Threats to sandy beach ecosystems: A review
, Anton McLachlan
, David S. Schoeman
, Thomas Schlacher
, Jenifer Dugan
, Mariano Lastra
, Felicita Scapini
Facultad de Ciencias. Unidad de Ciencias del Mar. Igua
´4225, 11400 Montevideo, Uruguay
Dean of Postgraduate Studies, Sultan Qaboos University, Muscat, Oman
School of Biological and Conservation Sciences, Howard College, University of KwaZulu-Natal, Durban, South Africa
Faculty of Science, Health, and Education, University of the Sunshine Coast, Maroochydore, Qld 4558, Australia
Marine Science Institute, University of California, Santa Barbara, CA 93106, USA
Marine Ecology, Australian Museum Sydney, 6 College Street, Sydney, NSW, Australia
Departamento de Ecologı
´a y Biologı
´a Animal, Universidad de Vigo, 36310 Vigo, Spain
Dipartimento di Biologia Animale e Genetica ‘‘Leo Pardi’’, Universita
`di Firenze, Florence, Italy
Received 7 March 2008
Accepted 23 September 2008
Available online xxx
We provide a brief synopsis of the unique physical and ecological attributes of sandy beach ecosystems
and review the main anthropogenic pressures acting on the world’s single largest type of open shoreline.
Threats to beaches arise from a range of stressors which span a spectrum of impact scales from localised
effects (e.g. trampling) to a truly global reach (e.g. sea-level rise). These pressures act at multiple
temporal and spatial scales, translating into ecological impacts that are manifested across several
dimensions in time and space so that today almost every beach on every coastline is threatened by
human activities. Press disturbances (whatever the impact source involved) are becoming increasingly
common, operating on time scales of years to decades. However, long-term data sets that describe either
the natural dynamics of beach systems or the human impacts on beaches are scarce and fragmentary. A
top priority is to implement long-term ﬁeld experiments and monitoring programmes that quantify the
dynamics of key ecological attributes on sandy beaches. Because of the inertia associated with global
climate change and human population growth, no realistic management scenario will alleviate these
threats in the short term. The immediate priority is to avoid further development of coastal areas likely to
be directly impacted by retreating shorelines. There is also scope for improvement in experimental
design to better distinguish natural variability from anthropogenic impacts. Sea-level rise and other
effects of global warming are expected to intensify other anthropogenic pressures, and could cause
unprecedented ecological impacts. The deﬁnition of the relevant scales of analysis, which will vary
according to the magnitude of the impact and the organisational level under analysis, and the recog-
nition of a physical–biological coupling at different scales, should be a suitable approach to quantify
impacts. Zoning strategies and marine reserves, which have not been widely implemented in sandy
beaches, could be a key tool for biodiversity conservation and should also facilitate spillover effects into
adjacent beach habitats. Setback and zoning strategies need to be enforced through legislation, and all
relevant stakeholders should be included in the design, implementation and institutionalisation of these
initiatives. New perspectives for rational management of sandy beaches require paradigm shifts, by
including not only basic ecosystem principles, but also incentives for effective governance and sharing of
management roles between government and local stakeholders.
Ó2008 Elsevier Ltd. All rights reserved.
The accelerating destruction of natural habitats and consump-
tion of natural resources by rapidly expanding human populations
has caused huge impacts to ecosystems across the globe. Ampliﬁed
by demographic population shifts towards the coast (Roberts and
Hawkins, 1999), many of these impacts are focussed at the world’s
coastlines, which are dominated by sandy shores (McLachlan and
Brown, 2006). Intense coastal development, the inevitable conse-
quence of economic progress, has resulted in widespread modiﬁ-
cation of sandy beach ecosystems. Human changes to sandy shores
began at least two centuries ago (Nordstrom, 2000), and are
predicted to intensify exponentially over the next few decades
(Brown et al., 2008). Global climate change, particularly sea-level
E-mail address: firstname.lastname@example.org (O. Defeo).
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rise, has added a new dimension to worldwide modiﬁcations of
shorelines (Jones et al 2007a; Schlacher et al., 2008b). It has
therefore become increasingly critical to understand how sandy
beach ecosystems and the goods and services they provide will
respond to this unprecedented environmental change.
Exposed sandy beaches are physically dynamic habitats,
inhabited by specialised biotic assemblages that are structured
mainly by physical forces (reviewed in Defeo and McLachlan, 2005).
Building on this perspective, documenting biotic responses to
modiﬁcations of the physical environment is a critical step in
predicting the consequences of global coastal change for beach
ecosystems. Evidence for ecological change in beach ecosystems,
which are exposed to human pressures at scales and intensities
unmatched in history, is accumulating worldwide (Brown and
McLachlan, 2002; Schlacher et al., 2007a).
Here we provide a brief synopsis of the unique physical and
ecological attributes of sandy beach ecosystems and review the
main anthropogenic pressures acting on the world’s single largest
type of open shoreline. Threats to beaches arise from a range of
stressors that span a spectrum of impact scales from localised
effects (e.g. trampling) to a truly global reach (e.g. sea-level rise).
These pressures act at multiple temporal and spatial scales (Fig. 1),
translating into ecological impacts that are manifested across
several dimensions in time and space so that today almost every
beach on every coastline is threatened by human activities.
2. Ecological features
2.1. The physical environment
Sandy beaches, deﬁned by their sand, wave and tidal regimes,
range from narrow and steep (reﬂective) to wide and ﬂat
(dissipative), as sand becomes ﬁner and waves and tides larger;
most beaches are intermediate between these extremes (Short,
1999; Finkl, 2004). Reﬂective beaches are coarse-grained and have
no surf zones, whereas dissipative beaches have ﬁner sediments
and extensive surf zones. Filtration volumes are higher on
permeable reﬂective beaches, mainly driven by wave action, and
lower on dissipative beaches, where tidal action drives most water
throughput. Thus ﬁltration is rapid on reﬂective beaches and
slower on dissipative beaches, but the sand body of all open bea-
ches is well ﬂushed and oxygenated; only under conditions of very
ﬁne sand, such as on some dissipative and low energy beaches, can
reducing conditions develop in the deeper sediment layers
(McLachlan and Turner, 1994).
Beaches are closely linked to surf zones and to coastal dunes
through the storage, transport and exchange of sand; therefore
impacts on beaches have consequences for these adjacent habitats
(Komar,1998). Sand transport, driven by waves on the wet side and
wind on the dry side, is highest in exposed surf zones, whereas
sand storage is often greatest in well-developed dunes. Sand tends
to move rapidly seawards across the beach and surf zone during
storms and to return more slowly landwards during calms. In this
way, stormwave energy is dissipated and the soft coast is protected
from permanent erosion (Short, 1999; Nordstrom, 2000).
2.2. Faunal components and ecosystem properties
The intertidal areas of beaches provide habitat for a diversity of
fauna. The lacunar environment between the grains harbours
interstitial organisms (bacteria, protozoans and meiofauna),
forming a distinct food web. Larger macrobenthic invertebrates
burrow actively and include representatives of many phyla, but
crustaceans, molluscs and polychaete worms are usually dominant
and encompass predators, scavengers, ﬁlter- and deposit feeders.
These macrobenthic invertebrates can reach high abundance
(ca. 100,000 ind m
) and biomass (>1000 g m
), particularly in
dissipative to intermediate beach types in temperate zones. Bea-
ches that receive signiﬁcant inputs of algae/seagrass wrack support
a rich supralittoral fauna of crustaceans and insects. Most beach
species are found in no other environment, their unique adapta-
tions for life in these dynamic systems include: mobility, burrowing
ability, protective exoskeletons, rhythmic behaviour, orientation
mechanisms and behavioural plasticity (Chelazzi and Vannini,
1998; Scapini et al., 1995; Brown, 1996; Scapini, 2006)
The composition and abundance of invertebrate assemblages
are controlled primarily by the physical environment, intertidal
swash and sand conditions, being harshest on reﬂective beaches
and more benign on dissipative beaches. Consequently, more
species can colonise dissipative beaches, but fewer species, mainly
robust crustaceans, can establish populations on reﬂective beaches
(McLachlan and Dorvlo, 2005). Whereas the effects of biological
interactions (e.g. competition, predation) are overshadowed by
physical factors on reﬂective beaches, they become more inﬂuential
in structuring communities on dissipative beaches (Defeo and
Faunal patterns change on either side of the intertidal zone.
Supralittoral zones are important nesting areas for turtles and
shorebirds, and provide afavourable habitat for invertebrates on
stable reﬂective beaches. The fauna of the lower beach may extend
their distribution seawards into the turbulent surf zone, where
zooplankton, shrimps and prawns can be abundant; surf zones are
also important nursery and foraging areas for ﬁshes.
Food webs of sandy beaches are based on marine sources, such
as phytoplankton, stranded algae, seagrasses and carrion
(McLachlan and Brown, 2006). In dissipative systems, high
productivity may be driven by surf phytoplankton and microor-
ganisms, supporting benthic macrofauna and zooplankton as
primary consumers and ﬁshes and birds as top predators. While
ﬁltering water, the porous sand body and its biota mineralise
organic matter and recycle nutrients, making beach ecosystems
a crucial element in the nearshore processing of organic matter and
10 100 1000
and sea-level rise
and urban development
Recreation and ORVs
Fig. 1. Conceptual model and schematic diagram showing the relative spatio-temporal
scales in which different impacts reviewed here generally operate on sandy beach
macrofaunal communities. Boxes/envelopes indicate the potential extent of individual
impacts in space and time with the lower curve reﬂecting the lower limit of impacts in
time and space, whereas the upper curve reﬂects the corresponding maximum.
Shorter term impacts (i.e., weeks to months) tend to be pulse disturbances and effects
are generally expected to last for shorter time periods, since sandy-beach species are
adapted to severe physical disturbances (e.g., storms). However, the temporal extent of
impacts from the anthropogenic pressures depicted here could be drastically altered if
the intensity of the disturbance is increased and/or its timing is more protracted.
Under such circumstances, sandy beach habitats could become unsuitable for
supporting macrofaunal communities and the ecosystem services they provide in the
medium or long term.
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Not only are beaches coupled to marine systems trophically, but
they also interact physically and biologically with coastal dunes.
Besides sediments, beaches and dunes exchange a variety of
organic materials, and animals from both habitats move across the
dune/beach interface to feed. Finally, beaches link terrestrial dune
aquifers with coastal seas through the discharge of nutrient-rich
2.3. Ecosystem services and sensitive features
Sandy shores provide a wide range of ecosystem services, many
of which are essential to support human uses of sandy coasts. The
most important ecosystem services include: (1) sediment storage
and transport; (2) wave dissipation and associated buffering
against extreme events (storms, tsunamis); (3) dynamic response
to sea-level rise (within limits); (4) breakdown of organic materials
and pollutants; (5) water ﬁltration and puriﬁcation; (6) nutrient
mineralisation and recycling; (7) water storage in dune aquifers
and groundwater discharge through beaches; (8) maintenance of
biodiversity and genetic resources; (9) nursery areas for juvenile
ﬁshes; (10) nesting sites for turtles and shorebirds, and rookeries
for pinnipeds; (11) prey resources for birds and terrestrial wildlife;
(12) provision of scenic vistas and recreational opportunities;
(13) supply of bait and food organisms; and (14) functional links
between terrestrial and marine environments in the coastal zone.
Certain features and processes on sandy shores are particularly
vulnerable to impacts or change: (1) large volumes of sand are
transported under high-energy conditions, so the consequences of
disturbing sand budgets are most severe in such situations;
(2) under low-energy conditions, water circulation is limited and
therefore dispersal of materials, such as pollutants, is slow;
(3) foredune vegetation is easily disturbed with resultant destabi-
lisation of dunes; (4) because many beaches rely on marine organic
subsidies, removal of such inputs (e.g. beach grooming) has serious
consequences for beach food webs; (5) vertebrates that nest on the
backshore are considered especially important in coastal conser-
vation, and both turtles and shorebirds are easily disturbed in their
Recreational seashore activities are overwhelmingly concen-
trated on sandy beaches. The effects of these pressures are partic-
ularly noticeable at scales ranging from weeks to months and from
<1to10km(Fig. 1). Burgeoning coastal populations, coupled with
more leisure time and improved mobility, have escalated the
intensity and spatial ambit of beach recreation over recent decades
(De Ruyck et al.,1997; Caffyn and Jobbins, 2003; Fanini et al., 2006).
Being the prime sites for human recreation, beaches underpin
many coastal economies (Klein et al., 2004). Beach management
therefore customarily focuses on maximising the recreational
experience for beach users, which often results in ecologically
harmful human interventions such as nourishment (Speybroeck
et al., 2006), beach grooming (Llewellyn and Shackley,1996; Dugan
et al., 2003), coastal armouring (Dugan and Hubbard, 2006; Dugan
et al.,2008), destruction of dunes to construct tourism infrastruc-
ture (Nordstrom, 2000), and light and sound pollution (Bird et al.,
2004; Longcore and Rich, 2004).
Impacts caused directly by recreational activities are emerging
as signiﬁcant environmental issues (Schlacher et al., 2008b). Dune
vegetation is vulnerable to mechanical impacts caused by
trampling (Liddle and Grieg-Smith, 1975), and modern beach
management practices progressively seek to restrict human access
to these sensitive areas (Scapini, 2002). Nevertheless, camping and
driving activities, which severely impact dune vegetation, both
continue unabated in many parts of the world (Luckenbach and
Bury, 1983; Hockings and Twyford, 1997; Groom et al., 2007).
Evidence is sparse about how sensitive intertidal invertebrates
might be to human trampling. Direct crushing of individual
organisms on the unvegetated beach has been documented (Mof-
fett et al., 1998), and macrobenthic populations and communities
respond negatively to increased human activity levels (Weslawski
et al., 2000; Fanini et al., 2005; Veloso et al., 2006), but not in all
cases (Jaramillo et al., 1996). It can be difﬁcult to separate the effect
of human trampling from habitat modiﬁcations (e.g. seawalls
replacing foredunes) because these often coincide in high-use areas
Impacts of trampling on supralittoral fauna are usually seasonal
at medium-high latitudes, enabling some recovery during periods
of low use (see e.g., Scapini et al., 2005). However, press distur-
bances are expected on low-latitude beaches where the intensity of
the disturbance is much higher than in temperate areas and its
timing is more protracted. Under such circumstances, beach habi-
tats could become unsuitable for sandy beach macrofauna in the
medium or long term. It is therefore necessary to establish the
ecological carrying capacity of beaches in terms of direct use, and to
develop mechanisms for controlling access, tasks that many
authorities neglect (De Ruyck et al., 1997).
Beaches and dunes are critical habitats for shorebirds and turtles
(Hosier et al., 1981; Hubbard and Dugan, 2003) and both taxa are
sensitive to disturbances. For example, human activities disturb
shorebirds, modifying key behavioural traits that are crucial to their
survival and reproduction (Burger, 1991, 1994; Lord et al., 2001;
Verhulst et al., 2001), including: (1) changes to foraging behaviour
resulting in less feeding time, temporal shifts in feeding times and
decreased food intake; (2) decreased parental care when disturbed
birds spend less time attending the nest, thus increasing exposure
and vulnerability of eggs and chicks to predators; and (3) decreased
nesting densities in disturbed areas and population shifts to less
Off-road vehicles (ORVs) are commonly used on beaches and
dunes worldwide (Godfrey and Godfrey, 1980; Priskin, 2003;
Schlacher and Thompson, 2007) and cause damage that includes:
(1) disturbing the physical attributes and stability of dunes and
beaches by deeply rutting the sand surface and destroying
embryonic foredunes in the tyre tracks (Anders and Leatherman,
1987; Kutiel et al., 1999; Priskin, 2003; Schlacher and Thompson,
2008); (2) destroying dune vegetation, leading to lower diversity
and less ﬂoral ground cover (Luckenbach and Bury, 1983; Rickard
et al., 1994; Groom et al., 2007); and (3) disturbing, injuring or
killing beach fauna (Van der Merwe and Van der Merwe, 1991;
Schlacher et al., 2007b, 2008a), including endangered vertebrates
such as turtles and shorebirds (Hosier et al., 1981; Buick and Paton,
1989; Williams et al., 2004). Whether direct mortality of beach
invertebrates caused by ORVs propagates to higher-order ecological
effects, such as disruptions of food web linkages to ﬁshes and
raptors, is currently unresolved. Nevertheless, positive population
responses by shorebirds following exclusion of ORVs (Williams
et al., 2004) suggest that ORV-impacts can extend beyond the
individual organism level of ecological organisation.
Cleaning or grooming, a common practice on beaches heavily
used for tourism (Poinar, 1977; Llewellyn and Shackley, 1996;
Engelhard and Withers, 1997; Dugan et al., 2003; Davenport and
Davenport, 2006), clears beaches of macrophyte wrack, litter and
other debris by raking and sieving the sand, often with heavy
equipment (Kinzelman et al., 2003). Grooming removes not only
unwanted material, but also propagules of dune plants and other
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species, and it perturbs resident organisms and roughens the sand,
thereby exposing a greater surface area to the erosive effects of
Wrack removal has signiﬁcant ecological consequences, espe-
cially in regions with high levels of marine macrophyte production.
The community structure of sandy beach macroinvertebrates can
be closely linked to wrack deposits (e.g. Grifﬁths and Stenton-
Dozey, 1981; Grifﬁths et al., 1983; Stenton-Dozey and Grifﬁths,
1983; McLachlan, 1985; McGwynne et al., 1988; Inglis, 1989;
Colombini and Chelazzi, 2003; Dugan et al., 2003; de la Huz et al.,
2005), which provide both a food source and a microhabitat refuge
against desiccation (Colombini and Chelazzi, 2003). Wrack-associ-
ated fauna, such as talitrid amphipods, oniscoid isopods and
insects, which can comprise up to 40% of the species and represent
important prey for higher trophic levels (Dugan et al., 2003), are
signiﬁcantly reduced in species richness, abundance, and biomass
by grooming (Dugan et al., 2003; Fanini et al., 2005). By contrast,
taxa with well-developed dispersal abilities, such as ﬂies, can be
more prevalent on groomed beaches (Dugan et al., 2003).
Upper intertidal meiofauna may also be impacted by grooming,
although these effects can be confounded with those from tram-
pling (Gheskiere et al., 2005). Meiofauna communities may recover
relatively quickly (24 h) from a single, short-term grooming event
(Gheskiere et al., 2006), but it remains unknown what the conse-
quences are of repeated, regular beach cleaning activities. Effects of
this stressor could be noticeable at scales ranging from weeks to
years and from <1 to 100 km (Fig. 1).
Because shorebird numbers are positively correlated with wrack
cover and the biomass of their invertebrate prey that feed on wrack
(Tarr and Tarr,1987; Hubbard and Dugan, 2003; Dugan et al., 2003),
grooming will lower bird numbers. The heavy equipment used in
beach grooming can also cause direct mortality of the eggs and
young of beach-nesting shorebirds, turtles and ﬁsh (Martin et al.,
2006); many groomed beaches therefore no longer support
breeding populations of these vertebrates. Furthermore, tracks
created by grooming can disorientate turtle hatchlings trying to
reach the sea (Hosier et al., 1981). Finally, grooming can result in
abnormally broad unvegetated zones that are inhospitable to dune
formation or plant colonisation, thereby enhancing the likelihood
Invertebrates may recover slowly from the effects of grooming,
especially if it is conducted on a daily or weekly basis throughout
the year (Fig. 1), as is the case along many developed coastlines. For
example, nearly half (>160 km) of the sandy beaches in southern
California are regularly groomed (Dugan et al., 2003), and some
beaches in Los Angeles are groomed twice a day (Dugan et al.,
2000). Although beach cleaning is generally undesirable from an
environmental conservation perspective, simple mitigation strate-
gies could entail alternating areas of natural (uncleaned) beach
with groomed sections, if grooming is deemed essential to provide
clean beaches for recreation.
More than 70% of the world’s beaches are experiencing erosion
(Bird, 1996). Because engineering solutions, such as seawalls,
breakwaters and groynes, are often ineffective to the point of
causing the loss of the intertidal beach (Pilkey and Wright, 1989,
Hsu et al., 2007), beach nourishment (also called beach replenish-
ment, restoration or renourishment), has increasingly been used to
combat shoreline erosion. Nourishment is preferred on both
economic and conservation grounds (Finkl and Walker, 2004), but
it can cause ecological damage to sandy beach habitats (Blott and
and biota (reviewed in Goldberg 1988; Nelson, 1988;
Peterson and Bishop, 2005; Speybroeck et al., 2006). The typical
scale of nourishment is at 1–10 km (Peterson and Bishop, 2005) and
at temporal scales ranging from weeks to years (Fig. 1). Impacts
occur both at sites from which sediment is extracted and at the
receiving environment. These impacts manifest at the population
(demography and dynamics), community (species richness) and
ecosystem (functional processes, nutrient ﬂux, trophic dynamics)
levels. Affected biota includes benthic micro-algae, vascular plants,
terrestrial arthropods, intertidal and subtidal invertebrates, other
marine zoobenthos and avifauna (Bishop et al., 2006; Peterson
et al., 2006; Speybroeck et al., 2006; Fanini et al., 2007, 2009).
Factors inﬂuencing the nature and extent of ecological impacts
of nourishment include the mechanical process itself, its timing,
and the quality and quantity of new sediment placed (Speybroeck
et al., 2006). Effects may be direct, such as mortality of organisms
when buried, or indirect, such as reduced prey availability for
shorebirds (Nelson 1993a, b; Bishop et al., 2006; Peterson et al.,
2006). Most research has targeted the effects of altered sediment
quantity and quality on macrofauna (Hayden and Dolan, 1974;
Rakocinski et al., 1996; Peterson et al., 2000; Menn et al., 2003;
Bilodeau and Bourgeois, 2004; Jones et al., 2007b). The immediate
impacts are usually large and may be caused either by burial (Menn
et al., 2003; Peterson et al., 2006; Jones et al., 2007b)orby
emigration (Hayden and Dolan, 1974). These effects may be com-
pounded by changes in beach morphology, particularly when
nourishment creates a steeper beach and reduces the habitat area
for some species (Peterson et al., 2006; Fanini et al., 2007, 2009).
Nourishment can also disturb the nesting and foraging of birds,
destroy dune vegetation and compact the sand (Speybroeck et al.,
2006). Compaction affects the interstitial spaces, capillarity, water
retention, permeability and the exchange of gases and nutrients.
Nourishment of sandy beaches usually acts as a short-term,
pulse disturbance (Peterson and Bishop, 2005) that elicits a pulse
ecological response (i.e. recovery occurs). This is expected since
sandy-beach species are adapted to severe physical disturbances,
storm events having been a frequent feature of their evolutionary
history (Hall, 1994). Recovery probably occurs in months rather
than years (Fig. 1), but the trajectory depends on sediment quality
(Nelson, 1988,1993a, b; Peterson et al., 2000, 200 6). However, if the
proﬁle of the nourished beach and the imported sediments do not
match the original conditions, recovery of the benthos is unlikely
(Goldberg, 1988; Peterson et al., 2000, 2006). Thus, where unnat-
urally coarse or ﬁne sediments are used severe ecological impacts
may occur and recovery is protracted (Rakocinski et al., 1996;
Peterson et al., 2000, 2006). In addition, artiﬁcially ﬂattened and
extended sand bodies can be colonised by rapidly moving oppor-
tunistic macrofauna; under these conditions, few species dominate
and biodiversity is reduced (Peterson and Bishop, 2005).
Mitigation of ecological impacts of nourishment is often
impeded by limited data about the life history of the affected
species, recovery rates and the cumulative effects of repeated
nourishment events (Speybroeck et al., 2006). Nevertheless, basic
management recommendations include: (1) the avoidance of
sediment compaction;(2) careful timing of operations to minimise
biotic impacts and enhance recovery; (3) the selection of locally-
appropriate techniques; (4) the implementation of several small
projects rather than a single large project, including repeated
application of sediment in shallow layers (<30 cm) rather than
single pulses that kill the fauna by deep burial; (5) interspersion of
nourished beach sections with unaffected areas; and (6) importing
sediments and creating beach proﬁles that match the original
beach conditions as closely as possible.
Pollution is a sensitive issue on beaches given their immense
value for recreation and tourism. Pollutants act at a variety of
spatial and temporal scales (Fig. 1) and include a variety of
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anthropogenic materials, ranging in size from molecules to large
debris, and can impair the physiology, survival, reproduction and
behaviour of species in all habitats of the beach from interstitial
environments (McLachlan, 1977) to the surf zone (Noble et al.,
2006). Pollution can also cause aesthetic disturbances, thereby
impacting tourism industries that depend on public perceptions of
clean beaches (Tudor and Williams, 2003).
Most solid waste stranded on the shore is allochthonous,
brought ashore by waves and currents. Plastic is extremely
persistent and it dominates the visible litter on sandy beaches
worldwide (Derraik, 2002; Shiber, 2007). The major dangers asso-
ciated with plastic include ingestion by and entanglement of
vertebrates, such as seals, seabirds and turtles (Mascarenhas et al.,
2004). In terms of human health, risk increases when medical
plastic wastes reach the shore from coastal dumping sites
(Sindermann, 1996). Economic losses can also arise when tourist
beaches are persistently contaminated by litter from land or ocean
Wastewater and sewage are signiﬁcant sources of beach pollu-
tion. Intertidal sands (Ulﬁg et al., 1997; Salvo and Fabiano, 2007)
and surf-zone waters (Bonilla et al., 2006, 2007; Noble et al., 2006)
can be contaminated by pathogens, including bacteria and ﬁla-
mentous fungi, which are delivered into the system by sewage
systems discharging directly into coastal waters or estuaries near
beaches (Mardon and Stretch, 2004; Araujo and Costa, 2007).
Bacterial levels in surf zone waters that exceed human health
standards are a frequent cause of beach closures and public
warnings on many developed coasts. Contamination at any given
beach is dependent on: (a) the appropriateness of waste-water
management practices; (b) the timing and intensity of local rainfall
events and subsequent runoff; and (c) the strength of mixing and
dispersion in the surf zone (Stretch and Mardon, 2005). Metal
pollution from wastewater and industry accumulates preferentially
on ﬁne-grained beaches (Ramirez et al., 2005), which usually
supports relatively high diversity and biomass, thus causing
signiﬁcant reductions both in biodiversity and in the population
density of economically valuable species (Castilla, 1983; Haynes
et al., 1997).
Freshwater efﬂuents arising from human activities can also
cause a broader deterioration in the quality of the surrounding
habitat (Lercari et al., 2002), with impacts at various levels of
ecological organisation, ranging from individuals and populations,
which might experience reduced survival, growth and fecundity
rates (Defeo and de Alava, 1995; Lercari and Defeo, 1999; Lozoya
and Defeo, 2006), through to the community, which can exhibit
changes in diversity and structure (Lercari et al., 2002; Lercari and
Defeo, 2003; Defeo and Lercari, 2004).
Oil spills are potentially the most destructive pollution source
impacting sandy beaches, affecting all trophic levels (Bodin, 1988;
Suderman and Thistle, 2003). Impacts can be acute and temporary,
but they can also be more chronic, lasting for many months or even
years (Irvine et al., 2006). Beach morphodynamics and exposure
strongly inﬂuence the duration of contamination: the coarser the
sediment (i.e., reﬂective beaches), the more rapidly and deeply oil
penetrates, sometimes even reaching below the groundwater table
(Bernabeu et al., 2006). Sheltered beaches are generally more
sensitive to pollution than exposed beaches, even when sediments
are ﬁne and oil does not penetrate deeply, because they are less
well ﬂushed by wave action and subsurface oil persists much longer
than surface oil (Sinderman, 1996). Persistence and breakdown of
stranded oil depends on particle size, wave energy, temperature
and other factors (Owens et al., 2008), including fungal degradation
(Elshaﬁe et al., 2007)
. Methods used for oil spill cleanup on beaches
may also cause ecological impacts. Public concern about oil spills
has focused on catastrophes like the IXTOC 1 in Mexico (Rabalais
and Flint, 1983), the Exxon Valdez in Alaska (Short et al., 2004) and
the Prestige in Spain (de la Huz et al., 2005) and many smaller
events tend to go unreported, butintroduce considerable quantities
of oil into the beach environment. Tar balls that originate from the
(often illegal) ﬂushing of ships’ bilges at sea are present on most
beaches, especially near oil shipping routes (Corbin et al., 1993;
Coles and Al-Riyami, 1996; Abu-Hilal and Khordagui, 2007).
In most cases, management measures against pollution have to
be taken without a preceding environmental impact assessment, as
has occurred been the case of oil spills affecting beaches all over the
world. Oil pollution impacts differ between chronic cases and
episodic events. The former seldom require direct action, whereas
the latter may need a major cleanup, which in turn causes impacts.
Management requires monitoring and control of water and sedi-
ment quality, particularly in urban areas (Pereira et al., 2003).
Artisanal invertebrate ﬁsheries are the commonest form of
exploitation on sandy beaches (McLachlan et al., 1996; Kyle et al.,
1997). Although these are generally of reasonably small scale, for
various reasons impacts can be signiﬁcant (Fig. 1). First, target
species tend to occur in patches (Defeo and Rueda, 2002; Schoeman
and Richardson, 2002), and ﬁshers can therefore easily target and
serially deplete dense aggregations (Caddy and Defeo, 2003; Pe
´vez, 2004). Most exploited sandy beach stocks are also
short lived and susceptible to recruitment overﬁshing in the short
term (Defeo, 1996b). Moreover, harvesting activities cause inci-
dental mortalities, both directly through physical damage of
organisms and indirectly when sediment disturbance lowers
habitat quality and suitability (Sims, 1997; Defeo, 1998). Such
incidental mortality is not limited to the exploited fraction of the
target populations, but also affects their unexploited fractions and
non-target species on ﬁshed beaches (Defeo 1996a, 1998).
Long-term, large-scale ﬁeld experiments have demonstrated
a predictable pattern of responses to ﬁshing (Defeo, 1996a, 1998;
Castilla and Defeo, 2001): elevated harvesting rates signiﬁcantly
reduce recruitment to the exploitable stock and to the population
as a whole, resulting in a declining spawning stock, a decreasing
age/size at maturity and a decreasing proportion of older individ-
uals in the catch. Several ﬁshery-related density-dependent
processes have also been identiﬁed. For example, ﬁshery closures
can result in elevated adult densities, which cause density-depen-
dent increases in natural mortality rates of older clams (Defeo,
1996a), as well as decreases in growth and survival rates of young-
of the-year (Defeo, 1998), the inhibition of recruitment of both
target and sympatric species, and the reduction of age-speciﬁc
fertility (Defeo and de Alava, 1995; Brazeiro and Defeo, 1999; Lima
et al., 2000).
Where ﬁshing-related disturbance is a pulse of limited spatial
extent, its impacts can be ameliorated relatively quickly (Schoeman
et al., 2000). However, chronic disturbance can cause large-scale
and persistent impacts for populations and communities, as
documented for Argentinean beaches where ﬁshing involving the
use of tractors caused lasting damage to populations of yellow clam
Mesodesma mactroides for decades (Defeo, 2003). Detrimental
effects of ﬁshing can also be exacerbated when they are super-
imposed on environmental change, irrespective of whether such
change is of anthropogenic origin, or not (Defeo, 2003). On a New
Zealand beach, clam populations depleted by artisanal ﬁsheries
between the mid-1960s and 1990 failed to recover following the
closure of the ﬁshery (McLachlan et al., 1996), apparently because
continuous erosion of beaches has dramatically reduced the clam
habitat (Beentjes et al., 2006). Similarly, the effects of ﬁshing can be
ampliﬁed or reinforced by mass mortalities associated with blooms
of toxic algae, by parasitism and by temperature anomalies (Arntz
et al., 1987; Defeo, 2003; Fiori et al., 2004).
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Technological interdependencies among user groups and
ecological interdependencies between stocks are common in sandy
beach ﬁsheries (Defeo, 2003). Even ﬁsheries in which juveniles and
adults are spatially segregated are not immune to this phenom-
enon, with separate user groups often targeting intertidal and
subtidal stocks using different gear types (McLachlan et al., 1996).
Moreover, commercial ﬁshers and recreational anglers, who collect
large specimens for bait, tend to be more selective than recreational
food collectors (Schoeman, 1996; Murray-Jones and Steffe, 2000).
In this way, different user groups, which are often difﬁcult to
distinguish, can inﬂict different patterns of ﬁshing mortality on the
resource. Thus, beach ﬁsheries tend to involve a complex blend of
users, requiring the explicit incorporation of cultural, social and
political dimensions in resource management (Castilla and Defeo,
2001). As harvesting not only affects the targeted species directly
through ﬁshing mortality, but it also has collateral impacts that can
alter ecosystem structure and functioning, sandy beach ﬁsheries
can have impacts disproportionate to their economic value. Being
vulnerable to ﬁshing, few beach ﬁsheries have scope for growth.
Because it is difﬁcult to restrict participation in these accessible
ﬁsheries, management failure and stock depletion seem to be the
norm rather than the exception (reviewed in McLachlan et al., 1996;
3.6. Biological invasions
Human activities that are vectors for species introductions to
beaches are not a recent phenomenon. Among the oldest-known
human agents of species introduction to beaches is the practice of
exchanging large quantities of ‘‘dry’’ ballast sand and gravel on
shore during ship maintenance, which dates back to the early days
of ocean-going sailing vessels (Carlton, 1989).
Nearshore and intertidal benthic habitats in many regions have
been invaded by alien macrophytes that displace native algae and
seagrasses (Russell, 1992; Boudouresque et al., 1995; Inderjit et al.,
2006). Although not invaders of beaches in the sense of establishing
there, they can form large wrack deposits on sandy beaches when
detached from their benthic habitats. In Hawaii, accumulations of
the invasive Hypnea musciformis of >9000 kg week
reported for pocket beaches. In Argentine Patagonia, the biomass
and composition of macrophyte wrack on beaches have been
changed by the invasive kelp Undaria pinnatiﬁda (Piriz et al., 2003).
The kelp Sargassum mutica is accumulating in large quantities on
beaches of northwestern Spain (Lastra,personal observation), and
the green algae Caulerpa taxifolia can replace the native seagrass
Posidonia oceanica as the main component of wrack on Mediter-
ranean beaches (Boudouresque et al., 1995; de Villele and Verlaque,
1995). The preference of wrack consumers (e.g., talitrid amphipods)
for macroalgae suggests the potential for impacts of invasive algae
on food webs and nutrient cycling on beaches (Lastra et al., 2008).
Few invasive invertebrates have been reported from sandy
beaches to date, but this may reﬂect poor sampling coverage rather
than a lack of biological invasions. The semi-terrestrial talitrid
amphipod Orchestia cavimana, originally from freshwater habitats
in the Black Sea and Mediterranean region, has been reported from
fresh to brackish habitats in the British Isles, and more recently has
been spreading rapidly on open exposed shores with low salinities
of the NE Baltic Sea (Herkul et al., 2006). The talitrid Platorchestia
platensis, which presently inhabits Europe, the Canaries, Japan,
India, Hawaii and other Paciﬁc islands, and which was recently
reported in the Baltic, may be an exotic species (Spicer and Janas,
2006). A few widespread polychaetes (e.g., glycerids and spionids)
reported from exposed beaches, are classiﬁed as cryptogenic,
implying that their origin is unknown and that they are neither
demonstrably native nor introduced (Carlton, 1996). The best
known of these are Glycera americana and Scololepis squamata,both
of which are reported from beaches and estuaries of several
continents (McLachlan and Brown, 2006). The possibility that some
cryptogenic forms could represent cryptic species rather than
a single widespread invasive species needs to be considered.
Invasive terrestrial insects can also impact beach ecosystems.
Red ﬁre ants (Solenopsis invicta) from Brazil, are now found in the
USA, Australia, Taiwan and China; they can prey on the eggs and
hatchlings of loggerhead and green sea turtles (Allen et al., 2001). A
congener, the black ﬁre ant (Solenopsis richteri) preys intensely on
intertidal polychaetes in Argentinean lagoons (Palomo et al., 2003),
which suggests that it could also consume intertidal invertebrates
on exposed beaches. The Argentine ant (Linepithema humile,
formerly Iridomyrmex humilis), a small, dark ant native to northern
Argentina, Uruguay, Paraguay, and southern Brazil, has invaded
a variety of habitats, including beaches, in South Africa, New
Zealand, Japan, Australia, Europe and the USA.
3.7. Coastal development and engineering
The expanding use of coasts by human populations characteris-
tically involves a gradual intensiﬁcation of urban development in
the littoral-active zone, with the ultimate consequence that coupled
surf–beach–dune systems must be managed (Nordstrom, 2000).
This management most commonly focuses on the sediment budget.
Modern coastlines are increasingly starved of sand as dams trap
sediments that would otherwise feed beaches; the sediment budget
is further disrupted by activities such as quarrying, land reclamation,
urbanisation, afforestation and agricultural use (Nordstrom, 2000;
Sherman et al., 2002). As a result, most modern coastlines are
experiencing accelerating rates of erosion (Awosika et al., 1993;
Innocenti and Pranzini, 1993; Cooper and McKenna, 2008).
Society’s response to beach erosion and shoreline retreat relies
heavily on engineering interventions that place armouring struc-
tures on beaches (Nordstrom 2000, Charlier et al., 2005, Griggs,
2005a, b). Hard structures, such as walls constructed of stone,
concrete, wood, steel or geotextiles, have been used for centuries as
a coastal defence strategy (Charlier et al., 2005), but this protection
is not achieved without ecological costs. Armouring structures alter
the natural hydrodynamic system of waves and currents, thereby
affecting sand transport rates, which in turn control the erosion-
accretion dynamics of beaches (Miles et al., 2001; Hsu et al., 2007).
Intertidal seawalls and other structures that reﬂect wave energy
and constrain the natural landward migration of the shoreline have
unplanned environmental impacts, such as ﬂanking erosion of
shorelines adjacent to those protected by engineering structures
(Hall and Pilkey,1991; Weigel 2002a, b, c; Griggs, 2005a, b). Passive
erosion occurs on the armoured beaches themselves; and because
seawalls arrest the landward retreat of the shoreline in the face of
erosion, the beach seawards of the structure is frequently drowned
(Hall and Pilkey, 1991; Fletcher et al., 1997; Griggs, 2005b).
Armouring might also enhance beach erosion on protected coasts,
although this remains controversial (Kraus and McDougal, 1996;
The impacts of seawalls and other coastal armouring structures
may cause signiﬁcant habitat changes, with attendant ecological
impacts (Sobocinski, 2003; Martin et al., 2005; Dugan and Hubbard,
2006; Bertasi et al., 2007) that can be difﬁcult to detect in the short
term (Jaramillo et al., 2002). As eroding beaches become narrower
after armouring, the reduced habitat can directly lower the diver-
sity and abundance of biota, especially in the upper intertidal zone
(Sobocinski, 2003; Dugan and Hubbard, 2006; Dugan et al., 2008).
This, in turn, can also be detrimental to higher trophic levels, e.g.,
coastal avifauna may be impacted both by reduced habitat area and
by declining intertidal prey resources. This phenomenon is reﬂec-
ted by observations of signiﬁcantly lower numbers and fewer
species of birds on armoured compared with unarmoured
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segments of Californian beaches (Dugan and Hubbard, 2006;
Dugan et al., 2008). The loss of dry-sand following armouring also
eliminates nesting habitats for turtles and some specialised species
of ﬁshes (e.g. California grunion, surf smelt in the USA).
Armouring may also affect the quality of beach habitats. For
example, the rates of deposition and retention of macrophyte
wrack, driftwood and other natural debris that can be important to
beach biota as food and habitat, are lower on armoured than
unarmoured beaches (Sobocinski, 2003; Dugan and Hubbard,
2006). By contrast, increased deposition of wrack can occur on
some beaches where offshore structures have been placed (Martin
et al., 2005). These offshore defences also alter physical character-
istics and benthic communities of beaches and surf zones (Martin
et al., 2005; Bertasi et al., 2007).
Because shore-protection structures often present novel
habitats on beaches, they could promote the introduction of exotic
species. Together, engineering structures and introduced species
could severely impact native assemblages, not only by modifying
selection pressures as a result of habitat transformation, but also by
introducing competition (Gonzalez et al. 2008). However, this issue
remains speculative and more work is needed to assess its impor-
tance for sandy beach ecosystems.
Sea-level rise and other large-scale and long-lasting effects of
global warming (Fig. 1) are expected to intensify beach erosion
(Slott et al., 2006). To protect human assets on soft-sediment
coastlines globally in the face of escalating erosion, the use of
coastal armouring will therefore expand substantially. For the
extensive armoured coastlines in many countries (Fletcher et al.,
1997; Griggs,1998; Nordstrom, 2000), changes in sea level will shift
the locations of existing armouring structures to lower positions on
the shore, amplifying interactions with waves and tides and further
accelerating beach erosion (Cooper and McKenna, 2008). Thus, the
combined effects of rising sea levels and coastal armouring are
predicted to cause ecological impacts to beaches on unprecedented
Beach sands have been mined in many places, including Sri
Lanka for building lime from coral sand (Clark, 1996), Tanzania,
Korea, Argentina, Italy and many other countries for building sand
(Masalu, 2002; Cho, 2006; Pousa et al., 2007), the Namibian coast
for diamonds (McLachlan, 1996; Theron et al., 2003), and several
areas for heavy minerals such as titanium and zirconium (Mulaba-
Bafubiandi et al., 2002; Panigrahi, 2005; Pirkle et al., 2005; Ghosh
et al., 2006). Further, mine tailings are discharged onto beaches in
many places. Diamond mining activity in Namibia has been shown
to negatively impact shorebirds (Simmons, 2005), and tailings from
copper mines cause lower meiofauna density and diversity on
Chilean beaches (Lee and Correa, 2005). Removal of sand disturbs
sediment budgets, possibly contributing to enhanced erosion
(Masalu, 2002; Thornton et al., 2006); it may also alter particle size,
changing the morphodynamic state of beaches and intertidal
benthic communities (McLachlan 1996). Thus, the general effects of
any form of mining on beaches are to damage the beach and dune
habitat, to alter the sediment budget and hasten erosion. It is now
illegal in many countries to remove sand from the littoral active
zone and the practice should in general be prohibited (Clark, 1996).
In many places sand is mined offshore, often on the outer shelf; the
ecological impacts of this activity are probably slight for beaches,
but not well understood (Byrnes et al., 2004).
3.9. Climate change
Although the exact magnitude of physical changes resulting
from global climate change is still uncertain (IPCC, 2007), ecological
responses (e.g., changes in phenology, physiology, range and
distribution, assemblage composition, species interactions) are
increasingly apparent in sandy beaches (Brown and McLachlan,
2002; Jones et al., 2007a). However, there are no direct studies of
effects of this long-lasting and large-scale stressor (Fig. 1) on beach
ecosystems; consequently, many of our predictions of the likely
impacts on beach ecosystems are derived from other systems.
Rising temperatures may have different implications at different
latitudes and for taxa with different dispersive abilities and ranges.
Because many sandy beach species (e.g. peracarid crustaceans) lack
dispersive larval stages, their rates of range extension might be
outpaced by changes in temperature, making these taxa particu-
larly vulnerable to climate-change effects. Narrow-range endemic
species would be at greatest risk (O’Hara, 2002) and some species
could eventually be replaced by species from lower latitudes. But
even migratory species, as well as those that have pelagic larval
stages, may be impacted directly by large-scale, climate-driven
changes to prevailing oceanographic systems.
Beach biota may also respond to indirect effects of temperature
change on beaches. For example, comparatively small increases
C) in temperature have been associated with major changes
in planktonic ecosystems in the North Atlantic over the past ﬁve
decades (Richardson and Schoeman, 2004). Given that plankton is
a key food source for suspension-feeding beach species, changes in
plankton communities will have unpredictable impacts on sandy
beach macrofauna. Changes may be signiﬁcant for semi-terrestrial
species (peracarids and insects), which will be affected both
directly and indirectly by changes in water and air temperatures.
A direct consequence of warming seas is sea-level rise. Current
estimates suggest that sea level is rising by an average of
1.7 0.5 mm yr
(IPCC, 2007), but there is potential for far greater
rates of change (Rahmstorf et al., 2007)
. Irrespective, rising sea level
pushes the high-water mark landward, causing beaches to migrate
slowly inland. Low-gradient dissipative shores, which house the
greatest biodiversity, are at most risk due to their erosive nature
and the much greater run-up of swashes on gentle gradients.
Moreover, warmer air and sea temperatures would translate into
more frequent and more severe storms (IPCC, 2007), thus escalating
beach erosion to the point where entire beaches could disappear,
removing habitat for the biota. The institution of setbacks is
currently international best practice. This involvesdeﬁning a shore-
normal zone within which new development is prohibited and
space is thereby provided for natural coast retreat. Complementary
to this is managed coastal retreat, which involves preparing to
abandon existing infrastructure within the setback zone and/or
developing plans to remove these existing structures that are too
close to the shore. Soft solutions, such as nourishment, should be
sought ﬁrst, with hard engineering the very last resort.
Predicted changes in ocean acidiﬁcation could further impact
sandy beaches. The continued reduction of pH in surface waters
(0.1 unit less than pre-industrial levels and predicted to decline by
a further 0.3–0.4 units by the end of the century) will reduce
calciﬁcation rates and calcium metabolism in marine organisms
(Feely et al., 2004), including several sandy beach molluscs and
crustaceans (Hall-Spencer et al., 2008).
Management to maintain beach habitat requires long-term
mitigation and/or adaptation strategies. Mitigation would arrest
climate change and its consequences via large reductions in
greenhouse emissions. Adaptation would provide setback zones,
allowing the beach to migrate inland as the sea rises. The latter
would have minimal ecological consequences for beaches, but
would be very expensive in urban areas. Alternatively, if engi-
neering solutions (e.g. seawalls) are used to defend societal assets,
intertidal sand habitat will be lost (Finkl and Walker, 2004).
Adaptive measures accept the reality of sea-level rise and coastline
retreat and seek to increase coastal resilience, a concept with
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ecological, morphological and socio-economic components
(Carpenter and Folke, 2006). Unfortunately, coastal resilience has
been impaired by anthropogenic effects on the sediment budgets of
beach-dune systems (Alonso and Cabrera, 2002; Sherman et al.,
2002; Tomlinson, 2002). Measures to promote resilience include
the protection, vegetation and stabilisation of dunes, the mainte-
nance of sediment supply and the provision of buffer zones, rolling
easements or setbacks that allow the landward migration of the
4. Discussion and conclusions
The major long-term threat facing sandy beaches worldwide is
coastal squeeze, which leaves beaches trapped between erosion
and rising sea level on the wet side and encroaching development
from expanding human populations on land, thus leaving no space
for normal sediment dynamics. Because of the inertia associated
with global climate change and human population growth, no
realistic management scenario will alleviate this threat in the short
term. Instead, management must be based on existing knowledge,
suggesting that the immediate priority is to avoid further devel-
opment of coastal areas likely to be directly impacted by retreating
shorelines. Under climate change, this zone must be shifted inland
to create substantial setbacks that can account for the predicted
inland movement of coastlines over the next decades. For example,
pending legislation in KawZulu-Natal (South Africa) formalises
setback lines at the 10-m elevation contour, which should account
not only for several decades of sea-level rise, but also for rates of
erosion that are disproportionately faster than sea-level rise.
Similar legislation exists elsewhere, and setbacks are widely
recognised as critical (Healy, 2002; Komar et al., 2002; Daniel and
Abkowitz, 2005; Jackson, 2005).
In the longer term, robust and efﬁcacious management inter-
ventions must be founded in a more comprehensive ecological
understanding of beaches. In this context, the spatial and temporal
scales used to evaluate impacts on sandy beaches are particularly
relevant. The scales of impacts to beach ecosystems (Fig. 1)
encompass multiple orders of magnitude in both time and space,
and include pulse and press drivers. Press disturbances (whatever
the source of impact) are becoming increasingly common; they
operate on time scales of years to decades and they may persist at
least as long as the impacts remain. Our summary of the estimated
spatio-temporal scales of human impacts on sandy beaches indi-
cates that many of the impacts manifest at the scale of tens to
hundreds of kilometres and last for months to years (Fig. 1). The
cumulative effects of these impacts may already extend to regional
and global scales for developed and urbanised coasts, as well as
coasts used intensively for recreation (Cooper and McKenna, 2008;
Halpern et al., 2008). However, long-term data sets describing
either the natural dynamics of beach systems or the magnitude of
human impacts on beaches are scarce and fragmentary. Thus, the
necessary research priority is to implement long-term monitoring
programmes that quantify the dynamics of key ecological attributes
on sandy beaches at different organisational levels ranging from
individuals to ecosystems, as well as to compile regional and global
databases of empirical measurements of ecosystem condition. This
should be useful to detect regime shifts and their corresponding
drivers, which have been increasingly documented for some
terrestrial, freshwater and marine systems (de Young et al., 2008),
but not for sandy beaches yet.
As the number and magnitude of global change drivers
increases over time, scaling up the ﬁndings of sandy beach research
to larger and longer scales is a short-term need. To this end, multi-
year, large-scale data sets should be built in order to assess the
extent of range contractions or expansions and even the potential
extinction of some species. Given that critical stressors faced by
sandy beach ecosystems, such as rising sea level and intensiﬁcation
of onshore winds, can overwhelm local patterns and processes, the
interplay between ﬁne-scale (e.g., erosion and local variations in
abundance and other population descriptors) and broad-scale
(e.g., macroscale variations in temperature and variations in
distribution ranges) phenomena (i.e., a cross-scale interaction
approach sensu Peters et al., 2008), is needed to provide a better
understanding of the dynamics of sandy beach ecosystems. This
should be operationalised through the creation of networks of
long-term observation sites that should be interdisciplinary in
nature (Carpenter, 2008 and references therein). Cooperative
interdisciplinary investigation is thus a pressing need to assess the
effects of human impacts on sandy beach ecosystems.
There is also scope for improvement in experimental design to
better distinguish natural variability from anthropogenic impacts
(Peterson and Bishop, 2005). It can be extremely difﬁcult to sepa-
rate the individual effects of different impact sources (e.g., human
trampling vs. seawalls replacing foredunes), or to ﬁnd unaffected
beaches that could act as truly independent controls in experi-
ments. In these situations, ﬁeld experiments could be improved by
interspersing sites across areas of beach with as full a spectrum of
impacts as is possible. Because good control sites are very difﬁcult
to ﬁnd (however, see Peterson et al., 2006 and Schlacher et al.,
2007b for illustrative examples on sandy beaches), distinguishing
natural variability from anthropogenic impacts can only be
attempted by modelling gradients away from the impact sources, as
shown by long-term (Defeo and de A
´lava, 1995) and large-scale
(Lercari and Defeo, 1999, 2003) ﬁeld studies. Therefore, the
unambiguous deﬁnition of the relevant scales of analysis, which
will vary according to the nature and extent of the impact (Fig. 1)
and the organisational level under analysis, will be critical in
developing a robust hypothesis testing framework.
Exposed sandy beaches have fewer invasive species than other
coastal habitats (e.g. Wasson et al., 2005). This raises the questions
of whether: (1) exposed beaches are more difﬁcult to invade and
colonise; and (2) the high frequency and intensity of disturbance on
ocean beaches impede the establishment and survival of invasive
species. In this way, sandy beaches may serve as a useful model
system in which to test general ecological theories about the rela-
tionship between physical disturbance regimes and invasion
success by alien species.
Because it is difﬁcult to completely exclude human activities
from sensitive coastal areas, zoning strategies will be important in
measuring and managing spatial gradients of human impact. In this
sense, areas of relatively natural, unaffected beach could be inter-
spersed with areas where recreational activities are permitted.
Here, the application of strategic management (Micaleff and Wil-
liams, 2002), systematic conservation planning (SCP, Margules and
Pressey, 2000) and marine spatial planning (MSP, Ehler, 2008)
approaches could assist beach managers in determining which
spatially explicit elements should receive conservation priority. SCP
and MSP could be used to determine the optimal along-shore
distribution of human activities, on the basis of knowledge about
the current distribution of ecological values and human threats,
coupled with an understanding of how these might change through
time and which tools are available to manage them.
Marine reserves (MRs) and marine protected areas (MPAs) have
not been widely implemented in sandy beaches. MRs could be a key
tool for biodiversity conservation and should also facilitate spill-
over effects into adjacent beach habitats. Given that most inverte-
brates of beaches are short-lived, the efﬁcacy of MRs in conserving
sandy beach ecosystems may be detectable quite rapidly. Design
and allocation of MR networks should therefore become a top
conservation priority for sandy beach ecosystems globally. Coastal
networks combining MRs, MPAs with limited level of human
activities, co-management and exploitation areas sustaining
O. Defeo et al. / Estuarine, Coastal and Shelf Science xxx (2008) 1–128
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exclusive community rights, could represent a short-to mid-term
management objective that balances bio-socio-economic factors
(Defeo and Castilla, 2005). Thus, management policies for beaches
should integrate the interacting natural, socio-cultural and
management systems to protect biological diversity and maintain
essential ecological processes and life-support systems (James,
2000; Ariza et al., 2008).
A common feature of setback and zoning strategies is that they
need to be enforced through legislation. Thus, all relevant stake-
holders should be included from the beginning in the design,
implementation and institutionalisation of spatial planning initia-
tives (Plasman, 2008). New perspectives for rational management
of sandy beaches require paradigm shifts, including not only basic
ecosystem principles, but also incentives for effective governance
and sharing of management roles between government and local
stakeholders (Castilla and Defeo, 2005). Legitimising the partici-
pation of stakeholders in the planning and surveillance of
management measures is a promising short-term solution to
current problems faced in sandy beach ecosystems, promoting
compliance with regulations.
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