Download Climate change impacts on connectivity in the ocean: Implications

CONCEPTS & THEORY
Climate change impacts on connectivity in the ocean:
Implications for conservation
LEAH R. GERBER,1, MARIA DEL MAR MANCHA-CISNEROS,1 MARY I. O’CONNOR,2
AND
ELIZABETH R. SELIG3
1
Ecology, Evolution and Environmental Science, School of Life Sciences, Arizona State University, Box 874501,
Tempe, Arizona 85287-4501 USA
2
Department of Zoology and Biodiversity Research Centre, University of British Columbia, Vancouver,
British Columbia V6T 1Z4 Canada
3
Betty and Gordon Moore Center for Science and Oceans, Conservation International, 2011 Crystal Drive, Suite 500,
Arlington, Virginia 22202 USA
Citation: Gerber, L. R., M. Del Mar Mancha-Cisneros, M. I. O’Connor, and E. R. Selig. 2014. Climate change impacts on
connectivity in the ocean: Implications for conservation. Ecosphere 5(3):33. http://dx.doi.org/10.1890/ES13-00336.1
Abstract. Effective spatial management in the ocean requires a network of conservation areas that are
connected by larval and adult dispersal. We propose a conceptual framework for including the likely
impacts of a changing climate on marine connectivity, and synthesize information on the relationships
between changing ocean temperature and acidification, connectivity and conservation tools. Our
framework relies on concepts of functional connectivity, which depends on an organism’s biological and
behavioral responses to the physical environment, and structural connectivity, which describes changes in
the physical and spatial structure of the environment that affect connectivity and movement. Our review
confirms that ocean climate change likely reduces potential dispersal distance and therefore functional
connectivity. Structural connectivity in the ocean will inevitably change with the spatial arrangement of
biogenic habitats resulting from disturbance as well as enhanced growth and mortality due to climate
change. Climate change will also likely reduce the spatial scale of connectivity, suggesting that we will
need more closely spaced protected areas.
Key words: climate change; connectivity; dispersal; marine reserves; marine spatial planning; ocean acidification;
warming.
Received 21 October 2013; revised 20 January 2014; accepted 22 January 2014; final version received 21 February 2014;
published 24 March 2014. Corresponding Editor: D. P. C. Peters.
Copyright: Ó 2014 Gerber et al. This is an open-access article distributed under the terms of the Creative Commons
Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the
original author and source are credited. http://creativecommons.org/licenses/by/3.0/
E-mail: [email protected]
INTRODUCTION
(IPCC 2007). Most areas of the ocean have
warmed, albeit with considerable spatial variation in the magnitude of change (Hansen et al.
2006, Rayner et al. 2006, Hoegh-Guldberg et al.
2007). Increases in ocean temperature are accompanied by changes in ocean chemistry—specifically ocean acidification (OA) resulting from
increasing dissolution of atmospheric CO2 in
the sea (Orr et al. 2005, Hoegh-Guldberg et al.
2007, Guinotte and Fabry 2008). Climate change
The rate of anthropogenic carbon dioxide
(CO2) emissions has increased from ;280 ppm
prior to the industrial revolution to more than
400 ppm in 2013 and is predicted to increase to
;700 ppm in 2100 (Sabine et al. 2004). Such
levels of atmospheric CO2 influence global sea
surface temperatures, which have already risen
by 0.768C, on average, since pre-industrial times
v www.esajournals.org
1
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
has also been implicated as a primary cause of
sea level rise, increased frequency and intensity
of storms, and changing ocean circulation patterns (Meehl et al. 2005, IPCC 2007). In coming
decades, the pace of these changes is expected to
accelerate, which will alter the processes underlying connectivity in marine populations. However, the implications of ocean change for marine
conservation planning are just beginning to be
explored (Gaines et al. 2010, McLeod et al. 2012,
Rau et al. 2012). In this paper, we propose a
general framework to understand the impacts of
warming and acidification on the connectivity of
marine organisms.
Over the past decade, the number of studies on
the effects of climate change on connectivity of
marine organisms has increased dramatically,
incorporating the effects of numerous physical
and biological climatic stressors as well as
multiple experimental, statistical, and modeling
approaches. Most studies have focused on
impacts on larval or juvenile stages of marine
animal populations (e.g., 81% of 110 studies
reviewed by Wernberg et al. [2012]), with benthic
invertebrates being the most studied groups.
Changes in ocean temperatures may influence
biological processes such as dispersal and survival (Munday et al. 2009c, Johnson and Welch
2010). For example, O’Connor et al. (2007)
demonstrated a relationship between temperature and a reduction in pelagic larval duration
(PLD) across 69 marine species. In addition,
ocean acidification has strong, negative effects on
calcified organisms (e.g., calcified algae, corals,
mollusks, and larval stages of echinoderms)
while other taxa (e.g., crustaceans, and fleshy
algae, seagrasses, and diatoms) are more resilient
to ocean acidification (Kroeker et al. 2010,
Kroeker et al. 2013).
Much of this work documents physiological
effects at the individual level, and focuses less on
ecological impacts at the population or community levels. Furthermore, 65% of experimental
studies conducted between 2000 and 2009 involve single-factor manipulations of either
warming or acidification (Wernberg et al. 2012),
with most temperature studies showing significant negative effects of warming, but acidification responses exhibiting a great degree of
variability (Kroeker et al. 2010, Wernberg et al.
2012, Kroeker et al. 2013). There have been fewer
v www.esajournals.org
studies examining the effects of two or more
climate change stressors at once (e.g., Wernberg
et al.’s [2012] review in which 30% of the studies
looked at two stressors and only ,5% of studies
looked at more than two stressors), and these
show even higher variation in responses. Increased ocean temperatures have been shown to
both increase and decrease the severity of ocean
acidification, and the potential interaction effects
are likely significant (Gooding et al. 2009).
In this paper, we develop a conceptual
framework based on the concepts of structural
and functional connectivity and use this framework to evaluate cumulative impacts of warming
and acidification on marine connectivity (Kindlmann and Burel 2008). We use this framework to
synthesize existing data on the effects of temperature and ocean acidification stress on marine
organisms. We then discuss how our findings
relate to developing effective marine conservation strategies to mitigate climate change impacts
on marine ecosystems.
A FRAMEWORK FOR UNDERSTANDING IMPACTS
OF CLIMATE CHANGE ON CONNECTIVITY
While there are many physical symptoms of
climate change on ocean ecosystems (e.g., sea
level rise, altered storm regimes [Meehl et al.
2005, Macreadie et al. 2011]), we focus specifically on changing temperature and pH for our
review due to the increased uncertainty in their
predictions and the lack of literature on their
impacts. We relate empirical estimates of effect
size of temperature and OA to functional connectivity, which describes the suite of biological or
behavioral responses of individuals to the physical environment, and structural connectivity,
which describes the physical and spatial structure of the environment (Kindlmann and Burel
2008, Turgeon et al. 2010). We use these concepts
to characterize our theoretical and empirical
understanding of the effects of climate on the
biological and spatial processes underlying connectivity. This framework unites two very different effects of climate change: direct effects on the
biology of dispersing individuals (functional
connectivity), and emergent changes to the
spatial structure of the landscape that influence
the probability of population persistence (structural connectivity). By integrating individual and
2
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Fig. 1. Climate change impacts on marine connectivity with implications for Marine Spatial Planning (MSP); (a)
Climate change could reduce functional connectivity by altering the duration of, trajectory and survival during the
larval life cycle. Bold arrows represent organism dispersal; (b) Climate change could reduce structural connectivity
by reducing habitat (dotted arrows). Grey patches represent pre-climate change habitat extent and blue patches
represent habitat after climate-related change.
population-level responses to climate change, our
framework can be populated with empirical data
that are relevant to conservation objectives, such
as probability of population persistence (Fig. 1).
and survival probability while dispersing), and
the aggregate of individual activity at the
population level. Ocean conditions influence the
timing and location of cues for reproduction and
settlement (Munday et al. 2009b), potential
dispersal distances and survival (O’Connor et
al. 2007, Munday et al. 2009c), and feeding and
predation opportunities during dispersal periods. Here we summarize empirical examples and
research approaches to studying the effects of
ocean temperature and acidification on functional connectivity.
FUNCTIONAL CONNECTIVITY: LINKING
ENVIRONMENTAL CONDITIONS WITH
INDIVIDUAL MOVEMENT
Functional connectivity may be used to describe an organism’s biological and behavioral
responses to individual elements in the physical
environment as well as the spatial configuration
of the entire landscape (Kindlmann and Burel
2008). This includes individual activity and
performance (e.g., time available for dispersal
v www.esajournals.org
Empirical examples
In marine systems, over 70% of species have a
planktonic larval stage that is subject to dispersal
3
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
(Lockwood et al. 2002, Levin 2006, Munday et al.
2009c). Furthermore, most dispersal and mortality occurs during this pelagic development stage
(Morgan 1995, Levin 2006). The eggs, larvae or
spores of benthic adult fish, invertebrates or
seaweeds may travel tens or hundreds of
kilometers before settling and transitioning to
juvenile life stages (Kinlan and Gaines 2003,
Siegel et al. 2003), yet mortality during the
pelagic stage is also extremely high (Morgan
1995). Early life stages in the organisms’ complex
life cycle thus can be critical for the maintenance
of functional connectivity in marine ecosystems.
Studying individual larval activity is notoriously
challenging because direct observations of dispersal (e.g., movement of larvae) for most species
are rare (but see Mora and Sale 2002, Planes et al.
2009). However, there has been a recent increase
in research on larval retention (self-recruitment)
within and dispersal (connectivity) among coral
reef populations, as well as a wide range of
techniques and approaches for their study, from
larval tagging and paternity analyses at small
scales, to biophysical models at larger scales
(Jones et al. 2009). Table 1 summarizes a few
examples of approaches with the potential to
examine the effects of climate change on life
history and functional connectivity.
Connectivity can also be mediated by behavioral traits that determine settlement time and
location, and these traits can be affected by
changing ocean conditions. The effect of ocean
acidification on larval behavior and their sensory
systems has been studied experimentally most
extensively on tropical coral reef fish fishes. High
CO2 levels may have significant effects on the
habitat selection process involved with attraction
to olfactory cues, which can bring about changes
in population connectivity patterns (Munday et
al. 2009b). At near-future levels of CO2, orange
clownfish (Amphiprion percula) tended to avoid
positive settlement cues that would normally
lead to the location of suitable adult settlement
sites. Instead, larvae were attracted to negative
stimuli that led to suboptimal locations for
settlement or even locations devoid of settlement
habitat, which they normally avoided in today’s
CO2 conditions (Munday et al. 2009b). Exposure
to elevated CO2 has also been shown to affect the
timing of settlement in damselfishes (Pomacentrus
spp.) with different patterns of habitat use,
v www.esajournals.org
potentially leading to detrimental effects on
larval survival and population replenishment if
settlement occurs at unfavorable times (Devine et
al. 2012a). Acidification has also been shown to
significantly affect adult fish homing behavior.
Devine et al. (2012b) tested the effects of nearfuture levels of CO2 on the ability of adult
cardinal fish (Cheilodipterus quinquelineatus) to
home to their diurnal resting site after nocturnal
feeding, and they detected an impaired ability to
distinguish between home and foreign sites as
well as increased activity level. These studies
provide evidence to suggest that ocean acidification has an effect on fish brain functioning and
fish behaviors (larval and adult) that may
negatively impact population connectivity patterns.
Statistical approaches
In addition to empirical approaches quantifying key aspects of functional connectivity, efforts
to characterize and model the probability of
dispersal can also be used to study potential
changes in the scale of functional connectivity
(Box 1). Dispersal kernel functions represent the
probability of dispersing a particular distance
from a source location over repeated dispersal
events (Lockwood et al. 2002), and capture
average patterns in pre-settlement processes,
but not post settlement mortality and selection
(Almany et al. 2007, Graham et al. 2008). Thus for
many species, dispersal kernels relate the spatial
scales of dispersal to those of demographic
population connectivity. Dispersal kernels can
also be used to relate distributions of dispersal
distances to the larval development period (Fig.
2). The average dispersal distance from one
population to another can determine the scale
of demographic processes, such as recolonization
after disturbance. In contrast, the long tails of a
dispersal distribution may be the most important
determinant of migration rate and genetic connectivity (Clark 1998, Neubert and Caswell 2000,
Palumbi 2004). Generalized dispersal kernels
have played a central role in the development
of marine reserve theory (Botsford 2001, Lockwood et al. 2002, Botsford et al. 2003, Gerber et
al. 2003) and are potentially a useful tool for
relating the effects of climate change to population-level responses (Munday et al. 2009a, Munday et al. 2009c). Dispersal kernels can also be
4
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Table 1. Review on the current state of knowledge of the effects of temperature and ocean acidification stress on
marine organisms with respect to functional and structural connectivity.
Climate change
impact
Functional
Ocean
acidification
(high CO2
conditions)
Increased ocean
temperatures
Approach
Taxa
Findings
Experimental
manipulation of CO2
levels þ settlement
cues
Experimental
manipulation of CO2
levels
Orange clownfish
(larvae)
Induced attraction to suboptimal or non-existent
settlement habitat
Munday et al. (2009c)
Damselfishes
(larvae)
Unfavorable timing of
settlement
Devine et al. (2012a)
Cardinalfish
(adult)
Altered homing behavior:
impaired distinction between
home and foreign sites
Significant changes in
spawning patterns in the
southern Benguela Upwelling
system.
Reduced PLD has large impact
on dispersal at large scales,
and spawning location and
temporal variability explains
connectivity patterns in the
Bay of Biscay.
Devine et al. (2012b)
Evolutionary physical
model
Small pelagic fish
(larvae)
Coupled physicalbiological model
(IBM) using
Lagrangian particle
tracking models and
dispersal kernel
descriptors
European
anchovy
(larvae)
Modeling framework
explicitly simulating
larval behavior
Coupled 3-D
hydrodynamic
model-IBM
(Lagrangian) and
dispersal kernels
Coupled oceanographic
model-IBM
Weakly
swimming fish
larvae
Generic benthic
invertebrate
(larvae)
Striped trumpeter
(larvae)
Shifts in major
ocean
currents
Passive particle
dispersion modeling
Generic larvae
Reduced (THC)
Coupled
hydrodynamic
model-IBM
Arcto-Norwegian
cod
Region-wide analysis
of changes in reef
architectural
complexity
Coral reefs
Regional analysis of
SST anomaly and
coral bleaching
reports
Coral reefs
Structural
Increased ocean
temperatures
v www.esajournals.org
Hydrodynamic spatio-temporal
variability causing stronger
stratification can concentrate
particles in the surface where
they can be affected by THC
changes.
Shortened PLD increased selfrecruitment rate in coastal
environments.
Decreased PLD induced selfretention and increased local
connectivity while decreasing
dispersal distance and thus
regional connectivity.
Positive minor effect on
survival by 2050, but
decreased performance and
dispersal outside thermal
niche by 2100, increasing
mortality in Tasmania,
Australia.
Southerly shifts in the South
Equatorial Current caused
particles to shift advection
patterns.
Larval and juvenile transport
patterns shift in response to a
slowed-down THC in the NE
Atlantic.
The architectural complexity of
coral reefs has declined
drastically in the last 40 years
due in part to climateinduced bleaching events,
affecting both coral
abundance and community
composition in the Caribbean
region.
Positive correlation between the
spatial extent and coral
bleaching intensity with SST
anomalies in the Caribbean
over the last two decades.
5
Reference
Lett et al. (2009)
Huret et al. (2010)
Lett et al. (2010)
Irisson (2008)
Ayata et al. (2010)
Tracey et al. (2012)
Munday et al. (2009c)
Vikebø et al. (2007)
Alvarez-Filip et al.
(2009)
McWilliams et al.
(2005)
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Table 1. Continued.
Climate change
impact
Ocean
acidification
(high CO2
conditions)
Approach
Taxa
Findings
Empirical fish
abundance estimates
coupled with on-site
temperature data
over 6 years and
long-term SST
analysis
Juvenile coral reef
fish
Meta-analyses of
experimental
evidence
Coral reefs
Natural in situ
investigations
Coral and
seagrass
communities in
cool volcanic
seeps
Experimental
manipulations of
CO2 conditions
Coral and
seaweed
communities
The increasing temperate
incursion of the East
Australian Current (EAC) is
shown to allow increased
connectivity with the Great
Barrier Reef. Present
warming trends are shown to
steadily increase the
frequency of warm,
survivable winters, thus
allowing potential range
expansions.
Coral calcification expected to
decline by ;15–22% on
average by the end of the
century.
Reduced coral diversity,
recruitment, abundances, and
shifts in competitive
interactions. Increase in
seagrass cover and noncalcareous macroalgae and
reduction in crustose
coralline algal cover.
Increased coral mortality
through enhanced seaweed
ability to kill and potentially
outcompete corals in contact
with seaweed communities.
Suppressed larval oxygen
consumption and reduced
metamorphosis rate.
Reduced coral settlement and
crustose coralline algal cover.
Increased sensitivity to OA
under warm conditions.
Coral reefs
(larvae)
Coral reefs
Both increased
ocean
acidification
and
temperature
Experimental
manipulation of CO2
and temperature
conditions
Crustose coraline
algae
Algal turfs and
kelp beds
In non-calcifying organisms.
Turfs had positive enhanced
abundance and were shown
to inhibit kelp recruitment.
Reference
Figueira and Booth
(2010)
Chan and Connolly
(2013)
Fabricius et al. (2011)
Diaz-Pulido et al.
(2011)
Nakamura et al.
(2011)
Doropoulos et al.
(2012)
Doropoulos et al.
(2012)
Connell and Russell
(2010)
Note: Abbreviations are: THC ¼ thermohaline circulation; IBM ¼ individual based model; PLD ¼ pelagic larval duration; OA ¼
ocean acidification.
applied to coupled circulation and population
models (Neubert and Caswell 2000).
mortality, and settlement processes of larvae, and
are particularly relevant to questions related to
the effects of climate change on connectivity
because connectivity patterns are usually determined by both physical and biological processes
(Werner et al. 2007). A small number of these
coupled models have been explicitly used to look
at the effects of climate change on connectivity
(Lett et al. 2010). For example, by using a
biophysical modeling approach and dispersal
kernels to describe spatial distributions of larvae,
Huret et al. (2010) found that variability in PLD
can have a major impact on dispersal at large
Analytical approaches: biophysical modeling
Biophysical modeling approaches are increasingly being used to understand the dispersal of
single-species pelagic larval stages by coupling
biological properties of the organism with
realistic hydrodynamic conditions (Miller 2007,
Werner et al. 2007, Treml et al. 2008, Cowen and
Sponaugle 2009, Lett et al. 2009, Metaxas and
Saunders 2009) (Box 1, Table 2). Such biophysical
models typically represent transport, behavior,
v www.esajournals.org
6
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Box 1
The Gulf of California as an Illustrative Case Study
To examine how temperature and pH change may influence conservation planning, we model
directional changes in functional connectivity with climate change by linking an abiotic change
(e.g. temperature, pH) to the rate of a biological process (larval development and related
dispersal). We examine the implications of a change in mean dispersal distance resulting from
climate change for spatial prioritization of conservation actions in the Gulf of California (GoC).
We focus on the GoC because (1) data on larval spawning locations are available (Soria et al.
2010), (2) a circulation model is available (Marinone et al. 2008), and (3) a spatial prioritization
effort has been conducted to identify conservation sites (Ulloa et al. 2006), but this effort did not
account for connectivity or climate change. Here we explore whether our framework suggests
that future adaptation of the prioritization scheme might be necessary.
We focus on how climate change may impact dispersal and connectivity of a key species in the
Northern Gulf, the rock scallop Spondylus calcifer. Rock scallops are highly sought after in
artisanal fisheries, and have been recently protected by a network of marine reserves (CudneyBueno et al. 2009). Planktonic larvae of S. calcifer drift with ocean currents for several weeks
following spawning and fertilization (Soria et al. 2010). Using our understanding of the
potential magnitude and spatial variability in changes in connectivity and abundance resulting
from potential changes in pH-value and temperature, we map potential changes in functional
connectivity that could result from local climate change using a published circulation model
(Marinone et al. 2008). We do not consider changes in structural connectivity because we have
no reason to forecast change to habitat for rock scallop. In the absence of a downscaled general
circulation model (GCM) for this region, we apply general climate change projections of 38C
warming and 0.4 pH units as a simple illustration of our approach (Soria et al. 2010). We
consider these changes in the context of the current configuration of priority areas identified in
the Gulf of California Ecoregional Assessment (Ulloa et al. 2006), specifically focusing on the
Rocas Consag priority area (Fig. 3).
Our results suggest that the current spatial extent of potential settling larvae is larger than the
range for future settling with ocean warming and acidification (Fig. 3). In the northern areas,
current settlement extends further south and in the eastern areas, settlement stretches further
west than in a þ38C warming scenario (Fig. 3). In addition, there is greater larval density within
a smaller area in the future temperature scenario. How the impacts of climate change on
functional connectivity affect the potential success of conservation areas depends on whether a
single priority area or a network is considered (Fig. 3). Rocas Consag (Graham et al. 2008) is the
biggest priority area in the Northern Gulf. Our climate change scenarios dramatically affect
scallop settlement in Rocas Consag, reducing it by nearly half. However, when the entire
priority area network in the Northern Gulf is considered, the impact of warming on settlement
locations is negligible (Fig. 3, Table 2). This suggests that the greater coverage and spatial
distribution of priority areas in the network would be more resilient to changes in mean
dispersal distance of rock scallop.
abilities by shortening the development time,
thus increasing local retention and self-recruitment rate (Irisson 2008).
The breadth of modeling approaches that have
been developed in the past 20 years represents a
significant increase in our understanding of the
scales and vertical behavior at smaller scales,
with a reduction in PLD having the potential to
reduce the dispersal distance. Other approaches
have explicitly simulated larval behavior and
focused on the effect of increased water temperatures accelerating development of swimming
v www.esajournals.org
7
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Fig. 2. Basic properties of dispersal kernels (A) vary predictably with the planktonic larval duration (PLD) and
mortality (Siegel et al. 2003). (B) Mean absolute displacement of larvae relative to parental source increases with
PLD, (C) variance in dispersal distances increases nonlinearly with PLD, (D) survival at the modal dispersal
distance declines with PLD and (E) daily mortality rate.
in oceanic currents as an additional variable (but
see Munday et al. 2009c). Nonetheless, the
modeling studies described in this section (see
Table 1) have proven useful tools to test
hypotheses about how climate change may
impact connectivity patterns (Lett et al. 2010).
interaction between the characteristics of the
physical environment (e.g., temperature and
circulation) and changes in organismal life
history traits that are ultimately responsible for
survival and dispersal (e.g., PLD). However, only
a few studies have explicitly considered changes
Table 2. Proportion of larval trajectoriesthat settle in conservation priority areas. Results from analysis of effects of
temperature and pH on potential larval dispersal distance in the Gulf of California (Fig. 3).
Rocas Consag
Network
Scenario
Number
Percentage of total
Number
Percentage of total
Current
þ38C
þ38C, 0.4 pH
465
270
273
4.0
2.3
2.3
4061
4054
3997
35
34.9
34.5
v www.esajournals.org
8
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Furthermore, an increasing number of approaches solely focusing on how modeling population
connectivity can be modified by ocean currents
(i.e., under current conditions) are becoming
available for specific regions (Marinone et al.
2008, Treml et al. 2008). Despite their widespread
application, however, using such approaches
remains difficult since assessing climate change
impacts still requires extensive computing power
and sophisticated models (e.g., coupled climate
ocean general circulation models). The capacity
to run them at the appropriate spatial resolution
and in conjunction with relevant biological
processes is still in its infancy (Munday et al.
2009c). The success of most of these models has
strongly relied on approaches that focus at small,
very local scales, which makes the use of these
models heavily site-specific.
substantially affect total settlement. Variation in
responses could also reflect species-specific sensitivities and local adaptation to different environmental conditions (Kroeker et al. 2013).
Nonetheless, the apparent trend of decreased
mean dispersal distance and reduced survival
probability remains crucial and useful when
considering the framework proposed here (Fig.
1a). Variability may be incorporated into estimates of functional connectivity for particular
cases using dispersal kernels.
Despite the fact that approaches quantifying
particular trajectories for the study of functional
connectivity are still under development, we can
nonetheless continue to study changes to the
dispersal process with the use of dispersal
kernels (Fig. 2). Even without specifically quantifying dispersal trajectories, we can consider
how changes to the ocean influence functional
connectivity by relating properties of the ocean
environment to outcomes of individual movement such as dispersal distance and survival.
Synthesis of functional connectivity framework
As seen from the few examples mentioned
here, increasing evidence suggests direct effects
of changes in ocean climate on individual
dispersal biology. Increasing ocean temperatures
are likely to accelerate larval and egg development, leading to reduced time in the water
column (Hirst and López-Urrutia 2006, O’Connor et al. 2007) allowing earlier habitat-seeking
behavior (Munday et al. 2009c), changes in
dispersal distance (Lett et al. 2009, Ayata et al.
2010, Huret et al. 2010, Lett et al. 2010), changes
in transport processes and thus settlement
location (Vikebø et al. 2007, Munday et al.
2009c, Huret et al. 2010, Tracey et al. 2012), and
modifications of local versus regional connectivity (Irisson 2008, Ayata et al. 2010). Additionally,
even though there remains considerable variability in responses among taxa, the combined effect
of increased ocean temperatures and acidification
shows a trend towards lower survival, growth,
and development (Findlay et al. 2010, Lischka et
al. 2011, Kroeker et al. 2013). These relationships
between temperature, acidification and mortality
could be mediated by direct physiological impacts, or indirectly through changes in size at
age, morphology or feeding efficiency (Kurihara
2008, Dupont and Thorndyke 2009, Kroeker et al.
2010, Kroeker et al. 2013). This combination of
stressors can also potentially impact other biological sources of mortality, such as predation
and starvation, and changes to mortality could
v www.esajournals.org
Structural connectivity: relating
environmental conditions to habitat area
Structural connectivity reflects the spatial
structure of the environment and the influence
of spatial structure on the likelihood of individuals from one habitat reaching another habitat
(Kindlmann and Burel 2008). In contrast to
functional connectivity, which can be used to
describe the effect of ocean conditions, such as
temperature, on the dispersal potential of individuals via physiological mechanisms, structural
connectivity describes the effect of habitat size,
proximity, and quality on the connectivity of
individuals inhabiting that habitat (Kindlmann
and Burel 2008). For some marine species, the
spatial arrangement of their habitat is determined by other species, referred to as foundation
species. In such biogenic habitats, species like
coral reefs, kelp, marine algae, and seagrasses
host a vast biodiversity of associated invertebrates, fish, and seaweeds (Thompson et al. 2002,
O’Hara et al. 2008, Hewitt and Thrush 2010,
Perry et al. 2011). Structural connectivity in the
ocean will inevitably respond to changes in the
spatial rearrangement of biogenic habitats resulting from climate change. The following section
summarizes a few examples of the impacts of
increased ocean temperatures and ocean acidifi9
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
cation on foundation species that maintain
structural connectivity. Although we focus mostly on coral reef systems given the breadth of
evidence available for them, the framework
could be applied to other systems with major
foundation species.
The effects of acidification could be modified by
concurrent warming. Some species of coralline
crustose algae have been shown to be particularly sensitive to ocean acidification under warm
conditions in manipulative experiments (Kleypas
and Yates 2009, Doropoulos et al. 2012). In
general, acidification is expected to reduce
calcification rates in crustose coralline algae,
potentially reducing the amount of suitable coral
settlement habitat (Kleypas and Yates 2009,
Doropoulos et al. 2012). Predicted end of the
century CO2 levels have also been demonstrated
through experimental manipulations to significantly reduce crustose coralline algal cover and
consequently coral settlement, suggesting acidification may also reduce coral population recovery (Doropoulos et al. 2012). These combined
impacts will likely affect structural connectivity
for species dependent on biogenic habitats like
coral reefs (Hoegh-Guldberg et al. 2007).
Many habitat-forming species are also likely to
be affected by ocean acidification through shifts
in relative abundances that could lead to changes
in community composition. Acidification is expected to increase productivity of autotrophic
foundation species such as seagrasses, especially
in cold climates (Guinotte and Fabry 2008), with
positive effects on meadow size, the number of
meadows, and their isolation. In tropical systems,
natural in situ field studies detected an eightfold
increase in seagrass cover and a doubling in noncalcareous macroalgae (Guinotte and Fabry
2008). Although structural connectivity of coral
reefs may be reduced, connectivity among
seagrass meadows seems likely to increase.
This emerging evidence suggests both positive
and negative responses of marine foundation
species to ocean change, with the direction of
effect depending in part on local environmental
conditions, species composition, acclimation and
adaptation capacity, physiological responses to
interacting factors such as temperature, acidification, and nutrients (Pandolfi et al. 2011, Chan
and Connolly 2013). The effects of temperature
increases and acidification on the spatial arrangement of foundation species will contribute to
changes in structural connectivity. Reduced
habitat area could decrease structural connectivity, reducing immigration or metapopulation
dynamics that might enable patch size to be
maintained. In addition, chances for successful
Empirical examples
Ocean warming may impact habitat patch area
directly through effects on foundation species’
growth, survival, and resilience potential. For
example, climate change is expected to increase
the frequency and severity of extreme temperature events, which may lead to increased rates of
coral bleaching and subsequent decreases in
coral cover (McWilliams et al. 2005, Vivekanandan et al. 2009, Anthony et al. 2011, HoeghGuldberg 2011, Hoeke et al. 2011). Even when
corals acclimatize to increasing temperatures,
they experience reduced growth rates (Cantin et
al. 2010, Jones and Berkelmans 2010). The
combination of mortality from bleaching and
slower growth rates can lead to an overall
reduction in coral habitat (Alvarez-Filip et al.
2009, Cantin et al. 2010, Pandolfi et al. 2011).
Reduced habitat area implies increased isolation
between some remaining habitats, and these
changes in area and configuration could reduce
dispersal, colonization and alter community
structure.
Much of the research on the effects of
acidification on foundation species has focused
on corals. The general mechanism by which
corals or other calcifying organisms are affected
by ocean acidification is through the increase in
atmospheric CO2 altering the current dissolved
inorganic carbon distribution in seawater and
reducing its pH (Cohen and Holcomb 2009,
Findlay et al. 2010). The reduction of carbonate
ions (CO32) slows the rate at which calcifying
organisms produce CaCO3 (calcification) to a
point where rates of erosion exceed rates of
skeletal accretion, most likely due to the fact that
calcification is very energetically costly (Cohen
and Holcomb 2009). Quantitative syntheses of
the available experimental evidence on the
sensitivity of coral calcification to ocean acidification have suggested that under business-asusual emissions scenarios, coral calcification
could decline by ;15–22% on average by the
end of the century (Chan and Connolly 2013).
v www.esajournals.org
10
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
recruitment and population persistence could be
reduced because of an increase in the number of
smaller habitat patches and increased distance
between them (structural connectivity; Fig. 1b;
Guinotte et al. 2003, Munday et al. 2009a,
Hendriks et al. 2010). Decreases in suitable
habitat may not only reduce overall coral reef
area, but also decrease resilience to future
disturbances because new larval recruits will be
less likely to reach new areas to settle.
Metapopulation theory suggests that changes
to habitat area will have disproportionate effects
on population size and persistence (Kininmoth et
al. 2010, Kininmonth et al. 2011). With less
available habitat, particularly of foundation
species, there will not only be less larvae
produced, but also potentially longer distances
for settlers to travel. Recent work has used
regionally downscaled climate change models
that include ocean circulation together with
models that include habitat type and foundation
species to estimate changes in connectivity
(Mumby 1999, Mumby et al. 2011). In the future,
these modeling exercises should incorporate
changes in functional connectivity for a more
complete estimation of connectivity changes due
to climate change (Fig. 1).
RESERVE SIZE
SPACING
Determining adequate reserve size depends on
management goals. One approach is to make the
reserve larger than the mean dispersal distance to
ensure that the reserve persists independent of
dynamics in other reserves or fished areas
(Botsford et al. 2001, White et al. 2010). With
increasing reserve size there is eventually a
tradeoff in terms of reduced spillover (Gaines et
al. 2010). Our review suggest that with reduced
structural connectivity, reserves may need to be
larger to compensate for habitat loss (Fig. 1b),
consistent with previous suggestions (McLeod et
al. 2009).
In terms of spacing, network designs must
consider the mean larval dispersal distance of
target species (Botsford et al. 2001, Lockwood et
al. 2002, Botsford et al. 2003). Marine reserves are
thought to benefit fisheries primarily through
adult spillover to unprotected areas and larval
export from protected areas to other protected
areas as well as unprotected areas (Gaines and
Bertness 1992, Botsford et al. 2001, Botsford et al.
2003, Gerber et al. 2003, Gaines et al. 2010). If
reserves are spaced so that there is sufficient
connectivity among them, then persistence of the
entire network is possible even if no reserve is
self-persistent (Kaplan 2006). With variability in
dispersal distances, reserve spacing is less important than size (Kaplan 2006); yet spacing
must be close enough relative to dispersal
distance to allow network persistence (Moffitt
et al. 2009). Ocean warming and acidification
may alter the spatial scale of connectivity,
potentially increasing connectivity among nearby
habitats and reducing connectivity among already distant habitats. Thus, marine reserve
networks may need to be designed either with
a spatial arrangement that can accommodate
changes in the scale of connectivity (McLeod et
al. 2009), or with future flexibility that could
allow for the creation of additional reserves to
serve as stepping stones (Treml et al. 2008) to
connect distant habitats that are increasingly
isolated. Optimal reserve size and spacing will
clearly depend on the dispersal distance of the
target species, which may be related to PLD.
Shorter PLDs will require reserves to be spaced
closer together, but may also mean that reserve
size can be relatively small and still maintain
IMPLICATIONS OF CHANGING CONNECTIVITY
FOR MARINE CONSERVATION
Understanding how to incorporate climate
change into marine conservation strategies such
as marine reserve design and marine spatial
planning is critical to sustainably managing
marine populations. One common approach to
marine conservation is the establishment of
marine reserves, which can be planned based
on three general principles: (1) reserve size and
spacing must be scaled to mean dispersal
distance of target species, (2) reserves must
capture an adequate fraction of larvae of target
species, and (3) networks of reserves have
distinct advantages over single reserves (Botsford et al. 2003, Gerber et al. 2003, Gruss et al.
2011). Here we discuss implications of our
conceptual framework and synthesis for marine
reserve theory.
v www.esajournals.org
AND
11
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Fig. 3. Weighted mean density of larvae at (A) current temperature, (B) current temperature þ38C, and (C) þ38C
and acidification of 0.4 pH units. Passive larval trajectories were modeled using a three-dimensional, baroclinic
numerical model, and final location was determined using a temperature-dependent larval duration (Eq. 1).
Conservation priority areas including Rocas Consag (Graham et al. 2008) are indicated by black lines. Values
were calculated on a 1-km grid with a 5 3 5 km smoothing algorithm.
export. Longer PLDs will likely require larger
reserves, but the space between them can also be
greater.
CAPTURING
OF LARVAE
AN
Identifying where these places may persist in a
changing climate is critical to future reserve
design. During spatial planning exercises, particular attention could be given to protecting places
that support high levels of larval export as the
climate changes (i.e., ‘‘core areas’’). Including
smaller, but spatially closer reserves may ensure
connectivity even to more distant habitats. In
places where warming is expected to be greater,
these distances may need be relatively short. In
Box 1, we use a coupled bio-physical model for
the Gulf of California to illustrate implications for
reserve design (Table 2).
ADEQUATE FRACTION
Our review suggests that warming and OA
could reduce larval survival, changing total
larval output for several taxa (Anthony et al.
2008, Kroeker et al. 2010, Talmage and Gobler
2010). Botsford et al. (2001) propose a numerical
threshold of 35% of natural larval settlement in a
population that must be captured by reserves.
This value was derived for a case of no
reproduction outside reserves and for a settlerrecruit relationship that has a slope of 1/0.35 at
the origin (reflecting groundfish stocks off the
California and Oregon coasts). While 35% is not a
universal or absolute number (White et al. 2010),
reduced larval survival from increasing pH could
represent a threat for any minimum threshold
based on a healthy, non-exploited or otherwise
impacted population. To accommodate climatedriven increases in mortality, targets for minimum larval retention should be greater than 35%.
Although larger reserves could compensate for
reduced larval retention, there would be tradeoffs in the amount of larval export. Reserves
designed to protect key life stages where the
target species aggregate may be sufficient.
v www.esajournals.org
IMPLICATIONS
FOR
MARINE SPATIAL PLANNING
In the context of a framework based on
structural and functional connectivity, the evidence we have reviewed suggests that connectivity could be reduced in warmer, more acidic
oceans. We illustrate how these changes could
impact the effectiveness of the design of marine
conservation areas for a benthic marine invertebrate in the Gulf of California (Box 1, Fig. 3).
Understanding the precise impacts of warming
and acidification on marine connectivity will
require detailed studies aimed at quantitatively
determining scaling relationships and thresholds
between climate factors and critical demographic
parameters. It is important to quantitatively
consider the potential for general changes in the
12
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
size and spacing of reserves (Almany et al. 2009,
McLeod et al. 2010), particularly for foundation
or target species that are experiencing changes in
ocean temperature and acidification. While it is
well-established that a network of marine protected areas is needed to manage marine species
with a bipartite life cycle, reserves should be
closer together and larger in size to mitigate
climate change impacts (McLeod et al. 2009). In
fact, climate change could lead to some predictable (directional) changes in patterns. For example, our review suggests that conservation areas
may need to be closer together and larger in area
than previously thought. Increasing the proportion of habitats covered by reserves by increasing
reserve size and putting reserves closer together
would increase the likelihood for success. Ensuring that connectivity is maintained through
stepping stones that may not necessarily increase
the total proportion of area under reserves is also
possible.
Protection of areas of greater resilience or
resistance is often cited as an element of
mitigating the effects of climate change for
marine ecosystems (West and Salm 2003, Grimsditch and Salm 2005), although identifying these
areas is challenging (Chollett et al. 2010).
Furthermore, because many fish species migrate
to warmer water for reproduction, shortened
PLDs could actually increase MPA retention and
therefore benefit single MPA persistence.
In recent years, Marine Spatial Planning (MSP)
has become a popular approach to marine
conservation (Foley et al. 2010) because of its
focus on sustainable use for a wide range of
stakeholders. Most MSP efforts are spatially
explicit to account for a range of biological,
social, and economic factors (Crowder and Norse
2008, Ogden 2010). Some efforts include connectivity in the design framework (Beger et al. 2010a,
Beger et al. 2010b), but they rarely consider how
connectivity may change in coming decades.
Given that the potential effects of climate change
and ocean acidification on connectivity are
significant, ignoring these variations may compromise the long-term effectiveness of management and conservation actions. In light of
uncertainty about how ecosystem services associated with multi-use marine zones are going to
change with climate change, flexible zoning will
be essential in a changing world. In the spirit of
v www.esajournals.org
adaptive management, regular (e.g., decadal)
assessment of social and ecological patterns
should be conducted and spatial management
can be modified based on the best available data.
In addition to considering habitat and species
in conservation planning, spatial planning
should also include areas that are predicted to
exhibit different responses to climate change
(Ackerly et al. 2010). Incorporating this type of
‘‘climate heterogeneity’’ into marine spatial planning offers the opportunity to zone for higher
levels of protection in key areas that are not
predicted to change while enforcing some restriction in places that may become key larval
settling areas over time. Considering the full
range of uses that are typically included in MSP
is important to understand the relevance of
climate change and ocean acidification for
conservation prioritization.
CONCLUSION
Even in the absence of desired data, developing a framework for predicting changes in the
spatial distribution of marine organisms not only
highlights insights from available data and
models (Fig. 1), but also underscores the importance of improving our fundamental understanding of climate change impacts on marine systems.
Data on the interactive effects of ocean acidification and temperature as well as habitat data at
fine-scales may enable us to better assess how
changes in functional and structural connectivity
may affect marine populations and ultimately,
design conservation strategies aimed at maintaining their persistence. Our framework can be
used to incorporate other dimensions of climate
change, such as altered circulation patterns, sealevel rise, and frequency and intensity of storms
as these data become available. Integrating
fundamental insights of climate change impacts
on marine organisms into conservation planning
is critical to the future of marine biodiversity.
ACKNOWLEDGMENTS
We are grateful to Guido Marinone for providing
outputs from his circulation model and to Robert
Wildermuth for insightful comments on this manuscript.
13
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Slows Coral Growth in the Central Red Sea. Science
329:322–325.
Chan, N. C. S. and S. R. Connolly. 2013. Sensitivity of
coral calcification to ocean acidification: a metaanalysis. Global Change Biology 19:282–290.
Chollett, I., P. J. Mumby, and J. Cortes. 2010. Upwelling
areas do not guarantee refuge for coral reefs in a
warming ocean. Marine Ecology Progress Series
416:47–56.
Clark, J. S. 1998. Why trees migrate so fast: Confronting theory with dispersal biology and the paleorecord. American Naturalist 152:204–224.
Cohen, A. L. and M. Holcomb. 2009. Why corals care
about ocean acidification: Uncovering the mechanism. Oceanography 22:118–127.
Connell, S. D. and B. D. Russell. 2010. The direct effects
of increasing CO2 and temperature on noncalcifying organisms: increasing the potential for
phase shifts in kelp forests. Proceedings of the
Royal Society B 277:1409–1415.
Cowen, R. K. and S. Sponaugle. 2009. Larval Dispersal
and Marine Population Connectivity. Annual Review of Marine Science 1:443–466.
Crowder, L. and E. Norse. 2008. The role of marine
spatial planning in implementing ecosystem-based,
sea use management. Marine Policy 32:772–778.
Cudney-Bueno, R., M. F. Lavin, S. G. Marinone, P. T.
Raimondi, and W. W. Shaw. 2009. Rapid effects of
marine reserves via larval dispersal. PLoS One
4:e4140.
Devine, B. M., P. L. Munday, and G. P. Jones. 2012b.
Homing ability of adult cardinalfish is affected by
elevated carbon dioxide. Oecologia 168:269–276.
Devine, B. M., P. M. Munday, and G. P. Jones. 2012a.
Rising CO2 concentrations affect settlement behaviour of larval damselfishes. Coral Reefs 31:229–238.
Diaz-Pulido, G., M. Gouezo, B. Tilbrook, S. Dove, and
K. R. N. Anthony. 2011. High CO2 enhances the
competitive strength of seaweeds over corals.
Ecology Letters 14:156–162.
Doropoulos, C., S. Ward, G. Diaz-Pulido, O. HoeghGuldberg, and P. J. Mumby. 2012. Ocean acidification reduces coral recruitment by disrupting
intimate larval-algal settlement interactions. Ecology Letters 15:338–346.
Dupont, S. and M. C. Thorndyke. 2009. Impact of CO2driven ocean acidification on invertebrates early
life-history: What we know, what we need to know
and what we can do. Biogeosciences Discussions
6:3109–3131.
Fabricius, K. E., C. Langdon, S. Uticke, C. Humphrey,
S. Noonan, G. De’ath, R. Okazaki, N. Muehllehner,
M. S. Glas, and J. M. Lough. 2011. Losers and
winners in coral reefs acclimatized to elevated
carbon dioxide concentrations. Nature Climate
Change 1:165–169.
Figueira, W. F. and D. J. Booth. 2010. Increasing ocean
LITERATURE CITED
Ackerly, D. D., S. R. Loarie, W. K. Cornwell, S. B.
Weiss, H. Hamilton, R. Branciforte, and N. J. B.
Kraft. 2010. The geography of climate change:
implications for conservation biogeography. Diversity and Distributions 16:476–487.
Almany, G. R., M. L. Berumen, S. R. Thorrold, S.
Planes, and G. P. Jones. 2007. Local replenishment
of coral reef fish populations in a marine reserve.
Science 316:742–744.
Almany, G. R., S. R. Connolly, D. D. Heath, J. D.
Hogan, G. P. Jones, L. J. McCook, M. Mills, R. L.
Pressey, and D. H. Williamson. 2009. Connectivity,
biodiversity conservation and the design of marine
reserve networks for coral reefs. Coral Reefs
28:339–351.
Alvarez-Filip, L., N. K. Dulvy, J. A. Gill, I. M. Cote, and
A. R. Watkinson. 2009. Flattening of Caribbean
coral reefs: region-wide declines in architectural
complexity. Proceedings of the Royal Society B
276:3019–3025.
Anthony, K. R. N., D. I. Kline, G. Diaz-Pulido, S. Dove,
and O. Hoegh-Guldberg. 2008. Ocean acidification
causes bleaching and productivity loss in coral reef
builders. Proceedings of the National Academy of
Sciences USA 105:17442–17446.
Anthony, K. R. N., J. A. Maynard, G. Diaz-Pulido, P. J.
Mumby, P. A. Marshall, L. Cao, and O. HoeghGuldberg. 2011. Ocean acidification and warming
will lower coral reef resilience. Global Change
Biology 17:1798–1808.
Ayata, S. D., P. Lazure, and E. Thiebaut. 2010. How
does the connectivity between populations mediate
range limits of marine invertebrates? A case study
of larval dispersal between the Bay of Biscay and
the English Channel (North-East Atlantic). Progress in Oceanography 87:18–36.
Beger, M., H. S. Grantham, R. L. Pressey, K. A. Wilson,
E. L. Peterson, D. Dorfman, P. J. Mumby, R.
Lourival, D. R. Brumbaugh, and H. P. Possingham.
2010a. Conservation planning for connectivity
across marine, freshwater, and terrestrial realms.
Biological Conservation 143:565–575.
Beger, M., S. Linke, M. Watts, E. Game, E. Treml, I. Ball,
and H. P. Possingham. 2010b. Incorporating asymmetric connectivity into spatial decision making for
conservation. Conservation Letters 3:359–368.
Botsford, L., A. Hastings, and S. D. Gaines. 2001.
Dependence of sustainability on the configuration
of marine reserves and larval dispersal distance.
Ecology Letters 4:144–150.
Botsford, L. W., F. Micheli, and A. Hastings. 2003.
Principles for the design of marine reserves.
Ecological Applications 13:S25–S31.
Cantin, N. E., A. L. Cohen, K. B. Karnauskas, A. M.
Tarrant, and D. C. McCorkle. 2010. Ocean Warming
v www.esajournals.org
14
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
temperatures allow tropical fishes to survive
overwinter in temperate waters. Global Change
Biology 16:506–516.
Findlay, H. S., M. A. Kendall, J. I. Spicer, and S.
Widdicombe. 2010. Post-larval development of two
intertidal barnacles at elevated CO2 and temperature. Marine Biology 157:725–735.
Foley, M. M., B. S. Halpern, F. Micheli, M. H. Armsby,
M. R. Caldwell, C. M. Crain, E. Prahler, N. Rohr, D.
Sivas, M. W. Beck, M. H. Carr, L. B. Crowder, J. E.
Duffy, S. D. Hacker, K. L. McLeod, S. R. Palumbi,
C. H. Peterson, H. M. Regan, M. H. Ruckelshaus,
P. A. Sandifer, and R. S. Steneck. 2010. Guiding
ecological principles for marine spatial planning.
Marine Policy 34:955–966.
Gaines, S. D. and M. D. Bertness. 1992. Dispersal of
juveniles and variable recruitment in sessile marine
species. Nature 360:579–580.
Gaines, S. D., C. White, M. H. Carr, and S. R. Palumbi.
2010. Designing marine reserve networks for both
conservation and fisheries management. Proceedings of the National Academy of Sciences USA
107:18286–18293.
Gerber, L. R., L. W. Botsford, A. Hastings, H. P.
Possingham, S. D. Gaines, S. R. Palumbi, and S.
Andelman. 2003. Population models for marine
reserve design: a retrospective and prospective
synthesis. Ecological Applications 13:S47–S64.
Gooding, R. A., C. D. G. Harley, and E. Tang. 2009.
Elevated water temperature and carbon dioxide
concentration increase the growth of a keystone
echinoderm. Proceedings of the National Academy
of Sciences USA 106:9316–9321.
Graham, N. A. J., T. R. McClanahan, M. A. MacNeil,
S. K. Wilson, N. V. C. Polunin, S. Jennings, P.
Chabanet, S. Clark, M. D. Spalding, Y. Letourneur,
L. Bigot, R. Galzin, M. C. Öhman, K. C. Garpe, A. J.
Edwards, and C. R. C. Sheppard. 2008. Climate
warming, marine protected areas and the oceanscale integrity of coral reef ecosystems. PLoS One
3:e3039.
Grimsditch, G. D. and R. V. Salm. 2005. Coral reef
resilience and resistance to bleaching. IUCN,
Gland, Switzerland.
Gruss, A., D. M. Kaplan, S. Guenette, C. M. Roberts,
and L. W. Botsford. 2011. Consequences of adult
and juvenile movement for marine protected areas.
Biological Conservation 144:692–702.
Guinotte, J. M., R. W. Buddemeier, and J. A. Kleypas.
2003. Future coral reef habitat marginality: temporal and spatial effects of climate change in the
Pacific basin. Coral Reefs 22:551–558.
Guinotte, J. M. and V. J. Fabry. 2008. Ocean acidification and its potential effects on marine ecosystems.
Pages 320–342 in Year in ecology and conservation
biology. Annals of the New York Academy of
Sciences.
v www.esajournals.org
Hansen, J., M. Sato, R. Ruedy, K. Lo, D. W. Lea, and M.
Medina-Elizade. 2006. Global temperature change.
Proceedings of the National Academy of Sciences
USA 103:14288–14293.
Hendriks, I. E., C. M. Duarte, and M. Álvarez. 2010.
Vulnerability of marine biodiversity to ocean
acidification: A meta-analysis. Estuarine, Coastal
and Shelf Science 86:157–164.
Hewitt, J. E. and S. F. Thrush. 2010. Empirical evidence
of an approaching alternate state produced by
intrinsic community dynamics, climatic variability
and management actions. Marine Ecology Progress
Series 413:267–276.
Hirst, A. and A. López-Urrutia. 2006. Effects of
ecolution on egg development time. Marine Ecology Progress Series 326:29–35.
Hoegh-Guldberg, O. 2011. Coral reef ecosystems and
anthropogenic climate change. Regional Environmental Change 11:S215–S227.
Hoegh-Guldberg, O., P. J. Mumby, A. J. Hooten, R. S.
Steneck, P. Greenfield, E. Gomez, C. D. Harvell,
P. F. Sale, A. J. Edwards, K. Caldeira, N. Knowlton,
C. M. Eakin, R. Iglesias-Prieto, N. Muthiga, R. H.
Bradbury, A. Dubi, and M. E. Hatziolos. 2007.
Coral reefs under rapid climate change and ocean
acidification. Science 318:1737–1742.
Hoeke, R. K., P. L. Jokiel, R. W. Buddemeier, and R. E.
Brainard. 2011. Projected changes to growth and
mortality of Hawaiian corals over the next 100
years. PLoS One 6.
Huret, M., P. Petitgas, and M. Woillez. 2010. Dispersal
kernels and their drivers captured with a hydrodynamic model and spatial indices: A case study
on anchovy (Engraulis encrasicolus) early life stages
in the Bay of Biscay. Progress in Oceanography
87:6–17.
IPCC. 2007. Summary for policymakers. IPCC, Cambridge, UK.
Irisson, J. O. 2008. Behavioural approach to larval
dispersal in marine systems. Ecole doctorale de
l’Ecole Pratique des Hautes Etudes.
Johnson, J. E. and D. J. Welch. 2010. Marine fisheries
management in a changing climate: a review of
vulnerability and future options. Reviews in
Fisheries Science 18:106–124.
Jones, A. and R. Berkelmans. 2010. Potential costs of
acclimatization to a warmer climate: growth of a
reef coral with heat tolerant vs. sensitive symbiont
types. PLos One 5:e10437.
Jones, G. P., G. R. Almany, G. R. Russ, P. F. Sale, R. S.
Steneck, M. J. H. Oppen, and B. L. Willis. 2009.
Larval retention and connectivity among populations of corals and reef fishes: history, advances and
challenges. Coral Reefs 28:307–325.
Kaplan, D. M. 2006. Alongshore advection and marine
reserves: consequences for modeling and management. Marine Ecology Progress Series 309:11–24.
15
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Kindlmann, P. and F. Burel. 2008. Connectivity
measures: a review. Journal of Landscape Ecology
23:879–890.
Kininmonth, S., M. Beger, M. Bode, E. Peterson, V. M.
Adams, D. Dorfman, D. R. Brumbaugh, and H. P.
Possingham. 2011. Dispersal connectivity and
reserve selection for marine conservation. Ecological Modelling 222:1272–1282.
Kininmoth, S., M. Drechsler, K. Johst, and H. P.
Possingham. 2010. Metapopulation mean life time
within complex networks. Marine Ecology Progress Series 417:139–149.
Kinlan, B. P. and S. D. Gaines. 2003. Propagule
dispersal in marine and terrestrial environments:
a community perspective. Ecology 84:2007–2020.
Kleypas, J. A. and K. K. Yates. 2009. Coral reefs and
ocean acidification. Oceanography 22:108–117.
Kroeker, K. J., R. L. Kordas, R. Crim, I. E. Hendriks, L.
Ramajo, G. S. Singh, C. M. Duarte, and J. P.
Gattuso. 2013. Impacts of ocean acidification on
marine organisms: quantifying sensitivities and
interaction with warming. Global Change Biology,
in press.
Kroeker, K. J., R. L. Kordas, R. N. Crim, and G. S.
Singh. 2010. Meta-analysis reveals negative yet
variable effects of ocean acidification on marine
organisms. Ecology Letters 13.
Kurihara, H. 2008. Effects of CO 2-driven ocean
acidification on the early developmental stages of
invertebrates. Marine Ecology Progress Series
373:275–284.
Lett, C., S. D. Ayata, M. Huret, and J. O. Irisson. 2010.
Biophysical modelling to investigate the effects of
climate change on marine population dispersal and
connectivity. Progress in Oceanography 87:106–
113.
Lett, C., K. A. Rose, and B. Megrey. 2009. Biophyisical
models. Pages 88–111 in D. M. Checkley, J. Alheit,
Y. Oozeki, and C. Roy, editors. Climate change and
small pelagic fish. Cambridge University Press,
Cambridge, UK.
Levin, L. A. 2006. Recent progress in understanding
larval dispersal: new directions and digressions.
Integrative and Comparative Biology 46:282–297.
Lischka, S., J. Büdenbender, T. Boxhammer, and U.
Riebesell. 2011. Impact of ocean acidification and
elevated temperatures on early juveniles of the
polar shelled pteropod Limacina helicina: mortality,
shell degradation, and shell growth. Biogeosciences
8:919–932.
Lockwood, D. R., A. Hastings, and L. W. Botsford.
2002. The effects of dispersal patterns on marine
reserves: does the tail wag the dog? Theoretical
Population Biology 61:297–309.
Macreadie, P. I., M. J. Bishop, and D. J. Booth. 2011.
Implications of climate change for macrophytic
rafts and their hitchhikers. Marine Ecology Pro-
v www.esajournals.org
gress Series 443:285–291.
Marinone, S. G., M. J. Ulloa, A. Parés-Sierra, M. F.
Lavı́n, and R. Cudney-Bueno. 2008. Connectivity in
the northern Gulf of California from particle
tracking in a three-dimensional numerical model.
Journal of Marine Systems:149–158.
McLeod, E., A. Green, E. Game, K. Anthony, J. Cinner,
S. F. Heron, J. Kleypas, C. E. Lovelock, J. M.
Pandolfi, R. L. Pressey, R. Salm, S. Schill, and C.
Woodroffe. 2012. Integrating climate and ccean
change vulnerability into conservation planning.
Coastal Management 40:651–672.
McLeod, E., R. Moffitt, A. Timmermann, R. Salm, L.
Menviel, M. J. Palmer, E. R. Selig, K. S. Casey, and
J. F. Bruno. 2010. Warming seas in the coral
triangle: coral reef vulnerability and management
implications. Coastal Management 38:518–539.
McLeod, E., R. Salm, A. Green, and J. Almany. 2009.
Designing marine protected area networks to
address the impacts of climate change. Frontiers
in Ecology and the Environment 7:362–370.
McWilliams, J. P., I. M. Côté, J. A. Gill, W. J. Sutherland,
and A. R. Watkinson. 2005. Accelerating impacts of
temperature-induced coral bleaching in the Caribbean. Ecology Letters 86:2055–2060.
Meehl, G. A., W. M. Washington, W. D. Collins, J. M.
Arblaster, A. Hu, L. E. Buja, W. G. Strand, and H.
Teng. 2005. How much more global warming and
sea level rise? Science 307:1769–1772.
Metaxas, A. and M. Saunders. 2009. Quantifying the
‘‘bio-’’ components in biophysical models of larval
transport in marine benthic invertebrates: advances
and pitfalls. Biological Bulletin 216:257–272.
Miller, T. J. 2007. Contribution of individual-based
coupled physical-biological models to understanding recruitment in marine fish populations. Marine
Ecology Progress Series 347:127–138.
Moffitt, E. A., L. W. Botsford, D. M. Kaplan, and M. R.
O’Farrell. 2009. Marine reserve networks for
species that move within a home range. Ecological
Applications 19:1835–1847.
Mora, C. and P. F. Sale. 2002. Are populations of coral
reef fish open or closed? Trends in Ecology &
Evolution 17:422–428.
Morgan, S. G. 1995. Life and death in the plankton:
Larval mortality and adaptation. In L. R. McEdward and P. L. Lutz, editors. Ecology of marine
invertebrate larvae. CRC Press, Boca Raton, Florida, USA.
Mumby, P. J. 1999. Can Caribbean coral populations be
modelled at metapopulation scales? Marine Ecology-Progress Series 180:275–288.
Mumby, P. J., I. A. Elliott, C. M. Eakin, W. Skirving,
C. B. Paris, H. J. Edwards, S. Enriquez, R. IglesiasPrieto, L. M. Cherubin, and J. R. Stevens. 2011.
Reserve design for uncertain responses of coral
reefs to climate change. Ecology Letters 14:132–140.
16
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Munday, P. L., N. E. Crawley, and G. E. Nilsson. 2009a.
Interacting effects of elevated temperature and
ocean acidification on the aerobic performance of
coral reef fishes. Marine Ecology Progress Series
388:235–242.
Munday, P. L., D. L. Dixson, J. M. Donelson, G. P. Jones,
M. S. Pratchett, G. V. Devitsina, and K. B. Doving.
2009b. Ocean acidification impairs olfactory discrimination and homing ability of a marine fish.
Proceedings of the National Academy of Sciences
USA 106:1848–1852.
Munday, P. L., J. M. Leis, J. M. Lough, C. B. Paris, M. J.
Kingsford, M. L. Berumen, and J. Lambrechts.
2009c. Climate change and coral reef connectivity.
Coral Reefs 28:379–395.
Nakamura, M., S. Ohki, A. Suzuki, and K. Sakai. 2011.
Coral larvae under ocean acidification: Survival,
metabolism, and metamorphosis. PLoS One
6:e14521.
Neubert, M. G. and H. Caswell. 2000. Demography
and dispersal: Calculation and sensitivity analysis
of invasion speed for structured populations.
Ecology 8:1613–1628.
O’Connor, M. I., J. F. Bruno, S. D. Gaines, S. E. Lester,
B. S. Halpern, B. P. Kinlan, and J. M. Weiss. 2007.
Temperature control of larval dispersal and implications for marine ecology, evolution, and conservation. Proceedings of the National Academy of
Sciences USA 104:1266–1271.
Ogden, J. C. 2010. Marine spatial planning (MSP) A
first step to ecosystem-based management (EBM)
in the Wider Caribbean. Revista de Biologia
Tropical 58:71–79.
O’Hara, T. D., A. A. Rowden, and A. Williams. 2008.
Cold-water coral habitats on seamounts: Do they
have a specialist fauna? Diversity and Distributions
14:925–934.
Orr, J. C., V. J. Fabry, O. Aumont, L. Bopp, S. C. Doney,
R. A. Feely, A. Gnanadesikan, N. Gruber, A. Ishida,
F. Joos, R. M. Key, K. Lindsay, E. Maier-Reimer, R.
Matear, P. Monfray, A. Mouchet, R. G. Najjar, G.-K.
Plattner, K. B. Rodgers, C. L. Sabine, J. L. Sarmiento, R. Schlitzer, R. D. Slater, I. J. Totterdell, M.-F.
Weirig, Y. Yamanaka, and A. Yool. 2005. Anthropogenic ocean acidification over the twenty-first
century and its impact on calcifying organisms.
Nature 437:681–686.
Palumbi, S. R. 2004. Marine reserves and ocean
neighborhoods: The spatial scale of marine populations and their management. Annual Review of
Environment and Resources 29:31–68.
Pandolfi, J. M., S. R. Connolly, D. J. Marshall, and A. L.
Cohen. 2011. Projecting coral reef futures under
global warming and ocean acidification. Science
333:418–422.
Perry, C. T., M. A. Salter, A. R. Harborne, S. F. Crowley,
H. L. Jelks, and R. W. Wilson. 2011. Fish as major
v www.esajournals.org
carbonate mud producers and missing components
of the tropical carbonate factory. Proceedings of the
National Academy of Sciences USA 108:3865–3869.
Planes, S., G. P. Jones, and S. R. Thorrold. 2009. Larval
dispersal connects fish populations in a network of
marine protected areas. Proceedings of the National Academy of Sciences USA 106:5693–5697.
Rau, G. H., E. McLeod, and O. Hoegh-Gudberg. 2012.
The need for new ocean conservation strategies in a
high-carbon dioxide world. Nature Climate
Change 2:720–724.
Rayner, N. A., P. Brohan, D. E. Parker, C. K. Folland,
J. J. Kennedy, M. Vanicek, T. J. Ansell, and S. F. B.
Tett. 2006. Improved analyses of changes and
uncertainties in sea surface temperature measured
in situ since the mid-nineteenth century: the
HadSST2 dataset. Journal of Climate 19:446–469.
Sabine, C. L., R. A. Feely, N. Gruber, R. M. Key, K. Lee,
J. L. Bullister, R. Wanninkhof, C. S. Wong, D. W.
Wallace, B. Tilbrook, F. J. Millero, T. H. Peng, A.
Kozyr, T. Ono, and A. F. Rios. 2004. The oceanic
sink for anthropogenic CO2. Science 305:367–371.
Siegel, D. A., B. P. Kinlan, B. HGaylord, and S. D.
Gaines. 2003. Lagrangian description of marine
larval disperson. Marine Ecology Progress Series
260:83–96.
Soria, G., J. Tordecillas-Guillen, R. Cudney-Bueno, and
W. Shaw. 2010. Spawning induction, fecundity
estimation, and larval culture of Spondylus calcifer
(Carpenter 1857) (Bivalvia: Spondylidae). Journal
of Shellfish Research 29:143–149.
Talmage, S. C. and C. J. Gobler. 2010. Effects of past,
present, and future ocean carbon dioxide concentrations on the growth and survival of larval
shellfish. Proceedings of the National Academy of
Sciences USA 107:17246–17251.
Thompson, R. C., T. P. Crowe, and S. J. Hawkins. 2002.
Rocky intertidal communities: past environmental
changes, present status and predictions for the next
25 years. Environmental Conservation 29:168–191.
Tracey, S. R., K. Hartmann, and A. J. Hobday. 2012. The
effect of dispersal and temperature on the early life
history if a temperate marine fish. Fisheries
Oceanography 21:336–347.
Treml, E. A., P. N. Halpin, D. L. Urban, and L. F.
Pratson. 2008. Modeling population connectivity
by ocean currents, a graph-theoretic approach for
marine conservation. Landscape Ecology 23:19–36.
Turgeon, K., A. Robillard, J. Gregoire, V. Duclos, and
D. L. Kramer. 2010. Functional connectivity from a
reef fish perspective: behavioral tactics for moving
in a fragmented landscape. Ecology 91:3332–3342.
Ulloa, R., J. Torre, L. Bourillón, A. Gondor, and N.
Alcantar. 2006. Planeación ecoregional para la
conservación marina: Golfo de California y costa
occidental de Baja California Sur. Informe final a
The Nature Conservancy. Comunidad y Biodiver-
17
March 2014 v Volume 5(3) v Article 33
CONCEPTS & THEORY
GERBER ET AL.
Werner, F. E., R. K. Cowen, and C. B. Paris. 2007.
Coupled biological and physical models: present
capabilities and necessary developments for future
studies of population connectivity. Oceanography
20:54–69.
West, J. M. and R. V. Salm. 2003. Resistance and
resilience to coral bleaching: implications for coral
reef conservation and management. Conservation
Biology 17:956–967.
White, J. W., L. W. Botsford, A. Hastings, and J. L.
Largier. 2010. Population persistence in marine
reserve networks: incorporating spatial heterogeneities in larval dispersal. Marine Ecology Progress
Series 398:49–67.
sidad, A. C., Guaymas, Mexico.
Vikebø, F. B., S. Sundby, B. Ådlandsvik, and O. H.
Otterå. 2007. Impacts of a reduced thermohaline
circulation on transport and growth of larvae and
pelagic juveniles of Arcto-Norwegian cod (Gadus
morhua). Fisheries Oceanography 16:216–228.
Vivekanandan, E., M. H. Ali, B. Jasper, and M.
Rajagopalan. 2009. Vulnerability of corals to warming of the Indian seas: a projection for the 21st
century. Current Science 97:1654–1658.
Wernberg, T. S., D. A. Smale, and M. S. Thomsen.
2012. A decade of climate change experiments on
marine organisms: procedures, patterns and problems. Global Change Biology 18:1491–1498.
v www.esajournals.org
18
March 2014 v Volume 5(3) v Article 33