral assemblage (77 species represented by
4569 colonies were recorded by the end of the experiment), and its location provided us with a baseline
comparison of an unusually intact fish fauna on heavily
grazed reef crests within an established no-take area
of the Great Barrier Reef Marine Park. This is the first
replicated herbivore-exclusion experiment that explicitly
examines herbivore-algae-coral interactions in the context of climate change. We demonstrate that exclusion
of larger fishes profoundly erodes the resilience of coral
reefs and their ability to regenerate after bleaching, with
Phase Shifts and Resilience of Coral Reefs
361
Figure 1. Experimental Phase Shifts on the Great Barrier Reef
(A) Roofless cages and partial cages constructed on the seaward edge of reef crest. Each structure is 5 3 5 m in area and 4 m tall. Note the 2 m
high door in the cage in the center of the photograph, for access at low tide.
(B) Growths of Sargassum up to 3 m tall dwarf understory corals inside a fish-exclusion cage.
(C) When fishes were experimentally excluded, a foliose coralline alga, Mesophyllum purpurescens, replaced shallow-water grazer-resistant
species.
(D) Coral recruits settled on dead corals killed 5 years earlier by thermally induced bleaching in 1998. Grazing of the dead substrate by herbivores
is crucial for settlement and early survival of corals and coralline algae.
major implications for reef ecology, conservation, and
management.
Experimentally Induced Phase Shift
Our experimental exclusion of fishes replicated the paucity of medium and large fishes that is characteristic of
chronically overfished reefs in S.E. Asia, the Caribbean,
and elsewhere [16–19]. The biomass of herbivorous
fishes inside cages (Figure 1A) was reduced to levels
seven to ten times lower than in adjacent partial cages
and open plots (0.45 6 0.08 [S.E.], 4.29 6 2.81, and
3.12 6 1.24 kg/m2 per hr of video observation, respectively, F = 7.79, p < 0.001; Figure S1 in the Supplemental
Data available with this article online). In response to
the experimental exclusion of larger herbivorous fishes,
benthic assemblages in the cages followed a fundamentally different trajectory over time, with upright fleshy
macroalgae rather than corals and algal turfs becoming
predominant, mimicking similar responses on many
overfished and polluted reefs worldwide [4–8, 20–21].
In the aftermath of massive loss of corals on Orpheus
Island in 1998 [10], roving herbivorous fishes continued
to suppress the biomass of macroalgae and thus facilitated the recruitment of corals (Movie S1). In the partial
cages and open plots where fish grazing was uninhibited, the cover of macroalgae (primarily the calcified
red alga, Galaxaura subfruticulosa) averaged only
4.1% and 1.7% during the experimental period (n = 16
censuses), ranging up to a maximum of 10% and 7%,
respectively (Figure 2A). In contrast, algal cover in the
cages far exceeded the two control treatments throughout the experiment, reaching up to 91%, and averaging
56% 6 21% (S.E.) after 30 months (repeated-measures
ANOVA, F = 3.82, p < 0.05; Figure 2). By the end of the
experiment, algal biomass in the cages was 9 to 20 times
higher than in partial cages and open plots (1363 6 234,
146 6 49, and 68 6 28 g wet weight per m2, respectively:
ANOVA, F = 20.8, p < 0.001). Over time, the species composition of macroalgae in the cages diverged dramatically from the other two treatments (Figure S2). Dense
thickets of Sargassum, previously absent on the reef
crest, grew to 3 m in height inside the cages, with maximum densities of greater than 1000 plants (holdfasts) per
25 m2 and a biomass of up to 8.55 kg wet weight per m2
(Figure 1B and Movie S2). Cover and species composition of crustose coralline algae also diverged in the three
experimental treatments (Figure 1C and Figure S3).
Herbivory Boosts the Resilience of Coral
Assemblages to Global Warming
In tandem with the changes in macroalgae and corallines, the trajectory of coral reassembly after the 1998
bleaching event diverged markedly in the fish-exclusion
cages compared to the partial cages and open plots
(Figure 2B). Initially, the most prevalent taxa (accounting
for >80% of coral cover) were branching Porites
Current Biology
362
Figure 3. Demographic Responses of Corals
Figure 2. Contrasting Trajectories of Macroalgae and Corals after
Exclusion of Fishes
(A) Macroalgal cover. Error bars are SE.
(B) Relative coral cover over time among three experimental treatments. Absolute coral cover after 130 weeks was 7.7% 6 1.0%
(S.E.), 19.2% 6 2.3%, and 20.2% 6 2.2% in the three treatments
(see text for analysis). Census dates were the same for all treatments
and are slightly staggered in the plots for clarity. Error bars are SE.
cylindrica, massive Porites spp. (especially P. lobata
and P. rus), and massive faviids (principally heads of
Goniastrea, Favia, and Montastrea spp.) that had survived the bleaching event 2 years prior to the initiation
of the experiment. Alcyonacean soft corals and branching hard corals, particularly a diverse suite of Acropora
species, were virtually eliminated from shallow sites by
bleaching [10], and only a few small recruits were present (<0.1% cover) when the experiment began in 2000.
In the fish-exclusion cages, total coral cover grew from
6.0% 6 0.8% (S.E.) to 7.7% 6 1.0% after 30 months.
Coral cover increased much more quickly inside the partial cages and open plots, reaching 19.2% 6 2.3% and
20.2% 6 2.2%, respectively (RM-ANOVA, F = 3.82, p <
0.05). In relative terms, coral cover increased by 28% inside the cages compared to 68% for partial cages and
83% for open plots (Figure 2B).
The divergence in coral cover among treatments was
attributable to both lower recruitment and higher mortality of established corals after the experimental reductions of fish biomass (Figures 3A and 3B). A total of
(A) Recruitment of corals into the three experimental treatments.
Error bars are SE.
(B) Mortality of coral colonies originally present in cages, partial
cages, and open plots. Error bars are SE.
1062 new recruits from 26 coral genera were recorded
in the three treatments at the end of the experiment (Figure 1D). Overall, coral recruitment in cages was approximately two-thirds lower (39 6 11 recruits per 25 m2,
compared to 108 6 26 for partial cages and 118 6 21
in the open plots; ANOVA, F = 150.9, p < 0.001; Figure 3A). Acropora, which was virtually eliminated in
1998 from the reef crest at Orpheus Island by bleaching
[10], accounted for 246 of the recruits, representing 23%
of the total. The dominant adult genus, Porites, had only
two recruits in the cages, compared to 45 elsewhere (19
in partial cages, 26 in open plots; F = 12.49, p = 0.003).
Similarly, Acropora recruits were three times more
abundant in partial cages and open plots (F = 7.7, p =
0.011). In contrast, Fungia and Euphyllia were more
abundant inside cages, where together they comprised
18% of the recruits compared to only 3% in each of the
two other treatments. A principal component analysis
summarizes the striking divergence in the composition
of coral recruits in cages compared to partial cages or
open plots (Figure 4). Recruit assemblages in the partial
cages and open plots were indistinguishable.
The suppression of coral recruitment inside cages is
unlikely to have been an experimental artifact for two
reasons. First, the considerable size of the cages and
Phase Shifts and Resilience of Coral Reefs
363
Figure 4. A Principal Component Analysis Showing the Divergent
Coral Assemblages in Cages versus Other Experimental Treatments
Cages are colored blue, partial cages are colored red, and open
plots are colored black. Each symbol represents one of the 4 3 25 m2
replicates in each experimental treatment. The first two axes
explain 63% of the variation among the 12 experimental replicates.
the absence of a roof minimized caging effects (e.g., because of reduced water flow or shading from the cage
structure). Light levels in the cages supported luxuriant
algal growth, and macroalgae and juvenile fishes recruited in great numbers into them. Second, the partialcage treatment did not show an intermediate reduction
in numbers of newly recruited corals (Figure 3B). Juvenile
Fungia and Euphyllia (and the coralline alga, Mesophyllum purpurescens) are normally found in deeper water
and on shaded vertical surfaces and are rare on shallow
reef crests. Consequently, the divergent response by
juvenile corals among the experimental treatments (Figure 4) is more likely to reflect a range of tolerances to
shading by the dense stands of Sargassum than differences among experimental treatments in the rate of
delivery of larvae by currents.
Mortality rates of older coral colonies, which had survived bleaching and were already established when the
experiment began, were more than double in the cages
(24.2% after 30 months compared to 9.8% for partial
cages and 11.3% for open plots (Figure 3B; RM-ANOVA,
F = 4.29, p < 0.05). Recruitment was insufficient to
counter these losses in the cages, where the total number of coral colonies decreased by an average of 72 6 32
per 25 m2 (a 26% decline). In contrast, counts of corals
increased by 43 6 21 (16%) and 39 6 24 (14%) per
25 m2 in partial cages and open plots, respectively. In
addition to changes in mortality of corals, we also recorded significant differences in sublethal indices of
coral condition attributable to indirect impacts of herbivorous fishes (see Supplemental Data).
Conclusions
Implications for Coral-Reef Management
The spatial and temporal scales of our experiment
(300 m2, 30 months) was sufficiently large that we successfully generated a phase shift to macroalgal dominance. The increased numbers of small fishes inside
the cages may be due partially to reduced rates of predation or to the enhanced settlement and migration of
juveniles into the dense algal canopies that formed
after the exclusion of large roving herbivores. Ironically,
the small herbivores and detritivores that dominated the
cages may have promoted blooms of fleshy seaweeds
by removing filamentous epiphytes and sediment from
the surfaces of macroalgal that were too large or welldefended for them to consume. These findings provide
robust experimental evidence for trophic cascades or
top-down control—changes in the structure of foodwebs and species composition (e.g., enhanced recruitment of fishes and increased algal biomass) due to reduction in the abundance of medium and large fishes
[18, 20, 22]. After 30 months, we removed the mesh
from cages to allow entry once more to roving herbivores and predators. Cover of macroalgae in the newly
accessible cages declined rapidly because of intense
grazing, from 53% to 13% after 12 days and to approximately 0 after 30 days [23]. Juvenile fishes in the former
cages declined much faster than the algae, by 98% after
only 3 days, presumably because of predation. In the
Caribbean, Mumby et al. [24] tested the potential importance of marine no-take areas for safeguarding parrotfish and their ability to control blooms of turf and fleshy
seaweeds. They found a greater biomass of parrotfishes
and less macroalgae inside a no-take reserve, consistent with the experimental results presented here (although the abundance of adult and juvenile corals was
not reported). Our large-scale experiment provides direct evidence that overfishing of herbivores affects
more than just the targeted stocks and can also influence the resilience of coral reefs to climate change.
Process-oriented research, exemplified by the experimental manipulations presented here, provides a more
rigorous basis for coral-reef management than conventional approaches. In particular, the current focus on
descriptive mapping and monitoring of reefs needs to
be substantially broadened for better understanding of
critical processes that underlie resilience. Our results
demonstrate that loss of coral-reef resilience can be
readily quantified with several metrics (e.g., depletion
of key functional groups of fishes, reduced rates of coral
recruitment and population regeneration, sublethal impacts, etc.). Furthermore, our findings show that local
management efforts in support of resilience can afford
significant protection against threats that are much
larger in scale. Preventing coral bleaching is not a tractable management goal at meaningful spatial or temporal
scales, and a long-term solution will require global reductions of greenhouse gases over decadal timeframes.
On the other hand, supporting resilience in anticipation
of bleaching and other recurrent disturbances can be
achieved locally by changing destructive human activities (e.g., overfishing and pollution) and thereby reducing the likelihood of undesirable phase shifts. Achieving
this outcome will require the linking of ecological resilience to social and governance structures and involve
scientists, other stakeholders, environmental managers, and policy makers [25–26]. A resilience-based
approach represents a fundamental change of focus,
from reactive to proactive management, aimed at sustaining the socioeconomic and ecological value of coral
reefs in an increasingly uncertain world.
Current Biology
364
Experimental Procedures
Study Site and Experimental Treatments
The fish-exclusion experiment was undertaken on the inner Great
Barrier Reef, in Pioneer Bay on the leeward coast of Orpheus Island
(18 360 S, 146 290 E), a high-island approximately 10 km offshore
from the Australian mainland. Like many continental reefs in Australasia, the reef fauna is highly diverse, with a benthos dominated by
massive and branching scleractinians and alcyonacean soft corals.
The water is turbid (typical horizontal visibility is 5–8 m), and the
typical tidal range is 3–3.5 m. The sheltered reef fringing the lee
of the island seldom experiences breaking waves except during
rare storms and cyclones. Fishing has been banned in Pioneer Bay
since 1987.
The three experimental treatments were (1) four 5 3 5 m fullymeshed roofless cages for excluding all large and medium fishes
(Figure 1C), (2) four partially meshed cage controls that afforded access along 50% of each perimeter to control for any effects of the
caging structure, and (3) four open plots. Each of the 12 replicates
was 25 m2 in area. The cage and partial cage framework consisted
of eight 4-m-tall vertical lengths of 50-mm-diameter tubular steel
(at each corner and midway along each side), three horizontal
lengths along each side at the bottom, middle, and top, and an internal cross of tubing that connected horizontally between the four
middle vertical uprights. We anchored the vertical tubes by sliding
them over a 2 m steel bar that was hammered halfway into the substrate and cemented in place. Eight stays were also attached to each
cage to prevent them from lifting. A door to each cage (2 3 0.8 m)
provided access at all tide levels. The 4 m height of the cages and
partial cages obviated the need for a roof because they extended
a few decimeters above water at high tide, and the base always
remained submerged. The plastic mesh on cages and partial cages
(1 cm2 for the bottom 2 m, and 2 cm2 for the top 3 m) was scrubbed
every 7–10 days to prevent fouling. A weighted net sealed the bottom. After 30 months, we removed the mesh from cages and partial
cages and closely followed the immediate response of fishes and
macroalgae. Diadema sea urchins are rare at this location. Three
were removed from the cages (100 m2) at the start of the experiment.
initially in September 2000 and again in April 2003, with a grid of
100 3 0.25 m2 quadrats covering each of the 12 experimental plots.
A comparison of the two censuses yielded data on recruitment
(arrival of new colonies) and coral composition.
Coral tissue thickness, an index of biomass and physiological
condition, was measured with calipers after 2 years in 64 colonies
of Porites cylindrica from within two cages and outside. Those colonies from within cages were (1) positioned at least 10 cm away
from the nearest clump of macroalgae, (2) shaded or (3) partially
overgrown by macroalgae. Reproductive output of corals was measured in 90 experimental fragments of Montipora digitata that were
placed 17 weeks before spawning outside and within two cages, the
latter either positioned in the open or beneath clumps of macroalgae
(principally Padina). After 14 weeks, fragments were collected and
decalcified for an estimation of egg size, number of eggs per polyp,
and number of reproductive polyps.
Supplemental Data
Supplemental Data include additional Experimental Procedures,
three figures and two movies and are available with this article online
at http://www.current-biology.com/cgi/content/full/17/4/360/DC1/.
Acknowledgments
We thank the Australian Research Council for providing financial
support, the Great Barrier Reef Marine Park Authority for granting
a research permit, and a small army of student volunteers for helping with routine scrubbing of the experimental cages. W. Evans,
R. Gibson, C. Hughes, J. Madin, and A. McDonald assisted with
the construction and removal of the experiment. We are grateful
also to M.J. Boyle, D. Gibson, and A. Hoey for technical assistance.
Received: November 16, 2006
Revised: December 18, 2006
Accepted: December 18, 2006
Published online: February 8, 2007
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Numbers and Sizes of Herbivorous Fishes
The abundance of herbivorous fishes in each treatment was measured after 28 months with 90 hr of high-resolution remote video.
In each of the 12 cages, partial cages, and open plots, five randomly
placed 1 m2 quadrats were censused for 90 min with remote video
cameras (so that diver effects could be eliminated). Recording was
undertaken within 90 min of high tide between 1000 and 1600, with
randomly allocated times among treatments. Fish were identified
to species, and their body lengths were recorded and converted
to biomass with standard length-weight regressions.
Response of Macroalgae and Coralline Algae
Fleshy macroalgae and noncoralline crusts were identified in situ to
genus, and their abundance in each of the 12 plots was estimated on
a six-point categorical scale (0 = ‘‘absent,’’ 1 = ‘‘rare,’’ 2 = ‘‘uncommon,’’ 3 = ‘‘common,’’ 4 = ‘‘abundant,’’ and 5 = ‘‘dominant’’). A total
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experimental areas. In addition, all corals greater than 1 cm in the experiment were identified (to species, where possible) and mapped
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Tourism Management 28 (2007) 570–579
www.elsevier.com/locate/tourman
Research article
Implications of climate and environmental change for nature-based
tourism in the Canadian Rocky Mountains: A case study of
Waterton Lakes National Park
Daniel Scott, Brenda Jones, Jasmina Konopek
Canada Research Chair in Global Change and Tourism, Faculty of Environmental Studies, University of Waterloo, 200 University Avenue West,
Waterloo, Ont., Canada N2L 3G1
Received 21 April 2006; accepted 26 April 2006
Abstract
In western North America, Rocky Mountain national parks represent a major resource for nature-based tourism. This paper examines
how climate change may influence park tourism in the Rocky Mountain region by focusing on both the direct and indirect impacts of
climate change for visitation to Waterton Lakes National Park (WLNP) (Alberta, Canada). A statistical model of monthly visitation and
climate was developed to examine the direct impact of climate change on visitation. The model projected that annual visitation would
increase between 6% and 10% in the 2020s and between 10% and 36% in the 2050s. To explore how climate-induced environmental
change could also indirectly affect visitation, a visitor survey was used (N ¼ 425). The environmental change scenarios for the 2020s and
2050s were found to have minimal influence on visitation, however the environmental change scenario for the 2080s (under the warmest
climate change conditions) was found to have a negative effect on visitation, as 19% of respondents indicated they would not visit the
park and 37% stated they would visit the park less often. The contrasting result of the two analyses for the longer-term impact of climate
change was a key finding. The management implications of these findings and methodological challenges associate with climate change
impact assessment for tourism are also discussed.
r 2006 Elsevier Ltd. All rights reserved.
Keywords: Climate change; Environmental change; National parks; Tourism; Canada
1. Introduction
Nature-based tourism is an important element of the
tourism industry in North America. National parks in
Canada and the United States are central components of
this tourism market (Jones & Scott, 2006). Eagles,
McLean, and Stabler (2000) estimated that there were over
2.6 billion visitor days in parks and protected areas in
Canada and the United States in 1996, 300 million of which
occurred in national parks.
A sizable share of park tourism in North America is
concentrated in national parks located in the mountainous
regions of western Canada and the United States. In
Canada, approximately 65% (or 10 million visits) of all
Corresponding author. Tel.: +519 888 4567x5497.
E-mail address: [email protected] (D. Scott).
0261-5177/$ - see front matter r 2006 Elsevier Ltd. All rights reserved.
doi:10.1016/j.tourman.2006.04.020
national park visits in 2003 occurred in the six national
parks located in the Rocky Mountains (Parks Canada,
2004); 10 million people visited the four US-based Rocky
Mountain national parks during the same year, representing 17% of all national park visits in the United States in
2003 (National Park Service, 2005). The economic impact
of tourism in the Rocky Mountain region is equally
significant. Tourism-related expenditures by visitors to
three mountain national parks in the Province of Alberta
(Banff, Jasper and Waterton Lakes) were estimated to
exceed US$765 million (Alberta Economic Development,
2000). In the United States, Yellowstone National Park
generates an estimated US$2 billion in economic benefits
for the states of Montana, Wyoming and Idaho (Gourley,
1997).
Climate directly affects nature-based tourism by limiting
when specific recreation and tourism activities can occur
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(e.g. season length with snow cover or open water),
recreation/tourism demand (e.g. proportion of people
willing to swim or camp under certain conditions), and
the quality of a recreation/tourism experience (utility) (e.g.
hiking in warm, sunny conditions versus a cold rain or
extreme heat).
In the Rocky Mountain region, this climatic influence
manifests itself through a marked seasonality in park
visitation; nearly two-thirds (64%) of annual person visits
in Canada’s six mountain parks occur during the traditional warm-weather months of May–September (Parks
Canada, 2004). Scott and Suffling (2000) and Scott (2003)
argue that any direct changes in the length and quality of
warm-weather tourism seasons induced by global climate
change could present opportunities to increase visitation in
national parks in this region. Visitor increases precipitated
by a warmer climate would have benefits for park revenues
and the economies of gateway communities near each park,
but could also exacerbate visitor-related environmental
pressures in some high-visitation mountain national parks
in Canada and the United States.
Climate indirectly affects nature-based tourism by
impacting the physical resources that define the nature
and quality of natural environments on which mountain
tourism depends (i.e. climate-induced biophysical change).
Any changes in the natural characteristics of mountain
environments could negatively influence tourism by reducing the perceived attractiveness of the region’s mountain
parks (Elsasser & Bürki, 2002; Scott, 2003; Wall, 1992).
For example, drought conditions during the summer of
1988 contributed to widespread forest fires in Yellowstone
National Park, which resulted in evacuations of campgrounds and seasonal visitor accommodations being closed
4 weeks earlier than normal (Franke, 2000). Total annual
visits to Yellowstone in 1988 were reduced 15% (compared
to 1987) and park officials estimated that the forest fires
resulted in a loss of tourism-related economic benefits of
US$60 million (Franke, 2000).
Progress has been made in documenting climate-induced
biophysical changes in the mountain region of western
North America and a number of studies have examined the
potential biophysical impacts of climate change to the end
of the 21st century. Analysis of biome-scale vegetation
modeling suggests that under climate change mountain
parks in this region would experience both latitudinal and
elevational environmental changes with the potential for
species reorganization and loss of biodiversity (Scott,
Malcolm, & Lemieux, 2002). In an analysis of Glacier
National Park (Montana), Hall and Farge (2003) projected
that forests would advance upslope approximately 20 m per
decade through 2050. While similar advancements in the
tree line are projected for Yellowstone National Park, the
results of vegetation modeling further suggested that the
range of high-elevation species would decrease, some tree
species would be regionally extirpated and new vegetation
communities not currently found in the park would emerge
(Bartlein, Whitlock, & Shafer, 1997).
571
Glaciers in western North America are important tourist
attractions for mountain parks, but they have been
retreating over the past century and are projected to
continue to do so under climate change. Glacier National
Park has lost 115 of its 150 glaciers over the past century
and scientists estimate that the remaining 35 glaciers will
disappear over the next 30 years (Hall & Farge, 2003).
Similar projections have been made for glaciers in
Canada’s Rocky Mountain parks with glaciers less than
100 m thick projected to disappear over the next 30–40
years (Brugman, Raistrick, & Pietroniro, 1997). If such
glacier retreat is realized in the Rocky Mountains, Glacier
National Park would lose its namesake and the very
resource that defines it. Scott (2005) also argued that the
projected glacial retreat in Canada could severely hinder
the snocoach tours (specially designed buses take visitors
onto the Athabasca glacier) that attract 600,000 visitors
annually to the Columbia Icefields between Banff and
Jasper national parks.
Mountain ecosystems depend on fire for regeneration,
but forest fires pose a threat for park tourism. Under
climate change, the frequency and severity of forest fires in
the mountainous region of western Canada is projected to
increase. Vegetation and fire behavior modeling by Li,
Flannigan, and Corns (2000) suggested that west-central
Alberta would experience a one-point shift in the fire fuel
moisture code under a mean temperature increase of 4 1C
(+7.2 F), which would contribute to an increase in the
frequency of fires that burn more than 1000 ha. Stocks
et al. (1998) projected that the geographic area in western
Canada currently designated as having an ‘extreme’ fire
danger would expand in the 2050s. Similar regional
projections were made by Flannigan et al. (2001) and
Weber and Flannigan (1997). If the fire season becomes
more severe under climate change, it is possible that visitors
to mountain parks may experience more restrictions on
their activities (e.g. campfire bans; trail and park closures).
As lakes and streams warm, temperature-induced
habitat loss and range shifts are projected to occur,
contributing to losses in recreationally valued fish populations. Research on the thermal habitat for salmonid species
in the Rocky Mountain region of the United States found
that a projected 4 1C (+7.2 F) summer warming in the
region would reduce habitat area by 62% (Keleher &
Rahel, 1996). Simulation studies of cold-water fish habitats
revealed that the southern boundaries of some cold-water
fish in the United States could move 500 km northward
under climate change (Magnuson, 1998). Increases in lake
and river temperatures could place pressure on cold-water
fish species in the Rocky Mountain region, thus providing
opportunities for the geographic expansion of cool- and
warm-water species that have higher temperature tolerances.
There remains a great deal of uncertainty as to how
potential climate change impacts on seasonality and
mountain landscapes would impact park visitation around
the world and in western North America. In the only
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empirical study to examine the potential impact of climate
change on mountain park tourism, Richardson and
Loomis (2004) assessed the direct and indirect impacts of
climate change on visitation to Rocky Mountain National
Park (RMNP) (Colorado, USA). Regression analysis of
historical monthly visitation data (1987–1999) and four
climate variables for the park’s peak and shoulder seasons
was used to model the current influence of climate on park
visitation and the projected changes under climate change
for the 2020s. Richardson and Loomis (2004) found that
visitation to RMNP would increase 7–12% in the 2020s as
the warm-weather tourism season was extended. Visitors to
the park were also surveyed to determine how their
visitation patterns (number of visits and duration of stay)
might change under a range of early environmental change
scenarios. Although hypothetical, the environmental
change scenarios were partially developed on the basis of
climate change studies of the potential environmental
impacts in the park. Environmental changes in the 2020s
were found to have minimal effect on visitor behavior, as
visitation was projected to increase between 10% and 14%
under the two environmental scenarios used in the visitor
survey.
With the results of their regression analysis of visitation
data and tourist survey projecting similar increases in park
visitation, Richardson and Loomis (2004) concluded that
climate change through the 2020s would have a positive
effect on park visitation in the Rocky Mountains, and
suggested that the results for RMNP would be representative for a number of mountain parks in the United States
(Glacier, Grand Teton and Yellowstone) and Canada
(Banff, Jasper and Waterton Lakes).
The Richardson and Loomis (2004) study only examined
the potential impacts of early stages of climate change
projected for the 21st century and thus the implications of
much greater warming and environmental change projected for the latter half of the century (IPCC, 2001) remain
an important knowledge gap. While climate change
scenarios for the 2020s have the most relevance for
contemporary tourism planning and there remains substantial uncertainty about the magnitude of long-term
climate change projections (i.e. 2050s, 2080s) (IPCC, 2001),
the potential implications of longer-term changes should be
explored for strategic relevance to park managers, the
tourism industry and broader climate change policy (i.e.
the costs of impacts verses mitigation). It remains uncertain
whether Richardson and Loomis’ (2004) findings would be
representative of the potential influence of long-term
climate change on mountain park visitation. Will climate
change-induced seasonality and environmental changes
continue to have a positive effect on visitation or will
environmental changes projected for mountainous regions
in western North America towards the end of the 21st
century begin to have a negative impact on visitation?
This paper presents an empirical assessment of climate
and visitation in the Canadian Rocky Mountains in order
to determine how visitors may respond to future climate
change and related environmental change. Using a similar
methodological approach to Richardson and Loomis
(2004), the specific objectives of the study were to: (1)
develop a model of climate and park visitation in order to
examine the implications of changed climatic conditions on
future visitation levels and seasonal visitation patterns; (2)
survey park visitors to determine how a range of potential
climate-induced environmental changes may influence their
intention to visit the park and visitation frequency; and, (3)
compare the implications of direct climatic changes and
indirect climate-induced environmental changes on park
visitation at three time steps commonly used in climate
change impact assessments (2020s, 2050s and 2080s).
Conceptually, the analyses of the potential direct and
indirect impacts of climate change are considered separate
and therefore the respective methodologies and results are
presented individually and then compared in the final
discussion section.
2. Data and research methods
The focus of this study is Waterton Lakes National Park
(WLNP), which is located along the foothills of the
Canadian Rocky Mountains in southwestern Alberta
adjacent to the Canada–US international border. WLNP
is described by Parks Canada as a place ‘where windswept
mountains rise abruptly out of gentle prairie grassland’
(Parks Canada, 2005). WLNP, the smallest of Canada’s six
Rocky Mountain national parks (525 km; 130,000 acres),
receives nearly 400,000 visitors annually (Parks Canada,
2004); 84% of annual visitation occurs between May and
September, clearly demonstrating the seasonal influence of
climate. WLNP is also bordered on the south by Glacier
National Park (GNP) in the state of Montana, and
together the two parks make up the Waterton-Glacier
International Peace Park (WGIPP). Established in 1932,
WGIPP protects a diverse ecosystem comprised of grasslands, pine and aspen forests, alpine meadows and glaciercovered mountains.
2.1. Climate and park visitation
Monthly recorded visitation data (number of person
visits) from WLNP for the 1996–2003 (January–December)
tourism seasons were used in this study to assess the
influence of climate on visitation. Analysis of the direct
influence of climate on visitor levels was undertaken using
data from the nearest Meteorological Service of Canada
climate station that contained suitable monthly observations for 1996–2003 for use in modeling the relationship
between current climate and visitation and a quality
historical climate record (i.e. 1961–1990)1 to establish a
climatic baseline for the climate change impact assessment.
1
Climate data met quality standards for inclusion in the Meteorological
Service of Canada’s national archive and there were no prolonged periods
of missing data.
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Monthly level temperature and precipitation data were
obtained for the airport station in the City of Lethbridge
(Alberta), located approximately 125 km northeast
(79 mile) of WLNP.
The climate change scenarios used in the analysis were
developed from monthly global climate models (GCM)
available from the Government of Canada’s Climate
Impacts and Scenarios (CCIS) Project. The scenario data
available from CCIS are constructed in accordance with
the recommendations set out by the United Nations
Intergovernmental Panel on Climate Change (IPCC) Task
Group on Scenarios. Climate change projections for three
future timeframes were examined, each of which were
based on a 30-year period of climate data (i.e. the 2020s
represent the period 2010–2039; the 2050s represent
2040–2069; and, the 2080s represent 2070–2099). All
scenarios represent climate changes with respect to the
30-year baseline climate (1961–1990).
In accordance with IPCC guidelines for climate change
impact assessments, more than one GCM and greenhouse
gas (GHG) emission scenario were used in this study to
represent uncertainty in future climatic conditions in the
study area. The two climate change scenarios utilized in
this study were the National Center for Atmospheric
Research (NCAR) GCM with a B2 emission scenario
(a low greenhouse gas emission future) and the Center for
Climate System Research (CCSR) GCM with an A1
emission scenario (a high greenhouse gas emission future).
The two climate change scenarios used in this study
(NCARPCM B21 and CCSRNIES A11) were selected
from among 19 available scenarios; the NCARPCM B21
scenario generally projects the smallest increase in annual
mean temperature in the region this century (+2 1C; 3.6 F),
while the CCSRNIES A11 scenario projects the largest
increase (+7 1C; 13 F).
Visitor projections for WLNP under a changed climate
were undertaken using statistical analysis. First, multivariate regression analysis using three monthly level
temperature variables (maximum, minimum and mean
temperature) and one monthly level precipitation variable
(total precipitation) was employed to establish the nature
and strength of the empirical relationship between climate
and monthly person visits during the 1996–2003 tourism
seasons. The resultant regression model was then applied
to monthly climate data from the Lethbridge climate
station for the 1961–1990 baseline to establish visitation in
a climatically ‘normal’ year, against which future climate
change scenarios (also averages of 30-year model runs)
would be compared. This step should not be interpreted as
an attempt to model past visitation to WLNP between
1961 and 1990. Rather, the purpose of running the
regression model with climate data from 1961 to 1990 is
to establish visits in a ‘climatologically average’ year during
the baseline period; 30 years is the standard used by
climatologists to establish ‘current’ climatic averages.
Under the assumption that current visitation patterns
would remain unchanged in the future (i.e. the analysis
573
does not account for other factors that influence visitation—population growth, transportation costs), the regression model was then run with both climate change
scenarios (NCARPCM B21 and CCSRNIES A11) for the
2020s, 2050s and 2080s to assess the potential impact of
climate change on the number and seasonal pattern of
visitation in WLNP with respect to the 1961–1990 baseline.
Visitor data made available for this analysis from Parks
Canada provided counts of total visits and did not
distinguish the origin of visitors (i.e. local, regional or
international) or the activities visitors participated in.
Without specific information correlating activities (e.g.
skiing, hiking) with visitors, the projections put forward in
this study will emphasize the implications of climate change
for seasonal visitation resulting mainly from a longer and
improved season for warm-weather tourism activities. The
winter tourism season is less important to WLNP (less than
10% of total annual visitation) than the warm-weather
tourism season, so although snow and ice-based recreation
is anticipated to decrease under a warmer climate (Scott,
2003, 2005), losses are expected to be compensated for
through larger gains in warm-weather recreation.
2.2. Environmental change and visitation
Any projected change in WLNP’s seasonal pattern of
visitation resulting from a changed climate will not occur in
isolation, as visitation will also be indirectly influenced by
climate-induced changes in the natural environment that
nature-based tourism depends on. Analysis of the indirect
impact of climate change on park visitation was undertaken in this study through a visitor survey that explored
how climate change-induced environmental change in
WLNP might affect visitor behavior.
Three environmental change scenarios were developed for
WLNP to reflect the types of environmental changes and the
magnitude of change anticipated to occur in the Canadian
Rocky Mountains over the next century. A range of
environmental changes identified in Scott and Suffling’s
(2000) climate change impact assessment of Canada’s
national parks were considered in each scenario, including
wildlife and vegetation compositions, forest fire occurrences,
the number of glaciers lost, water temperatures and the
probability of campfire bans. The scenarios were hypothetical, but were based on available scientific literature on
documented environmental changes in the Rocky Mountains
of western North America and projected biophysical changes
from region-specific climate change impact assessments in the
scientific literature (Brugman et al., 1997; Hall & Farge, 2003;
Stocks et al., 1998), and other studies of environmental
changes in the Rocky Mountains (Harding & McCullum,
1997; McCarty, 2001; McDonald & Brown, 1996; Rhemtulla,
Hall, Higgs, & MacDonald, 2002).
Once created, each scenario provided a plausible story
about the anticipated environmental changes in WLNP in
the 2020s, 2050s or 2080s (under the warmest climate
change scenarios) with respect to current conditions.
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D. Scott et al. / Tourism Management 28 (2007) 570–579
Fig. 1. WLNP visitor survey—environmental change scenarios. Note: Respondents were not provided any time frame for the three scenarios, but scenario
1 approximates the 2020s, scenario 2 approximates the 2050s and scenario 3 approximates the 2080s.
Unlike Richardson and Loomis (2004) who examined
implications of potential environmental changes into the
2020s, this study continued to further examine how
projected climate change into the 2050s and 2080s might
affect tourist behavior. Survey participants were asked to
reflect on each scenario as a holistic package of environmental changes and consider whether they would still visit
WLNP if the identified changes occurred, and, if so,
whether they would visit more or less frequently. Participants were also asked whether they would visit a mountain
park other than WLNP if the types and magnitude of longterm environmental changes identified for WLNP were not
occurring elsewhere (i.e. destination substitution). Participants could not respond differently to individual potential
changes (i.e. warmer lake temperatures were desirable, but
loss of glaciers was unacceptable). The scenarios provided
in the survey instrument are presented in Fig. 1.
The presentation of the environmental change scenarios
in the visitor survey was similar to Richardson and Loomis
(2004) to provide consistency for comparison of results and
because pre-tests suggested that graphics and numerical
change estimates would be easier for respondents to
interpret than a detailed text-based scenario. The participants in the WLNP survey were not informed of the time
period that each environmental change scenario represented (i.e. changes under the warmest climate change
scenario for the 2020s, 2050s or 2080s). This was done to
avoid biasing the participant’s responses (e.g. ‘I will not be
alive in 2080, so these impacts are less important or will not
change my intention to visit’). Before the survey was
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D. Scott et al. / Tourism Management 28 (2007) 570–579
administered, it was pre-tested (N ¼ 30) and appropriate
revisions were undertaken.
The visitor survey was administered in WLNP during the
summer of 2004 (July 24–August 20) at a variety of
locations throughout the park including campgrounds,
scenic rest stops, public beaches, backcountry hiking areas
and visitor parking lots. A range of survey locations was
selected with the purpose of identifying visitors engaged in
different recreation and tourism activities. At each location, visitors were approached randomly (i.e. every third
person or group) and informed of the study’s purpose and
asked if they were willing to complete the survey. If visitors
were in a group, the person with the birthday closest to the
survey date was asked if they would participate. The
random sampling approach ensured that each visitor had
an equal opportunity to participate. Willing participants
took the survey with them and were asked to return it by
mail in the pre-paid return envelope provided. Mail-return
surveys were used in order to give respondents adequate
time to consider the three environmental change scenarios
and so as not to interfere with people’s activities. A total of
800 surveys were distributed in WLNP; 425 usable surveys
were returned (53% response rate).
3. Results
3.1. Direct influence of climate and climate change on
visitation
Similar to other national parks in the Rocky Mountains
of western North America, there is marked seasonality to
WLNP’s visitation. Fig. 2 illustrates monthly visits to
WLNP for the 2000–2003 tourism seasons. Person visits
were found to be highest during the summer months of July
and August (110,000 per month) when most Canadians
have school or work-related vacations (i.e. institutional
seasonality). The winter period between November and
March traditionally experienced the lowest levels of
visitation (6000 visits per month).
Comparison of WLNP’s tourism seasons revealed some
notable inter-annual variability in visitation. Annual
person visits declined 8% in 2003 over 2002 with the most
significant monthly reductions occurring in July (7%),
Monthly person visits
120,000
2000
2001
2002
2003
100,000
80,000
60,000
40,000
20,000
0
J
F
M
A
M
J
J
A
S
Fig. 2. WLNP visitation, 2000–2003.
O
N
D
575
August (17%) and September (15%). Officials with
Parks Canada have suggested that the reductions were in
response to forest fires in a number of mountain parks
during the hot and dry summer of 2003 in western North
America. Although WLNP did not experience fires, Parks
Canada acknowledged that fire-related park closures (e.g.
trails, access roads) and wide-spread media coverage of the
fires in nearby Canadian (e.g. Kootenay, Jasper) (Parks
Canada, 2003; Winks, 2003) and US (e.g. Glacier National
Park) (Mann, 2003; Newhouse, 2003) mountain parks
deterred visitors from WLNP.
Although institutional seasonality (i.e. summer school
holidays) plays an important role in seasonal visitation to
WLNP, Fig. 2 demonstrates that climate (i.e. natural
seasonality) is also important factor. A primary objective
of this study was to estimate the potential direct impact of
climate change on annual person visits to the park. In order
to assess the potential impact of climate change on
visitation, a multivariate linear regression analysis involving four monthly climate variables (maximum, minimum
and mean temperature and total precipitation) was undertaken. The regression analysis resulted in a one-variable
model (r2 ¼ 0:67) with minimum temperature being identified as the strongest predictor of monthly visits
t-statistic ¼ 13.83; po0:005). Fig. 3 (Graph A) demonstrates the nature and strength of the relationship between
total monthly visits and minimum temperature in WLNP.
Closer examination of the linear regression model in
Fig. 3 (Graph A) revealed two important limitations. First,
the regression model continued to project an increase in
visits as minimum temperature increased. Conceptually, at
some critical temperature conditions would likely become
too uncomfortable for most people (i.e. heat stress), and
rather than increasing, visitation would begin to stabilize
and then decline. Similarly, at some minimum temperature,
visitation would decline to zero. Second, the relationship
between temperature and visitation to WLNP is clearly not
linear and distinct climate–visitor relationships are evident
at different times of the year. July and August are
characterized by visitation levels in excess of 80,000 and
little variation with temperature, while visitation in all
other months was less than 60,000 but varied substantially
with temperature. Due to the combined effect of natural
and institutional seasonality, a second regression analysis
was undertaken in which visitation in WLNP’s peak
(defined as July and August; visits 480,000) and shoulder
(defined as September–June; visits o80,000) tourism
seasons was modeled separately.
Fig. 3 (Graphs B and C) illustrates the nature and
strength of the regression models for WLNP’s peak and
shoulder seasons. Minimum temperature was found to be
the strongest predictor of visits in each season, with a strong
relationship in the shoulder season (r2 ¼ 0:94), but a weak
relationship during the peak season (r2 ¼ 0:01). Since the r2
value in the peak summer season was low (0.01), future
changes in minimum temperature are projected to have
minimal impact on visitation to WLNP during the peak
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D. Scott et al. / Tourism Management 28 (2007) 570–579
Monthly person visits
120,000
linear, year-round model
July and
August
100,000
80,000
60,000
40,000
20,000
r2 = 0.67
0
-25
-20
-10
-5
0
5
10
CCSRNIES A11
100,000
80,000
60,000
40,000
20,000
J
F
M
A
M
J
J
A
S
O
N
D
Fig. 4. Projected changes in seasonality in WLNP under climate change
(2050s).
60,000
50,000
40,000
30,000
20,000
10,000
r2 = 0.94
0
-25
-20
-15
-10
-5
0
5
10
15
Minimum temperature (degrees C)
(B)
125,000
Monthly person visits
NCARPCM B21
0
15
cubic, shoulder season model
70,000
linear, peak season model (July and August)
110,000
95,000
80,000
65,000
r2= 0.01
50,000
5
(C)
1961-90
Minimum temperature (degrees C)
(A)
Monthly person visits
-15
120,000
Monthly person visits
576
6
7
8
9
10
11
12
13
Minimum temperature (degrees C)
Fig. 3. Linear and cubic regression models for WLNP.
season. The regression expressions used in this study were:
Shoulder season ðSeptember2JuneÞ ¼ 3:95x3 þ 194:34x2
þ 2903:20x þ 19; 536,
Peak season ðJuly and AugustÞ ¼ 230:59x þ 102; 253.
When the regression models were run with the two climate
change scenarios, the resulting projections suggested that
WLNP would expect to experience an increase in visitation
under a warmer climate relative to baseline conditions. In
the 2020s, total annual visits were projected to increase 6%
(NCARPCM B21) to 10% (CCSRNIES A11) over the
modeled 1961–1990 baseline average (418,358 person visits
annually). These results were consistent with Richardson
and Loomis’ (2004) regression analysis of visitation (+7%
to 12%) in RMNP under their 2020s climate change
scenarios. In the 2050s, increases were projected to range
between 10% (NCARPCM B21) and 36% (CCSRNIES
A11), which translates into an additional 42,000–152,000
person visits to WLNP annually. The strong growth in
visitation in the shoulder season resulting from climate
conditions becoming more suitable for warm-season tourism
during the spring (April–June) and fall (September–November) could lead to an extended peak tourism season in
WLNP (Fig. 4). In the 2080s, the number of people visiting
WLNP was projected to increase 11% over current baseline
conditions under the least-change scenario (NCARPCM
B21) and increase 60% under the warmest scenario
(CCSRNIES A11).
If these findings are suggestive of the long-term effects of
a more suitable climate for tourism in WLNP, to say
nothing of future increases from population growth in
Canada and the United States (which were not included in
this analysis), the implications for visitation and park
management are substantive, but can be interpreted either
positively or negatively. As for the benefits, elevated visitor
levels would result in higher revenues for Parks Canada,
primarily from the collection of additional entrance and
recreation service fees in WLNP. The Town of Waterton
(population of 150), located inside WLNP, and communities around the park would also benefit from higher
visitor levels as long as growth could be achieved in a
sustainable manner. Higher visitation and a longer tourism
season would also contribute to higher operating costs and
enhance existing visitor-related environmental pressures.
Increased need for services (e.g. campgrounds) and visitor
management could put stress on existing staff resources in
WLNP. The additional stress placed on existing park
infrastructure such as roads, trails, water systems and
waste management could also lead to increased maintenance costs, as it is anticipated that infrastructure will
need more frequent upgrades and/or repair with higher use.
Implications of higher visitation for the park’s ecological
integrity mandate remain uncertain.
3.2. Influence of climate change-induced environmental
change on visitation
The results of the above analysis of the potential direct
impact of a changed climate suggested that WLNP could
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D. Scott et al. / Tourism Management 28 (2007) 570–579
experience substantial increases in visitation as the climate
in the Canadian Rocky Mountains becomes more favorable for warm-weather tourism and recreation. The
regression model only considered the implications of a
changed climate for visitation to WLNP, yet the park is
also anticipated to undergo concurrent biophysical changes
as the climate warms. To explore how park visitation may
be affected by climate change-induced environmental
change a survey was administered in WLNP that asked
visitors to consider the frequency of future visits under
three hypothetical scenarios of environmental change in the
park.
More than 75% of survey respondents indicated that
viewing mountain landscapes and viewing wildlife were
either ‘important’ or ‘extremely’ important to their decision
to visit WLNP. By contrast, less than 15% of respondents
indicated that recreational activities including golf, fishing
and boating were important for motivating them to come
to WLNP. These results suggest that WLNP’s mountain
landscape is a critical factor in attracting visitors, and any
environmental changes that diminish that landscape could
have a negative effect on park tourism.
Fig. 5 summarizes the projected impact of climate
change-induced environmental change in WLNP on park
visitation. After considering the environmental changes
outlined in scenario 1 (2020s), the majority of respondents (99%) indicated that they would still visit WLNP and
10% indicated that they would visit more often. These
results were consistent with Richardson and Loomis’
(2004) analysis of visitation in RMNP under their 2020s
scenarios. In WLNP, a majority of respondents (97%)
again indicated that they would visit the park if the
environmental changes in scenario 2 (2050s) were
realized, however 14% of those who said they would still
visit, would visit less often.
For many respondents, the environmental changes
described in scenario 3 (2080s) represented a critical
threshold of acceptability. Of the 425 respondents, 19%
indicated that they would no longer visit WLNP if the
environmental changes in scenario 3 were realized. An
additional 37% of respondents indicated they would visit
less often. With most respondents indicating that they
would not visit or would visit the park less often, it is
possible that the considerable environmental changes
projected to occur later this century may contribute to
More often
Same amount
Less often
Won't visit
Scenario 3 (~2080s)
Scenario 2 (~2050s)
Scenario 1 (~2020s)
0
10
20
30
40
50
60
70
80
90
100
Percent (%)
Fig. 5. Projected impact of environmental change on WLNP visitation.
577
reduced annual visitation at WLNP. This finding represents a notable contrast to the large increase in visitation
projected by the previous analysis that only considered the
impact of a direct change in climate later in the century (i.e.
14–60% increase in the 2080s). The results of the survey
suggest that long-term environmental changes may diminish the attractiveness of WLNP’s landscape to some
visitors and offset some of the potential gains in visitation
made possible by an extended and climatically improved
warm-weather tourism season.
The projected environmental changes in WLNP later this
century could negatively impact visitation by reducing the
attractiveness of the park landscape and thereby providing
a comparative advantage to other mountain parks where
the impacts of climate change may be less pronounced.
When respondents were asked if they would visit another
mountain park in the region not experiencing the changes
in scenario 3 (2080s), 34% of respondents indicated they
would go to the alternative destination.
4. Discussion and concluding comments
This case study examined the potential impact of climate
change on nature-based tourism in WLNP, one of
Canada’s Rocky Mountain national parks. The study used
two separate methodological approaches to examine the
potential direct and indirect impacts of climate change on
park visitation. A multi-year data set of monthly observed
visitation were used to develop regression-based climate–
visitation models in order to explore how a changed
climate could directly impact the timing and number of
annual visitors to WLNP. A visitor survey was then used to
explore how climate change-induced changes to WLNP’s
natural landscape might indirectly influence future
visitation.
The climate-visitation model projected that annual
visitation to WLNP would increase under all of the climate
change scenarios examined, but particularly under the
warmer scenario (CCSRNIES A11). The direct affect of a
changed climate was projected to increase visitation to
WLNP by 6–10% in the 2020s. The findings for the 2020s
are consistent with Richardson and Loomis (2004) and
together are likely representative of other Rocky Mountain
national parks in Canada (e.g. Banff, Jasper and Waterton
Lakes) and the United States (e.g. Glacier, Grand Teton
and Yellowstone). If climate change scenarios for the 2050s
were realized, visitation to the park was projected to
increase 10–36%. For the same time periods (2020s and
2050s), the visitor survey revealed that climate changeinduced environmental changes would likely have minimal
impact on visitation, as most respondents (490%) indicted
that they would not change their intentions to visit WLNP
or the frequency of their visits. As such, the direct effect of
climate change in lengthening and improving the warmweather tourism season would appear to be the dominant
impact on visitation in the early to mid-decades of the 21st
century.
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D. Scott et al. / Tourism Management 28 (2007) 570–579
A key finding was the contrast between the climatevisitation model and the visitor survey with regard to the
impact of climate change late in the 21st century. The
climate-visitation model projected the direct impact of a
changed climate would increase visitation to the park
(+11% to 60%—2080s), but the survey found that the
indirect impact of climate-induced environmental change
in the park might reduce visitation, with 56% of
respondents indicating that they would no longer come to
the park or would visit less often if the environmental
changes in scenario 3 (2080s warmest scenario) were
realized.
Although climate-induced environmental change was
found to have a potentially important negative impact on
visitation under scenarios for the latter part of the 21st
century, caution needs to be taken when interpreting these
survey results. There is much greater uncertainty in longerterm scenarios because of greater uncertainties in climate
change projections (IPCC, 2001) and how society and
tourist preferences may evolve over such long timeframes.
As such, while the environmental changes posed in scenario
3 may result if the warmest (high emission) climate change
scenarios for the 2080s were realized, if a low-emission
future were realized, environmental change closer to that
outlined in scenario 2 may occur. Notably, scenario 2 has
virtually no impact on visitors’ intentions to visit the park.
The long-time frames involved also present critical
methodological challenges. Although respondents were
not given the time frames of the scenarios in the survey,
the large magnitude of environment change portrayed in
scenario 3 would take several decades to manifest (2080s
or later). The majority of people visiting the park in the
2080s would not be born until the 2040s. By this time, some
of the environmental changes projected (e.g. disappearance
of the glaciers) may have already occurred and consequently the sense of loss felt by a respondent in 2005 may
not be shared by a visitor born in the 2040s. It currently
remains uncertain whether these ‘future visitors’ will be
deterred from visiting mountain parks in western North
America if they have never experienced the landscape
attributes that current visitors use to define and measure
the quality of mountain experiences. The lived experience
of visitors in the 2080s will be very different (perhaps
unimaginably so) than those in 2005, and therefore it
remains uncertain whether the negative impact of environmental change on visitation later in this century would
occur to the extent the survey results suggest.
Understanding the behavior of ‘future tourists’ is an
important conceptual barrier for climate change impact
and adaptation studies in the tourism sector (and other
economic sectors—e.g. farmers, planners) to overcome.
Regardless of whether respondents in 2005 can accurately
represent the behavioral intentions of visitors in the 2050s
or 2080s, the salient findings of the survey component of
this study stand—the magnitude of environmental change
required to become meaningful for altering people’s
intentions to visit Rocky Mountain national parks is very
substantial, and based on current understanding, is several
decades away even under the warmest climate change
scenarios. That is not to suggest climate change and
climate-induced environmental change should not be a
concern to park managers, but rather the near-term focus
of climate change adaptation in mountain parks should be
on managing the impacts of climate change for conservation mandates (Hannah et al., 2002; Scott & Lemieux,
2005) and increased visitation over the next 20–30 years,
resulting from an extended and climatically improved
warm-weather tourism season.
Mountain tourism is clearly influenced by the climate
(Scott, 2005), and this study has provided important
insight into the potential direct and indirect impacts of
climate change on visitation in one Rocky Mountain
national park in western North America. Given the
economic importance of national parks to nature-based
tourism in North America, it is hoped that this study will
encourage further research into the impact of climate
change on park tourism in other regions of Canada, the
United States and world-wide.
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