Recreational impacts on the fauna of Australian coastal marine ecosystems
Nigel Hardiman, Shelley Burgin*
School of Natural Sciences, College of Health and Science, University of Western Sydney, Locked Bag 1797, South Penrith DC 1797, Australia
a r t i c l e i n f o
Received 4 December 2009
Received in revised form
21 May 2010
Accepted 11 June 2010
Available online 7 July 2010
Nature-based tourism impacts
Recreational infrastructure impacts
a b s t r a c t
This paper reviews recent research into the ecological impacts of recreation and tourism on coastal
marine fauna in Australia. Despite the high and growing importance of water-based recreation to the
Australian economy, and the known fragility of many Australian ecosystems, there has been relatively
limited research into the effects of marine tourism and recreation, infrastructure and activities, on
aquatic resources. In this paper we have reviewed the ecological impacts on fauna that are caused by
outdoor recreation (including tourism) in Australian coastal marine ecosystems. We predict that the
single most potentially severe impact of recreation may be the introduction and/or dispersal of non-
indigenous species of marine organisms by recreational vessels. Such introductions, together with other
impacts due to human activities have the potential to increasingly degrade recreation destinations. In
response, governments have introduced a wide range of legislative tools (e.g., impact assessment, pro-
tected area reservation) to manage the recreational industry. It would appear, however, that these
instruments are not always appropriately applied.
? 2010 Elsevier Ltd. All rights reserved.
Outdoor recreation, including nature-based tourism, is growing
in popularity worldwide (Buckley, 2002; Cole, 1996; IUCN, 1996;
Pickering and Hill, 2007). In parallel with the growth in visitor
numbers, natural resource managers are increasingly required to
demonstrate ecological sustainability. This requirement is typically
driven by legislative obligations, often originating in the United
States of America (USA; e.g., Cole and Wright, 2004; Cole et al.,
2005; Haas, 2004). Despite such growth, knowledge of visitors’
ecological impacts and cost-effective methods for capturing and
using relevant information for their management are often inade-
quate or simply lacking (Buckley, 2003; Cole and Wright, 2004;
Hadwen et al., 2007; Pickering and Hill, 2007). This contrasts
with the importance of tourism, especially nature-based tourism,
tothe Australian economy in recent decades. Directlyand indirectly
the recreational industry produces an estimated AUD$58.3 billion
in Gross Value Added annually to the national economy and
employed approximately 810,000 people in 2005e2006 (TRA,
2007). Despite such economic importance, and Australia’s recog-
nised ecological fragility, knowledge of recreational impacts tends
to lag behind other countries, notably the USA and the United
Kingdom (Hadwen et al., 2006; Pickering and Hill, 2007; Sun and
Australia’s population is more concentrated on the coast than
anycountry of comparable size. An estimated 64% of the population
reside in the country’s eight capital cities, seven of which have
a coastal location, and account for 66% of annual population growth
(ABS, 2008). Coastal tourism is a major industry of the country with
significant on-going growth. Since the coast is the industry’s
primary asset, impacts on this resource will ultimately affect the
growth and sustainability of tourism, and coastal developments
more generally (DEWHA, 2007).
Previous Australian reviews of the ecological impacts of recre-
ation (e.g., Pickering and Hill, 2007; Sun and Walsh, 1998) have
tended to focus on terrestrial protected areas, and especially on the
effects on plant biodiversity and vegetation. A review of ecological
impacts on coastal marine ecosystems due to recreation has not
been previously undertaken, and data on such impacts are scat-
tered, and span multiple disciplines. In this paper we review the
ecological impacts on fauna that are caused by outdoor recreation
(including tourism) in Australian coastal marine ecosystems. This
will complement an earlier review of the impacts of recreation on
freshwater ecosystemsin Australia
submitted for publication).
* Corresponding author. Tel.: þ61 2 45 701029; fax: þ61 2 45 701403.
E-mail address: email@example.com (S. Burgin).
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Journal of Environmental Management 91 (2010) 2096e2108
2. Recreational impacts on coastal marine ecosystems in
Coastal marine recreation may be viewed as comprising leisure
activities that are focused at the interface between the land and the
marine environment, either in or affected by waters which are
saline or tide-affected (Orams, 1999).
Generally, recreation impacts on aquatic ecosystems are
considered to be less well understood than their terrestrial coun-
terparts, despite their disproportionately high importance as focal
points for recreation (Hadwenet al., 2006). This is probably because
changes in such environments are not as immediately obvious as in
terrestrial systems (Kuss et al., 1990; Liddle, 1997) where clear
delineation of recreational infrastructure impacts are often specific
and obvious, for example vegetation clearing for camp sites and/or
road construction for access (Pickering and Hill, 2007). Aquatic
threshold stress levels are difficult to define and measure (Davis
and Tisdell, 1996) and, if used for multiple activities, attributing
impacts to a specific activity may be problematic (Liddle, 1997;
Liddle and Scorgie, 1980). Similarly, attributing effects of recrea-
tion-specific infrastructure and that built for other purposes are
also not necessarily readily apparent in aquatic ecosystems. For
example, sewage from coastal residential or tourism properties, or
air pollution from commercial or recreational vessels using
a harbour may be indistinguishable (Warnken and Byrnes, 2004).
Aquatic recreational impacts may be either abiotic or biotic and
emanate from infrastructure (e.g., harbours, jetties, marinas) or
activities (e.g., boating, whale watching, reef walking, swimming,
fishing, SCUBA diving). They may cause decreased water quality,
introduction of non-indigenous
(including greater predation), and may be either direct or indirect.
In aquatic ecosystems linking these impacts is inherently prob-
lematical. Unlike their terrestrial counterparts, aquatic environ-
ments do not necessarily have clear boundaries, estuaries and
shorelines are linear with extensive ‘edge effects’, and their health
often depends on adjoining catchment and land management
practices that frequently occur outside of the adjacent zone. For
example, the same boat may be used on lakes, estuaries and at sea,
and some recreational impacts (e.g., chemical pollution, aquatic
weeds) can therefore be transported substantial distances from the
point source. Changes in biotic assemblages may also not be as
immediately obvious as for terrestrial areas that have been
denuded of vegetation or suffered track erosion. The effects may
also be indirect, for example, on biota that are dependent on the
aquatic ecosystem (e.g., water birds for nesting, feeding, and/or
breeding). Such impacts may also be localised (e.g., via predation)
or dispersed, for example via pollution, especially due to
construction of infrastructure.
biota, physical disturbance
2.1. Impacts associated with coastal marine tourism and recreation
Heavy metals such as copper, lead and zinc associated with
urban runoff (including recreation infrastructure) have been found
to change aquatic animal assemblages and reduce diversity of
macrobenthic fauna in intertidal soft sediments (Stark, 1998).
When coastal or estuarine wetlands that contain acid sulphate soils
are disturbed, for example by tourism (or agricultural or industrial
development), release of sulphuric acid and the associated toxic
metal ions can potentially degrade adjacent aquatic ecosystems,
cause marine cyanobacteria algal blooms (e.g., Lyngbya majuscala)
and may result in episodic assemblage effects and/or death of fish,
corals and aquatic plants, in addition to impacting on associated
terrestrial flora and water-dependent animals and birds. With an
estimated 666,000 ha of acid sulphate soils within the Great Barrier
Reef catchment, this represents major ecological and economic
problems in managing marine recreation infrastructure develop-
ment (Powell and Martens, 2005). Such problems are not restricted
to the environments of the Great Barrier Reef coast. Treatment
technologies to avoid the formation of acid drainage is associated
with tourism resort development on the Gold Coast (Queensland)
were estimated to cost in excess of AUD$100 million annually in
2000 (NSWA, 2000).
The most extensive primary water pollutant from recreation
infrastructure, however, is probably sewage, together with some
detergent and/or chemical toilet discharge, for example from
houseboats and shore-based facilities. Although there are many
records of high levels of faecal coliform bacteria at bathing beaches
these are, however, often the consequences of nearby urban
discharge and/or the level of the associated sewage treatments. The
relative quantity of sewage discharged that is directly associated
with recreation activities mayonly constitute a small fraction of the
total discharge, and resulting concentrations may be typically low.
Under such conditions, sites close to a low-volume, tertiary-treated
domestic sewage effluent outfalls have not displayed consistent
differences in hard-bottom macrofauna assemblages inhabiting
kelp holdfasts (Smith, 1994). This situation may change substan-
tially, however, during periods of restricted flow and/or if they are
located close to specific recreation infrastructure developments
such as boat marinas or localised tourist development/s (Liddle,
1997). Seasonal trends in tourism can also mediate the timing
and effects of eutrophication from sewage effluent. With increased
loads caused by periodic and often predictable booms in visitor
numbers at specific times of the year, impacts may spread over
a wide area and have effects over substantial distances away from
the main areas of tourist aggregation. Terrestrial sewage treatment
facilities, such as septic tanks at popular coastal tourism spots, are
often unable tocopewith massive surgesin visitornumbers at peak
periods (e.g., long weekend holidays, Christmas period). Such peaks
cause the infrastructure to overload/malfunction and often lead to
marine pollution and ecological degradation (Hadwen et al., 2006).
Secondary-treated sewage effluent (including from recreation) has
also been found to reduce abundance of some species. For example,
under such circumstances the rate of mortality increased in the
mussel Brachidontes rostratus (Hindell and Quinn, 2000).
Many types of coastal ecosystems such as mangrove swamps,
tidal salt marshes and seagrass beds are important nursery
habitats for fish and other organisms. Clearing these habitats,
together with dredging channels, dredging/blasting for marinas
and other coastal recreation infrastructure (e.g., resorts) directly
impacts on coastal marine ecosystems, and associated offshore
productivity. Such disturbance may affect fish abundance and
overall biological diversity due to increased sedimentation, and
changed hydrology. The removal of such protective vegetation
may also render the shoreline prone to storm surge erosion (Hall,
2001; Walker, 1991). Indirect impacts caused by infrastructure
have also been recorded. For example, on Heron Island (Great
Barrier Reef Marine Park, Queensland), resort development
reduced species diversity and altered the community composition
of breeding sea-birds due to post-construction terrestrial weed
infestation combined with a large influx of silver gulls Chroico-
cephalus novaehollandiae attracted by food scraps originating from
the resort (Walker, 1991).
In contrast to the many detrimental effects of recreational
infrastructure such as marinas, some have been found to increase
fish abundance and diversity, and change the composition of
species assemblages associated with them (Clynick, 2008). For
example, pilings used in the construction of such infrastructure
affect food availability, shelter and habitat for fishes potentially at
all depths of the water column (e.g., Rilov and Benayahu, 1998), in
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
contrast to artificial reefs, which tend only to affect ocean floor
species (e.g., Clynick, 2008; Russell, 1975).
Sessile organisms, such as mussels, often grow on submerged
artificial structures and provide a secondary substratum, and these
niches may be taken up by a wide range of taxa. Assemblages of
mobile and sessile organisms inhabiting mussel beds have
however been found to differ among natural and various types of
artificial structures. Assemblages associated with mussels that
have settled on pontoons have been consistently observed to be
different from those on other structural substrates such as pilings,
seawalls, and natural reefs (People, 2006). Manmade recreational
structures (e.g., seawalls, pontoons, pilings) have also been found
to affect subtidal epibiota such as macroalgae, invertebrates and
fish, either by changing assemblage structure (Glasby, 1999) or by
increasing their abundance and diversity (Connell and Glasby,
1999). Although differences in type of artificial substrate may
cause changes in assemblages from adjacent natural habitat
(Connell and Glasby, 1999; Glasby, 1999), the type of structure
appears to be more important than its surface characteristics
Such infrastructure impacts are not restricted to the marine
foreshore but may also extend further offshore. Mid-sea mooring
pontoons located over reefs to facilitate SCUBA diving are an
increasingly popular form of marine tourism infrastructure. On the
Great Barrier Reef these structures have evolved from small, simple
platforms, initially for coral viewing, swimming and SCUBA diving
to large, complex structures that may comprise multiple storeys,
underwater observatories, dive platforms, overnight accommoda-
tion, sewage treatment, waterslides, and other features that are
more commonly associated with land-based resorts. Impacts of
such structures include physical damage to coral during installa-
tion, on-going damage from periodic movement of mooring blocks
and chains, and further subsequent loss of coral due to sunlight
shading, biological impacts from sewage discharge, oil and fuel
pollution, either from use and/or from pontoons sunk in cyclones
(Smith et al., 2005).
Artificial reefs are another form of recreational structure
deployed in many countries across the world, including Australia
(e.g., Talbot et al., 1978), Europe (e.g., Jensen, 2002; Wilding and
Sayer, 2002), USA (e.g., Ambrose and Anderson, 1990; Bohnsack
et al., 1994; Grossman et al., 1997; Ogden and Ebersole, 1981),
and Japan (e.g., Polovina, 1989). By 1985 artificial reefs were
installed across all Australian states (Pollard and Matthews, 1985).
While more recently such reefs have been developed to enhance
incident waves for recreational surfers (e.g., The Cables, Western
Australia, Pattiaratchi, 2007), historically the major recreational use
of artificial reefs has been to promote increased fish abundance and
diversity for fishing and diving, including SCUBA and/or spear-
fishing (e.g., McGlennon and Branden, 1994; Pears and Williams,
2005; Seaman, 2002).
Despite their widespread use, numerousresearchers (e.g., Guiral
et al., 1995; Grossman et al.,1997; Pickering and Whitmarsh,1997;
Wilding and Sayer, 2002) have questioned whether artificial reefs
impact detrimentally on local animal communities. The State with
the largest number of officially deployed artificial reefs (South
Australia), no longer considers them a mechanism for fisheries
enhancement, in part because the authorities considered that there
was insufficient evidence to suggest that the target species pop-
ulations were not detrimentally impacted due to over-exploitation
by fishers (Pears and Williams, 2005). Internationally, researchers
(e.g., Grossman et al., 1997; Guiral et al., 1995; Polovina, 1989;
Wilding and Sayer, 2002), together with researchers with a focus
on Australia reefs (e.g., Pears and Williams, 2005; Pickering and
Whitmarsh, 1997) have documented concern for the potentially
deleterious affects on reef fish populations due to the concentration
of previously unexploited segments of the target population, that
subsequently provide opportunities for increased fishing effort and
catch rates, and thus encourage over-exploitation of species tar-
geted by recreational (and commercial) fishers. Pears and Williams
(2005) suggested that in areas such as the Great Barrier Reef with
‘relatively healthy’ ecosystems, any recreational benefits derived
from fishing on artificial reefs are likely to be short-term due to
increased fishing pressure. This is because the reefs are at known
destinations and typically there is increased participation in fishing
in such areas, and there is an associated increase in catchability and
catch which ultimately leave the artificial reef and the source areas
over-exploited. Conversely, the productivity of the artificial reef
may increase diversity on a local scale by increasing biotic and
abiotic habitat complexity, although the scale of such influence may
be limited (Davis et al.,1982). For example, Ambrose and Anderson
(1990) found that around artificial reefs, although some inverte-
brate species increased in density, others decreased and they sup-
ported the concept that the overalleffect was limited toa small area
close to the reef.
Wilding and Sayer (2002) suggested that large inshore artificial
reefs may also have the potential to significantly change hydro-
logical and biological attributes in the surrounding waters
including, for example, tidal current dynamics over the complete
spring-neap tidal cycle including granulometric changes in the
sediments, seasonal changes in oxygen levels throughout the
sediment column, and changes to species composition. Ambrose
and Anderson (1990) observed such differences only in the
immediate vicinity of an artificial reef, however there is a greater
potential for such issues to arise when the artificial reef is inshore,
for example those developed for surfing (cf. Pattiaratchi, 2007).
Shark nets (cf. protective gillnets) are another form of recreation
infrastructure that is widely used on the east coast of Australia (e.g.,
Burgin, 1986; Gribble et al., 1998; Krogh and Reid, 1996; Paterson,
1979), and off South Africa (e.g., Allen and Cliff, 2000; Cliff and
Dudley, 1992; Dudley, 1997; Dudley and Cliff, 1993) to protect
swimmers on popular bathing beaches (Allen and Cliff, 2000;
Heinsohn et al., 1977). Most attacks on swimmers along Austral-
ia’s east coast are by great white Carcharodon carcharias, bull
Carcharhinus leucas, and tiger Galeocerdo cuvier sharks and there-
fore these are the target species for such protective netting (Dudley,
Analysis of the data on non-target species captured in shark
nets in Australian waters (e.g., Krogh, 1994; Paterson, 1979; Reid
and Krogh, 1992) has revealed that many non-target animals,
including protected species, also perish annually in the nets. For
example, the vulnerability of the dugong Dugong dugon to shark
netting has long been recognised. In 1972, Heinsohn reported that
the population off Magnetic Island (Queensland) was completely
destroyed in 1 year post-deployment of nets, and populations in
the vicinity of Townsville were ‘severely decimated’ when shark
netting was introduced in the area: 82 dugongs were killed in the
first year of netting, and by 1965 most of the animals caught in
the area were young females thought to be immigrants. The
numbers that drown annually in shark nets has declined over
time (Marsh, 1986). For example, Gribble et al. (1998) reported
that the long-term, pre-1992, dugong mortality rate from the
Queensland Shark Control Program on the Southern Great Barrier
Reef averaged 20 individuals annually. Modifications to netting
occurred between 1992 and 1995. This apparently has reduced the
overall annual loss (Gribble et al., 1998), and more recently
(Marsh, 2006) there has been an evaluation of management
initiatives to further reduce mortality in the Great Barrier Reef
World Heritage Area.
In addition to dugongs, other protected species of whales,
dolphins and turtles are also caught incidentally in shark nets off
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
Queensland (e.g., Gribble et al., 1998; Paterson, 1979) and New
South Wales (Krogh,1994; Reid and Krogh,1992). Forexample, over
16 years (1962e1978), approximately 165 turtles were caught
annually, including the ‘critically endangered’ leatherback turtle
including the humpback Megaptera novaeangliae (Gribble et al.,
1998; Paterson, 1979), pilot Globicephala spp., minke Balaenoptera
spp., and false killer Pseudorca crassidens have also been found
entangled in shark nets (Gribble et al., 1998).
In Queensland, dolphin captures in these nets occur mainly in
the southern waters (Paterson, 1979). Two species, the Irrawaddy
River Orcaella brevirostris, and Indo-Pacific humpbacked Sousa
chinensis, both threatened species, have been captured in the nets
protecting swimmers from sharks (Gribble et al., 1998). Of all
species captured, Paterson (1979) suggest that dolphins were
potentially the most probable of the marine mammals to avoid
capture, however, Shaughnessy et al. (2003) suggested that marine
mammals that are active predators of fish, cephalopods and crus-
taceans may not detect the presence of fisheries equipment and/or
are attractedtopotential prey. This provides an explanation for why
even animals as intelligent as dolphins are routinely captured in
Paterson (1979) also considered the loss of unprotected species
was ‘severe’, and included among these species were a number of
species of rays, including the cowtail ray Dasyatis sephen, long-
tailed ray Himantura uarnak, brown stingray Dasyatis fluviorum,
coachwhip ray Himantura granulata, spotted eagle-ray Aetobatus
narinari, and the pigmy devil ray Mobula diabolus.
Despite the abovementioned potentially detrimental effects on
larger marine animals, shark netting appears to have beneficial
effects on some species. For example, two species of seahorses
(Hippocampus spp.) are abundant on the shark nets around
permanent swimming enclosures in Sydney Harbour. Since colo-
nisation of new netting takes at least four months, removal and/or
disturbance of this artificial habitat for maintenance and/or
replacement may impact these fishes’ survival (Clynick, 2007).
The spatial placement of nets may also influence the impact on
particular species. An analysis of New South Wales shark catch data
between 1950 and 1990 (Krogh, 1994; Reid and Krogh, 1992)
revealed differences in catch rates due to the location of the shark
net. Where the netted region was central on the beach shoreline
there tended to be higher catch rates than when the nets where
deployed at each end of the beach. By-catch species composition
also varied due to beach topography, for example hammerhead
sharks Sphyrna lewini were caught on long open beaches while
whaler Carcharinus spp., white pointer C. carcharias, and tiger
Galeocerdo cuvier sharks were captured when deep water was close
inshore. Gribble et al. (1998) suggested that despite large annual
fluctuations in the capture rates, there were no discernible trends
in the by-catch from shark nets, although post-1973 catches of all
taxonomic groups decreased substantially, reportedly due to
changes in net specifications and their spatial deployment. Overall
they considered the impact of incidental mortality on populations
was ‘probably minor’. Since 1974 the number of sharks caught in
shark nets and catch per unit effort have declined substantially
(Krogh, 1994). The greatest reduction in species catch has been the
whaler and white pointer sharks. There has also been an on-going
catch of non-target sharks and non-shark species including rays
(approximately one ray to every 2e3 sharks), and dolphins
(approximately one to every 30e40 sharks). Since shark pop-
ulations are in decline worldwide and they are long-lived, slow
growing and have low fertility, they are vulnerable to overfishing
(Liddle, 1997; Walker, 1998). Such decline in catch rate is therefore
likely, at least in part, to be due to the declining populations of such
2.2. Impacts associated with coastal marine tourism and recreation
2.2.1. Recreational fishing
Recreational fishing is effectively a universal activity that is
growing in popularity (McPhee et al., 2002). In Australia it may
equal or exceed the commercial catch in many areas (Lynch, 2006;
Lynch et al., 2004; West and Gordon, 1994). For example, penaeid
prawning is an economically important seafood resource, and these
prawns are also the basis of a substantial recreational fishery. It has
been estimated that recreational prawning occurs in 128 of the 136
estuaries in New South Wales, and accounts for 31% (weight;
value ¼ 24%) of the annual commercial prawn catch (Reid and
Montgomery, 2005). The recreational ‘abalone season’, off Perth
(Western Australia) occurs over six Sundays in November and
December. During permissible times within this period collectors
may wade, snorkel or SCUBA along Perth’s metropolitan rock
platforms and each harvest has a maximum of 20 Roe’s abalone
Haliotis roei over 60 mm. In the 2005 season it was estimated that
3000 people collected abalone in a single day, with an estimated
240,000 abalone harvested (Morton, 2005).
Such focused attention on a single target species in a specific
area is typical of recreational fishing. As a result impacts on the
local population may be highly concentrated (West and Gordon,
1994). Species composition of the recreational catch also often
differs substantially from the commercial catch (Reid and
Montgomery, 2005; West and Gordon, 1994). This may result in
reduced abundance and/or breeding success of the target species.
For example in one survey of recreational anglers 16e56% of the
catch was below legal minimum length (West and Gordon, 1994).
The impact of recreational fishing may also extend beyond the
target species. The widespread practice of ‘bait-pumping’ on tidal
mud flats (i.e., digging up invertebrates for subsequent use as fish
bait) has been shown to reduce populations of the ghost shrimp
Trypaea australiensis (Contessa and Bird, 2004). Such activity,
undertakenfrequentlyandoverextensive areasof themud flats can
degrade ecosystem health and biodiversity in the area. Specific
effort in a relatively restricted habitat can also be highly destructive
(Lynch, 2006). The more severe impacts of fishing, such as reef
destruction, have been observed even in remote areas such as the
North Kimberley in the north of Western Australia (Hercock,1999).
Where specific species are targeted, angling may also indirectly
cause cascade effects within the assemblage. For example, as
observed in the Ningaloo Marine Park (Western Australia) differ-
ences in composition of species within the local marine assemblage
may change, even at the family and generic level of taxonomy
(Westera et al., 2003).
Direct impacts may also occur, even to endangered non-target
species such as the grey nurse shark Carcharias taurus. Although
historically subjected to commercial fishing, grey nurse sharks are
often mistaken for a dangerous species. This has resulted in
recreational spearfishers targeting them with explosive harpoons
and individuals continue to be illegally fished, accidentally
captured by recreational (and commercial) fishers, and killed in
shark nets (EPA, 2009; DEWHA, 2009a).
Although there are a few exceptions, for example the trans-
location of native black bream Acanthopagrus butcheri to adjacent
estuaries and saline lakes in southern Western Australia (Smith
et al., 2008), the basis of recreational fishing in Australian marine
environments is based on the take of native species of the area. This
is in contrast to recreational fishing in freshwater habitats where
the target is often based on exotic species such as the salmonids
(Lintermans, 2004), sometimes with devastating effects (e.g.,
European carp Cyprinus carpio; Liddle,1997). Recreational fishing in
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
marine environments does however have other impacts, for
example on species and/or habitat, and this has been acknowl-
edged in the design of aquatic protected areas such as marine parks.
In such reserves recreational fishing is largely concentrated near-
shore, using small craft, while commercial fishing zones are
typically situated further offshore. Based on a 14 year study,
comparisons between no-take and recreational fishing zones
located over near-shore reefs of the Great Barrier Reef Marine Park
(GBMP) revealed that density, abundance and biomass of species
targeted in non-take zones was significantly higher than in fished
zones (Evans and Russ, 2004; Williamson et al., 2004). Compari-
sons of no-take and recreationally fished zones in the Ningaloo
Marine Park showed changes in assemblages of families and
genera. There was also lower biomass, and fewer legal-sized fish of
some targeted species in the fished zones (Westera et al., 2003). In
an equivalent investigation in sub-tropical eastern Australian
waters, the same trend was revealed for nekton (fish and inverte-
brates) in mean size, evenness, and community composition, but
not necessarily species richness or diversity (Pillans et al., 2007).
Such results indicate the effects that recreational fishing may
have on ecosystems although other studies that have compared
fished and un-fished marine habitats have returned equivocal
results. For example, a study of the effects of protection zoning on
the abundance of two reef fish species (sub-tropical wrasse
Choerodon rubescens and coral trout Plectropomus leopardus)
commonly targeted by recreational fishers in water adjacent to the
sub-tropical Houtman Abrolhos
revealed that some species responded positively to protection
while others did not (Nardi et al., 2004). Spatial and temporal scales
of the effectiveness of zoning within marine park reserves may
therefore vary among species and/or the overall area.
An additional issue is that most models used in designing
marine parks are based on the assumption that fishing effort is
a dynamic pool, homogenously distributed over areas. It is also
assumed that area and fishing effort are directly correlated. Subsets
of reserves may therefore be selected on this basis. Conversely,
research in the Jervis Bay Marine Park (New South Wales) has
revealed the need to consider the incorporation of subducted
fishing effort into decisions on zoning in marine protected areas, as
opposed to decisions being based on simple percentage targets/
mean allocation of effort by area (Lynch, 2006). This is because
recreational fishers’ effort may be heavily skewed towards specific
locations (Lynch et al., 2004), for example, artificial reefs (e.g., Pears
and Williams, 2005; Pickering and Whitmarsh, 1997; Lynch et al.,
2004). Design based on mean fishing effort over a broad area
could therefore be inappropriate and underestimate recreational
fishing impacts. Despite these observations, few published marine
park area plans appear to have considered spatial allocation of
fishing effort, probably because the design of such reserves is often
heavily influenced by political, rather than scientific argument.
Any plan for recreational fishing in marine protected areas is
further complicated by the need to accommodate multiple recre-
ation uses. During preparation of the draft zoning plan for the
gazettal of the Jervis Bay Marine Park in 2000, Lynch et al. (2004)
investigated the conflict between SCUBA divers and anglers,
whose activities temporally and geographically overlapped. As part
of the study they also investigated the trends in recreation activi-
ties. These data showed that over the previous decade the number
of divers had remained constant while angler numbers had
substantially increased, and had typically doubled over the decade,
and tripled for some months of the year.
2.2.2. Physical disturbance to the environment
Among the three basic types of coastal terrestrial ecosystems e
rocky shores, tidal wetlands and sandy beaches, the latter are the
most-visited and highly valued (Schlacher et al., 2007). Apart from
providing key ecological services such as filtering large volumes of
seawater and recycling nutrients, they provide critical habitat for
a diverse array of invertebrates, birds and turtles. Caught in
a squeeze between increasing human population and associated
infrastructure development, recreation, and rising sea levels, sandy
beaches are under increasing ecological pressure. One high-impact
recreational activity popular on sandy beaches is four-wheel
driving (4WD). For example, in Australia in 2010, one in five of the
new vehicles purchased are 4WD (FCAI, 2006-2010). Although
national figures are unavailable, participation in 4WD in south east
Queensland, a region famed for its beach lifestyle, the numbers
have increased from 20% to 23% between 1997 and 2007, and
activity events have more than doubled from 672,700 to 1,516,634
(Hales, 2008). Obvious impacts include visual (aesthetic) degrada-
tion, noise, and air pollution. Overseas research has shown that
4WD can cause ecological damage indirectly via habitat destruction
and behavioural modification, for example of birds, reduced
survival rate in newly hatched turtles, and direct impacts due to
crushing of flora and fauna. Despite its extensive coastline of sandy
beaches and the popularity of 4WD there is, however, limited
empirical research of the ecological effects of this activity in
Australia. In the few studies undertaken, vehicle tracking has been
shown to be extensive, typically covering 54e61% of the beach area
between dunes and swash, with corrugating reaching a deep of
28 cm, with associated disruption in a single day to as much as 9.4%
of the available invertebrate faunal habitat matrix (typically the top
30 cm) (Schlacher and Thompson, 2008). Crab abundance and
densities have also been found to be significantly lower on beaches
where 4WD is permitted than on beaches where it does not occur
(Moss and McPhee, 2006).
Although in more temperate regions of Australia 4WD associ-
ated with beaches is likely to be seasonal, in more northern sub-
tropical and tropical clines the activity and its associated impacts
occur year-around (pers. obs.), and it is an increasingly popular
activity (Hales, 2008). Since human trampling in, for example,
mangrove forests may take several years to recoverafter the impact
has been removed (Ross, 2006) we assume that even seasonal use
of beaches for 4WD would have an on-going impact on the biota of
the impacted beach areas.
Physical impacts can also emanate from recreational vessels,
particularly power boats. These include wash, turbulence and
turbidity, along with direct and indirect propeller impact on fish,
macroinvertebrate communities (disturbance of bottom sedi-
ments), mammals (nesting holes) and birds (nesting sites). All of
these disturbances have the potential to impact on marine and/or
estuarine fauna (Liddle, 1997). In most coastal habitats, such
impacts are dispersed or mitigated by sandy beaches or rocky
foreshores, but in estuaries traffic may be highly concentrated in
shallow waters and/or narrow channels, for example around
marinas, and boat launching ramps. Although there is apparently
limited Australian marine recreation ecological research specific to
estuarine habitats (e.g., Ellis et al., 2005), effects are likely to be
similar to those known for lakes and reservoirs (Burgin and
Hardiman, submitted for publication; Mosisch and Arthington,
Coastal marine physical disturbance impacts may also occur as
a result of recreation even without the effects of infrastructure or
vehicles/vessels but apparently these do not always have a strongly
detrimental outcome. Forexample, Gladstoneet al. (2006) reported
that almost 50% of the estuaries in South-east Australia were
classified as ‘intermittently open’ typically due to artificial
breachingof the sand barriers sealing the mouth of the estuary. One
justification for this engineering is to enhance the recreational
amenity, for example for swimming and boating. Over time the
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
entrance barriers are reformed naturally by wave action. Griffiths
(1999), Griffiths and West (1999) reported that such intermittent
opening of estuaries had no substantial effect on species richness or
fish abundance, however assemblage structure was changed due to
the migration of the recruits of marine-spawning fishes into the
estuary. The intermittent breaching was also not found to impact
on species richness, the density of the amphipod Paracalliope aus-
tralis, or the gastropod mollusc Aschoris victoriae. In breached
estuaries there was also no significant overall change in the mac-
roinvertebrate assemblage structure above natural variation. Indi-
vidual species and the overall macroinvertebrate assemblage of the
estuaries where barriers are intermittently breached therefore
appear to be resilient to this artificial manipulation (Gladstone
et al., 2006).
A recreation activity that may be assumed to be potentially less
detrimental than breaching estuaries is walking in seashore inter-
tidal habitats, for example on rock platforms. Studies undertaken to
investigate the impact of trampling on intertidal algal and marine
invertebrate communities have revealed that most molluscs tendto
be resistant totrampling, and some mayeven increase in density. In
contrast, algae were found to be less resistant and may suffer
reduced abundance and cover, and recover slowly. Some species,
for example, brown alga Hormosea banksii take up to 4 years to
recover (Keough and Quinn, 1998; Povey and Keough, 1991). Ross
(2006) also investigated the effects of trampling in the intertidal
zone. Her research was undertaken in temperate mangrove forests
in south-eastern Australia and trampling was observed to cause
significant changes to microhabitat features and macrofaunal
assemblages. With only 25 trampling passes in the forest, therewas
at least a 3 year impact on the fauna (the length of the study). In
parallel with a decrease in the number and height of mangrove
trychiaeCaloglossa algal association was reduced by 50%, and the
gastropod macrofauna (e.g., Ophicardelus spp., Assiminea bucci-
noides) associated with the pneumatophores/algal assemblages
were most impacted. One year after trampling, recruitment of
Ophicardelus spp., A. buccinoides and Salinator solida were still
significantly reduced. In contrast to the detrimental effects on some
species, the gastropods and crabs that were more closely associated
with the sediment surface than the mangrove pneumatophores
were not significantly affected. Even modest trampling may
therefore cause substantial changes in floral and faunal assem-
blages in seashore or estuarine habitats (Ross, 2006).
Trampling may occur in habitats further offshore, for example
on coral reefs. Australian near-shore waters support some of the
world’s most extensive and diverse coral reef habitat (e.g., Ningaloo
Marine Park, Great Barrier Reef World Heritage Area) which draw
largenumbers of domestic and internationalvisitors annually. Early
research (Kay and Liddle, 1987, 1989) on trampling impacts of
visitors ‘reef walking’ showed that corals are fragile and their
resistance to physical damage generally low, especially for those
species with branching morphology. Survival rate of detached
pieces was dependant on species and fragment size. For example,
larger fragments had higher survival rates than smaller ones, and
recovery was slow, typically more than three months, and species-
Coral reefs also attract snorkelers and SCUBA divers in
substantial numbers. There were more than 17.5 million certified
SCUBA divers worldwide in 2008 and participation in the sport
showed a double-digit annual growth between 1980 and 2000,
although this growth has since plateaux (PADI, 2009). SCUBA
diving has traditionally been considered to have a low environ-
mental impact. Data on its long-term impacts are lacking however,
and there are none on Australian reef habitats (Davis and Tisdell,
1996; Dixon et al., 1993; Harriott et al., 1997; Hawkins and
Roberts, 1992, 1993). Critical threshold levels for sustainability
have also not been investigated (Dixon et al., 1993; Davis and
Tisdell, 1996). Although increased siltation may lead to suffoca-
tion of corals, the main impact from diving is considered to be
physical damage tothe substrate and/or the coral, mostly as a result
of fin kicks or inadvertent contact with the coral with hands or
equipment (Lynch et al., 2004; Rouphael and Inglis, 2001).
Although damage by an individual is likely to be minor, the
cumulative effects may cause substantial localised destruction of
coral over time. Impacts appear to be spatially and temporally
heterogeneous however, and also vary widely among divers. While
most divers do not damage the coral a minority do, generally due to
poor buoyancy control and finning (Harriott et al., 1997; Roberts
and Harriott, 1995; Rouphael, 1995). Significant behavioural
differences have also been found among purpose of dive and
gender groups. For example, specialist photographers are more
likely to cause damage than non-specialist photographers, and
male divers typically do more damage than female divers
(Rouphael and Inglis, 2001). Although some studies (e.g., Roberts
and Harriott, 1995) have found an association between increased
diver experience and decreased damage, others (Harriott et al.,
1997; Rouphael and Inglis, 2001) have found no such correlation.
The morphological composition of benthic assemblages damaged is
consistent with reef walking impacts, with substantial breakage
(lower resistance) observed among branching corals compared to
massive/flat growth species (Kay and Liddle, 1987, 1989; Rouphael
and Inglis, 1997).
Physical damage to coral assemblages may also occur as a result
of boats anchoring directly above them, for example to fish or dive.
In such cases, damage may not be limited to the point of impact by
the anchor but may also be caused by the anchor chain dragging
along the sea bed and/or as the boat shifts with wind, tide and
current (Davenport and Davenport, 2006).
Another recreational impact in the marine environment is the
accidental lethal collision between boats, particularly power boats
and cetaceans, seals, turtles and dugong D. dugon. Slow-swimming
species such as turtles and dugongs are particularly vulnerable.
Similarly, because green turtles Chelonia mydas often browse in
near-shore seagrass meadows, where recreational boat traffic is
concentrated, individuals of this species are frequently injured. The
dugong also browses on the inshore seagrass beds. It has small
populations, and is highly endangered even within marine pro-
tected areas (Davenport and Davenport, 2006). Dugongs are also
vulnerable even to sailing boats, and may be at risk of displacement
from its natural habitat because of sensitivity to, and avoidance of
boat-generated noise (Preen, 2001). The introduction of new
technologies such as propellerless jet skis that can travel at high
speeds (noisily) in very shallow water or even over low banks
through salt marsh vegetation, even in protected marine areas, has
further extended such impacts into habitats that were previously
refuges for animals from boat traffic (Burger, 1998; Preen, 2001),
and has potentially increased stress levels of individual dugongs
that are considered to be particularly susceptible to stress (Marsh
and Anderson, 1983).
Although vessel strike is known to be a threat to turtles there is
no published quantification of such impacts on turtles in Australian
waters (Hazel and Gyuris, 2006) although it is likely to have been
substantial. For example, between 1999 and 2002 at least 65
turtles, mostly green and loggerhead Caretta caretta were killed
annually byvesselstrike on the Queensland east coast. This figure is
broadly comparable to the deaths due to commercial trawl nets
before the introduction of mandatory turtle-exclusion net devices.
Since these figures were based on the numbers of stranded mori-
bund animals, they are likely to be an underestimate of mortality
(Hazel and Gyuris, 2006).
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
2.3. Indirect impacts associated with coastal marine tourism and
2.3.1. Water quality/pollution
The impact on native species from boat traffic is not restricted to
collision. Recreational marine pollution occurs in the form of fuel,
oil and other chemicals discharged from power boats. Such pollu-
tion is worst from 2-stroke engines (which use a mix of petrol and
oil) traditionally used on small personal watercraft such as jet skis
and speedboats with outboard engines, although the more wide-
spread introduction of small 4-stroke engines which are cleaner
and more efficient, together with the increased use of modern
unleaded fuels have reduced this pollution (Davenport and
Such pollution results in reduced water quality and/or is accu-
mulated in the sediments, and exhaust fumes contribute to air
pollution. The main chemical contaminants are methyl tertiary
butyl ether (MTBE) and polycyclic aromatic hydrocarbons (PAH).
While the effects of MTBE appear to be unstudied, some
compounds of PAH are potentially carcinogenic (Davenport and
shallow confined areas such as estuaries and harbours (typically
favoured recreational areas), where the pollution in the water
column and sediments mayincrease, evenwhen the boat is moored
(Davenport and Davenport, 2006; Mosisch and Arthington, 2001,
Boats used in seawater are subject to hull fouling by encrusting
organisms such as barnacles and limpets. Such fouling causes drag,
reduces speed and increases fuel consumption and, potentially,
associated pollution. To remove this biota from the hull anti-fouling
paints are applied regularly, most commonly with copper-based
paints or those containing the chemical Tributyltin (TBT). While
TBT has proved effective, it is highly toxic to marine organisms,
even in minute quantities (Mosisch and Arthington, 1998, 2001,
2004). In 2001 the International Maritime Organisation adopted
a treaty to globally ban the use of TBT in anti-fouling paints to
protect the marine environment (Champ, 2003; Kotrikla, 2009).
Impacts of this chemical will therefore gradually decline over time.
Between 1999 and 2009 the total number of registered recrea-
tional boats in Australia increased from between 589,346 and
803,788, an increase of 36.4%. Of the total number registered in
2009, 56.2% was registered in the three eastern mainland States
(Data collated from information supplied by Recreational Boating
Units of all State/Territory Marine Safety Departments). This
reflects the country’s coastal-concentrated population and the
popularity of recreational fishing (boat ownership and participa-
tion in fishing are strongly related). Each vessel is a potential source
of human- or machine-generated waste. For example, it was esti-
mated that over 1 year across 20 popular anchor sites in the eastern
part of the Moreton Bay Marine Protected Area (Southern
Queensland), approximately 10,000 recreational vessels (>6 m
length) produced an estimated annual load of 141 þ46 kg of copper
(from copper-based anti-fouling paints), 22.2 ? 13.3 ton of human
faecal matter, 102.1 þ 13.3 ton of urine, and 1.17 þ 0.38 ton of
nitrogen (from sewage). This load is broadly equivalent to the load
from a modern sewage treatment plant serving a 4.5 km2, mostly
residential development and associated infrastructure (Leon and
Warnken, 2008). Such vessel activity (and hence pollution load)
is also likely to be highly variable, with spikes of higher activity on
weekends and school/public holidays, most notably in the
Christmas/New Year and Easter periods (Widmer and Underwood,
Small, privately-owned boats are not the only vessels employed
in recreation and tourism. Large cruise ships, now estimated to be
approximately 250 ships globally, carry an estimated 12 million
is potentiallygreatest in
passengers annually (Davenport and Davenport, 2006), and
increasingly transport tourists to hitherto largely unvisited desti-
nations such as the Antarctic Treaty Area. Since the first recorded
tourist cruise in 1958 (Reich, 1980), seaborne tourists making
landings on the Area increased from 6704, carried on 12 vessels
over 59 voyages by 10 commercial operators in 1992e1993 to
32,198 tourists, 55 vessels, 308 voyages by 48 operators in
2007e2008 (IAATO, 2009). Increase visitation has been particularly
strong since 1985e1986, with the introduction of larger ships and
more frequent cruises (Enzenbacher, 1992). Although such cruises
may arguably impose a light ecological footprint as they require
little or no land-based infrastructure, each passenger on such ships
nevertheless produces an estimated 3.5 kg of garbage and solid
waste daily and each ship discharges roughly 1 million litres of
sewage on a weekly voyage (EPA, 2000). As solid waste cannot
legally be dumped at sea, it contributes to landfill at shore-based
recreation destinations or home ports, further degrading habitat
and increasing pollution (Davenport and Davenport, 2006). Food is
often the largest single component of a ship’s garbage stream, with
per capita estimates ranging between 1.4 and 2.4 kg/day (IGMCO,
1978), increasing as a function of crew size (Polglaze, 2003), and
up to 46% of total garbage (Horsman, 1982). In most cases, such
waste is still simply discharged at sea, although increasingly
restricted under the International Convention for the Prevention of
Pollution from Ships 73/78. Although little-researched, such waste
can diminish water and sediment quality, impact marine biota and
change nutrient levels, potentially causing changes in species
composition and diversity (Polglaze, 2003). Major accidents are
also a risk. For example when the cruise ship MS Explorer hit ice and
sank in the Antarctic in 2007 it dumped 180,000 l of diesel fuel into
the sea (Travel Weekly, 2007) and the Norwegian Crown caused
significant localised contamination of reef sediments bycopper and
zinc due to the ship’s anti-fouling hull paint when it grounded in
Bermuda in 2006 (Jones, 2007).
2.3.2. Stress disturbance/changed behaviour
whale, dolphin and/or bird watching, and glass-bottomed boat
trips (Duffus and Dearden, 1990) has increased substantially in
popularity worldwide in recent years. Although watching wildlife
may be regarded as benign because typically animals are not
physically interfered with, animal behaviour may be altered as
a result of human behaviour around them which may, in turn,
result in threats to the long-term sustainability of the target pop-
ulation. For example, in a study undertaken on the diversity and
relative abundance of non-breeding (over-wintering) terns on
sandbanks in Moreton Bay (Ramsar site), researchers noted that the
terns’ visits coincided with the peak human recreational use of the
adjacent waterway. Recreational activities included boating, diving,
fishing, bait gathering, and kite surfing, together with the use of jet
skis and similar motorised personal watercraft. Each of these
activities had the potential to impact on bird behaviour (Chan and
Recreation may also cause impacts on the behaviour of aquatic
mammals and fish (e.g., whales, dolphins and porpoises, seals and
sea lions, dugong, whale sharks; Liddle, 1997). Although observa-
tion of such animals is primarily from boats, aerial observation from
plane or helicopter also occurs, and at Monkey Mia (Western
Australia) the Indo-Pacific bottlenose dolphin Tursiops aduncus
moves into the shallows to waiting tourists. The main perceived
negative impact of such interaction is that the animals may be
driven from nurseries and/or mating areas, or become habituated
to humans and change their behaviour, for example, for playful
interaction, curiosity and to seek food, at times with detrimental
outcomes (Liddle, 1997).
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
Whale watching is particularlyattractive to recreationists. There
has been an estimated increase in this activity from four million
people in 31 countries and territories in 1991 to more than nine
million watchers in 87 countries and territories in 1998. The
industrygenerated approximately $US1 billionworldwide annually
in 2001 (Hoyt, 2001). In Australia, there are at least 175 whale-
watching commercial tour permits (Birtles et al., 2001; Valentine
et al., 2004), and the number of watchers had increased by 15%
annually to more than 1.5 million in 2003, and this activity
contributed approximately AUD$300 million to the Australian
economy (IFAW, 2004). Since the ‘swim with the whale shark’
experience commenced at Ningaloo in 1989, the Department of
Conservation and Land Management (CALM, 2005) reported that
the industry increased to 15 licensed operators and was worth AUD
$15 million annually to the region.
Tourists are increasingly offered an opportunity to swim with
whales. For example, during the 1999e2000 season over 450
passengers swam with dwarf minke whales in the Great Barrier
Reef Marine Park (Valentine et al., 2004). Although a relatively new
tourism opportunity with whales, swimming with whale sharks in
Western Australia has existed since 1989, however Catlin and Jones
(2010) suggested that while the main attraction for visiting Nin-
galoo Marine Park in 1995 was to experience the wildlife
(predominantly whales and sharks) they noted that by 2005 the
demand for this type of recreation had plateaux. Visitor expecta-
tions had apparently changed. This was indicated by the wider
distribution of age groups among tourists, fewer were qualified
divers, there was a higher tolerance of crowding, and there was
a stronger focus on the non-wildlife components of the experience.
Managementof suchencounters varies due tospecies. Watching
and interaction codes of behaviour for commercial tour operators
apply differently even among marine mammals such as small (e.g.,
dolphins, porpoises) and large (e.g., dwarf minke Balaenoptera
acutorostrata, humpback M. novaeangliae) whales, and the whale
shark Rhincodon typus (Catlin and Jones, 2010; Davenport and
Davenport, 2006; Valentine et al., 2004). Such codes are intended
to reduce animal stress and, as far as possible, allow the animals to
control the level of interactions with humans. In practice, enforcing
such codes is difficult, and non-compliance is often recorded. For
example, Scarpaci et al. (2003) reported that none of the
commercial dolphin swimming and watchingtouroperators in Port
Phillip Bay (Melbourne) complied with all permit conditions.
One issue is that such restrictions may affect visitor satisfaction.
A significant correlation has been found between visitor satisfac-
tion and closeness of approach by animals, total number of animals
observed, and total time spent with the animals (Davis et al., 1997;
Valentine et al., 2004). Valentine et al. (2004) suggested that these
findings have implications for sustainable management of such
activities. They suggested that one approach to better protection of
the target animals could be achieved by restricting the length of
season of access by visitors, and limiting the length of encounters.
They also suggested that visitor’s expectations of encounters with
the animals need to be managed.
Although limited, Australian-based studies have not revealed
obvious behavioural changes or long-term detrimental effects on
the animals. For example, although dwarf minke whales have been
reported to show possible threat displays towards swimmers (e.g.,
bubble blasts, ‘gaping’, engulfment behaviour), no aggression
resulted, even by cows with calves (Birtles et al., 2002). Some
behaviours of whale sharks (e.g., diving away, porpoising, eye
rolling, banking, shuddering) may be reactions to human activities
(e.g., snorkeler/vessel proximity, flash photography, physical
interaction, obstruction of the shark’s path, SCUBA; Catlin and
Jones, 2010), however equivalent behaviours have also been
observed in the absence of snorkelers and vessels (Gunn et al.,
1999). Less equivocal are the results of several decades of obser-
vations on the response of the Indo-Pacific bottlenose dolphin to
vessel activity in Shark Bay Marine Park (Western Australia), where
increased numbers of tour operator vessels were associated with
a significant decline in dolphin abundance (Bejder et al., 2006).
3. Recreational impacts of non-indigenous species
One conclusion that Pickering and Hill (2007) reported in their
review of the impact of recreation on the terrestrial flora of pro-
tected areas was that there were no apparent recreation-specific
impacts on any threatened terrestrial plant species. They did,
however note the potential role of recreationists in the dispersal of
the introduced and destructive soil-borne pathogen Phytophthora
cinnamomi. It was suggested that while recreation was not the
source of its introduction to Australia, its spread could be exacer-
bated by recreation activities. A marine analogy is that of non-
indigenous species of flora and fauna.
Introduction of such exotic species poses significant risks,
including direct competition for resources (e.g., habitat, food),
predation, disease, toxic species, and degradation of habitats. Once
established in a particular location, they can be subsequently
dispersed to other locations and habitats, ultimately altering native
species assemblages, diversity and abundance, potentially over
a very wide geographic area. More than 250 non-native species
now occur in Australian coastal waters, ranging from plankton and
algae to fish, and there is the on-going potential for further invasion
(DEWHA, 2009b). For example, in Port Phillip Bay (Melbourne) it
was proposed that an exotic species becomes established every
three to six months (Thresher, 1999). Approximately 50 species of
marine organisms are considered to have been introduced into
New South Wales coastal waters, often with significant deleterious
effects. Examples include the invasive alga Caulerpa taxifolia
recorded in several estuaries of New South Wales, where it has
replaced seagrass beds that were important fish habitats. Several
toxic dinoflagellates have also been recorded, including Alexan-
drium catenella, a causative agent of paralytic shellfish poisoning.
Major fish kills have also been linked to the introduction of exotic
disease pathogens. Other introductions include the European shore
crab Carcinus maenas, the European fanworm Sabella spallanzani
and the Pacific oyster Crassostrea gigas (DPI, 2005).
Discharge of foreign-sourced ballast water from commercial
ships has long been recognised as a major sourceof non-indigenous
species in marine/estuarine ecosystems (Carlton, 1985; Low, 1999;
Williams et al., 1988). Such ballast tanks are generally restricted
to large freight vessels (Low, 1999; Thresher, 1999) rather than the
relatively small boats typically used for recreation. Increased size,
number and diverse destinations of cruise liners with such ballast
tanks can, however extend the potential route for morewidespread
invasion. Privately-owned recreational vessels or car ferries have
also been shown to be vectors of non-indigenous species on fouled
hulls, in sea chests, and/or seawater pipes (Coutts et al., 2003;
Davenport and Davenport, 2006; NIMPIS, 2002a).
One such documented invasion is the black-striped mussel
Mytilopsis sallei, a close relative of the zebra mussel Dreissena pol-
ymorpha. It was accidentally introduced to marinas near Darwin
(Northern Territory), probably originating on fouled hulls or in
seawater piping of visiting international yachts (Thresher, 1999;
NIMPIS, 2002a). Although detected ‘early’, its eradication necessi-
tated the efforts of 270 people and the decontamination of
hundreds of recreational and fishing boats at a cost of more than
AUD$2 million, together with the destruction of all taxa within
three enclosed marinas (Bax et al., 2001; Goggin, 2004). A similar
eradication operation was required in 2001 following the discovery
of the invasive Asian green mussel Perna viridis on a recreational
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
vessel’s hull in Cairns Harbour (North Queensland; Goggin, 2004;
Neil, 2002; NIMPIS, 2002b).
Although not highly invasive, the Caribbean tubeworm Hydoides
sanctaecrucis has also been introduced to the Cairns Harbour, and
other Australian ports and marinas. It has become a hull-fouling
nuisance species. A more serious threat is posed by the northern
Pacific seastar Asterias amurensis (NIMPIS, 2002c). Introduced from
Japan into Tasmanian waters in the 1980s, this exotic species had
spread to Port Phillip Bay by 1998. Eradication attempts have failed
and due to that port’s importance as a major hub for recreational
(domestic and commercial) traffic it is likely to be dispersed to
other Australian ports (Thresher, 1999).
The degree of hull fouling depends on a hull’s susceptibility to
such recruitment, and the local availability of competent plank-
tonic propagules. Management strategies to combat the problem
have typically focused on the use of toxic paints such as TBT.
Studies of marinas in Cairns and Townsville (North Queensland)
have, however, indicated that the degree of fouling may also be
influenced by the design of the harbour in which the boat is
moored. For example, recruitment of sessile invertebrates to
available surfaces has been shown to be several orders of magni-
tude more abundant in enclosed marinas (e.g., near breakwaters)
than in unenclosed marinas. Barnacles recruit most densely in
open locations (Floerl and Inglis, 2003). These differences may be
because although entrapment of water in enclosed marinas may
limit the dispersal of planktonic propagules, it also increases
propagule recruitment pressure to available surfaces, including
residentboat hulls. Harbour
a substantial influence on the size and frequency of exotic species
Using deductive hazard assessment, Hayes and Sliwa (2003)
identified potential ‘next pests’ for Australian marine habitats.
They used selection criteria that they considered transparent and
consistent with other biosecurity initiatives elsewhere in the
world: (a) the species had been previously reported as a shipping
vector or had a ship-mediated invasion history; (b) the vector
existed; (c) the species has been shown to be responsible for
economic or environmental harm; and, (d) it was not native to
Australian waters or currently present in Australia but subject to
official control. From over 850 species identified worldwide to be
invasive, they identified 31 that fitted all of their criteria. With an
estimated 2e4 species annually being introduced just to one
Australian harbour (Port Phillip Bay; Thresher, 1999), the threat of
a recreation-mediated introduction is real e there is the potential
for a large, ocean-going yacht or cruise ship to act as a vector for the
introduction of exotic species, and smaller recreation vessels could
then realistically contribute to their subsequent dispersal among
ports and marinas.
An Intergovernmental Agreement on a National System for the
Prevention and Managementof Marine Pest Incursions was mooted in
April 2005. This proposed System may be seen as an aquatic
analogy to Australia’s National Weeds Strategy, originally proposed
in 1997 and adopted in revised form in November 2006 (DEWHA,
2009c). It specifically acknowledged the role of recreational boat-
ing as a vector in the introduction and/or spread of non-indigenous
species. The aims of the proposed intergovernmental agreement
are to (i) prevent the arrival of non-indigenous species into
Australian waters or to stop their spread to new areas; (ii) provide
a coordinated emergency response to new exotic species; and
(iii) control and manage those already present where eradication is
not feasible. The Australian, Victorian, Tasmanian, Northern Terri-
tory and South Australian governments are signatories to this
Agreement. Western Australia and the two States with the highest
recreational vessel ownership, Queensland and New South Wales,
are yet to sign the document (DEWHA, 2009b).
Tourism is now the largest single industry in the world
(Davenport and Davenport, 2006), and one of the three largest in
financial and employment terms within Australia’s national
economy (TRA, 2007). The coast is the industry’s primary asset
(DEWHA, 2007). As the only island continent, there is an extensive
coastline including areas with the densest human population (e.g.,
Sydney coastal region) and some of the most sparsely populated
(e.g., sections of the north-west coast), together with the draw of
the iconic Great Barrier Reef. Issues associated with management of
such a large and varied geography are exacerbated by the frag-
mented distribution of jurisdiction among states,territories andthe
Coastal marine recreation has a wide range of sources and
types of impacts emanating from both infrastructure and activi-
ties, direct and indirect. Internationally, most research on the
impacts of marine recreational activities has tended to focus on
water pollution such as those due to boating activities (e.g.,
Mosisch and Arthington, 2001, 2004; Davenport and Davenport,
2006; Kotrikla, 2009), and physical damage associated with
SCUBA diving activities and infrastructure (e.g., Lynch et al., 2004;
Pears and Williams, 2005; Rouphael and Inglis, 2001; Smith et al.,
2005). Evidence is mounting, however that many other, perhaps
more subtle impacts occur such as the result of changes in animal
behaviour (e.g., Davis et al., 1997; Liddle, 1997; Valentine et al.,
2004), and recreational fishing. The scale of such impacts may
also be higher than previously recognised, for example from
fishing, both in overall terms (% of total catch) or disproportionally
on particular species and/or ecosystems (e.g., Lynch, 2006; West
and Gordon, 1994; Westera et al., 2003). Sustainable manage-
ment of even the most long-established and researched activity
(commercial fishing) of marine ecosystems has proved elusive,
reflected in it becoming a major conservation issue globally
(Botsford et al., 1997; Cooke and Cowx, 2004; Roberts, 1997). The
impacts of recreational fishing are less well known but are
probably substantial, based on the Canadian figures for recrea-
tional fishing of Cooke and Cowx (2004). They estimated that
recreational fishing represented more than 10% of the commercial
catch worldwide, but unlike commercial fishing it is more likely to
Ward et al. (2001) suggested that while Australian fisheries are
not over-fished, ‘many’ are fully exploited and, although empirical
data are limited, ‘no-take’ marine reserves offer a potential
management tool. Conservation through marine reservation is
becoming increasingly popular (e.g., Agardy et al., 2003; Crosby
et al., 2000), however Agardy et al. (2003) warn that they need
to be underpinned with effective conservation science based on
a ‘factual foundation’ and an understanding of the long-term
implications. Difficulties arise with attempting to ensure such
a basis for recreation management because ecological impacts
from recreation are often gradual and difficult to quantify, even in
terrestrial environments (Buckley, 2003; Hadwen et al., 2007; Kuss
et al., 1990; Newsome et al., 2002), and in the aquatic environ-
ments changes are even less immediately obvious (Kuss et al.,
1990; Liddle, 1997). Threshold stress levels are also difficult to
define and measure (Davis and Tisdell,1996). However even where
there are legislative mechanisms in place to monitor such impacts,
at least historically, the opportunity has not been seized. For
example, many forms of recreation infrastructure, including
marinas and pontoons, are ‘designated developments’ under
Australian laws and require an Environmental Impact Assessment
as part of the planning process. Although the impacts of localised
high-intensity tourism developments such as hotels and marinas
arewellknown comparedto the impacts of dispersed
N. Hardiman, S. Burgin / Journal of Environmental Management 91 (2010) 2096e2108
non-mechanised activities (which are not) the scientific quality of
the Australian environmental impact procedures for such tourism
developments is apparently weak. For example, of 76 tourism
projects EIA-approved between 1987 and 1993 and subsequently
developed, post-development monitoring was only conducted for
17. Most only required monitoring of water quality and none
monitored biota. Only four projects in the Great Barrier Reef
Marine Park and an alpine resort development were monitored for
aquatic biota (Warnken and Buckley, 1998, 2000).
By January 2005 there were 37 permitted semi-permanent or
permanent pontoons in the Great Barrier Reef Marine Park (Smith
et al., 2005). Their construction requires an initial Environmental
Impact Assessment and, subsequently, on-going monitoring,
insurance cover, and a posted performance bond for cleanup and
site remediation in the case of accident. Such bonds may be
lodged by the operator as a cash deposit or, more commonly,
a guarantee from a financial institution. All such bonds that are in
excess of $50,000 are adjusted annually against the Brisbane
consumer price index (Lal and Brown, 1996; Smith et al., 2005).
Equivalent bonds are widely used throughout Australia in the
recreational industry, although even more commonly in the
mining and property rental market sectors. Between their intro-
duction in 1987 and 1996, the Great Barrier Reef Marine Park
Authority called upon bonds only four times. It prefers to use the
operator’s insurance cover, especially when cleanup and removal
is required due to a natural disaster. Their effectiveness in
providing adequate rehabilitation and efficient use of resources
depends on the specification of the original resource conditions,
and on the form of the bond posted. Only the level, and not the
method of site rehabilitation is stipulated, and this is site-specific,
rather than an average across multiple sites which is a common
situation in many industries. Such site-specific rehabilitation
requirements encourage operators to consider possible rehabili-
tation costs early in their project planning and adoption of the
most efficient scale and mode of operation. Experience has proved
them administratively simple and enforceable, provided the level
of rehabilitation is clearly defined in a legal document (Lal and
The greatest risk of negative ecological impact from coastal
marine recreation is, however, arguably the spread of non-indige-
nous species which are capable of fundamentally restructuring
entire ecosystems. Recreational vessels have the potential to
introduce and rapidly spread such pests. Despite this, appreciation
of the risk among recreationists is probably low. Prevention and
control has the potential to be enhanced, however, with the current
initiative to implement a national aquatic pest strategy similar to
the National Weeds Strategy (DEWHA, 2009c).
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and the increasing recognition of the associated environmental
impacts (IUCN, 1996; Pickering and Hill, 2007), the long-term
viability of the ecosystems and the associated biodiversity that the
industry relies on are increasingly vulnerable to collapse due to the
impacts of the humans who visit the areas. In response govern-
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