andrés hugo arias, maría clara menendez-marine ecology in a changing world-crc press (2013)

Marine Ecology in a Changing World

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Marine Ecology in a Changing World

Editors

Andrés Hugo AriasInstituto Argentino de Oceanografía

Bahía BlancaArgentina

María Clara Menendez Instituto Argentino de Oceanografía

Bahía BlancaArgentina

A SCIENCE PUBLISHERS BOOKp,

GL--Prelims with new title page.indd ii 4/25/2012 9:52:40 AM

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PrefacePreface

The world is rapidly changing. In recent decades, technological progress has been impressive in fi elds such as communications, computers, robotics, development of high precision acoustic instruments, diving equipment, etc. This modern technology has undoubtedly improved our ability to explore oceans and coasts, and get solid and reliable information about their ecology. At the same time, we have been gradually experiencing the effects of the global climate change: sea ice declination, receding of glaciers and permafrost, increased snow melt and runoff, shifted ranges for plants and animals, changes in populations, timing of many life-cycle events—such as blooms and migration-, decoupling of species interactions, damages due to droughts and fl oods, etc. The global ocean is no exception, and due to its extent, it is the largest, though silent ecosystem(s) under change.

In a broad sense, ecology is the study of organisms in relation to their surroundings. This book aims to cover the classic topics on marine ecology and the changes and deviations induced by climate change that modify the preexistent natural laws that govern the entire spectrum from species to ecosystem. With contributions from an impressive group of Argentinean and German oceanographers, Marine Ecology in a Changing World brings a comprehensive analysis of a discipline facing a turning point in recent history. The book begins with an overview of the fundamentals of marine ecology: ecosystem stability, water quality and biodiversity in the context of the documented world changes. The following chapters are organized in accordance with the major biological orders, from primary producers to large marine mammals, through to the primary consumers, benthic communities, seaweeds, wetlands and fi sheries. This information will provide students and researchers from the international scientifi c community with a wide view and present cutting-edge information about the marine life presently facing deviations from the classical theory.

Chapter 1 introduces general aspects of physical and chemical oceanography, dealing with the stressing changes affecting the stability and water quality of the oceans.

Chapter 2 deals with coastal marine biodiversity in the general context of the global change, considering some of the consequences of climate change

vi Marine Ecology in a Changing World

on the physical and chemical properties of coastal environments to later discuss how these changes affect the biotic components of the ecosystem.

Chapter 3 discusses both direct and indirect effects of ocean-climate change on the phytoplankton productivity, providing examples of proximate impacts on individuals, populations and communities by reviewing fi eld observations at different latitudes, empirical approaches and data modeling.

Chapter 4 summarizes the observed and potential future responses of zooplankton populations to climate change, focusing mainly on the effects that global warming, ocean acidifi cation and UV-radiation.

Chapter 5 describes the variability of benthic organisms in relation to climate change, mainly in the context of increasing temperatures and ocean acidifi cation. It also describes these effects on coral reefs and rocky intertidal habitats.

Chapter 6 discusses deviations affecting coastal wetland environments through the world, including changes driven by global atmospheric and climate alterations, coastal changes induced by human use of water on land, increased erosion of terrestrial sediments and direct human destruction of coastal habitats.

Chapter 7 introduces basic concepts of the seaweeds’ ecology, emphasizing their role in the climate change phenomenon. The chapter also illustrates some of the evidence for changes in the seaweed community, focusing on studies related to changes in temperature, UV-radiation, sea-level rise and ocean acidifi cation.

Chapter 8 summarizes the current and future impacts of climate-driven changes on the physiology and ecology of marine fi shes, and how world fi sheries are responding to these changes.

Finally, Chapter 9 deals with the natural history of marine mammals, analyzing how they were affected by climate change and also considering the anthropogenic causes.

The preparation of this book was significantly facilitated by the collaborative efforts of each of the authors. We are indebted to them, main players in the realization of this book, and the many other colleagues who provided suggestions and help during the entire process of development of the book. An acknowledgement is also given to the main editorial board and all the editorial staff who provided us with the confi dence and help to accomplish this project which started in late 2011.

August 2013 Andrés Hugo Arias María Clara Menendez

ContentsContents

Preface v

1. Potential Effects of Climate Changes on the Marine 1 Ecosystem Stability: Assessment of the Water Quality

Jorge Eduardo Marcovecchio, Natalia Sol Buzzi, Matías Nicolás Tartara, Carla Vanesa Spetter and Pia Simonetti

2. Coastal Marine Biodiversity Challenges and Threats 43 Jerónimo Pan, M. Alejandra Marcoval, Sergio M. Bazzini,

Micaela V. Vallina and Silvia G. De Marco 3. Climate Change Effects on Marine Phytoplankton 68 Valeria Ana Guinder and Juan Carlos Molinero 4. Climate Change and Marine Zooplankton 91 María C. Menéndez, Melisa D. Fernández Severini,

Florencia Biancalana, María S. Dutto, María C. López Abbate and Anabela A. Berasategui

5. Benthic Community and Climate Change 121 Sandra Marcela Fiori and María Cecilia Carcedo

6. The Ecology of Coastal Wetlands 140 Paula Daniela Pratolongo

7. Seaweeds Ecology and Climate Change 165 Gauna M.C., Croce M.E. and Fernández C.

8. World Fisheries and Climate Trend 194 Ana Laura Delgado and Maria Celeste Lopez Abbate 9. Marine Mammals in a Changing World 219 Cappozzo, Humberto Luis, Panebianco María Victoria and

Juan Ignacio Túnez

Index 261

CHAPTER 1

Potential Effects of Climate Changes on the Marine

Ecosystem StabilityAssessment of the Water Quality

Jorge Eduardo Marcovecchio,1,2,3,* Natalia Sol Buzzi,1,4,a Matías Nicolás Tartara,1,b Carla Vanesa Spetter,1,5,c and

Pia Simonetti1,d

Introduction

A huge amount of carbon is being annually released into the Earth’s atmosphere, reaching levels of gigatonnes (Jongen et al. 2011, Zhang et al. 2012). These accumulative post-industrial emissions have caused different effects, including increasing global temperature, rising sea level, changes

1 Área Oceanografía Química, Instituto Argentino de Oceanografía (IADO – CONICET / UNS). C.C. 804, 8000 Bahía Blanca, Argentina.

a Email: [email protected] Email: [email protected] Email: [email protected] Email: [email protected] 2 Fac.de Ingeniería, Universidad FASTA, Gascón 3145, 7600 Mar del Plata, Argentina.3 Universidad Tecnológica Nacional, Facultad Regional Bahía Blanca (UTN-FRBB), 11 de Abril

461, 8000 Bahía Blanca, Argentina.4 Depto.de Biología, Bioquímica y Farmacia, Universidad Nacional del Sur (UNS), San Juan

670, 8000 Bahía Blanca, Argentina.5 Depto.de Química, Universidad Nacional del Sur (UNS), Av Alem 1253, 8000 Bahía Blanca,

Argentina.* Corresponding author: [email protected]

2 Marine Ecology in a Changing World

in regional weather patterns, acidifi cation of oceans, variations in nutrient loads or alteration in ocean circulation (Brierley and Kingsford 2009). All these changes and others that may be occurring, affect biological processes taking place in the ocean at all levels, from the molecular to the ecosystemic one (Drinkwater et al. 2010). There is broad consensus that contemporary global climate change is a reality, and that much of the ongoing change is a direct result of human activity (IPCC 2007a). In particular, burning fossil fuels, making cement and changing land use have driven atmospheric carbon dioxide concentrations up from a pre-industrial value of about 280 ppm to 385 ppm in 2008 (Meure et al. 2006) (Fig. 1). Annual increases are now exceeding 2 ppm, an emission trend that exceeds the worst case scenario discussed at the Intergovernmental Panel on Climate Change (IPCC 2007b). There is a direct link between global temperature and CO2 concentration (IPCC 2007a). The increased heating in the lower atmosphere/Earth’s surface (radiative forcing) resulting from the “greenhouse” effect caused by increasing atmospheric CO2, methane and other gases (at a value of about 3 W.m–2, following IPCC 2007c) is unprecedented in at least the last 22,000 years (Joos and Spahni 2008) and has already had direct physical consequences for the marine environment and organisms living there. These include increases in mean global sea surface temperature by 0.13ºC per decade since 1979, and ocean interior temperature by >0.1ºC since 1961, increasing wind velocity and storm frequency, changes in ocean circulation, vertical structure and nutrient loads (IPCC 2007c), as well as rising sea level by more than 15 cm in the last century (Rahmstorf 2007) (Fig. 1), and presently by a mean of about 3.3 mm per year. Because the oceanic and atmospheric gas concentrations tend towards equilibrium, increasing CO2 pressure drives more CO2 into the ocean, where it dissolves forming carbonic acid (H2CO3) and thus increases ocean acidity; ocean pH has dropped by 0.1 (a 30% increase in H+ ion concentration) in the last 200 years (The Royal Society 2005) (Fig. 1).

Marine ecosystems clearly respond to changes in ocean variability and climate over a wide range of spatial and temporal scales (Mann and Lazier 1996, Southward et al. 2005, Drinkwater et al. 2010). The processes through which the physical environment affects the factors controlling primary production have long been known (Sverdrup 1953). These include infl uence on upper layer nutrient levels through mixing or upwelling, light levels through the effects of cloudiness or sea-ice coverage, and stratifi cation through changes in mixing or heat and salt fl uxes (Lavoie et al. 2009). For example, the relationship between mixing and production of phytoplankton in the North Atlantic depends upon the ratio of Sverdrup’s critical depth in spring to the mixed-layer depth at the end of the winter (Dutkiewicz et al. 2001). Where this ratio is near 1, as in the subtropical gyre, increased mixing reduces stratifi cation which tends to increase primary production

Climate Change & Marine Ecosystem Stability 3

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due to a rise in near surface nutrient concentrations. On the other hand, decreased production occurs in the waters within the subpolar gyre due to phytoplankton cells being mixed out of the euphotic zone because of the deeper mixed layer (Follows and Dutkiewicz 2002).

Numerous processes have been proposed to explain how climate forcing infl uences zooplankton and higher trophic levels (e.g., Bakun 2010, Brander 2010, Ottersen et al. 2010), and it is largely reported that climate impacts population dynamics of marine organisms indirectly through multi-step processes in food webs under “bottom-up” and “top-down” controls (Beaugrand et al. 2003, Ware and Thomson 2005, Frank et al. 2006, Perry and Schweigert 2008). In addition, it must be considered that climate also regulates metabolic factors (e.g., activity and feeding rates, swimming speeds, reproduction, etc. Pörtner et al. 2001, Pörtner 2002a, b). As a consequence, plankton and fi sh are often found in a limited range of hydrographic conditions, and large-scale shifts in water mass boundaries can lead to distributional changes of the fl ora and fauna (Brander 2010).

Accordingly, the current warming trends, largely thought to be caused by anthropogenic CO2 accumulation (IPCC 2007c), have resulted in poleward shifts in the geographical distribution of phytoplankton, macroalgae and marine-ectothermal animals and increased the risk of extinction of local species or even whole ecosystems such as coral reefs (Lüning 1990, Southward et al. 1995, Hoegh-Guldberg 1999, Harrington et al. 1999, Walther et al. 2002, Parmesan and Yohe 2003, Root et al. 2003, Thomas et al. 2004, Genner et al. 2004, Perry et al. 2005, Helmuth et al. 2006). Such changes are often related to thermal extremes such as maxima or minima, more than to the changing mean temperatures (Stachowicz et al. 2002, Stenseth and Mysterud 2002). Also, the recent decreasing frequency of colder winters and increased occurrence of warmer summers have impacted the population structure and community composition, as observed in the German Wadden Sea (Kröncke et al. 1998, Günther and Niesel 1999, Pörtner and Knust 2007).

The Large Climate Change Concern

Global climate change is a shift in the long-term weather patterns that characterize the regions of the world. Scientists state unequivocally that the Earth is warming. Natural climate variability alone cannot explain this trend. Human activities, especially the burning of coal and oil, have warmed the Earth by dramatically increasing the concentrations of heat-trapping gases in the atmosphere (Vijaya Venkata Raman et al. 2012). The more of these gases humans put into the atmosphere, the more the Earth will warm in the decades and centuries ahead. The impacts of warming can already be observed in many places, from rising sea levels to melting snow and ice to


changing weather patterns (Hinzman et al. 2005). Climate change is already affecting ecosystems, freshwater supplies, and human health. Although climate change cannot be avoided entirely, the most severe impacts of climate change can be avoided by substantially reducing the amount of heat-trapping gases released into the atmosphere (VijayaVenkataRaman et al. 2012).

Numerous studies related to different aspects of global climate change have been published in the last decades (e.g., Crane 1985, Crowley 1992, Norberg and DeAngelis 1997, Francis et al. 1998, Najjar et al. 2000, Rabalais et al. 2001, Moss et al. 2003, Straile et al. 2003, Ohring et al. 2005, Occhipinti-Ambrogi 2007, Adrian et al. 2009, Bardají et al. 2009, Coma et al. 2009, Collins et al. 2010, Eissa and Zaki 2011, Hollowed et al. 2012, Norris 2012). Nevertheless, many quite different topics are included within this literature, and consequently several concepts could be alternatively used in different ways and scenarios. In order to avoid this kind of problem various central topics must be clearly defi ned.

One signifi cant aspect of this topic is the understanding of how far can environmental changes modify the sensitivity of marine systems… In this sense, Perry et al. (2010a) have defi ned “sensitivity” as a measure of the strength in the relation between the biotic and the climate variables; for example, increasing sensitivity implies an increasing correlation between fl uctuations in population abundance (or another characteristic) and some climate signal, regardless of the mechanism by which climate variability affects the ecosystem functioning or structure (Lehodey et al. 2006, Drinkwater et al. 2010).

Variability is an inherent characteristic of marine ecosystems (e.g., Drinkwater et al. 2010). This variability is due to climate forcing, internal dynamics such as predator–prey interactions, and anthropogenic forcing such as fi shing. The latter has occurred for centuries (Jackson et al. 2001, Poulsen 2010), but is recognized as being globally more intensive since the onset of industrial fi shing in the 1950s (Pauly et al. 2002, Perry et al. 2010). Focusing on the issue at hand, it is highly advisable to distinguish between the two primary components of climate forcing of marine systems: (i) variability, and (ii) change (trend). How can each of them be defi ned?

The term “climate variability” is often used to denote deviations of climate statistics over a given period of time (such as a specifi c month, season or year) from the long-term climate statistics relating to the corresponding calendar period (Smit et al. 2000). In this sense, climate variability is measured by those deviations, which are usually termed anomalies. According to Overland et al. (2010) climate variability occurs on a wide range of time scales from seasonal periods, to 1–3 year oscillating but erratic periods (e.g., ENSO), to decadal aperiodic variability like 5–50 years, to centennial and longer periods. Climate variability includes extreme events,


such as the one in one hundred year storm, that may suffi ciently disrupt the system and cause it to move to a new state (Perry et al. 2010).

On the other hand, “climate change” is a significant and lasting change in the statistical distribution of weather patterns over periods ranging from decades to millions of years. It may be a change in average weather conditions, or in the distribution of weather around the average conditions (e.g., more or fewer extreme weather events) (Smit et al. 2000). Climate change is caused by factors that include oceanic processes (such as oceanic circulation), variations in solar radiation received by Earth, plate tectonics and volcanic eruptions, and human-induced alterations of the natural world (Brierley and Kingsford 2009). Climate change (trend) is defi ned as the secular change which at present, in the case of temperature, appears to be increasing and largely anthropogenically-driven, and whose rate is small as compared to that of the variability at the shorter time scales (Kelly and Adger 2000). Climate change may also affect climate variability, for example the frequency of El Niño or extreme events, although large uncertainties remain (e.g., Collins 2000).

Is Climate change a new story?

Earth’s climate has changed (Zachos et al. 2001), and will likely continue to change (Crowley and Hyde 2008), over multiple time scales. Temperature change is apparent in the existing instrument record, and numerous proxies enable past temperature variations to be reconstructed (Mann et al. 2008).

The geological record is punctuated by numerous abrupt changes in temperature. These discontinuities (for example, the Paleocene-Eocene Thermal Maximum 56 million years ago, when global temperatures rose by 6ºC in 20,000 years) defi ne boundaries between epochs of more consistency lasting tens of millions of years. During the Paleocene-Eocene Thermal Maximum 1500 to 2000 gigatonnes of carbon was released into the atmosphere in just 1,000 years; however, that rate is less than the one at which carbon is being now released through anthropogenic activity (The Royal Society 2005). Temperatures fell after the Paleocene-Eocene Thermal Maximum perhaps because of prolifi c growth of marine ferns Azolla (Brinkhuis et al. 2006), which reduced atmospheric carbon dioxide concentrations dramatically from 3500 ppm to 650 ppm (Pearson and Palmer 2000), switching Earth from “greenhouse” to “icehouse”. This switch well illustrates the power of marine biological infl uences on global climate.

Variations in solar activity and Earth’s orbit cause cyclical changes in temperature over tens to hundreds of thousands of years (so called Milankovitch cycles, according to Lisiecki et al. 2008). Feedback mechanisms involving greenhouse gases, ocean circulation and ice extent, which in turn infl uences albedo (the fraction of incoming solar radiation refl ected back to


space) interaction with Milankovitch cyclicity to provoke the Quaternary cycles of glaciation (c. 10ºC change with c. 100,000 year periodicity) that have persisted for the past 2.5 million years (Crowley and Hyde 2008). The last glaciation ended 12,000 years ago and Earth is presently in a warm period. Climatic changes have also occurred at higher frequencies (stadials/interstadials), but these changes are not necessarily global (Brierley and Kingsford 2009). In the north Atlantic region, for example, Dansgaard-Oeschger and Bond events (Bond et al. 1997) occur roughly every 1500 years, and include the beginning of the Younger Dryas and the Little Ice Age. Fluctuating ocean circulation and associated greenhouse gas variations are implicated in these climate oscillations (Schmittner and Galbraith 2008).

The climate history of the early Paleocene is marked by long-term global warming, beginning in the Late Paleocene (Selandian, ~59 Ma) and fi nishing in the Early Eocene (Ypresian, ~50 Ma) (Zachos et al. 2001, 2008). In addition to this long-term warming trend, a short term hypothermal event (ca 200 kyr) at the Paleocene–Eocene boundary (P–E) known as the Paleocene–Eocene Thermal Maximum (PETM) had a signifi cant impact on marine and terrestrial biota (Zachos et al. 2005, Tripati and Elderfi el 2005).

In addition, both the abrupt environmental change and extinction events may also result from a discontinuous climate response to slowly varying terrestrial boundary conditions; that is, under certain conditions, instabilities in the climate system can be triggered by small changes in force (Smith A. et al. 2001). Theoretical support for the hypothesis of abrupt climate change is based on climate model results that suggest the presence of multiple equilibrium climate states for a given level of forcing. Transitions between states at “critical points” can be rather sudden and can be caused by small changes in forcing (Crowley and North 1988).

There are some particularly good examples of abrupt climate change in records from the Quaternary: the terminations of Pleistocene glaciations (Corliss et al. 1984), the “Younger Dryas” cool oscillation during the last deglaciation (Brauer et al. 2008), evidence for rapid climate swings in the interstadial preceding the last glacial maximum (Denton et al. 2010), the abrupt initiation of glaciation during the early stages of a glacial cycle (Zacos and Kump 2005), and a relatively abrupt transition in the dominant period of glaciations during the mid-Pleistocene (Sosdian and Rosenthal 2009).

There is also evidence of signifi cant changes in the evolution of climate for the last 100 million years (Ma) (Fig. 2a). The long-term trend involves the evolution of climate from an ice-free earth in the mid-Cretaceous (100 Ma) to a bipolar glacial state with periodic glacial expansion into northern mid-latitudes (Poulsen et al. 2001). There have also been signifi cant increases in aridity during the last 30 Ma (Wolff et al. 2006). Each stage in the isotopic curve presumably involves one stage in the evolution of this


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process—for example, the development of a cold deep-water circulation, separate development of the East and West Antarctic Ice Sheets, initiation of Arctic Ocean ice cover and glaciation on Greenland, and onset of signifi cant mid-latitude Northern Hemisphere glaciation.

When compared with the long-term paleoclimate record, the Cretaceous-Tertiary (K-T) extinction stands out as somewhat different from the other extinctions (Crowley and North 1988). The background oxygen isotope record is relatively stable over a 10 to 15 Ma interval bracketing the event (Fig. 2a), so there is no step-function change in the climate. There was a general fall in sea level between the late Cretaceous and early Tertiary (Miller et al. 2003), but with little geological evidence that it may have been associated with an ice-growth event (Miller et al. 2008). The effect of abrupt climate change on organisms can be evaluated in more detail by comparing the oxygen isotope record of the last 100 Ma (Fig. 2a) with extinction events in marine invertebrates (Regan et al. 2001) over the same interval (Fig. 2b). First three of the extinction events coincide to some degree with the three major steps in the evolution of Cenozoic climate: the onset of mid-latitude Northern Hemisphere glaciation at about 2.4 to 3.0 Ma (Schaefer et al. 2006); expansion of ice on Antarctica between about 10 and 14 Ma (Shevenell et al. 2004); and major cooling between about 31 and 40 Ma (Bond et al. 1993).

A fourth extinction event at about 90 Ma coincides with a major environmental change not manifested in the oxygen isotope record: an ocean anoxic event (Leckie et al. 2002) that correlates with the highest sea level of the last 200 Ma (Miller et al. 2005) and with an abrupt change in carbon isotopes in pelagic carbonates (Hesselbo et al. 2007). Changes in organic carbon burial may have signifi cantly affected atmospheric pCO2 levels at this time (Royer et al. 2004). This last event is therefore also a candidate for an abrupt environmental change due to slowly changing boundary conditions. Some of the second-order trends in the oxygen isotope record also correlate with smaller extinction events (Wing et al. 2005).

However, the 18O event at 36 Ma (Fig. 2a) represents only one of at least three stages of climate change that resulted in an overall transition from the warm climates of the Early Tertiary to the cool climates of the Late Tertiary: Late Eocene cooling (36 to 40 Ma), abrupt bottom water cooling with some ice growth at about 36 Ma, and a major sea level fall and presumed ice growth event at about 31 Ma (Crowley and North 1988).

A different time-scale observation: the decadal climate variability

The effects of anthropogenically forced climate change are expected to continue through the twenty-fi rst century and beyond. However, on a timescale of a few years to a few decades ahead, future regional changes in weather patterns and climate, and the corresponding impacts, will also


be strongly infl uenced by natural unforced climate variations (Folland et al. 2009). Numerous studies linked with this kind of processes have been reported in the international scientifi c literature, and deserve to be highlighted. In this sense the review by Murphy et al. (2010) showed in a very integral way several remarkable examples of sustained (decadal-scale) climate variations with signifi cant impacts on society: the United States 1930s dust bowl droughts (Seager et al. 2008); unusual rainfall in India (Mehta and Lau 1997) and China (Hameed et al. 1983); fl oods in the Nile river (Kondrashov et al. 2004); droughts in the Northeast region of Brazil (Mehta 1998); the current drought in the south-western United States (Barnett et al. 2008); Sahel drought of the 1970s and 1980s (Lu and Delworth 2005); variability in Atlantic hurricane activity (Goldenberg et al. 2001, Zhang and Delworth 2006); Arctic warming in the 1930s–1940s (Semenov and Bengtsson 2003, Johannessen et al. 2004); the mid-1970s climate shift in the Pacifi c (Meehl et al. 2009); rapid warming in European winter temperatures from the 1960s to the 1990s (Scaife et al. 2005); variations of the Caspian Sea level (Rodionov 1994); and others.

The decadal timescale is widely recognized as a key planning horizon for governments, businesses, and other societal entities (Vera et al. 2009), and its importance is fully recognized by the Intergovernmental Panel on Climate Change (IPCC 2007a).

On decadal timescales, regional anthropogenically forced changes can be expected, but will typically be smaller than internal variability. There is emerging evidence, however, that some aspects of internal variability could be predictable for a decade or longer in advance (Murphy et al. 2010). These studies address the possibility of achieving skill in multi-year means of global or large-scale regional surface temperature.

To achieve an adequate prediction of decadal climate variations, reasonably well designed ecosystemic models should be applied to solve different problems which are simultaneously acting (Sohngen et al. 2001). The aspects that must be considered include some which have been treated by different authors and deserve to be highlighted. For example, Meehl et al. (2005) emphasized the commitment to future climate change arising from incomplete adjustment to past changes in external forcing. In addition, Stott and Kettleborough (2002) and Lee et al. (2006) have commented on the effects of future changes in anthropogenic forcing, noting that the effects of explosive volcanic eruptions are also potentially important (Mann et al. 2005), but cannot be predicted in advance. Moreover, analyses by Hurrel et al. (2009) have focused in the potential predictability of internal variability arising from initialization of slowly varying components of the climate system. In this sense, and directed to major enhancements of observational networks, particularly in the oceans, this will require further development in initialization techniques (Glenn et al. 2000). Consequently, more ambitious


strategies will be needed for the design of ensemble climate model projections in order to understand and quantify decadal predictability and how it may be affected by forced climate change (Murphy et al. 2010).

According to Murphy et al. (2010), several of the main indexes of DCV (decadal climate variability) which deserve to be considered in the present analysis are:

The North Atlantic Oscillation and the Atlantic Multidecadal Oscillation

Sir Gilbert Walker of the India Meteorological Department fi rst discovered a north-south atmospheric pressure “seesaw” which he termed as the North Atlantic Oscillation (NAO) in the late 1920s (Walker and Bliss 1932). This north-south pattern oscillates at a variety of timescales, among them decadal and longer periods (Hurrell 1995, Hurrell and van Loon 1997). In the last 10–15 years, the Arctic and Antarctic Oscillations (AO and AAO, respectively) have been associated with climate variability over the two respective high latitude regions (Thompson and Wallace 2000). The NAO is believed to be the North Atlantic component of the AO (Marshall et al. 2001).

The Atlantic Multidecadal Oscillation (AMO) (Delworth and Mann 2000, Knight et al. 2005) is a broad hemispheric pattern of multidecadal variability in surface temperature, centred on the North Atlantic basin (Fig. 3a).

The Tropical Atlantic SST Gradient Oscillation

The tropical Atlantic Sea Surface Temperature (SST) gradient (TAG) across the equator is known to vary at the 12 to 13 year period (Chang et al. 1997, Sutton et al. 2000). Variability of many atmosphere and ocean variables are associated with the TAG variability, such as winds in the lower troposphere; heat transferred between the Atlantic Ocean and the overlying atmosphere; cloudiness; rainfall in North-east Brazil and West Africa; Atlantic hurricanes; and water vapour infl ux and rainfall in the southern, central, and mid-western United States (e.g., Mehta 1998, Hurrell et al. 2006, Murphy et al. 2010).

The North Pacifi c Oscillation, the Pacifi c Decadal Oscillation and the Interdecadal Pacifi c Oscillation

Sir Gilbert Walker also discovered a phenomenon which he termed as the North Pacifi c Oscillation (NPO) in the 1920s (Walker 1925). The NPO is a seesaw in atmospheric pressure between sub-polar and sub-tropical latitudes in the North Pacifi c region (Murphy et al. 2010). Subsequently, when longterm SST data in the Pacifi c Ocean became available in the 1990s, a number of researchers found that the dominant pattern of SST variability in the extra-tropical North Pacifi c varied at timescales of one or


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2010

).


more decades, and that this SST pattern was associated with the NPO in the atmosphere (Kushnir et al. 2002). This SST pattern is called the Pacifi c Decadal Oscillation (PDO) (Mantua et al. 1997). The Interdecadal Pacifi c Oscillation (IPO) (Power et al. 1999) is a Pacifi c-wide SST pattern covering both hemispheres, showing a similar pattern of variability to the PDO in the North Pacifi c (Folland et al. 2002). The IPO is characterized by year-to-year and longer-term, predominantly decadal-to-multidecadal, variability of the Pacifi c Ocean SSTs, with opposite phases between the tropical-subtropical Pacifi c Ocean and the mid-latitude Pacifi c Ocean in both hemispheres (Bridgman and Oliver 2006) (Fig. 3b).

Decadal modulation of higher frequency phenomena

There is evidence that shorter-term phenomena, such as El Niño–Southern Oscillation (ENSO) events, heavy rainfall events and occurrences of tropical cyclones undergo signifi cant decadal modulation. In particular, the frequency, intensity, spatial pattern and predictability of interannual El Niño–Southern Oscillation (ENSO) events have been found to undergo decadal–multidecadal variability (Kestin et al. 1998, Torrence and Webster 1999, Rajagopalan et al. 2000, England and Huang 2005, Murphy et al. 2010). Predictability of ENSO impacts on Australian climate was found to be modulated by the IPO such that in the warm IPO phase, there is no robust relationship between year-to-year Australian climate variations and ENSO. In the cold IPO phase, year-to-year ENSO variability is closely associated with year-to-year variability in rainfall, surface temperature, river fl ow and the domestic wheat crop yield in Australia (Power et al. 1999, Arblaster et al. 2002). Moreover, ENSO impacts on North American climate were also found to be modulated by the NPO (Bonsal et al. 2001, Di Lorenzo et al. 2010).

However, it is very important to clearly understand that all these signals can be expressed simultaneously and not in an isolated way. As an example, and according to Hunt Jr. and Stabeno (2002) the Bering Sea, as a marginal ice zone, should be particularly sensitive to climate change, because small changes in wind velocities can make large differences in the extent, timing and duration of wintertime sea ice. Although such far-reaching signals as El Niño/Southern Oscillation (ENSO) on occasion may affect the climate of the Bering Sea (e.g., Overland et al. 2001), the climate of the southeastern Bering Sea is most strongly infl uenced by the Pacifi c North American pattern (PNA) (with which the Pacifi c Decadal Oscillation—PDO—is correlated), and by the Arctic Oscillation (AO) (Overland et al. 1999). Recent work has shown that ecosystem responses to decadal-scale changes in these and other indices of North Pacifi c Ocean and Bering Sea climate have been pervasive and of great economic importance (Francis et al. 1998, Hare and Mantua 2000, McFarlane et al. 2000, Hollowed et al. 2001).


What does “stability of the ecosystem” Mean?

By definition, stability is the ability of an ecosystem to return to an equilibrium state after a temporary disturbance (Holling 1973). MacCillivray and Grime (1995) updated this defi nition considering both the community’s ability to (1) resist change in order to maintain an ecosystem function (resistance), and (2) recover to normal levels of function after disturbance (resilience).

Many marine ecosystems of the world share a similar confi guration of their biological community structure, characterized by a crucial intermediate trophic level often occupied by a small plankton-feeding pelagic species (Bakun 1996). The major control of trophic dynamics in these wasp-waist ecosystems (sensu Rice 1995) is neither “bottom-up” nor “top-down” but rather “both up and down from the middle”, as variations in size of these populations may have major effects on the trophic levels above, which depend on the wasp-waist species as their major food source, and also on the trophic levels below, which are fed upon by massive wasp-waist populations (Bakun 1996).

In the ecosystem development theory of Odum (1969), stability is viewed as one property of mature ecosystems, which tend, over time, to increase in size and diversity within the constraints of available resources. Hence, along with other characteristics, mature ecosystems, according to Odum (1969), should incorporate a high information content, attain high biomass, and have a high capacity to entrap and hold nutrients for cycling within the system.

System recovery time, defi ned as the time required for all elements of a system to return to their baseline biomass values following a perturbation, is used here as a measure of ecosystem internal stability, or resistance to external changes. This approach seeks to identify the ecosystem attributes directly involved in the stability and to address their relation to ecosystem maturity (Christensen 1995, Ludovisi et al. 2005). A comparative analysis of systems behavior was carried out using a set of ecosystem goal functions previously identifi ed as representative of Odum’s attributes of ecosystem maturity (Christensen 1995). The attributes were chosen to represent three different aspects of ecosystem development: (i) complexity in community structure; (ii) community energetics; and, (iii) overall community homeostasis.

According to Holling et al. (1995) the structure of biological communities is therefore controlled through the balance of destabilizing and stabilizing forces. While destabilizing forces are important in maintaining diversity, resilience and opportunity, stabilizing forces, such as nutrient recycle, are important in maintaining productivity and biogeochemical cycles. The role of destabilizing forces may be particularly important in pelagic marine


ecosystems. Thus, for example, in the sea, short-term variability is damped out by very large heat capacity of the ocean. In turn, this large thermal capacity and the long period exchange rates between deep and near-surface waters leads to relatively large-amplitude changes at the long term scales (Steele 1985). As a result, less robust internal ecosystem processes are needed to handle the smaller amplitude variability at short periods. The possible absence of such mechanisms, combined with increase variance with period, can mean that pelagic marine populations or ecosystems have to continually adapt to physical variability in the short as well as the long term (Holling et al. 1995).

The absence of well structured recycling routes, the low recycling and reduced stability of upwelling ecosystems can be considered a result of a longer-term adaptation of biological community to the physical variability and transitory nature of these systems. Bakun (1996) considered variability itself as a key asset for the massive small pelagic wasp-waist populations inhabiting upwelling systems, which must rely on pulsing its abundance to cope with the temporal and spatial patterns presented by their prey, while simultaneously presenting patterns to their predators that overcome growth of intolerable levels of predation.

For instance, while the internalization of system activity by recycling renders resistance to change (increasing stability), the lack of redundancy in the recycling pathways could make it very diffi cult for a highly organized system to reestablish broken pathways (Ulanowicz and Wulff 1991). In this sense, the environmental price for stability would be a decrease in the resilience of the studied ecosystem (Holling 1973), that is of their ability to absorb changes and still persist in a state of high biomass.

The vulnerability of marine ecosystems, the value of the ecosystem services they provide, and the need for different approaches in understanding and managing human activities that affect oceans have recently received much attention (Levin and Lubchenco 2008). Reports from the Pew Oceans Commission (2003), the US Commission on Ocean Policy (2004), the Joint Ocean Commission Initiative (2006), the Millennium Ecosystem Assessment (2006), among others, draw attention to the seriously disrupted state of marine ecosystems, a result of climate change, coastal development, overexploitation of ocean resources, nutrient and chemical pollution from the land, and other anthropogenic infl uences. Disruption of marine ecosystems diminishes ecosystem services such as the provision of fi sh and other seafood, the maintenance of water quality, and the control of pests and pathogens (Levin and Chan 2012). The collective conclusion of these reports is that if people wish to have safe seafood, stable fi sheries, abundant wildlife, clean beaches, and vibrant coastal communities, priority must be given to protecting and restoring the coupled land-ocean systems that provide these services (Levin and Lubchenco 2008).


How the stability of an ecosystem can be measured

According to Grimm and Wissel (1997), the stability concept is a collective notion or term, which is defi ned via three fundamental properties: constancy (a system staying essentially unchanged), resilience (the ability of a system to return to the reference or dynamic state after a temporary disturbance), and persistence (the ability of a system to persist through time).

Several attempts have been undertaken to investigate the relationship between biodiversity and the stability properties of an ecosystem, using different proxies, habitats or types and levels of disturbance (e.g., Loreau et al. 2001, Balvanera et al. 2006, Isbell et al. 2009, Campbell et al. 2011, Godbold et al. 2011). One of the hypotheses tested states that “higher biodiversity promotes higher stability” (e.g., MacArthur 1955, Odum 1959, Margalef 1969). Several decades after its formulation, there is still a lack of comprehension regarding the relation between biodiversity and stability (e.g., Worm et al. 2006, Ives and Carpenter 2007, Baraloto et al. 2010). One of the major diffi culties relies on the selection and use of tools and measures able to correctly “quantify” the system stability properties.

A few studies (e.g., Srivastava and Vellend 2005, Tilman et al. 2006, Bodin and Wiman 2007) have tried to assess the connection between ecosystem stability and services provision. In addition the results from several authors (e.g. Hooper et al. 2005) have suggested that ecosystem functions are more stable through time at relatively high levels of biodiversity.

Some authors (e.g., Winfree and Kremen 2009, Haines-Young and Potschin 2010) have suggested that both the level and stability of ecosystems tend to improve with increasing biodiversity through space and time, importantly, although most of these studies were conducted in terrestrial ecosystems (e.g., Kremen et al. 2002, Tilman et al. 2005), and there are very few cases where this relationship has arisen for aquatic ecosystems (e.g., Valdivia and Molis 2009). Transitional habitats, like estuaries, are particularly challenging for many reasons all over the world (Pinto et al. 2013). Most important are: (1) biological communities under naturally stressful conditions (Elliott and McLusky 2002); (2) biota under multiple anthropogenic pressures (Wilkinson et al. 2007); (3) estuarine communities characterized by low number of species and high species abundance (Elliott and Quintino 2007), although their number is increasing due to invaders (Nehring 2006).

In this sense, Tilman (1999), Lehman and Tilman (2000) and Tilman et al. (2006) proposed the use of “temporal stability” (TS), defi ned as the ratio of mean abundance to its standard deviation, to test the diversity–stability hypothesis. Within this framework Pinto et al. (2013) suggest that the diversity–stability relationships are neither linear nor monotonic in estuaries due to their complexity. The observed stability results appeared to be more


associated to species abundance than to species richness, suggesting that biodiversity may act not only as a measure of biophysical integrity (Smith 1994), but also as a contributor to overall stability. TS has been estimated using the coeffi cient of variation [CV = 100/(standard deviation/mean)], for which smaller values represent greater stability (Tilman 1999). For example, the TS of a system could be quantifi ed as mean macroinvertebrate biomass (b, gC m−2) divided by the standard deviation of community biomass production through time:

TS = b / σxi xj

The ecosystem concept cuts through the myriad of complex interactions at a species level by focusing on a small subset of average or integrated properties of all the populations within the area of study. Its big advantage is that it can identify emergent properties such as energy fl ow and nutrient cycling and study the stability of function of this abstract structure (Allen 2010, Allen and Fulton 2010). The weakness lies in its ability to explain the relative stability of ecological systems in a changing environment; the focus on a self regulating system leading to a focus on local and short term stability (i.e., recovery from disturbance) rather than fl exibility in the sense of maintaining variability in space and time as conditions change (O’Neill 2001). The result of the ecosystem approach has been the development of the current generation of coupled bio-physical models, with a limited ability to respond to environmental change. However, there is a requirement to understand the impact of multiple climatic and anthropogenic drivers on the whole ecosystem, which requires the development of a new generation of end to end models (Parkes et al. 2003). Another driver in model development has been the increase in knowledge of “previously non considered processes”, e.g., the microbial loop (Azam et al. 1983), iron limitation (Martin and Fitzwater 1988) and ocean acidifi cation (Raven et al. 2005). This has led to increasing model complexity, but often at a rate where the speed with which processes are included in models outstrips the modelling community’s ability to realistically parameterize them (Anderson 2005). This is despite the increased availability of integrated data sets, such as the World Ocean Atlas (http://www.nodc.noaa.gov/OC5/WOA05/pr_woa05.html).

One point that engages the effects of climate on the ecosystem and its characteristics is the “regime shift”. The term regime shift has been used to describe large, decadal-scale switches in the abundance and composition of organisms within the ecosystem (e.g., plankton or fi sh) (Reid et al. 2001). As an example, Venrick et al. (1987) reported an increase in phytoplankton biomass before and after this period, which they attributed to a modifi cation of atmospheric circulation. Spatial gradients in sea level pressure increased the shift. Increased strength and frequency of storminess and westerly winds


allowed a deeper mixing and a transfer of more nutrients to the surface. This climatic forcing modifi ed the carrying capacity of the central North Pacifi c gyre, contributing to an increase in the abundance of fi shes such as the Alaskan salmon and cod, and a decrease in the abundance of shrimps.

It has been diffi cult to demonstrate shifts between alternative stable dynamic regimes in the real world (Scheffer et al. 2001). To demonstrate that an ecosystem regime shift may have actually happened stepwise changes should be detected (1) across different trophic levels, (2) at the level of the community structure, (3) for key species, (4) in attributes of ecosystems such as diversity, and (5) one should expect that ecosystem changes would refl ect hydro-climatic variability.

Effects of Climate Change on Marine Ecosystems

The functioning of marine ecosystems is supported by the fl ow of energy going from primary producers to intermediate consumers, top predators (including humans) and pathogens, and then back through the process of decomposition and generation of debris (Moore and de Ruiter 2012). So, it is clearly understood that marine communities are biological networks where the success of species is directly or indirectly linked through various biological interactions (e.g., predator-prey relationships, competition, facilitation, mutualism) to the performance of other species within the community (Werner and Peacor 2003). Within this theoretical framework, Doney et al. (2009) emphasized that the aggregate effect of these interactions constitutes ecosystem function (e.g., nutrient cycling, primary and secondary productivity), through which ocean and coastal ecosystems provide the wealth of free natural benefi ts that society depends upon, such as fi sheries and aquaculture production, water purifi cation, shoreline protection and recreation.

Climate change pressures are having profound and diverse consequences for marine ecosystems. Rising atmospheric CO2 is one of the most critical problems because its effects are globally pervasive and irreversible on ecological timescales (Raven et al. 2005). The primary direct consequences are increasing ocean temperatures (IPCC 2007c) and acidity (Doney et al. 2009). Climbing temperatures create a host of additional changes, such as rising sea level, increased ocean stratifi cation, decreased sea-ice extent, and altered patterns of ocean circulation, precipitation, and freshwater input. In addition, both warming and altered ocean circulation act to reduce subsurface oxygen (O2) concentrations (Keeling et al. 2010). In recent decades, the rates of change have been rapid and may exceed the current and potential future tolerances of many organisms to adapt. Further, the rates of physical and chemical change in marine ecosystems will


almost certainly accelerate over the next several decades in the absence of immediate and dramatic efforts toward climate mitigation (IPCC 2007c).

Direct effects of changes in ocean temperature and chemistry may alter the physiological functioning, behavior, and demographic traits (e.g., productivity) of organisms, leading to shifts in the size structure, spatial range, and seasonal abundance of populations. These shifts, in turn, lead to altered species interactions and trophic pathways as change cascades from primary producers to upper-trophic-level fi sh, seabirds, and marine mammals, with climate signals thereby propagating through ecosystems in both bottom-up and top-down directions. Changes in community structure and ecosystem function may result from disruptions in biological interactions (Doney et al. 2012).

Considering these comments it is important to use models of evaluation which allow to understand and predict the effects of climate variability on marine food webs and marine productivity which are of great importance. This is especially true with respect to potential consequences of climate change on commercially important fi sheries. The use of hydrographic models coupled to Nutrient–Phytoplankton–Zooplankton–Detritus (NPZD) ones to describe and predict future marine ecosystem dynamics has demonstrated to be a useful approach that is becoming increasingly widespread (Gibson and Spitz 2011). This kind of model includes the processes linking the different components of the water column, which determine both the functioning and stability of the system: nutrients and phytoplankton + microzooplankton + mesozooplankton + jellyfi sh + detritus. In addition, particular “submodel conditions” (e.g., benthic biogeochemical submodel, ice submodel, etc.) can be input on the described biological scenario. Finally, this module is fully coupled with a strong physical model (like ROMS: Regional Ocean Modelling System) which includes the main processes of the water column and atmosphere (e.g., currents, physical-chemical properties of seawater, radiation, air pressure and temperature, etc.). This kind of numeric tool allows to simulate the marine environment functioning on a real data basis, as well as to modify scenarios (e.g., climate changes) to evaluate corresponding consequences on the biological system (Perry 2010, Jørgensen et al. 2012).

Which are the main parameters to be considered within the evaluation?

The most useful parameters within the marine environment have been mentioned in the previous paragraphs, and their corresponding signifi cance has been highlighted. Even though, several of them deserve to be clearly pointed out as the principal triggers of new processes due to changing


scenarios. According to the review by Brierley and Kingsford (2009) the following parameters must be kept in mind:

- Temperature (seawater, air, ocean—atmosphere boundary layer) - Salinity - CO2 concentration and partial pressure - pH and alkalinity - Dissolved oxygen concentration - Sea level rise - Timing of plankton blooms - Strengthened stratifi cation/mixed layer depth

In this sense it might be included several brief comments might be included which could help to enlighten these points. For example, CO2 and ocean pH represent a great threat to many marine organisms and ecosystems (Doney et al. 2009). Over the past 200 years, the oceans have absorbed approximately half of the anthropogenically-generated CO2 and at present a further approximately 1 million tonnes of CO2 diffuses into the world ocean per hour (Joos and Spahni 2008). The rate of decreasing pH, 0.1 units in the last 200 years and an expected drop of 0.3 to 0.5 units by 2100, is more than 100 times as rapid as at any time over the past hundreds of millennia (IPCC 2007c). Rates of oceanic CO2 absorption vary regionally as a function of wind strength and temperature. Colder waters can accommodate more dissolved CO2 than warm waters and are, therefore, more prone to acidifi cation (Guinotte and Fabry 2008). One of the main impacts of ocean acidifi cation on marine life arises because of interactions between acidity and carbonate availability. A taxonomically diverse array of marine organisms, including tiny coccolithophores (a type of phytoplankton), pelagic and benthic mollusks, fi st-sized starfi sh and urchins, as well as massive corals, require calcium carbonate for their skeletons, and others have key carbonate rich structures (e.g., fi sh otoliths). All of these are likely to suffer as increasing acidity reduces carbonate availability, and impacts at the species level may cascade through to widespread community change (Hoegh-Guldberg 1999). At present, shallow waters are generally saturated with carbonate ions, but dissolution increases with depth (Orr et al. 2005).

Another well known confl ictive scenario in the marine environment is the occurrence of a reducing dissolved oxygen concentration one. Low oxygen concentrations transform compartments of the world ocean in inadequate habitats for most of the marine organisms. Oxygen solubility in seawater is a function of temperature, and O2 availability in the world ocean has been declining since the 1950s (García et al. 2005) as the ocean has warmed. Over a range from 0 to 15ºC, dissolved oxygen concentration in seawater is related approximately linearly to temperature, and will decline by about 6% per one degree rise (Brierley and Kingsford 2009).


Ongoing warming together with rising CO2 will see an expansion of low oxygen zones, perhaps by more than 50% of their present volume by the end of the century (Oschlies et al. 2008). These expansions will affect some of the world’s most productive regions in terms of fi sheries, so there could be economic as well as ecological consequences. Furthermore, coastal eutrophication resulting from increased riverine run-off of fertilizers and increases in sea level will bring further accumulations of particulate organic matter and increased microbial activity that consumes dissolved oxygen (Díaz and Rosenberg 2008). Mobile organisms are able to avoid low oxygen concentrations, but sedentary ones have little choice but to tolerate low oxygen concentrations or die. Those which are able to tolerate hypoxic conditions might, paradoxically, benefi t from reduced predation if predators are themselves excluded (Altieri 2008).

Bakun (2010) has reported that the available data series are short compared to the relevant time scales of variation. So, how can the multiple realizations of the controlling processes be assessed in order to confi dently identify the basic dynamics? Moreover, if one surrenders the assumption of system stationarity, how can one hope to parameterize any sort of predictive model? The same author states, “if one is confronted with a complex adaptive system, wishing (or pretending) it were otherwise may yield answers”. But these probably will not be the correct, comprehensive or useful answers that are needed.

Main Manifestations of Climate Change Effects on the Marine Environment

Ocean climate is variable and there have been warm periods previously, notably from the mid-1920s to the 1960s in the North Atlantic (Jensen 1939). The changes in species and ecosystems which took place then were very similar to those occurring now and we can use these past warm periods as analogues. However, these warm periods during the 20th century are examples of natural climate variability, whereas present climatic trends are expected to continue and conditions are moving outside the bounds of previous experience due to climate change. The trend will not be smooth and will continue to have large interannual and decadal variability superimposed on it (Sutton and Hodson 2005, Smith et al. 2007). Decadal variability in ocean climate is one of the major causes of regime shifts, when the biology of large areas such as the North Sea changes quite rapidly to a different state (altered species dominance, production and seasonality) (Beaugrand 2004). Understanding the way in which climate change may affect decadal and shorter time scale variability, is therefore essential in predicting future climate impacts on marine ecosystems and fi sheries.


Climate change and climate variability have occurred throughout history and natural systems have developed a capacity to adapt, which will help them to mitigate the impact of future changes. However, two factors will limit this adaptive capacity in future: (i) the rate of future climate change is predicted to be more rapid than previous natural changes; and, (ii) the resilience of species and systems is being compromised by concurrent pressures, including fi shing (Planque et al. 2010), loss of biodiversity (including genetic diversity), habitat destruction, pollution, introduced and invasive species and pathogens.

Recent fi ndings on climate change impacts on marine ecosystems and fi sheries can be divided into observational studies of past and current effects of climate change, and modelling studies of future impacts (Roesig et al. 2004). By the way, Brander (2010) presented some studies which included both observation and modelling, as well as some empirically based models (e.g. they use functional relationships which are statistically derived from observations).

There is now a wealth of evidence of impacts of recent climate change on distribution, species composition, seasonality and production in marine and freshwater systems. A very small selection from the large number of recent papers which analyze climate effects on a variety of taxa includes: phytoplankton (Richardson and Schoeman 2004), global primary production (Schmittner 2005), krill in the Southern Ocean (Atkinson et al. 2004), plankton in the North Atlantic (Richardson and Schoeman 2004), tropical tuna (Lehodey et al. 2003), sardine and anchovy in Eastern Boundary currents (Chavez et al. 2003) and fi sh species in North European shelf seas (Perry et al. 2005).

Several of the most important effects on the marine environment linked to Climate Changes which have been recognized include: (i) Changes in global marine primary production; (ii) Cascade effect of changes in primary and secondary production on future biological production; (iii) Regional consequences of changes in primary poduction; (iv) Regional changes in distribution and phenology of plankton and other organisms; (v) Spread of pathogens; among others. Brief comments on each of them are included in the following paragraphs.

Changes in global marine primary production

According to Brander (2010), three groups of factors govern the biological response: (i) warming, (ii) light, as determined by ice cover, cloudiness and surface mixed-layer thickness, and (iii) altered nutrient supply due to changed vertical stability and nutrient fl ux. Predicted climate induced


alterations in nutrient supply and production are predominantly negative, due to reduced vertical mixing. In high latitude regions the resultant increased stability of the water column may, however, have a positive effect on production in spite of reduced nutrient supply because phytoplankton will no longer be mixed down to depths greater than their compensation depth (the depth at which respiration loss exceeds photosynthetic gain) (Behrenfeld et al. 2006).

Sarmiento et al. (2005) have performed a comparative study on primary production using empirical models for a set of seven biomes (marginal sea ice; subpolar; subtropical seasonally stratifi ed; subtropical permanently stratifi ed; low latitude upwelling; tropical upwelling; tropical downwelling), which are further subdivided into a total of 33 biogeographical provinces resembling those of Longhurst (1998). A small global increase in marine chlorophyll and primary production is predicted (< 10%) for 2050 and 2090, compared with the pre-industrial control scenario, but with quite big regional differences. Decreases in the North Pacifi c and the area adjacent to the Antarctic continent are slightly more pronounced than offset by increases in the North Atlantic and the open Southern Ocean. The most robust part of the outcome is the change in biome areas, with reductions in the marginal sea–ice biome and increases in the permanently stratifi ed subtropical gyre biome.

Other critical factor in determining the change in primary production is temperature sensitivity of primary production for a given chlorophyll level. This in itself determines whether primary production increases or decreases at low latitudes, and whether there would be no change or quite large increases in primary production at high latitudes (Brander 2010).

Satellite observations of ocean chlorophyll indicate that global ocean annual primary production has declined by more than 6% since the early 1980s (Gregg et al. 2003). Global blended chlorophyll seasonal climatologies were used as inputs to the Vertically Generalized Production Model or VGPM (Behrenfeld and Falkowski 1997) to compute seasonal ocean primary production. Nearly 70% of the global decline occurred in the high latitudes. In the northern high latitudes, these reductions in primary production corresponded with increases in sea surface temperature and decreases in atmospheric iron deposition to the oceans, e.g., the processes involve both direct and indirect effects on nutrient supply. In the Antarctic, the reductions were accompanied by increased wind stress. It must be noted that these declines in primary production at high latitude have been offset by increases at low latitudes, and that three of the four low latitude basins exhibited decadal increases in annual primary production.


Cascade effect of changes in primary and secondary production on future biological production

Brander (2010) has reported that changes in primary and secondary production will obviously have a major effect on fi sheries production, but the complexity of the trophic systems leading from primary production to fi sh makes it diffi cult to establish reliable predictive relationships. Although global aggregated marine primary production is not expected to change substantially over the next 4 or 5 decades, there is a stronger basis for predicting changes in production at regional level and also good observational evidence, particularly for the North Pacific and North Atlantic (Jennings and Brander 2010). In both cases changes in production are driven mainly by regime-scale and event-scale (e.g., El Niño) changes (Brander 2010).

In the Arctic Ocean, the reduction in ice cover will allow light to penetrate in new areas and therefore increase the productive area, but the retreat of the highly productive marginal sea–ice zone will disrupt the existing food web (Santhi Pechsiri et al. 2010). In the “new” ice-free areas of the Arctic Ocean production is likely to be limited by nutrient supply due to the increased freshwater input from Arctic rivers (Prowse et al. 2006). Brander (2010) highlighted that this will increase vertical stratifi cation and hence reduce the vertical fl ux of nutrients. The riverine input is also nutrient poor.

Qualitative changes in production may have major impacts on food chains leading to fi sh regardless of changes in the absolute level of primary production (Jarre et al. 2006). In this sense, Atkinson et al. (2004) presented results including the observed switch from krill to salps as the major nektonic species in parts of the Antarctic, while Daskalov (2002) did with the ascendance of gelatinous species to a dominant position in areas such as the Black Sea. In the former case climate change was probably a major factor, but in the latter it was not (Brander 2010).

Regional consequences of changes in primary production

A large amount of information is available on this topic, and several of the most remarkable examples are as follows:

Tropical Pacifi c

Tuna species fi sheries (e.g., skipjack Katsuwonus pelamis, yellowfi n Thunnus albacares, albacore Thunnus alalunga) are the most important within the area, representing more than 3.5 million tons/year (Bigelow and Maunder 2007). The catches and distribution of these species are governed by variability in primary production and location of suitable habitat for spawning and


for adults, which in turn are linked to varying regimes of the principal climate indices El Niño–La Niña Southern Oscillation Index (SOI) and the related Pacifi c Decadal Oscillation (PDO). The tropical tuna species, skipjack and yellowfi n have higher recruitments during El Niño events, whereas the subtropical albacore has low recruitment during El Niño and high recruitment during La Niña (Brander 2010). Both statistical and coupled biogeochemical models have been developed to explore the causes of regional variability in catches and their connection with climate.

The model area includes the Pacifi c from 40°S to 60°N and includes the Kuroshio extension east of Japan (Taguchi et al. 2007). This is one of the best examples linking processes and scales from climate related upwelling and primary production to large geographic regions and decadal regime shifts. The model captures the slowdown of Pacifi c meridional overturning circulation and decrease of equatorial upwelling, which has caused primary production and biomass to decrease by about 10% since 1976–77 in the equatorial Pacifi c (McPhaden and Zhang 2002).

North Atlantic

Plankton samples collected between 1958 and 2002 showed an increase in phytoplankton abundance in the cooler regions of the Northeast Atlantic (north of 55°N) and a decrease in warmer regions (south of 50°N) (Richardson and Schoeman 2004). The likely explanation for this apparently contradictory result is that although both areas have undergone warming over this period, with consequent reduction of vertical mixing, the nutrient supply in the cooler, more turbulent regions remains suffi cient and plankton metabolic rates benefi t from the increased temperature. In the warmer regions reduced supply of upwelled nutrients limits production. The effects of these changes in phytoplankton propagate up through herbivores to carnivores in the plankton food web (bottom-up control), because of tight trophic coupling (Brander 2010).

Another study attributed the observed decadal variability in phytoplankton biomass in the Northeast Atlantic to climate forcing, as expressed by the NAO (Edwards et al. 2001). In the North Sea this resulted in a shift in seasonal timing of the peak in phytoplankton colour from April to June which may have been accompanied by a taxonomic shift from diatoms to dinofl agellates, with consequences for the food webs dependent on them (Brander et al. 2006).

Antarctic

Antarctic krill (Euphausia superba) is among the most abundant animal species on earth, providing the main food supply for fi sh, birds and whales. They have declined since 1976 in the high latitude SW Atlantic sector,


probably due to reduction in winter sea-ice extent around the western Antarctic Peninsula (Atkinson et al. 2004). Krill are dependent on the highly productive summer phytoplankton blooms in the area east of the Antarctic Peninsula and south of the Polar Front. Salps, by contrast, which occupy the extensive lower productivity regions of the Southern Ocean and tolerate warmer water than krill, have increased in abundance. These changes have had profound effects within the Southern Ocean food web. Penguins, albatrosses, seals and whales have wide foraging ranges but are prone to krill shortage. Thus the wide areal extent of change in krill density—not just its magnitude—is important (Brander 2010).

Regional changes in distribution and phenology of plankton and other organisms

There are many examples of distribution changes in marine ecosystems throughout the world (Beare et al. 2002, Beaugrand et al. 2003). The planktonic ecosystem is dependent on the properties and movement of the water. The life cycles of most marine fi sh have a planktonic phase, which often involves transport over long distances. The potential for rapid distribution change is therefore inherent, but requires favorable conditions for survival, particularly if the developing juveniles settle to the bottom (Altieri 2008).

Survival of fi sh larvae during the planktonic stage is thought to depend strongly on the availability of suffi cient suitable food (match-mismatch hypothesis) (Frank and Leggett 1982, Stenseth et al. 2002). Therefore, in addition to effects of changes in production, described in the previous section, climate induced changes in distribution and phenology of fi sh larvae and their prey can also affect recruitment and production of fi sh stocks (Durant et al. 2007, Rijnsdorp et al. 2009).

Spread of pathogens

Pathogens have been implicated in mass mortalities of many aquatic species, including plants, fi sh, corals, and mammals, but lack of standard epidemiological data and information on pathogens generally makes it diffi cult to attribute causes (Harvell et al. 1999). An exception is the northward spread of two protozoan parasites (Perkinsus marinus and Haplosporidium nelsoni) from the Gulf of Mexico to Delaware Bay and further north, where they have caused mass mortalities of Eastern oysters (Crassostrea virginica). Winter temperatures consistently lower than 3ºC limit the development of the MSX disease caused by Perkinsus (Hofmann et al. 2001), and the poleward spread of this and other pathogens can be


expected to continue as such winter temperatures continue becoming rarer. This example also illustrates the relevance of seasonal information when considering the effects of climate change, since in this case it is winter temperature which controls the spread of the pathogen (Brander 2010).

Finally, the impact of climate changes on the ecosystem services must also be considered. It is well recognized that services derived from ecosystems are essential to human welfare (Dobson et al. 2006, Halpern et al. 2007) and could be critically affected through climate change (Menzel et al. 2006, Paterson et al. 2009).

Smith et al. (2001) synthesized all possible information on climate-change impacts to evaluate which impact level would constitute a dangerous climate change. They used global mean temperature increase (GMTI) in 2100, which is a widely accepted climate change indicator, and determined related risk levels (low, medium and high) for 5 different ‘Reasons for concern’ characterized by specifi c entities (e.g., unique and rare species, extreme events, regional distribution, aggregated impacts and large-scale singularities). These large-scale impacts on species, landscapes, ecosystems and many of the services they provide (e.g., water purifi cation, slope stabilization, carbon sequestration and many cultural and aesthetical values) are mostly non-market impacts. Even though the value of specifi c ecosystem services (e.g., crops and timber) can be estimated in dollars (e.g. Balmford et al. 2002), we believe that the actual damages or benefi ts of changes in species, ecosystems and landscapes are not satisfactorily characterized in monetary terms.

Oceanic and coastal areas provide important environmental goods and services to the human population, such as food production, fi ltration and cleaning of fresh waters, the shoreline stabilization, regulation of the hydrological regime, dioxide carbon storage and oxygen production, and many more. They also have a tremendous biological richness: of the 82 recognized phyla, 60 include marine representatives, when exclusively animals are considered, 36 of the 37 recognized phyla are present in ecosystems from the oceans and coastal areas, ranging from coral reefs to seagrass communities, mangroves, coastal lagoons and estuaries (Rohde 1992, May and Godfrey 1994).

Doney et al. 2012 made a full summary of these phenomena. So, disruptions of existing biological interactions can occur through asynchronous shifts in the seasonal phenologies of interacting predator and prey populations (e.g., the match-mismatch hypothesis); biogeographic reorganizations, leading to changes in community composition and biodiversity; and, loss of functionally prominent species (Forrest and Miller-Rushing 2010). Furthermore, and following Doney et al. (2012), these processes can be expressed through bottom-up impacts such as declines in water-column primary production (e.g., O’Reilly et al. 2003) and/or shifts


toward smaller cells in planktonic communities (Winder et al. 2009), as well as through top-down impacts that cascade down from the losses or gains of ecologically dominant consumers (Schmitz et al. 2000). Alterations in biogeochemical cycling can occur because of the replacement of functional groups (e.g., calcifi ers) even if overall productivity and diversity remain approximately constant.

Climate-driven impacts on keystone and foundation species may be especially important. Some critical habitat-forming marine benthic species, such as oysters or corals, appear sensitive to CO2 and climate change both directly and through pathogens. As has been previously mentioned, in oyster populations within Delaware Bay (USA) the protistan parasite Perkinsus marinus (which causes the disease Dermo) proliferates at high water temperatures and high salinities, and epidemics followed extended periods of warm winter weather; these trends in time are mirrored by the northward spread of Dermo up the eastern seaboard as temperatures warmed (Cook et al. 1998). Similarly, corals on the Great Barrier Reef showed more infections by the emerging disease “white syndrome” in warmer than normal years (Bruno et al. 2003). These processes and others resulting from altered species composition will likely have important rippling affects through ecosystems.

In addition, climate change and altered ocean circulation may change organism dispersal and the transport of nutrient and organic matter that provide important connectivity across marine ecosystems (Walther et al. 2002). If species dispersal is disrupted by climate-induced thermal blocks or shifts in currents carrying larvae, both species and community dynamics will be altered (Parmesan 2006).

As a brief summary it can be commented that the Earth’s climate has changed throughout history, showing different processes, effects and consequences at different times (Pearson and Dawson 2003). Just in the last 650,000 years there have been seven cycles of glacial advance and retreat, with the abrupt end of the last ice age about 7000 years ago marking the beginning of the modern climate era—and of human civilization (VijayaVenkataRaman et al. 2012). Most of these climate changes are attributed to very small variations in Earth’s orbit that change the amount of solar energy our planet receives (Bard and Frank 2006).

The evidence for rapid climate change (IPCC Fourth Assessment Report) is compelling: sea level rise; global temperature rise; warming oceans; shrinking ice sheets; declining Arctic sea ice; glacial retreat; ocean acidifi cation (IPCC 2007a).


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CHAPTER 2

Coastal Marine Biodiversity Challenges and Threats

Jerónimo Pan,1,2,* M. Alejandra Marcoval,1,2 Sergio M. Bazzini,2,3,5 Micaela V. Vallina2,3,6 and

Silvia G. De Marco3,4

Global Change and Coastal Marine Ecosystems

Global change, a term that adequately fi ts into the focus of this book, is broader than climate change and comprises the major anthropogenic forcings that produced a signifi cant change or impact on the natural environment during the last ~ 200 years. Global change issues have to be addressed with a planetary perspective, and as part of a time continuum, running from a few centuries ago, increasing its rates in the present and with implications in the near future. It is usually said that global change is an unprecedented human experiment on the planet, and as any experiment its consequences and reaches are to a certain degree, unpredictable.

Marine systems are highly responsive to alterations in the physical environment (particularly those with decadal scales), and also highly adaptable to such changes (Steele 1998), which makes it diffi cult to defi ne

1 Departamento de Ciencias Marinas & Estación Costera «J. J. Nágera», Facultad de Ciencias Exactas y Naturales (FCEyN), Universidad Nacional de Mar del Plata (UNMdP), Argentina.

2 Consejo Nacional de Investigaciones Científi cas y Técnicas (CONICET), Argentina.3 Departamento de Biología, FCEyN, UNMdP.4 Facultad de Ingeniería, Universidad FASTA, Mar del Plata, Argentina.5 Instituto de Geología de Costas y del Cuaternario, FCEyN, UNMdP.6 Instituto de Investigaciones Marinas y Costeras (IIMyC), FCEyN, UNMdP.* Corresponding author: [email protected]


a baseline state for marine ecosystems upon which to evidence changes by comparison. This chapter will narrow down to the discussion on some of the anthropogenic disturbances that impact coastal marine ecosystems. The coastal zone is the region that runs from the inner shelf to the coastline, usually defi ned as an interface where the land meets the ocean, encompassing shoreline environments as well as adjacent coastal waters (Post and Lundin 1996). It is a geologically-young and dynamic area with changing biological, chemical and geological attributes. Coastal ecosystems are highly productive and biologically diverse and can act to moderate the impacts of pollution originating from land (Post and Lundin 1996).

Coastal areas provide critical ecological services such as nutrient cycling, fl ood control, shoreline stability, beach replenishment and genetic resources (Post and Lundin 1996, Scavia et al. 2002). Some estimates by Boesch (1999), mention that the ocean and coastal systems contribute 63% of the total value of Earth’s ecosystem services (worth $21 trillion year–1). Population growth is a major concern for coastal areas with more than 50% of the world population concentrated within 60 km of the coast (Post and Lundin 1996); in the United States the expected tendency for the next decades is that the coastal population will increase by ~ 25% (Scavia et al. 2002). The continued growth of human population and of per capita consumption have resulted in unsustainable exploitation of Earth’s biological diversity, exacerbated by climate change, ocean acidifi cation, and other anthropogenic environmental impacts. The effective conservation of biodiversity is essential for human survival and the maintenance of ecosystem processes.

Our discussion of coastal zone impacts will necessarily be diverse (e.g., coastal zones vary from place to place), fragmentary (e.g., not all coastal regions have been studied) and incomplete (e.g., not all coastal regions have been studied with the same scope and extent). Also important to bear in mind is that even though we will try to focalize on the impacts to coastal locations, the distinction is not real or absolute, since coastal ecosystems are not separate compartments of the world’s oceans and no boundaries exist. In fact they are better defi ned by their interconnections with other regions of the ocean, and are better interpreted as open-sided systems (Steele 1998).

Climate change per se is not the foremost pervasive human disturbance to coastal ecosystems, nor is it usually the original cause of major impacts, but its effects constitute a major concern for coastal ecosystems in the long run (Jackson et al. 2001). In that sense, it is important to remember that more direct human disturbances do not have isolated or punctual impacts, but they rather combine in a synergistic fashion to amplify their pervasive effects to coastal ecosystems. Jackson et al. (2001) proposed a sequence of historical events of disturbances for coastal ecosystems, in which overfi shing precedes other phenomena like pollution, eutrophication, outbreak of disease or climate change. In their scheme, climate change constitutes a

Coastal Marine Biodiversity Challenges and Threats 45

fairly recent major impact to coastal ecosystems, whose effects are most remarkable due to the synergy established with previous impacts that have been taking place for longer historical periods. In our discussion, we will demonstrate that the enchainment of these factors produces major impacts on coastal ecosystems.

Biodiversity: Defi nitions and Conceptual Approaches

Biological diversity or biodiversity (a term introduced a few decades ago) comprises the variety of life on Earth, from genes and organisms to larger units such as ecosystems and landscapes. This concept not only encompasses the biotic components of ecosystems, but also makes reference to specifi c temporal and spatial dimensions and the complex species interactions that arise as a product of natural selection, adaptation and other evolutionary processes.

As much as a quantifi cation of the intrinsic value of biodiversity would contribute to the argumentation for its conservation, assigning a value to biodiversity is a diffi cult task and it necessarily implies a value assessment from a multidimensional perspective (González Barberá and López Bermúdez 2000). Numerous studies indicate the fundamental role of biodiversity in the modulation of ecosystem functioning and stability (Emmerson et al. 2001, Singh 2002, Hooper et al. 2005).

The ecosystem approach is a perspective that emerged from 1992 Rio’s Earth Summit, which deals with ecosystem functioning and management from a holistic perspective (Beaumont et al. 2007). Biodiversity clearly affects the way ecosystems function (Hooper et al. 2005), and in that sense, identifying ecosystem services facilitates the incorporation of biodiversity into management discussions and planning. Simply put, such services are the benefi ts that humans obtain from ecosystems. These include provisioning services such as food and water; regulating services that affect climate, fl oods, and water quality; cultural services that provide recreational, aesthetic, and spiritual benefi ts; and supporting services such as soil formation, primary productivity, and nutrient cycling. Most, if not all, human endeavors are directly or indirectly dependent upon ecosystem services (Hooper et al. 2005, UN Millennium Ecosystem Assessment 2007).

Despite some successful conservation efforts (mostly at local scales) biodiversity continues to decline (Rands et al. 2010). Marine ecosystems are not exempt to this global trend, and despite their adaptive nature, they are vulnerable to rapid changes in diversity and function (Palumbi et al. 2008). To name a few examples, the increasing pressure on marine biodiversity is evidenced by the steep declines in fi sh populations and loss of marine habitats that resulted from overexploitation of fi sh protein from the top of the food chain; poorly managed aquaculture practices;


and direct habitat destruction from coastal development and pollution (Allsopp et al. 2009). Biodiversity also faces new pressures and threats in the form of anthropogenic climate change. Climate change forces species to shift their ranges and disrupts ecological communities (Lemoine and Böhning-Gaese 2003).

With this framework in mind, we will now consider some of the consequences of climate change on the physical and chemical properties of coastal environments to later discuss how and to what extent these changes affect biotic components of the ecosystem.

Forces of Global Change on Coastal Environments

Nutrient enrichment

Nutrient enrichment is defi ned as the addition of inorganic or organic N and P carried from land through river runoff or sewage inputs (sometimes also from aquaculture practices). It is a relatively recent phenomenon that began to get noticeable in coastal waters and experienced a most remarkable increase from the 1960s to the 1980s, probably linked to a dramatic increase in agriculture, changes in land use and the introduction of industrial fertilizers in agricultural practices (Boesch 2002). It is mostly a phenomenon that has impacts on coastal waters (estuaries, embayments and semi-enclosed seas) of developed countries in the Northern Hemisphere, even though some smaller-scale enrichments are also observed in coastal areas of developing countries. Global change will likely infl uence the vulnerability of estuaries and other semi-enclosed coastal environments to eutrophication (Scavia et al. 2002) by introducing changes in mixing characteristics and the exchange with the ocean, altering freshwater runoff, changing surface temperature and rising sea level.

The addition of fi xed N and P triggers a series of phenomena, including increased primary production, decrease in water clarity, alteration of food chains and the occurrence of harmful algal blooms with increased frequency (Boesch 2002, Martin and LeGresley 2008). Some algal species that are not normally toxic may become so when exposed to altered nutrient regimes from over-enrichment (Burkholder 1998). Not only the increment in nutrients matters, but rather the changes in the ratios produced by differential additions is what determines the resource selection by different groups of primary producers (Cloern 2001). A selective enhancement in the loadings of N and P but not silicon has been taking place in coastal waters; increasing N usually shows decreasing trends in the Si:N ratios, in that way, the occurrence of diatom-dominated blooms give way to small phytofl agellates and dinofl agellates involved in the production of toxins or harmful in other ways (Cloern 2001). Turner et al. (1998) documented


silicate:nitrate ratio fl uctuations over the past 40 years in coastal plankton food webs through the analysis of the variations registered in silicate:nitrate ratios.

The approach to the problem of eutrophication is becoming more holistic. Current hypotheses include questions such as the interaction of nutrient enrichment with other stressors (e.g., contaminants, introduced species, habitat loss, hydrologic manipulation, regional climate change), but most importantly, scientists have started to delve into how the responses to multiple stressors are linked (Cloern 2001).

Potential effects of rising water temperature

Rising global temperature has been both a topic of intense study, and a debate in the media. It is not surprising that, when issues related to global change started to be treated in the press, they were usually referred to as “global warming”, probably because the increase in temperature and the associated rise in sea level are perceived as very close and “tangible” threats by the public. The 20th century has been the warmest in historical records, with the 1990s being the warmest decade of the millennium in the Northern Hemisphere (Trenberth et al. 2007). The global ocean has warmed signifi cantly since the late 1940s and more than half of the increase in heat content has occurred in the upper 300 m since the late 1950s (Bindoff et al. 2007). The mean surface temperature has increased by 0.6 ± 0.2ºC during the 20th century (Gitay et al. 2002).

An increase in surface water temperature is likely to affect most metabolic rates of marine organisms and be translated into signifi cant changes in biological processes and biodiversity (Hall 2002). It is not possible to make a valid generalization in this respect, but it is reasonable to expect an increase in the occurrence of physiological stress in organisms and disease outbreaks (Scavia et al. 2002, Roessig et al. 2004). The effects of temperature increase on coastal organisms are summarized in Hiscock et al. (2004); these authors point that environmental temperature might especially have an indirect infl uence on populations, acting on reproductive processes (e.g., development of gonads, release of propagules, survival and settlement of larval stages).

Other than that, an increase in temperature has several enchained effects on the physical and chemical properties of the environment. For shallow coastal waters thermal stratifi cation combined with nutrient enrichment can lead to the occurrence of hypoxia (i.e., a defi cit in dissolved oxygen). There seems to be a series of stages linking the presence of excessive decomposing of organic matter, stratifi cation and the development of hypoxia and anoxia. Excess nutrients lead to increased primary production and accumulation of excess organic matter in the bottom, which reduces oxygen levels; when


combined with water column stratifi cation (which prevents the exchange of bottom water with oxygen-rich surface water) hypoxic conditions arise. In that way, the increasing input of nutrients to coastal areas has been suggested as the main contributor to declining bottom water oxygen concentrations (Diaz 2001). It is expected that global warming of coastal waters accelerates this series of events and enlarges current areas with hypoxic conditions.

A near-future scenario with increased (sea and atmospheric) temperatures will probably lead to signifi cant changes in surface current patterns, which in turn will lead to shifts in the geographic distribution of coastal organisms. It has been extensively documented that coastal organisms make use of currents as means of dispersal. Therefore, changes in the distribution and dispersal patterns of organisms are to be expected. On the other hand, there may be changes in species abundance near the limits of their current distribution (Thompson et al. 2002). A note of caution is needed here, since range expansions may not be as rapid as range contractions because the former require numerous factors acting in conjunction for successful establishment (e.g., transport, absence of predation, low incidence of disease; Hall 2002).

Several variables are to be considered in order to determine future distributions. Hiscock et al. (2004) tried to assess future scenarios of changing environmental parameters, and how those deviations from present conditions might impact the distribution of coastal organisms. For instance, these authors made projections for distributions and relative abundances of subtidal and intertidal benthic invertebrates and macroalgae from coastal waters around Britain and Ireland assuming an estimated 2.1ºC increase in inshore sea temperature by the 2050s. They paid particular attention to dominant or key species (in structural and functional terms), and how changes in these species might affect other members of the ecosystem.

Effects of ozone depletion and increased UV radiation fl uxes

The increase in atmospheric greenhouse gases has caused a depletion in stratospheric ozone in recent years, which has resulted in an increased fl ux of ultraviolet radiation (UVR), a natural component of solar radiation, to the Earth’s surface (McKenzie et al. 2010). Paradoxically, the anthropogenic emissions of greenhouse gases that tend to cause a temperature increase at the Earth’s surface also produce a decrease in stratospheric temperatures. The decrease in stratospheric temperatures leads to enhanced formation of polar stratospheric clouds and may serve to increase ozone loss in Polar regions (Shindell et al. 1998). This results in a greater change in UVR fl uxes in Polar and high-latitude regions, which are more susceptible to the formation of an “ozone hole” during spring. Increased UVR represents a


relatively new problem to marine organisms (Whitehead et al. 2000, Häder et al. 2010). In that sense, the recognition of the Antarctic ozone hole, initially prompted considerable research on the effects of UVR over phytoplankton communities, owing to their importance in primary production. Subsequent investigations were extended to the microbial loop and higher trophic levels.

The role of UVR as an environmental stressor has been demonstrated for aquatic animals including corals, zooplankton and fi sh (Häder et al. 2007, 2010). A meta-analysis revealed negative effects of ambient UV-B on growth and survival of a range of aquatic organisms (Bancroft et al. 2007).

Increasing atmospheric CO2 levels and ocean acidifi cation

Atmospheric levels of carbon dioxide (CO2) have steadily increased through anthropic sources (Forster et al. 2007). This excessive atmospheric CO2 is uptaken by the world’s oceans to maintain the balance of the carbonate buffer system (Libes 1992). In this naturally-occurring process, atmospheric CO2 readily dissolves into seawater; dissolved CO2 reacts with water to produce carbonic acid (H2CO3). In turn, carbonic acid dissociates into H+ and bicarbonate (HCO3

–) ions. Bicarbonate further dissociates into more H+ and carbonate (CO3

=) ions. However, the recent uptake of CO2 is too rapid for the supply of CO3

= ions, and therefore, H+ and bicarbonate levels are increasing, while carbonate levels are decreasing, with the ultimate result of an increased acidity of ocean waters at a global scale, a phenomenon termed as ocean acidifi cation (Orr et al. 2005). Ocean acidifi cation (OA) is a global threat to marine ecosystems and its long term implications for the diversity of marine organisms and ecosystem functions are diffi cult to predict (Doney et al. 2009).

There is some experimental evidence that the severity of the impacts of OA could be dependent upon factors related to an organism’s lifestyle and activity (e.g., infaunal vs. epifaunal, deep vs. shallow, and deposit feeder vs. suspension feeder) rather than to its phylogeny (Widdicombe and Spicer 2008). There is also uncertainty over the extent to which organismal adaptation or acclimation might mitigate the long term effects of OA.

A recent meta-analysis quantifying the variability of biological responses of marine organisms to OA, revealed a strong negative effect on calcifi cation and growth, despite the variability in the sensitivity of taxonomic groups and developmental stages (Kroeker et al. 2010). However, differential sensitivities may still have important implications for marine ecosystems in those cases where individual species play disproportionately strong roles in structuring communities (Shurin et al. 2002, Borer et al. 2005). Additionally, differential sensitivities will infl uence species interactions and


could drive unforeseen impacts on marine communities and ecosystems (Kroeker et al. 2010).

A sound knowledge of organismal biology and physiological mechanisms can help elucidate the larger ecosystem changes that arise in response to climate change forcings, such as population collapses or local extinctions (Pörtner and Knust 2007), disruptions in large-scale animal migrations (Farrell et al. 2008), changes in phenology (Wiltshire and Manly 2004), and changes in food availability and food web structure (Pörtner and Knust 2007, Farrell et al. 2008). All organisms live within a range of optimal body temperatures and climate change will differentially favor species with wide thermal windows, short generation times, and genotypic variability among their populations. The specifi c effects of CO2, hypoxia, salinity change, and eutrophication reduce the overall fi tness of organisms, especially at extreme temperatures, therefore narrowing thermal windows and biogeographic ranges (Pörtner and Farrell 2008).

Laboratory experiments demonstrated that declining pH can negatively impact calcifi cation in marine organisms like corals, mollusks, coralline algae and phytoplankton (Kleypas et al. 1999, Riebesell et al. 2007). However, results from laboratory experiments can be diffi cult to extrapolate to ecosystem responses because pH may affect other aspects of a species biology besides calcifi cation, and also because interspecifi c relationships can enhance or counteract the effects of environmental impacts (Schindler et al. 1985, Hall-Spencer et al. 2008). In that sense, modeling approaches provide a means of linking changes in environmental parameters with the in situ dynamics of complex ecosystems and then predict long-term impacts on community structure. Using such approach Wootton et al. (2008) demonstrated that coastal ocean pH is unexpectedly dynamic given the high buffer capacity of oceans, and revealed strong links between in situ benthic species dynamics and variation in ocean pH, with calcareous species generally performing more poorly than non-calcareous species in years with low pH.

In the face of OA, there is considerable interest in understanding how the loss of species and the alteration of communities will affect ecosystem function. A study by Kroeker et al. (2011) examined how changes of multispecies assemblages of marine invertebrates were affected by OA with respect to community composition and structure, density compensation among taxa, and aggregate biomass and trophic structure. They found divergent and compensatory responses of marine invertebrates to OA, and concluded that these do not offset the effects on aggregate biomass or trophic structure, suggesting that acidifi cation will likely affect ecosystem function and the services they provide (Kroeker et al. 2011).


Regime shifts

Regime shifts arise when a combination of climatic, biological and physical changes lead to persistent new sets of ecosystemic characteristics that represent deviations or shifts from the historic record. Even though regime shifts have an extensive record in geologic time, the temporal and spatial scales at which these shifts have occurred recently is what concerns scientists the most. For instance changes in precipitation frequency and intensity, ocean acidifi cation, water temperature increase, changing wind patterns, hydrology fl uctuations and alterations, combined with anthropogenic pollution by nutrients and toxins, all can affect water quality in estuarine and coastal waters (Hall 2002, Gitay et al. 2002).

It has been demonstrated that for the past 20 to 30 years (in comparison to the previous 100 years), El Niño-Southern Oscillation (ENSO) events have increased their frequency, persistence and intensity (Gitay et al. 2002). This basin-scale phenomenon has well-documented effects not only on coastal regions, but effect on several teleconnections to distant areas on land and on other ocean basins have also been established.

Global Change Effects on Biotic Components of Coastal Ecosystems

Invasive species in coastal marine ecosystems

Non-indigenous species (NIS; synonyms: alien, exotic, non-native, allochthonous) are species or lower taxa introduced outside of their (past or present) natural range and beyond their natural dispersal potential. This includes any (vegetative or reproductive) structure that might survive and subsequently reproduce (Council Regulation 2007). Their presence in a new system is due to intentional or unintentional introduction resulting from human activities, and might have implied various pathways and/or vectors. Invasive alien species (IAS) are a subset of established NIS with the potential or actual ability to spread elsewhere, and have an adverse effect on biological diversity, ecosystem functioning, socio-economic values and/or human health in invaded regions. Species of unknown origin which cannot be ascribed as being native or alien are termed cryptogenic species (sensu Carlton 1996). Biological pollution is defi ned as the adverse impacts that IAS can cause on one or more levels of biological organization (Elliott 2003, Olenin et al. 2010).

IAS represent an increasing problem in marine coastal waters (Olenin et al. 2011). In contrast to more enclosed water bodies, the openness of marine systems indicates that once a species is in an area, then eradication is usually impossible. When the number of species involved in the pathway


is greater than the number of species which managed to survive transport and establish a population, then a bioinvasion occurs. Therefore, marine biological invasions are increasingly altering coastal biota, generating changes in the chemical and/or physical properties of an ecosystem, ecosystem functioning and ultimately result in adverse effects on economy and human health (Convention on Biological Diversity 2004, Lodge et al. 2006, European Commission 2008, Nunes and Markandya 2008, Pyšek and Richardson 2010). Even when it is very diffi cult to predict which non-indigenous species (NIS) may result in detrimental effects on environmental quality, not all NIS will necessarily cause harm to the environment (Olenin et al. 2011).

The fi rst requirement for NIS to become successfully established is to have the physiological ability to survive in a new environment (Ignacio et al. 2012). Besides survival, resource acquirement capability is another major factor affecting the establishment, range expansion and invasiveness potential of a NIS (Shea and Chesson 2002). The ecological changes that arise as a consequence of introduced species have been established as one of the most serious environmental concerns nowadays (Pederson and Blakeslee 2008). Marine bioinvaders have sometimes been overlooked and are certainly less well studied and documented than terrestrial and fresh-water invasions, mainly because marine organisms are less conspicuous, and not as easily sampled.

The occurrence of marine bioinvaders in advective environments (Byers and Pringle 2008), and other characteristics linked to their life histories such as open spawning, pelagic larval stages and large larval output, make them successful and persistent over time. Moreover, basin-scale physical events such as ENSO (Yamada and Gillespie 2008), and global climate change phenomena may collaborate in spreading native and non-native species (Cordell et al. 2008) with unknown resultant impacts (Pederson and Blakeslee 2008). Human activities such as commercial shipping and recreational boating, coastal urbanization and mariculture not only offer new transport opportunities for NIS (Carlton 1996), but also new substrates for colonization (Bulleri and Chapman 2010, Farrapeira 2011, Gittenberger and Stelt 2011), and their spread (Glasby et al. 2007, Tyrrell and Byers 2007).

Biological invasions by NIS are currently spread to several coastal areas worldwide. A comprehensive review of these is beyond the scope of this chapter, but we have provided a few examples from the Northern and Southern Hemispheres, and planktonic and benthic realms.

Initial records of seaweed introductions for the Northwest Atlantic date back to the late 1800s to the early 1900s. Currently, the number of introduced seaweeds is increasing worldwide, counting ~ 120 taxa, some of which aggressively dominate marine habitats (Mathieson et al. 2008).


Seemingly similar habitats may be differently impacted by invaders. For example, manipulating the densities of the non-indigenous snail Littorina littorea in two Gulf of Maine salt-marshes with diverging physical conditions (e.g., with reference to inundation, elevation, drainage and sediment characteristics), Tyrrell et al. (2008) found that stressing environments for Spartina alternifl ora favored grazing by the snail, resulting in declined cordgrass productivity.

Cordell et al. (2008) reported the introduction of nine Asian calanoid and cyclopoid copepods into the Northeast Pacifi c, some of which moved upstream and invaded the Columbia-Snake River system (USA), illustrating the potential of some species to colonize new areas, against environmental gradients (see Byers and Pringle 2008).

A study by Orensanz et al. (2002), presented the fi rst exhaustive review of marine NIS for the SW Atlantic Ocean (particularly Argentina and Uruguay). Considered at a regional scale, these authors cited the occurrence of 31 introduced and 46 cryptogenic species in coastal and shelf areas. Some of the latter are currently considered as IAS (S. Obenat, pers. comm.). It is noteworthy that some of the introductions are relatively recent (~ 30 years), but nonetheless showed striking ecological impacts in the area. For example, the barnacle Balanus glandula developed calcareous belts on rocky intertidals; Limnoperna fortunei (a macrofouling bivalve) and Ficopomatus enigmaticus (a reef-building polychaete) strongly modifi ed estuarine ecosystems; while the Pacifi c oyster Crassostrea gigas established well-developed reefs which rapidly expanded along shallow confi ned bays. The case of the Asian kelp Undaria pinnatifi da is remarkable. This phaeophyte modifi ed the benthic communities and signifi cantly changed the seascape of the Patagonian coasts where it became established, within a few years. Its fi rst record dates back to the mid-1990s (Piriz and Casas 1994). A decade ago, Orensanz et al. (2002) reported its distributional range from northern Peninsula Valdés (42°05’S) to Camarones (44°48’S). However it has recently been recorded in Mar del Plata harbor (38º02’S; Meretta et al. 2012), expanding its range in 4 degrees of latitude in 10 years. The above-mentioned F. enigmaticus also deserves a special mention for the area. This highly successful invader currently dominates the benthos of Mar Chiquita coastal lagoon, a UNESCO MAB Reserve. Within this lagoon, the polychaete builds calcareous reefs that range in size from a few cm to ~ 7 m in diameter (Schwindt and Iribarne 2000), and accordingly it has impacted the system by modifying circulation patterns and sedimentation rates (Schwindt et al. 2001, 2004). By becoming the dominant suspension feeder, it also plays a signifi cant role in benthic-pelagic coupling, with the capacity to drive changes in plankton community structure through selective grazing (Pan and Marcoval 2012).

Summarily, IAS represent one of the primary threats to biodiversity, and the risks of potential invasions may be increasing due to increasing global


trade, tourism transport and climate change (Convention on Biological Diversity 2004). Several international organizations and programs have established guidelines for the management of IAS and have urged the different parties to promote and implement them. In that sense, there are several options for managing marine invaders, ranging from comprehensive agency-based programs to focused approaches that target vectors and specifi c activities for eradicating species. Management actions are generally based on determination of the economic costs and benefi ts of the eradication (Pederson and Blakeslee 2008). Risk analysis or organism impact assessment, is a management tool that is becoming increasingly common in biosecurity (Campbell 2008). The method incorporates the ecological, cultural, social, and economic impacts of a target introduced species and aids in the management decision-making process by establishing a relative risk ranking. Other models have been used to predict the vulnerability of marine and estuarine ecosystems to invasion by NIS over several spatial scales (e.g., from a whole estuary scale down to a habitat-within-an-estuary scale; Reusser and Lee 2008). However, and regardless of the different models and management strategies available, certain gaps and inconsistencies in the international regulatory framework still persist, resulting in an ineffi cient management of IAS (Convention on Biological Diversity 2004).

Habitat loss

The extent of structurally complex marine habitats is gradually decreasing at local, regional and global scales (Suchanek 1994, Duarte 2002, Thrush and Dayton 2002, Reise 2005, Lotze et al. 2006). Habitat (i.e., the predominant features that create structural complexity in the environment) loss (i.e., the measurable reduction in habitat abundance and distribution; sensu Airoldi and Beck 2007) has been pointed out as one of the major threats to marine biodiversity (Beatley 1991, Gray 1997). However, this has not been a core topic in marine science and conservation (Airoldi et al. 2008).

Coastal communities are subject to increasing pressure from multiple anthropic stressors resulting in habitat loss (Lotze et al. 2006, Hoegh-Guldberg et al. 2007, Halpern et al. 2008, Fraschetti et al. 2011). Coastal urbanization, the dredging, fi lling and isolation of salt-marshes, eutrophication and decreasing water quality, are among the human activities that produce dramatic changes in marine coastal areas. Habitat destruction is bound to be a major problem for coastal wetlands, estuaries and intertidal environments in the near future (Hall 2002). For instance, some estimates presented in Gitay et al. (2002) indicate that if the global trend in sea level rise for the 20th century (a 1 to 2 mm average annual rate) is to be maintained, by the year 2080 ~ 20% of coastal wetlands could be


lost due to the combined actions of sea-level rise, more intense monsoonal rains and larger tidal or storm surges.

Airoldi et al. (2008) propose three major categories of habitat loss:

1) loss of native resident species; particularly those ecosystem engineers that have narrow distributions or are exclusive to certain habitats (Thrush et al. 2006).

2) loss of food resources; habitats dominated by primary producers (e.g., seagrass meadows, salt-marshes, kelp forests) are highly productive (Duffy 2006, Hosack et al. 2006) and responsible for the exportation of signifi cant amounts of C, N and P to adjacent coastal areas (Graham 2004); the loss of these food sources may have a direct or indirect negative effect on the productivity of the whole system (Dobson et al. 2006).

3) loss of environmental complexity and related ecosystem functions; the loss of habitat complexity carries a loss of numerous functions that shape the physical environment (e.g., light conditions, hydrodynamics, sedimentation, attenuation of disturbance; Jones et al. 1994, Dobson et al. 2006).

A 15-year study of the irreversible environmental consequences of unplanned coastal development in the Mediterranean Sea, offers a synthetic example of what has been previously exposed (Fraschetti et al. 2011). Habitat fragmentation and human-induced changes in sedimentation, ultimately resulted in a loss of > 50% of seagrass beds (Posidonia oceanica), a decline in macroalgal cover (Cystoseira spp.) and a loss in associated faunal assemblages, which impacted negatively on the goods and services provided for local human population.

Ecological shifts in phytoplankton and harmful algae

Ecological shifts in primary producers for coastal ecosystems have a long historic record, spanning ~ 200 years. An often cited example is that of Chesapeake Bay (Jackson et al. 2001) in which there is a record of a gradual shift in the organisms responsible for primary productivity since the 18th century (the trend is characterized by the decline of seagrasses and benthic diatoms, to give way to planktonic diatoms and other phytoplankton).

An increase in sea surface temperature in the surface mixed layer of the oceans is bound to alter circulation and increase density stratifi cation (Bindoff et al. 2007). Increased stratifi cation in the coastal ocean will have an effect in the dominance of groups of phytoplankton (Huisman et al. 2004), likely favoring motile species such as dinofl agellates (Peperzak 2003, Peperzak 2005), over dominant components of the early spring bloom (such


as diatoms). Such shifts in the plankton can produce a trophic mismatch in the food web (Edwards and Richardson 2004).

The occurrence of harmful algae in coastal waters is of most importance from several points of view. They impact coastal fi sheries and are known for their detrimental effects on human health (approximately 10% of all foodborne disease outbreaks in the US result from the consumption of poisoned seafood; van Dolah 2000). Other than the very obvious consequences from poisoning or sub-lethal effects, harmful algae can have more subtle infl uences on fi sh populations, such as habitat quality deterioration and may exert changes in the food web structure (Burkholder 1998).

In the past ~ 35 years, harmful or toxic algal incidents have increased in frequency and geographical extent. Even though paleontological evidence indicates that red tide-producing dinofl agellates occurred throughout the Holocene with characteristic periodicities, the variability and concurrence of species in the past 60 years is unmatched in the past and suggests an ecosystemic disequilibrium (Mudie et al. 2002). Such observations have led researchers to agree that human impacts on the environment are responsible for the increase in frequency and expansion of harmful algae (van Dolah 2000).

The increased geographical extent of harmful algal blooms, can be explained by unintentional introductions by ballast water and fouling, and the transport of stocks for aquaculture (Burkholder 1998, Thompson et al. 2002). As an example, Martin and LeGresley (2008) reported that together with the expansion of salmonid aquaculture industry in the Bay of Fundy, harmful algal blooms increased in intensity, frequency and geographic distribution. Additionally, a number of non-indigenous phytoplankton species have been detected since 1995. From an ecosystem viewpoint, introduced harmful microalgae have been involved in the loss of habitat for other phytoplankton and benthic algae species, and the disruption of the microbial food web. The introductions and the ecosystemic changes they trigger are in most cases irreversible.

Other factors have been invoked to explain the increase in frequency of harmful blooms. Eutrophication in estuaries and coastal waters (already discussed) is surely the most obvious (Burkholder 1998). Anomalous weather events (e.g., the increase in frequency, persistency and intensity of ENSO events; van Dolah 2000) and the warming trend of surface waters are also related to harmful blooms (Burkholder 1998, Mudie et al. 2002, van Dolah 2000).

The experimental evidence also points that toxic algal blooms will probably increase as a result of global change (Peperzak 2003, Peperzak 2005). Simulations of future environmental conditions for the coastal zone of The Netherlands for the year 2100 (i.e., increased stratifi cation of shallow


waters and 4ºC increase in the maximum summer temperature) led to a doubling of growth rates of harmful dinofl agellates.

Impacts on benthic communities

Global climate change projections indicate that coastal systems will be particularly vulnerable to future shifts (McLean et al. 2001), and that plankton and benthic communities (among other biotic components) will likely be affected (Barry et al. 1995, Sagarin et al. 1999, Clark and Frid 2001, Schiel et al. 2004, Beukema and Dekker 2005, Smith et al. 2006). Benthic communities in coastal ecosystems are probably most affected by global change. This, on the one hand, has to do with the reduced mobility of benthic organisms, but also with the location of coastal benthic environments in proximity to both terrestrial and marine disturbances. While all these circumstances make benthic environments more susceptible, it is also true that these environments are remarkably resilient and recovery of biological resources can occur rapidly due to recruitment from adjacent unaffected areas (Thompson et al. 2002). In this section of the chapter, we will discuss a few studies to illustrate the major global change-related impacts on different benthic communities and organisms.

Ascophyllum nodosum is a macroalga that extends its southernmost distribution range (for the western Atlantic) to Long Island Sound. Even at its range limit A. nodosum represents a major primary producer and habitat-forming organism (i.e., a key species in ecological terms). In recent years, it has been observed (Keser et al. 2005) that Ascophyllum experiences thermal stress in the area mentioned above (as evidenced by a rapid decrease in growth rates at temperatures above 25ºC and complete mortality at temperatures exceeding 27–28ºC). In view of recent continued warming of Long Island Sound, it is possible that Ascophyllum became locally extinguished for this region, which in turn is likely to have signifi cant impacts on the ecosystem (Keser et al. 2005).

Another example from coastal Gulf of Mexico, is the fl uctuation in the stocks of the eastern oyster, Crassostrea virginica, correlated with changes in freshwater infl ow (Hoffmann and Powell 1998). Suspension feeders such as mussels, clams and oysters are commercially important species susceptible to changes in temperature and salinity (Scavia et al. 2002). The freshwater input to Galveston Bay, regulates the salinity and fl uctuates between maxima and minima with a ~ 7–10 year periodicity (in turn, governed by climatic factors such as unusual spring storms). Salinity is the principal environmental factor that determines the spatial distribution and productivity of oyster reefs. Therefore, dramatic episodic decreases in salinity are followed by substantially reduced landings occurring about once every ~ 7–10 years (Hoffmann and Powell 1998).


Effects on coastal and estuarine fi shes

Nektonic communities (e.g., fish, squids and some large schooling crustaceans) are usually less impacted by local changes in the coastal zone due to their displacement ability, which enables them to migrate to other areas where they fi nd optimum ranges in environmental parameters. That explains why over the past few decades several changes in migration patterns have been recorded (Roessig et al. 2004). However migration to better grounds is not always the case and changes in environmental parameters are known to have affected the spawning patterns and larval drift of some species (Boesch 1999).

The general effect of projected human-induced climate change is that due to global warming the distributional ranges of many species (both terrestrial and marine) will move poleward from their current locations or expand their ranges (Kennedy 1990, Gitay et al. 2002), although it is impossible to make generalizations or reach a consensus on this respect. Some fi sheries and aquaculture enterprises would benefi t from the results of these range expansions, while others would most probably suffer losses (Kennedy 1990); some assessments suggest that overall productivity will be unaffected (Hall 2002) since local extinctions will generally be matched by colonization of new areas (Thompson et al. 2002).

Fish communities are just one component of the coastal ecosystem (probably the most important from a commercial viewpoint) and to study the extent to which global changes impact them is not necessarily a straightforward process. For example, Heath (2005) studied changes in the food chain structure and function (with especial attention to fi sh communities) for the North Sea for the past 30 years. He found that fi shing pressure initially caused a change in fi sh secondary production. Additionally, climatic variability for the North Sea combined with fi shing pressure to produce shifts in the proportion of piscivorous to planktivore fi sh species in a ~ 30 year period.

From the previous sections, it becomes apparent that the relationships among processes and factors involved in global change and coastal marine biodiversity are extremely complex. Figure 1 sumarizes the processes and relationships treated in the above sections.

Concluding Remarks: What Until Now, What for the Future

For reasons concerning social, economic and cultural circumstances, it is impossible to stop or revert the trends imposed by global change in the past ~ 200 years. Steele (1998) mentions that there is no longer a choice between pristine and managed coastlines, but only between priorities of


use, implying that at this stage it is important to take measures to regulate or mitigate the impacts of global change in the coastal environment.

With this in mind we would like to close this chapter with some fi nal remarks and ideas for the future. From what we have discussed, we can state that we are still far from reaching conclusive remarks on most of the global change threats on coastal areas. The best we can do, is try to draw similarities and congruencies in patterns of change, keeping in mind that coastal areas are interconnected and, at the longest time scales, the problems are certainly regional and fi nally global (Steele 1998).

There are pressing issues requiring stronger inclusion of science in ocean governance (Boesch 1999) in a globalized context; after all marine ecosystems rarely coincide with national boundaries and are affected by international economic, social and legal decisions, and therefore international agreement on policies and action is needed (Holdgate 1994).

Even though some gaps in knowledge still remain, the scientific community has gathered a body of evidence large and consistent enough

Fig. 1. Conceptual relationships among concepts involved in global change and coastal marine biodiversity. More links among the different concepts may be built, and many more concepts may be linked to the ones depicted here; the ones chosen are treated in the text. Boxes include biological components of ecosystems, physico-chemical forcings and processes, and ecological processes, that are affected in their rates or intensity by global change, with an ultimate effect on coastal biodiversity.


to prove the occurrence of global change on the coastal zone. Increasing levels of atmospheric CO2, UVR fl uxes and seawater temperatures can synergistically interact with each other and have profound effects on primary producers in marine ecosystems, which cascade up to higher trophic levels. There are multiple ways in which habitat loss affects marine species diversity. The loss of habitat structure generally leads to a decline in species richness and biomass (Airoldi et al. 2008). In turn, habitat complexity and biodiversity loss impacts negatively on the functioning of ecosystem and the services they offer.

Fig. 2. This scheme depicts the close relationships between the general public, the scientifi c community, policy makers, and politicians and decision makers, cartooned here as instrumental pieces in a clockwork. All four are key actors in the interplay of humans and the impacts on marine coastal ecosystems raised by global change. Surrounding these players are environmental policy reinforcement and education, as key strategies of urgent implementation for the amelioration of the effects of global change on coastal environments.


In our opinion the next 10 to 20 years will be instrumental to put into practice this knowledge on the broad spectrum of impacts affecting the coastal zone. Considering the non-linearities that arise from the interactions among multiple environmental stressors is of paramount importance (Beardall et al. 2009, Wootton et al. 2008) and so we believe that now is the time to move on to the next phase in global change research and start looking into the synergistic effects of multiple stressors (Boesch 2002, Cloern 2001).

Finally, it is of utmost importance that the key human actors (i.e., the general public, the scientifi c community, policy makers, and politicians and decision makers; Fig. 2) establish close relationships among each other and attain a common objective of environmental policy reinforcement and education. Such strategy needs urgent implementation if we are to ameliorate the impacts of global change on coastal marine ecosystems.

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CHAPTER 3

Climate Change Effects on Marine Phytoplankton

Valeria Ana Guinder1,* and Juan Carlos Molinero2

Introduction

Phytoplankton play a central role in marine ecosystems by yielding ca. 50% (~ 50 Gt C/year) of the global primary production (Longhurst et al. 1995, Field et al. 1998). By their central role at the base of the food web these communities shape biogeochemical cycles, carbon export from the euphotic zone to the deep ocean and energy fl uxes through food web networks (Finkel et al. 2010, van de Waal 2010). Whilst in coastal areas, microphytobenthos, macroalgae and halophytes also contribute in carbon fi xation (Kromkamp et al. 2006, Connell and Russell 2010), in the open ocean, phytoplankton constitute the only source of primary production to sustain pelagic food webs (Falkowsky and Oliver 2007, Chavez et al. 2011). In fact, phytoplankton blooms are an essential condition for fi sheries and for the benthic-pelagic coupling in coastal systems (Legendre 1990). These rising biomass events mainly occur in response to changes in light and nutrients driven by the seasonal cycles of radiation, temperature and water column stability; while the end phase has been ascribed to nutrient depletion and zooplankton grazing pressure (Sommer et al. 2012). Spring blooms are

1 Área de Oceanografía Química, Instituto Argentino de Oceanografi a (IADO-CONICET), Camino La Carrindanga km 7.5, CC 804 B8000FWB Bahía Blanca, Argentina.

2 Helmholtz Centre for Ocean Research Kiel (GEOMAR), Marine Ecology/Food Webs, Duesternbrooker Weg 20, D-24105 Kiel, Germany.

Email: [email protected]* Corresponding author: [email protected]

Climate Change Effects on Marine Phytoplankton 69

ubiquitous in temperate coastal systems and are recognized as the most constant interannual biomass events (Smayda 1998, Winder and Cloern 2010). A classic paradigm earlier advanced by Sverdrup (1953) explains the annual recurrence of phytoplankton spring blooms, and has been used as a baseline pattern to evaluate changes among ecosystems.

Over the last 50 years, reports on climate-related changes in marine ecosystems have noticeably increased (Hays et al. 2005, Harley et al. 2006, Parmesan 2006, Yang and Rudolf 2010). Phytoplankton responses to climate variations have been examined at different spatiotemporal scales, both in empirical (e.g., Feng et al. 2008, Huertas et al. 2012, Rossoll et al. 2012) and fi eld investigations (e.g., Wiltshire et al. 2008, Guinder et al. 2010, Wetz et al. 2011), as well as using modeling approaches (e.g., Sarmiento et al. 2004, McNeil and Matear 2006, Boyce et al. 2010). Climate modifi cations, such as the rise in atmospheric CO2 and warming, affect the marine biosphere through modifi cations in pH, carbonate availability, water column stability, nutrient and light regimes. These changes directly impact small-sized (ca. < 1 to > 100 µm) phytoplankton organisms, whose short-term life cycles make them amenable to quickly respond to subtle environmental variations. Therefore, tracking changes in the phytoplankton community structure can be an accurate indicator of ecosystem perturbations (Beaugrand 2005, Hays et al. 2005, Irwin et al. 2006). Modifi cations at the bottom of the food web are likely to permeate the trophic network due to trophic amplifi cation and the subsequent cascading effects (Fig. 1). Understanding how climate interacts with the marine environment from global to local scales is therefore critical to assess consequences on marine biota at all organization levels, from individuals (e.g., physiology, growth rate and cell size) to communities (e.g., structure and phenology).

In this chapter we review recent advances in the understanding of the physical and chemical nature of ocean-climate change and the implications for phytoplankton ecology. We fi rst introduce current global ocean threats, i.e., ocean acidifi cation and warming. We address both direct and indirect effects of these environmental changes on phytoplankton productivity and provide examples of proximate impacts on individuals, populations and communities by reviewing fi eld observations at different latitudes, empirical approaches and data modeling. We further examine broader ecological responses that emerge from these proximal impacts: alteration in the cycle of elements and plankton stoichiometry, changes in food webs structure and societal repercussions. We conclude by identifying future research foci that might help gaining a thorough understanding of phytoplankton responses to climate change.


Climate-driven Abiotic Changes in Marine Ecosystems

Rise in atmospheric carbon dioxide and ocean acidifi cation

Atmospheric carbon dioxide levels (CO2) have increased from pre-industrial levels by nearly 40%, from circa 280 ppmv to nearly 384 ppmv in 2007 (Solomon et al. 2007). Such rate of increase is of pressing concern, as it is at least one order of magnitude faster than the rate observed over the past centuries. Rising atmospheric CO2 is tempered by ocean uptake; however the diffusion into the water causes major impact on C chemistry, as dissolved CO2 reacts with H2O molecules to form carbonic acid (H2CO3), which dissociates into bicarbonate (HCO3

–), releasing a proton H+ and reducing pH levels. These chemical processes are collectively known as ocean acidifi cation and cause a decrease in the concentration of carbonate ions (CO3

2–), increasing the solubility of calcium carbonate (CaCO3) (Caldeira and Wickett 2003, Doney et al. 2009). Ocean acidifi cation has decreased the pH of surface waters by ~0.1 units over the last two centuries to a present pH average of 8.1 (Orr et al. 2005), and the projected concentrations of CO2

Tem

pora

l sca

le

Spatial scalekm μm

deca

des

min

utes

Interdecadalvariability

Interannualvariability

Regional/localclimate

Regional/localclimate

Seasonality

Behaviour

Physiology

Water column

Air temperatureAtmospheric pressureSolar radiation - UVGreenhouse gases

Globalclimate

PrecipitationWinds - stormsTemperatureCloudiness

Temperature, salinity, pHEstratification/mixLight availability/turbidityDissolved nutrients, oxygen

Life cycles (phenology)Ecological nichesSpecies interactionsCell size, methabolism

Phytoplankton

Fig. 1. Cascading climate-related changes throughout the atmosphere-sea interfaces, the pelagic environment and phytoplankton. Hydroclimatic variability from the large-scale and long-term to the local and short-term scales affects water properties and consequently the ecology of microalgae which follow the biophysical rules. Modifi cations in species physiology and behavior (e.g., ecological traits and trade-off) may restructure the phytoplankton community composition. These changes can alter the seasonal timing of blooms and eventually can result in shifts of interannual and/or interdecadal biomass patterns.


up to 750 ppm for the year 2100 implicate a decrease in pH of 0.3 units, which falls in a potential scenario outside the natural range over the past 20 million years (IPCC 2007).

Global warming and rising ocean temperature

Global air and sea temperatures have risen in the past century by 0.4–0.8°C. Global circulation models predict warmer conditions of additional ca. 3°C in some areas of the global ocean by the end of the 21st century, from a mean sea surface temperature (SST) of 18°C today to 21.5°C (McNeil and Matear 2006, IPCC 2007). Rising SST enhances the water column stratifi cation and decreases nutrient supply in the euphotic zone (Behrenfeld et al. 2006, Doney 2006). It further induces alterations in the underwater light regime (Sarmiento et al. 2004). Along with the growing temperatures, global circulation models forecast a potential freshening of mean sea surface salinity, presumably as a result of increased precipitation and ice-melt in the poles offsetting increased evaporation from the surface of ocean in low latitudes (McNeil and Matear 2006).

Additional abiotic consequences of climate warming are the thermal expansion of the world ocean, which coupled with freshwater input from ice-melt causes sea level rise (IPCC 2007). Thermal expansion enhances water column stratifi cation and a deepening of the thermocline preventing cool, nutrient-rich waters from being upwelled (Roemmich and Mc Gowan 1995). Owing to the fundamental importance of upwelling in the productivity of coastal marine systems, further elucidation of the linkage between these events and climate is a high research priority (Harley et al. 2006). In a more local scale, changes in atmospheric circulation also affect storm frequency and wind and precipitation patterns, which eventually may yield changes in coastal salinity, turbidity, light attenuation and inputs of terrestrial and bottom sediments-derived nutrients and pollutants (Nixon et al. 2009, Noyes et al. 2009, Wetz et al. 2011).

Overall, the main modifi cations that result from rising atmospheric CO2 and global warming in the physical and chemical nature of pelagic environments can be synthesized by 1) carbon enrichment and acidifi cation and 2) thermal stratifi cation and associated changes in nutrient and light regimes. These changes shape seawater chemical speciation, nutrient supply and biogeochemical cycles, and ultimately structure of ecosystems. In the following sections we assess the implications of such climate-related modifi cations on phytoplankton ecological performance and the repercussion on pelagic food webs.


Phytoplankton Responses to Hydroclimatic Changes

Ocean acidifi cation

Calcifi cation: Compared to the vast studies about warming effects on marine phytoplankton growth and production rates, data are relatively scarce on the responses to increased CO2 and low pH (Berge et al. 2010). This is in part related to the fact that the chemical reactions involved in ocean acidifi cation are complex, and therefore diffi cult to reproduce under laboratory conditions (Iglesias-Rodriguez et al. 2008). The general mechanisms can be explained as follows: the decrease in carbonate ions concentration (CO3

2–) linked to ocean acidifi cation causes higher solubility of calcium carbonate (CaCO3) (Caldeira and Wickett 2003), and thus aragonite, the metastable form of CaCO3, becomes less available for organisms that need it to build their skeletons, such as coccolithophores, corals, foraminifers and mollusks (Orr et al. 2005, Hoegh-Guldberg et al. 2007). Coccolithophores have received special attention because their calcite precipitation plays a signifi cant role in alkalinity fl ux to the deep ocean (i.e., inorganic carbon pump). For instance, malformation of CaCO3 skeletons and reduced cell size in response to high CO2 levels have been reported in monocultures of two marine dominant calcifi ed phytoplankton species: Emiliania huxleyi and Gephyrocapsa oceanica (Riebesell et al. 2000). In agreement to this, empirical studies using mesocosms have evidenced a decrease in E. huxleyi calcifi cation and enhanced loss of organic carbon from the water column when exposed to high CO2 (Delille et al. 2005). However, other laboratory experiences using the same coccolithophore species under elevated CO2 showed an increase in calcifi cation and net primary production which agrees with fi eld evidence based on geological records in the deep ocean over the past two centuries of anthropogenic CO2 rise (Iglesias-Rodriguez et al. 2008). Likewise, a recent theoretical model provided by Irie et al. (2010) on the growth schedule of coccolithophores forecasted how natural selection alters phenotypes as ocean acidifi cation increase. Assuming that the formation of exoskeleton is a defensive strategy to reduce the instantaneous mortality rates, the model predicts that natural selection favors constructing more heavily calcifi ed exoskeleton—and slows down the growth strategy—in response to increased acidifi cation-driven costs. This raises fundamental questions regarding the plasticity of phytoplankton species responses to Global Ocean threats, and calls for an evolutionary perspective to assess climate changes effects on phytoplankton.

The studies described above show a variety of responses to acidifi cation by coccolithophore species. It is worth noting that the experiment design can induce variability in the phytoplankton reactions to the dissociation of carbonate species and pH. Manipulation of the pH levels by the addition


of acid and/or base or by bubbling CO2-enriched air through the seawater have different effects on the water bio-chemistry. According to Iglesias-Rodriguez et al. (2008), the latter pH manipulation may be a more realistic representation of the ocean response to anthrophogenic change: the relative proportion of the carbonate species are controlled by decreasing pH at the same time when concentration of DIC increases. However, as recently reviewed by Andersson and Mackenzie (2012), addition of acid, base or CO2 gas to seawater can all be useful techniques to manipulate seawater chemistry in ocean acidifi cation experiments.

Interactive effects on phytoplankton: Ecophysiological responses of marine phytoplankton to acidifi cation not only depend on community composition but on acclimation and growth conditions as well. The interactive effects of CO2 rise and warming-related changes, i.e., shifts in mean irradiance exposure, nutrient inputs, sinking rates and organic carbon exportation from the euphotic zone, lead to large uncertainties related to phytoplankton physiological assumptions (Feng et al. 2008, Tagliabue et al. 2011, Gao et al. 2012). For instance, Feng et al. (2008) empirically demonstrated that the combined effects of CO2 levels, temperature and irradiance on Emiliania huxleyi enhanced photosynthesis by increasing both CO2 and temperature regardless of the irradiance regime, and calcifi cation decreased by a combined effect of CO2 and light. In contrast, natural phytoplankton assemblages exposed to rising CO2 and increased light (Gao et al. 2012) reduced their growth and photosynthesis rates and the community shifted away from diatoms. These recent works highlight the synergistic effects of key modulating factors of phytoplankton primary production and community composition, which can signifi cantly impact higher trophic levels and carbon cycle in the ocean.

In natural environments, seawater pH directly affects the phytoplankton growth rate and therefore the timing and abundance of coastal species (Hinga 2002). Accordingly, direct CO2-related effects on E. huxleyi growth, calcification and elemental stoichiometry of uptake and production processes have been reported from outdoor mesocosms experiments with eventual implications for the marine biogeochemistry (Engel et al. 2005). The oceanic CO2 enrichment and acidifi cation also perform indirect effects at the base of the pelagic food webs through modifi cations in the elemental composition of the marine water. On the one hand, a recent study in a coastal system has shown that nitrifi cation rates were highest at low pH indicating that nitrifying organisms tolerate a wide range of pH levels (Fulweiler et al. 2011). The impact of pH range on nutrient cycles depends upon the local environmental conditions and active biological processes (Hendriks et al. 2010). For instance, under eutrophic conditions, productivity and C sink are predicted to increase, and nuisance phytoplankton blooms may


be aggravated at elevated atmospheric CO2 concentrations (Schippers et al. 2004). On the other hand, combined ocean CO2 enrichment with nutrient depletion in the upper layers due to thermal stratifi cation may cause C-to-nutrient ratios imbalance with signifi cant implications on phytoplankton stoichiometry, food quality and on the structure of the pelagic food webs (van de Waal et al. 2010), as we will discuss later. Accordingly, recent experiments demonstrated that the composition and structure of fatty acids in the diatom Thalassiosira pseudonana change signifi cantly when cultivated under high CO2 levels, and such changes are likely to permeate the food web as they constrain somatic growth and eggs production of the copepod Acartia tonsa (Rossoll et al. 2012).

Overall, contrasting responses-stimulation (Schippers et al. 2004) and reduction (Steinacher et al. 2009), have been acknowledged concerning the effects of ocean acidifi cation on global marine primary production, particularly by shell-forming and calcifying organisms (Riebesell et al. 2000, Kroeker et al. 2010). Indeed, two recent meta-analysis of empirical and fi eld assessments yielded rather contrasting conclusions on the responses of marine biota to increased CO2 (Hendriks et al. 2010, Kroeker et al. 2010). Meanwhile, other studies reported that marine phytoplankton in general appear resistant to ocean acidifi cation, showing no increase or decrease in responses in growth rates under ecological relevant ranges of pH and CO2 (Berge et al. 2010). Hence, the extent to which rising atmospheric CO2 will enhance or reduce global primary production in the oceans remains equivocal. Further fi eld research and accurate empirical representation of future projections of the carbonate systems are needed to assess both direct and indirect effects of ocean acidifi cation on marine phytoplankton.

Water warming

Direct effects on phytoplankton: Temperature is a key parameter that directly affects physiological rates of marine biota at multiple scales, e.g., enzymatic reactions, respiration, body size, generation time, ecological interactions, community metabolism, etc. (Peters 1983). Phytoplankton experience an increase in enzymatic activity and growth rates over a moderate range of temperature rise with an average Q10= 1.88 (Eppley 1972), which suggest that an increase in SST from 18°C today to 21.5°C in 2100 (McNeil and Matear 2006), may lead to an increase of ~25% in growth rate assuming that there are no other factors (Finkel et al. 2010). Nevertheless, considering the polyphyletic complexity of the phytoplankton community, the temperature impact on metabolic rates is intricate by individual species’ vulnerability to warming (Huertas et al. 2011). Further consequences of rising temperatures are related with the germination of resting spores in sediments (Shikata et al. 2008). The increase in both water temperature and


underwater light intensities are also recognized as important environmental triggers for summer and spring diatom blooms in shallow environments (Eilertsen et al. 1995).

Salinity closely co-varies with temperature and also has important implications in plankton physiology, affecting germination of resting stages, growth rates and development of blooms in coastal waters (McQuoid 2005, Shikata et al. 2008). The effect of the climate warming on water temperature and salinity appears stronger in shallow and semi-enclosed areas where evaporation is high and river runoff is low (Guinder et al. 2010). Phytoplankton species have different tolerances to variations in salinity and temperature (Gebühr et al. 2009, Huertas et al. 2011), which defi ne the water density and viscosity, and shape nutrient diffusion and cell motility (Falkowsky and Oliver 2007, Finkel et al. 2010). Therefore, changes in these parameters affect ecological niches and species-specifi c interactions leading to shifts in community structure and composition (Litchman et al. 2007). Changing conditions also favor the development of fast-growing opportunistic species, able to exploit open niches and establish dominance in the system (Cloern and Dufford 2005). Particular examples are the cases of Paralia sulcata at Helgoland Roads in the North Sea (Gebühr et al. 2009) and Thalassiosira minima in the Bahía Blanca Estuary, Southwestern Atlantic coast (Guinder et al. 2012).

Indirect effects on phytoplankton: Enhanced growth of primary producers is expected under future trends of temperature increase. The projected scenario, however, becomes complex when considering indirect effects of warming, such as grazing acceleration, which play key modulating roles of phytoplankton biomass accumulation (Aberle et al. 2012, Sommer and Lewandowska 2011). Hence, food web interactions need to be considered when assessing temperature effects on phytoplankton (Klauschies et al. 2012).

Sea-surface warming lead to higher mean underwater irradiances and nutrient depletion from the upper layers due to intensifi ed stratifi cation (Doney 2006). The establishment of a shallow pycnocline (Fig. 2) acts as a barrier for vertical mixing and upward transport of nutrients and resting stages, and exposes the phytoplankton cells to more intensifi ed irradiances, both PAR and UVR. This affects cellular photochemistry, eventually hampering the photosynthetic carbon assimilation (Litchman and Neale 2005, Marcoval et al. 2008), and ultimately may yield a decline in phytoplankton productivity under the above conditions (Behrenfeld et al. 2006, Doney 2006, Boyce et al. 2010). On the contrary, positive effects of warming on photoautotrophic production and phytoplankton biodiversity have been registered related to upwelling events (Chavez et al. 2011), earlier stratifi cation and extension of the growing season with adequate resource supply (Winder and Schlinder 2004, Sommer et al. 2012b).


The documented contrasting results highlight complex non-linear responses of phytoplankton to global ocean climate and cascading consequences linked to predator-prey interactions, resource limitation and community species composition. Moreover, predicting future distributional shifts in accordance to warming—i.e., temperature gradients—(Rühland et al. 2008, Morán et al. 2009) requires additional knowledge of species’ range boundaries (population plasticity) and their driving factors. Particular attention should be paid to alongshore where warming-associated weakening of advection could break down certain marine biogeographical barriers that currently prevent range expansions (Harley et al. 2006). We describe hereinafter the common ecophysiological responses of phytoplankton to modifi cations in the pelagic environment driven by warming, including: size structure, bloom phenology, elemental stoichiometry and food quality.

Cell size: Increasing evidence of changes in plankton size structure has been reported worldwide in relation to global warming (Forster et al. 2012 and references therein). Temperature effects on the size structure (i.e., a reduction in the mean size) have been detected in microzooplankton (Molinero et

Fig. 2. Contrasting scenarios of vertical mixing, light and nutrient regime. Solar radiation is exponentially attenuated through the water column. The euphotic zone is defi ned as the depth at which the underwater irradiance reaches the 1% of the incident irradiance at the surface, and mainly depends on turbidity. The pycnocline is the vertical gradient in water density caused by differences in temperature (thermocline) or salinity (halocline) and defi nes the limit of the mixing depth. The critical depth in a) is smaller compared to the one observed in b), but a weaker and deeper pycnocline allows the upward transport of bottom nutrients through strong vertical mixing. In b), a shallow and strong pycnocline acts as a barrier against mixing causing nutrient depletion in the upper layers. Furthermore, the phytoplankton cells in b) could be negatively affected by prolonged exposures to high irradiance, causing photoinhibition or photodamage. Further consequences of strong vertical stratifi cation are higher sinking rates and cell loses beyond the mixing depth.

a) b)


al. 2006), stream fi sh communities (Genner et al. 2010) and suggested in pelagic marine copepods (Beaugrand et al. 2003). Further fi eld studies (Gómez and Souissi 2007, Winder et al. 2008, Guinder et al. 2010, 2013), and empirical investigations (Sommer and Lengfellner 2008, Lewandowska and Sommer 2010) reported a restructuring of the phytoplankton community composition toward a dominance of small species along with increasing water temperature. The reduction in cell/body size of planktonic organisms and displacements of species ranging to higher latitudes have been mainly ascribed to global warming (Morán et al. 2009, Daufresne et al. 2009, Beaugrand et al. 2010), which might affect food web networks, with potential negative effects on the biological carbon pump.

Phytoplankton cell size follows biophysical rules (e.g., nutrient uptake, motion, sinking rates, kinetics of metabolism) that affect growth rates, the biogeochemical cycling and trophodynamics (Finkel et al. 2010). The replacement of large cells by smaller ones under warming conditions is likely related to shifts of the species’ environmental optimum growth and the higher competitive skills of small cells (Winder et al. 2008). A non-exclusive hypothesis suggests changes in grazing rate or selectivity of zooplankton (Sommer and Lengfellner 2008, Klauschies et al. 2012). The climate-change scenario of warming and nutrient depletion in the euphotic zone favor the dominance of small-sized phytoplankton species (e.g., Rodriguez et al. 2001), as they present higher surface to volume ratios and thus lower sinking velocities (Huisman and Sommeijer 2002) and small diffusion boundary layers, i.e., more effi cient nutrient uptake and superior ability to harvest light (Litchman et al. 2007). In agreement to this, the appearance of small phytoplankton species followed by the persistence, perennial predominance or even establishment as dominant species has been increasingly documented worldwide (Hays et al. 2005, Beaugrand et al. 2010). For instance, the abundance of Cyclotella taxa has increased in lakes since the nineteenth century (Rühland et al. 2008, Winder et al. 2008) linked to enhanced thermal stratifi cation. Similarly, the phytoplankton size-structure in some estuaries has shifted towards the dominance of smaller diatoms, e.g., Cyclotella sp. and Thalassiosira minima (Guinder et al. 2010) in relation to complex interactive effects of increase in water temperature and salinity, as well as changes in precipitation regime and modifi cations in trophic interactions (Guinder et al. 2013).

Bloom phenology and magnitude: Rising temperatures in both marine and freshwater systems are related to the advancement of seasonal ecological events in phyto- and zooplankton (Parmesan and Yohe 2003, Edwards and Richardson 2004, Winder and Schindler 2004). The phenology is the study of the seasonal cycles, i.e., phytoplankton blooming events, zooplankton hatching eggs, and their link with environmental variations. In coastal


systems, phenology has been associated to both climate and anthrophogenic disturbances on the marine biota (Yang and Rudolf 2010), and in many species phenology is biased in the directions predicted from global warming in the last few decades (Parmesan 2006). The synchronization of the phenological cycles of phyto- and zooplankton is crucial for the matter and energy transfer through the food web (Beaugrand et al. 2003, Edwards and Richardson 2004, Chassot et al. 2010). In fact, a mismatch scenario between food availability and heterotrophic demand might profoundly affect population of superior predators (Durant et al. 2007, Yang and Rudolf 2010).

Spring phytoplankton blooms are ubiquitous in temperate systems, but growing evidence showed changes in phenology, magnitude and composition both in the fi eld (Cloern et al. 2007, Wiltshire et al. 2008, Guinder et al. 2010) and in mesocosms experiments (Sommer and Lengfellner 2008, Lewandowska and Sommer 2010). Increasing temperature advances the spring phytoplankton bloom, and the degree of advance depends on resource dynamics, predator-prey interactions and taxonomic phytoplankton groups according to their physiological characteristics. In deep systems with thermal stratification, spring blooms are triggered by correlated increases in temperature and seasonal light availability (Edwards and Richardson 2004). Conversely, in shallow, well-mixed systems, phytoplankton blooms can occur coupled to external light regime and independently of temperature change (Sommer and Lengfellner 2008). For instance, in shallow estuaries, changes in turbidity, i.e., light attenuation, salinity and nutrient supply along the land-sea transition can signifi cantly affect the magnitude of the phytoplankton bloom and the community structure (e.g., Struyf et al. 2004). Similarly, changes in phytoplankton phenology and species composition of the winter-early spring bloom have been observed in the Bahía Blanca Estuary (Argentina) in relation to long-term decreasing trends in local precipitations and warmer conditions over the last decades (Guinder et al. 2010). In the eutrophic Neuse River Estuary (USA), phytoplankton production has been signifi cantly reduced in response to droughts events (Wetz et al. 2011). In the Narragansett Bay (USA), changes in the phytoplankton annual pattern over the last 50 years (i.e., decrease in the winter-spring bloom and occurrence of relatively short diatom blooms in spring, summer and fall) have been related to warming water especially in winter, cloudiness and a signifi cant decline in the wind speed (Nixon et al. 2009), together with a shift from eutrophic to oligotrophic conditions due to wastewater treatments. A further example is given by the increase in phytoplankton summer blooms in the Bahía Blanca Estuary in recent years. The combination of dredging operations together with changes in the wind pattern have induced the resuspension of nutrients and resting stages of diatoms (Guinder et al. 2012), that have


subsequently germinated under the current warmer, more saline and highly turbid conditions (Guinder et al. 2013).

Summer blooms have notably increased in magnitude and frequency in several coastal systems worldwide, although different underlying factors have been identifi ed (Carstensen et al. 2007, Shikata et al. 2008, Guinder et al. 2013). In a shallow coastal ecosystem in the north of Europe, the Kattegat strait, summer phytoplankton blooms are thought to be related to short and strong nutrients pulses associated to river discharge, resuspension from the bottom and anthropogenic inputs (Carstensen et al. 2004). In contrast, in the Hakata Bay, Japan (Shikata et al. 2008) and in the Gullman Fjord, Sweden (McQuoid 2005) the occurrence of summer blooms seems to result from the germination of resting stages of different phytoplankton species (mainly diatoms) in response to environmental stimulus, i.e., increase in radiation and water temperature and climate related changes in sea-surface salinity. These cases evidence that the responses of phytoplankton bloom phenology to climate change largely depend on the life strategies of the community.

The metabolism of heterotrophic organisms is more sensitive to temperature than photosynthesis rates. In consequence, the zooplankton grazing activity will be more affected than the primary production as warming progresses, thereby enhancing the top-down control on the timing and magnitude of phytoplankton blooms (Irigoien et al. 2005, Aberle et al. 2012, Klauschies et al. 2012). For instance, in mesocosms studies, Aberle et al. (2012) demonstrated that an increase in the winter temperature produces accelerated growth and large ciliate biomass, altering the specifi c composition and creating an asynchrony between the components of the plankton. Additionally, Sommer and Lengfellner (2008) found higher grazer activities in the warmer mesocosms due to enhanced metabolic demand of copepods at higher temperatures, which could explain both the decreased phytoplankton biomass during the spring bloom and the shift towards smaller phytoplankton at higher temperatures. It is therefore plausible that released predation pressure on small phytoplankton cells under warmer conditions may promote their outburst with a potential reduction in the matter transfer through the trophic chain (Sommer and Lewandowska 2011, Winder et al. 2008, Guinder et al. 2012). The dominance of smaller phytoplankton may cause a shift in the pelagic food web away from the biological pump dominated by copepods and rapid sedimentation of particulate matter towards rapid carbon cycling in the microbial loop (Finkel et al. 2010). The path of carbon fl ow between primary producers and mesozooplankton may become longer through heterotrophic fl agellates and ciliates, which can reduce productivity of higher predators.

Elemental stoichiometry and food quality: As previously described, ongoing anthropogenic increases in atmospheric CO2 levels and global warming of


the world’s oceans will modify water chemistry through carbon enrichment and nutrient depletion from the upper layers due to thermal stratifi cation. Because taxonomic groups (i.e., diatoms and non-diatoms) have different nutrient concentration/ratios requirements and different sensitivity to light levels, shifts in species dominance and size-structure are expected under modifi cations of the surrounding environment (Tilman 1982, Litchman et al. 2007). The environmental stoichiometry affects the metabolic rate of photosynthetic organisms, because the rate at which they acquire energy and materials for maintenance, growth and reproduction depends on both, their specifi c cellular requirements, as well as external bio-availability of elements (Sterner and Elser 2002, Finkel et al. 2010). Accordingly, element imbalance in the seawater has signifi cant implications in the phytoplankton stoichiometry and therefore in the quality of food available for higher trophic levels (Fig. 3) (van de Waal et al. 2010, Klauschies et al. 2012). It is worth noting that the food quality encompasses all features of the food that make it suitable for ingestion and for fulfilling the consumer’s nutritional requirements (Sommer et al. 2012). Therefore, quality properties not only include stoichiometric composition and biochemical make-up but also morphological characteristics such as presence of setae or cell projections and life-styles concerning solitary or chain forming and motile (e.g., fl agellates) or non-motile cells.

Fig. 3. Changes in the plankton elemental stoichiometry driven by the interactive effects between the rise in the atmospheric carbon dioxide (CO2) and sea-surface warming. Higher CO2 levels are available for primary producers and vertical thermal stratifi cation causes nutrient depletion in surface layers. Under these conditions, the phytoplankton stoichiometry shifts towards high C-to-nutrient ratios and the community structure towards dominance of smaller species. These changes in phytoplankton affect the zooplankton composition promoting shifts towards species with low nutrient requirements and high recycling effi ciency of nutrients. To compensate the low food quality, zooplankton excrete high carbon levels acting as a feedback mechanism for the imbalanced C-to-nutrient ratios in the environment. The high loads of dissolved organic carbon (DOC) are in turn transformed into CO2 and liberated into the atmosphere through microbial decomposition. Storms and eutrophication may partially compensate the water column stratifi cation and nutrient depletion.


The elemental composition of phytoplankton affects the cellular concentrations of proteins, fatty acids and other important constituents for zooplankton growth. The most widely used stoichiometric relationship in marine systems is the Redfi eld ratio of around 106:16:1 for the molar C:N:P ratio (Redfield 1934). In general, individual species in natural assemblages display signifi cant variability in this proportion depending upon the concentration of bio-available N and P and changes in irradiance, temperature and carbon dioxide (e.g., Hessen et al. 2002, Finkel et al. 2006, Fu et al. 2007). In particular, the predicted excess of C in relation to N and P, and the increase in underwater irradiance, associated to projections of CO2 and warming (i.e., thermal stratifi cation), will shift the phytoplankton cellular stoichiometry towards higher C-to-nutrient ratios (Fig. 3), with negative cascading effects on herbivores performance (e.g., high C:P ratios, Hessen et al. 2002) and eventually on upper trophic levels through trophic linkages amplifi cation (Hessen and Anderson 2008). In agreement to this, negative changes in algal fatty acid composition (i.e., decrease of polyunsaturated and increase of saturated fatty acids) have been shown under acidifi ed conditions with detrimental effects on the reproduction of copepods (Rossoll et al. 2012). Changes in algae stoichiometry further drive shifts in zooplankton community composition (i.e., microzooplancton vs. mesozooplankton dominance) and in the whole pelagic food web in relation to phytoplankton community size-structure and food quality and quantity (Fig. 3) (Sterner and Elser 2002). The zooplankton community shifts towards organisms with high effi ciency for nutrient recycling regarding the elevated C-to-nutrient stoichiometry of their food and consequently high amounts of POC are excreted (van de Waal et al. 2010). The offset between food quality and quantity depends strongly upon facilitation via grazing and recycling by grazers, and this effect is more important in systems with low renewal rates (Hessen and Anderson 2008).

Overall, both CO2 water enrichment and the strengthening of thermal stratification will enhance elemental imbalance between phyto- and zooplankton generating a feedback mechanism of increasing atmospheric carbon dioxide and climate warming driven by plankton activity (Fig. 3) (van de Waal et al. 2010). It is important to consider that the increase in storms frequency and eutrophication of coastal environment may in part prevent the projected shift to more stratifi ed and oligotrophic sea-surface waters.

Harmful Algal Blooms (HABs) and Parasitism of Phytoplankton

Coastal ecosystems are highly dynamic in terms of hydro-climatic variability, biogeochemical processes, occurrence of phytoplankton blooms and food webs structure (Cloern and Jassby 2008, Winder and Cloern 2010).


Additionally, increasing human settlements in the near-shore modify the environment trough eutrophication and pollution processes, and signifi cantly affect the marine biota and human health (e.g., McMichael et al. 2006, Moore et al. 2008, Noyes et al. 2009). The synergistic effects of anthropogenically driven temperature rise (Barnett et al. 2005) and eutrophication (Cloern 2001) can enhance the occurrence of harmful algae blooms HABs in the coasts (Edwards et al. 2006, Moore et al. 2008, Paerl and Huisman 2008). HABs have signifi cant negative implications for the marine ecosystem functioning, as they can cause detriment to biodiversity and eventual death of predators, causing severe impacts on fi sheries and resource availability with serious repercussion for human health and economy. In addition, parasitic infection of phytoplankton has been increasingly documented in relation to enhanced temperature, changes in pH, salinity and turbulence (Kühn and Hofmann 1999, Kühn and Köhler-Rink 2008); and parasitism has been suggested as a natural control of HABs (Elbrächter and Schnepf 1998). Parasitoid protists of phytoplankton comprise diverse taxonomic groups, such as euglenozoa, dinofl agellates, cercomonads, plasmodiophorids, oomycetes and chytrids (fungi). Parasitism is often highly host-specifi c, and the rates of infection and transmission increase with host population density, which is drastically reduced when conditions allow epidemic outbreaks of disease (Tillmann et al. 1999). How climate change and anthropogenic infl uence will affect parasite-host assemblages in coastal system needs further investigation (Brooks and Hoberg 2007, Colin and Poulin 2012). Particular attention must focus on parasitism of phytoplankton because it constitutes an important detrimental factor of biomass at the base of the food webs (Kagami et al. 2007).

Climate-driven modifi cations in coastal environments are expected to increase and continue, as well as human settlements in the near-shore and marine resources exploitation (IPCC 2007). Hence, addressing the alterations at the base of the food webs and the repercussion on the ecosystems structure and dynamics (e.g., carbon fl uxes, trophic interactions, HABs, parasitism) is crucial for safeguarding habitat sustainability and developing a sustainable management of ecosystem services. For this purpose, long-term phytoplankton records are essential to understand how coastal environments respond to variations in climate and anthropogenic perturbations or their synergies.

Summary of Common Phytoplankton Responses to a Changing Climate

Climate-driven modifi cations in marine water physics and chemistry impact phytoplankton from the individual to ecosystem levels (Fig. 4) through alterations in both bottom-up and top-down controls, namely resources


availability -light and nutrients- and zooplankton grazing pressure and selectivity. At the individual level phytoplankton is affected via changes in physiology, morphology and behavior, whereas the population level by shifting tolerance ranges and thus amplitude of ecological niches, which in turn infl uence the dispersion and recruitment of the species. The community level is affected by changes in structure (size, composition, diversity), phenology and the biogeographical distribution of species that derive in new inter-specifi c interactions and trophodynamics. Among the most common direct and indirect universal responses of phytoplankton to climate change we can summarize: 1) alterations in growth and photosynthesis rates and in calcifi cation processes related to CO2 rise and ocean acidifi cation, 2) dominance of smaller species under warmer conditions (related to both water stability and/or grazing pressure), 3) changes in the phenology, magnitude and species composition of phytoplankton spring bloom due to earlier thermal stratifi cation, overwintering and/or enhanced zooplankton activity, 4) occurrence of biomass peaks or miniblooms in other seasons (e.g., summer, autumn) and HABs in coastal systems related to warmer conditions, eutrophication and/or changes in wind patterns, and 5) changes

Fig. 4. Marine phytoplankton is affected by climate-driven hydrological modifi cations at every level of ecological organization (from organisms to ecosystem). The direct effects of increasing CO2 and temperature on phytoplankton are related with cell physiology (e.g., photosynthesis, growth rates and range shifts). The indirect effects are related with modifi cations of the pelagic environmental conditions (e.g., pH, light, nutrients and grazing activity), affecting phytoplankton size-structure, stoichiometry, sedimentation rates, species interactions and bloom phenology. In turn, these changes result in substantial alterations of the structure of pelagic food webs and ecosystem functioning.


in elemental stoichiometry and food quality in response to interactive effects of CO2 enrichment and nutrient depletion from upper layers due to sea-surface warming.

Final Remarks

Concerning the wide temporal- and spatial-scale effects of climate change on marine phytoplankton, the integration of continuous monitoring programs with empirical research in the laboratory arises as an imperious necessity for a better interpretation of current responses and accurate projections of future scenarios. On one hand, field studies provide information about in situ environmental conditions and natural regulator factors of phytoplankton succession. Moreover, they allow addressing the ecosystem functioning and the underlying controlling forces that emerge from the interaction among the atmosphere, the sea and the bottom sediments. On the other hand, experimental research under controlled conditions complements fi eld observations because proximal phytoplankton responses (e.g., cell physiology and autoecology) can be straightforwardly measured. Experimental investigation is useful to test hypotheses that come out from observations in the natural ecosystems, with a simple interplay of variables. The integration of both approaches –natural and experimental simulation—together with predictive data modeling will provide new insights of the interactive effects of abiotic and biotic forces affecting the primary producer’s ecology under different climate change scenarios. Accordingly, emergent responses that could be masked by considering only one type of approach can be revealed.

Coastal areas deserve particular attention as they have been recognized as the most productive ecosystems on the Earth. These environments are exposed to the synergic effects of climate modifi cations and antrophogenic impacts, whose consequences provoke substantial changes on the benthic-pelagic habitat and the associated biota. The implementation of rigorous programs of coastal management and resource exploitation is imperative. We must improve our ability to disentangle the phytoplankton ecological responses and the fl uctuations in biomass in order to predict and mitigate potential detrimental effects throughout the food webs towards ecosystem functioning and habitat sustainability.


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CHAPTER 4

Climate Change and Marine Zooplankton

María C. Menéndez,* Melisa D. Fernández Severini, Florencia Biancalana, María S. Dutto, María C. López Abbate

and Anabela A. Berasategui

Introduction

Humans infl uence climate mainly through fossil-fuel, industrial, agricultural, and other land-use emissions that alter atmospheric composition (Doney et al. 2012). Long-lived, heat-trapping greenhouse gases (CO2, CH4, N2O, tropospheric ozone, and chlorofl uorocarbons) warm the surface of the planet, whereas shorter-lived aerosols can either warm or cool at a lower spatial scale (Doney et al. 2012). CO2 is particularly important for the Earth’s climate system. Its worldwide output is enormous, entailing a ~40% increase of its atmospheric concentration over the past 250 years (Danovaro et al. 2011). According to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change (IPCC), the global mean surface air temperature increased by 0.74°C whereas the global mean sea-surface temperature rose by 0.67°C over the last century (Trenberth et al. 2007).

The oceans cover approximately 70% of the surface of the Earth and have the potential to store >1000 times more heat than the atmosphere (Levitus et al. 2005). Oceans play a key role in regulating climate by storing, distributing

Instituto Argentino de Oceanografía, Consejo Nacional de Investigaciones Científi cas y Técnicas (IADO-CONICET). Camino La Carrindanga km 7.5, B8000 FWB Bahía Blanca, Argentina.

* Corresponding author: [email protected].


and dissipating energy from solar radiation and exchanging heat with the atmosphere (Danovaro et al. 2011). They are also the main reservoir of heat and salt and regulate the evaporation and precipitation rates (Danovaro et al. 2011). Moreover, oceans are able to store large quantities of CO2 (Danovaro et al. 2011). Since the beginning of the 19th century, the oceans are estimated to have taken up 50% of fossil fuel emissions and 30% of all anthropogenic emissions (including those from land-use activities), thereby reducing the build-up of CO2 in the atmosphere (Danovaro et al. 2011). The direct and indirect impacts of the increase of greenhouse gas concentration on the oceans will include increasing temperatures, acidifi cation, changes in the density structure of the upper ocean and alteration of vertical mixing of waters, intensifi cation/weakening of upwelling winds, and changes in the timing and volume of freshwater runoff into coastal marine waters, among others (Fig. 1) (Moore et al. 2008).

Zooplanktonic organisms are key components of marine ecosystems as integral links between primary producers and upper trophic levels. Zooplankton communities are highly diverse and thus perform a variety

Fig. 1. Important abiotic changes in the oceans associated with climate change. Human activities infl uence climate mainly through fossil-fuel, industrial, agricultural, and other land-use emissions that alter atmospheric composition; CO2 is particularly important for the Earth’s climate system. The direct and indirect impacts of these increase of greenhouse gas concentration on the oceans leads to a suite of physical and chemical changes in coastal ecosystem. See text for details.

Climate Change and Marine Zooplankton 93

of ecosystem functions (Richardson 2008). One of the most important roles of zooplankton is as the major grazers in ocean food webs, providing the main pathway for energy from primary producers to consumers (Richardson 2008). Consequently, climate-induced changes on pelagic ecosystems are effectively transferred by zooplankton to upper trophic levels as commercially important fi sh populations (Mollman et al. 2008). The goal of this chapter is to summarize the observed and potential future responses of zooplankton communities to climate change. It focuses mainly on the effects that global warming, ocean acidifi cation and UV radiation have on zooplankton communities.

Zooplankton as Indicator of Climate Change

The observed climate modifi cations in the last decades have led to an increased effort on monitoring the environmental conditions of aquatic ecosystems. The detection of sentinel organisms may provide an early warning of climate-related environmental degradation. Zooplanktonic organisms are particularly valuable bioindicators of climate-driven change in marine environments, since they present various particular attributes:

• They are poikilothermic, so their physiological processes (e.g., ingestion, respiration, reproduction) are highly sensitive to changes in temperature.

• Zooplanktonic species have a short life cycle; therefore, population size is less infl uenced by the persistence of individuals from previous years (Richardson and Kunz 2006). This leads to a close connection between environmental changes and population dynamics (Richardson 2008).

• Few species of zooplankton are commercially exploited; consequently, any long-term variation in response to environmental change is generally not confounded with trends in exploitation (Richardson 2008).

• Zooplanktonic organisms are better indicators of change than environmental variables themselves, because the non-linear responses of the organisms can amplify environmental perturbations (Taylor et al. 2002).

• Considering their free floating habits, zooplankton can show dramatic changes in distribution and may respond easily to changes in temperature and oceanic currents by expanding/contracting their distributional ranges (Richardson and Kunz 2006).


Marine Zooplankton and Global Warming

Global warming and increase in sea surface water temperature

The increase in production of greenhouse gases has contributed to the observed warming of the atmosphere and the oceans in the last decades. Rising atmospheric concentrations have increased global mean sea-surface temperature by approximately 0.2°C per decade over the past 30 years (Guldberg and Bruno 2010). Consequently, the heat content of the upper 700 m of the global ocean has increased by 14 x 1022 J since 1975 (Guldberg and Bruno 2010), resulting in many cascading environmental changes (Fig. 1).

The warming of the upper layers as well as mid- to high-latitude freshening (Doney et al. 2012) promotes greater stratification of the water column, reducing mixing of the ocean and consequently affecting nutrient cycling and primary production. The relationship between nitrate concentrations and sea-surface temperature suggests that global nitrate supply to the surface might have decreased in the 20th century as a result of climate-driven changes in ocean stratifi cation and circulation (Kamykowski and Zentara 2005). The decrease in surface nitrate as a result of water stratifi cation, however, does not consider anthropogenic changes in riverine nutrient inputs, which might also affect global ocean uptake of atmospheric CO2 (Schultz 2008). Recent observations indicate that ocean warming and increased stratifi cation also entails serious consequences on dissolved oxygen concentrations (O2) (Keeling et al. 2012). Systematic deoxygenation of the ocean will have dramatic implication for ocean productivity since O2 saturation affects the functioning of marine ecosystems as it drives the biogeochemical cycles of most seawater constituent and determines the fate of all aerobic marine life (Keeling et al. 2012).

As a consequence of increasing temperatures in high latitudes, which are rising faster than the global temperature average, sea-ice extent has declined dramatically in the Arctic (7.4% per decade since 1978) (Bindoff et al. 2007) and along the western Antarctic Peninsula (Stammerjohn et al. 2008). Thermal expansions of the oceans as well as increased meltwater and discharged ice from terrestrial glaciers and ice sheets have increased ocean volume and hence, sea level (Guldberg and Bruno 2010). Climate warming affects regional wind patterns and thus ocean circulation in multiple dimensions (Doney et al. 2012). A shift in the balance between evaporation and precipitation regime has led to an increase of sea-surface salinity in low latitude regions, while high latitude waters have become fresher due to both increased precipitation and melting of the ice (Bindoff et al. 2007). Warmer oceans also drive more intense storm systems and other changes in the hydrological cycle, increasing the vulnerability of coastal habitats.


Effects of global warming on marine zooplankton

Temperature is one of the most important physical variable structuring marine ecosystems (Richardson 2008). Changes in temperature may alter the physiological performance, behavior, and demography of organisms, leading to shifts in the size structure, distributional range, and seasonal abundance of planktonic populations (Doney et al. 2012) (Fig. 2). Additionally, these shifts alter species interactions and trophic pathways from primary producers to upper trophic levels (Doney et al. 2012). As any environmental condition shift, organisms initially respond based on physiological and behavioral plasticity (Somero 2012). The new condition may be physiologically acceptable, allowing acclimatization (the process by which an organism adjusts to a gradual change in its environment) or adaptation (the evolutionary process whereby an organism becomes better able to live in its habitat), or may be intolerable, promoting migration, changes in phenology and local extinction (Parmesan 2006). Environmental change may benefi t some organisms or populations due to greater availability of food or nutrients, reduced physiological costs of maintenance (e.g., energy used for respiration, acid-base balance, calcifi cation), or reduced competition or predation (Parmesan 2006). Such species may experience higher survival, growth, and reproduction. In many cases, however, a shift can be stressful for some organisms, causing suboptimal physiological performance, higher mortality, reduced growth, and reduced reproduction (Parmesan 2006).

Fig. 2. Summary of the main effects of climate-dependent changes on marine zooplankton.


Metabolic rate of ectothermic organisms rises exponentially with temperature, leading to higher rates of physiological processes, including photosynthesis and respiration, within the range of temperatures that an organism tolerates (Doney et al. 2012). The effect of temperature on a biological process is traditionally expressed as Q10, which quantifi es temperature dependence across a limited temperature range (i.e., 10°C) (Gillooly et al. 2001). It might be expected that primary production, as well as the growth rates of ectothermic animals will increase in a warmer ocean (Doney et al. 2012). However, nutritional status, thermal tolerance, oxygen availability, elemental stoichiometry, food availability, among other factors may limit growth and production, or other biological processes, regardless of metabolic rate (Doney et al. 2012). In heterotrophic organisms, warmer temperatures raise basal metabolic rates but can also raise respiratory demand, potentially reducing their aerobic capacity (e.g., feeding, predator avoidance, digestion) and leading to less energy for growth and reproduction (Portner and Knust 2007).

At the population and community levels, individual physiological responses to global warming are evident as shifts in structure and abundance, spatial distribution of organisms and timing of annually recurring events (e.g., phenology) (Doney et al. 2012):

Changes in zooplankton community structure and abundance. Hydroclimatic changes can exert signifi cant effects on the size structure, taxonomic composition and diversity of zooplankton communities since these features are regulated by their physical and chemical environment (Richardson 2008). On a global scale, plankton community would exist in a continuum of states between two extremes, the cold, well-mixed, high-nutrient environment and the warm, stably stratifi ed, nutrient-poor environment (Schultz 2008). Falkowski (2003) used the terms “perturbed regime” and “balanced regime” to distinguish between these two systems. In cool waters with relatively strong turbulence and well-mixed conditions, surface waters are full-up of nutrients. In this perturbed regime, plankton community is dominated by large centric diatoms and large crustaceans like copepods. The food chain is short and highly effi cient, and supports a large number of planktivorous and piscivorous fi shes, seabirds and mammals (Ryther 1969, Iverson 1990, Pauly and Christensen 1995, Richardson 2008). In the balanced regime, the warmer and more stratifi ed waters have limited concentrations of nutrients. Increased heating can enhance existing stratifi cation, reducing the availability of nutrients in the surface (Richardson and Schoeman 2004). Under such conditions, plankton community is dominated by picoplankton and fl agellates, which are mostly grazed by heterotrophic protist, small crustaceans and gelatinous zooplankton (Ryther 1969, Iverson 1990, Pauly and Christensen 1995, Richardson 2008). This long and ineffi cient foodweb


has lower nutritional quality, supporting less production at higher trophic levels. Summarizing, nutrients concentration in marine environment is the main factor defi ning the confi guration of local food web, and water temperature is a valuable proxy for nutrient enrichment. If the predictions are true and the global temperature rises 1° to 2°C in the next 40 years increasing stratifi cation, the impact on biological communities could be devastating (Roemicch and McGowan 1995).

Climatic fl uctuations have had profound impacts on the abundance of planktonic species (Mackas et al. 1998, Beaugrand et al. 2002, Stenseth et al. 2002, Parmesan and Yohe 2003, Richardson and Schoeman 2004, Perry et al. 2005, Chiba et al. 2006). The coincidence of oceanic temperature rise and the decline in zooplankton densities in diverse aquatic systems are suggestive of a direct causal relationship. Changes in the abundance of some planktonic organisms off the coast of California have been well documented over the past few decades (Hughes 2000). The surface waters of the California Current have warmed by 1.2–1.6°C in approximately 40 years. This warming was accompanied by an 80% decline in zooplankton abundance (Roemmich and McGowan 1995), possibly because increased surface temperatures reduced the upwelling of cold, nutrient rich waters (Hughes 2000). As a consequence, Puffi nus griseus, one of the top predators in the system, suffered a 90% reduction in abundance off western North America (1987–1994) (Hughes 2000). In the Northeast Atlantic, Richardson and Schoeman (2004) also evidenced the effect that sea surface warming has on stratifi cation and plankton dynamics. Phytoplankton abundances were higher with warming of cool, windy, and well-mixed regions. Warmer temperatures increase metabolic rates and water stratification, thus increasing the residence time of phytoplankton cells in the euphotic zone (Richardson and Schoeman 2004). In contrast, phytoplankton abundances decreased in warm regions that become even warmer, probably because heater surface water blocks further nutrient-rich deep water from rising to the euphotic layer (Richardson and Schoeman 2004). The increased phytoplankton abundances in cooler regions and the opposite trend in warm regions was thus highly correlated with changes in the densities of primary (herbivores) and secondary consumers (carnivores) (Richardson and Schoeman 2004).

Although most of the evidence of climate impacts on zooplankton community structure and abundance is from the Northern Hemisphere, there are dramatic changes documented in the Southern Ocean (Richardson 2008). The Antarctic krill Euphausia superba is the primary prey for many predators in Antarctic waters, supporting commercial fi sheries (Atkinson et al. 2004). It has a key status in the Southern Ocean and occupies a central place in commercially valuable food webs (Meyer et al. 2003). Since the 1970s, there has been a decline in krill density and a concomitant increase


in salps, which occupy less productive and warmer regions (Atkinson et al. 2004). These changes are likely the consequence of global warming. Summer phytoplankton blooms and the extent of winter sea ice are the key factors triggering the high krill densities observed in the southwest Atlantic Ocean (Atkinson et al. 2004). In fact, krill larvae as well as the recruitment to adult stocks depend on phytoplankton blooms at the margins of sea ice cover (Atkinson et al. 2004, Richardson 2008). As waters have warmed, the extent of winter sea ice and its persistence have declined, leading to lower larval survival and explaining the observed decline in krill density. As krill densities decreased, salps appear to have synchronously increased in the southern part of their range distribution. These changes have had profound effects within the Southern Ocean food web, especially the populations of baleen whales, fi shes, penguins, seabirds, and seals that depend upon krill as their primary food source (Richardson 2008).

Population outbreaks of gelatinous zooplankton have been increasingly detected in recent years in many marine ecosystems (Mills 2001, Attrill et al. 2007, Brotz et al. 2012). Jellyfi sh and ctenophore blooms are part of the natural seasonal cycle of these species (Boero et al. 2008). Nevertheless, climate warming has been suggested as one of the main driving forces for changes in the abundance of gelatinous plankton, given that warmer temperatures can trigger greater and more rapid production of many species (Purcell 2005). As gelatinous organisms are key predators of other zooplankton species, including fi sh eggs and larvae (Purcell and Arai 2001), an increase in their populations could implicate the disruption of pelagic ecosystems (Mills 2001, Oguz et al. 2008).

The effects of physical/chemical changes due to climate change are transmitted through networks of interacting organisms to shape the structure of communities and the dynamics of ecosystems (Shurin et al. 2012). Biological systems are generally controlled by their top predators through top-down control, by their producers through bottom-up control, or by a number of key species in the middle through wasp-waist control (Cury et al. 2000). Strong bottom-up control results in a positive correlation between predator and prey whereas strong top-down control, results in a negative correlation (Richardson and Schoeman 2004). A variety of evidences suggest that increased temperatures may affect the sensitivity of food webs to top-down and bottom-up forcing (Shurin et al. 2012). For example, organisms at different positions within aquatic food webs have a specifi c sensitivity to temperature, leading to imbalanced responses to temperature change among trophic levels (Shurin et al. 2012). More active primary consumers may exert stronger top-down effects on producers; however, their greater metabolic demands may intensify resource limitation and reduce their abundances, leading to weaker effects at the long-term population level (Shurin et al. 2012). The close coupling between trophic


levels in pelagic ecosystems implies that the impact of future climate change will permeate the entire marine food webs. Ocean temperature is likely to be further affected by anthropogenic climate change. The IPCC predicts a rise in temperature of between 2 and 4°C in the northeast Atlantic by 2100 (Richardson and Schoeman 2012). The effect of climate change will have severe impacts on phytoplankton community, herbivorous copepods and carnivorous zooplankton, thereby affecting ecosystem services, such as oxygen production, carbon sequestration, and biogeochemical cycling (Richardson and Schoeman 2012). Finally, fi shes, seabirds, and marine mammals will need to adapt to a changing spatial distribution of primary and secondary production within pelagic marine ecosystems (Richardson and Schoeman 2012).

Changes in zooplankton distributional ranges. Distributions and/or abundances of numerous species have been extensively altered by human activities (e.g., habitat loss, ecosystem alteration) (Hughes 2000). However, some distributional shifts are explained more by an association with changing climatic conditions, especially when the shift has been towards the poles (Hughes 2000). Fossil evidence shows that marine organisms shifted polewards as sea surface temperatures raised, e.g., during the Pleistocene-Holocene transition (Harley et al. 2006). Although relatively few in number, long-term ocean biological data indicate that zooplanktonic organisms exhibit fast and large shifts in their ranges in response to global warming (Richardson 2008). Most cases correspond to species whose distributions are mainly driven by climate or organisms that are highly mobile at some stage of their life cycle (Hughes 2000).

Several copepod species have already modifi ed their habitat ranges in response to climate warming. Most of the examples are from the North Atlantic, where the Continuous Plankton Recorder survey (CPR) has been operating since 1931 (Richardson et al. 2006). The CPR survey provides a unique long-term dataset of oceanic plankton abundance in the North Atlantic and North Sea (Warner and Hays 1994). It has been running for almost 70 years sampling at a depth of 10 meters. In 1998, Calanus hyperboreus was recorded at its farthest southern limit in the CPR survey, 39°N, off the Georges Bank shelf edge (Johns et al. 2001). These authors suggested a direct response of this species to a cooling of the surrounding environment. Ocean climate in the Northwest Atlantic is driven by thermohaline mechanisms, and these infl uence the south-fl owing Labrador Current (Richardson 2008). The Labrador-Newfoundland area experienced abnormally cold temperatures during the late 1980s and early 1990s (Prinsenberg et al. 1997) which increased the production of Labrador seawater and thus the strength of the Labrador Current (Dickson 1997). This cold water has spread farther


south, bringing colder conditions to an area formerly unfavorable for C. hyperboreus (Johns et al. 2001, Richardson 2008).

Large-scale changes in the biogeography of calanoid copepods in the Northeast Atlantic Ocean and adjacent seas were also attributed to regional changes in sea surface temperature (Beaugrand et al. 2002). Strong distributional shifts in all copepod assemblages have occurred with a northward extension of more than 10° latitude of warm-water species associated with a decrease in the number of colder-water species (Beaugrand et al. 2002). The biological associations showed consistent long-term changes that appear to refl ect a movement of marine ecosystems towards a warmer dynamical regime (Beaugrand 2005). As an example, the cool-water assemblage is dominated by Calanus fi nmarchicus, a large calanoid copepod species (Richardson 2008). As water warmed over recent decades and the assemblage retracted northward, this species has been replaced by Calanus helgolandicus, the dominant species of the warm-water assemblage (Richardson 2008). Given that larval stages of Atlantic cod feed on C. fi nmarchicus, the replacement of this species could have a detrimental effect on cod stocks because both copepods are abundant at different times of the year (Beaugrand et al. 2003). In fact, cod recruitment decreased from the mid-1980s, coincident with unfavorable change in the plankton ecosystem. All of these changes in the plankton ecosystem may be the cause of temporal predator-prey decoupling and hence, reduced cod recruitment (Beaugrand 2005).

The distribution of two individual copepod species in the Northeast Atlantic has also been studied in relation to ocean warming (Lindley and Daykin 2005). Centropages chierchiae and Temora stylifera both moved north from the vicinity of the Iberian Peninsula in the 1970s and 1980s to the English Channel in the 1990s (~ 6° of latitude). Concurrent with the expansion polewards of warm-water copepods, the Arctic assemblage has retracted to higher latitudes (Beaugrand et al. 2002). Although these translocations have been associated with regional warming of up to 1°C, they may also be partially explained by stronger north-fl owing currents on the European shelf edge. These shifts in distribution have had dramatic impacts on the foodweb of the North Sea (Beaugrand et al. 2003).

Biogeographical shifts may have severe consequences for exploited resources, especially fi sheries. During the last 20 years, there has been an increasing interest in the scientifi c community in understanding the relationship between zooplankton and climate change due to the fact that several marine fi sh and invertebrates feed on zooplankton at some stage of their life (Drinkwater et al. 2003). If changes continue, it would lead to important modifi cations in the abundance of fi sh, with a decline or even a collapse in the stock of species (Beaugrand et al. 2002).


Changes in zooplankton phenology. Phenology is a sensitive indicator of global warming (Edwards and Richardson 2004, Richardson 2008). Since the level of response to climate change varies across functional groups and trophic levels, changes in annually recurring life cycle events may be of vital importance to ecosystem functioning (Edwards and Richardson 2004). The decoupling of annually recurring events will have severe consequences for trophic interactions, changing food-web structures and leading to eventual ecosystem-level changes (Edwards and Richardson 2004). In the case of temperate marine environments, where the recruitment success of higher trophic levels is highly dependent on synchronization with planktonic pulses, these changes can dramatically affect community connectivity (Edwards and Richardson 2004, Costello et al. 2006). More crucial than any change in timing of a single species is the potential disruption of coordination in timing between the life cycles of predators and their prey (Parmesan 2006). A fundamental concept in aquatic ecology establishes that the fi tness of a predator depends upon its temporal and spatial synchrony with the production of its prey (Cushing 1990). Ecologists have also observed drastic population decline in predators when predator-prey relationships are disrupted through climate-related perturbations (Winder and Schindler 2004).

In the Narragansett Bay, USA, phenological alterations concerning the copepod Acartia tonsa and the ctenophore Mnemiopsis leidyi have provided an opportunity to examine the mechanisms that underlie species-specifi c responses to climate warming in estuarine ecosystems (Costello et al. 2006). The change in seasonal timing of population growth by M. leidyi relative to A. tonsa has altered summer zooplankton dynamics in the central region of the bay (Costello et al. 2006). The advance in M. leidyi’s seasonal appearance (59 days between 1951 and 2003) has shifted the predator’s peak abundance into a time period during which A. tonsa has historically enjoyed a temporal refuge from ctenophore predation (Costello et al. 2006). Before climatic warming, A. tonsa was the dominant secondary producer in the estuary and its main period of production occurred before the seasonal appearance of M. leidyi. However, since 2000, the seasonal peak abundances of the two species have overlapped, intensifying the predator–prey relationship and resulting in the near extirpation of the once-abundant copepod from the estuary (Costello et al. 2006).

In the surface waters of the Subarctic North Pacifi c Ocean, the copepod Neocalanus plumchrus, which dominates the zooplankton biomass, has a seasonal cycle of abundance that is tightly coupled with sea surface temperature (Mackas et al. 1998). Its vertical distribution and development are strongly seasonal, with an important and relatively short (<60 days) annual maximum in spring and early summer. This peak has shifted dramatically between 1956 and the present (Mackas et al. 1998). Population


development was very late in the early 1970s, early in the late 1950s, and very early in the 1990s (Mackas et al. 1998). These changes in timing were strongly correlated with large-scale year-to-year and decade-to-decade ocean climate fl uctuations, as refl ected by spring season temperature anomalies in the surface mixed layer in which juvenile copepodites feed and grow (Mackas et al. 1998). The change in developmental timing is probably a consequence of both increased survival of early cohorts in warm years and physiological acceleration (Mackas et al. 1998).

Using long-term data (1958–2002), Edwards and Richardson (2004) detected signifi cant phenological changes in zooplankton community in the central North Sea. The timing of temporary members of the zooplankton (meroplankton, organisms that are planktonic for only a part of their life cycle), might be affected more by warming sea temperatures than permanent members (holoplankton) (Edwards and Richardson 2004). During summer, meroplankton as a whole (larvae of cirripeds, cyphonautes, decapods, echinoderms, fi sh larvae, and lamellibranchs) has anticipated their appearance in the plankton by 27-days over the 45-years study period (Edwards and Richardson 2004). In the case of copepods and non-copepod holoplankton, they have both moved forward only by 10-days. Organisms that are dependent upon temperature to stimulate physiological developments and larval release have signifi cantly moved forward in their seasonal cycle in response to temperature (Edwards and Richardson 2004). The level of response differed throughout the community and the seasonal cycle, leading to a mismatch between trophic levels and functional groups (Edwards and Richardson 2004). The different extent to which functional groups are moving forward in time in response to warming has led to a mismatch between successive trophic levels and a change in the synchrony of timing between primary, secondary and tertiary production (Edwards and Richardson 2004).

Marine zooplankton and ocean acidifi cation

Acidifi cation of the ocean is another effect of global change linked to CO2 emissions. Almost 50% of anthropogenic CO2 emitted in to the atmosphere diffuses passively into the ocean, and when CO2 dissolves in the seawater, it causes alterations in fundamental chemical balances that together are commonly referred to as ocean acidifi cation (Fig. 1) (e.g., Caldeira and Wickett 2003, Sabine et al. 2004, Doney et al. 2009, Feely et al. 2010). Such changes comprise increase in the concentration of carbonic acid (H2CO3), bicarbonate ion (HCO3

–) and hydrogen ion (H+), and decreases in the concentration of carbonate (CO3

2–) in surface waters, where the bulk of oceanic production occurs. Thus, production of H+ lowers the pH and causes the phenomenon called ocean acidifi cation. Carbonate ions can react with the excess H+ to


form HCO3–, and the carbonate buffering system would allow oceans to

retain a stable pH despite rising emissions. However, as the partial pressure of carbon dioxide (pCO2) increases, the buffering capacity of seawater decreases, in fact the ocean pH has been reduced by 0.1 since the start of the industrial revolution, representing a 30% increase in the concentration of H+ ions (Caldeira and Wickett 2003, Key et al. 2004, Doney et al. 2009).

Effects of ocean acidifi cation on marine zooplankton

It is considered that the concentration of carbon dioxide in the surface ocean was more or less in equilibrium with overlying atmosphere CO2; however, researchers largely dismissed the potential impact on the ocean biota because calcite, the CaCO3 mineralogy of most marine calcifying organisms, would remain supersaturated in the surface ocean (Doney et al. 2009). Since then, multiple studies consider ocean acidifi cation as a threat to marine biota. Most of them found that:

• The calcifi cation rates of many shell-forming organisms respond to the degree of supersaturation of CaCO3 minerals (e.g., Smith and Buddemeier 1992, Kleypas et al. 1999).

• Aragonite, a more soluble CaCO3 mineral in calcifying organisms, may become undersaturated in the surface ocean within the early 21st century (Feely and Chen 1982, Feely et al. 1988, Orr et al. 2005).

• The biological effects of decreasing ocean pH reach far beyond limiting calcifi cation (e.g., Mayor et al. 2007, Fabry et al. 2008).

Therefore, in marine organisms, physiological processes such as growth, development, metabolism, ionoregulation and acid-base balance, can be affected by increases in CO2 (Fabry et al. 2008, Pörtner and Farrell 2008, Pörtner et al. 2004, Widdicombe and Spicer 2008) (Fig. 2). However, most of the studies concerning ocean acidifi cation on marine organisms are focused on calcifying organisms which produce CaCO3 shells or skeletons (Gattuso et al. 1998, Kleypas et al. 2006, Riebesell et al. 2000) and the reduced availability of carbonate ions which affects calcifi cation processes. These works include the dissolution of calcifying plankton but also the reduced growth and shell thickness in gastropods and echinoderms and declining growth of reef-building corals.

Effects on plankton food chain. Ocean acidifi cation can produce indirect impacts on zooplankton, since it may change the biochemical composition of phytoplanktonic preys, affecting its nutritional food quality. Elevated concentrations of CO2 can encourage carbon fi xation in primary producers and thus reduce the nutrient content relative to carbon (Rossoll et al. 2012). Under elevated CO2 scenario, enhanced carbon consumption relative to nutrients, can produce changes in the phytoplankton stoichiometry thus


altering the food demand of predators. Additionally, fatty acid (FA) is an important factor that regulates the transfer of energy from phytoplankton through zooplankton since heterothrophic organisms cannot synthesize FAs and have to incorporate them through the diet. Some long-chain polyunsaturated FAs (PUFAs) are essential for growth, development and reproduction success. Acidifi cation may affect the production of FA by phytoplankton because several physiological parameters that infl uence enzyme activity are affected. Although the direct response of copepods exposed to increased CO2 concentration is weak (Mayor et al. 2007, Kurihara and Ishimatsu 2008), it seems likely that CO2 indirectly affect zooplankton growth through its impact on the nutritional quality of phytoplankton. For instance, elevated CO2 concentrations signifi cantly changed the FA concentration and composition of the diatom Thalassiosira pseudonana that constrained the growth and reproduction of A. tonsa, leading to decreased somatic growth and egg production (Rossoll et al. 2012).

On the other hand, when the nutritional quality of phytoplankton declines, copepod’s diet might switch to be dominated by microzooplankton (ciliates and heterotrophic dinofl agellates) which can upgrade the food offer for higher trophic levels by buffering nutritional imbalances at the interface between primary production and consumers (Malzahn et al. 2010). This phenomenon becomes more evident at the end of the phytoplankton bloom when dissolved nutrients decay and zooplankton like copepods change from autotrophic to microheterotrophic diets.

Additionally, as pCO2 increase in seawater, dissolved CO2 readily diffuses across animal surfaces and equilibrates in both intra- and extracellular spaces. As in seawater, CO2 reacts with internal body fl uids causing H+ to increase and, therefore, pH to decrease (Fabry et al. 2008). Mechanisms available to counteract this acidifi cation are limited and relatively conserved across animal phyla. These changes can have positive and negative effects on the growth of marine plankton, with a corresponding impact on their role as a net source or sink of CO2 to the atmosphere. The production of organic matter during phytoplankton photosynthesis predominantly utilizes CO2 dissolved in seawater and so provides a sink for atmospheric CO2. It is expected that phytoplankton photosynthesis will be stimulated by the increases in dissolved CO2 associated with the predicted doubling of atmospheric CO2. Phytoplankton have differing sensitivities to CO2 concentration and have a variety of mechanisms for carbon utilization. Thus, increases in seawater CO2 concentration will not only change the activity of individual phytoplankton species, but will also tend to favor some species over the others. These shifts in phytoplankton community structure will infl uence the community structure of the higher trophic levels that are reliant upon phytoplankton as food and will also infl uence the cycling of elements that differ between species (e.g., carbonate by calcifying organisms and


silicate by non-calcifying organisms). Furthermore, the activity of bacteria (which produce CO2) and the zooplankton (which consume phytoplankton) might also be affected by pH, resulting in changes in the structure and functioning of the marine ecosystem as a whole.

Effects on early developmental stages. Ocean acidifi cation can also produce effects on early developmental stages, and studies on this subject are very important since larvae and juveniles are generally more vulnerable to environmental perturbations, and their survival will largely determine population abundance, distribution and community confi guration. For instance, Kawaguchi et al. (2011) found that high concentrations of CO2 impacted severely on the embryonic development of krill (E. superba) embryos and larvae experimentally exposed to different pCO2. They found that at high concentrations development was disrupted before gastrulation in 90% of embryos, and no larvae hatched successfully. In addition, they pointed out an urgent need for understanding the pCO2-response of later stages of krill, in order to predict the possible fate of this key species in the Southern Ocean ecosystems.

Mac Donald et al. (2009) studied the effect of pH on planktonic larval stages of the barnacle Amphibalanus amphitrite, and revealed no effects of reduced pH on larval condition, cyprid size, cyprid attachment and metamorphosis, juvenile to adult growth or egg production. However, barnacles exposed to pH 7.4 showed overcalcifi cation at the lower, active growth regions of the wall shells. Despite this enhanced calcification, further studies revealed that the central shell wall plates required significantly less force to be penetrated than those of individuals raised at pH 8.2. Therefore, barnacles with weakened wall shells are more vulnerable to predators.

Some copepod species appear more tolerant to increased CO2 than other marine organisms (i.e., sea urchins, bivalves, barnacles, amphipods). Kurihara and Ishimatsu (2008) found that high CO2 exposure through all life stages of the 1st generation of Acartia tsunesis copepods did not signifi cantly affect survival, body size or developmental speed. Egg production and hatching rates were also not signifi cantly different between the initial generation of females exposed to high CO2 concentration and the 1st and 2nd generation females developed from eggs to maturity in high CO2. Mayor et al. (2007) also found that ocean acidifi cation did not affect the survival of Calanus fi nmarchicus adults. In fact, growth and egg production of adult females was not affected by experimentally-simulated ocean acidifi cation. In contrast, a maximum of only 4% of the eggs successfully yielded nauplii after 72 h in the experimental treatment. The authors demonstrated that environmental risk assessments for marine CO2 disposal must look beyond adult mortality as an endpoint. Furthermore, they conclude that if CO2 is to be disposed of in the deep sea, the location and timing of such activities must take into consideration the overwintering populations of C. fi nmarchicus.


Effects on other non-skeletal, calcifi ed secretions of marine fauna. The use of calcium minerals in gravity sensory organs is widespread in many zooplankton organisms that possess statoliths, statocysts, or statocontia. Thus, changes in the carbonate chemistry of seawater would affect mineralization of the various types of gravity receptors and so might impact the overall fi tness of organisms. Fabry et al. (2008) hypothesized that the potential impacts would depend on the ability of the organisms to regulate the acid–base balance in the tissues surrounding those structures. Other carbonate secretions of marine fauna that could be impacted include gastroliths, mineralized structures formed in the lining of the cardiac stomachs of some decapods that serve as storage sites for calcium during moult intervals (cf. Lowenstam and Weiner 1989).

Effects on other physiological processes. Acidosis and hypercapnia occur as a consequence of the disruption of the acid-base balance of marine animals, and are the detrimental result of low seawater pH and high pCO2, respectively (Pörtner et al. 2004). These homeostatic disorders influence most physiological processes (Roos and Boron 1981), such as neural signaling (Waggett and Buskey 2008), development, reproduction (Kikkawa et al. 2004), metabolism and even gene expressions (Pörtner et al. 2010) as well as behavior (Thistle et al. 2007). Regulation of body acid–base depends upon energy and acidosis or hypercapnia can affect the energy acquisition and allocation (Whiteley 2011). Recently, Li and Gao (2012) analyzed the possibility that marine secondary producers increase their respiration and feeding rate in response to ocean acidification to balance the energy cost against increased acidity and CO2 concentration. They found that the copepod Centropages tenuiremis can perceive chemical changes in seawater under elevated CO2 concentration with avoidance strategies. The copepod’s respiration and its feeding rate increased at the elevated pCO2 (1000 uatm) and its associated acidity (pH 7.83). The ability to perceive the chemical changes in seawater and to escape from the adverse circumstance is the key for C. tenuiremis to survive in coastal waters where pH changes are intensifi ed by eutrophication and atmospheric CO2 rise (Cai et al. 2011). To respond to and cope with the increased external acidity, the copepod C. tenuiremis increased its food acquisition to compensate the extra energy demand via enhanced respiration.

Effects on zooplankton behavior. Studies relative to acidifi cation effects on zooplankton reproduction reveal changes on swimming and mating behaviors as well as mating success. Specifically, ocean acidification decreases the ability of male copepods to detect, track and capture a female. Seuront (2010) found that the level of ocean acidifi cation expected to occur in 2100 (i.e., pH = 7.8 to 7.6) signifi cantly modifi es the stochastic properties of successive displacements of the copepods Eurytemora affi nis


and Temora longicornis. This phenomenon caused lower mate encounter rates since copepods cannot rely on female pheromone plumes (i.e., under turbulent conditions) and also impaired the ability of males to detect females pheromone trails, to accurately follow trails and to successfully track a female. These results indicate that ocean acidifi cation decreases the ability of male copepods to detect, track and capture a female, hence suggest an overall impact on population fi tness and dynamics.

Effects on pollutants toxicity. In addition, acidifi cation may increase the potential effects of pollutants in marine organisms, especially in coastal areas. For example, ocean acidifi cation will change the organic and inorganic speciation of metals and will modify interactions of metals with marine organisms such as zooplankton. As mentioned above, a consequence of ocean acidifi cation is a decreased concentration of OH– and CO3

2–. These anions form strong complexes in ocean water with divalent and trivalent metals (Millero et al. 2009). This reduction is expected to change the speciation of numerous metal ions in seawater (Byrne 2002). Experiments carried out with two copepods species in laboratory experiments showed that they were more sensitive to acidity by increasing CO2 concentration (Pascal et al. 2010). These authors also found that CO2 enrichment increased the free-ion concentration of some metals such as Cu, altering its toxicity, since free-ion forms of metals are generally more toxic than complex forms. In addition, the same authors found antagonistic toxicities of CO2 with Cd, Cu and Cu free-ion in harpacticoid copepod like Amphiascoides atopus and they concluded that this interaction could be due to a competition for H+ and metals for binding sites.

Marine zooplankton and UV radiation

The Earth’s surface is protected from solar ultraviolet radiation (UV) by the stratospheric ozone. The anthropogenic emission of chlorofl uorcarbons (CFCs) has been recognized as the principal factor that produces the depletion of the ozone layer (Kerr and McElroy 1993). As a consequence of this decrease of the stratospheric ozone, UV radiation increases its incidence on the Earth’s surface (De Mora et al. 2000).

The ultraviolet spectral region is constituted by wavelengths in the range of 100–400 nm. Wavelengths range of 280–315 nm corresponds to UV-B whereas UV-A is defi ned by a range of 315–400 nm. Both wavelengths ranges are absorbed by O3 being UV-B more harmful than UV-A to Earth’s surface (Madronich et al. 1995, 1998). UV-B radiation can lead to substantial biological effects since biological responses to UV-exposure are far greater at shorter wavelengths (Madronich et al. 1998). Global long-term data observations on UV-B levels established a signifi cant increase at mid and


high-latitude areas of the Northern and Southern Hemisphere (Kerr and McElroy 1993, Browman et al. 2000, McKenzie et al. 2003). UV-B effects on Earth’s surface are especially pronounced in the Antarctic and Arctic regions but are not signifi cant in tropics (Madronich et al. 1998).

The penetration of UV into the water column depends upon the concentration and composition of dissolved organic matter (DOM) along with non-living particles (De Mora et al. 2000). Solar radiation is absorbed only by specifi c portions of DOM. Within DOM, the most important substances involved in the absorption of biological harmful UV, are humic substances, which are largely composed of chromophoric dissolved organic matter (CDOM) (Wetzel 2003). Aquatic environments differ enormously in their UV attenuation. Coastal areas and shallow continental shelf waters, including coastal Arctic waters, are characterized by low penetration of UV due to the very high CDOM concentration. In these waters, photochemical processes (photodegradation of DOM) may dominate photobiological processes. In open ocean, the low CDOM concentration allows higher UV penetration. In this system, photobiological processes (i.e., effects of UV on DNA), which can occur at signifi cant depth (20 m), may dominate photochemical processes (photodegradation of CDOM). Signifi cant penetration of UV is observed in Antarctic waters, particularly during episodes of thinning of the ozone layer, resulting in enhanced photochemical/photobiological processes (Tedetti and Sempere 2006).

Variability in system’s steady state as changes in turbidity, vertical mixing, migration rates, and seasonality and location of spawning can influence substantially the actual UV-damage on fresh and marine organisms. Several studies have demonstrated that solar UV, mostly UV-B radiation has a wide range of harmful effects on aquatic ecosystems (Häder and Worrest 1991, Häder et al. 1998, Kouwenberg et al. 1999a, b, Browman et al. 2000, Browman 2003, Tedetti and Sempere 2006, Häder et al. 2007). In aquatic organisms, UV radiation produces several effects as a reduction in productivity, decrease in reproduction, and development and increase in DNA mutations and damages (Browman 2003, Häder et al. 1998, 2003, 2007). Changes of solar UV radiation produce a decrease in biomass productivity which is transferred through the entire food webs and may also affect biogeochemical cycling within aquatic systems (Häder et al. 1998, 2003, 2007, Browman 2003).

Direct effects of UV radiation on zooplankton

UV-B radiation has been reported to have negative effects on zooplankton, inducing mortality, especially in early stages, reducing survival and fecundity in females, and changing sex ratios (e.g., Karanas et al. 1979, Chalker-Scott 1995, Scott et al. 1999, Alonso Rodriguez et al. 2000, Browman et al. 2000, Browman 2003) (Fig. 2). Several studies have also documented


UV-A deleterious effects on aquatic organisms (Alonso Rodriguez et al. 2000, Browman et al. 2000, Ban et al. 2007).

UV-B radiation cause negative effects on various copepods species (Karanas et al. 1979, Naganuma et al. 1997, Lacuna and Uye 2001). For instance, experiments on Acartia clausii, a dominant copepod of coastal and estuarine zooplankton community of the Northern Hemisphere, revealed that UV-B affects both survival and reproduction rate of this copepod (Karanas et al. 1979). The embryogenesis, survival of nauplii and copepodites, feeding and egg production of adult Acartia omorii were affected by artifi cial UV-B radiation. Moreover, the relatively high UV-B exposure of eggs produced malformations of the nauplii (Lacuna and Uye 2001). In contrast, no harmful effect was detected under UV-A radiation on this copepod (Lacuna and Uye 2001). Similar results of the UV-B effects were observed on Calanus sinicus (Naganuma et al. 1997), Calanus fi nmarchicus (Kouwenberg et al. 1999b) and Sinocalanus tenellus (Lacuna and Uye 2000).

Negative effects of UV radiation on C. fi nmarchicus eggs have been observed on fi eld experiments in the Estuary and the Gulf of St. Lawrence (Alonso Rodriguez et al. 2000, Browman et al. 2000). The results suggested that under the current levels of exposure, UV-B radiation is negatively affecting C. fi nmarchicus eggs occurring in the upper layers of the ocean. Furthermore, the hatching of eggs of C. fi nmarchicus exposed to UV-B and UV-A was not different from that exposed only to UV-A. In contrast, UV-A radiation seemed to be more detrimental on embryos of C. fi nmarchicus than UV-B (Alonso Rodriguez et al. 2000, Browman et al. 2000). In an analogous experiment, embryos of the Atlantic cod Gadus morhua exposed to UV-B radiation exhibited a higher rate of mortality, presumably as a result of DNA damage (Browman et al. 2000). In accordance, both calyptosis larvae of E. superba and copepodites of Calanoides acutus and Calanus propinquus were susceptible to natural UV-B and UV-A (Ban et al. 2007). The effects of UV-B radiation in different stages of these species were related to DNA damage, while the UV-A radiation effects consisted of the formation of hydroxyl radicals that accumulated in cells and caused oxidative damage to membrane lipid and other cellular components (Beer et al. 1993). The modifi cation of the DNA structure is at the base of all detrimental UV effects in living organisms. The DNA damage generated by UV-B is produced by the accumulation of cyclobutane pyrimidine dimers (CPDs) (Malloy et al. 1997). High levels of DNA damage detected in eggs and larvae of icefi sh Chaenocephalus aceratus have been attributed to UV-B irradiance, and CPD content was correlated with the daily incidence of UV-B irradiance (Malloy et al. 1997). Additionally, Brownam et al. (2003) observed that the copepod C. fi nmarchicus was much more susceptible to CPD formation than the cod


G. mohua. Eggs in both species were less sensitive to accumulated CPD and thus resulted in less damage than hatched larvae (Brownam et al. 2003).

Evidence of the harmful effect of UV-B radiation was also observed on early life stages of marine decapods and fi shes (Hovel and Morgan 1999, Wübben 2000). The larval survival of three estuarine crab species (Uca pugnax, Sesarma reticulatum and Dyspanopeus say) decreased by the exposure of UV-B radiation (Hovel and Morgan 1999). Short-experiments on Crangon crangon demonstrated that moderate levels of UV-B caused lethal effects on zoea I. Long-term exposure (0–10 days) of this crab to UV-B radiation produced no signifi cant moulting and only 10% of these larvae moulted from zoea I to zoea II (Wübben 2000). Lethal effects of UV-B on larvae and embryos of the fi sh species Engraulis mordax and Scomber japonicus were also reported by Hunter et al. (1979).

In the Southern Hemisphere, studies on the effect of UV radiation and its consequences on zooplankton were performed. Several studies have been carried out on freshwater zooplankton reporting lethal effects and photo-repair activity (Gonçalves et al. 2002, 2007, 2010). Nevertheless, research on marine southern zooplankton concerning UV-B impact needs to be done.

Indirect effects of UV radiation on zooplankton

Most of the studies focus on the direct effects of UV-B radiation in specifi c organisms. Recently, a few studies in both marine and freshwater environments have examined the indirect effects of UV radiation and how long exposure to low-level UV induces changes in food web interactions (Hessen et al. 1997, Browman et al. 2000, 2003, Gonçalves et al. 2010, Williamson et al. 2010) (Fig. 2).

Indirect effects produced by UV exposure are related to changes in phytoplankton cell wall and size and species composition that affect the ingestion and digestion of zooplankton (Mostajir et al. 1999, De Langer and Lürling 2003, Grem et al. 2004). Changes on food quality as the increase of carbohydrates and the decrease of fatty acids due to the phytoplankton exposure to UV-B (Arts and Rai 1997, Goes et al. 1997), caused negative effects on growth and reproduction rates of herbivorous zooplankton (De Lange and Van Donk 1997, Scott et al. 1999). Increased mortality, decreased growth and decreased overall fecundity were observed in an experimental study on the cladoceran Daphnia pulex feeding on UV-B irradiated algae (Scott et al. 1999).

The reduction of the total lipid content of some microalgae including the PUFAs is another consequence of the exposure to UV-B radiation, even at low dose rates (Goes et al. 1994, De Lange and Van Donk 1997). Zooplankton and fi sh larvae cannot synthesize those fatty acids and they must be obtained from their prey (Tuner and Rooker 2005). The reduction


of PUFA content in microalgae by UV-B radiation will be transferred to herbivorous zooplankton that graze on them, thus decreasing the availability of this essential fatty acid. As a consequence of the reduction in the nutritional quality of preys, the productivity and health of aquatic ecosystem could be signifi cantly affected (Brett and Müller-Navarra 1997, Browman et al. 2003).

Other indirect effects of UV radiation, particularly UV-B, have been related to suppression of both systematic and local immune to a variety of antigens (Salo et al. 1998, Browman et al. 2000, Jokinen et al. 2000). All these indirect (and/or longer-term) effects of UV-B radiation need to be further investigated.

Zooplankton responses to UV radiation effects

Several strategies to avoid UV radiation such as increase in resting eggs production, vertical migration, accumulation of photoprotective compounds, and photoenzymatic repair have been already reported for zooplankton in both marine and freshwater systems (McFadyen et al. 2004, Gonçalves et al. 2010, Hylander and Hansson 2010, Hylander and Jephson 2010, Zengling et al. 2010).

Zooplankton has the ability to migrate vertically and horizontally to defend against UV radiation (Wold and Norrbin 2004, Hylander and Hansson 2010, Zengling et al. 2010). For instance, the cladocerans genus Daphnia, Chydorus and Eurycercus showed vertical migration in response to UV radiation (Hylander and Hasson 2010). Daphnia showed the strongest response to UV whereas Chydorus and Eurycercus, displayed a weak response remaining mainly at the bottom during daytime (Hylander and Hasson 2010). Vertical migration in response to UV radiation has also been observed on female and nauplii of C. fi nmarchicus which migrated downwards when were exposed to radiation (Wold and Norbbin 2004). The behavioral response to UV in these species has been attributed to the amount of photo-protective pigmentation in the organisms (Wold and Norbbin 2004, Hylander and Hasson 2010).

It is also known that zooplankters can synthesize or accumulate photo-protective compounds such as melanin, carotenoids and mycosporine-like amino acids (MAAs), which are associated with UV resistance in copepods (Hansson et al. 2007, Hansson and Hylander 2009, Hylander and Jephson 2010, Zengling et al. 2010). Melanin is the most important photo-protective compound found in cladocerans (Hansson et al. 2007). Copepods may not have the capacity of synthesize these compounds and have to obtain these photo-protective compounds, carotenoids and MAAs, mainly from their preys. Since carotenoids and MAAs are more abundant in phytoplankton, the accumulation levels of these photo-protective compounds in copepods


would be dependent on their different feeding habits (Hansson et al. 2007, Hansson and Hylander 2009, Rautio et al. 2009, Ma et al. 2012).

The synthesis and accumulation of MAAs in response to UV radiation have been simultaneously studied in both phytoplankton (dinofl agellates) and their zooplankton grazer (copepods as A. tonsa) (Hylander and Jephson 2010). The results showed that the MAAs content increased approximately four times in dinofl agellates exposed to UV compared to non-UV radiation treatment. Moreover, the elevated MAAs level in dinofl agellates was refl ected in the copepods, which accumulated more MAAs when exposed to UV in comparison to the non-UV radiation treatment (Hylander and Jephson 2010). However, protective compounds, like carotenoids, were not accumulated by these marine copepods. This last fi nding differed from other studies which showed that several species of freshwater copepods had the ability to synthesize those pigments (Zagarese et al. 1997, Hansson 2000, Hansson et al. 2007, Hylander et al. 2009). High concentrations of MAAs in Antarctic krill have also been observed after feeding on algae that had been grown under PAR-supplemented UV radiation (Newman et al. 2000).

Other mechanisms which organisms possess to prevent DNA damage induced by UV-B or to repair it after UV-exposure, are photo-enzymatic repair (PER; ‘‘light repair’’) and nucleotide excision repair (NER; ‘‘dark repair’’) (Malloy et al. 1997). DNA repair mechanisms have been studied in both freshwater and marine zooplankton (Zagarese et al. 1997, Grad et al. 2001, Browman et al. 2003, MacFadyen et al. 2004). Moreover, several studies have demonstrated the ability of marine and freshwater crustaceans to repair DNA damage via PER (Malloy et al. 1997, Zagarese et al. 1997). Additionally, it has been reported that NER is found in all taxa and is not specifi c of UV-induced DNA damage. Conversely, PER is specifi c to UV-induced DNA damage and it is not present in all taxa (Sinha and Häder 2002). Although several studies have demonstrated the deleterious effects of UV-A on aquatic organisms (Williamson et al. 1997, Alonso Rodriguez et al. 2000, Browman et al. 2000, Ban et al. 2007), the role of UV-A radiation is not as clearly defi ned as the UV-B, and appears to be involved in the photo-repair of UV-B-induced damage (Sutherland 1981, Sutherland et al. 1992).

Photorepairing of UV-B-induced damage to the DNA was found in both Daphnia menucoensis and the copepod Metacyclops mendocinus, which showed a signifi cant decrease of mortality when exposed to visible radiation, PAR in addition to UV-B (Gonçalves et al. 2002). A study on Antarctic zooplankton has indicated that during periods of increased UV-B, the accumulation of CPD levels was signifi cant and the DNA damage was largely repaired by the photoenzymatic repair system (Mallloy et al. 1997). Another study compared the vulnerability to UV-B radiation of three copepod species (Boeckella brevicaudata, Boeckella gibbosa, and Boeckella gracilipes) and showed that the potential photoprotection (i.e., resistance to UV-B in the absence


of recovery radiation) of B. gracilipes and B. gibbosa was lower than that of B. brevicaudata (Zagarese et al. 1997). The capacities to repair DNA damage induced by UV radiation differ among species and within species among developmental stages. Moreover, these differences could be related to the presence of MAAs and/or carotenoides that function as protected compounds against UV radiation damage (Malloy et al. 1997, Zagarese et al. 1997).

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CHAPTER 5

Benthic Community and Climate Change

Sandra Marcela Fiori* and María Cecilia Carcedo

Introduction

Benthic marine communities are composed of a diversity of species belonging to different taxa that live in association with the sea bottom. They can be partially or totally buried in the sediment, adhered to the bottom or move without departing too much from the substrate. Depending upon their association with substrate type (hard/soft) and depth, these communities settle and develop in a broad range of areas, from the high tide line to the bottom of the deep ocean trench. Benthic species are important for a variety of reasons:

• Although the best-known reefs are generated by corals, the skeletal remains of species like molluscs, echinoderms, polychaetes and other invertebrates are also used in reef-building (Kirtley 1968). These biogenic constructions are considered local hotspots of biodiversity: they function as important spawning, nursery, breeding and feeding areas for a multitude of organisms and provide refuge and substrate to an array of organisms including invertebrates and fi shes (Kirtley 1968, Nelson and Demetriades 1992, Lindeman and Snyder 1999, Moberg and Folke 1999).

Instituto Argentino de Oceanografía (IADO), Complejo CCT-BB-CONICET, C.C. 804, 8000 Bahía Blanca, Argentina; and Departamento de Biología, Bioquímica & Farmacia, Universidad Nacional del Sur, San Juan 670, 8000 Bahía Blanca, Argentina.

Email: [email protected]* Corresponding author: sfi [email protected]


• Benthic organisms are also important members of the lower food web, consuming organic matter and phytoplankton. Constituting a source of food for larger organisms such as many fi sh, shorebird and marine mammal species, they link primary production with higher trophic levels (Carlson et al. 1997). In addition, a number of benthic invertebrates, particularly clams, are consumed by humans and are used for recreational purposes such as fi shing bait (McLachlan et al. 1996, Thrush and Dayton 2002).

• The infauna of nutrient-poor tropical carbonate sediments play a crucial role in bioturbation, oxygenation, nutrient cycling and transport, and processing of pollutants (Snelgrove 1997, 1998, Uthicke and Klumpp 1998, Uthicke 1999). Many benthic organisms, including fi lter feeders like clams, scallops and mussels, obtain their food by taking in sea water. As the water fl ows through their bodies, sediments, organic matter, and pollutants are fi ltered out and ingested.

• The interest in benthic indicators for soft-bottom marine communities has increased dramatically as the need for new tools to assess the status of marine waters has grown (Dauvin 2012). There are a number of advantages of using benthic invertebrate fauna for assessing ecological quality: they are sedentary; they have relatively long life-spans; they comprise diverse species that exhibit different sensitivities or tolerances to stress; and they play an important role in the cycling of nutrients and materials between bed sediments and the overlying water column (Borja et al. 2000, 2009, Rees et al. 2006, Dauvin 2007, Dean 2008).

Descriptions of benthic variability and its relation to climate change and other global stressors are still evolving as more evidence and time-series observations are becoming available. Climate change may modify population dynamics over time and space and change the geographical distribution of communities, sometimes resulting in habitat loss and species extinction, with repercussions for ecosystem functioning and biodiversity (Birchenough et al. 2011).

Effects of Increasing Temperatures

Temperature affects all biological structures and physiological processes ranging from protein damage and membrane fl uidity to organ function (Hochachka and Somero 2002). Organisms may have to pay a high toll for their thermal sensitivities, particularly in terms of repairing and replacing heat-denatured proteins and in the adaptive alterations of cellular structures and processes during acclimatization (Somero 2002). The biological importance of rising temperature varies within and among species and it is known that different ontogenetic stages are differently susceptible

Benthic Community and Climate Change 123

to environmental stress. Certain young benthic stages, for example, are more vulnerable to stress than are adults (Harley et al. 2006). Changes in temperature may directly infl uence mortality, reproduction, the onset of spawning and the embryonic and gonad development of benthic species, leading to phenological changes (Birchenough et al. 2011) that directly affect trophic interactions, alter food-web structures and can lead to changes at the ecosystem level. Since the recruitment success of higher trophic levels is highly dependent upon synchronization with pulsed planktonic production, temperate marine environments may be particularly vulnerable to these changes (Edwards and Richardson 2004). For example, the spawning of Macoma balthica in north-western Europe is timed in accordance with the temperature; warmer trends in recent years have led to earlier spawning but the timing of spring phytoplankton blooms has remained unchanged, resulting in a temporal mismatch between larval production and food supply (Philippart et al. 2003). Furthermore, the peak abundance of shrimp has advanced to coincide more closely with the arrival of vulnerable spat, thus intensifying shrimp predation on juvenile Macoma balthica (Philippart et al. 2003).

Early biogeographic studies have already established a link between the distribution of marine species and mean sea surface isotherms such that increases in ocean temperature can be expected to change the latitudinal distribution of species (Birchenough et al. 2011). Over decades, climate warming may alter the composition of the resident biota by facilitating the poleward spread of species characteristic of warmer temperature regimes (Southward et al. 1995, Sagarin et al. 1999). However, climate change could also produce a signifi cant expansion of the range of species across ocean basins or continents (Stachowicz et al. 2002). The relationship between temperature and distribution shifts, however, gets complicated by the effects of other environmental parameters, such as physical barriers to movement and human usage of the coastal zone (Birchenough et al. 2011). Southward et al. (1995) reported changes in the abundance of Northeast Atlantic taxa ranging from kelps to barnacles and from zooplankton to fi sh. The local abundance of warm-water species increased and that of cold–water species decreased during periods of ocean warming, whereas the opposite occurred during a cooling period. Most changes are initially observed at the edge of ranges, where the level of physiological stress to which organisms are exposed is likely to be higher, but local and regional heterogeneity within biogeographic ranges has also been observed, with infi lling of gaps or loss of site occupancy away from range limits (Birchenough et al. 2011). In order to predict future distributional shifts, closer attention needs to be paid to species range boundaries and the factors that determine them (Harley et al. 2006). Rather than—or in addition to—shifts in species ranges, several researchers have proposed that ocean warming may result in cascading


community changes such as decreases in the abundance of higher trophic level species due to changes in ecological interactions and biological processes (Holbrook et al. 1997, Schiel et al. 2004).

Global warming is expected to result in an acceleration of current rates of sea level rise, inundating many low-lying coastal and intertidal areas. Although predictions vary across models and regions, the sea level is expected to rise 0.18–0.79 m during the 21st century, owing to thermal expansion and polar ice melt (IPCC 2007). The most obvious consequence of a rise in sea level will be an upward shift in species distribution. With the exception of some slow-growing species such as many corals, most species are expected to be able to keep pace with predicted rates of sea level rise (Harley et al. 2006). However, decreased habitat availability within a particular depth zone can lead to striking ecological changes such as the reduction of the intertidal habitat area where steep topography and anthropogenic structures (e.g., sea walls) prevent the inland migration of mudfl ats and sandy beaches (Schlacher et al. 2007). In addition, changes in sea level will shift the locations of existing anthropogenic structures to lower positions on the shore, amplifying interactions with waves and tides and further accelerating beach erosion (Cooper and McKenna 2008). The combined effects of rising sea levels and coastal armouring are therefore expected to have an unprecedented ecological impact on beaches (Defeo et al. 2009).

Effects of Ocean Acidifi cation

Today, surface waters are saturated with respect to calcium carbonate, but increasing levels of atmospheric carbon dioxide are reducing ocean pH and carbonate ion concentrations, and thus the level of calcium carbonate saturation (Orr et al. 2005). There is every indication that these changes will have a signifi cant impact on species that produce hard structures such as skeletons, shells, and on tests of biogenic calcium carbonate (e.g., molluscs, crustaceans, echinoderms, protists, algae). Recent work suggests that benthic adult molluscs and echinoderms are sensitive to changes in seawater carbonate chemistry (Shirayama and Thornton 2005). There was a signifi cant reduction in the growth rate, size and body weights and shell dissolution in specimens of the mussel Mytilus galloprovincialis and sea urchins Hemicentrotus pulcherrimus and Echinometra mathaei when they were exposed to low pH (Shirayama and Thornton 2005). Calcifi cation rates in Mytilus edulis were observed to decline linearly with increasing CO2 levels, and 70% of oyster Crassostrea gigas larvae reared under pH 7.4 were either completely non-shelled, or only partially shelled, in contrast to 70% successful development in control embryos (Kurihara 2008). In addition, higher levels of CO2 partial pressures (hypercapnia) will


affect the physiology of water breathing animals, causing physiological stress, decreased reproductive potential, slower growth and increased susceptibility to diseases (Pörtner et al. 2004). Those organisms will continue to be affected by ocean acidifi cation is well established, but less attention has been paid to the impact of the ecosystem at a higher trophic-level: organisms that depend on these calcifi ers for shelter, nutrition, and other core functions (Fabry et al. 2008, Guinotte and Fabry 2008).

Effects of Climate Change on Coral Reefs

Coral reefs are one of the most productive and diverse ecosystems on Earth, providing critical habitats that support approximately 25% of marine species (Connell 1978). In addition to human reliance on coral reefs for food supply, livelihoods and tourism (Salvat 1992, Wilkinson and Souter 2008, Pandolfi et al. 2011), they also provide important ecosystem services such as coastline protection from storm damage, erosion and fl ooding, by reducing the action of waves along the shore. The protection afforded by coral reefs also enables the formation of associated ecosystems such as sea-grass beds and mangroves (Moberg and Folke 1999). Coral reefs are also areas of spawning, nursery, breeding and feeding for a multitude of organisms; they export fi sh and invertebrate larvae to adjacent ecosystems (mangroves, sea-grass beds) and support pelagic food webs (Hoegh-Guldberg 1999, Moberg and Folke 1999).

There has been a dramatic decline in tropical coral reefs over the centuries, the pace of coral mortality and reef degradation accelerating in particular over the past 20–50 years (Pandolfi et al. 2003). Coral reefs around the world have suffered both gradual and chronic stress as a result of anthropogenic activities in the form of over- and destructive fi shing, pollution, coral diseases, mining of coral rock and sand and coastal developments that have modifi ed the reefs; however, the main threat now lies in global climate change (Hoegh-Guldberg 1999, Pandolfi et al. 2003).

Coral bleaching occurs when coral colonies under physiological stress expel their symbiotic algae (zooxanthellae), which provide much of the energy for coral and coral reef growth. Among the different kinds of stress leading to bleaching, by far the most signifi cant cause over the past two decades has been related with the increase in sea surface temperature (Wilkinson and Souter 2008). Corals may partially or fully recover from bleaching events, but may also die (Lough 2000). Even in cases of apparent recovery, thermal stress can have long-term effects in terms of reduced reproduction, reduced growth rates and increased susceptibility to other disturbances, such as coral diseases (Lough 2008). The increased frequency of bleaching reduces the capacity of coral reefs to recover and repeated mass coral bleaching events since 1970 have caused a decline in coral


cover globally (Wilkinson 2002, Gardner et al. 2003, Hughes et al. 2003), a trend that can be expected to continue as ocean temperatures rise (Hoegh-Guldberg 1999).

Climate change is therefore likely to reduce local and regional coral reef biodiversity, the elimination of sensitive species causing alterations in the community structure of the reef (Graham et al. 2007). Most researchers predict that the abundance of the more susceptible coral such as the branching Acropora that form much of the habitat complexity of Indo-Pacifi c reefs will decline as compared to less sensitive species such as slower-growing genera with massive or encrusting growth forms (Pratchett et al. 2011). When herbivores are absent or avoid macroalgal species, mass bleaching can be followed by increases in macroalgae, thus reducing the space available for coral recruitment (Pratchett et al. 2011). Following mass bleaching events there is typically a decline in the abundance of fi shes and invertebrates that consume or inhabit corals at least during some part of their life cycle, accompanied by an increase in roving herbivores (Bellwood et al. 2004). The limited evidence available of systematic changes in the abundance of mesopredators or apex predators or of declines in fi sheries’ yields associated with bleaching suggests that such effects are likely to manifest themselves in the long term (Pratchett et al. 2011).

Projections of the saturation levels of aragonite, a metastable form of calcium carbonate used by many marine organisms, indicate a possible 30% decrease in coral calcifi cation rates over the next century (Gattuso et al. 1998, Langdon and Atkinson 2005), resulting in the slower growth of corals as they become less able to compete for space, or weaker coral skeletons, thus increasing their vulnerability to erosion, storm damage and predation. Furthermore, since most reef habitats are of carbonate construction, any loss of coral diversity will have a major knock-on effect on all reef-dwelling taxa (Veron 2011).

The lesser known deep-sea corals living in cold waters below the photic zone, at depths of 50–1000 m, are widely distributed, very long-lived (several 100s of years old), forming reef frameworks that persist for millennia, and are considered to undergo relatively little environmental variability (Turley et al. 2007). These biodiversity hotspots play an important role as a refuge, feeding ground and nursery for deep-sea organisms, including commercial fi sh (Rogers 1999, Fossa et al. 2002, Turley et al. 2007); however, their slow growth and limited ability to recover make them particularly vulnerable to anthropogenic activities such as bottom trawling, seabed mining, cable and pipe laying, and oil and gas exploration (Turley et al. 2007). The growing problem of ocean acidifi cation is also threatening their existence (Guinotte et al. 2006) since the depth is in effect moving upwards as CO2 concentrations rise and ocean acidifi cation increases, meaning that many deep-sea coral ecosystems will soon be immersed in under-saturated waters.


Today the distribution of cold water corals appears to be limited to depths above the saturation depth, suggesting that the deeper coral ecosystems will eventually disappear (Guinotte et al. 2006).

Predicted rises in sea-level will probably have little effect on coral reefs since they will merely provide more space for corals to grow upwards without being exposed to the air. However, combined with an increase in the frequency of tropical storms, sea level rises will hinder the development of coral islands and atolls (Veron 2011).

Certain climate change impacts, particularly when combined with other infl uences, are likely to reduce the overall resilience of coral reefs. Changes in a coral community such as reduced biodiversity may severely undermine system resilience, resulting in a phase shift to a non-coral reef community. The loss of fi sh and invertebrates for instance may expose a coral reef to outbreaks of pests or invading species. Such effects are usually unpredictable but likely to increase (Wilkinson and Souter 2008).

The worldwide decline of coral reefs makes apparent the urgent need to adopt appropriate management techniques based on a greater understanding of the ecological processes that underlie reef resilience and of the manner in which human activities contribute to shaping ecosystems. Only in this way can we be better prepared for future changes (Bellwood et al. 2004).

Effects of Climate Change on Rocky Intertidal Habitats

The intertidal zone constitutes the interface between marine and terrestrial environments, where rocky shores are the most common littoral habitat on open wave-exposed coasts (Thompson et al. 2002). Determining factors for the occurrence of the biota found on these shores are the ability to colonize sites and tolerate a variety of stresses as well as smaller-scale physical infl uences and interactions with other organisms (Lewis 1964, Connell 1972, Stephenson and Stephenson 1972, Little and Kitching 1996, Raffaelli and Hawkins 1996). Rocky shores constitute an important functional link with other inshore habitats and the land itself and provide feeding, resting, spawning and nursery ambiences for a variety of mobile marine animals, including fi shes and crustaceans, as well as birds, reptiles and mammals (Thompson et al. 2000).

Rocky shores are relatively simple ecosystems and the ecologies of many of their species are well known, making them a good model system for detecting changes in abundance, species distribution and biodiversity (Hawkins et al. 2003, Harley et al. 2006, Helmuth et al. 2006a), ultimately furthering our knowledge of the consequences of climate change for community and ecosystem processes (Hawkins et al. 2008). Some of the most profound and best-documented changes in accelerated warming of


the planet have been seen on rocky shores of Europe (Mieszkowska et al. 2006, Herbert et al. 2007, Lima et al. 2007), United States (Zacherl et al. 2003, Harley et al. 2006) and South America (Rivadeneira and Fernandez 2005).

Temperature is an important determinant of rocky intertidal species survival, and organisms living in this habitat exist at or near the edges of their thermal tolerance limit (Davenport and Davenport 2005). Temperature is also an important factor affecting metabolism, growth, feeding behavior, reproduction and rates of larval development (Anil et al. 2001, Sanford 2002, Luppi et al. 2003, Philippart et al. 2003, Phillips 2005). The physiological function and geographic distribution of rocky intertidal species are determined by both aerial and aquatic body temperature (Helmuth et al. 2006a, Helmuth et al. 2006b), such that organisms living on rocky shores will show strong responses to changes associated with sea level rise, which affect their emersion time (Harley et al. 2006); the behavior of organisms living higher on the shore may be less predictable, whereas organisms that live at median and low tide, exposed to periods of immersion and exposure, will be more affected by the changes (Helmuth et al. 2006a). The body temperature of intertidal invertebrates during aerial exposure depends upon a number of interacting factors such as the absolute tidal height of the organism on the shore, the amount of wave splash it receives and the local tidal cycle, all of which affect the timing and duration of exposure to terrestrial conditions at low tide; during exposure, the substratum angle plays a major role in determining the amount of solar radiation received (Helmuth and Hofmann 2001). Experimental manipulation of aerial body temperature by shading has enabled the documentation of changes in mortality, rates of predation, relative competitive ability and species zonation patterns (Harley and Lopez 2003). The importance of biotic interactions and behavior has also been revealed: species are geographically limited by physiological stresses related to aerial exposure such as when the upper limits of an organism are squeezed down to the upper limit of a predator or dominant competitor, resulting in the elimination of the subordinate or prey species from the intertidal zone (Harley 2003, Harley and Helmuth 2003). The very same climatic conditions can lead to widely differing body temperatures in two different organisms, and hence different levels of physiological stress. For example, Helmuth (2002) has shown that at low tide, the predatory seastar Pisaster ochraceus may be up to 10ºC cooler than its prey, Mytilus californianus.

Many intertidal organisms are expected to display strong responses to changes in terrestrial climatic conditions (Somero 2002). A well-studied case documenting changes in the abundance of intertidal rocky shore species in response to climatic change involves two groups of acorn barnacles: Semibalanus balanoides, a boreoarctic form that reaches its main southern limit in the south of England and Brittany and two warm-water species of


Chthamalus that reach their northern limits in Scotland and dominate shores in southern Europe and North Africa (Hawkins et al. 2003). Observations since 1985 indicate that warm-water barnacle species have become more abundant (Southward 1991) (Fig. 1) and that there is a strong correlation between barnacle abundance and sea surface temperature: the abundance of the warm-water species (Chthamalus spp.) is positively correlated with inshore temperature whereas the abundance of the cold-water species (Semibalanus balanoides) is negatively correlated. Abundances of other southern intertidal species such as Patella depressa and B. perforatus have increased in the region of the biogeographic boundary in southwest England since the 1980s (Herbert et al. 2003, Mieszkowska et al. 2005) and those of many northern species have decreased (e.g., Semibalanus balanoides and Patella vulgata). It is likely that these increases/decreases in abundance are associated with increasing sea temperatures (Harris et al. 1998, Helmuth et al. 2006a, Hawkins et al. 2009).

Increasing sea temperatures have facilitated the range extension of warm-water species and numerous studies have focused on the role of water temperature in setting species distributional limits. In order to effectively determine the overall impact of climate change it is considered necessary to carry out studies throughout species ranges rather than just at the margins of species range (Helmuth et al. 2002, Sagarin and Somero 2006). Changes in water temperature can have the effect of eliminating species from intertidal regions, resulting in contractions in species geographic distribution as climatic conditions exceed the species’ physiological threshold of tolerance. Conversely, range expansion occurs when new individuals colonize at sites that become physiologically tolerable for the fi rst time as environmental

Fig. 1. Abundance of Chthamalus spp. and Semibalanus balanoides at Cellar Beach, Yealm estuary, Devon. Points represent average abundance during each autumn. Taken from: Hawkins et al. 2003 (Copyright Clearance Center: License Number: 3087770882741).


conditions change (Helmuth et al. 2006b). The impact of climate change on intertidal invertebrate distribution has been the focus of a growing number of recent studies. For instance, the blue mussel Mytilus edulis has been reported to have extended its distribution north by 1000 km between 1977 and 2002, from the Norwegian mainland to Svalbard Island (Salvigsen et al. 1992, Weslawski et al. 1997, Berge et al. 2005), possibly as a consequence of the dispersal of planktonic larvae from a source population on the Norwegian mainland by a northward transport of warm Atlantic water into the Greenland Sea region in recent years (Berge et al. 2005).

Hawkins et al. (2008) have reported a general polewards retreat and decrease in the abundance of cold water-adapted species and an advance and increase in the abundance of warm water species. Although climate change has quite clearly had and will continue to have a large impact on at least some species in rocky intertidal ecosystems, this impact is not likely to be evident everywhere (Helmuth et al. 2006b). For example, in a study conducted along the Chilean coast, where the last 57 years have evidenced a weak warming trend, the range boundaries of species did not change beyond chance expectation (Cane et al. 1997). These fi ndings suggest the impossibility of making generalizations about poleward shifts in species’ ranges because of differences in regional warming trends (Rivadeneira and Fernandez 2005). Responses to climate change are also species specifi c and depend upon life history and other ecological traits (Hawkins et al. 2008). As underlined by Helmuth et al. (2006b), ecological responses to climatic variability within the intertidal ecosystem can only be fully elucidated through an integrated approach linking changes in a variety of environmental parameters to the physiological and ecological responses of organisms over a range of temporal and spatial scales and within a hierarchy of biological organization.

Can Climate Change Trigger Biological Invasions?

Marine ecosystems have been subject to changes in species composition; many of the species moved by humans—deliberately or otherwise—beyond their native range become established and spread in their new habitat (Vitousek et al. 1997). Some species have been introduced accidentally via ballast water, soil, or as crop seed “contaminants” and others have been intentionally introduced as ornamentals, food, or fi ber products (Vilà et al. 2007). Intercontinental shipping and the commercial transport of aquaculture products from one coast to another have also played an important role (Wolff and Reise 2002). However when they are introduced, only about 10% of these species actually become established and spread in their new environments, and only a small fraction are likely to induce changes to the recipient ecosystem (Williamson and Fitter 1996). Nowadays,


natural migration of organisms due to climatic variations is becoming superimposed by anthropogenic vectors, facilitating a much faster and wider distribution into new habitats (Diederich et al. 2005).

Biological invasions can have many ecological impacts, including the modifi cation of benthic communities and displacement of native species, and for this reason are considered to be one of the major causes of biodiversity loss (Soulé 1991). Some invasions may affect the overall functioning of an ecosystem in terms of material fl ow between trophic groups, primary production, the relative extent of organic material decomposition, and extent of benthic-pelagic coupling (Occhipinti-Ambrogi 2007). The decline in existing populations of native species can sometimes be overlooked because of taxonomic mistakes or lack of available information on species richness, diversity and composition (Geller 1999).

The past few decades have witnessed an accelerated rate of establishment of introduced species in coastal waters (Ruiz et al. 2000). The enhanced global transport of species together with increasing coastal ocean temperatures could provide a fuller explanation for these increasing rates of invasion by nonindigenous species (Stachowicz et al. 2002). Mass mortalities in benthic organisms as a result of anomalous temperature stress (Cerrano et al. 2000, Pérez et al. 2000, Garrabou et al. 2001) open up niches for new colonizers (Occhipinti-Ambrogi 2007). The competition for open space on the substrate is heavily infl uenced by the timing of recruitment, which in turn is highly dependent on temperature; changing seasonal patterns of temperature may therefore favor the settlement of invasive species at a particular time of the year and have long-lasting consequences in preventing the recruitment of native species (Occhipinti-Ambrogi 2007). Increasing coastal ocean temperatures may therefore accelerate the homogenization of global biota by favoring the growth and recruitment and hence dominance of non-native species over native species (Stachowicz et al. 2002, Ricciardi 2006).

Sharp differences have been found over the past 12 years in the response of native and introduced ascidians (sea squirts) to interannual variations in temperature in the eastern Long Island Sound (North America). Total annual recruitment of introduced ascidian, the solitary Ascidiella aspersa and the colonial Botrylloides violaceous and Diplosoma listerianum, was positively correlated with the mean winter water temperature (Fig. 2; slope=0.280, R2=0.02, R2=0.46) and a strong negative correlation was found between the timing of the initiation of recruitment and winter water temperatures. The magnitude of native ascidian recruitment on the other hand, was negatively correlated with mean winter temperature (more recruitment in colder years) (Fig. 2; slope=0.151, R2=0.08, R2=0.30) and its timing was unaffected (Stachowicz et al. 2002), suggesting that invaders arrived earlier in the season in years with warmer winters. Ocean warming therefore facilitates


the establishment and spread of introduced species, particularly those from warmer or less extreme climates. The spread of exotic species is known to depend upon temperature regimes and may profi t from global warming (Walther et al. 2002). For instance, recent studies suggest that winter mortality is the main factor limiting population increases of the introduced American slipper limpet (Crepidula fornicata) in the Wadden Sea of Germany and that the milder winters caused by global warming combined with a northward shift of the negative ecological and economic effects may help increase the abundance of northern populations (Thieltges et al. 2004).

A well-known case of biological invasion worldwide is that of the Pacifi c oyster Crassostrea gigas; this bivalve originated in Japan and has been distributed in oyster cultures all over the world since the early twentieth century (Andrews 1980, Quayle 1988, Arakawa 1990, Chew 1990). In most regions it reproduced and dispersed successfully in new environments such as British Columbia, Australia, New Zealand, Chile, Peru, Ecuador, Argentina, and Brazil (Qualye 1988, Ayres 1991, Dinamani 1991, Möller et al. 2001, Orensanz et al. 2002, Ruesink et al. 2005, Melo et al. 2009). Recent studies found that high recruitment corresponded with higher than average water temperatures in late summer, when spawning occurs, larvae are

Fig. 2. Total annual recruitment of introduced (nonnative) species is positively correlated with mean water temperature during the preceding winter, whereas native species recruitment is negatively correlated with winter temperatures. Copyright (2013) National Academy of Sciences, USA.

annual


dispersed and juveniles settle on hard substrates (Diederich et al. 2005). The fi rst natural recruitment of C. gigas, introduced into the Oosterschelde (The Netherlands) in 1964, was observed in 1975 during exceptionally warm summers with water temperatures above 20ºC (Drinkwaard 1999). In New Zealand, the dramatic increase in C. gigas in 1978, superseding that of the native rock oyster Saccostrea glomerata, was attributed to a marked rise in temperature during the main spatting period (Dinamani 1978). Global warming may therefore increase the success of C. gigas spatfall in summer and the survival of spat during the following winter, leading to increased rates of population increase of the Pacifi c oyster and an expected decline in native bivalves due to increased predation rates in the subtidal and lower intertidal (Troost 2010).

It is therefore clear that changes in atmospheric circulation, precipitation patterns and ocean circulation, increases in global mean temperatures, elevated CO2 and a higher frequency of hot summers or warmer late-summer water temperatures, can have a profound impact on dispersion routes, leading to changes in the abundance and distribution of introduced species around the world (Diederich et al. 2005, Occhipinti-Ambrogi 2007). Responding to the growing human reliance on goods and services derived from nature and natural processes calls for a better understanding of the dynamics of invasive species and their long term consequences for marine ecosystems. In the context of a fast-changing marine landscape, deeper insight is required not only to monitor the state of the environment and assess corrective measures but also to predict the behavior of invasive species under an altered climate and the emergence of new invasive species. Looking into the future, the successful management of ecosystem goods and services will involve new tools that integrate species invasion and climate change (Occhipinti-Ambrogi 2007, Hellmann et al. 2008).

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CHAPTER 6

The Ecology of Coastal Wetlands

Paula Daniela Pratolongo

Introduction

Coastal wetlands have always changed in time and space, with climate and sea level driving their long term evolution. The last ice age, which occurred from approximately 110,000 to 10,000 years before present, buried much of the current temperate coasts and continental shelves under thick layers of ice. During the maximum extent of glaciations, about 20,000 years ago, the mean sea level decreased to about 120 m below the present as a consequence of the massive storage of water on the continents. At that time, coastal wetlands would only have existed along those coasts that were free of permanent ice. Even at warmer latitudes, wetlands would not have occupied the present shorelines, but a fringe on the upper continental slope (Wolanski et al. 2009). After the last glacial maximum, the mean sea level rose rapidly, reaching its present world averaged (eustatic) level about 6,000 years ago. Coastal environments responded to these changing conditions and the land elevation related to sea level determined, at any time, the presence and location of wetlands.

Sea level controlled wetlands comprise a wide variety of environments from tidal flats to freshwater swamps, fens or barren salt flats, in a continuum of increasing elevation from a shoreline to the upland. In the case of the intertidal zone, the primary abiotic control on wetland structure and

IADO-CONICET-UNS, Cno. La Carrindanga km 7,5 (8000) Bahía Blanca-Argentina. Email: [email protected]

The Ecology of Coastal Wetlands 141

function is a combination of tidal inundation frequency, depth, and duration, known as hydroperiod (French 1993). Wetland environments below the highest astronomical tide experience direct tidal inundation, with decreasing frequency and duration as a function of increasing elevation within the tidal frame. For coastal wetlands landward, the water table is linked to the sea level infl uence, which is an important control on the groundwater position that provides the waterlogged conditions necessary for their development (Hageman 1969). The result of the interaction between hydrodynamics and elevation is a shore-parallel zonation of plants, where each zone tend to move both vertically and horizontally in response to changing sea level and associated stressors (Hayden et al. 1995).

Coastal wetlands developing under rising relative sea level during the past 10,000 years (the Holocene) have been largely studied along the eastern coast of North America, as well as marshes in northern Europe. The process of wetlands migration under a rising sea level was earlier described by Dutch geologists. Hageman (1969) termed the area where freshwater wetlands persist under the control of relative sea-level as the perimarine zone and studied the evolution of freshwater swamps in the western Rhine/Meuse delta, in response to the rise in sea level during the Holocene. There are examples of sedimentary records (Kirby 2001, Waller 1994) showing that peat-forming perimarine wetlands accumulated deep layers of organic matter between around 6,000 and 2,000 years BP, and palynological analysis of these peat deposits showed sequences of salt marshes, reed swamps, fens and woodland carr communities developing under a rising sea-level, which maintained a near surface watertable (Waller et al. 1999).

In contrast to these well studied examples of continuous rising in relative sea level, little is known about wetland response in coastal environments that developed under different conditions after the last glacial age. While the globally averaged sea level has been rising from the Last Glacial Maximum (LGM) to the present, the relative height of the sea with respect to land (relative sea level) can vary from place to place due to local tectonic and hydrographic effects (Fig. 1). As the mass of the continental ice melted, a huge weight was released from continental shelves, which rose by isostatic rebound of the land. In those areas where the ice load was the greatest and the largest rebound occurred, the land rose faster than the sea, the relative sea level decreased, the coast prograded, and new land emerged over the last 10,000 years (relative sea level curve type A in Fig. 2). In other areas, the coast initially receded from a rising sea until the relative sea level reached a maximum about 5,000 years ago. After the transgressive maximum the coast prograded, as the relative sea level decreased to its present elevation (relative sea level curve type C in Fig. 2). Where the relative sea fell rapidly, new land constantly emerged, the coastal wetlands continuum migrated seaward, and the Holocene estuarine environments became part of the terrestrial


landscape. Where sea level rose, Holocene estuaries were drowned and new wetlands formed landward.

In coastal environments, continuous change is the paradigm. Whatever the Holocene relative sea level curve is, coastlines worldwide are still evolving, estuaries are not at a steady state and coastal wetlands continue to migrate. Through this chapter, we will consider these changes in the geological time scale as the reference state or “normal” pattern, and we will focus on those changes that are affecting the world’s coastal environments within time scales of days to centuries. Human activities can modify rates of natural changes, and there are evidences of anthropogenic actions that have signifi cantly enhanced changes driven by natural agents. As we cannot separate human activities from natural changes, we will consider not just ecological changes driven by global atmospheric and climate alterations, but also coastal change created by human use of water on land, and increased erosion of terrestrial sediments, as well as direct human destruction of coastal habitats.

Fig. 1. Approximate distribution, according to Pirazzoli (1991), of the typical relative mean sea-level curve types A, B and C as shown in Fig. 2.

Fig. 2. Approximate relative mean sea-level curves for the zones A, B, and C shown in Fig. 1.


Climate Change

The temperature of the earth has largely changed through time. Atmospheric temperature records are available from the late 1880s and reconstructions of the temperature regimes across the globe show that in the latter part of the 20th century the atmosphere of the earth has warmed considerably faster than at any time during the previous millennium. A time series of global mean temperatures show that during the 20th century we experienced relatively cooler global temperatures up to the 1940s, followed by increasing warming, particularly after the 1980s (Jones et al. 2001). There are large uncertainties surrounding the prediction of changes in world’s climate, but there is considerable agreement that temperatures will continue to rise, with average global surface temperature projected to increase by between 1.4 and 5.8°C above 1990 levels by 2100.

The net primary production of wetland plants is related to latitude and temperature, with greater productivity occurring at lower and warmer latitudes (Turner and Gosselink 1975). Warmer temperatures are also expected to change the geographical distribution of salt marshes and forested wetlands. For many species, specially mangroves, the limiting factor for the geographic distribution is low temperature or freezing events that exceed tolerance limits (Snedaker 1995), and the limits of tropical and subtropical mangrove communities are expected to migrate to higher latitudes. A warmer climate might also favor highly opportunistic exotic species to take advantage over native species (Malcolm and Markham 1996), and salt marsh plant composition could also affect ecosystem productivity.

There are large uncertainties regarding the impacts of a warmer climate on coastal wetlands, and the reciprocal effects of wetlands response on climate, potentially moderating or enhancing warming. The increase in atmospheric CO2 is a major driver of climate change and coastal wetlands have shown a great potential as carbon sinks (Chmura et al. 2003). Increased atmospheric CO2 could increase net primary production and carbon sequestration, if nutrients, precipitation, and other factors are not limiting to plant growth. Increased CO2 has shown to produce higher growth rates and greater biomass in different salt marsh plants and mangrove seedlings (Rozema et al. 1991). Related research on the effects of elevated atmospheric CO2 on different wetland systems suggests that growth enhancement is a common consequence, but climate-related changes to the carbon cycle are likely to alter the sequestration service provided by salt marshes, mangroves and other coastal ecosystems in ways that are still unclear.

There are different types of photosynthetic pathways. Most plants use C3 photosynthesis, in which the CO2 is fi rst incorporated into a 3-carbon compound, but the C4 photosynthetic carbon cycle is an elaborated addition


to the C3 photosynthetic pathway. It evolved as an adaptation to high light intensities, high temperatures, and dryness. Therefore, C4 plants dominate biomass production in the warmer climates of the tropical and subtropical regions (Edwards et al. 2010). Due to differences in the photosynthetic pathway of CO2, C3 plants are expected to respond differently to CO2 increases than C4 plants (Ainsworth and Long 2005). In a brackish tidal marsh in the Chesapeake Bay, an experimental enrichment with CO2 produced a significant biomass increase in Scirpus olneyi, a C3 sedge (Erickson et al. 2007). The C4 grasses S. patens and Distichlis spicata did not grow better under elevated CO2 conditions. This evidence, along with other experimental results (Lenssen et al. 1993, Rozema et al. 1991) suggests that the response to elevated CO2 depends upon plant composition, and that higher concentrations of CO2 would benefi ciate C3 plants.

As evaporation increases exponentially with rising surface water temperatures, it was predicted that hurricanes and tropical storms could increase their frequency and intensity as a consequence of global warming (Emanuel 1987, IPCC 2007, Raper 1993). Although shorelines are greatly affected by storms, there is little evidence that hurricanes produce long term detrimental impacts in natural coastal wetlands (Michener et al. 1997) probably because, unlike upland biota, salt marsh and mangrove species are highly adapted to salt stress and inundation (Valiela et al. 1998). However, coastal storms have short term effects capable of accelerating, disrupting and reversing numerous geomorphic events and ecological processes. Storm surges temporarily raise salinities in brackish and freshwater tidal marshes that promote shifts in plant community compositions. Floodwaters can drown salt marshes, having detrimental effects on animal populations, and causing brief and localized population declines (Michener et al. 1997). Storm-delivered sediments are also essential to the sediment budgets of many coastal systems, having varying effects on marsh elevation, including increases due to sediment deposition and stimulation of root growth, as well as decreases due to erosion and compaction of soils (Cahoon 2006).

Although changes in storm patterns would have an impact in sedimentation rates, global climate models produce mixed results related to the effects of global warming on the future incidence of hurricanes. It has been suggested that current models cannot adequately predict changes in storm distribution and frequency (Gates et al. 1990, Lighthill et al. 1994, Mitchell et al. 1990). Moreover, there is no evidence that the frequency and intensity of tropical storms has increased as the ocean warmed over the past decades (Folland et al. 1990). Regardless of any effects of global climate trends, changes in storm frequency and intensity are also expected to occur as a response to long term meteorological cycles including the multidecadal Sahel rainfall (Landsea and Gray 1992), the quasi biennial oscillation (Gray 1984), and El Niño Southern Oscillation (Wu and Lau 1992).


Accelerated Sea Level Rise

A major concern related to climate change is rising sea level associated to the melting of ice-sheet, land ice, and thermal expansion of the ocean (Webb III et al. 1993, Wigley and Raper 1993). During the last 100 years, eustatic sea level has risen was about 1.5 to 2 mm per yr (Miller and Douglas 2004), but local rates of relative sea level rise are highly variable, due to regional differences in groundwater and oil withdrawal, compaction of muddy soils, subsidence, isostatic rebound, and tectonic uplift. The rate of rise is predicted to accelerate in the 21st century, but the actual amount of increase is diffi cult to tell precisely. While predictions based on empirical relationships between temperature and sea level suggest rates of sea level rise of 1 m or more in the 21st century (Rahmstorf 2007), IPCC projections span from 9 to 88 cm rise by 2100. Model averages range more narrowly from 28 to 43 cm above the global sea level at the beginning of the century (IPCC 2007), but uncertainties for regional predictions are about 50% greater than for the global average.

There has been considerable discussion as to how coastal wetlands will develop in the future under climate enhanced sea level rise (Reed 1990, Simas et al. 2001). Early studies (Boorman et al. 1998, Titus 1987) predicted the large-scale loss of coastal wetlands as a consequence of sea level rise exceeding sediment supply (Temmerman et al. 2004), or the infl uence of sea-level rise on marsh productivity (Morris et al. 2002). However, there is some evidence to suggest that, at some locations, the geomorphic response of salt marshes is not sediment limited. Many temperate salt marshes built from allochthonous sediment show a signifi cant excess of vertical sediment accretion relative to sea level rise (French 2006, Stupples and Plater 2007). In the Mississippi Delta, accretion rates greater than 10 mm year−1 have been measured where there is suffi cient sediment input from the river (Cahoon et al. 1995, Conner and Day Jr. 1991, Day et al. 2000, Hatton et al. 1983), and mangroves in many estuaries in northern Australia tolerated sea level rise of 8–10 mm per year in the early Holocene (Woodroffe 1995). These accretion rates are higher than most projections, and suggest that coastal wetlands can persist at a given location, in spite of high rates of sea level rise, if there is suffi cient mineral and organic soil formation.

Nevertheless, human activities alter the ability of wetlands to accrete both at local and regional scales, and enhanced sea level rise has led to signifi cant changes on coastal systems, mainly associated with salinity intrusion in estuaries and altered sediment transport. There are numerous examples of detrimental effects of accelerated sea level rise on coastal wetlands around the world, including Chesapeake Bay, the Mississippi Delta and other Atlantic estuaries in North America (Day et al. 2007, Day et al. 2003, Hackney and Cleary 1987, Stevenson et al. 1985), Rhone, Ganges,


Indus, Nile, and Ebro deltas (Day et al. 2006, Ibañez et al. 1999, Milliman et al. 1989, Pont et al. 2002, Snedaker 1984, Stanley and Warne 1993), and Venice Lagoon (Day et al. 1999, Pirazzoli 1987). Deltaic regions are particularly vulnerable to a relative sea level rise because of a rapid subsidence. Under this scenario, river sediment supply and human-induced changes in sedimentary fl uxes are critical agents in shaping deltaic evolution. Dam construction and the increase in water demand for agriculture, industry, and tourist development have dramatically reduced the sediment load of rivers, and are thought to be a major cause of deltaic degradation when coupled with subsidence (Stanley and Warne 1993).

Vertical accretion and progradation resulting from sedimentation are not the only processes supporting wetland persistence within the zone of hydrologic infl uence of sea level. For coastal areas experiencing a relative rise in sea level, the different plant associations within the coastal wetlands continuum are expected to migrate landward, and the future extent of the wetland zone will depend upon the combined effect of seaward vertical accretion, disturbance, and landward transgression (Christian et al. 2000).

There are good examples of coastal marshes and mangroves throughout the world, set against the land as a fringe parallel to the shore that seem capable of responding to sea level rise by moving inland, but there are also some exceptions. Wetlands growing on islands within estuaries have no land to migrate (Kearney and Stevenson 1991, Wray et al. 1995). Similarly, the migration of wetlands inland may also be prevented in places where the landward slope is too steep, or where people have built hard barriers landward of the wetlands. In these cases where transgression stalls, low sediment supply results in an eroding seaward margin, and wetland communities may disappear by erosion over time (Brinson et al. 1995).

Erosion of salt marsh sediments as a response to higher tidal fl ow velocities and wave energy also raises concerns. Although eroded sediments supply mud to the tidal sediment budget, these resuspended muds contain pollutants accumulated over the past years, many of which are now banned or strictly regulated due to their detrimental environmental and health impacts (Valette-Silver 1993, Williams et al. 1994). For example, the erosion of sediment from the Mersey Estuary will release DDT, mercury, lead, and radionuclides into suspension (Fox et al. 1999). As a side effect of sea level rise and marsh erosion, pollutants that are currently locked away in sedimentary reserves may become a major source of environmental pollution in the near future.


Alterations in Sediment Transport

Humans have altered fl uvial transport of sediment to the coast, mainly through agricultural practices for food production. The cultivation of land makes soil more susceptible to erosion, and some of the eroded soil reaches the coastal zone. Deforestation and cultivation may cause up to eight fold increase in sediment loads in small rivers downstream, and 3.5 fold increase in larger rivers (Douglas 1990), and the exponential growth of human population has led to an increased downstream transport of sediments. At the same time, the expansion of human settlements and more sophisticated agricultural practices required an increasing management of fresh water, and triggered widespread construction of dams, dikes, levees, and canals. The net result of waterworks has been that sediments now accumulate behind dams, instead of being exported to the coastal zone.

The human-mediated changes in sediment transport do have the potential to locally alter coastal environments. Before 200 BC the Huanghe (Yellow River) drained a forested steppe, and the sediment load was about an order of magnitude lower than during the 20th century (Milliman et al. 1987). As agricultural pressure grew, soil eroded from fi elds increasingly entered the river, and the coastline expanded seaward 50 km across an interval of 130 years. The sediment loads were high and varied seasonally and interannually drastically altering the course of the river, but delta expansion continued.

The delta of the Nile River shows a similar pattern. Natural and human alterations of the course of the Nile created the Rosetta Branch through the delta between 500 and 1000 AD, which brought an increased fl ow of Nile water through the area, and silty sediments. The increased input of silt, and subsequent trapping within the delta, caused the Rosetta Promontory to advance into the Mediterranean Sea during the years 1500 to 1900 (Fanos 1995). After 1900 irrigation of fi elds on the Nile fl ood plain lowered the fl ow of water through the Nile delta, what resulted in a steady recession of the Rosetta Promontory and a shoreline that was back to where it might have been between 1700 and 1800.

In the catchment area of the Mississippi River, intensive agriculture expanded through the 1800s and into the 1900s, and the enhanced soil erosion encouraged the formation of the Soil Conservation Service in 1935 (Turner and Rabalais 2003). In the early 20th century, intensive river engineering, especially reservoir construction, greatly reduced sediment transport within the basin (Meade and Moody 2010), and dikes built after World War 2 to control fl oods also contributed to trap suspended sediments and lowered the sediment transport downriver. Changes in area of the


Mississippi delta during these decades clearly refl ect changes in sediment loads arriving at the mouth of the Mississippi, which drastically reformed the structure and functions of associate coastal wetland environments (Wells 1996). The earliest maps of the Mississippi River delta show channels bounded by narrow marshes, but in the mid-1800s a period of rapid land formation began, resulting in 560 km2 of new land in 100 yr (Wells and Coleman 1987). Maps of the 1930s delta show large wetland areas, much of which have been more recently eroded.

There seems to be a common pattern in which, as agricultural land use intensifi es, sediment loads carried by rivers draining the area increase. As it becomes necessary to strengthen agricultural production, water interception works are built, and sediment loads entering coastal waters decrease. The rivers discussed here are examples that demonstrate how anthropogenic changes in sediment transport have become the dominant factor determining whether relative sea level rise is positive or negative in many deltas (Syvitski and Kettner 2011). These well known cases may be used as indicators of what may occur in other developing watersheds around the world. Because of the positive relationship between suspended sediments and land area in these deltas, it may be inferred that future variation in suspended sediments delivered to the mouth of rivers, higher or lower, would likely lead to proportional wetland changes in the coastal zone.

Species Introduction

Wetland plants have been introduced to coastal areas worldwide, to make use of their commonly high productivity, tolerance to inundation, and peat-building capabilities. Spartina anglica is a common halophyte in lower marshes of northern Europe. This fertile allopolyploid arose by the end of the 1880s in Southampton Water, UK, by chromosomal doubling from Spartina townsendii, a hybrid between the introduced Spartina alternifl ora and the native Spartina maritime (Ayres and Strong 2001). Although S. anglica invaded large areas during the fi rst 30 years after being reported (Raybould 1997), during the late 1920s and early 1930s the species began to show signs of loss of vigor at some locations, leading to widespread diebacks. This process of recession, fi rst described by Goodman et al. (1959), has continued up to the present. However, the concept of comprehensive spread followed by sudden decline cannot be generalized, and there are some locations where S. anglica has maintained its vigor and capacity for colonization of mudfl ats, displacing eel grass (Zostera marina) and algae (Raybould 1997).

In addition to a rapid natural spread following its appearance, S. anglica was also extensively planted to stabilize soft sediments, resulting


in a considerable expansion throughout the British Isles and nearby Europe over a relatively short period. S. anglica has also displaced native Puccinellia maritima on Dutch coasts (Daehler and Strong 1996), and mudfl at, seagrass, mangrove, and marsh habitats in southern Australia. The rapid colonization of S. anglica over extensive fl ats can be viewed as an autogenic control whereby plants may cause a rapid buildup of sediment, and thereby signifi cantly affect coastal dynamics (Allen 2000).

The intertidal zone in the Yangtze Delta extended over about 1,550 km2, but an area of 750 km2 has been reclaimed in the past half century, and approximately one third remains as salt marsh (Gao and Zhang 2006). In 1979, S. alternifl ora was transplanted into coastal China to stabilize tidal fl ats. Since its introduction, this species has gradually invaded the former P. australis communities, and the upper Scirpus mariqueter zone, aggressively replacing native species (Cheng et al. 2006). At present, the plant zonation of coastal wetlands in the Jiangsu Province typically includes S. alternifl ora in the lower limit of vegetation, dominating the upper intertidal zone (Liu et al. 2007).

Spartina foliosa has a narrow range of distribution, limited to the Pacifi c coast of North America, from Humboldt Bay to Baja California. Similar to S. alternifl ora, S. foliosa takes on a tall or “robust” form, which grows at lower elevations, and a “dwarf” form at elevations closer to mean high water. Further north, exotic species are invading including the hybrid S. alternifl ora x S. foliosa in the San Francisco Estuary, where it has been reported to increase in area a 100-fold since the 1970s, and Spartina densifl ora, a South American native, ranking second in area covered, also appearing in Humboldt Bay. Finally, S. patens and S. anglica, the latter a hybrid species of S. alternifl ora and the European S. maritima, are also found, but have not yet dispersed far beyond their introduction sites (Ayres et al. 2004).

In Willapa Bay, nearly one-third of the area originally covered by mudfl ats is now infested with S. alternifl ora. This species was introduced into Willapa Bay during the late 1800s but was not identifi ed until the 1940s. During the fi rst 50 years, the population expanded slowly, but from 1945 to 1988 the plant spread rapidly throughout the bay, resulting in severe habitat alteration as unvegetated mudfl ats were converted to salt marshes (Simenstad and Thom 1995). Salt cedars, Tamarix spp., have also invaded wetlands on the Pacifi c coast of North America (Whitcraft et al. 2007). The trees are salt tolerant and notably disrupt water cycling in riparian wetlands (Zedler and Kercher 2004). In salt marshes, Tamarix invasion converts succulent plant communities less than 1 m in height into woody-dominated communities with a 3-m canopy (Whitcraft et al. 2007).

Phragmites australis, a European reed, has increased in abundance in North American marshes over the past century, outcompeting native plants (Minchinton et al. 2006, Orson 1999, Saltonstall 2002). This large reed is a


common species historically appearing at lower abundances in nonsaline marshes. Over the past 150 years, the species has increased its cover in freshwater and brackish tidal wetlands along the middle Atlantic, displacing other well-established wetland plant (Saltonstall 2002). Although P. australis expansion was thought to be limited by high soil salinities (Chambers et al. 1999), observations in southern New England indicate that this reed is actually invading salt marshes, and the spreading could be related to human induced changes in nutrient regimes (Bertness et al. 2002). Disturbance and periodic long-term changes in tidal regime, in addition to nutrient enrichment, appear to play an important role in P. australis expansion (Chambers et al. 1999).

Salt marsh plant invasions commonly affect animal communities, given the effects of vegetation on the physiochemical and hydrodynamic environment, as well as detritus quality and availability (Neira et al. 2006). For example, P. australis reduces aquatic habitat quality and it is less palatable to animal consumers than native species (Zedler and Kercher 2004), and S. alternifl ora × foliosa invasion in California has been associated to reduced biodiversity of invertebrate species (Brusati and Grosholz 2006).

Similar to plant invasions, some invasive animals have had powerful effects on native communities. Musculista senhousia, an exotic mussel to the Pacifi c coast of North America, colonizes marshes and mudfl ats and creates hard substrate in previously soft sediment environments. This change in structure has detrimental effects on several invertebrate communities, especially of native clams (Crooks 2001). In San Francisco Bay, the invasive burrowing isopod Sphaeroma quoyanum caused erosion at the marsh edge, and is considered a signifi cant cause of marsh loss (Talley et al. 2001). Myocaster coypus, introduced to the Gulf coast for fur farming, also contributes to marsh loss through its burrowing and foraging activities, which depress soil accretion processes, ultimately leading to submergence of marshes with low sediment supply (Ford and Grace 1998).

These are just a few examples of plant and animal species that can signifi cantly alter the physical structure of the environment. In their non native range, species such as S. alternifl ora, P. australis, and M. senhousia, have predictably strong effects on other organisms, and are expected to have a deep impact on the original ecosystem functions.

Eutrophication and Pollution

Eutrophication is an increase in supply or production of organic matter in an environment (Nixon 1995), which forces a large array of ecological and biogeochemical consequences (Cloern 2001). One major human modifi cation of coastal environments has resulted from activities that change the natural


cycling of nitrogen and phosphorus through land clearing, applications of fertilizer, discharge of human waste, animal production, and combustion of fossil fuels (e.g., Nixon 1995). At some locations, waters reaching the coastal zone have dramatically increased its contents of N and P compared to concentrations in the middle of the 20th century (Cloern 2001). This human driven fertilization of coastal wetlands has serious environmental consequences, because it stimulates plant growth and disrupts the balance between the production and metabolism of organic matter in the coastal zone.

Nutrients reaching the coastal zone are captured and transformed by sediments and wetland plants, which function as long-term nutrient sinks (Valiela and Teal 1979). Most salt marsh plants are nitrogen limited under natural conditions, and relief of nitrogen limitation results in an increase in aboveground plant height and biomass (Kiehl et al. 1997, Valiela et al. 1976). However, the limiting role of nitrogen is not always the case for every primary producer, in every coastal wetland, and nutrients retained in the produced biomass will eventually be released, at a longer time scale.

Coastal wetlands can be generally considered as sinks for the excess of nutrients, but eutrophication has the potential to alter performance and competitive ability of wetland plant species, changing patterns of plant zonation (Levine et al. 1998). In New England salt marshes, the effects of nutrient-rich runoff from human activities has been pointed as a major cause in the observed landward expansion of S. alternifl ora and the seaward invasion of P. australis (Silliman and Bertness 2004). In the Wadden Sea, Elymus athericus, a typical high marsh species, expands over marshes formerly dominated by Festuca rubra, and this long-term vegetation change has also been attributed to eutrophication (Rozema et al. 2000).

In addition to nutrients, rivers reaching the coastal zone may carry many other contaminants, including metals, petroleum hydrocarbons, and synthetic organochlorines. Salt marshes and mangroves are depositional environments for suspended particulate matter and associated contaminants. In anoxic soils, metal ions precipitate as sulfi des of low solubility, making deeper sediments stable storage for pollutants in the absence of bioturbation or other sources of oxidation (Weis and Weis 2004). Heavy metals have rarely been found to have negative effects on plants (Giblin et al. 1980, Suntornvongsagul et al. 2007). The adaptations to salt stress, such as cellular compartmentalization of solutes and excretion through salt glands, may allow for higher tolerance of metals (Windham et al. 2001). There are experimental results to support the idea that sediments and vegetation can absorb and transform a large portion of deposited metals, working as sinks for metal contamination (Leendertse et al. 1996). Though salt marsh plants have proved resistant to metal pollution, there is concern that resuspension during erosion events pollute marine systems and that


plant uptake of metals introduces contaminants into estuarine food webs (Giblin et al. 1980).

Since the industrial revolution, fossil fuels contribute the largest portion of world energy demand, in particular for transportation. In addition, petroleum products are increasingly used in the synthesis of plastics. A major consequence of use of oil is the release of petroleum-generated compounds into coastal waters, and transport accidents contribute substantially to marine pollution. Studies of the ecological consequences of accidental spills are diffi cult to compare because it is hard to assess the relative exposure and the types of compounds involved. Although catastrophic oil spills are a more widely recognized source of marine pollution, these events occur at local spatial scales, and show recovery in the long term. A more global problem is the local chronic release of petroleum around areas with much human activity. The effects of each new input will recede as the oil weathers, but there are always new little inputs, each of which impacts the same or a nearby area. This pattern of local, but widespread, repeated perturbations could have more detrimental and wide-ranging effects than major spills (Valiela 2006).

In the mid 20th century, synthesis of chlorinated hydrocarbon compounds led to great advances in public health, agriculture and industries. In general, human-made chlorinated compounds of low molecular weight like dichlorethane, vinyl chloride, tetrachloride, trichlorethane, and trichlorethylene are rather volatile, and do not accumulate in coastal environments. In contrast, synthetic chlorinated hydrocarbons of larger molecular weight, particularly DDT and PCBs, raise concern because of their ability to reach and alter coastal environments. Chlorinated compounds have been found widely distributed throughout the entire world (Atlas et al. 1986). About 1,000,000 tons were produced before 1976, and perhaps 100,000 tons have managed to enter natural environments (Axelman and Broman 2001).

The concentrations of persistent chlorinated hydrocarbons in organisms are affected by the degree of contamination, biomagnifi cations (the transfer of contaminant from food to consumer), bioconcentration (storage within consumers), depuration (the relative ability to metabolize chlorinated hydrocarbons), and the kind and age of the organisms. The better known effect of DDT is the thinning of bird eggshells, observed after the 1940s, when DDT became broadly used (Risebrough 1989). In the west coast on North America, shell-thinning affected double-crested cormorants (Phalacrocorax auritus), and was responsible for the near extinction of brown Pelicans (Pelecanus occidentalis). PCBs reduce growth and photosynthesis in algae and plants (Mahanty 1986), and create sublethal conditions in invertebrates, fi sh, and birds, which involve lower reproduction, malformations, altered liver, thyroid, and circulatory functions, suppression of immunity, and


increased nutritional defi ciencies, among others (Harding and Addison 1986, Peakall 1986).

With the global reductions in use of DDT and PCBs during recent decades, residues in sediments and concentrations in animals have diminished in many coastal areas. However, a low but long-term concentration of the degradation products persists in organisms in most places in the world. Besides, there are of long term reservoirs in soils and aquatic sediments that will slowly release DDT and PCB residues to the environment, and it is not clear yet whether these low-concentration residues will have long-term biological consequences.

Direct Human Transformations

There always has been an intense relationship between humans and wetlands. Human impacts on wetlands date back to the end of the last glacial age, when the combination of changing climates and expanding and migrating human populations extinguished a considerable number of wetland species (Martin and Wright 1967). In a study of 12 of the world’s largest estuaries, Lotze et al. (2006) found a 67% loss of coastal wetlands during human history. Historical records describe people who depended almost entirely on wetlands, like Fen Slodgers in the English Fenlands (Wheeler 1896), or more recently, the Marsh-Arabs (Madan) of Southern Iraq (Thesinger 1964). The use of salt marshes for fi shing and livestock grazing date to the Neolithic in the North and Wadden Seas (Knottnerus 2005, Meier 2004), and the Mesopotamian tidal marshes are thought to be birthplace of agriculture (Sanlaville 2002).

Grazing lands for cattle, sheep, goats, and horses has been probably the most common use of salt marshes around the world (Knottnerus 2005), but large scale wetland modifi cation and the conversion of wetland to upland was historically undertaken for agricultural purposes (Glover and Higham 1996). Humans have directly converted wetlands into drylands to win areas for intensive agriculture and forestry (Appleton et al. 1995). Moreover, wetland destruction may result also from changing land use outside the wetland boundaries. In the Netherlands, upland deforestation started about 3,000 years ago, affecting river discharges of the Rivers Rijn, Maas and Schelde. The consequent widening of the estuaries gave opportunity for saline intrusion and led to massive erosion of the perimarine peatlands (Pons 1992).

Salt marsh reclamation for agriculture began in the Netherlands and France by the eleventh century, and probably earlier in China (Yoshinobu 1998), through diking for fl ood protection (Knottnerus 2005, Reise 2005). Diking results in the conversion of wetland to upland. The biogeochemistry of salt marsh soils, characterized by slow decomposition and tidal fl ushing


of metabolic toxins (Mendelssohn and Morris 1999), is heavily altered by diking. While drained marsh soils become acidifi ed and organic matter is more rapidly oxidized (Portnoy 1999), toxic sulfi des accumulate in undrained areas behind dikes. Lack of sediment inputs and increased decomposition also causes subsidence of the marsh surface (Portnoy and Giblin 1997). Many areas that were originally reclaimed for agriculture were transformed into urban, residential, and industrial land. Large coastal cities such as Boston, San Francisco, Amsterdam, Rotterdam, Venice, and Tokyo expanded on former coastal wetlands (Pinder and Witherick 1990).

Land reclamation is the most striking transformation of wetland areas, but there are other human disruptions of natural hydrology that signifi cantly change the structure and functions of wetland communities. Ditching is a common practice used by farmers to accelerate tidal fl ushing and to increase salt marsh plants production (Sebold 1998). The invention of ditch-digging machines accelerated ditch construction, and many marshes were extensively ditched throughout the world with different purposes (Bourn and Cottam 1950, Resh 2001). The immediate effect of ditches is to ameliorate waterlogging stress in the high marsh, and to drain organic soils up to 5 m away from ditches. By increasing tidal fl ushing frequency, ditches ameliorate anoxic stress and increase plant productivity near ditch banks (Valiela 2006). Ditches have caused shifts in salt marsh vegetation that favor high marsh species, which are better below-ground competitors when anoxic and saline stresses are reduced (Bertness and Ellison 1987). In response to intensive ditching, it is expected that low marsh grasses, well adapted to anoxia, give place to species often found on well-drained soils (Bourn and Cottam 1950).

Although the landscape level effects of ditches have been well described, ecosystem level impacts of ditching on biogeochemical cycles and marsh accretion processes have not been thoroughly explored. Tidal restrictions are another common hydrologic alteration, occurring as a consequence of roads and railway crossings that limit tidal infl uence on upriver wetland areas and restrict sheet fl ow across marshes (Buchsbaum 2001). Tidal restrictions limit marine infl uence upriver, and impounded marshes become fresher and receive less input of inorganic sediments. Complete restriction converts impounded salt marshes to freshwater wetlands (Roman et al. 1984). Tidal restrictions have negative impacts on biogeochemical cycling, leading to whole marsh subsidence in similar ways to dikes, endangering animal wetland species (Soukup and Portnoy 1986) and ecosystem functioning (Portnoy 1999).

The combination of land claim and hydrologic alterations has had an important impact on the long-term persistence of coastal wetlands. The fact that many former and extensive perimarine wetlands are largely absent from European coastal lowlands is primarily due to land claim (e.g.,


the Somerset Levels, Romney Marsh, and Fenland systems in the United Kingdom). Former marshes in northern Europe have been extensively diked over the past 2,000 years, and the few surviving areas are reduced in size and strongly modifi ed. On the German coasts of the Wadden Sea, approximately 1,000 km2 of former coastal marshes have been diked over the last millennium. Continuous embankments of newly accreted land have shaped the mainland coasts of the northern Netherlands, where anthropogenic salt marshes extend over 190 km2, and virtually no natural tidal marshes remain (Lotze 2004). In the Severn Estuary, about 840 km2 of marshes have been impounded since the end of the Roman occupation, whereas only 14 km2 of active marsh remains (Allen and Duffy 1998) and similar relations between active and embanked marshes would also apply for France and Denmark (Allen 2000).

In Eastern Asia, several human activities, such as dredging and deepening of navigation channels, water diversions to northern China and large-scale land reclamation are thought to remarkably change the Yangtze Delta environment in the next few decades (Xiqing 1998). The Three Gorges Dam is the largest hydroelectric dam in the world. Its construction began in 1994, and the reservoir began to fi ll in 2003. Since cultivation in this region dates back to more than 7,000 years, human activities have played a major role in shaping present landscapes. With more than 50% of the world’s population living in Asia, and most of the world’s ricefi elds occurring in Asian deltas (Galloway and Melillo 1998), human pressures on the contributing watersheds have greatly infl uenced the past and present coastal processes. The future development of big Asian deltas may be driven, to a great extent, by large scale engineering projects.

Prior to European settlement in the Bay of Fundy, marshes covered wide areas of the Minas Basin and the upper reaches of Chignecto Bay, where large amounts of fi ne sediments accumulate. However, during the past 400 years lowlands have been intensely diked and reclaimed, with an estimated reduction of about 70% of the former marsh area (Gordon and Cranford 1994). Agricultural expansion is the major cause of wetland losses. Since the early 1800s, wetland conversion to agriculture is estimated at over 20 million hectares, including 65% of the coastal marshes of Atlantic Canada.

The temperate coasts of Australia have also been greatly affected by human activities. Although losses have been small, compared to the original wetland area (Adam 1990), reclamation for industrial and human development has been concentrated on the southeastern coasts. Besides reclaimed areas, the main threats to salt marshes are ecosystem degradation due to alterations in hydrologic regimes, pollution, and weed invasion. The presence of large wetland complexes, comparatively little affected by human alterations, is an attribute that distinguishes South America (Pratolongo et al. 2009). Salt marshes have commonly been used as pasture for livestock


and the seasonal burning of vegetation to improve fodder quality is the most common management practice. However, the rapidly growing coastal population, the massive clearance of upland forests for agriculture, the poorly controlled industrial effl uents, and the untreated sewage discharge from coastal cities are major threatens to the coastal environment.

Concluding Remarks

Coastal wetlands have naturally evolved in response to global changes. Sea level, the major environmental driver on coastal wetlands evolution, has fl uctuated by an order of 100 meters over the past 18,000 years, and salt marshes still persist. The examples shown in this chapter reveal the complex nature of the responses of coastal wetlands to global changes, as well as the need for long-term studies to separate the effects of short term catastrophic events and widespread chronic human alterations from ‘‘natural background change”.

Numerous studies show the resilience of coastal wetlands to natural disturbances. However, changes in climate and sea level, coupled with anthropogenic changes on sediment loads, species introduction, nutrient enrichment, and other human alterations are likely to have a disproportionate impact on mangroves and salt marshes. Projected rates of human-induced change are unprecedented, and there is no outcome of global change that we can reliably predict (Schneider 1993). We actually have a poor understanding of species’ tolerances and the relationships between species diversity and ecosystem function (Ray et al. 1992). Ecosystem processes are often characterized by complex nonlinear interactions involving numerous biological, chemical, and physical components, and predicting the effects of global change on complex ecosystem functions may be a real challenge given the little knowledge that we have about thresholds of nonlinearity in the responses of coastal wetlands to future stresses (Mintzer 1992).

Most coastal wetlands export reduced nitrogen compounds, organic matter and other energy-rich substances to deeper waters supporting the receiving ecosystems, as suggested in the “outwelling” hypothesis (Odum 1980). Many commercially important species use coastal wetlands as nurseries and foraging areas during their early life stages (Robertson and Duke 1987, Van der Welde et al. 1992). Shallow protected waters fringed by wetlands are commonly rich in phytoplankton, which supports dense populations of suspension feeders and other consumers. All these functions make wetlands valuable environments for humans that harvest shellfi sh and other highly appreciated stocks. As shown in previous section, salt marsh plants and sediments retain contaminants, including heavy metals and chlorinated hydrocarbons. As long as any pollutant from land is buried in marsh or mangrove sediments, these wetlands are preventing


contamination of coastal waters. The roots and rhizomes of macrophytes consolidate otherwise loose sediments, and salt marsh canopies dissipate wave energy, lowering erosion of vegetated platforms (Leonard and Luther 1995, Stumpf 1983). Coastal wetlands may also provide a substantial capture of land-derived nitrogen because of their high rates of denitrifi cation and burial (Mitsch and Gosselink 2007).

There have been major losses of salt marshes and mangroves worldwide. Widespread and chronic human induced modifi cations are responsible for their complete destruction, in many cases, but also for the permanent change in their valuable ecosystem functions. The effects of anthropogenic factors are often diffi cult to distinguish from “natural” agents of change. For example, anthropogenic greenhouse gases are being released into the atmosphere in the context of a solar driven climatic change, and changes in sediment and nutrient supplies to coastal waters, owing to human water use, may be confused with changes in runoff, associated to long term meteorological cycles like El Niño Southern Oscillation. However, human responsibility for global environmental change cannot be denied, and we are urged to develop a scientifi c basis for management, protection, and sustainable use of coastal wetlands in this changing world.

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CHAPTER 7

Seaweeds Ecology and Climate Change

Gauna M.C.,* Croce M.E. and Fernández C.

The goal of this chapter “Seaweeds Ecology and Climate Change” is to introduce basic concepts on seaweeds’ ecology and to emphasize their ecologic role, within the phenomena of climate change occurring on Earth. We will illustrate some of the evidences for changes in the seaweed communities attributed to climate change, focusing mainly on studies about changes in temperature, UV radiation, sea-level and ocean acidifi cation.

General Aspects of Seaweed Communities

Seaweeds are major components of intertidal and subtidal communities, contributing signifi cantly to marine primary production and structuring habitats and nursery grounds for a diverse benthic fauna (Figs. 1 and 2) (Lüning 1990). Both intertidal and subtidal zones occupy a narrow coastal area and account for less than 1% of the Earth’s surface, and the productivity of this zone can equal or exceed that of most terrestrial communities, contributing to about 10% of the marine primary production (Dawes 1998).

CONICET–CCTBBca, Laboratorio GIBEA, Instituto Argentino de Oceanografía (I.A.D.O.), Camino Carrindanga 7.5 km, B8000FWB, Bahía Blanca, Argentina.

* Corresponding author: [email protected]


The term seaweed traditionally includes only macroscopic, multicellular red, green, and brown marine algae. These algae can grow from a few millimetres in size up to more than 60 m, but all of them have microscopic life stages.

Seaweeds are used in eastern countries as sea-vegetables for food consumption, as Porphyra (Nori), Rhodymenia (Dulse), Laminaria saccharina (Kombu), Chondrus and Ulva blades, because they are rich in vitamins, minerals, and proteins, and also have been reported as potential sources of bioactive compounds, and act as antibacterial and antiviral agents (Mabeau and Fleurence 1993, Vo and Kim 2010).

Seaweed assemblages grow mainly on rocky shores, being one of the most complex and compact habitats (Morton 1991). The most impressive feature of rocky shores is the zonation of the available space into distinct vertical bands (Lobban and Harrison 1997). These vertical subdivisions are separated on the basis of light penetration in the euphotic region where submarine light supports plant growth, and the aphotic region where light does not penetrate (Lüning 1990). There is also a horizontal gradient associated with exposure to wave action with greater wave force and frequency at exposed headlands than in bays or inlets (Santelices et al. 2009). Along with the depth gradient, the seaweeds are exposed to other environmental parameters. Species settling in the supralittoral zone are exposed to desiccation, high solar radiation and also atmospheric changes in temperature while species settling in the eulittoral zone (intertidal fridge) are exposed to regular and extreme changes in abiotic conditions, based on tidal infl uence (Davison and Pearson 1996). In the upper intertidal, seaweeds are usually controlled by abiotic factors that result in desiccation due to high temperatures and intense solar radiation (Davison and Pearson 1996). In constrat, in the lower intertidal, seaweed species are more often infl uenced by biotic factors such as competition or grazing (Dawes 1998).

Figs. 1 and 2. Intertidal and subtidal communities at Puerto Madryn, Chubut Province, Argentina. 1. Dense, interwoven branches of Corallina offi cinalis, Dictyota sp., Ulva spp. and Leathesia difformis macroalgae form dense turfs in the low intertidal zone. 2. Subtidal turf formed by red (Polysiphonia spp., Ceramium spp.), brown (Dictyota sp.) and green (Codium spp., Ulva spp.) macroalgae.

Seaweeds Ecology and Climate Change 167

The euphotic zone of the continental shelf that is not uncovered by tides is called the subtidal zone. In shallow subtidal areas, benthic algae can form extensive communities such as the giant kelp forests, if hard substrata is available. At greater depths, light becomes limited and algal diversity and abundance are limited to shade-tolerant forms (Dawes 1998). Subtidal seaweeds communities form important habitats and are the primary source of food for many organisms (Littler and Littler 1984).

Seaweed Communities under Stress

Rocky shores are subjected to considerable natural environmental change, the most obvious one is tidal amplitude that varies on short (daily to monthly) and long-term (annual and decadal) time scales. Occasional natural catastrophic events such as cold winters, hot summers, extreme storms, toxic algal blooms, earthquakes and volcanic events also occur (Denny and Paine 1998). Similarly, increased freshwater input from fl ooding rivers can also infl uence adjacent coastlines (Thompson et al. 2002). Rocky shore organisms have mechanisms to deal with stresses associated with alternating submersion and emersion in air (desiccation, temperature extremes, osmotic stress) (Karsten et al. 1996). The tolerance mechanisms involved are also effective for withstanding some anthropogenic stresses. Anthropogenic stress is the response of a biological entity to an anthropogenic disturbance. Many coastal ecosystems are subjected to a variety of stresses caused by human activities ranging from subsistence collection of food through the discharge of domestic and industrial effl uents and occasional catastrophic oil spills. These anthropogenic stresses are superimposed on the stress caused by natural environmental factors (Harvell et al. 1999).

Anthropogenic stresses infl uence the metabolism, activity patterns, respiration, growth, reproductive output and immune responses of intertidal organisms (Harvell et al. 1999). Numerous studies have examined mortality of intertidal organisms in relation to environmental stresses and attempted to relate these to their distribution (Lüning 1984). For most species, the distribution falls well within their ultimate physiological tolerances and is determined by interactions among several physical and biological factors (Raffaelli and Hawkins 1996). Hence the consequences of anthropogenic stress are often expressed by sub-lethal effects, which alter the competitive balance between species and indirectly infl uence their population distribution and abundance.

Climate change and global warming occurring on Earth are recognized phenomena within the public and scientific communities. The IV Intergovernmental Panel on Climate Change (IPCC 2007) reported changes in the abundance of seaweed and phytoplankton, a higher acidity of the


ocean waters and an increase in the surface seawater temperature. Another phenomenon associated to climate change is the increase of UV radiation on aquatic ecosystems.

Thermal stress

Temperature and distribution

Temperature is one of the most fundamental abiotic factors for organisms as it affects all levels of biological organization from molecular to community level, e.g., the physiology, ontogeny, trophic interactions, biodiversity, phenology and biogeography of organisms.

The geographical distribution of seaweeds depends upon habitat temperature regime and temperature requirement for growth, reproduction and survival (van den Hoek 1982a, b, Breeman 1988). Four critical temperatures are important determining factors in this distribution: the minimum temperature for survival, the minimum and maximum temperature for reproduction, and the maximum temperature for survival (Hutchins 1947). Such temperatures vary widely between species; as a general rule, seaweeds with a wide geographical range are eurythermal, that is, they are able to tolerate a wide range of temperature conditions, whereas stenothermal species have relatively limited geographical ranges since they are unable to tolerate large fl uctuations in temperature. For example, the stenothermal Antarctic endemic red seaweed Palmaria decipiens, grows at temperatures from 0 to 10°C (Wiencke and tom Dieck 1989). On the contrary, the eurythermal brown seaweed Fucus vesiculosus shows a wide geographical distribution from –2.1°C to 27.8°C, and has been recorded in several shores worldwide.

Seaweed communities are important indicators of climate change not only on a long time-scale but also on a short time-scale because of several reasons: a) generally, intertidal species respond more rapidly than their terrestrial equivalents to environmental changes because they usually have a shorter life-span associated with sessile adult stages (Southward et al. 2004); b) they are periodically exposed to temperature and weather extremes so many intertidal species live close to their thermal tolerance (Helmuth et al. 2002, Tomanek and Helmuth 2002); c) even when temperatures are not suffi ciently high to directly cause the death of the organisms (Denny et al. 2006), they may have sublethal effects that eventually lead to death, limited growth and reproduction (Breeman 1988, Somero 2002), increase in desease’s susceptibility or indirectly set limits on distribution by determining the outcome of biotic interactions (Sanford 1999, 2002); d) because of their position in the base of the marine food webs, changes in the composition of seaweed communities can substantially restructure food-web interactions amplifying their effect on the entire ecosystem.


Temperature and composition

Since sea temperatures govern global biogeographical distributions of seaweeds any temperature change is expected to affect the distribution range and seasonality of seaweeds (van den Hoek et al. 1990, Adey and Steneck 2001, Schils and Wilson 2006). Predicted increase in global temperature (IPCC 2007) is likely to have dramatic effects on the structure and function of ecosystems worldwide (Carlton 2001, Walther et al. 2002). Signifi cant changes in biogenic habitat structure including increase in monospecifi c kelp patches, decrease in mixed canopies, and change in fucoid species were found along a latitudinal gradient in ocean temperature equivalent to projected temperature increases for the coming 25–50 years in the southwest coast of Western Australia (Wernberg et al. 2011b).

In recent decades, global climate change (Hansen et al. 2006) has caused profound biological changes across the planet (Walther et al. 2002, Wernberg et al. 2011a). For example, the rockweed beds of southwestern Nova Scotia, which have been almost 99% pure Ascophyllum nodosum with a minor component of Fucus vesiculosus are undergoing a steady increase in F. vesiculosus since 2004 in coincidence with an increase in surface seawater temperature in the maritime region since 2000 (Ugarte et al. 2009). Along the Korean Peninsula, Sargassum, Laminaria and Ecklonia forests were abundant until the end of 1980 but since the beginning of 1990s, these forests had been decreasing due to various reasons including global warming (Kang 2010).

The impact of the rising temperature on other links in the food web also affects indirectly the composition of seaweed community. In Tasmania the strengthening of the East Australia Current (Ridgway 2007) and warming ocean temperatures, resulting in ocean temperatures exceeding the threshold for successful reproduction of the herbivorous sea urchin Centrostephanus rodgersii, has facilitated the poleward expansion of their populations (Ling et al. 2008). As a result, subtidal reefs that formerly supported dense stands of macroalgae have been intensively grazed and transformed into urchin barrens, with considerable loss of biological diversity (Johnson et al. 2005, Ling 2008, Ling and Johnson 2009).

Climatic warming on the time scale of decades may also alter the composition of the resident biota by facilitating the poleward spread of species characteristic of warmer temperature regimes (Southward et al. 1995, Holbrook et al. 1997, Sagarin et al. 1999). Since seaweeds are sedentary, they are incapable of a rapid response to climate variation by altering the pattern of individuals’ movements; however, changes in distribution would occur at the level of the population through changes in the ratios of extinction to colonization at the northern and southern boundaries of the range. For these species, responses to the warming trend should be slower, refl ected


in poleward shifts of the range. Therefore, a northward range shift occurs when there is net extinction at the southern boundary or net colonization at the northern while a southward range shift occurs when there is a net extinction at the northern boundary or a net colonization at the southern. Wernberg et al. (2011c) found that, on both the Indian and Pacifi c Oceans sides on the Australian continent, seaweed communities in the southern (poleward) part of the tropical-temperate transition progressively came to resemble past macroalgal communities farther north, the change was of similar magnitude on both coasts showing a rearrangement of entire local communities rather than mere shifts of a few individual species. In Sogn og Fjordane, Norway, a study of the macroalgal community at 22 sites between 1994 and 2004 showed a signifi cant increase in the abundance of southern species (Husa 2007).

Temperature and invaders

Non-native marine species are considered a threat to biodiversity because of their potential to compete with native species and preempt their resources, to alter energy fl ow through communities, to facilitate the introduction of other non-native species, and to homogenize regional biotic diversity (McKinney and Lockwood 1999, Grosholz 2005, Bulleri et al. 2008). While the threat from invasive species is not directly climate related, climate has often been proposed to facilitate the establishment, further spread and impact of invasive species on temperate reefs (Thresher et al. 2003, Scheibling and Gagnon 2009) since warming can allow warm water species to extend to or invade previously nonhospitable regions altering the competitive interactions between introduced and native species (Occhipinti-Ambrogi 2007). In Australia, for example, it was suggested that the spread of the shore crab Carcinus maenas from Victoria into Tasmania was facilitated by increasing ocean temperatures in response to a strengthening of the EAC (Thresher et al. 2003), C. maenas is voracious predator in intertidal and shallow subtidal reef habitants and exerts strong top down control of reef communities (Bertness et al. 2002). In Nova Scotia (Canada) warmer water increased fouling of kelp fronds by an invasive bryozoa, causing reduced reproductive output and defoliation of the kelps, which led to a switch in reef community structure (Scheibling and Gagnon 2009).

The rate of marine introductions, including introductions of seaweeds, has increased over the last 20 years. Marine macroalgae are also a signifi cant component of marine alien taxa (Schaffelke et al. 2006) with current global estimates of introduced species ranging from 163 to 260 species (Ribera Siguan 2002). The invasion of the green algae Caulerpa in the Mediterranean has profound effects both on the habitat and on ecosystem functioning. Caulerpa taxifolia originated from aquarium release on the coast of France (Jousson et al. 1998) and has expanded to very large areas of the Ligurian


and Tyrrhenian Seas (Meinesz et al. 2001). It has been described as ‘‘killer algae’’ (Meinesz et al. 2002) for its alleged property to cover extensively the substrata where it had been introduced without leaving any other previously established vegetation, and profoundly affecting all the biota. It has been regarded as the winner in the competition with the autochtonous Posidonia oceanica that forms the characteristic Posidonia meadows of the Mediterranean, which is now undergoing a rapid decline. The other introduced species is Caulerpa racemosa var. cylindracea (Verlaque et al. 2003), a south-western Australian variety, which is currently achieving dramatic and continuous expansion throughout most of the Mediterranean Sea and the Atlantic (Piazzi et al. 2001, Verlaque et al. 2000, 2004). The growth rate of Caulerpa spp. is correlated with temperature (Komatsu et al. 1997) and this is probably one of the reasons for the success of this species in the Mediterranean (Raniello et al. 2004, Ruitton et al. 2005).

Temperature and resilience

The range of temperature which a species can tolerate is determined by (a) genetic adaptation and (b) the ability to acclimate to changes in temperature (phenotypic adaptation) (Kuebler et al. 1991). Genetic differences in temperature tolerance and optimum growth temperatures are important because they expand the geographical range available to individual species. Acclimation is also important because it allows seaweeds to optimize photosynthesis, and hence growth, in response to seasonal changes in water temperature (Davison 1987, Egan et al. 1989). Kelps adjust their physiological performance in response to seasonal variation in temperature through metabolic or structural changes (Davison 1991), similar metabolic adjustments may, at least partially, offset physiological constraints of chronically warm environments (Stæhr and Wernberg 2009).

Acclimatization is, however, energetically costly (Clarke 2003) and is often achieved at the expense of processes that maintain reproduction and growth (Clarke 2003, Wernberg et al. 2010). Projected increases in temperature will therefore continue to reduce population resilience (Wernberg et al. 2010) and increase the likelihood to sudden range shifts in synergy with other stressors such as nutrient input and coastal development (Brook et al. 2008, Wernberg et al. 2011a). For example, the dominant habitat-forming seaweed in temperate Australasia, kelp Ecklonia radiata, is known to exhibit substantial physiological adjustment to seasonal (Fairhead and Cheshire 2004) and latitudinal changes in environmental conditions (primarily temperature and light conditions) (Stæhr and Wernberg 2009). While these adjustments enable the alga to maintain a positive metabolic balance, the cost appears to be reduced recruitment success and recruit performance with subsequent suppressed ability to recover from physical disturbances in relatively warm conditions (Wernberg et al. 2010). In other


words, increasing temperatures per se may not topple the algae, but rather gradually reduce their resilience to natural perturbations such that impacts manifest either abruptly when the physiological threshold of existence is fi nally exceeded or progressively as the cumulative effect of localized failure to recover.

UV stress

In their natural environment, seaweed are exposed to excessive solar PAR (photosynthetic active radiation, 400–700 nm) as well as to ultraviolet radiation (UV-B, 280-315 nm and UV-A, 315–400 nm), especially in the upper eulittoral and supralittoral zones (Hanelt 1998). UV radiation is usually considered harmful at either organism or community levels (Häder et al. 2007), specially the UV-B range which is increasing at the Earth’s surface as a consequence of the decrease in stratospheric ozone concentration (Björn et al. 1999, Bischof et al. 2006).

The increase in irradiance and light quality can promote photosynthesis, but also inhibit many biological processes if it becomes excessive (Barber and Andersson 1992), or if short wavelength radiation with high energy content, such as UV-B radiation, is absorbed by biomolecules (Vass 1997). Consequently, damage results in reduced photosynthesis and general metabolic activity leading to a decrease in biomass production (Helbling et al. 2003). Most of the seaweeds photoinhibition is due to PAR, as this waveband has a high proportion of solar radiation energy reaching the Earth’s surface. However, in the upper meters of the water column, a signifi cant percentage of photoinhibition is caused by UV-B, and to a lesser extent by UV-A (Dring et al. 1996, Häder 1997).

Positive effects of UV-A have also been reported. UV-A enhances carbon fi xation under reduced solar radiation (Barbieri et al. 2002), even in absence of PAR, and allows photorepair of UV-B induced DNA damage (Buma et al. 2003).

Biological effects of UV-B

The effects of UV-B exposure on biological systems range from molecular to organism level, thereby affecting growth and reproduction, and consequently the ecosystem structure and function (Bischof et al. 2006).

DNA is one of the most UV-sensitive molecules and UV-induced damage occurs directly by the absorption of UV-B quanta by aromatic residues, inhibiting replication or causing mutations, thereby affecting gene expression. UV-B is also absorbed by aromatic residues present in certain amino acids, and therefore affects proteins (Bischof et al. 2006). Lipids are also affected, which may be destroyed by UV-B in the presence of oxygen,


having a direct effect on membranes structure. Also, the generation of lipid peroxy radicals can induce further damage by participating in free radical cascades (Murphy 1983).

The effects of UV-exposure in seaweeds are multiple, including a decrease in the CO2-fi xation rate, oxygen evolution (Allen et al. 1997, Vass 1997), and photosynthesis activity due to the photodestruction of pigments (Bischof et al. 2002, Poppe et al. 2003). Also, the UV-B induces a decline in the Rubisco activity, which is related to the decreasing amount of its subunits as well as the corresponding mRNA levels (Bischof et al. 2000, 2002). Another effect of UV-B is the inactivation of the chloroplastic ATPase. Effects have also been observed on the ultrastructural level, producing changes of the fi ne structure of thylakoids and mitochondria (Holzinger et al. 2004).

Mechanisms to mitigate the UV radiation effects

Adaptation to UV radiation has equipped macroalgae with defensive mechanisms to minimize UV-induced damages. Seaweeds are able to protect themselves via avoidance, repair, and screening mechanisms (Karentz 2001). An important mechanism to reduce the damaging impact of UV radiation is the synthesis and accumulation of UV-absorbing compounds (UVAC). These compounds as mycosporine-like amino acids (MAAs), scytonemin, and phlorotannins have been found in many photosynthetic organisms. They function as passive shielding solutes by dissipating the absorbed short wavelength radiation energy in form of harmless heat generating photochemical reactions (Bandaranayake 1998). Synthesis of UVAC has been induced by UV-B in Chondrus crispus (Karsten et al. 1998), Porphyra columbina (Korbee-Peinado et al. 2004), and Ulva pertusa (Han and Han 2005). MAA compounds also have other ecophysiological functions such as protectors against desiccation or osmotic regulators, antioxidants, accessory pigments (biosynthesis of several carotenoids) (Korbee et al. 2006), and intracellular nitrogen storage (Korbee-Peinado et al. 2004, Korbee et al. 2006). De la Coba et al. (2009) reported the potential antioxidant capabilities of purifi ed aqueous extracts of the MAAs isolated from Porphyra rosengurttii, Gelidium corneum and Ahnfeltiopsis devoniensis.

Supra-and eulittoral Antarctic species experience the strongest solar radiation and synthesize and accumulate very high MAA contents, which are positively correlated with the natural UV doses (Huovinen et al. 2004) or affected by osmotic stress (Klisch et al. 2002) and/or nutrient availability (Zheng and Gao 2009). Red algae have the highest percentage of species that synthesize MAAs (Huovinen et al. 2004), followed by brown and green algae. Porphyra contain high levels of MAAs, being up to 1% of their dry weight (Hoyer et al. 2001).

The function of MAAs as intracellular screening agents has been inferred from a decrease in their intracellular concentration with increasing


depth (Hoyer et al. 2001, 2003). Many taxa growing in the sublittoral are not physiologically capable to produce MAAs, which explains their strong sensitivity to solar radiation. Different portions of seaweed thallus do not respond uniformly to the solar conditions. It was observed that young apical or marginal zones, with growing cells, synthesize and accumulate MAAs leading to cross sectional and longitudinal concentration gradients (Hoyer et al. 2001). Older tissue regions exhibit thicker cell walls and a leathery texture, and are therefore optically well protected. In contrast, higher MAAs concentration in the most exposed outer cortex is essential guarantee protection of the delicate meristematic cells (Bischof et al. 2006).

Other compounds that absorb UV radiation are phlorotannins, polymers of phloroglucinol, which can be found in physodes and cell walls of brown algae (Schoenwaelder and Clayton 1998, 1999). Phlorotannins have multiple roles (Arnold and Targett 2002, 2003, Lüder and Clayton 2004, Amsler and Fairhead 2006). Bischof et al. (2006) suggested four points to consider phlorotannins as UV-protecting compounds: (1) high tissue concentration which absorbs harmful radiation and prevents cell damages, observed in outer cell layers in Hormosira banksii (Schoenwaelder 2002); (2) harmful radiation induces its synthesis, performed in Ascophyllum nodosum after exposure of UV-B (Pavia and Brock 2000); (3) exudation in the surrounding medium shielding harmful radiation, observed in Macrocystis integrifolia (Swanson and Druehl 2002), in Eisenia bicyclis and Ecklonia kurome (Shibata et al. 2006); and (4) an excess inclusion of phlorotaninns in cell walls shielding harmful radiation. It was also demonstrated by Swanson and Druehl (2002) that seawater containing phlorotannin exudates of Macrocystis increased survivorship of germinating Laminaria groenlandica spores exposed to UV-B. Added to these, Schoenwaelder et al. (2003) linked higher numbers of physodes in Fucus spiralis embryos with a greater tolerance to elevated levels of UV-A and UV-B.

DNA damage can be repaired photoenzymatically in the presence of UV-A or blue light. This repair mechanism is known as “photoreactivation” or “photoenzymatic repair” and it reverses the photodimer products (cyclobutane-pyrimidine dimers; CPDs) (van de Poll et al. 2002). The dimers are cytotoxic because they block DNA and RNA polymerase and consequently inhibit genome replication and expression (Jordan 1996). Experiments in Palmaria palmata and Chondrus crispus after the exposure to artifi cial UV radiation evidenced no accumulation of DNA damage but decrease in CPD concentration (van de Poll et al. 2002). The accumulation of CPDs is accompanied by reduce growth rates in several phytoplankton and macroalgal species (van de Poll et al. 2002, Buma et al. 2003).


The effects of UV radiation on early settlement stages

Different stages of macroalgae show different sensitivity to UV radiation stress (Altamirano et al. 2003, Roleda et al. 2004, Véliz et al. 2006). UV-susceptibility of DNA damage is highly depending on the developmental stage of the species. Among the different stages in the life-cycle of seaweeds the unicellular propagules are clearly the most susceptible stages to UV radiation. The impact of UV-B on the early life stages of macroalgae is important in shaping up the structure of the community and zonation pattern (Bischof et al. 2006). Spores seem to be the most sensitive life history stage found in Laminaria digitata, Laminaria saccharina and Alaria esculenta and are strongly affected by increased UV-B radiation, both with respect to their photosynthesis performance and their susceptibility to DNA damage (Wiencke et al. 2000). Also juvenile stages of red and green algae showed a pronounced UV sensitivity (Han et al. 2003).

Swarmers of brown seaweeds use light directed movement (phototaxis) to assemble at the water surface improving the chances of fi nding a mating partner; however, that phototactic response is drastically inhibited by solar UV (Wiencke et al. 2006). Flores-Moya et al. (2002) found that the response of photomovement in Scytosiphon lomentaria and Petalonia fascia swarmers was negatively infl uenced by UVR. In the same way, Makarov and Voskoboinikov (2001) also observed that the zoospores of Laminaria saccharina ceased their movement, release and germination after treatment with UV-B. As a consequence of the UV-B effects, if the swarmers are the sexual gametes, fertilization is affected and life history may be blocked to a certain extent, but if they are zoospores recruitment of new thalli by asexual reproduction can be diminished (Wiencke et al. 2000, 2004).

Roleda et al. (2004) observed that the DNA damage in carpospores of Mastocarpus stellatus and Chondrus crispus was lower compared to zoospores of three species of Laminaria. Also, less genetic damage was observed in diploid carpospores compared to haploid zoospores (Roleda 2006), as the last ones are known to be less effi cient in DNA damage repair. In a sexual organism, during the diploid state, DNA damage can be repaired since there are two copies of the gene in the cell and one copy is presumed to be undamaged (Long and Michod 1995). This was observed in Laminariales, where haploid zoospores were more sensitive to DNA damage compared to the diploid young sporophytes (Bischof et al. 2006).

The effects of UV radiation at the celular level

UV-B radiation causes clear effects in cellular ultrastructure, especially in chloroplast and mitochondria of marine macroalgae. In red seaweeds,


such as Palmaria decipiens, Palmaria palmata, Phycodrys austrogeorgica and Bangia atropurpurea, thylakoids form abnormal vesicles upon exposure to artifi cial UV radiation (Poppe et al. 2002, 2003). UV-induced changes in the membrane structure of mitochondria were observed in Palmaria decipiens and Palmaria palmata. Protein crystals occurring in the cytoplasm of Phycodrys austrogeorgica showed degradation after UV radiation. These fi ndings gave fi rst insight into the fi ne structural changes during and after UV exposure and confi rmed the results on inhibition of the photosynthesis performance by UV (Poppe et al. 2002, 2003).

Other negative effects of UV radiation especially on macroalgal spores concern the cytoskeleton. Nuclear division and the activity of the fl agellar apparatus depend on a functional microtubular system, which might be damaged by UV. No polarization has been observed in UV exposed zygotes, they remained spherical and there was no further development. In zygotes of two species of Fucus, actin inhibitors prevented polarization, cross wall formation and vesicle movement after UV treatment (Schoenwaelder and Clayton 1999).

Sea level rise

Changes in sea-level are natural phenomena that periodically occur in geological time of Earth as a consequence of geological and oceanographic process such as melting of glaciers and ice caps (Church et al. 2001). These processes have been molding marine and coastal ecosystems as well as the communities that live in them over thousands of years. There is a general consensus that this process is today increasing at a signifi cantly faster rate compared to earlier centuries (Hekstra 1989, Church et al. 2001). The possible effect of the increase in sea-level on seaweed communities has not been deeply studied, however it can vary according to the environment they inhabit and largely depends on the species. The expected consequences of the increase in sea-level on seaweeds range from changes in the substrate available for propagules to changes in the distribution limits of entire communities. While the effect of rising sea level on seaweed begins at individual and species level, it is expected that the ultimate long-term consequence would be the change in the spatial distribution and species composition of assemblages.

Many authors have correlated the vertical distribution of seaweed with tide levels. In this regard, the sea-level rise is a direct infl uence on the distribution of seaweeds in the intertidal and subtidal habitat. Predictions suggest that the increase in sea-level could permanently submerge some intertidal areas while others could be created, thus changing the mosaic of seaweed communities along the coast. On the other hand, in shallow coastal regions, the fl ood of new substrates would increase the availability of


substrate for propagules settlement allowing colonization and development of new subtidal populations, increasing the range of spatial distribution of seaweeds. However, subtidal seaweed communities would also be affected due to increased depth and the consequent reduction of the incidence of light on the water column would limit photosynthesis. These changes would lead to the displacement of their communities from the subtidal towards the intertidal resulting in changes of zonation patterns (Harley et al. 2006).

Another less direct consequence, but with great long-term effect of sea-level rise comes from the change in the relationship between vertical and horizontal surfaces of the substrate, especially in areas where the tide amplitude is small (Vaselli et al. 2008). In regions where most of the rocky coast is horizontal, a rapid increase in sea-level could cause fl ooding of most of the intertidal zone, transforming the platforms into subtidal reefs.

Rising sea level can also affect the diversity of seaweed communities in coastal areas through changes in the characteristics of the substrate. Field measurements on the rocky coast of the Mediterranean Sea showed that an increase in sea level in the range of 5 to 50 cm may increase the availability of steeped substrate up to 58% compared to current conditions and suggested that a change in the tilt of the substrate affects the ability of assemblages to recover after a disturbance; leading to an expansion of assemblages dominated by crustose algae and reduced abundance of fi lamentous forms (Vaselli et al. 2008).

A valuable and highly diverse coastal ecosystem that is affected and threatened by the rise in sea-level is the coral reef, which exhibits high diversity of seaweed species (Wilkinson 2004, Hoegh-Guldberg et al. 2007). One of the direct effects of rising sea level on seaweed that live on coral reef is the habitat loss due to subsidence of these reefs as sea level rises over them (Diaz-Pulido et al. 2007). Another equally important consequence is that the rise in sea level could induce an acceleration of vertical growth rates of coral to a maximum level, although this would not be enough to keep off the rate of increase in sea level and therefore less protected reefs will decrease their growth rate as a result of fl ooding and erosive effect that the waves could produce grades in them (Buddemeier and Smith 1988).

Ocean acidifi cation

Carbon dioxide (CO2) levels in the atmosphere have strongly increased during the last millennium as a result of human activities related to the use of fossil fuel, deforestation and agriculture (IPCC 2001). This sudden increase in atmospheric CO2 concentration levels brings several global side effects, and one of the most important is the ocean acidifi cation.


Ocean acidifi cation is a consequence of the incorporation of high amounts of atmospheric CO2 to the seawater through physical processes such as waves and rainfall. Once the CO2 is absorbed by the seawater, it becomes carbonic acid (HCO3H), a weak acid which results in an increase of H+ and HCO3- and an increase of dissolved CO2, leading to a reduction of carbonate ions (CO3

2–) and a consequent change in pH. The ocean surface is slightly alkaline with pH values of approximately 8.2, however estimates indicate that if CO2 emissions to the atmosphere continue towards the end of this century, the pH of seawater is expected to drop others from 0.3 to 0.4 units, resulting in an increase of 150% in the H+ ions and a corresponding increase in the availability of CO2 from 300% to 400% (Arnold et al. 2012).

The pH is a critical variable in marine systems. It is known that small changes in pH can have a deep impact on the dynamics of chemical compounds, nutrients and trace metals in oceans. Because the incorporation of high levels of CO2 in the oceans directly affects the pH of the seawater, the fi nal impact of acidifi cation on marine biological systems is extremely high. Many organisms are sensitive to changes in carbonate chemistry and their responses to these changes can lead to profound ecological changes in marine ecosystems (Doney et al. 2009, Rilov and Trebes 2010).

Seaweed responses to ocean acidifi cation

Seaweeds play a fundamental role in carbon cycle of coastal ecosystems (Reiskind et al. 1989). Their response to changes in seawater CO2 can be very different for each species; this is because different species have different strategies to obtain the carbon needed for photosynthesis (Kroeker et al. 2010, Zou and Gao 2010). Several studies on the potential effect of seawater acidifi cation on seaweeds have shown highly variable responses, including positive, negative or neutral responses (Hall-Spencer et al. 2008, Wu et al. 2008, Roleda and Hurd 2012). Understanding the sensitivity of seaweed to ocean acidifi cation is important to understand the impacts that this phenomenon may have on coastal ecosystems.

The consequences of ocean acidifi cation on seaweed-based coastal ecosystems range from organism to community levels. On one hand, this process may be able to infl uence physiological processes of seaweeds. On the other hand, ocean acidifi cation may infl uence the process of tissue calcifi cation in those species that use calcium carbonate to synthesize structures. Changes in any of these metabolic processes result in a direct impact on seaweed growth.

Photosynthesis and growth in the context of seawater acidifi cation

CO2 is the substrate for the process of photosynthesis, thus it is expected that an increase in its availability as a result of ocean acidifi cation would


result in an increase in photosynthesis rates of seaweeds, and consequently increased growth rates. This response to acidifi cation of the seawater would be positive in the case of CO2-limited seaweeds; however the response may be very different depending on the species.

According to the results of various experiments, the response of seaweed photosynthesis against high concentrations of CO2 in water is highly variable and species-specifi c. Some species have lower photosynthetic rates under culture conditions with high concentrations of CO2 (Johnston and Raven 1990, Garcia-Sanchez et al. 1994). Other studies have demonstrated higher photosynthesis rates and increases in growth rate of red algae in presence of double levels of CO2 than today’s atmospheric concentrations (Gao et al. 1991, Andría et al. 1999, Kübler et al. 1999, Zou 2005); however, the opposite effect has been found in other species of red seaweeds (Garcia-Sanchez et al. 1994, Mercado et al. 1999, Israel et al. 1999). Furthermore, studies on species from the three main groups of seaweed have shown that growth rates in the presence of CO2 enriched seawater could be similar to those obtained in seawater without enrichment (Israel and Hophy 2002). Many seaweeds have carbon concentration mechanisms, allowing them to use bicarbonate for photosynthesis (Gao and McKinley 1994, Beardall et al. 1998). The general consensus is that seaweeds having these mechanisms would not increase their productivity in the future conditions of high CO2 (Beardall et al. 1998). Some experiments have also found that seaweeds can acclimate to achieve high concentrations of CO2 decreasing the content of their pigments (bleaching), similarly to the response of seaweeds to high irradiance conditions (Garcia-Sanchez et al. 1994, Mercado et al. 1999, Andría et al. 1999, 2001).

Intertidal seaweeds experience the daily exposure to air and during this emersion period they must use the atmospheric CO2 in the photosynthesis process which diffuses much faster than in water (Raven 1999). For these species living in the intertidal, the increase in atmospheric CO2 levels may have advantages over other species by increasing their photosynthetic production during the periods of emersion (Gao et al. 1999, Zou and Gao 2002, 2010, Zou et al. 2007). In these cases, growth responses to increases in the availability of CO2 could be much lower compared to subtidal species (Beardall et al. 1998).

There is also evidence that increased CO2 concentrations in the seawater may increase nitrogen assimilation rates in the form of nitrate in red, brown and green algae, which could be related to higher nitrogen requirement to withstand high growth rates (Gao et al. 1993b, Gordillo et al. 2001, Zou et al. 2001, Zou 2005).


Effects of seawater acidifi cation on seaweed population and community ecology

Unlike the projections about the potential effects of ocean acidifi cation at physiological level of seaweeds, which can be estimated using controlled culture experiments of individuals, the predictions for populations and communities from an ecological point of view are more complex. Attempts to determine whether the predictions based on laboratory experiments translate to the conditions of the natural environment have been hampered by the diffi culty to simulate the conditions of ocean acidifi cation and to test the hypotheses in experiments that require the manipulation of CO2 concentration for long periods of time. However, some studies have drawn these diffi culties by taking advantage of high CO2 conditions present in natural environments that may simulate future conditions such as pH gradient that occur as a result of volcanic discharges or leaks (Hall-Spencer et al. 2008, Porzio et al. 2011). In example, a study about the impact of decreased pH on seaweeds communities have predicted changes in species richness, decrease of turf forming species, inhibition of reproduction, decrease in coverage and number of calcifi ed seaweeds and dominance of a few species of non-calcifi ed seaweeds. These results were exacerbated when the pH reached very low levels (Porzio et al. 2011). Another recent study at ecosystem-level showed that communities of rocky shores where calcareous organisms were abundant shifted to communities with scarcity of corals and signifi cant reductions in coverage and abundance of Corallinaceae under conditions of low pH. On the other hand, some seaweed genera were resilient to high levels of CO2, such as Caulerpa, Cladophora, Asparagopsis, Dictyota and Sargassum, some of which are considered alien species (Hall-Spencer et al. 2008).

Ocean acidifi cation and calcifi ed seaweeds

Many studies concerning the effect of ocean acidification on marine organisms have shown that calcifi ed organisms would be more susceptible due to the reduced availability of the chemical constituents that they use to synthesize calcifi ed structures (Kroeker et al. 2010). Seaweeds that have calcium carbonate structures include crustose seaweed and seaweed belonging to the group Corallinaceae. The calcifi ed seaweeds use aragonite or calcite as a form of calcium carbonate to synthesize their thalli (Lowenstam 1981). The direct effect of seawater acidifi cation on these compounds is the reduction of their concentration and therefore its availability in the form of carbonate ions (Díaz-Pulido et al. 2007). These conditions may predict a reduction in the rate of calcifi cation of the structures of calcifi ed seaweeds facing CO2 enrichment conditions due to


dissolution of their carbonate structures. Some experiments have tested this hypothesis and have obtained signifi cant results (Borowitzka and Larkum 1976, Gao et al. 1993a, Gattuso et al. 1999, Orr et al. 2005). Studies by means of controlled experiments have shown bleaching processes in the crustose seaweed Porolithon onkodes and a consequent sharp decrease in productivity with reductions in calcifi cation rate of up to 190% (Anthony et al. 2008). On the other hand, Johnson et al. (2012) found a positive response to seawater acidifi cation on calcifi ed brown algae Padina from temperate and tropical regions despite the reduction in their content of CaCO3 with CO2 increases. Contrary to other studies of calcifi ed seaweed, they found that there was an increase in the abundance of the species of Padina spp. under conditions of acidifi cation. This phenomenon could be explained by reduced grazing pressure by sea urchins and the signifi cant increase in the rates of photosynthesis (Johnson et al. 2012).

There is evidence that photosynthesis can stimulate calcification in calcifi ed seaweeds, thus it is expected that the increase in rates of photosynthesis due to the increased availability of CO2 in the seawater could also act as a buffer mechanism against the negative effects of acidifi cation on the calcifi cation process of calcifi ed seaweed (Gatusso et al. 1999). Recent studies have found that those seaweeds that secrete an external organic layer between the skeleton and the seawater, such as Corallinaceae and calcareous green algae, increase net calcifi cation under intermediate levels of CO2 and generally exhibit high resilience, which may suggest that the direct use of CO2 through photosynthesis can infl uence the response of seaweed to calcifi cation. Despite the complexity of the relationships between the processes of calcifi cation and photosynthesis, an increase in CO2 in water may increase the rate of photosynthesis (Bowes 1993, Iglesias-Rodriguez et al. 2008) and potentially increase the amount of energy available to convert HCO3 to CO3

2– through pH regulation in the sites where calcifi cation occurs (Ries et al. 2009). The studies of Hofmann et al. (2012) support the above mentioned hypothesis since they found that Corallina offi cinalis decreased its growth rate under high CO2 levels, while the carbonic anhydrase activity was stimulated and negatively correlated with the inorganic content of the algae.

Connell and Russell (2009) studied the effect of elevated CO2 levels on non-calcareous seaweeds and tested the hypothesis of an increase in the abundance of turfs under the combined conditions of elevated CO2 and temperature. The results showed an interaction between temperature and CO2 levels on turf coverage and a positive effect of both factors on turf biomass. They also found that the loss of kelp canopies may be exacerbated by these positive effects of temperature and CO2 on species turfs, inhibiting the recruitment of kelp propagules.


Interaction between seaweeds and corals in the context of ocean acidifi cation

General predictions suggest that the ocean acidifi cation can lead to reduced calcifi cation rates of corals (Kleypas and Langdon 2006, Anthony et al. 2008). Reducing the potential for coral growth would lead to changes in the competitive process between corals and seaweeds and promote changes leading to the dominance of seaweeds (Carilli et al. 2009).

A recent study has attempted to predict that the increase of CO2 modifi es the interactions between seaweeds and corals. The results showed that the CO2 level is signifi cantly correlated with the decline in coral growth rates and also with coral mortality which was also increased by contact of the coral with the seaweeds (Diaz-Pulido et al. 2011). In general, an increase in growth rate of seaweed coupled with a decrease in coral growth rate and increased coral mortality with increasing atmospheric CO2 would give rise to changes in the competitive interaction between corals and seaweeds and therefore the dynamics of reef communities would be altered (Diaz-Pulido et al. 2011). Similar results have been found by other authors who have predicted that ocean acidifi cation will have a signifi cant reduction in the resilience of coral specifi cally because it will reduce the growth rate due to accretion, which in addition to interaction processes between corals and seaweed, would lead the community to states of seaweed dominance (Hoegh-Guldberg et al. 2007, Anthony et al. 2011).

The effects of ocean acidifi cation on seaweed epibionts

Seaweeds are key species of some benthic coastal ecosystems in which they constitute the main substrate for many epibionts such as bryozoans and tubeworms. Epibionts of seaweeds may also be affected by acidifi cation of the oceans, either directly due to changes in abundance of seaweed or indirectly through changes in the physiology of the host. A recent study has shown that biological activity of the host could modulate the pH and therefore the conditions of CO2 in the boundary layer of the thallus where the epibionts live due to a dominance of the photosynthetic activity over acidifi cation (Saderne and Wahl 2012).

General Conclusions

Several aspects of the biology and ecology of seaweeds are infl uenced by global environmental changes such as increase in superfi cial seawater temperature, solar UV-B irradiance, atmospheric CO2 concentration, and sea level (Fig. 3). Elevated temperature negatively affects most aspects of the biology of kelps and fucoids (e.g., reproduction, recruitment and growth, and resilience to disturbances). Hence, projected global warming


is causing serious concern for permanent changes in habitat structure and loss of seaweed habitats. Moreover, not only the biodiversity of the algae themselves is under threat, they are foundation species which provide and modify resources available to other organisms both in terms of environmental conditions, habitat structure and food. Because of their central role, changes in seaweed population are likely to affect associated species populations and the ecological function of the entire ecosystem.

UV radiation is usually known to harm physiological process in macroalgae. During their long history of evolution, seaweeds have developed protective strategies against harmful UV irradiances, such as synthesizing and accumulating UVAC and the repair of DNA damage. Different life stages of seaweeds show different sensitivity to solar UV radiation, with less-differentiated forms being more sensitive. Species distributed at different depths in the intertidal zone also show different responses to solar UVR; upper species, that are usually exposed to higher levels of solar radiation and accumulate higher contents of UVAC (such as MAAs) are more tolerant of UVR.

Ocean acidifi cation as a consequence of high levels of CO2 in seawater could affect seaweed populations and communities in different aspects, and seaweed responses are highly variable. The main effects at organismic level would include increased photosynthesis and growth rates and disruption of the calcifi cation processes in calcifi ed seaweeds. Among the effects of ocean acidifi cation at community level, the most important include changes in interaction processes between corals and seaweeds leading to states of seaweed dominance. It is expected that the responses of seaweeds to reduced

Fig. 3. Schematic diagram of the possible consequences of climate change on seaweed benthic communities.


pH have considerable impacts on the functioning of the community where they inhabit, including the epibionts that live on them.

Acknowledgments

The authors want to thank The Secretary of Science and Technology of the Universidad Nacional del Sur and CONICET, and the editors for their useful comments.

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Zou, D.H., K.S. Gao and Z.X. Ruan. 2007. Daily timing of emersion and elevated atmospheric CO2 concentration affect photosynthetic performance of the intertidal macroalga Ulva lactuca (Chorophyta) in sunlight. Bot. Mar. 50: 275–279.

CHAPTER 8

World Fisheries and Climate Trend

Ana Laura Delgado1,* and Maria Celeste Lopez Abbate2

Introduction

Climate oscillations have always occurred in history and natural systems have developed a capacity to adapt, by migrating to nearby stable domains. This ability will enable ecosystems to mitigate the impact of future changes, however, there are two factors which will limit the ecosystem adaptability, 1) the current exacerbated rate of climate change compared to previous natural changes, and 2) the restricted resilience of species as a result of human pressure which has caused overfi shing, loss of biodiversity, habitat destruction, pollution and introduction of invasive species and pathogens (Brander 2010).

The influence of human activities upgrade the impact of climate change over marine ecosystems, thus, recognizing the additive effect of both sources of extrinsic variability will help reduce the uncertainty when predicting the response of marine communities (Perry et al. 2010a). The concept of social-ecological systems (fi rstly proposed by Berkes and Folke 1998) can be thought of as the interplay between two subsystems: the biophysical (including biology and climate) and the human (including

1 Instituto Argentino de Oceanografía (IADO-CONICET), Camino La Carrindanga, km7, (8000) Bahía Blanca, Argentina.

2 Instituto Argentino de Oceanografía (IADO-CONICET), Camino La Carrindanga, km7, (8000) Bahía Blanca, Argentina.


World Fisheries and Climate Trend 195

cultural, management, economic, socio-political and ethical aspects) and has proved to be useful to assess the complex responses of ecosystems to global climate change (Perry et al. 2010a, Perry et al. 2010c).

Fishing productivity has grown rapidly from 1950 uptil 1980 when maximum yield was reached; however, the global landings are currently stacked and might be decreasing (FAO 2007, Dow et al. 2009). Indeed, many commercially important fi sh populations have been declining in the past several decades (Myers and Worm 2003, Hutchings and Reynolds 2004), though the exact role of fi shing and environmental change as underlying factors driving the changes is not yet clarifi ed (Hsieh et al. 2006).

The goal of this chapter is to summarize the current and future impacts of climate-driven changes on the physiology and ecology of marine fi shes, and how world fi sheries are responding to the observed changes. The interaction between fi sh stocks and climate change is analyzed on the context of human disturbance. Finally, we present three case studies of fi sheries in South-America, illustrating the vulnerability, possible mitigation and future perspectives in the climate change context.

Climate Change and Fishes

Effects of climate warming

Individual-level responses. Marine organisms are capable of synchronizing developmental, reproductive and migratory cycles to periodic climate oscillation (Overland et al. 2010). Within climatic variables, temperature is recognized to infl uence the fate of biochemical reactions and to dictate the rate of almost every biological activity. Indeed, metabolism increases exponentially with temperature, and the metabolic rate of individuals determines the rate of resource uptake in a given habitat, which affects the ecological processes in all levels of organization (Brown et al. 2004). However, higher metabolic rates involve higher oxygen demand, which in turn becomes less available as the temperature increases (Keeling et al. 2010). As a consequence, aquatic species in a warmer environment must confront the paradox of higher tissue oxygen demand and lower habitat oxygen saturation.

The mechanism by which seawater warming affects the physiology and ultimately the local adaptation of fi sh species was described by Pörtner and Knust (2007). The authors suggested that once aerobic capacity (i.e., the capacity of an organism to fulfi ll tissue oxygen demands) starts to be restricted, the threshold critical temperatures are surpassed (Fig. 1). This value might not be the same for the entire population, since smaller specimens present a wider thermal tolerance and thus will continue to grow after higher critical temperatures are reached. Beyond specifi c critical


temperature, the fi tness of the individuals starts to decrease and ultimately, fi sh population experience enhanced mortality of larger animals (Fig. 1). It is worth noting that critical temperature is well below the temperature needed for the onset of anaerobic metabolism or for cell damage, thus the consequences of slightly above normal temperature are expected to yield low population abundance (Pörtner and Knust 2007).

Body size is another specifi c trait of ectotherms that can be shaped by temperature (Atkinson 1994). Within a single species, body size tends to decrease in the warmer region of its geographical distribution (James 1970), thus, it is expected that climate warming benefi ts the smaller species in aquatic environments. In fact, Daufresne et al. (2009) compiled data from several fi sh species in environments that experienced signifi cant warming in the last 14–27 years, and found that mean body size decreased with gradual

Fig. 1. Temperature infl uence on aerobic metabolism of marine fi shes. The concentration of dissolved oxygen ([O2]) in seawater varies negatively with temperature, and both variables determine the fate of aerobic metabolism (solid curve) of fi shes. The aerobic capacity starts to be restricted after a threshold temperature (optimal T) is surpasses, after the onset of the anaerobic metabolism (dashed curve). When temperature continues to grow, the concentration of dissolved oxygen becomes limiting and after the second threshold is reached (detrimental T) the anaerobic metabolism is activated, reducing species fi tness. Beyond the upper tolerance limit (critical T), the aerobic capacity is null and the individuals are sustained by anaerobic metabolism and the cells starts to be damaged. Additionally, larger animals are expected to experience temperature-dependent aerobic metabolism earlier than smaller individuals (body size effect).


environmental warming. The shift occurred at every level of organization. At the individual-level, the authors found that the fi nal size of adults was lower at higher temperatures. At the population level, a dual response was found, since intra-specifi c size changed negatively with warming and the age-structure of the population also modifi ed by an increase in the number of juveniles. In addition, temperature can shape the population structure by infl uencing the reproductive traits, for instance, the spawning period, age at sexual maturity, and egg and larval size are determined by temperature in most species (e.g., Crozier et al. 2008, Hutchings and Myers 1994, Jonsson and Jonsson 2004). Finally, at the community-level warming benefi ted the smaller species over the bigger, leading to a shift on the community structure (Daufresne et al. 2009).

Special attention has been given recently to the interplay between climate warming and the occurrence of marine diseases and parasitism. Warmer temperatures are expected to benefi t the development, dispersal and transmission of pathogens (Harvell et al. 2002, Parmesan 2006, Walther et al. 2009), while climate-driven changes and human activities are expected to accelerate the emergence and geographical expansion of infectious diseases and parasitism (Brooks and Hoberg 2007). However, the long-term response of pathogens to climate warming in marine systems seems to involve several species-specifi c and host-specifi c aspects and is under considerable debate (e.g., Ward and Lafferty 2004).

Changes in community abundance, structure and phenology. Temperature might have a direct infl uence on population abundance by affecting individual´s physiology. Furthermore, population abundance and biomass are indirectly infl uenced by temperature since it determines the magnitude of growth and recruitment rate, as well as the geographical distribution of species. Also, by changing the distribution and availability of plankton, climate warming might change carrying capacity of an ecosystem, leading to changes in the abundance and structure of local fi sh communities (Beaugrand et al. 2008). For example, temperature infl uences the population abundance of the Atlantic cod (Gadus morhua) by affecting larval recruitment, however, as temperature rises cod stocks increase in the lower temperature range and decrease in the upper temperature range (e.g., Sundby 2000). On the other hand, the long-term temperature increase had a net positive effect on the recruitment of pipefi sh (Entelurus aequoreus) in the Northeastern Atlantic since 2002, probably as a result of the benefi cial effect on the reproductive trait of the species (Kirby et al. 2006).

The occurrence of community reorganization and ecological regime shifts has also been associated to climate change (Stenseth et al. 2002). Fish communities on the North Sea have experienced profound modifi cations since 1970, given that smaller species benefi ted by climate warming and


partially replaced bigger species (Hiddink and Hofstede 2008). Community reorganization may also occur as a result of the positive effects of climate warming on diverse marine organisms, for example, in the last decades an apparent increase in jellyfi sh blooms and a shift from fi sh-dominated to jellyfi sh-dominated trophic webs has been registered in many aquatic habitats (e.g., Mills 2001). The factors leading to this shift include climate variations, overfi shing, eutrophication and habitat modifi cation (Purcell et al. 2007). One of the most concerning consequences of jellyfi sh outbreaks and the replacement of top predators, is that the carbon of planktonic origin is rerouted into respiration instead of being effi ciently transferred into upper trophic levels (Condon et al. 2011). Thus, the reorganization of local food webs as a consequence of global warming, increasing climatic uncertainty and anthropogenic activities, might lead to profound ecosystem modifi cations by altering the functioning of the biological pump.

The seasonal recurring cycle or phenology of marine organisms is mainly driven by environmental signals and the seasonal pattern of resources availability, and is already being modifi ed by climate change (Edwards and Richardson 2004). Plankton represents a key food item for early stages of fi shes, thus the trophic coupling between early stages of fi shes and mass productivity events in aquatic ecosystems sets the magnitude of annual recruitment and population abundance and determines the survival success of fi sh species. Changes in the plankton phenology as a consequence of temperature shifts lead to trophic mismatch and subsequently, to larval starvation and decreasing recruitment (Cushing 1990, Durant et al. 2007, Edwards and Richardson 2004). The trophic mismatch occurs after the temporal decoupling cannot be overcome by phenotypic plasticity of predators, and results in prey biomass accumulation due to ineffi cient exploitation (Donnelly et al. 2012). Hence, the effects of climate change on the timing of primary producers can be transferred up in the food chain, impairing the effi ciency of local food webs.

Changes in the distribution range. Climate change is gradually changing the mapping of ocean biodiversity as a consequence of changes in species distribution. As temperature continues to rise, marine species will migrate pole-ward in search for suitable environmental conditions and resources (Parmesan and Yohe 2003). Functional traits and adaptability of species will set the potential for colonizing new geographic areas, and many will inevitably face local extinction (e.g., Root et al. 2003, Thomas et al. 2004, Cheung et al. 2009). Fish species are highly susceptible to experience distributional shifts as a result of climate warming given that their spreading ability is promoted by the migratory capacity of juveniles and adults, eggs and larval dispersal strategies and the lack of geographical barriers in oceans (Metcalfe et al. 2002). Those species with short developmental times and


small body size are particularly well adapted to change their distributional range (Perry et al. 2005). Behavioral adaptation of fi sh species to climate warming also includes vertical migration in the water column in order to encounter deeper and colder waters (Perry et al. 2005, Dulvy et al. 2008).

Several fi sh species have already extended their range distribution in response to climate warming. For instance, in the North Sea temperature rising matched with an increase of species richness, as a result of a range expansion of small species from lower latitudes (e.g., Brander et al. 2003, Wynn et al. 2007, Hiddink and Hofstede 2008). In the southern hemisphere (Tasmania, southeastern Australia), several inshore species experienced pole-ward movements since 1970 in response to sea temperature rising (Last et al. 2011). Species from cold waters in high latitudes are also undergoing distributional shifts. For example, as a result of sea ice retreat in the Bering Sea, a new habitat has emerged northward which used to be permanently covered by ice, and has been colonized by several invertebrate and fi sh species (Mueter and Litzow 2008). The distribution and diversity of top predators in the North Pacifi c are expected to change under the current climate trend, since some predator species (e.g., the shark guild) might be impaired by habitat compression, while others (e.g., tuna) might benefi t by gaining core habitat (Hazen et al. 2012). However, the synergistic effect of climate change and human activities might exacerbate population vulnerability and lead to irreversible shift on community confi guration.

Besides the direct effect of temperature, the shifts in the geographical pattern of preys as a result of climate driven changes, may indirectly affect the range distribution of many fi shes. Due to their small size and rapid turnover rates, phyto- and zooplankton experience rapid response to warming. Many studies conducted in diverse habitats have registered shifts on the distribution map of diverse planktonic communities (e.g., Mackas et al. 2007, Richardson and Schoeman 2004, Beaugrand et al. 2002) as well as changes in the pattern of ocean productivity (Behrenfeld et al. 2006). These major changes permeate the entire ecosystem through cascading effects in the trophic web, and determine the fate of fi sheries worldwide (Brown et al. 2010). In the North Sea, the abundance and survival of cod larvae (Gadus morhua), are directly linked to the cold water copepod Calanus fi nmarchicus (Beaugrand and Kirby 2010). However, in recent years this species have retracted toward the North Pole and is thought of as highly vulnerable to climate stress due to shrinkage of its ecological niche (Helaouët et al. 2011). This copepod species represents a key food item for various fi sh species, including the cod, and it has been hypothesized that the reduced geographical availability of copepods leads to enhanced cod larvae mortality in the North Sea (Beaugrand et al. 2003). In the future, the cumulative and probably intensifi ed effects of climate warming will have an increased


impact on fi sh community, which will rely on geographic gradient, species-specifi c plasticity and the strength of community interactions (Roessig et al. 2004).

Effects of other climate-driven changes in aquatic systems

Besides temperature rising, several abiotic changes have permeated the marine ecosystems as a result of global climate change. These include increase of atmospheric CO2 concentration, ocean deoxygenation, enhanced occurrence of extreme weather events, and changes in ocean circulation patterns.

CO2 increase. The increase of CO2 emission has generated an increase on the concentration of CO2 in the world’s oceans, which is evidenced by a reduction of seawater pH and ocean acidifi cation (e.g., Caldeira and Wickett 2003, Orr et al. 2005, Wootton et al. 2008). Lower ocean pH impacts negatively on calcifying organisms such as phytoplankton, corals, mollusks and crustaceans (Orr et al. 2005); produces severe tissue damage of fi sh larvae (Frommel et al. 2011), and impairs food quality and transfer efficiency in aquatic communities (Rossoll et al. 2012). Physiological responses of aquatic organisms to decreasing seawater pH remains almost unknown, however, recent studies demonstrated that olfactory and auditory capacity of fi shes can be disrupted when exposed to pH levels close to the ones expected to occur in 2100. Tropical fi sh larvae proved to be unable to discriminate olfactory and auditory cues for detecting predators and suitable settlement habitats when exposed to lower pH levels (Munday et al. 2009, 2010, Simpson et al. 2011). The modifi cation of such essential behavioral trait during early life history stages might ultimately lead to reduced adult survival and species recruitment. Additionally, fi sh otoliths are vulnerable to increasing CO2 concentration as they are composed of aragonite. These constituent of fi sh ears are essential to adequately detecting external auditory stimuli and body orientation, and when exposed to high CO2 concentrations, otolith become overcalcifi ed and lose functionality (Checkley et al. 2009, Manjela et al. 2012a, Munday et al. 2011).

Brain function is also affected by the concentration of CO2 by altering fi sh lateralization, which is a key mechanism involved in behavioral activities including predator avoidance (Domenici et al. 2012). Moreover, the swimming behavior of the Atlantic cod was also shown to respond to increasing CO2 concentration (Manjela et al. 2012b). Behavioral shifts after high CO2 exposures might decrease predator avoidance and species fi tness.

Deoxygenation. Deoxygenation of certain oceanic regions is expected to occur as a consequence of a decrease in oxygen solubility due to elevated


seawater temperature, and a lower oxygen exchange with underlying layers due to enhanced stratifi cation (Keeling et al. 2010). Moreover, the emergence of oxygen minimum zones in coastal environments (Diaz and Rosenberg 2008), tropical oceans (Bograd et al. 2008, Stramma et al. 2008), and the North Pacifi c Ocean (Whitney et al. 2007, Pierce et al. 2012) have increased in the last decades resulting in habitat compression and mass mortalities. Dissolved oxygen concentration strongly constrains the development of most fi sh species, which are recognized as especially vulnerable to hypoxic conditions (Vaquer-Sunyer and Duarte 2008). Although much research is being currently undertaken, some studies have already highlighted the profound infl uence of decreasing oxygen on marine fi sh. For example, hypoxic conditions impair predator avoidance capacity of mesopelagic fi shes in the California Current, since fi sh are forced to move vertically from deep water to better oxygenated-surface waters, where they are highly vulnerable to visually-orientated predators (Koslow et al. 2011). Other authors predicted a considerable habitat loss for benthic fi shes as a consequence of the shoaling of the upper hypoxic zone in the Southern California Bight (McClatchie et al. 2010). This phenomenon is likely to affect most marine habitat worldwide, and is expected to be more severe in shallow areas where the hypoxic conditions can easily be transferred to the entire water column. In fact, the shoaling of hypoxic layers triggered a substantial vertical habitat loss in the tropical Atlantic Ocean, which could lead pelagic species to be highly vulnerable to surface fi shing gear and thus could be potentially exposed to overfi shing (Prince et al. 2010, Stramma et al. 2012).

Local weather variability. Modifi cations in local weather as a result of climate warming are refl ected in the increasing occurrence of episodic extreme events, such temperature and precipitation extremes, storms and droughts (Easterling et al. 2000). For example, the magnitude of El Niño-Southern Oscillation (ENSO) has increased in the last decades, leading to more extreme El Niño events (Dai et al. 1997). The productivity of the pelagic habitat of the Tropical Pacifi c is tightly coupled to the El Niño-Southern Oscillation, and the fi sh community is deeply modifi ed during strong El Niño episodes (Lehodey et al. 2006). The intensifi cation of climate variability exposes marine organisms to an increased environmental uncertainty, which lead to unpredictable ecosystem responses.

The modifi cations of the Atlantic meridional overturning is expected to impact the evolution of the global climate change (Bryden et al. 2005). Further simulations predict that the slowing of the Atlantic Ocean circulation can cause the collapse of plankton as a result of the reduced upwelling of deep, nutrient-rich waters (Schmittner 2005), and the deleterious effect might spread up in the trophic web causing severe reduction of fi sh stocks. The


high sensitivity of pelagic communities to climate-driven changes will lead to variable responses at all levels of organization (Fig. 2), in which direction and magnitude, would be diffi cult to predict, thus the implementation of proper management and mitigation policies are urgently needed.

Fig. 2. Abiotic consequences of global climate change (left boxes) and the associated impact on the different levels of ecological organization of fi shes (right boxes).

How Fisheries Increment the Effects of Climate Change on Marine Ecosystems

It is expected that fi shes and ecosystems in general will be adapted to the changes in climate variability as they did in the past. However, there are new factors like overfi shing that will alter this natural ability, especially if the rate of climate change grows rapidly (Perry et al. 2010b). Overfi shing has become an ocean-wide problem, since it simplifi es the general characteristics of marine populations making them more sensitive to climate change (FAO 2004, Perry et al. 2010b). Some of the fi sh species that are the main target of fi sheries are in serious danger of being extinguished (Sadovy and Cheung 2003). Meanwhile a loss of marine biodiversity is expected as a consequence of the poor specifi city of fi shing gear that leads to the incidental capture of non-target species (Allan 2005, Worm et al. 2006, Daw et al. 2009).

The main issue concerning the mortality produced by fi shing and its selective nature, is that it leads to restricted adaptation potential to climate change (Brander 2010). Fishing generates an alteration on the demographic


structure, since it removes preferentially the larger (older) individuals. In fact, the length of fi sh species landed from coastal areas has been reduced by 25% (Trites et al. 2006). For example, in the North Atlantic the maximum length of the species landed from 1950 to 2000 decreased approximately from 95 cm to 65 cm (Planque et al. 2010). As a consequence, the buffering capacity of populations is constrained since a fi sh population with many year classes can survive during long periods of adverse environment conditions until the circumstances became favorable (Planque et al. 2010). On the other hand, a population with few cohorts may collapse before the onset of favorable conditions (Formentin and Fonteneau 2001).

One of the main repercussions of selecting the individuals at the edges of the species range is that they are likely to have particular adaptations to extreme conditions, therefore the genetic pool and the population plasticity is greatly reduced (Brander 2010). The loss of sub-populations, alleles and genotypes will reduce their surplus productivity potential and make them more vulnerable to fi shing and climate variability (Brander 2010). The alteration of life-history traits is another consequence of the fi shing-associated acceleration of population turnover rates (Law and Rowell 1993, Law 2000). As a consequence, the population replacement time is reduced because there is an increment in the growth rate and a decrease in the age-at-maturity (Perry et al. 2010b). It has been experimentally proved, that fi sh species which present more rapid turnover of generations have faster demographic responses to climate forcing and also have a faster tendency of spatial redistribution (Perry et al. 2005).

The reproductive potential of fi sh populations is also affected by the selectivity of fi sheries, since recruitment capacity is greater if larger/older individuals are more abundant (Planque et al. 2010). Reducing the average length and age of individuals may increase the recruitment variability by diminishing the capacity of the population to cope with short-term unfavorable environmental conditions (Barkley et al. 2004, Hutchings and Reynolds 2004, Hutchings and Baum 2005, Hsieh 2006). In addition, a population with fewer age classes could diffi cultly survive after several years of poor recruitment (Perry et al. 2010b). On the other hand, larger/older females present various adaptive advantages, such as longer spawning period and over a larger vertically and horizontally area, higher probability of egg survival and higher tolerance of their post-hatch larvae to long periods of starvation (Barkeley et al. 2004, Secor 2007, Planque et al. 2010).

Many fi sh species use bet-hedging strategies to increment the survival of larvae in variable environmental conditions (Lambert 1987, Hutchings and Myers 1993, Marteinsdottir and Steinarsson 1998). If the age structure is truncated by fi shing, fi sh population become more unstable because bet-hedging strategies are undetermined and are more infl uenced by short-


term environmental variability and forcing (Hutchings and Myers 1993, Marteinsdottir and Steinarsson 1998, Hsieh et al. 2006).

Through the social learning, the individuals use information provided by others to make decisions related to feeding, mating, habitat selection for spawning or migrating (Brown and Laland 2003, Planque et al. 2010), thus the loss of older individuals by fi shing could affect their behavioral and migration pattern (Corten 2002, Petitgas et al. 2006, Planque et al. 2010). What is more, they transmit their knowledge on migration and location of suitable habitat of spawning to the younger fi shes (Warner 1988, Corten 2002). If the older individuals are removed, it could derivate in a loss of tradition and it is possible that they would never recover their migration pattern (Planque et al. 2010). The spatial distribution of fi shes is also conditioned by the range and abundance of a population. These two characteristics are directly linked so if the abundance falls the range distribution of the population will contract (MacCall 1990, Fisher and Frank 2004). At low population abundance, the population has a tendency to inhabit only the most suitable spaces (hot spots), reducing its encountering rate with prey and competitive species, thus benefi ting the development of other populations less impacted by fi shing (Planque et al. 2010).

The ecological consequences of overfi shing leads to a reduction of the top-down predatory control, and bottom-up processes are expected to play a relatively more important role in the dynamics of marine communities. The predominance of bottom-up processes suggests that climate variability will have a greater impact on the structure of these communities, leading to population instability and shifts in the community structure (Perry et al. 2010b, Planque et al. 2010).

To conclude, it is worth to highlight that heavy fi shing not only reduces population’s levels, it also modifi es the age structure and the geographical distribution. As a result, the population is less resilient to unfavorable environmental conditions (Brander 2005, Brander 2010, Planque et al. 2010).

How Climate Change is Affecting and Will Affect World Fisheries

In the marine systems the environmental conditions play a major role in the distribution of fi sh populations. The climate change, which leads to an increase of sea water temperature and an alteration of ocean currents and coastal upwelling patterns, is producing a change in the distribution of the pelagic species. The situation is further impaired by overfi shing, which affects the resilience of fi sh communities and ultimately leads to ecosystem degradation. As a result, climate change will have a large impact on the


magnitude and distribution of maximum catch potential of marine fi sheries (Cheung et al. 2009a).

The variability of the fi sh stocks has major economic consequences for human societies. The increase of the distant water fl eets during the last century, has not only reduced the dependence of part of the fi shing industry on a particular area or species, but has resulted overfi shing of certain populations which consequently resulted in a higher variability of the stocks and the economical revenue (Brander 2010). However, the most vulnerable to the stock variability will be the millions of small-scale fi sher-folk (fi shers, fi sh processors, traders and ancillary workers) in the developing world, who depend upon a few local populations (Allison et al. 2009, Brander 2010). The vulnerability of natural populations has become a key concept in light of current and projected climate scenarios, and is defi ned as the susceptibility of species to the disrupted climate-driven (Dow et al. 2009).

Although most fi sheries are negatively affected by climate change, it is expected that the landings result benefi ted in certain latitudes. This information is crucial when planning strategies for adapting and to effi ciently mitigate the impacts (Brander 2010). For example, the warmer oceanic conditions are expected to have a positive ecological effect on Greenland, which has been affected by a cold period since 1960’s (IPCC 2007). On the contrary, some of the most vulnerable systems will be the big deltas of rivers in Asia, where millions of people depend on fi shing activities as well as on the hydrological cycle and land-sea interface use, which are highly vulnerable to climate-driven changes (Easterling et al. 2007, Brander 2010).

Cheung et al. (2009a) projected the trend of global catch potential from 2005 to 2055 under biophysical climate change scenarios. The highest impacts on catch potential will be in the Indo-Pacifi c region, where the catch potential will decrease by a 50% in the next 10 years. Fish landings will also decrease in coastal regions as well as in the tropics and the Antarctic zone. On the other hand, the catch potential in high latitude and offshore regions of the North Atlantic, the North Pacifi c, the Arctic and the northern edge of the southern Ocean will increase, as well as the continental margins of South Africa, southern Argentina and Australia. The melting of sea ice in the Arctic, will lead to a new habitat that could be colonized by species from lower latitude, increasing the maximum catch potential in the Exclusive Economic Zones of the Nordic Countries (Norway, Greenland and Iceland) by an 18–45%. The catches in Alaska and Russia are also expected to increase by around 20% (Cheung et al. 2009a).

The prediction of the economic and social impact that will have the climate change on fi sheries, depends not only on recognizing the places susceptible to climate warming and its ecosystem-level responses, but also


will depend on understanding the social and economic dynamics of fi shing fl eets and fi shing communities and their capacity to adapt to change (Allison et al. 2009). According to Brander (2010), it is very diffi cult to judge who the winners and losers will be, though there are obvious advantages for those countries that are well-informed, well-capitalized and able to shift to alternative areas as the circumstances change.

Allison et al. (2009) compared the vulnerability of 132 national economies to potential climate change impacts on their fi sheries. Although warming will be most pronounced at high latitudes, the most vulnerable countries will be the poorest, since their inhabitants are twice dependent upon fi sh and seafood as those of other nations. Thus, the most vulnerable regions lie in the tropics in central and western Africa (e.g., Malawi, Guinea, Senegal, Uganda), northwestern South-America (Peru, Colombia), and Asia (Bangladesh, Cambodia, Pakistan and Yemen), which presents different combinations of climate exposure, fi sheries dependence and adaptive capacity (Allison et al. 2009).

Climate change will favor some fractions of fi sh populations that may not be favored today and thereby change the biogeography of fi sh stocks and their abundance (Cheung et al. 2008, Cheung et al. 2009b), which highlights the importance of conserving biodiversity (Planque et al. 2010). For example in Peru, the increase of sea temperature has a negative infl uence in the pelagic artisanal fi sheries, but at the same time it attracts different tropical and subtropical species, generating new opportunities for the fi shermen and their community (Daw et al. 2009). An extreme would be the creation of a complete new fi shery in the open sea as a result of the infl ow of thaw waters of the Arctic Ocean. The adaptation capacity of the local economies, fi sheries, communities, individuals and governance systems will determine how viable the new fi sheries will be (Daw et al. 2009).

South-American Fisheries: Between the Climate Change, the Variability of Climate and the Fisheries Pressure

In the present section we are going to introduce three cases (as examples) of climatically-impacted fi sheries in South-America. The fi sheries from South America are extremely vulnerable to climate variability (Southern Oscillation events) since they present a combination of high economical dependence on primary production, a lack of appropriate public polices and ecosystem-based management, weak governance and little institutional support (e.g., Allison et al. 2009, Pitcher et al. 2009, Brander 2010).


Industrial fi sheries of anchoveta peruana (Peru and Chile)

The fi sheries from northwestern South America (Chile and Peru) are highly vulnerable to climate variability, since a climate-sensitive upwelling of nutrient-rich waters support huge catches of anchovy and sardines (Pauly and Tsukayama 1987, Bakum and Broad 2003). These fi sheries represent a signifi cant input of the countries’ economies. In Peru, the production of fi shmeal and fi sh processing represents the second highest income source after mining, and the annual exportations are valued at 1124 million of American dollars (FAO 2003, FAO 2004). Furthermore, Peru is the major producer of fi sh fl our and fi sh oil in the world (Daw et al. 2009).

The main issue concerning fi sheries of Peru and Chile, is the tight coupling of fi sh productivity and the Southern Oscillation, since El Niño warm events cause changes on the upwelling dynamics, turbulence pattern and water temperature, which ultimately leads to a decline in the concentration of nutrients and pelagic productivity (Jacobson et al. 2001, Chavez et al. 2003, Lehodey et al. 2006). The inter-annual variability produced by El Niño events shape the magnitude of the annual captures of anchovy which fl uctuated between 1.7 and 11.3 millions of tones during the last decade (Daw et al. 2009). During the 1997/1998 El Niño event, the fi sheries landings fell nearly 70% in the coastal area of Chile (Avaria et al. 2004), while in Peru, anchovy and sardine landings fell about 55% compared to the previous year, causing a loss in the revenues of more than $26 million (CAF 2000). On the contrary, La Niña events, which are associated with cooling sea surface temperature around Peru and Chile, benefi t these fi sheries (Ordinola 2002, Badjeck et al. 2010).

The climate variability can favor in this system some species over the others, as it happened in Peru, where the 1997/1998 El Niño event benefi ted the growth of the scallop (Argopecten purpuratus). In response to this, the fi sherfolks rapidly changed their fi shing methods to catch the scallop resulting in a record harvest (Badjeck et al. 2009, Badjeck et al. 2010). Drastic changes occur in peruvian and chilenian fi sheries when El Niño occurs, since they change from being monospecifi c (anchovy) to become multispecifi c fi sheries based on sardine, jack mackerel, pacifi c mackerel, longnose anchovy and scallop between others (Ñiquen and Bouchon 2004, Badjeck et al. 2009). The variation in the stock size of anchovy and the changes in the species landings is a central problem for determining an optimal organization of the fi shing industry and how the governments can best regulate the fi shing effort (Pontecorvo 2001).


Growing evidences predict that with the climate change and the augmentation of greenhouse gases in the atmosphere, the Pacifi c trade wind system will tend to decline its intensity (Vecchi et al. 2006, Vecchi and Soden 2007). It was suggested by Bakun and Weeks (2008) that there is a tendency of less intense El Niño events and also they are becoming more frequent (Moller et al. 2009). On contrary there are other investigations that proposed that climate warming will lead to stronger and more frequent Southern Oscillation events (i.e., Dai et al. 1997, Timmerman et al. 1999, Hansen et al. 2006). In this context, the future of the fi sheries that are supported by the Southeast Pacifi c upwelling system is highly uncertain.

The artisanal Shellfi sheries in South America (Chile, Uruguay and Peru)

In Latin America and the Caribbean, artisanal shellfi sheries constitute social-ecological systems critically important for thousands of coastal communities, integrated by vulnerable people who depend upon this activity (Defeo and Castilla 2012). This economic activity takes place directly from the shore or deckles boats generating incomes for fi shers and in some cases, they create export earnings for their countries (Mahon et al. 2003, Gelcich et al. 2010). The climate change represents a critical stressor for shellfi sh at multitemporal and spatial scales (Defeo et al. 2009, McCay et al. 2011, Perry et al. 2011, Defeo et al. 2012).

The interannual (e.g., Sothern Oscillation) and multidecadal (e.g., Atlantic Multidecadal Oscillation) environmental variability associated with fi shing and weak governance under open access policies, intensifi es climate-induced changes on shellfi sh. For example in Peru and Chile, El Niño events caused the mass mortality of the surf clam Mesodesma donacium, which was exacerbated by unsustainable harvest levels and weak governance (Defeo and Castilla 2012). As a consequence, in Peru the fi shery was closed in 1999 and the stock has not been recovered yet, probably because the ecosystem has exceeded critical thresholds (Scheffer et al. 2009). Clam mass mortality has became a current issue in South America (e.g., Argentina, Uruguay), mainly due to climate variability. The events are usually attributed to increase in temperature, harmful algal blooms and parasitism and generated variations in community composition, simplifi cation of food webs, introduction of non native species and distributional shifts (Fiori et al. 2004, Defeo et al. 2009, Beck et al. 2011, Riascos et al. 2011).

As mentioned previously, the interactions between climate change and fi shing affect the ecosystems and governance structures at different temporal and spatial scales. In this sense, sustaining shellfi sh resources will require the implementation of resilient management and governance policies in a changing world full of uncertainty (Defeo and Castilla 2012).


Lagoa dos Patos: an example of the small folk fi sheries (Brazil)

It is predicted that from 2040 the Patos Laggon estuary (Brazil) will be smaller, less salty and warmer, with stronger outfl ow and weaker infl ow currents at the mouth of the estuary, which is expected to bring positive and negative impacts on fi sh stocks (Schroeder and Castello 2010). Furthermore, the fi sheries landings in this area are markedly infl uenced by strong ENSO events, which directly affect the rainfalls of the area and in consequence, the availability of resources for artisanal fi sheries (Kalikoski et al. 2010). The pink shrimp is the economically most important population of the estuary, and its productivity pattern is highly unpredictable due to its reliance on climate variability (ENSO) (Castello and Moller 1978, Moller et al. 2009). There is an important fi shermen community of approximately 3500 people, who depend upon the resource (Haimovici et al. 2006, Schroeder and Castello 2010), and most of them do not have a livelihood strategy in case the pink shrimp harvest fails (Kalikioski et al. 2010).

Kalikoski et al. (2010) describe how the lack of institutional support for small-scale fi shing communities combined with an erosion of traditional resource systems and declining fi sh stocks led to increasing vulnerability of fi shing communities to the climate change. The fi shing resources are under intense pressure, from both the industrial and artisanal fi sheries, increasing the risk of collapse. They conclude that the fi shing communities that diversifi es (i.e., varying the species caught, working on industrial fi shing vessels, having alternative sources of income) and have a higher degree of self-organization are less vulnerable to adverse conditions. In this study, the authors pointed out that the artisanal fi shing sector is marginalized and has little institutional support. In addition, even governmental programs support exists to assist fi sheries which do not consider the impact of the climate change (Kalikoski et al. 2010).

Summary and Future Perspectives

Climate change has happened before, but nowadays the velocity of change has no historical precedents (e.g., Duvly et al. 2008, Badjeck et al. 2010, Brander 2010). The increase of water temperature has direct implications on the ecological processes at all levels of organization, since it accelerates the metabolism of individuals which determines the rate of resource uptake in aquatic habitats (Brown et al. 2003). Temperature has also deep infl uence on the reproductive traits of most fi shes (i.e., Hutshings and Myers 1993, Jonsson and Jonsson 2004, Crozier et al. 2008) and on the body sizes, which tends to decrease with the warming of the environment (i.e., James 1970, Atkinson 1994, Daufresne et al. 2009). As a consequence of climate change, ecological regime shifts, community reorganization (Stenseth et al. 2002)


and geographical redistribution occurs (i.e., Parmesan and Yohe 2003, Perry et al. 2005, Duvly et al. 2008, Last et al. 2011).

In addition to the direct consequences of the natural environmental changes, the human pressure by overfi shing, habitat destruction and pollution (among others) is compromising the resilience of the marine fi sh species (Brander 2010). The overfi shing reduces population’s levels, modifi es the age structure and the geographical distribution, making the population less resilient to unfavorable environmental conditions (Brander 2005, Brander 2010, Planque et al. 2010). It simplifi es the general characteristics of marine populations making them more sensitive to climate change. The high variability of declining populations resulting from overfi shing and the growing climate uncertainty, represents a double-edge issue impending natural communities in marine ecosystems worldwide (Hsieh et al. 2006).

Combined with environmental changes and human pressure, climate change will have a large impact on the distribution and magnitude of maximum catch potential of marine fi sheries (Cheung et al. 2009). The variability of fi sh stocks has major economic consequences for human societies and has been increasing over time, though the most vulnerable will be the small-scale fi sher-folk in the developing world that depend upon few local populations (Halley and Stergiou 2005, Allison et al. 2009, Brander et al. 2010). South America is considered to be highly vulnerable to climate variability and change for presenting a combination of high economical dependency on primary production, lack of appropriate public polices and ecosystem based management, weak governance and little institutional support (i.e., Pitcher et al. 2009, Allison et al. 2010, Brander et al. 2010).

The uncertainty is inherent of fi sheries management (Miller and Fluharty 1992) mainly because there is lack of information about the consequences of climate change, especially in the southern-hemisphere (McIlgorm et al. 2010). Adaptation has become one of the major issues in planning, and is based on revising the core functions of fi shing governance, that can be accomplished by collecting data on stocks and environmental variables, adjusting fi shing effort, reducing catch and emphasizing management tools based on fl exibility (Badjeck et al. 2010). Up to now, the fi sheries managers have tried to distinguish between the impacts due to climate change and those produced by overfi shing with the objective to predict the fi rst one and to control the second one. Nowadays, the challenge of the fi sheries management is to understand the sources of variability in landings and their interactions with external forcing (Perry et al. 2010b).

It is essential to identify ecological mechanisms capable of stabilizing marine communities and strengthen their resilience in order to effi ciently design management policies. It will be necessary to adopt a social-ecological approach on the fi sheries management, which will require recognizing that


the reduction in fi shing will relieve the pressure in marine systems, but on the other hand, will increase the pressure in the human system, since nowadays fi sh is an important economical source in the developing world (Perry et al. 2010a).

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CHAPTER 9

Marine Mammals in a Changing World

Cappozzo, Humberto Luis,1,2,* Panebianco María Victoria1,3,a and Juan Ignacio Túnez4

Introduction

Planet Earth is the most unusual one among the known planets since 71% of its surface is covered by oceans full of life. Life originated and diversifi ed in the oceans along millions of years. At present, the oceans are largely responsible for sustaining planet life and in regulating the climate (IUCN 2009, Hoegh-Guldberg and Bruno 2010).

Climate change is thoroughly accepted as a global concern (IPCC 2007). Rising atmospheric gas concentrations have increased global average temperatures by approximately 0.2°C per decade over the past 30 years

1 Laboratorio de Ecología, Comportamiento y Mamíferos Marinos (LEC y MM), Museo, Argentino de Ciencias Naturales Bernardino Rivadavia (MACN–CONICET); Av. Ángel, Gallardo 470, C1405DJR, Buenos Aires, Argentina.

a Email: [email protected] Fundación de Historia Natural Félix de Azara, Departamento de Ciencias Naturales y,

Antropología - CEBBAD - Instituto Superior de Investigaciones, Universidad Maimónides, Hidalgo 775, C1405BDB, Buenos Aires, Argentina.

3 Laboratorio de Química Marina, Área de Oceanografía química, Instituto Argentino de Oceanografía (IADO-CONICET),Florida 8000 (Camino La Carrindanga km 7,5), B8000FWB, Bahía Blanca, Argentina.

Email: [email protected] Grupo de Estudios en Ecología de Mamíferos, Departamento de Ciencias Básicas, Universidad

Nacional de Luján and CONICET, Rutas 5 y 7 (6700), Luján, Provincia de, Buenos Aires, Argentina.



(Hansen et al. 2006) with most of this added energy being absorbed by the world’s oceans (Hoegh-Gouldberg and Bruno 2010). And in addition, to acting as the planet’s heat sink, the ocean had absorbed approximately one-third of the carbon dioxide produced by human activities. Increases in the heat content of the ocean have driven other changes. Thermal expansion of the oceans as well as increased meltwater and discharged ice from terrestrial glaciers and ice sheets has increased ocean volume and hence sea level (Rahmstorf et al. 2007). Warmer oceans also drive more intense storm systems (Knutson et al. 2010) and other changes to the hydrological cycle (Trenberth et al. 2007). The warming of the polar oceans also has important ramifi cations for the stability of continental ice sheets, such as those in Greenland and Western Antarctica, which are sensitive to small increases in temperature (Naish et al. 2009). Marine air and sea temperatures have risen over the northeast Atlantic in the last 25 years. The largest increase in sea surface temperatures occur in the southern North Sea and eastern English Channel, at a rate of 0.5 and 0.8°C per decade (Rayner et al. 2003, IPCC 2007). Changes in the West Antarctic Peninsula are profound: mid-winter surface atmospheric temperatures have increased by 6°C (more than fi ve times the global average) in the past 50 years (Skvarca et al. 1999, Vaughan et al. 2003). Eighty seven percent of the Western Antarctic Peninsula glaciers are in retreat (Cook et al. 2005), the ice season has shortened by nearly 90 days, and perennial sea ice is no longer a feature in this environment (Martinson et al. 2008, Stammerjohn et al. 2008).Variation in temperature also has impacts on key biological processes. The distribution and abundance of phytoplankton communities throughout the world, as well as their phenology and productivity, are changing in response to warming, acidifying and stratifying oceans (Doney et al. 2009, Polovina et al. 2008). The annual primary production of the world’s oceans has decreased by at least 6% since the early 80’s, with nearly 70% of this decline occurring at higher latitudes (Gregg et al. 2003). These changes in the primary production of the oceans have overall profound implications for the marine biosphere and biochemistry of the Earth (Falcowski et al. 2000). The shift in phytoplankton biomass and size has direct consequences for grazer communities, especially Antarctic krill (Euphausia superba), whose spawning behavior depends on sea ice (Quetin and Ross 2001). Because krill form a critical trophic link between primary producers and upper-level consumers such as whales and other species this has profound consequences (Fraser and Hofmann 2003, Schofi eld et al. 2010).

Sea ice, coral reefs and kelp forests play a critical role in structuring the biodiversity of polar oceans (Hoegh-Gouldberg and Bruno 2010). Many arctic mammals face serious declines, with polar bears projected to lose 68% of their summer habitat by 2100. Antarctic species, such as penguins and

Marine Mammals in a Changing World 221

seals are declining and, in some cases, face an escalating risk of extinction under current projections for Antarctic warming (Barbraud et al. 2008).

Marine mammals comprise several orders that are very well adapted to life in the water, among these mammals being cetaceans including whales and dolphins, carnivores with polar bears, sea otters, sea lions, seals and walruses between this particular group and fi nally the sirenians including manatees and dugongs (Bastida and Rodriguez 2009). Fossils show that cetaceans arose from terrestrial ancestors more than 50 millon years ago (Fordyce 2002). Pinnipeds predecessor inhabit coastal regions approximately 23 to 29 million years ago, during Oligocene. Polar bears and sea otters are a group of modern species with the last emergence of aquatic carnivores, recently during the Pleistocene of the Quaternary Era (Bastida and Rodriguez 2009). Marine mammals and bats are probably the mammals whose anatomy and physiology have undergone the greatest modifi cations. Marine Mammals adapt very well to living in oceans and seas. Some of them (cetaceans) adapt better than others (carnivores and sirenians). Discussion of evolution depends on an agreed phylogenetic approach. On the other hand, evolutionary processes at the species level have likely a different mechanism as natural selection, sexual selection, co-evolution and others. The processes involved were different across different marine mammals. On geological timescales evolution and paleoecology are inextricably linked, for evolution occurs through natural selection and adaptation to the environment (Fordyce 2002).

The impact of the climate change induced changes in marine ecosystems is increased by anthropogenic reasons. Human activities are polluting, warming, and acidifying the oceans, melting sea ice, overharvesting fi sheries, and altering entire food webs (Davidson et al. 2012, Halpern et al. 2008, Hoegh-Gouldberg and Bruno 2010). Fisheries’ by-catch causes deaths of more than 650,000 marine mammals each year (Read 2008). Overfi shing has depleted food supplies by reducing fi sh populations by 50–90%, and industrial-scale krill harvesting will likely further deplete food resources (Myers and Worm 2003, Hilborn et al. 2003, Schiermeier 2010). In addition, polar oceans are warming at rates twice as fast as the global average (Hoegh-Gouldberg and Bruno 2010); this has already altered whale migrations, reduced prey populations, and caused declines in seals and polar bears (Ursus maritimus) whose lifestyles are dependent on sea ice (Moore 2008). The International Union for the Conservation of Nature (IUCN) Red List currently classifi es 25% (32 of 128 species) of marine mammals as threatened with extinction. Examination of the threats on the basis of the Red List shows that nearly half of all species are threatened by two or more human impacts, with pollution being the most pervasive, followed by fi shing,


invasive species, development, hunting, and climate change (Davidson et al. 2012). Marine mammals accumulate large concentrations of contaminants in organs and tissue, such as organochlorine contaminants (OCs), heavy metals and polycyclic aromatic hydrocarbons (PAHs) (Leonzio et al. 1992, Marcovecchio et al. 1994, Fossi and Marsili 1997, Lailson-Brito et al. 2008). Evidence links pollutants exposure to several biological effects in marine mammals, such as, immune and endocrine system dysfunction, increased risk of infection or diffi culties in reproduction (De Guise 1995, Jepson et al. 2005, Hall et al. 2006).

These changes in the marine environment in turn have direct and indirect potential impacts on marine mammals. Direct impacts include the mentioned effects of reduced sea ice and rising sea levels on seal haul-out sites, while indirect impacts include changes in prey availability affecting distribution, abundance and migration patterns, community structure, susceptibility to disease and contaminants (Learmonth et al. 2006).

Moreover, rapid climate change is likely to impose strong selection pressures on traits important for fi tness (Gienapp et al. 2008). One of the potential long term effects of climate change will be alterations of the genetic structure of wildlife populations (Lynch and Lande 1993). The knowledge of current genetic structure and the effect of past climate on patterns of geographical differentiation at the molecular level may reveal valuable information on the underlying evolutionary processes and past demographic events (Milá et al. 2000) and would light up on the future genetic consequences of present climate change (Waltari et al. 2007). Molecular markers, such as mitochondrial DNA genes and nuclear microsatellites, have been extensively used in marine mammals in the last decade for the study of these topics (Trujillo et al. 2004, Weber et al. 2004, Baker et al. 2005, Hoffman et al. 2006, 2009, Matthee et al. 2006, Fontaine et al. 2007, Túnez et al. 2007, 2010).

The earth’s climate is changing, the planet is warming, sea ice and glaciers are in retreat, sea level is rising, and pollutants are accumulating in the environment and within organisms (Moore 2008). The severity of damaging human-induced climate change that takes place due to increase in carbon dioxide concentration is largely irreversible (Solomon et al. 2009). All these changes affect marine ecosystem and the different marine species suffer the consequences. All those species that depend upon the ice, such as polar bear (Ursus maritimus) and the ringed seal (Phoca hispida) are good examples of these changes and they are very sensitive at population level to climate change. Carnivores are the clearest examples, and the response of cetaceans to climate change are more diffi cult to evaluate because they are more extended in their home range and are totally aquatics, besides bears and seals which depend upon ice or land to complete their life cycles. The use of marine mammals as ecosystem sentinels is a function of their


ecological diversity and the variability that present the marine ecosystem (Moore 2008). Moore (2008) mentioned the use of behavioral ecology of gray whales (Eschricthius robustus) in the eastern North pacifi c and Arctic to illustrate the importance of ecological scale and considered the use of marine mammals as sentinels of hotspots in ocean production, changes to food webs, contaminant levels and disease pathways. We agree with Moore (2008) that marine mammal’s research must be included in local to large scale ocean studies.

This chapter will address the challenges for marine mammals in a changing world. The objective is to walk the natural history of marine mammals and understand how they were affected by climate change-related reasons, considering anthropogenic causes.

Climate Change and Population Genetic Structure in Marine Mammals

Climate change and potential impacts on marine mammals

Climate change is considered as one of the main threats to biodiversity at a global scale. The direct impacts of climate change have been documented on every continent, in every ocean, and in most taxonomic groups (Parmesan 2006). Predicted impacts of climate change on the marine environment include, among others, an increase in water temperature, a rise in sea levels, and a decrease in sea-ice cover (IPCC 2001). These changes in the marine environment can in turn have direct or indirect potential impacts on marine mammals. Direct impacts include the effects of reduced sea ice and rising sea levels on seal haul-out sites, while indirect impacts include changes in prey availability affecting distribution, abundance and migration patterns, community structure, susceptibility to disease and contaminants (Learmonth et al. 2006). Moreover, rapid climate change is likely to impose strong selection pressures on traits important for fi tness (Gienapp et al. 2008). One of the potential long term effects of climate change will be alterations of the genetic structure of wildlife populations (Lynch and Lande 1993). The knowledge of current genetic structure and the effect of past climate on patterns of geographical differentiation at the molecular level may reveal valuable information on the underlying evolutionary processes and past demographic events (Milá et al. 2000) and would light up on the future genetic consequences of present climate change (Waltari et al. 2007). Molecular markers, such as mitochondrial DNA genes and nuclear microsatellites, have been extensively used in marine mammals in the last decade for the study of these topics (Trujillo et al. 2004, Weber et al. 2004, Baker et al. 2005, Hoffman et al. 2006, 2009, Matthee et al. 2006, Fontaine et al. 2007, Túnez et al. 2007, 2010). In this section, we will review some papers


published in recent years that study the genetic structure and diversity in different species of marine mammals and discuss its fi ndings in relation to the potential direct or indirect effects of climate change on the conservation of genetic resources. Finally, we will review recent genetic studies in the Steller sea lion, Eumetopias jubatus, possibly the marine mammal species for which more genetic data are available, which highlight the importance that effective population size has on how species respond to climate change.

Direct impacts—reduced sea ice cover and rising sea levels

Ringed seals, Phoca hispida, have a circumpolar distribution from approximately 35°N to the North Pole (King 1983). The species depend upon snow and sea ice to breed. In recent years, ringed seals face a rapid alteration of their habitat as Arctic sea ice and snow, for which the species depend to breed, responds to climate warming. Early snow melts have been associated with increased mortality from premature exposure to severe weather and predators (Ferguson et al. 2005). Moreover, ringed seals population level effects of habitat change will depend upon its population structure. The potential for local depletions or extinctions is increased if ringed seals are distributed in many demographically independent subpopulations. Thus, knowing how the populations are structured is important to understand the likely impacts of the rapidly changing sea ice cover. Kelly et al. (2009) investigated the population structure of the ringed seal in 11 locations along the northern coast of Alaska, western Canada, the Bay of Bothnia in the Baltic Sea, and the Lake Saimaa, in Finland. They sampled 354 individuals that were analyzed with 9 microsatellite loci and a 359 bp segment of the mitochondrial Cytochrome Oxidase I gene. Their nuclear and mtDNA data were consistent in suggesting little divergence among sample sites in the Arctic Ocean. However, they found evidence of genetic differentiation or isolation between the Arctic Ocean and Baltic Sea seals. The authors conclude that the rapidly changing snow and ice cover of the Arctic will inevitably force changes in habitat use and in the patterns of gene fl ow and that the degree to which genetic diversity will be lost will depend upon current population structure and the rate of environmental changes.

Another pinniped species that would suffer the direct impact of global warming is the South American sea lion, Otaria fl avescens. The species is distributed along the Atlantic and Pacifi c coasts of South America (King 1983). In its Atlantic distribution, breeding colonies are located in three different areas: the coast of Uruguay, north-central Patagonia and southern Tierra del Fuego (Túnez et al. 2008a). In north-central Patagonia, the segment of coast with the highest number of sea lions in Argentina, distribution of colonies is associated with islands availability, and at a local scale, breeding colonies are positively associated with slight slope coasts (Túnez et al. 2008a).


In the last 5 years, three genetic studies analyzed the genetic structure and historical population dynamics of the species in the Atlantic coast. Túnez et al. (2007) and Feijoo et al. (2011), analyzing respectively cytochrome b and D-loop sequences, found that Uruguayan and Patagonian colonies are genetically different and should be considered as different conservation units. However, this population structure was not found when microsatellite markers were used (Feijoo et al. 2011). The authors interpret this result as a consequence of female phylopatry and high male dispersion. Three years later of the fi rst work, Túnez et al. (2010), using mitochondrial D-loop sequences, found that the historical population dynamics of O. fl avescens in north-central Patagonia appears to be closely related with the dynamics of the Late Pleistocene glaciations. The genetic analyses suggested that sea lions population in Patagonia suffered a demographic contraction during the glacial period, followed by a population expansion when the glaciers retracted. Northern Patagonian colonies, as Punta Norte or Puerto Pirámide, which did not show clear evidence of any expansion at that moment, could have acted as refugia during glaciations, and individuals coming from these colonies would have recolonized southern colonies during the posterior interglacial period. As was stated, predicted impacts of climate change on the marine environment include a rise in sea levels (Learmonth et al. 2006). In this context, it is possible that future rises in sea levels would act as glaciations in the past, making some regions of the coasts of southern Argentina and Tierra del Fuego uninhabitable by coastal breeding animals. Beaches with slight slope coasts and islands currently occupied by sea lion colonies genetically distinct from other colonies would be covered by the sea or drastically altered by changing sea levels, resulting in the loss of an important genetic resource and increasing the degree of isolation between remaining colonies. Another example of possible consequences of rising sea levels is provided by the Mediterranean monk seal, Monachus monachus, a critically endangered species of which only two genetically different populations remain (Pastor et al. 2007). In this case, rising sea levels could also eliminate already scarce haul-out sites of the species, especially by the fl ooding of caves that provide the only refuges for some groups (Harwood 2001).

Indirect impacts—changes in prey availability

Recent climate change has triggered profound reorganization in northern latitude ecosystems, with substantial impact on the distribution of marine assemblages from plankton to fi shes (Richardson and Schoeman 2004, Perry et al. 2005). These shifts in marine habitat and community structure are expected to drive major changes in the distribution, density and dispersal of predators such as marine mammals (Learmonth et al. 2006, Marx and


Uhen 2010). One example is the harbour porpoise, Phocoena phocoena, which is currently distributed throughout cold to temperate coastal waters of the North Pacifi c, the North Atlantic and the Black Sea, being absent in the Mediterranean (Read 1999). The south-eastern part of the distribution range in the North Atlantic is of particular interest in regard to recent climate change, as it covers the biogeographic transition between arctic/boreal species and subtropical species (Southward et al. 1995). Fontaine et al. (2010) used data previously obtained from 10 microsatellite loci (Fontaine et al. 2007) to test whether the genetic divergence observed among harbour porpoises populations in the Northeast Atlantic is compatible with the recent changes reported in the seascape and marine assemblages, or is more likely owing to ancient vicariance processes. Their results showed that the isolation of porpoises in Iberian waters from those further north occurred approximately 300 years ago, matching with a warming trend that started after the Little Ice Age period and with the ongoing retreat of cold-water fi shes from the Bay of Biscay. Results of Fontaine et al. (2010) work show how past climate change can produce population fragmentation. In this context, it is expected that future increases in sea water temperature, a predicted impact for the ongoing climate change, will force this marine predators to adapt to the changing spatial distribution of their prey, as they did in the past. The survival of the current genetic units will depend upon the ability of this top predator to adapt to rapidly changing habitat conditions.

The dusky dolphin, Lagenorhynchus obscurus, is another species for which prey availability played an important role along its evolutionary history. The species is widely distributed in the Southern Hemisphere, with a disjoint distribution restricted to temperate and sub-Antarctic waters (Würsig et al. 1997). Dusky Dolphins eat a wide diversity of prey, including Southern Anchovy, and mid-water and benthic prey such as squid and laternfi shes, species tightly associated with temperate ocean sea surface temperatures. The species is thought to be particularly vulnerable to broad-scale changes in the marine environment because of their high positions in marine food chains (Bannister et al. 1996) and has been suggested that its timing of foraging may be seasonal and be correlated with the movements of their prey (Bannister et al. 1996, Würsig et al. 1997). Previous studies on the genetics and distribution of dolphin genera Cephalorhynchus and Lagenorynchus (Pichler et al. 2001, Cassens et al. 2003) have suggested that the dusky dolphin dispersed in the Southern Hemisphere eastward from Peru via a linear, temperate dispersal corridor provided by the circumpolar west-wind drift. Harlin-Cognato et al. (2007) using a multi-locus approach analysis, that included mitochondrial and nuclear sequence data, proposed the alternative hypothesis that the phylogeographic history of the dusky


dolphin was structured by paleoceanographic conditions that repeatedly altered the distribution of its temperate prey species during the Plio-Pleistocene. They found that dispersal of the dusky dolphin into the Atlantic is correlated with the history of anchovy populations, including multiple migrations from New Zealand to South Africa. The authors suggest that moderate, short-term changes in sea surface temperatures and current systems have a powerful effect on anchovy populations; thus, is feasible that repeated fl uctuations in anchovy populations continue to play an important role in the phylogeographic history of coastal dolphin populations.

Populations of the Harbour seal, Phoca vitulina, have recently experienced periods of marked decline (Mathews and Pendleton 2006, Lonergan et al. 2007) that have been attributed to climate change induced regime shifts and other causes (Andersen et al. 2011). Using 15 polymorphic microsatellites and a segment of the mitochondrial D-loop region, Andersen et al. (2011) determined the genetic diversity, the population genetic structure and the degree of gene fl ow in four populations of the species on the northern edge of their distribution, Greenland, Iceland, northern Norway, and Svalbard. The latter is the world’s northernmost population of the species and the only protected population of the Nordic countries. Direct counts in the area suggest that its population size is approximately 1000 individuals (Lydersen and Kovacs 2010). Additionally, the authors investigate whether the observed population declines can be detected as genetic bottlenecks, and determine the genetic history of the Svalbard population. Results of the study showed that each of the four locations was a genetically distinct population and that the Svalbard population was highly genetically distinct, had reduced genetic diversity, received limited gene fl ow, had a rather low effective population size, and showed an indication of having experienced a bottleneck resulting from a recent population decline. Moreover, the authors suggest that the signifi cant heterozygote excess observed in the Svalbard sample might be attributed to the low effective population size, which could initiate future population inbreeding effects. If this is true, the observed decline in Svalbard could be accelerated due to reduced resilience to future climate change, which makes it necessary to immediately implement the monitoring of this vulnerable population.

On the southern hemisphere, one of the examples of climatic changes affecting food availability and genetic diversity came from the study carried out by Matthee et al. (2006) in the Cape fur seal, Arctocephalus pusillus pusillus. This pinniped species is currently distributed in 25 breeding and 9 non-breeding colonies along the southern African coastline (Butterworth et al. 1995). Analyzing a 361 bp segment of the mtDNA control region, Matthee et al. (2006) examined the partitioning of genetic variability and the


historical population dynamics of the species across six breeding colonies that represent a range of geographic localities. The observed mismatch distribution obtained by the authors showed evidences of a population expansion that would date back to at least 18,000 years before present and could possibly be as old as 38,000 years. This date coincides with the height of the last glacial maximum in the area. The authors suggested that the observed population expansion could be attributable to the high ocean productivity that characterizes maxima glacial periods that made resources abundant in the South Atlantic (Martin 1990). On the other hand, the nutrient content of oceanic waters is thought to decrease substantially during warmer interglacial periods, particularly in the southern oceans between 22°S and 41°S (Oppo and Horowitz 2000). Thus, current and future global warming would affect the abundance of Cape fur seals by means of a diminution of the ocean productivity and prey abundance.

Another case that deserves attention is that of the South American fur seal, Arctocephalus australis. As the South American sea lions, A. australis is distributed along the Atlantic and Pacifi c coasts of South America (King 1983). Breeding colonies are not homogeneously distributed along the species distribution. They are only found in three areas of the northern and southern extremities of the species range: the Uruguayan islands; the southern coast of Chile and Isla de los Estados; and the northern coast of Chile and Peru (Túnez et al. 2008b). Morphological and genetic studies support the idea that Uruguayan and Peruvian colonies can be considered as different Evolutionary Signifi cant Units (ESUs) (Túnez et al. 2007, Oliveira et al. 2008). South American fur seals have been intensively exploited by humans (Ximenez and Langguth 2002). However, the most important threat to the conservation of the species in Peru is the mortality during El Niño Southern Oscillation (ENSO) events (Oliveira 2011). During the severe ENSO in 1997–1998, the Peruvian population of fur seal declined a 72% (Oliveira 2011), as a result of low food availability as a consequence of warming of sea-surface temperatures and primary productivity reduction. Considering both the intense harvesting and the severe population reduction caused by the 1997–1998 ENSO, Oliveira et al. (2009) carried out the fi rst bottleneck test for the Pacifi c and Atlantic populations of A. australis based on the analysis of seven microsatellite loci. Results of the test indicated that the Peruvian population of the species may have experienced a genetic bottleneck, while there is no suggestion of the Atlantic population having experienced a similar phenomenon. The authors conclude that the detection of the suggestion of a genetic bottleneck in the Peruvian fur seal population, combined with its small effective population size (Oliveira et al. 2006) and global warming models that predict stronger and more frequent ENSO events in the future (NCDC–NOOA 2007), are enough reasons for concern regarding the survival for the species, and should be taken into account in


future management plans for the conservation of this species on the Pacifi c coast of South America.

Demographic responses to climate change depend on population size

The biology of the Steller sea lion, Eumetopias jubatus, has been the subject of intense scientifi c investigation that has resulted in signifi cant advances in the understanding of its evolutionary history and population structure (Baker et al. 2005, Bickham et al. 1996, 1998, Harlin-Cognato and Honeycuth 2006, Hoffman et al. 2006, 2009). Harlin-Cognato and Honeycuth (2006) examined the phylogeographic history of the species in relation to the presence of Plio-Pleistocene insular refugia. Their results suggest that the genetic structure of Steller sea lions is the result of Pleistocene glacial geology, which caused the elimination and subsequent reappearance of suitable rookery habitat during glacial and interglacial periods. Five years later, Phillips et al. (2011) analyzed a large genetic data set consisting of 3 mitochondrial regions for more than 1,000 individuals in order to better understand the historical processes that have culminated in the extant populations of E. jubatus. The results of their study describe the role of climate change in shaping the population structure of the species. Climatically associated historical processes apparently involved differential demographic responses to ice ages dependent on population size. Ice ages during times of small effective population size promoted restricted gene fl ow and fragmentation, and ice ages occurring during times of large population size promoted gene fl ow and dispersal. These results illustrate that effective population size has a profound effect on how species respond to climate change. According to the authors, the most important aspect of its study is the implication that the ongoing anthropogenically caused climate change has the potential to affect distributions and demography of contemporary species in ways similar to that documented in their study.

Pollutants in Marine Mammals

Wildlife toxicology and sentinel organisms

The fi eld of wildlife toxicology can be traced to the late nineteenth and early twentieth century. This discipline focuses on the study of the effects that xenobiotics could have on wildlife and man and led to many investigations on those pollutants—heavy metals and synthetic pollutants. The use of sentinel organisms or bioindicator has emerged as the most common method to determine the presence of pollutants in a given environment. Such sentinels are used to gain early warnings about current or potential


negative impacts on individual- and population-level animal health (Bossart 2011). In turn, such warnings permit better characterization and management of these impacts that ultimately affect human and animal health associated with the oceans. The effects of exposures to xenobiotics in the marine environment have focused on marine mammals for several reasons (Reddy et al. 2001): 1) they have relatively long life spans that permit the expression of chronic diseases including cancer, abnormalities in growth and development, and reproductive failure; 2) are long-term coastal residents; 3) feed at a high trophic level; 4) and have large blubber stores that can serve as depots for anthropogenic chemicals and toxins (Bossart 2011, Reif et al. 2008). Because all of these, marine mammals integrate and refl ect ecological variation across large spatial and long temporal scales, they are prime sentinels of marine ecosystem change (Moore 2008). As a high level predator of the marine food chain many marine mammals (such as odontocetes and pinnipeds) tend to bioaccumulate high concentrations of anthropogenic contaminants, such as organochlorine contaminants (OCs), heavy metals and polycyclic aromatic hydrocarbons (PAHs) (Leonzio et al. 1992, Marcovecchio et al. 1994, Fossi and Marsili 1997, Lailson-Brito et al. 2008). A large body of evidence links pollutants exposure to a range of deleterious biological effects (e.g., immune, endocrine system dysfunction, increased risk of infection, impaired reproduction) in marine mammals (De Guise 1995, Jepson et al. 2005, Hall et al. 2006). The increased interest in studying these effects is because they could lead to decline in population of these mammals.

For the study of global trend of pollutants, we have selected three marine mammal species: the bottlenose dolphin (Tursiops truncatus), the harbour porpoise (Phocoena phocoena), and the franciscana dolphin (Pontoporia blainvillei). These species were chosen because the number of surveys focused on them covered a comparatively wider geographical range. The harbor porpoises are found in cold temperate to sub-polar waters of the Northern Hemisphere and they are usually found in continental shelf waters (Hammond et al. 2008a). The franciscana dolphins inhabit shallow coastal waters of tropical and temperate regions of the western South Atlantic Ocean (Crespo 2002) and they are found only along the east coast of South America (Brazil, Uruguay, and Argentina) (Reeves et al. 2008). On the other hand, the Common bottlenose dolphins Tursiops truncatus are distributed worldwide through tropical and temperate inshore, coastal, shelf, and oceanic waters (Hammond et al. 2008b) and is, by far, the species for which the most comprehensive data exist pertaining to contaminants —heavy metals and COs (Figs. 1 and 2). The factor of declining trend (FD, Borrell and Aguilar 2007) was used to evaluate the global trend of Hg, Cu, Cd, Zn, PCBs and DDTs.


Fig. 1. Global temporal trends of concentrations of Hg (total), Cu and Zn (µg/g dry weight) in the liver of bottlenose dolphin, Tursipos truncatus. * Year of publication.

Fig. 2. Temporal trends in the concentrations of PCBs and DDT (µg/g lw) in the A. Northern and, B. Southern hemispheres. Concentrations were determined from blubber tissue of bottlenose dolphin, Tursipos truncatus. (F) female (M) male dolphins. ˠ Samples of dolphins from Gulf of Mexico and ͌ from Northwestern Mediterranean (Spain).* Year of publication.


Heavy metals

Heavy metals are considered very important and highly toxic pollutants in the various environmental departments. The metals that have received more attention are Hg, Cd, and Pb, because of their highly toxic properties and their effects on the environment and the living organisms. Thousand of tons of these non essential and some essential elements have been input to the environment from anthropogenic sources: Pb: 450, Zn: 320, Ni: 47, Cu: 56, Cd: 7.5 (thousand tn yr–1). Atmospheric and river inputs, dredging spoil, direct discharges, industrial dumping and sewage sludge are some of the important contributors to metal pollution, which lead to the release of metals to the marine environment.

In surface and ground water, sediment and air, bioavailability of pollutants is a complex function of many factors including total concentration and speciation (physical-chemical forms) of metals, mineralogy, pH, redox potential, temperature and total organic content (both particulate and dissolved fractions). Many of these factors vary seasonally and temporally, and most factors are interrelated. Consequently, changing one factor may affect several others. In addition, generally poorly understood biological factors seem to strongly infl uence bioaccumulation of metals and severely inhibit prediction of metal bioavailability (Luoma 1989). The factors affecting bioaccumulation and bioavailability of metals jointly with the climate change, that is one of the most signifi cant threats to global ecosystems and biodiversity (King 2004, Thomas et al. 2004, MEA 2005), are the phenomena to which we must pay more attention when studying trends in global distribution and concentration of pollutants. In this sense it is worth pointing out that the increase in global temperature between 1900 and 2000 was the largest for any century during the past 1000 years (Huntley et al. 2006, Osborn and Briffa 2006, IPCC 2007); even if greenhouse and other gaseous emissions do not increase further, this warming is expected to continue (Meehl et al. 2005). There is compelling evidence that animals and plants have been affected by recent global changes in temperature (Walther et al. 2002, Parmesan and Yohe 2003), potentially leading to extinction of species in a wide range of taxa (Thomas et al. 2004), so impacts are harder to predict.

The results of the FD of Hg (µg/g dw) in liver of Tursiops truncatus from Mediterranean Sea (including Adriatic and Ligurian Sea, Fig. 1; Table 1) are in agreement with the trend found by Kakuschke and Prange (2007) where a fall in the concentration of Hg for 1994 to 1999 was observed (FD=1.8; Table 1) and for 1999 and 2002 Hg concentration remained constant (FD = 0.1). This trend was followed by a sustained decrease in concentration up to 2009 (FD = 5.7) in contrast to the fi ndings of Kakuschke and Prange (2007). The general Hg trend for the Mediterranean Sea—for 1994 to 2009—

Marine Mammals in a Changing World 233Ta

ble

1. M

etad

ata

for

met

al p

ollu

tant

s, in

clud

ing

tota

l Mer

cury

(H

g), C

adm

ium

(C

d),

Cup

per

(Cu)

and

Zin

c (Z

n) (

µg/

g d

ry w

eigh

t) in

sel

ecte

d

ceta

cean

’s s

peci

es r

epor

ted

from

197

8 to

201

2.

Sp

ecie

sC

onse

rvat

ion

S

tatu

sL

ocat

ion

(C

oun

try,

Sta

te)

No.

Sam

ple

s (s

ex)

Year

H

g (l

iver

)C

d

(liv

er)

Cd

(K

idn

ey)

Cu

(l

iver

)Z

n (l

iver

)R

efer

ence

Pon

topo

ria

blai

nvill

eiV

ulne

rabl

eSA

O(A

rgen

tina

)2

1988

–198

915

± 6

13 ±

640

± 1

664

± 9

330

± 6

0M

arco

vecc

hio

et

al. 1

990a

SAO

(Arg

enti

na)

18–

30b

9b27

b76

b11

9bG

erpe

et a

l. 20

02a

SAO

(Arg

enti

na)

4020

01–2

010

–8

± 1

0–

27 ±

24

139

± 6

1Pa

nebi

anco

201

0,

Pane

bian

co e

t al.

2012

a–

nd-4

0–

nd-1

2657

–301

SAO

(Arg

enti

na)

3720

01–2

010

––

23 ±

30

––

Pane

bian

co e

t al.

2011

a–

–nd

-119

––

SAO

(Bra

zil)

n Cu-

Zn=

1719

87–1

996

–2

± 2

1.2

± 2

33 ±

18

145

± 5

6L

ails

on–B

rito

et

al. 2

002

n Cd=

14

n Hg=

152–

18nd

-7nd

-516

–84

44–2

44

SAO

(Bra

zil)

2319

97–1

999

4 ±

20.

4 ±

1–

45 ±

89

152

± 8

2K

unit

o et

al.

2004

3–9

0.1–

4–

18–8

811

1–15

6

SAO

(Bra

zil)

44–

–2

± 1

–31

± 5

0–

Dor

nele

s et

al.

2007

a–

0.2–

17–

4–22

9–

SAO

(Bra

zil)

1819

98–2

005

2.6

± 2

.30.

6 ±

1–

26 ±

41

–Se

ixas

et a

l. 20

07,

2008

, 200

9c0.

7–9.

7nd

-3–

3.13

–189

SAO

(Bra

zil)

1320

03–2

004

111.

5 ±

1.3

––

1.2–

520.

1–4.

1–

–

Tabl

e 1.

con

td...

.


Sp

ecie

sC

onse

rvat

ion

S

tatu

sL

ocat

ion

(C

oun

try,

Sta

te)

No.

Sam

ple

s (s

ex)

Year

H

g (l

iver

)C

d

(liv

er)

Cd

(K

idn

ey)

Cu

(l

iver

)Z

n (l

iver

)R

efer

ence

Tabl

e 1.

con

td.

SAO

(Bra

zil)

719

98–2

000

50.

4–

–15

0C

arva

lho

et a

l. 20

08a

1–11

0.2–

1–

–87

–186

Uru

guay

n Hg=

19

n Cd=

1520

067

± 9

1.9

± 1

.8–

–Fo

glia

200

8a

0.3–

29

0–6

––

Pho

coen

a ph

ocoe

naL

east

Con

cern

NA

O

(Can

ada)

10 (m

ales

)19

6929

± 3

5–

––

–G

aski

n et

al.

1979

a

16 (f

emal

es)

1969

–197

372

± 1

53–

––

–

NA

O

(Can

ada)

(fem

ales

)19

69–1

971

2–36

5–

––

–

(mal

es)

2–73

––

––

NA

O

(Can

ada)

10 (m

ales

)19

72–1

974

16 ±

15

––

––

Nor

th S

ea

(Eas

t Sco

tlan

d)

17 (m

ales

)19

7414

± 1

30.

6 ±

0.5

4 ±

329

± 1

2–

Falc

oner

et a

l. 19

83

1–42

nd-4

1–12

11–5

1–

6 (f

emal

es)

24 ±

24

1.0

± 1

.411

± 1

129

± 1

1–

1–64

nd-4

1–30

11–4

1–

NA

O

(Can

ada)

21 (m

ales

)19

75–1

976

66 ±

44

––

––

Gas

kin

et a

l. 19

83a

Bal

tic

Sea

319

77–

0.3

± 0

.42.

0 ±

2.4

33 ±

24

176

± 3

6H

arm

et a

l. 19

78a

n Cd

-K=

2–

0.1–

10.

1–4

16–6

013

6–20

0

Nor

th S

ea

23–

1–64

–1–

30–

–Fa

lcon

er e

t al.

1983

Iris

h se

a (B

riti

sh Is

land

)36

1988

–199

182

± 1

620.

8 ±

0.5

–78

± 9

925

8 ±

152

Law

et a

l. 19

91,

1992

a2–

760

nd-2

–11

–480

80–6

00


Nor

th S

ea

(Bel

gium

-Fr

ance

)

n L=49

nK

=48

n Hg-

L=

2719

94–2

001

23 ±

66

0.5

± 0

.63

± 3

39 ±

38

234

± 1

72D

as e

t al.

2004

1–56

nd-2

.5nd

-12

9–25

740

–684

Nor

th S

ea

(Ger

man

y)n L

=14

n K

=12

14 ±

18

0.7

± 1

.34

± 9

58 ±

49

219

± 1

81

1–12

nd-5

nd-3

320

–169

77–7

27

NA

O

(Ice

land

)11

16 ±

14

6 ±

11

19 ±

17

30 ±

11

135

± 2

6

1–44

0.4–

39–

13–5

311

4–18

7

NA

O

(Nor

way

)n L

=21

nK=

20

n Cu-

L=

2014

± 1

00.

5 ±

0.5

6.0

± 4

.531

± 3

198

± 2

0

1–32

nd-2

nd-1

612

–161

69–1

40

nd-2

22–1

1988

–661

Pho

coen

a ph

ocoe

naL

east

Con

cern

NA

O

(Can

ada)

10 (m

ales

)19

6929

± 3

5–

––

–G

aski

n et

al.

1979

a

16 (f

emal

es)

1969

–197

372

± 1

53–

––

–

NA

O

(Can

ada)

(fem

ales

)19

69–1

971

2–36

5–

––

–

(mal

es)

2–73

––

––

NA

O

(Can

ada)

10 (m

ales

)19

72–1

974

16 ±

15

––

––

Nor

th S

ea

(Eas

t Sco

tlan

d)

17 (m

ales

)19

7414

± 1

30.

6 ±

0.5

4 ±

329

± 1

2–

Falc

oner

et a

l. 19

83

1–42

nd-4

1–12

11–5

1–

6 (f

emal

es)

24 ±

24

1.0

± 1

.411

± 1

129

± 1

1–

1–64

nd-4

1–30

11–4

1–

Tabl

e 1.

con

td...

.


Sp

ecie

sC

onse

rvat

ion

S

tatu

sL

ocat

ion

(C

oun

try,

Sta

te)

No.

Sam

ple

s (s

ex)

Year

H

g (l

iver

)C

d

(liv

er)

Cd

(K

idn

ey)

Cu

(l

iver

)Z

n (l

iver

)R

efer

ence

Tabl

e 1.

con

td.

NA

O

(Can

ada)

21 (m

ales

)19

75–1

976

66 ±

44

––

––

Gas

kin

et a

l. 19

83a

Bal

tic

Sea

319

77–

0.3

± 0

.42.

0 ±

2.4

33 ±

24

176

± 3

6H

arm

et a

l. 19

78a

n Cd

-K=

2–

0.1–

10.

1–4

16–6

013

6–20

0

Nor

th S

ea

23–

1–64

–1–

30–

–Fa

lcon

er e

t al.

1983

Iris

h se

a (B

riti

sh Is

land

)36

1988

–199

182

± 1

620.

8 ±

0.5

–78

± 9

925

8 ±

152

Law

et a

l. 19

91,

1992

a2–

760

nd-2

–11

–480

80–6

00

Nor

th S

ea

(Bel

gium

-Fr

ance

)

n L=49

nK

=48

n Hg-

L=

2719

94–2

001

23 ±

66

0.5

± 0

.63

± 3

39 ±

38

234

± 1

72D

as e

t al.

2004

1–56

nd-2

.5nd

-12

9–25

740

–684

Nor

th S

ea

(Ger

man

y)n L

=14

n K

=12

14 ±

18

0.7

± 1

.34

± 9

58 ±

49

219

± 1

81

1–12

nd-5

nd-3

320

–169

77–7

27

NA

O

(Ice

land

)11

16 ±

14

6 ±

11

19 ±

17

30 ±

11

135

± 2

6

1–44

0.4–

39–

13–5

311

4–18

7

NA

O

(Nor

way

)n L

=21

nK=

20

n Cu-

L=

2014

± 1

00.

5 ±

0.5

6.0

± 4

.531

± 3

198

± 2

0

1–32

nd-2

nd-1

612

–161

69–1

40

nd-2

22–1

1988

–661

Nor

th a

nd B

alti

c Se

as22

2004

–200

6–

0.2

–37

152

Fahr

enho

ltz

et a

l. 20

09a

– n

d-2

–22

–119

88–6

61

Turs

iops

tr

ucat

usL

east

Con

cern

NA

O

(Fra

nce)

519

77–1

990

421

± 3

711

± 0

.55

± 5

15 ±

10

117

± 3

6H

olsb

eek

et a

l. 19

9822

–783

nd-1

1–11

4–30

65–1

55

SAO

(A

rgen

tina

)1

1988

–198

934

4 ±

29

3 ±

1–

311

± 1

578

5 ±

136

Mar

cove

cchi

o et

al

. 199

0a

Marine Mammals in a Changing World 237Ir

ish

Sea

219

8982

± 3

0.4

± 0

.2–

28 ±

714

8 ±

28

Law

et a

l. 19

91a

80–8

40.

3–0.

5–

23–3

312

8–16

8

Gul

f of M

exic

o9

(mal

es)

1990

182

2–

––

Kue

hl a

nd H

aebl

er

1995

a20

–351

0.4–

5–

––

5 (f

emal

es)

104

1.3

––

–

24–1

950.

4–3

––

–

Gul

f of M

exic

on K

=21

n L

=29

1990

–199

4–

0.2

1.3

––

Woo

d a

nd V

an

Vle

et 1

996

–nd

-2nd

-6–

–

Sout

h A

ustr

alia

1119

88–2

004

856

± 9

6516

± 2

5–

85 ±

109

161

± 4

7L

aver

y et

al.

2008

e

10–3

088

nd-8

0–

20–3

4010

5–27

2

Gul

f of M

exic

o (F

lori

da)

13

n Cd

-L=

1119

91–1

992

304

1.6

4.4

––

Mea

dor

et a

l. 19

99

18–1

312

nd-2

nd-5

–

–

Nor

th a

nd B

alti

c Se

a57

1991

–199

339

± 7

7–

––

–Si

eber

t et a

l. 19

99

1–44

9–

––

–

Sout

h C

hina

Sea

319

94–1

995

299

1.4

± 0

.812

± 6

10 ±

263

± 1

9Pa

rson

and

Cha

n 20

01nd

-299

1–2.

45–

168–

1250

–84

Med

iter

rane

an

Sea

119

9542

50 ±

11

9–

––

–Fr

odel

lo e

t al.

2000

NA

O

(Mia

mi)

6–

280

± 6

1–

––

–M

acke

y et

al.

1995

a

2–15

4–

––

–

Tabl

e 1.

con

td...

.


SPO

(A

ustr

alia

)2

1995

–199

665

± 8

88

± 1

060

53 ±

27

370

± 2

91L

aw e

t al.

2003

a

3–12

80.

3–15

–34

–72

164–

576

NA

O

(Sou

th C

arol

ina)

34–

71 ±

111

0.2

± 0

.3–

43 ±

59

228

± 2

04B

eck

et a

l. 19

97a

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86nd

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648

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4

Nor

th E

aste

rn

Atl

anti

c10

219

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003

60 ±

93

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± 8

44 ±

921

4 ±

156

Lah

aye

et a

l. 20

07 a

1–56

9–

nd-5

27–

669

53–9

93

NA

O

(Por

tuga

l)2

–13

2 ±

120

––

20 ±

615

2 ±

48

Car

valh

o et

al.

2002

23–2

41–

–16

–25

118–

186

Med

iter

rane

an

(Isr

ael)

1419

94–1

999

388

± 5

962

± 1

4 ±

736

± 2

217

6 ±

120

Rod

iti-

Ela

sar

et a

l. 20

03a

4–19

640.

5–4

0.2–

1717

–96

60–4

60

Sout

h A

dri

atic

Se

a13

1996

–199

715

73 ±

5.

28–

11 ±

16

33 ±

0.6

211

± 3

Stor

elli

and

M

arco

trig

iano

20

02

Lig

uria

n Se

a2

1999

–200

218

75 ±

26

333

5 ±

758

± 5

226

6 ±

32

Cap

elli

et a

l. 20

08

Ital

ian

Coa

st12

2000

–200

933

2 ±

450

––

––

Bel

lant

e et

al.

2012

10–1

404

––

––

NA

O N

orth

Atl

anti

c O

cean

; SA

O S

outh

Atl

anti

c O

cean

; Wet

wei

ght b

asis

con

cent

rati

on w

as c

onve

rted

to d

ry w

eigh

t bas

is c

once

ntra

tion

ass

umin

g th

at m

oist

ure

cont

ent w

as a 7

5% o

r c 6

9.7%

(Yan

g an

d M

iyaz

aki 2

003)

; b Rep

orte

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um c

once

ntra

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s fo

und

by

the

auth

ors;

d a

nim

als

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d

from

infe

ctio

us d

isea

se; e

FC

0.2

9 fo

r liv

er a

nd 0

.23

for

kid

ney

used

to c

onve

rt to

dry

wei

ght b

asis

for

the

auth

or. –

No

info

rmat

ion

avai

labl

e.

Sp

ecie

sC

onse

rvat

ion

S

tatu

sL

ocat

ion

(C

oun

try,

Sta

te)

No.

Sam

ple

s (s

ex)

Year

H

g (l

iver

)C

d

(liv

er)

Cd

(K

idn

ey)

Cu

(l

iver

)Z

n (l

iver

)R

efer

ence

Tabl

e 1.

con

td.


was a slight increase (FD= 0.7). Zn and Cu, also showed a steady increase in their concentrations until the late 90’s followed with a slight decrease by 2002 (FD = 0.1) in the same coastal regions. Cd concentration showed an important decrease in their levels (FD = 7) from 1994 to 1999 followed by an increase (FD = 0.1) towards 2002. As a general trend, once again, Cd showed increased levels in the Mediterranean Sea (FD = 0.5).

The North Sea ecosystem was highly loaded with both organic and metal pollutants introduced by various anthropogenic activities within the coastal zones in the past century. Until the middle of the 80’s the yearly input of metal pollution was around 340 tn of Cd, 75 tn of Hg and 11.000 tn of Pb (Rachor and Rühl 1990, Kakuschke and Prange 2007). Currently several studies have shown a diminishing trend in the input of pollutants into the ecosystem, this general tendency was especially confi rmed for metal pollutants (Kakuschke and Prange 2007). The Quality Status Report of the Trilateral Monitoring and Assessment Program (TMAP) concluded that major reductions in the input and the concentrations of metals in the Wadden Sea occurred mainly in the late 1980s until the early 1990s and continued moderately until 2002 (Kakuschke and Prange 2007). Our results are in agreement with those fi ndings, FD show an increase from the late 70’s and 80’s (FD Hg = 0.3, Cu = 0.4, Zn = 0.2). The Hg concentration trend continued with a steady decline in its levels in the North Sea (FD Hg1993–2001= 1.7, Table 1, Fig 1). Whereas Cu and Zn showed an increase in their levels for late 80’s until 2001 (FD1989–2001 Cu= 0.7, Zn = 0.6). Finally, Hg levels exhibited an increase (FD = 0.2) in North Atlantic Ocean (Gulf of Mexico, France, Portugal, Spain, Grand Britain coasts, Iceland and Norway) during late 60’s to 90’s followed by a sharp decrease during early 90’s to 2000 (FD = 26.3) and a smooth increasing trend from 2000 to 2003 (FD = 0.3).

Extensive studies of metal concentrations in cetaceans have been carried out in the Northern Hemisphere but relatively little is known about contaminant levels in Southern Hemisphere cetaceans. Hg levels showed an increase during 90’s decade (FD1988–1996= 0.2) in the Southern Atlantic Ocean (Argentina, Australia, Brazil and China coasts). This trend was followed by a sharp decrease up to 2000 (FD1996–2000= 13) and a mild increase until 2004 (FD2000–2004= 0.5). Cu, Cd and Zn showed similar trends, a decrease from 1988 to late 90’s (FD Cd= 3.6, Cu = 1.4, Zn = 2.2). The concentrations of Cd and Zn continued that trend by increasing their levels to the present (FD Cd = 0.2, Zn = 0.9). Meanwhile Cu levels tend to decrease (FD = 1.2).

The average levels of Hg (liver) of the Northern Hemisphere (North and Irish sea) were higher than those found in the Southern Hemisphere (South Atlantic Ocean: Brazil and Uruguay) (Student test: SAO vs North seal: t = 2.67, df = 35, P = 0.01; SAO vs Irish Sea: t = 2.3, df = 57, P = 0.02;


Table 1). Hg levels recorded during late 90’s in the Ligurian Sea (1875 ± 2633 µg/g, Capelli et al. 2008) and Adriatic Sea (1573 ± 5.3 µg/g, Storelli and Marcotrigiano 2002) were also higher than the levels found in South Atlantic Oceans (4 ± 2 µgg–1, Kunito et al. 2004, Student test: t = 4.62, df = 23, P < 0.01 and t = 1282.6626 df = 35, P < 0.01 respectively).

Our results are in agreement with previous review and papers (Aguilar et al. 2002, Kakuschke and Prange 2007) where a global increase trend of some heavy metals (Hg, Cu, Zd and Zn) was observed.

Persistent organic pollutants

With the advent of synthetic pesticides in the 1930s and 1940s, effects of DDT and other pesticides were investigated in free-ranging and captive wildlife (Rattner 2009). In the mid-1960s, Koeman and van Genderen (1966) reported, for the fi rst time, the presence of organochlorine compound residues in the tissues of a wild marine mammal. They found a few parts per million of dieldrin and various forms of DDTs in blubber and fat of three common seals (Phoca viulina) from the Wadden Sea. Soon after, Jensen and others (1969), in a toxicological survey of Swedish marine fauna, were the fi rst to detect polychlorinated biphenyls (PCBs) in the tissues of marine mammals, this time in common seals (Phoca vitulina), grey seals (Halichoerus grypus) and ringed seals (Pusa hispida). Since then, the number of references on the subject has increased exponentially, reaching a ceiling of 30–40 refereed articles per year in the 1990s (Aguilar and Borrell 1996).

As a global trend numerous works have revealed that marine mammals from the northern hemisphere, which inhabit the mid-latitudes of Europe and North America, show the greatest organochlorine loads whereas tropical and equatorial areas of the northern hemisphere and all over the southern hemisphere show low or very low loads. The polar regions of both hemispheres showed the lowest concentrations of DDTs and PCBs, although levels of HCHs, chlordanes and HCB were moderate to high in the cold waters of the North Pacifi c. During recent decades, concentrations have tended to decrease in the regions where pollution was initially high but they have increased in regions located far from the pollution source as a consequence of atmospheric transport and redistribution. It is expected that the Arctic and, to a lesser extent, the Antarctic, will become major sinks for organochlorines in future (Aguilar et al. 2002).

Organochlorine concentrations referred to in this paper were always calculated on the basis of the sample content of extractable lipids (lipid basis). This expression was preferred to that calculated on the basis of the fresh weight of the sample because it corrects, at least partially, for heterogeneity of nutritive condition and laboratory methods used for lipid extraction. These two sources of heterogeneity are likely to introduce additional variation in the determination of levels of organochlorine


compounds in tissues. When source papers expressed concentrations on a fresh weight basis, we calculated the lipid-based concentrations by dividing them by sample lipid content (Aguilar 1986). If this latter value was not available, the concentrations in fresh weight basis were divided by 0.7, a fi gure that was considered to be a representative means of lipid richness for blubber (Aguilar et al. 2002). The mean, maximum and minimum values were directly extracted from the source papers or calculated from the raw data when available.

During the 1970s and 1980s, research concentrated in Western Europe and Northern America, clearly refl ected the technical strength of researchers from these regions. Since the mid 1980s however, the scientifi c production on the subject in Asia, particularly Japan, and in South America has increased. Nevertheless, to date, research levels in Australia, the Indopacifi c region and Africa still remain minimal (Aguilar and Borrell 1996, Aguilar et al. 2002; Table 2).

Not surprisingly, a previous review of the geographic patterns of OCs pollution in marine mammals showed that the Mediterranean Sea and the waters of California were the marine environments with the highest Levels of PCBs and DDTs worldwide (Aguilar et al. 2002, Aguilar and Borrell 2004, Storelli and Marcotrigiano 2004; Table 2). The highest average level of both DDTs and PCBs (400 and 1204 µg/g lw) was recorded in fi ve males Tursiops truncatus of Mediterranean Sea in 1992 (Corsolini et al. 1995). Nevertheless, a decreasing trend was confi rmed by FD values. DDTs and PCBS for 1978 to 2002 decreased by factors of 23.7 and 6.1 for Tursiops truncatus of Mediterranean Sea (Borrell and Aguilar 2007). This negative trend was also recorded for T. truncatus from 1980 to 1990 in the Indian Ocean (FD PBCs = 12, DDTs = 2, Table 2). Similar trend was recorded for Phocoena phocoena; DDTs and PCBs decreased by factors of 3.8 and 1.4 (for 1989 to 2000 and 1989 to 2010 respectively) in North and Belgian Seas (Fig. 2; Table 2). While the same negative trend is repeated (FD1994–2004 PCBs = 1.4, DDTs = 2.2) for Pontoporia blainvillei from Rio Grande do Sul, in southern Brazil. The trend was positive particularly for PCBs (FD1991–2007 = 0.1) in Sao Paulo and Parana to the north of the country. The concentrations found between 2001 and 2007 were the highest recorded for the last millennium in the Southwestern Atlantic Ocean (13 µg/g lw; Table 2). Finally, PCBs showed an increase in their levels during 1980s in Gulf of México (FD1983–1987= 1.7) and after the early 90’s PCBs levels gradually declined until late 90’s (FD1991–1996= 2.7), from 1996 to 2001 the levels increased in a factor of 0.2. Johnson-Restrepo and col. (2005) reported the highest concentration of PCBs ever recorded for the Gulf of Mexico (240 µg/g lw; Table 2). DDTs also showed a decrease in a factor of 2.3 from 1987 to 1991 (Fig. 2; Table 2).

242 Marine Ecology in a Changing WorldTa

ble

2. M

etad

ata

for

PCB

s an

d D

DTs

in b

lubb

er (µ

g/g li

pid

bas

is w

eigh

t) o

f sel

ecte

d c

etac

ean’

s sp

ecie

s re

port

ed fr

om 1

980

to 2

012.

Sp

ecie

sC

onse

rvat

ion

Sta

tus

Loc

atio

n

(Cou

ntr

y, S

tate

)N

o. S

amp

les

(sex

)Ye

ar∑

PC

Bs

∑D

DTs

Ref

eren

ce

Pon

topo

ria

blai

nvill

eiV

ulne

rabl

eSA

O(A

rgen

tina

)74

1988

–199

22

± 1

1.6

± 1

.6B

orre

l et a

l. 19

95

SAO

(Arg

enti

na)

819

91–1

996

3.4

± 2

.0C

aste

llo e

t al.

2000

SAO

(B

razi

l)32

1991

–199

61.

3 ±

0.5

Cas

tello

et a

l. 20

00

SAO

(B

razi

l)10

–3

± 2

1.6

± 1

6L

ails

on-B

rito

et a

l. 20

111–

60–

6

SAO

(B

razi

l)5

1999

–200

04

2Yo

gui 2

002

1–6

0.8–

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SAO

(B

razi

l)8

1997

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34

± 3

2 ±

2Yo

gui e

t al.

2010

0.5–

10.0

0–7

SAO

(B

razi

l)26

(ad

ults

)19

94–2

004

4 ±

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0 ±

0.8

Leo

nel e

t al.

2010

1–11

0.2–

3.5

SAO

(B

razi

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(mal

es)

1999

34

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iwar

a et

al.

2004

10 (f

emal

es)

22.

5

SAO

(B

razi

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(mal

esa )

2001

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08

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11 (f

emal

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7

0.3–

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2.4

Marine Mammals in a Changing World 243P

hoco

ena

phoc

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Lea

st C

once

rnN

AO

(D

enm

ark)

29 (m

ales

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80–1

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116

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n 19

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SA)

6 (m

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986

2337

O’S

hea

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80

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6

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(C

alif

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1985

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alam

boki

dis

198

6

6–13

2

NA

O

(Far

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land

s)3

(mal

es)

1987

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813

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orre

ll 19

93

10–1

66–

8

3 (f

emal

es)

94

8–10

4–5

NA

O

(Nor

way

)27

1987

–199

122

16

Kle

ivan

e et

al.

1995

4–65

3–45

NA

O

(Bri

tish

C

olum

bia)

719

87–1

988

5 ±

17

3 ±

24

Jarm

an e

t al.

1996

NPO

(C

alif

orni

a)3

4.5

± 4

2.0

16 ±

25

NA

O

(Uni

ted

K

ingd

om)

50 (m

ales

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88–1

992

267

Law

199

4

0.3–

110

0–33

47 (f

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185

0–13

90–

34

Tabl

e 2.

con

td...

.


Bal

tic

Sea

(Pol

onia

)3

(fem

ales

)19

89–1

990

3413

Kan

nan

et a

l. 19

93a

26–4

711

–15

Nor

th S

ea

(Eur

opa)

1 (a

dul

ts)

1990

–199

882

Wei

js e

t al.

2010

3 (a

dul

ts)

1998

–200

825

15–3

5

Bac

k Se

a (T

urke

y)9

1993

21 ±

10

Tana

be e

t al.

1997

6–39

Bac

k Se

a11

1998

13W

eijs

et a

l. 20

09

9–25

NA

O

(Gre

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nd)

46 (m

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962

3B

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1–8

54 (f

emal

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22

0.5–

3.0

0–5

Bel

gian

Nor

th S

ea

Coa

st21

1997

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036

± 2

63

± 2

Cov

aci e

t al.

2002

2–40

50–

44Tu

rsio

ps

truc

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Lea

st C

once

rn

Ind

ian

Oce

an

(Sou

th A

fric

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(mal

es)

1980

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714

20C

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roft

et a

l. 19

89

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0.5–

86.0

52 (f

emal

es)

8.5

6

0–48

0–50

NA

O

(USA

)8

(mal

es)

1983

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598

32K

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1987

6–19

54–

64

7 (f

emal

es)

22.5

5

2–41

1–8

Sp

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onse

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(Cou

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tate

)N

o. S

amp

les

(sex

)Ye

ar∑

PC

Bs

∑D

DTs

Ref

eren

ce

Tabl

e 2.

con

td.

Marine Mammals in a Changing World 245N

AO

(U

SA)

3 (m

ales

)19

87–1

988

138

Kue

hl e

t al.

1991

47–1

95

5 (f

emal

es)

219

4–40

2–16

Med

iter

rane

an

Sea

5 (m

ales

)19

87–1

992

4515

Mar

sili

and

Foc

ard

i 19

970–

175

1–58

2 (F

emal

es)

304

29–3

143–

5

Ind

ian

Oce

an

(Ind

ia)

2 (m

ales

)19

901

9Ta

nabe

et a

l. 19

93

1–2

13–5

2 (f

emal

es)

115

0.6–

0.9

9–21

Gul

f of M

exic

o (U

SA)

9 (m

ales

)19

90–1

991

9338

Kue

hl a

nd H

aebl

er

1995

6 (f

emal

es)

74

Med

iter

rane

an

Sea

5 (m

ales

)19

9212

0440

0C

orso

lini

et a

l. 19

95

250–

2100

48–1

100

Sout

h C

hina

Sea

319

94–1

995

13 ±

479

± 6

8Pa

rson

and

Cha

n 20

018–

1614

–149

SPO

(A

utra

lia)

219

95–1

996

0.6

± 0

.3L

aw e

t al.

2003

Gul

f of M

exic

o (U

SA)

3115

Sala

ta e

t al.

1995

0.4–

75

Tabl

e 2.

con

td...

.


NA

O

n PCB

s=3

n DD

Ts=

5 (m

ales

)19

95–2

000

5352

Han

sen

et a

l. 20

0416

–52

11–8

7n PC

Bs=

2 n D

DTs

=6

(fem

ales

)12

433

–41

2–19

4 (m

ales

)19

99–2

001

5033

24–8

518

–54

5 (f

emal

es)

84

3–31

1–22

9 (m

ales

)20

0220

1315

–28

8–24

2 (f

emal

es)

94

5–17

2–10

Gul

f of M

exic

o (U

SA)

319

91–1

996

36Jo

hnso

n-R

estr

epo

et

al. 2

005

6–11

0G

ulf o

f Mex

ico

(USA

)11

2000

–200

124

010

–112

0N

AO

620

01–2

004

963–

303

Med

iter

rane

an

(Spa

in)

719

78–1

988

604

± 4

1026

1 ±

162

Bor

rell

and

Agu

ilar

2007

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1377

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The negative trend observed in the waters of Europe, USA and Brazil is consistent with the chemical history of use. DDT was first synthesized in 1874. However, its commercial use in agricultural and sanitary pest control started only after World War II, particularly after the 1950s. The amount of DDT produced increased exponentially since 1950 until the 1970s. Manufacture ceased in the western world in the 70’s, due to environmental concerns which led to restrictive legislation. The results obtained for PCB also point to a decrease in levels. PCB production began in 1929. These chemicals were used in hundreds of industrial and commercial applications, as they are non-flammable, chemically stable, and have electrical insulating properties. Their use in Europe was first regulated in 1976 when they were restricted to closed circuits. However, PCBs were not fully banned until 1987 (UNEP 2002).Total PCB production in some of the European countries (France, Italy and Spain) was in the range of 300.000 tn for the period 1954–84 (Tolosa et al. 1997). This was about 20% of the total worldwide production (1500000 tn) (De Voogt and Brinkman 1989). The few studies on temporal PCB trends in different matrices (mussels, sediment, vertebrates) from the Northwestern Mediterranean (Tolosa et al. 1997, Villeneuve et al. 1999, Aguilar and Borrell 2004), suggest that levels are tending to decline. The highest levels of PCBs compared to those of DDTs in northern Brazil could be due to a several sources of this pollutant on the coast of Sao Paulo (outputs outfall Santos Bay and an extensive industrial area (Alonso 2008)). The exposure to considerable PCBs levels of bottlenose dolphins from Florida coastal waters was mainly through the food web (Johnson-Restrepo et al. 2005).

These findings indicate that apex marine predators represented by dolphins that live in coastal and oceanic waters continue to sustain considerable PCB exposure in the environment. Our results on bottlenose dolphins, franciscana dolphins and harbor porpoise confi rm that PCBs and DDTs pollution in marine environment is declining only slowly. However, PCB decline in blubber of the species studied was found to be much less pronounced than that of DDTs.

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